Air Quality Criteria for Lead Volume I of II

United States October 2006
EPW600
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EPA/600/R-05/144aF
October 2006
Air Quality Criteria for Lead
Volume I
National Center for Environmental Assessment-RTF Division
Office of Research and Development
U.S. Environmental Protection Agency
Research Triangle Park, NC
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PREFACE
National Ambient Air Quality Standards (NAAQS) are promulgated by the United States
Environmental Protection Agency (EPA) to meet requirements set forth in Sections 108 and 109
of the U.S. Clean Air Act. Those two Clean Air Act sections require the EPA Administrator
(1) to list widespread air pollutants that reasonably may be expected to endanger public health or
welfare; (2) to issue air quality criteria for them that assess the latest available scientific
information on nature and effects of ambient exposure to them; (3) to set "primary" NAAQS to
protect human health with adequate margin of safety and to set "secondary" NAAQS to protect
against welfare effects (e.g., effects on vegetation, ecosystems, visibility, climate, manmade
materials, etc); and (5) to periodically review and revise, as appropriate, the criteria and NAAQS
for a given listed pollutant or class of pollutants.
Lead (Pb) was first listed in the mid-1970's as a "criteria air pollutant" requiring NAAQS
regulation. The scientific information pertinent to Pb NAAQS development available at the time
was assessed in the EPA document Air Quality Criteria for Lead; published in 1977. Based on
the scientific assessments contained in that 1977 lead air quality criteria document (1977 Lead
AQCD), EPA established a 1.5 |ig/m3 (maximum quarterly calendar average) Pb NAAQS in
1978.
To meet Clean Air Act requirements noted above for periodic review of criteria and
NAAQS, new scientific information published since the 1977 Lead AQCD was later assessed in
a revised Lead AQCD and Addendum published in 1986 and in a Supplement to the 1986
AQCD/Addendum published by EPA in 1990. A 1990 Lead Staff Paper, prepared by EPA's
Office of Air Quality Planning and Standards (OPQPS), drew upon key findings and conclusions
from the 1986 Lead AQCD/Addendum and 1990 Supplement (as well as other OAQPS-
sponsored lead exposure/risk analyses) in posing options for the EPA Administrator to consider
with regard to possible revision of the Pb NAAQS. However, EPA chose not to revise the Pb
NAAQS at that time. Rather, as part of implementing a broad 1991 U.S. EPA Strategy for
Reducing Lead Exposure, the Agency focused primarily on regulatory and remedial clean-up
efforts to reduce Pb exposure from a variety of non-air sources that posed more extensive public
health risks, as well as other actions to reduce air emissions.
The purpose of this revised Lead AQCD is to critically assess the latest scientific
information that has become available since the literature assessed in the 1986 Lead
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AQCD/Addendum and 1990 Supplement, with the main focus being on pertinent new
information useful in evaluating health and environmental effects of ambient air lead exposures.
This includes discussion in this document of information regarding: the nature, sources,
distribution, measurement, and concentrations of lead in the environment; multimedia lead
exposure (via air, food, water, etc.) and biokinetic modeling of contributions of such exposures
to concentrations of lead in brain, kidney, and other tissues (e.g., blood and bone concentrations,
as key indices of lead exposure).; characterization of lead health effects and associated exposure-
response relationships; and delineation of environmental (ecological) effects of lead. This final
version of the revised Lead AQCD mainly assesses pertinent literature published or accepted for
publication through December 2005.
The First External Review Draft (dated December 2005) of the revised Lead AQCD
underwent public comment and was reviewed by the Clean Air Scientific Advisory Committee
(CASAC) at a public meeting held in Durham, NC on February 28-March 1, 2006. The public
comments and CASAC recommendations received were taken into account in making
appropriate revisions and incorporating them into a Second External Review Draft (dated May,
2006) which was released for further public comment and CASAC review at a public meeting
held June 28-29, 2006. In addition, still further revised drafts of the Integrative Synthesis
chapter and the Executive Summary were then issued and discussed during an August 15, 2006
CASAC teleconference call. Public comments and CASAC advice received on these latter
materials, as well as Second External Review Draft materials, were taken into account in making
and incorporating further revisions into this final version of this Lead AQCD, which is being
issued to meet an October 1, 2006 court-ordered deadline. Evaluations contained in the present
document provide inputs to an associated Lead Staff Paper prepared by EPA's Office of Air
Quality Planning and Standards (OAQPS), which poses options for consideration by the EPA
Administrator with regard to proposal and, ultimately, promulgation of decisions on potential
retention or revision, as appropriate, of the current Pb NAAQS.
Preparation of this document has been coordinated by staff of EPA's National Center for
Environmental Assessment in Research Triangle Park (NCEA-RTP). NCEA-RTP scientific
staff, together with experts from academia, contributed to writing of document chapters. Earlier
drafts of document materials were reviewed by scientists from other EPA units and by non-EPA
experts in several public peer consultation workshops held by EPA in July/August 2005.
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NCEA acknowledges the valuable contributions provided by authors, contributors, and
reviewers and the diligence of its staff and contractors in the preparation of this document. The
constructive comments provided by public commenters and CASAC that served as valuable
inputs contributing to improved scientific and editorial quality of the document are also
acknowledged by NCEA.

DISCLAIMER
Mention of trade names or commercial products in this document does not constitute
endorsement or recommendation for use.
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Air Quality Criteria for Lead

VOLUME I

EXECUTIVE SUMMARY E-l

1. INTRODUCTION 1-1

2. CHEMISTRY, SOURCES, AND TRANSPORT OF LEAD 2-1

3. ROUTES OF HUMAN EXPOSURE TO LEAD AND OBSERVED
ENVIRONMENTAL CONCENTRATIONS 3-1

4. TOXICOKINETICS, BIOLOGICAL MARKERS, AND MODELS OF LEAD
BURDEN IN HUMANS 4-1

5. TOXICOLOGICAL EFFECTS OF LEAD IN LABORATORY ANIMALS
AND IN VITRO TEST SYSTEMS 5-1

6. EPIDEMIOLOGIC STUDIES OF HUMAN HEALTH EFFECTS
ASSOCIATED WITH LEAD EXPOSURE 6-1

7. ENVIRONMENTAL EFFECTS OF LEAD 7-1

8. INTEGRATIVE SYNTHESIS: MULTIMEDIA LEAD EXPOSURE,
HUMAN HEALTH EFFECTS, AND ECOSYSTEM EFFECTS 8-1
VOLUME II
CHAPTER 4 ANNEX (TOXICOKINETICS, BIOLOGICAL MARKERS, AND
MODELS OF LEAD BURDEN IN HUMANS) AX4-1

CHAPTER 5 ANNEX (TOXICOLOGICAL EFFECTS OF LEAD IN
LABORATORY ANIMALS AND IN VITRO TEST SYSTEMS) AX5-1

CHAPTER 6 ANNEX (EPIDEMIOLOGIC STUDIES OF HUMAN HEALTH
EFFECTS ASSOCIATED WITH LEAD EXPOSURE) AX6-1

CHAPTER 7 ANNEX (ENVIRONMENTAL EFFECTS OF LEAD) AX7-1
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Table of Contents
Page

Preface I-ii
Disclaimer I-iv
List of Tables I-xxii
List of Figures I-xxviii
Authors, Contributors, and Reviewers I-xxxvii
U.S. Environmental Protection Agency Project Team 1-1
U.S. Environmental Protection Agency Science Advisory Board (SAB)
Staff Office Clean Air Scientific Advisory Committee (CASAC) I-lii
Abbreviations and Acronyms I-liv

EXECUTIVE SUMMARY E-l

1. INTRODUCTION 1-1
1.1 LEGAL AND HISTORICAL BACKGROUND 1-1
1.1.1 Legislative Requirements 1-1
1.1.2 Criteria andNAAQS Review Process 1-3
1.1.3 Regulatory Chronology 1-4
1.2 CURRENT LEAD CRITERIA ANDNAAQS RE VIEW 1-7
1.2.1 Procedures and Key Milestones for Document Preparation 1-7
1.3 ORGANIZATIONAL STRUCTURE AND CONTENT OF
THE DOCUMENT 1-10
1.3.1 Ascertainment of Literature and General Document Format 1-10
1.3.2 Organization and Content of the Document 1-11
REFERENCES 1-13

2. CHEMISTRY, SOURCES, AND TRANSPORT OF LEAD 2-1
2.1 PHYSICAL AND CHEMICAL PROPERTIES OF LEAD 2-1
2.2 SOURCES OF LEAD 2-13
2.2.1 Natural Sources 2-13
2.2.2 Lead Emissions in the United States 2-16
2.2.3 Stationary Sources 2-20
2.2.4 Mobile Sources 2-44
2.3 TRANSPORT WITHIN THE ENVIRONMENT 2-52
2.3.1 Atmospheric Transport of Lead Particles 2-52
2.3.2 Deposition of Airborne Particles 2-55
2.3.3 Resuspension of Lead-Containing Soil and Dust Particles 2-62
2.3.4 Runoff from Impervious Surfaces 2-66
2.3.5 Leaching of Soil Lead 2-70
2.3.6 Transport in Aquatic Systems 2-74
2.3.7 Plant Uptake 2-77
2.3.8 Routes of Exposure for Livestock and Wildlife 2-78
2.4 METHODS FOR MEASURING ENVIRONMENTAL LEAD 2-80
2.5 SUMMARY 2-81
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Table of Contents
(cont'd)
REFERENCES 2-85

3. ROUTES OF HUMAN EXPOSURE TO LEAD AND OBSERVED
ENVIRONMENTAL CONCENTRATIONS 3-1
3.1 EXPOSURE: AIR 3-3
3.1.1 Routine Monitoring of Lead in U.S. Ambient Air 3-4
3.1.2 Observed Concentrations- Indoor Air 3-14
3.1.3 Observed Concentrations - Occupational 3-15
3.2 EXPOSURE: SOIL AND DUST 3-18
3.2.1 Concentrations of Soil Lead in Urban Areas 3-19
3.2.2 Soil-Lead Concentrations Near Stationary Sources 3-21
3.2.3 Observed Concentrations - House Dust 3-27
3.2.4 Concentrations of Lead in Road Dust 3-32
3.3 EXPOSURE: DRINKING WATER 3-33
3.4 EXPOSURE: DIETARY INTAKE 3-41
3.5 LEAD-BASED PAINT 3-49
3.6 OTHER ROUTES OF EXPOSURE 3-50
3.6.1 Calcium Supplements 3-50
3.6.2 Glazes 3-50
3.6.3 Miniblinds 3-51
3.6.4 Hair Dye 3-51
3.6.5 Other Potential Sources of Lead Exposure 3-51
3.7 MEASUREMENT METHODS 3-52
3.8 SUMMARY 3-53
REFERENCES 3-55

4. TOXICOKINETICS, BIOLOGICAL MARKERS, AND MODELS OF
LEAD BURDEN IN HUMANS 4-1
4.1 INTRODUCTION 4-1
4.2 TOXICOKINETICS OF LEAD 4-2
4.2.1 Absorption of Lead 4-3
4.2.2 Distribution 4-13
4.2.3 Metabolism 4-17
4.2.4 Excretion 4-18
4.3 BIOLOGICAL MARKERS OF LEAD BODY BURDENS
AND EXPOSURE 4-19
4.3.1 Lead in Blood 4-19
4.3.1.1 Summary of Key Findings from the 1986 Lead
AQCD 4-19
4.3.1.2 Analytical Methods for Measuring Lead in Blood 4-20
4.3.1.3 Levels of Lead in Blood 4-21
4.3.1.4 Blood Lead as aBiomarker of Lead Body Burden 4-24
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Table of Contents
(cont'd)
4.3.1.5 Blood Lead as aBiomarker of Lead Exposure 4-29
4.3.1.6 Summary of Blood Lead as a Biomarker of Lead
Body Burden and Exposure 4-34
4.3.2 Lead in Bone 4-34
4.3.2.1 Summary of Key Findings from the 1986 Lead
AQCD 4-34
4.3.2.2 Methodology of Bone Lead Analysis 4-35
4.3.2.3 Bone Lead as aBiomarker of Lead Body Burden 4-41
4.3.2.4 Distribution of Lead from Bone into Blood and
Plasma 4-43
4.3.2.5 Mobilization of Lead From Bone 4-47
4.3.2.6 Summary of Bone Lead as a Biomarker of Lead
Body Burden and Exposure 4-52
4.3.3 Lead in Teeth 4-52
4.3.3.1 Summary of Key Findings from the 1986 Lead
AQCD 4-52
4.3.3.2 Analytical Methods for Measuring Lead in Teeth 4-53
4.3.3.3 Tooth Lead as aBiomarker of Lead Body Burden 4-54
4.3.3.4 Relationship Between Tooth Lead and Blood Lead 4-54
4.3.3.5 Mobilization of Lead from Teeth 4-55
4.3.3.6 Summary of Tooth Lead as a Biomarker of Lead
Body Burden and Exposure 4-56
4.3.4 Lead in Urine 4-56
4.3.4.1 Summary of Key Findings from the 1986 Lead
AQCD 4-56
4.3.4.2 Analytical Methods for Measuring Lead in Urine 4-56
4.3.4.3 Levels of Lead in Urine 4-57
4.3.4.4 Urine Lead as aBiomarker of Lead Body Burden 4-58
4.3.4.5 Relationship Between Lead in Blood and Urine 4-60
4.3.4.6 Summary of Urine Lead as a Biomarker of Lead
Body Burden and Exposure 4-63
4.3.5 Lead in Hair 4-63
4.3.5.1 Summary of Key Findings from the 1986 Lead
AQCD 4-63
4.3.5.2 Analytical Methods for Measuring Lead in Hair 4-64
4.3.5.3 Levels of Lead in Hair 4-64
4.3.5.4 Hair Lead as aBiomarker of Lead Body Burden 4-64
4.3.5.5 Hair Lead as aBiomarker of Lead Exposure 4-65
4.3.5.6 Summary of Hair Lead as a Biomarker of Lead
Body Burden and Exposure 4-65
4.4 MODELING LEAD EXPOSURE AND TISSUE DISTRIBUTION
OF LEAD 4-65
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Table of Contents
(cont'd)
4.4.1 Introduction 4-65
4.4.2 Slope Factor Models 4-69
4.4.3 Empirical Models of Lead Exposure-Blood Lead
Relationships 4-72
4.4.4 Historic Overview of Mechanistic Models of Lead
Biokinetics 4-88
4.4.4.1 Rabinowitz Model 4-88
4.4.4.2 Marcus Model(s) 4-90
4.4.4.3 Bert Model 4-91
4.4.4A Contemporary Models 4-92
4.4.5 Integrated Exposure Uptake Biokinetic (IEUBK) Model for
Lead in Children 4-95
4.4.5.1 Model Structure 4-95
4.4.5.2 Model Calibration and Evaluation 4-101
4.4.5.3 Model Applications 4-105
4.4.5.4 Implementation Code 4-106
4.4.6 Leggett Model 4-106
4.4.6.1 Model Structure 4-106
4.4.6.2 Model Calibration and Evaluation 4-111
4.4.6.3 Model Applications 4-111
4.4.6.4 Implementation Code 4-112
4.4.7 O'Flaherty Model 4-112
4.4.7.1 Model Structure 4-112
4.4.7.2 Model Calibration and Evaluation 4-115
4.4.7.3 Model Applications 4-116
4.4.7.4 Implementation Code 4-117
4.4.8 EPA All Ages Lead Model 4-117
4.4.8.1 Model Structure 4-117
4.4.9 Model Comparisons 4-119
4.4.10 Conclusions and Future Directions 4-127
4.5 SUMMARY 4-130
REFERENCES 4-135

5. TOXICOLOGICAL EFFECTS OF LEAD IN LABORATORY ANIMALS
AND IN VITRO TEST SYSTEMS 5-1
5.1 INTRODUCTION 5-1
5.2 EFFECTS OF LEAD ON HEME SYNTHESIS 5-2
5.2.1 Effects of Lead on Erythrocyte Biology and Function 5-2
5.2.2 Effects of Lead on Erythrocyte Functions 5-3
5.2.3 Effects of Lead on Erythrocyte Heme Metabolism 5-9
5.2.4 Effects of Lead on Other Hematological Parameters 5-11
5.2.5 Effects of Lead on Erythrocyte Enzymes 5-12
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Table of Contents
(cont'd)
5.2.6 Erythrocyte Lipid Peroxidation and Antioxidant Defense 5-16
5.2.7 Summary 5-17
5.3 NEUROLOGIC/NEUROBEHAVIORAL EFFECTS OF LEAD 5-18
5.3.1 Introduction 5-18
5.3.2 Neurochemical Alterations Resulting from Lead Exposure 5-20
5.3.3 Actions of Lead Exposure Defined by Neurophysiologic
Approaches 5-28
5.3.4 Lead Exposure and Sensory Organ Function 5-33
5.3.5 Neurobehavioral Toxicity Resulting from Lead Exposure 5-36
5.3.6 Lead-Induced Changes in Cellular Development and
Disposition of the Metal 5-61
5.3.7 Susceptibility and Vulnerability Factors Modifying Lead
Exposure and Thresholds for CNS Effects 5-66
5.3.8 Summary 5-71
5.4 REPRODUCTIVE AND DEVELOPMENTAL EFFECTS OF LEAD 5-74
5.4.1 Summary of Key Findings on the Developmental and
Reproductive Effects of Lead in Animals from the 1986
LeadAQCD 5-74
5.4.2 Effects on Male Reproductive Function 5-76
5.4.2.1 Effects on Male Sexual Development and
Maturation 5-77
5.4.2.2 Effects on Male Fertility: Effects on Sperm
Production and Function 5-81
5.4.2.3 Effects on Male Sex Endocrine System 5-82
5.4.2.4 Effects on Morphology and Histology of
Male Sex Organs 5-83
5.4.3 Effects on Female Reproductive Function 5-84
5.4.3.1 Effects on Female Sexual Development and
Maturation 5-85
5.4.3.2 Effects on Female Fertility 5-88
5.4.3.3 Effects on the Female Sex Endocrine System
and Menstrual Cycle 5-88
5.4.3.4 Effects on Morphol ogy and Hi stol ogy of
Female Sex Organs and the Placenta 5-89
5.4.4 Effects on Embryogenesis 5-90
5.4.4.1 Embryo/Fetal Mortality 5-90
5.4.4.2 Effects on embryo/fetal morphology 5-91
5.4.5 Effects on Growth and Endocrine Regulation of Growth 5-97
5.4.6 Effects on Other Endocrine Systems during Development 5-98
5.4.7 Effects on Other Organ Systems during Development 5-98
5.4.7.1 Developmental Effects on Blood and Liver 5-98
5.4.7.2 Developmental Effects on Skin 5-99
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Table of Contents
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5.4.7.3 Developmental Effects on the Retina 5-100
5.4.8 Summary 5-100
5.5 CARDIOVASCULAR EFFECTS OF LEAD 5-102
5.5.1 Introduction 5-102
5.5.2 Lead Exposure and Arterial Pressure in Experimental
Animals 5-103
5.5.2.1 Effect of Lead on Production of Reactive
Oxygen Species and Nitric Oxide Metabolism 5-103
5.5.2.2 Protein Kinase C, Inflammation, NFKB
Activation, and Apoptosis 5-109
5.5.2.3 Effect of Lead Exposure on the Adrenergic
System 5-112
5.5.2.4 Effects of Lead on the Renin-Angiotensin-
Aldosterone (RAAS) and KininergicSystems 5-113
5.5.3 Effects of Lead Exposure on Vasomodulators 5-114
5.5.4 Effects of Lead on Vascular Reactivity 5-116
5.5.5 Lead-Calcium Interactions in Vascular Tissue 5-117
5.5.6 Cardiotoxicity and Atherogenesis 5-118
5.5.7 Effects of Lead on Endothelial Cells 5-119
5.5.8 Effects of Lead on Vascular Smooth Muscle Cells 5-122
5.5.9 Summary 5-123
5.6 GENOTOXIC AND CARCINOGENIC EFFECTS OF LEAD 5-125
5.6.1 Introduction 5-125
5.6.2 Carcinogenesis Studies 5-125
5.6.2.1 Human Studies 5-125
5.6.2.2 Laboratory Animal Studies 5-127
5.6.2.3 Cell Culture Studies 5-129
5.6.2.4 Organ-Specific Studies 5-132
5.6.2.5 Carcinogenesis Summary 5-132
5.6.3 Genotoxicity Studies 5-132
5.6.3.1 Human Studies 5-132
5.6.3.2 Laboratory Animal Studies 5-134
5.6.3.3 Cell Culture Studies 5-136
5.6.3.4 Animal Cell Cultures 5-139
5.6.3.5 Cell-Free Studies 5-140
5.6.3.6 Organ-Specific Studies 5-141
5.6.3.7 Genotoxicity Section Summary 5-141
5.6.4 Genotoxicity as it Pertains to Potential Developmental Effects 5-141
5.6.5 Epigenetic Effects and Mixture Interactions 5-142
5.6.5.1 Gene Expression 5-142
5.6.5.2 DNA Repair 5-143
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Table of Contents
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5.6.5.3 Mitogenesis 5-144
5.6.5.4 Epigenetic Mechanisms Summary 5-145
5.6.6 Summary 5-145
5.7 LEAD AND THE KIDNEY 5-147
5.7.1 Review of Earlier Work 5-147
5.7.2 Markers of Renal Toxicity 5-149
5.7.3 Biochemical Mechanisms of Lead Toxicity 5-150
5.7.4 Animal Studies 5-152
5.7.4.1 Lead Toxicokinetics 5-152
5.7.4.2 Pathology, Ultrastructural, and Functional Studies 5-153
5.7.4.3 Biochemical Mechanisms of Lead Toxicity 5-159
5.7.4.4 Effect of Age on Lead Toxicity 5-176
5.7.5 Summary 5-177
5.8 EFFECTS ON BONE AND TEETH 5-179
5.8.1 Biology ofBone and Bone Cells 5-179
5.8.2 Summary of Information Presented in the 1986 Lead AQCD 5-180
5.8.3 Bone Growth in Lead-Exposed Animals 5-181
5.8.4 Regulation of Bone Cell Function in Animals-
Systemic Effects of Lead 5-184
5.8.4.1 Hypercalcemia/Hyperphosphatemia 5-184
5.8.4.2 Vitamin D[1,25-(OH)2D3] 5-184
5.8.4.3 Parathyroid Hormone 5-185
5.8.4.4 Growth Hormone 5-186
5.8.5 Bone Cell Cultures Utilized to Test the Effects of Lead 5-187
5.8.5.1 Bone Organ Culture 5-187
5.8.5.2 Primary Cultures of Osteoclasts and Osteoblasts 5-187
5.8.5.3 Rat Osteosarcoma Cell Line (ROS 17/2.8) 5-187
5.8.5.4 Human Osteosarcoma Cells (HOS TE 85) 5-191
5.8.5.5 Chick Chondrocytes 5-191
5.8.6 Bone Lead as a Potential Source of Toxicity in Altered
Metabolic Conditions 5-192
5.8.6.1 Pregnancy and Lactation 5-193
5.8.6.2 Age/Osteoporosis 5-197
5.8.6.3 Weight Loss 5-198
5.8.7 Teeth - Introduction 5-199
5.8.8 Uptake of Lead by Teeth 5-201
5.8.9 Effects of Lead on Enamel and Dentine Formation 5-201
5.8.10 Effects of Lead on Dental Pulp Cells 5-203
5.8.11 Adverse Effects of Lead on Teeth—Dental Caries 5-204
5.8.12 Lead from Teeth as a Potential Source of Toxicity 5-205
5.8.13 Summary 5-206
5.9 EFFECTS OF LEAD ON THE IMMUNE SYSTEM 5-208
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Table of Contents
(cont'd)
5.9.1 Introduction 5-208
5.9.2 Host Resistance 5-208
5.9.2.1 Viral Diseases 5-209
5.9.2.2 Bacterial Diseases 5-209
5.9.2.3 Parasitic Diseases 5-210
5.9.2.4 Tumors 5-211
5.9.3 Humoral Immunity 5-211
5.9.3.1 General Effects on B lymphocytes and
Immunoglobulins 5-212
5.9.3.2 IgE Alterations 5-213
5.9.4 General Effects on Thymocytes and T lymphocytes 5-216
5.9.4.1 Delayed Type Hypersensitivity 5-217
5.9.4.2 Other T-Dependent Cell-Mediated Immune Changes 5-219
5.9.5 Lymphocyte Activation and Responses 5-221
5.9.5.1 Activation by Mitogens 5-221
5.9.5.2 Activation via Other Receptors 5-222
5.9.5.3 Cytokine Production 5-224
5.9.6 Macrophage Function 5-227
5.9.6.1 Nitric Oxide (NO) Production 5-228
5.9.7 Granulocytes and Natural Killer (NK) Cells 5-235
5.9.8 Hypersensitivity and Autoimmunity 5-236
5.9.9 Mechanism of Lead-Based Immunomodulation 5-238
5.9.10 Age-Based Differences in Sensitivity 5-240
5.9.11 Summary 5-243
5.10 EFFECTS OF LEAD ON OTHER ORGAN SYSTEMS 5-246
5.10.1 Effects of Lead on the Hepatic System 5-246
5.10.1.1 Hepatic Drug Metabolism 5-247
5.10.1.2 Biochemical and Molecular Perturbations
in Lead-Induced Liver Tissue Injury 5-251
5.10.1.3 Effects of Lead Exposure on Hepatic
Cholesterol Metabolism 5-253
5.10.1.4 Effect of Lead on Hepatic Oxidative Stress 5-254
5.10.1.5 Lead-Induced Liver Hyperplasia:
Mediators and Molecular Mechanisms 5-256
5.10.1.6 Effects of Lead on Liver Heme Synthesis 5-261
5.10.2 Gastrointestinal System and Lead Absorption 5-262
5.10.2.1 Lead and In Vitro Cytotoxicity in Intestinal Cells 5-264
5.10.2.2 Alterations in Intestinal Physiology and
Ultrastructure 5-264
5.10.2.3 Intestinal Uptake and Transport 5-265
5.10.2.4 Alterations in Gastrointestinal Motility/
Gastrointestinal Transit and Function 5-266
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5.10.2.5 Lead, Calcium, and Vitamin D Interactions
in the Intestine 5-266
5.10.2.6 Lead and Intestinal Enzymes 5-268
5.10.3 Summary 5-268
5.11 LEAD-BINDING PROTEINS 5-270
5.11.1 Lead-binding Proteins within Intranuclear Inclusion
Bodies in Kidney 5-270
5.11.2 Cytoplasmic Lead-binding Proteins in Kidney and Brain 5-272
5.11.3 Lead-binding Proteins in Erythrocytes 5-275
5.11.4 Lead-binding Proteins in Rat Liver 5-278
5.11.5 Lead-binding Proteins in Intestine 5-278
5.11.6 Lead-binding Protein in Lung 5-281
5.11.7 Relationship of Lead-binding Protein to Metallothionein 5-281
5.11.8 Is ALAD an Inducible Enzyme and Is It the Principal
Lead-binding Protein in the Erythrocyte? 5-282
5.11.9 Summary 5-283
REFERENCES 5-285

6. EPIDEMIC-LOGIC STUDIES OF HUMAN HEALTH EFFECTS
ASSOCIATED WITH LEAD EXPOSURE 6-1
6.1 Introduction 6-1
6.1.1 Approach to Identifying Lead Epidemiologic Studies 6-2
6.1.2 Approach to Assessing Epidemiologic Evidence 6-2
6.1.3 Considerations in the Interpretation of Epidemiologic
Studies of Lead Health Effects 6-4
6.1.4 Approach to Presenting Lead Epidemiologic Evidence 6-6
6.2 NEUROTOXIC EFFECTS OF LEAD in children 6-7
6.2.1 Summary of Key Findings on Neurotoxic Effects of
Lead in Children from 1986 Lead AQCD and Addendum,
and 1990 Supplement 6-8
6.2.2 Introduction to Neurotoxic Effects of Lead in Children 6-9
6.2.3 Neurocognitive Ability 6-11
6.2.3.1 Prospective Longitudinal Cohort Studies of
Neurocognitive Ability 6-11
6.2.3.2 Cross-sectional Studies of Neurocognitive
Ability 6-31
6.2.3.3 Meta-analyses of Studies of Neurocognitive
Abilities 6-34
6.2.4 Measures of Academic Achievement 6-36
6.2.5 Measures of Specific Cognitive Abilities 6-41
6.2.6 Disturbances in Behavior, Mood, and Social Conduct 6-44
6.2.7 Sensory Acuities 6-50
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6.2.8 Neuromotor Function 6-51
6.2.9 Brain Anatomical Development and Activity 6-52
6.2.10 Gene-Environment Interactions in the Expression of Lead-associated
Neurodevelopmental Deficits 6-54
6.2.11 Persistence of Lead-Related Neurodevelopmental Deficits Associated
with Prenatal and Postnatal Exposure 6-56
6.2.12 Periods of Enhanced Developmental Susceptibility to
Central Nervous System Effects of Environmental Lead 6-60
6.2.13 Effect of Environmental Lead Exposure on
Neurodevelopment at the Lower Concentration Range 6-64
6.2.14 Selection and Validity of Neuropsychological Outcomes
in Children 6-71
6.2.15 Confounding, Causal Inference, and Effect Modification
of the Neurotoxic Effects of Lead in Children 6-73
6.2.16 Summary of the Epi demi ol ogi c Evi dence for the
Neurotoxic Effects of Lead in Children 6-76
6.3 Neurotoxic Effects of Lead in Adults 6-77
6.3.1 Summary of Key Findings on the Neurotoxic Effects of
Lead in Adults from the 1986 Lead AQCD 6-77
6.3.2 Overview of Cognitive and Psychomotor Tests Used to
Assess Adult Lead Exposure 6-78
6.3.3 Adult Environmental Lead Exposure Effects 6-80
6.3.3.1 Neurobehavioral Effects Associated with
Environmental Lead Exposure 6-80
6.3.3.2 Summary of Adult Environmental Lead
Exposure Effects 6-83
6.3.4 Adult Occupational Lead Exposure Effects 6-83
6.3.5 Amyotrophic Lateral Sclerosis and Other Neurological
Outcomes Associated with Lead in Adults 6-84
6.3.6 Summary of the Epi demi ol ogi c Evi dence for the
Neurotoxic Effects of Lead in Adults 6-87
6.4 RENAL EFFECTS OF LEAD 6-88
6.4.1 Summary of Key Findings on the Renal Effects of Lead
from the 1986 Lead AQCD 6-88
6.4.2 Renal Outcome Definitions 6-89
6.4.3 Lead Exposure Measure Definitions 6-90
6.4.4 Lead Nephrotoxicity in Adults 6-90
6.4.4.1 General Population Studies 6-90
6.4.4.2 Occupational Studies 6-99
6.4.4.3 Patient Population Studies 6-101
6.4.4.4 Mortality Studies 6-102
6.4.5 Lead Nephrotoxicity in Children 6-103
I-xv
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Table of Contents
(cont'd)
6.4.5.1 Studies in Adults Following Childhood Lead
Poisoning 6-103
6.4.5.2 Lead Body Burden in Children with Chronic
Renal Disease 6-104
6.4.5.3 Environmental Studies in Children 6-104
6.4.6 Mechanisms for Lead Nephrotoxicity 6-105
6.4.7 Susceptible Populations for Lead Nephrotoxicity 6-107
6.4.7.1 Chronic Medical Diseases 6-107
6.4.7.2 Age 6-107
6.4.7.3 Genetic Polymorphisms 6-108
6.4.8 Confounding of the Renal Effects of Lead by Other
Potential Risk Factors 6-110
6.4.8.1 Cadmium 6-110
6.4.9 Summary of the Epi demi ol ogi c Evi dence for the Renal
Effects of Lead 6-112
6.5 CARDIOVASCULAR EFFECTS OF LEAD 6-114
6.5.1 Summary of Key Findings of the Cardiovascular Effects
of Lead from the 1986 Lead AQCD and Addendum, and
1990 Supplement 6-114
6.5.2 Effects of Lead on Blood Pressure and Hypertension 6-116
6.5.2.1 Introduction 6-116
6.5.2.2 Blood Pressure and Hypertension Studies Using
Blood Lead as Exposure Index 6-118
6.5.2.3 Blood Pressure and Hypertension Studies Using
Bone Lead as Exposure Index 6-133
6.5.3 Other Cardiovascular Outcomes 6-138
6.5.3.1 Ischemic Heart Disease 6-138
6.5.3.2 Cardiovascular/Circulatory Mortality 6-140
6.5.3.3 Other Cardiovascular Effects 6-143
6.5.4 Lead and Cardiovascular Function in Children 6-144
6.5.5 Potential Confounding of the Cardiovascular Effects of Lead 6-146
6.5.5.1 Confounding by Copollutants 6-146
6.5.5.2 Confounding by Smoking Status 6-147
6.5.5.3 Confounding by Alcohol Consumption 6-147
6.5.5.4 Confounding by Dietary Calcium Intake 6-148
6.5.5.5 Summary of Potential Confounding of the
Lead Effect on Cardiovascular Health 6-150
6.5.6 Gene-lead Interactions 6-150
6.5.7 Summary of the Epi demi ol ogi c Evi dence for the
Cardiovascular Effects of Lead 6-153
6.6 Reproductive and Developmental Effects of Lead 6-155
I-xvi
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Table of Contents
(cont'd)
6.6.1 Summary of Key Findings of the Reproductive and
Developmental Effects of Lead from the 1986 Lead AQCD 6-155
6.6.2 Placental Transfer of Lead 6-156
6.6.3 Effects of Lead on Reproductive Function 6-158
6.6.3.1 Effects on Male Reproductive Function 6-158
6.6.3.2 Effects on Female Reproductive Function 6-166
6.6.4 Spontaneous Abortion 6-166
6.6.4.1 Spontaneous Abortion and Maternal Exposure to Lead 6-166
6.6.4.2 Spontaneous Abortion and Paternal Exposure to Lead 6-169
6.6.5 Fetal Growth 6-170
6.6.6 Preterm Delivery 6-175
6.6.7 Congenital Abnormalities 6-177
6.6.8 Summary of the Epi demi ol ogi c Evi dence for the
Reproductive and Developmental Effects of Lead 6-179
6.7 Genotoxic and Carcinogenic Effects of Lead 6-180
6.7.1 Summary of Key Findings from the 1986 Lead AQCD 6-180
6.7.2 Summary of Key Findings by the International Agency
for Research on Cancer and the National Toxicology
Program 6-181
6.7.3 Genotoxicity of Lead 6-183
6.7.4 Meta-analyses of Lead and Cancer 6-185
6.7.5 Review of Specific Studies on the Carcinogenicity of Lead
Since the 1986 Lead AQCD 6-186
6.7.5.1 Introduction 6-186
6.7.5.2 Key Studies of Occupational Populations in the
United States 6-187
6.7.5.3 Key Studies of the General Population 6-189
6.7.5.4 Other Lead Studies 6-192
6.7.6 Confounding of Occupational Lead Studies Due to Other
Occupational Exposures: Arsenic, Cadmium 6-192
6.7.7 Confounding of Lead Studies: Smoking and Other Factors 6-193
6.7.8 Summary of Epidemiologic Evidence for the Genotoxic
and Carcinogenic Effects of Lead 6-194
6.8 Effects of Lead on the Immune System 6-195
6.8.1 Summary of Key Findings of the Effects of Lead on the
Immune System from the 1986 Lead AQCD 6-195
6.8.2 Host Resistance, Hypersensitivity, and Autoimmunity 6-196
6.8.3 Humoral Immunity 6-197
6.8.4 Cell-Mediated Immunity 6-204
6.8.5 Lymphocyte Function 6-210
6.8.6 Phagocyte (Macrophage and Neutrophil) Function 6-211
I-xvii
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Table of Contents
(cont'd)
6.8.7 Summary of the Epidemiologic Evidence for the Effects
of Lead on the Immune System 6-213
6.9 EFFECTS OF LEAD ON OTHER ORGAN SYSTEMS 6-214
6.9.1 Biochemical Effects of Lead 6-214
6.9.1.1 Summary of Key Findings of the Biochemical
Effects of Lead from the 1986 Lead AQCD 6-214
6.9.1.2 Heme Biosynthesis 6-216
6.9.1.3 Effects on Blood Lipids: Cholesterol 6-220
6.9.2 Effects of Lead on the Hematopoietic System 6-221
6.9.2.1 Summary of Key Findings of the Effects of Lead
on the Hematopoietic System from the 1986
Lead AQCD 6-221
6.9.2.2 Blood Hemoglobin Levels 6-222
6.9.2.3 Erythrocyte Volume and Number 6-224
6.9.2.4 Erythropoiesis 6-227
6.9.2.5 Other Effects on Erythrocyte Metabolism and
Physiology 6-231
6.9.3 Effects of Lead on the Endocrine System 6-232
6.9.3.1 Summary of Key Findings of the Effects of
Lead on the Endocrine System from the 1986
Lead AQCD 6-232
6.9.3.2 Thyroid Endocrine Function 6-233
6.9.3.3 Reproductive Endocrine Function 6-237
6.9.3.4 Pituitary and Adrenal Endocrine Function 6-241
6.9.3.5 Calcitropic Endocrine Function 6-241
6.9.4 Effects of Lead on the Hepatic System 6-244
6.9.4.1 Summary of Key Findings of the Effects of
Lead on the Hepatic System from the 1986
Lead AQCD 6-244
6.9.4.2 Nonspecific Hepatic Injury 6-245
6.9.4.3 Hepatic Cytochrome P450 Function 6-245
6.9.5 Effects of Lead on the Gastrointestinal System 6-247
6.9.5.1 Summary of Key Findings on the Effects of
Lead on the Gastrointestinal System from the
1986 Lead AQCD 6-247
6.9.5.2 Gastrointestinal Colic 6-247
6.9.6 Effects of Lead on Bone and Teeth 6-248
6.9.6.1 Summary of Key Findings of the Effects of
Lead on Bone and Teeth from the 1986
Lead AQCD 6-248
6.9.6.2 BoneToxicity 6-249
6.9.6.3 Dental Health 6-250
I-xviii
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Table of Contents
(cont'd)
6.9.7 Effects of Lead on Ocular Health 6-253
6.9.7.1 Summary of Key Findings of the Effects of Lead
on Ocular Health from the 1986 Lead AQCD 6-253
6.9.7.2 Ocular Effects 6-254
6.9.8 Summary of the Epidemiologic Evidence for the Effects
of Lead on Other Organ Systems 6-255
6.10 EPIDEMIOLOGIC CONSIDERATIONS AND SUMMARY OF
EVIDENCE FOR LEAD HEALTH EFFECTS 6-257
6.10.1 Introduction 6-257
6.10.2 Exposure and Outcome Assessment in Lead Epidemiologic
Studies 6-257
6.10.2.1 Assessment of Lead Exposure and Body Burdens
Using Biomarkers 6-257
6.10.2.2 Assessment of Health Outcomes 6-262
6.10.3 Confounding of Lead Health Effects 6-263
6.10.3.1 Methods Used to Adjust for Confounding in
Epidemiologic Studies of Lead 6-263
6.10.3.2 Effects of Confounding Adjustment on Lead
Health Effect Estimates 6-265
6.10.4 Inferences of Causality 6-267
6.10.5 Summary of Key Findings and Conclusions Derived from
Lead Epidemiology Studies 6-268
REFERENCES 6-273

7. ENVIRONMENTAL EFFECTS OF LEAD 7-1
7.1 TERRESTRIAL ECOSYSTEMS 7-1
7.1.1 Methodologies Used in Terrestrial Ecosystem Research 7-2
7.1.2 Distribution of Atmospherically Delivered Lead
in Terrestrial Ecosystems 7-4
7.1.3 Species Response/Mode of Action 7-8
7.1.4 Exposure/Response of Terrestrial Species 7-11
7.1.5 Effects of Lead on Natural Terrestrial Ecosystems 7-14
7.2 AQUATIC ECOSYSTEMS 7-17
7.2.1 Methodologies Used in Aquatic Ecosystem Research 7-18
7.2.2 Distribution of Lead in Aquatic Ecosystems 7-21
7.2.3 Species Response/Mode of Action 7-24
7.2.4 Exposure/Response of Aquatic Species 7-29
7.2.5 Effects of Lead on Natural Aquatic Ecosystems 7-31
7.3 CRITICAL LOADS FOR LEAD IN TERRESTRIAL AND
AQUATIC ECOSYSTEMS 7-32
7.3.1 Definitions 7-33
7.3.2 Historical Perspective 7-33
I-xix
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Table of Contents
(cont'd)
7.3.3 Application of Critical Loads to Terrestrial and
Aquatic Ecosystems 7-35
7.3.4 Calculation of Critical Loads 7-37
7.3.4.1 Critical Limits 7-37
7.3.4.2 Models 7-38
7.3.5 Critical Loads in Terrestrial Ecosystems 7-42
7.3.6 Critical Loads in Aquatic Ecosystems 7-44
7.3.7 Limitations and Uncertainties 7-45
7.3.8 Conclusions 7-46
REFERENCES 7-47

INTEGRATIVE SYNTHESIS: MULTIMEDIA LEAD EXPOSURE,
HUMAN HEALTH EFFECTS, AND ECOSYSTEM EFFECTS 8-1
8.1 INTRODUCTION 8-1
8.1.1 Historical Background 8-1
8.1.2 Chapter Organization 8-3
8.2 OVERVIEW OF MULTIMEDIA LEAD, SOURCES, EMISSIONS,
AND CONCENTRATIONS IN THE UNITED STATES 8-3
8.2.1 Sources of Lead Emissions into Ambient Air 8-4
8.2.2 Transport and Secondary Dispersal of Atmospheric Lead 8-7
8.2.3 Ambient Air Lead Concentrations 8-10
8.2.4 Non-air Environmental Lead Exposure Routes 8-12
8.3 TOXICOKINETICS, BIOLOGICAL MARKERS, AND MODELS OF
LEAD BURDEN IN HUMANS 8-15
8.3.1 Biokinetics of Lead Uptake and Internal Distribution 8-16
8.3.2 Selection of Blood-Lead Concentration as Key Index of
Lead Exposure 8-17
8.3.3 Trends in U.S. Blood Lead Levels 8-19
8.3.4 Approaches to Predictive Estimation of Pb-Exposure Impacts
on Distribution to Internal Tissues 8-21
8.4 LEAD-INDUCED TOXICITY: INTEGRATION OF TOXICOLOGIC
AND EPIDEMIOLOGIC EVIDENCE 8-24
8.4.1 Introduction 8-24
8.4.2 Neurotoxic Effects 8-25
8.4.2.1 Neurocognitive Ability 8-27
8.4.2.2 Behavior, Mood, and Social Conduct 8-31
8.4.2.3 Neurophysiologic Outcomes 8-33
8.4.2.4 Neuromotor Function and Vocalization 8-35
8.4.2.5 Neurochemical Alterations 8-36
8.4.2.6 Assessment of Dose-Response Relationships for
Neurotoxic Effects of Lead Exposure 8-37
I-xx
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Table of Contents
(cont'd)
8.4.2.7 Susceptibility and Vulnerability to Neurotoxic
Effects from Lead Exposure 8-39
8.4.2.8 Persistence/Reversibility of Neurotoxic
Effects from Lead Exposure 8-43
8.4.2.9 Summary of Toxi col ogi c and Epi demi ol ogi c
Evidence of Lead-Induced Neurotoxicity 8-44
8.4.3 Cardiovascular Effects 8-45
8.4.4 Heme Synthesis and Blood Effects 8-47
8.4.5 Renal System Effects 8-48
8.4.6 Lead-Associated Immune Outcomes 8-50
8.4.7 Reproduction and Development Effects 8-53
8.4.8 Bone and Teeth Effects 8-54
8.4.9 Hepatic and Gastrointestinal System Effects 8-56
8.4.10 Genotoxicity and Carcinogenicity 8-57
8.5 KEY LOW-LEVEL LEAD EXPOSURE HEALTH EFFECTS AND
IDENTIFICATION OF FACTORS THAT AFFECT SUSCEPTIBILITY
TO LEAD TOXICITY 8-63
8.5.1 Concentration-Response Relationships for Neurotoxicity Effects 8-63
8.5.2 Persistence/Reversibility of Lead Neurotoxic Effects 8-67
8.5.3 Factors Affecting Susceptibility to Lead Toxicity 8-70
8.6 POTENTIAL PUBLIC HEALTH IMPLICATIONS OF LOW-LEVEL
LEAD EXPOSURE 8-75
8.6.1 Introduction 8-75
8.6.2 Potential Implications of Lead Effects on Intelligence 8-78
8.6.3 Potential Implications of Cardiovascular Effects of Lead 8-83
8.6.4 Potential Implications of Renal Effects of Lead 8-89
8.6.5 Potential Implications of Lead-Induced Immune System Effects 8-90
8.7 KEY LEAD ECOSYSTEM EFFECTS AND POTENTIAL
IMPLICATIONS 8-90
8.7.1 Terrestrial Ecosystems 8-90
8.7.2 Aquatic Ecosystems 8-101
8.7.3 Application of Critical Loads to Terrestrial and
Aquatic Ecosystems 8-110
REFERENCES 8-115
I-xxi
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List of Tables

Number Page

1-1 Key Milestones and Proj ected Schedule for Development of Revised Lead
Air Quality Criteria Document 1-8

2-1 Lead Alloys and Their Industrial Applications 2-3

2-2 Physical Properties of Elemental Lead 2-4

2-3 Lead Salts: Names, Formulae, Physical Characteristics, and Uses 2-6

2-4 Lead Oxides: Names, Formulae, Physical Characteristics, and Uses 2-7

2-5 Lead Compounds Observed in the Environment 2-8

2-6 Worldwide Annual Emissions of Lead from Natural Sources 2-13

2-7 Naturally Occurring Lead Concentrations in Maj or Rock Types 2-15

2-8 Annual Air Emission Rates for U.S. Lead Sources (20 Tons per Year or
Greater) for 1990 and 2002, Ordered by Emissions Levels in 2002 2-18

2-9 Mass-Median Aerodynamic Diameters for Particles from Various Processes
at Primary Lead Smelters 2-22

2-10 Emissions of Lead from Nonlead Metallurgical Processes 2-24

2-11 The Range of Lead Concentrations in Coal Lithotypes 2-28

2-12 Lead Emission Factors for Coal Combustion in Three Different Furnaces 2-30

2-13 Lead Emissions from Industrial, Commercial, and Residential
Coal Combustion 2-31

2-14 Lead Emission Factors for Oil-Fired Utility Boilers 2-32

2-15 Lead Concentrations in Biomass, Char, and Ash Samples from Spruce,
Beech, Oak, Pine, and of Ailanthus Trees 2-34

2-16 Emission Factors for Processes Used in Cement Manufacture by
Control Device 2-41

2-17 Rate of Lead Compound Emissions from Glass-Melting Furnaces 2-42
I-xxii
-------
List of Tables
(cont'd)

Number Page

2-18 Lead Emission Factors During Summer Versus Winter for Automobiles
withModel Years Between 1971 and 1996 2-46

2-19 Lead Emission Factors for Different Driving Phases for Automobiles with
Model Years Between 1971 and 1996 2-47

2-20 Lead Concentration in Particulate Matter Emissions and Lead Emissions
Factors for Buses and Trucks Fueled with Diesel No. 2 and Jet A Fuel 2-49

2-21 Dry Deposition Velocities for Lead Particles 2-58

2-22 Lead Concentrations in Rainwater in the United States 2-61

2-23 Observed Percentages of Lead in Resuspended Particulate Matter 2-66

2-24 Lead Concentrations Observed in Runoff From Building Surfaces 2-69

2-25 Soil/Water Partition Coefficients for Several Different Soils and Conditions 2-73

3-1 Descriptive Statistics for Lead Measurements (in |ig/m3) from Monitors
Using Different Size Fractions of PM for Recent Years 3-12

3-2 Maximum Quarterly Mean and Overall Average Quarterly Mean Lead
Measurements (in |ig/m3) from U.S. Monitors using Different Size Fractions
of PM for Recent Years 3-13

3-3 Airborne Lead Concentrations in Areas Surrounding Residential Lead-Based
Paint Abatement Activities 3-17

3-4 Concentrations of Soil Lead with Distance from Lead Smelters 3-22

3-5 Soil Lead Concentration Profile Measured Near a Lead Smelter in
Northern France 3-23

3-6 Soil Concentrations Measured Near Mining Sites 3-25

3-7 Concentrations of Lead in Soils Grouped by Soil Grain Size 3-26

3-8 Examples of Lead Concentrations and Dust Lead Loadings in Indoor Dust 3-30

3-9 Examples of Observed Road Dust Lead Concentrations 3-34
I-xxiii
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List of Tables
(cont'd)

Number Page

3-10 Examples of Tap Water Concentrations of Lead 3-40

3-11 90th Percentile Tap Water Lead Concentrations for a Selection of U.S. Cities
Exceeding the EPA Lead Action Level 3-41

3-12 Examples of Lead Concentrations in Food Products 3-44

4-1 Blood Lead Concentrations in United States by Age, NHANES IV
(1999-2002) 4-22

4-2 Blood Lead Concentrations in United States by Gender, NHANES IV
(1999-2002) 4-22

4-3 Blood Lead Concentrations by Occupation, NHANES III (1988-1994) 4-23

4-4 Urine Lead Concentrations in U.S. by Age, NHANES IV (1999-2002) 4-57

4-5 Recommended parameter values for the Adult Lead Methodology (ALM) and
corresponding risk-based soil Pb concentrations (RBCS) 4-71

4-6 General Linear Model Relating Blood Lead Concentration in Children
and Environmental Lead Levels—Bunker Hill Superfund Site 4-74

4-7 Structural Equation Model (1) Relating Blood Lead Concentration in Children
and Environmental Lead Levels—Bunker Hill Superfund Site 4-76

4-8 Structural Equation Model (2) Relating Blood Lead Concentration in Children
and Environmental Lead Levels—Bunker Hill Superfund Site 4-77

4-9 General Linear Model Relating Blood Lead Concentration in Children and
Environmental Lead Levels—Coeur d'Alene Basin 4-78

4-10 Multivariate Regression Model Relating Blood Lead Concentration in Children
and Environmental Lead Levels—Multi-study Pooled Analysis 4-79

4-11 Children's Predicted Blood Lead Levels for Floor Dust Lead Loading (|ig/ft2)
and Exterior Lead Exposures (ppm) 4-81

4-12 Likelihood of a Child's Blood Lead > 10 |ig/dL for Floor Dust Lead Loadings
and Exterior Exposure Levels (ppm) 4-82
I-xxiv
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List of Tables
(cont'd)
Number
4-13 Meta-analysis of the Relationship Between Log-transformed Blood Lead
and Various Environmental Lead Sources 4-84

4-14 Structural Equation Models Relating Blood Lead Concentration in Children
and Pre-abatement Environmental Lead Levels—Lead in Urban Soil
Abatement Demonstration Project 4-89

4-15 Comparison of Observed and Predicted Geometric Mean Blood Lead for
Three Community Blood Lead Studies 4-102

4-16 Comparison of Observed and Predicted Probability of Exceeding a Blood
Lead Concentration of 10 |ig/dL Lead for Three Community Blood
Lead Studies 4-102

4-17 Summary of Models of Human Exposure that Predict Tissue Distribution
of Lead 4-120

4-18 Inputs and Results of Simulations Comparing the U.S. EPA Adult Lead
Methodology (ALM) With Multicompartmental Models 4-128

5-1 Chronic Lead Exposure and LTP 5-29

5-2 Mechanisms of Lead-Induced Impairment of Retinal Function 5-36

5-3 Selected Studies Showing the Effects of Lead on Reproductive Function
in Males 5-78

5-4 Selected Studies Showing the Effects of Lead on Reproductive Function
in Females 5-86

5-5 Selected Studies Showing the Effects of Lead on Mammalian Embryogenesis
and Development 5-92

5-6 Recent Studies Reporting Lead-Induced Increase in IgE 5-214

5-7 Studies Reporting Lead-Induced Shifts in Thl versus Th2 Cytokines 5-225

5-8 Suggested Mechanisms of Lead-Induced Immunotoxicity 5-239

5-9 Immunomodulation Associated with Low Blood Lead Levels in Animals 5-241
I-xxv
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List of Tables
(cont'd)

Number

5-10 Comparisons of Age-Based Sensitivity to Lead-Induced Immunotoxicity 5-243

6-1 Summary of Studies with Quantitative Relationships of IQ and Blood Lead
for Blood Lead Levels Less than 10 |ig/dL 6-66

6-2 Summary of Studies with Quantitative Relationships of Systolic Blood
Pressure and Blood Lead 6-119

6-3 Results of Meta-Analyses Addressing the Association Between Lead
Exposure and Cancer 6-184

6-4 Results of Epidemiologic Studies on the Genotoxicity of Lead Exposure 6-185

6-5 Summary of Results of Selected Studies of Associations Between Lead
Exposure and Serum Immunoglobulin Levels 6-198

6-6 Summary of Results of Selected Studies of Associations Between Lead
Exposure and Serum Lymphocyte Abundances 6-206

6-7 Blood Lead-Response Relationships for Heme Synthesis Biomarkers in
Adults and Children 6-217

6-8 Summary of Results of Selected Studies of Associations Between Lead
Exposure and Blood Hemoglobin Levels 6-223

6-9 Summary of Results of Selected Studies of Associations Between Lead
Exposure and Serum Erythropoietin 6-228

6-10 Summary of Results of Selected Studies of Associations Between Lead
Exposure and Thyroid Hormone Levels 6-235

6-11 Summary of Results of Selected Studies of Associations Between Lead
Exposure and Male Sex Hormone Levels in Adults 6-238

6-12 Summary of Results of Selected Studies of Associations Between Lead
Exposure and Calcitropic Hormones 6-242

7-1 Summary of Lead Ambient Water Quality Criteria for Freshwater
Organisms at Different Hardness Levels 7-19
I-xxvi
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List of Tables
(cont'd)

Number Page

7-2 Summary of Sediment Quality Benchmarks and Guidelines for Lead 7-20

7-3 Summary of Lead Concentrations in United States Surface Water,
Sediment, and Fish Tissue 7-23

8-1 Descriptive Statistics for Lead Measurements (in |ig/m3) from Monitors
Using Different Size Fractions of PM for Recent Years 8-12

8-2 Estimated Dietary Lead Intake in U.S. Population Groups in 1982-1984
versus 1994-1996 8-15

8-3 Blood Lead Concentrations in United States by Age,
NHANES IV (1999-2002) 8-20

8-4 Blood Lead Concentrations in United States by Gender,
NHANES IV (1999-2002) 8-20

8-5 Summary of Lowest Observed Effect Levels for Key Lead-Induced
Health Effects in Children 8-61

8-6 Summary of Lowest Observed Effect Levels for Key Lead-Induced
Health Effects in Adults 8-62

8-7 Summary of Studies with Quantitative Relationships for IQ and Blood Lead 8-79

8-8 Summary of Lead Concentrations in United States Surface Water,
Sediment, and Fish Tissue 8-103
I-xxvii
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List of Figures

Number Page

E-l Simplified diagram of environmental pathways contributing to multimedia
Pb exposure of human populations E-4

2-1 Multiple possible molecular targets for interference by Pb2+ ion at
nerve synapses 2-12

2-2 Annual lead production and use in the United States (1968 - 2003) 2-17

2-3 Percentage volatility of lead during combustion of plastics at
four temperatures 2-37

2-4 Lead concentrations in sediment samples in 12 Michigan lakes 2-56

2-5 The deposition velocity plotted against the geometric mean Stokes diameter
for particles with a density of 6 g/cnT3 (i.e., lead) 2-59

2-6 Modeled soil concentrations of lead in the South Coast Air Basin of
California based on four resuspension rates 2-67

2-7 Modeled and measured airborne concentrations of lead in the South Coast
Air Basin of California based on two resuspension rates 2-67

2-8 Trends in U.S. air lead emissions, 1982-2002 2-81

2-9 Transport pathways for lead in the environment 2-84

3-1 Principal pathways of lead from the environment to humans 3-2

3-2 Airborne Pb concentrations measured at FRM sites, averaged across the
United States for the years 1983 through 2002 3-4

3-3 United States Lead TSP monitoring sites from 2000-2006 3-5

3-4a Locations monitored by the Speciation Trends Network 3-7

3-4b The average maximum quarterly mean Pb concentrations observed in PM2.5
bytheSTN 3-7

3-5a The Interagency Monitoring of Protected Visual Environments network of
PM2.5 monitors 3-8
I-xxviii
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List of Figures
(cont'd)

Number Page

3-5b IMPROVE sites with Pb PM2.5 concentrations at or above 0.0008 |ig/m3
between 2000 and 2005 3-9

3-6a The National Air Toxics Trends Stations network 3-10

3-6b Arithmetic mean of maximum quarterly average Pb concentrations measured
in PMio at NATTS network sites during 2002 through 2005 3-11

3-7 The changes in lead concentration with depth in two peat cores 3-27

3-8 The change in lead concentration versus stagnation time 3-38

3-9 Change in lead concentration versus stagnation time 3-39

4-1 Relative bioavailability (RBA) is the bioavailability of the lead in the test
material compared to that of lead acetate (test material/lead acetate) 4-10

4-2 Estimated relative bioavailability (RBA, compared to lead acetate) of ingested
lead in mineral groups, based on results from juvenile swine assays 4-11

4-3 Blood lead concentrations in U.S. children, 1-5 years of age 4-24

4-4 Simulation of relationship between blood lead concentration and body burden
in adults 4-26

4-5 Simulation of relationship between blood Lead concentration and body burden
in children 4-28

4-6 Simulation of temporal relationships between lead exposure and blood lead
concentration in children 4-31

4-7 Simulation of relationships between lead intake and blood lead concentration
in adults and children 4-33

4-8 Cortical lead to blood lead ratios for occupationally-exposed subjects (both
active and retired) and referents 4-45

4-9 Tibia lead to blood lead ratios for environmentally-exposed pregnancy-related
subj ects, middle-aged to elderly subj ects, and younger subj ects 4-47
I-xxix
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List of Figures
(cont'd)

Number Page

4-10 Simulation of relationship between urinary lead excretion and body burden
in adults 4-59

4-11 Simulation of relationship between lead intake and urinary lead excretion
in adults and children 4-61

4-12 Adult lead model predictions of the relationship between soil lead
concentration and 95th percentile fetal blood lead concentration 4-72

4-13 Structural equation model for relationships between dust and soil lead and
blood lead concentration in children, based on data collected at the Bunker
Hill Superfund Site (1988-1999) 4-75

4-14 Structural equation model for relationships between dust and soil lead and
blood lead concentration in children 4-83

4-15 Structural equation model for relationships between dust and soil lead and
blood lead concentration in children, based on data collected in the
Rochester (NY) Lead in Dust Study 4-85

4-16 Structural equation model for relationships between dust and soil lead and
blood lead concentration in children, based on data collected in the
Cincinnati (OH) Prospective Child Study 4-86

4-17 Structural equation model for relationships between dust and soil lead and
blood lead concentration in children, based on pre-abatement cross-sectional
data collected in the Urban Soil Lead Abatement Demonstration Project 4-88

4-18 Lead biokinetics based on Rabinowitz et al. (1976) 4-90

4-19 Lead biokinectics based on Marcus (1985a) 4-91

4-20 Lead biokinetics based on Marcus (1985b) 4-92

4-21 Lead biokinetics based on Marcus (1985c) 4-93

4-22 Lead biokinetics based on Bert etal. (1989) 4-94

4-23 Structure of the integrated exposure uptake biokinetics model for lead
in children 4-96
I-xxx
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List of Figures
(cont'd)

Number Page

4-24 Age-dependency of absorption fraction for ingested lead in the IEUBK
model for lead in children 4-99

4-25 Comparisons of IEUBK model predictions and observed blood lead
concentrations 4-103

4-26 Comparison of IEUBK model predictions and observed blood lead
concentrations 4-104

4-27 Structure of the Leggett Lead Biokinetic Model 4-107

4-28 Age-dependency of absorption fraction for ingested lead in the Leggett and
O'Flaherty models 4-110

4-29 Structure of the O'Flaherty Lead Exposure Biokinetics Model 4-113

4-30 Bone growth as simulated by the O'Flaherty Lead Exposure Biokinetics
Model 4-114

4-31 Structure of the All Ages Lead Model 4-118

4-32 Model comparison of predicted lead uptake—blood lead concentration
relationship in children 4-123

4-33 Model comparison of predicted lead uptake—blood lead concentration
relationships in adults 4-124

4-34 Model comparison of predicted of lead uptake—bone and soft tissue lead
burden relationship in adults 4-125

4-35 Comparison of model predictions for childhood lead exposure 4-126

4-36 Comparison of model predictions for adult lead exposure 4-126

5-1 Schematic presentation of the enzymatic steps involved in heme
synthesis pathway 5-10

5-2 Time course and magnitude of response of extracellular GLU concentration
as a result of chronic lead exposure 5-22

5-3 Simplified diagram showing the actions of lead at a synapse 5-26
I-xxxi
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List of Figures
(cont'd)

Number Page

5-4 PKC activity as a function of Ca2+ andPb2+ concentrations 5-27

5-5 Difference score measure of population spike amplitude 5-30

5-6 Dose-effect function for lead-induced changes in fixed-interval performance 5-42

5-7 Data from male and female experimental animals suggests that lead has
multiple targets in the hypothalmic-pituitary-gonadal axis 5-75

5-8 This illustration depicts some of the potential mechanisms by which
oxidative stress may participate in the pathogenesis of lead-induced HTN
and cardiovascular complications 5-110

5-9 Changes in GFR of experimental high-dose lead and control animals
with duration of exposure to lead 5-154

5-10 Correlation between GFR and blood lead during the first 6 months
of high-dose lead exposure 5-154

5-11 GFR in high-lead and low-lead experimental discontinuous (ED6)
and DMSA-treated rats (DMSA) as compared to controls (C12) 5-156

5-12 Changes in GFR in experimental and control rats, at various time periods 5-156

5-13 Urinary NAG concentration in experimental and control rats at
various time periods 5-157

5-14 Kidney, liver, brain, and bone lead levels in 56 Pb-exposed rats 5-158

5-15 Percentage of moderate and severe hypertrophy and vacuolization lesions
in small and medium sized arteries in the kidney of lead-exposed rats 5-160

5-16 Percentage of moderate and severe muscular hypertrophy lesions in
arterioles of the kidney in lead-exposed rats 5-160

5-17 Windows during prenatal development (days postconception for rat)
or embryonic development (days postincubation initiation for chicken)
during which sensitivity of DTH to lead emerges 5-220

5-18 This figure shows the fundamental alterations to the immune system and to
immunological response and recognition induced by exposure to lead 5-244
I-xxxii
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List of Figures
(cont'd)

Number Page
5-19 Flow diagram indicating the lead effects on the cholesterol synthesis pathway 5-255

5-20 Schematic diagram illustrating the mode of lead-induced lipid peroxidation 5-255

5-21 Hypothesis of chemical-induced liver injury generated primarily on the basis
of different types of inhibitors 5-260

5-22 Sephadex G-75 gel filtration of RBC hemolysate from lead-exposed
individual. Ultraviolet absorption and radioactivity of 210Pb are plotted
against elution volume 5-275

5-23 SDS-polyacrylamide gel electrophoresis of RBC hemolysates from
normal control (A) and lead-exposed individuals (B), and of low-mol-wt.
lead-binding protein (C) stained with coomassie blue 5-276

5-24 Chromatographic profiles of protein, ALAD activity and lead in human
erythrocytes incubated with 5% glucose solution containing lead acetate 5-279

5-25 Chromatic profiles of protein, ALAD activity, lead, and Se in the
erythrocytes of lead-exposed workers 5-280

6-1 Unadjusted and adjusted relationships between average lifetime blood lead
concentrations and Wechsler Scale performance IQ 6-16

6-2 Linear models for the 7 cohort studies in the pooled analysis, adjusted for
maternal IQ, HOME score, maternal education, and birth weight 6-30

6-3 Golgi-stained section of human cerebral cortex taken from equivalent areas
of the anterior portion of the middle frontal gyrus at different ages 6-61

6-4 Full scale IQ test scores by previous or concurrent blood lead concentration 6-63

6-5 Restricted cubic splines and log-linear model for concurrent blood
lead concentration 6-68

6-6 Log-linear model (95% CI shaded) for concurrent blood lead concentration
adjusted for HOME score, maternal education, maternal IQ, and birth weight 6-69
I-xxxiii
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List of Figures
(cont'd)

Number Page

6-7 Log-linear model for concurrent blood lead concentration along with linear
models for concurrent blood lead levels among children with peak blood lead
levels above and below 10 |ig/dL 6-69

6-8 Creatinine clearance versus blood lead slope at a blood lead of 5 |ig/dL 6-97

6-9 Effect on associations between lead dose and renal function depending
on whether effect modification (age in this example) is assessed 6-101

6-10 Change in the systolic pressure (effect estimate in mm Hg) associated
with a doubling of the blood lead concentration 6-131

6-11 Change in the diastolic pressure (effect estimate in mm Hg) associated
with a doubling of the blood lead concentration 6-132

6-12 Effect of doubling mean blood lead on estimate of blood pressure change
with95%CIs 6-134

6-13 Five-knot cubic spline regression models of total cancer mortality and blood
lead level by gender, based on analyses of theNHANES II cohort 6-190

6-14 Relationship between blood lead concentration (PbB), age, and serum IgE
level in children 6-200

6-15 Relationship between blood lead concentration and serum IgE level
in children 6-201

6-16 Relationship between blood lead concentration and serum
immunoglobulin (Ig) levels in children 6-202

6-17 Relationship between blood lead concentration and serum IgE level in
lead workers 6-203

6-18 Relationship between blood lead concentration and T- and B-cell
abundances in children 6-207

6-19 Relationship between lead exposure and T- and B-cell abundances in
firearms instructors 6-209
I-xxxiv
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List of Figures
(cont'd)

Number Page

6-20 Effects of lead on heme biosynthesis 6-215

6-21 Relationship between blood lead and hematocrit in children 6-226

6-22 Relationship between blood lead and serum erythropoietin in children 6-229

6-23 Association between blood lead concentration and serum erythropoietin
in pregnant women 6-231

7-1 The predicted development of metal concentrations in ecosystems for
four cases of exceedance or non-exceedance of critical limits and critical
loads of heavy metals, respectively 7-36

7-2 The relationship between the critical limit of Pb in soil as a function of
organic matter and pH 7-38

8-1 Principal pathways of lead from the environment to human consumption 8-4

8-2 Trends in U.S. air lead emissions during the 1982 to 2002 period 8-6

8-3 Lead concentrations in sediment samples in 12 Michigan lakes 8-8

8-4 Airborne Pb concentrations measured at FRM sites, averaged across
the United States for the years 1983 through 2002 8-11

8-5 Blood lead concentrations in U.S. children, 1-5 years of age 8-21

8-6 Comparison of a linear and log-linear model to describe the relationship
between exposure and response 8-64

8-7 Concentration-response relationships of IQ to blood lead for the individual
studies and the pooled analysis by Lanphear et al. (2005) 8-80

8-8 Mean blood lead levels adjusted for HOME score, maternal education,
maternal IQ, and birth weight from the pooled analysis of seven studies
by Lanphear et al. (2005) 8-81

8-9 Effect of blood lead on fraction of population with IQ levels
<80 or <70 points (A) and IQ levels >120 or >130 points (B) 8-82
I-xxxv
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List of Figures
(cont'd)

Number Page

8-10 Distribution of systolic blood pressure in women and men aged
35 to 64 years from the Framingham Heart Study (Kannel, 2000a) 8-85

8-11 Relationship of serious cardiovascular events (coronary disease, stroke,
peripheral artery disease, cardiac failure) to systolic blood pressure in
women and men aged 35 to 64 years from the Framingham
Heart Study (Kannel, 2000a) 8-86

8-12 Effect of blood lead on expected annual risk of cardiovascular events
per 1,000 person-years 8-87

8-13 The predicted development of metal concentrations in ecosystems for
four cases of exceedance or non-exceedance of critical limits and critical
loads of heavy metals, respectively 8-111
I-xxxvi
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Authors, Contributors, and Reviewers

CHAPTER 1 - INTRODUCTION
Principal Author

Dr. Lester D. Grant—National Center for Environmental Assessment (B243-01),
U.S. Environmental Protection Agency, Research Triangle Park, NC 27711
CHAPTER 2 - CHEMISTRY, SOURCES, TRANSPORT OF LEAD
Chapter Managers/Editors

Dr. Brooke L. Hemming—National Center for Environmental Assessment (B243-01),
U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

Dr. Mary Ross—National Center for Environmental Assessment (B243-01),
U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

Principal Authors

Dr. Brooke L. Hemming—National Center for Environmental Assessment (B243-01),
U.S. Environmental Protection Agency, Research Triangle Park, NC 27711 (Section 2.1)

Ms. Allison Harris—Department of Civil and Environmental Engineering, Carnegie-Mellon
University, Pittsburgh, PA 15213 (Sections 2.2 - 2.4)

Professor Cliff Davidson—Department of Civil and Environmental Engineering,
Carnegie-Mellon University, Pittsburgh, PA 15213 (Sections 2.2 - 2.4)

Contributors and Reviewers

Professor Brian Gulson—Graduate School of the Environment, Macquarie University
Sydney, NSW 2109, Australia

Professor John W. Winchester (Emeritus)—Dept. of Oceanography, Florida State University,
Tallahassee, FL 32306-4320

Ms. Rosemary Mattuck—Gradient Corporation, 20 University Road, Cambridge, MA 02138

Professor Russell Flegal— Department of Environmental Toxicology, University of California,
Santa Cruz, 1156 High Street, Santa Cruz, CA 95064
I-xxxvii
-------
Authors, Contributors, and Reviewers
(cont'd)
Contributors and Reviewers
(cont'd)

Ms. Beth Hassett-Sipple—Office of Air Quality Planning and Standards, U.S. Environmental
Protection Agency, Research Triangle Park, NC 27711

Dr. Zachary Pekar—Office of Air Quality Planning and Standards, U.S. Environmental
Protection Agency, Research Triangle Park, NC 27711

Ms. Anne Pope—Office of Air Quality Planning and Standards, U.S. Environmental Protection
Agency, Research Triangle Park, NC 27711

Mr. Douglas Solomon—Office of Air Quality Planning and Standards, U.S. Environmental
Protection Agency, Research Triangle Park, NC 27711

Dr. Joseph Touma—Office of Air Quality Planning and Standards, U.S. Environmental
Protection Agency, Research Triangle Park, NC 27711

Dr. John Vandenberg—National Center for Environmental Assessment (B243-01),
U.S. Environmental Protection Agency, Research Triangle Park, NC 27711
CHAPTER 3 - ROUTES OF HUMAN EXPOSURE AND OBSERVED
ENVIRONMENTAL CONCENTRATIONS
Chapter Managers/Editors

Dr. Brooke L. Hemming—National Center for Environmental Assessment (B243-01),
U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

Dr. Mary Ross—National Center for Environmental Assessment (B243-01),
U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

Principal Authors

Ms. Allison Harris—Department of Civil and Environmental Engineering, Carnegie-Mellon
University, Pittsburgh, PA 15213 (Sections 3.1-3.5)

Professor Cliff Davidson—Department of Civil and Environmental Engineering, Carnegie-
Mellon University, Pittsburgh, PA 15213 (Sections 3.1-3.5)
I-xxxviii
-------
Authors, Contributors, and Reviewers
(cont'd)
Contributors and Reviewers
Mr. Kevin Cavender—Office of Air Quality Planning and Standards, U.S. Environmental
Protection Agency, Research Triangle Park, NC 27711

Professor Brian Gulson—Graduate School of the Environment, Macquarie University
Sydney, NSW 2109, Australia

Professor John W. Winchester (Emeritus)—Department of Oceanography, Florida State
University, Tallahassee, FL 32306-4320

Ms. Rosemary Mattuck—Gradient Corporation, 20 University Road, Cambridge, MA 02138

Professor Russell Flegal— Department of Environmental Toxicology, University of California,
1156 High Street, Santa Cruz, CA 95064

Dr. Sharon Harper—National Exposure Research Laboratory (D205-05), U.S. Environmental
Protection Agency, Research Triangle Park, NC 27711

Mr. Tom Helms—Office of Air Quality Planning and Standards, U.S. Environmental Protection
Agency, Research Triangle Park, NC 27711

Mr. James Hemby—Office of Air Quality Planning and Standards, U.S. Environmental
Protection Agency, Research Triangle Park, NC 27711

Ms. Beth Hassett-Sipple—Office of Air Quality Planning and Standards, U.S. Environmental
Protection Agency, Research Triangle Park, NC 27711

Mr. Phil Lorang—Office of Air Quality Planning and Standards, U.S. Environmental Protection
Agency, Research Triangle Park, NC 27711

Mr. David Mintz—Office of Air Quality Planning and Standards, U.S. Environmental Protection
Agency, Research Triangle Park, NC 27711

Dr. Zachary Pekar—Office of Air Quality Planning and Standards, U.S. Environmental
Protection Agency, Research Triangle Park, NC 27711
I-xxxix
-------
Authors, Contributors, and Reviewers
(cont'd)
Contributors and Reviewers
(cont'd)

Dr. Michael Rizzo—Office of Air Quality Planning and Standards, U.S. Environmental
Protection Agency, Research Triangle Park, NC 27711

Mr. Mark Schmidt—Office of Air Quality Planning and Standards, U.S. Environmental
Protection Agency, Research Triangle Park, NC 27711

Dr. John Vandenberg—National Center for Environmental Assessment (B243-01),
U.S. Environmental Protection Agency, Research Triangle Park, NC 27711
CHAPTER 4 - TOXICOKINETICS, BIOLOGICAL MARKERS, AND MODELS
OF LEAD BURDEN IN HUMANS
Chapter Managers/Editors

Dr. James Brown—National Center for Environmental Assessment (B243-01),
U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

Dr. Robert Elias—National Center for Environmental Assessment (B243-01),
U.S. Environmental Protection Agency, Research Triangle Park, NC 27711 (Retired)

Principal Authors

Dr. Gary Diamond—Syracuse Research Corporation, 8191 Cedar Street, Akron, NY 14001
(Section 4.2-4.4)

Dr. Brian Gulson—Graduate School of the Environment, Macquarie University
Sydney, NSW 2109, Australia (Section 4.3)

Contributors and Reviewers

Dr. Lester D. Grant—National Center for Environmental Assessment (B243-01),
U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

Ms. Rosemary Mattuck—Gradient Corporation, 20 University Road, Cambridge, MA 02138
I-xl
-------
Authors, Contributors, and Reviewers
(cont'd)
Contributors and Reviewers
(cont'd)

Professor Russell Flegal—Department of Environmental Toxicology, University of California,
1156 High Street, Santa Cruz, CA 95064

Dr. Zachary Pekar—Office of Air Quality Planning and Standards, U.S. Environmental
Protection Agency, Research Triangle Park, NC 27711

Ms. Beth Hassett-Sipple—Office of Air Quality Planning and Standards, U.S. Environmental
Protection Agency, Research Triangle Park, NC 27711

Dr. John Vandenberg—National Center for Environmental Assessment (B243-01),
U.S. Environmental Protection Agency, Research Triangle Park, NC 27711
CHAPTER 5 - TOXICOLOGICAL EFFECTS OF LEAD IN LABORATORY ANIMALS,
HUMANS, AND IN VITRO TEST SYSTEMS
Chapter Managers/Editors

Dr. Anuradha Mudipalli—National Center for Environmental Assessment (B243-01),
U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

Dr. Srikanth Nadadur—National Center for Environmental Assessment (B243-01),
U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

Dr. Lori White—National Center for Environmental Assessment (B243-01),
U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

Principal Authors

Dr. Anuradha Mudipalli—National Center for Environmental Assessment (B243-01),
U.S. Environmental Protection Agency, Research Triangle Park, NC 27711 (Sections 5-2, 5-10)

Dr. Stephen Lasley—Dept. of Biomedical and Therapeutic Sciences, Univ. of Illinois College of
Medicine, PO Box 1649, Peoria, IL 61656 (Section 5.3)

Dr. Lori White—National Center for Environmental Assessment (B243-01),
U.S. Environmental Protection Agency, Research Triangle Park, NC 27711 (Section 5.3)
I-xli
-------
Authors, Contributors, and Reviewers
(cont'd)
Principal Authors
(cont'd)

Dr. Gary Diamond— Syracuse Research Corporation, 8191 Cedar Street, Akron, NY 14001
(Section 5.4)

Dr. N.D. Vaziri—Division of Nephrology and Hypertension, University of California - Irvine
Medical Center, 101, The City Drive, Bldg 53, Room #125. Orange, CA 92868 (Section 5.5)

Dr. John Pierce Wise, Sr.—Maine Center for Toxicology and Environmental Health,
Department of Applied Medical Sciences, 96 Falmouth Street, PO Box 9300, Portland, ME
04104-9300 (Section 5.6)

Dr. Harvey C. Gonick—David Geffen School of Medicine, University of California at
Los Angeles, CA (201 Tavistock Ave, Los Angeles, CA 90049) (Sections 5.7, 5.11)

Dr. Gene E. Watson—University of Rochester Medical Center, Box 705, Rochester, NY 14642
(Section 5.8)

Dr. Rodney Dietert—Institute for Comparative and Environmental Toxicology, College of
Veterinary Medicine, Cornell University, Ithaca, NY 14853 (Section 5.9)

Contributors and Reviewers

Dr. Lester D. Grant—National Center for Environmental Assessment (B243-01),
U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

Dr. Paul Reinhart—National Center for Environmental Assessment (B243-01),
U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

Dr. Michael Davis—National Center for Environmental Assessment (B243-01),
U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

Dr. David A. Lawrence—Dept of Environmental and Clinical Immunology, Empire State Plaza
P.O. Box 509, Albany, NY 12201

Dr. Michael J. McCabe, Jr.—Dept of Environmental Medicine, University of Rochester,
575 Elmwood Avenue, Rochester, NY 14642
I-xlii
-------
Authors, Contributors, and Reviewers
(cont'd)
Contributors and Reviewers
(cont'd)
Dr. Theodore I. Lidsky—New York State Institute for Basic Research, 1050 Forest RD,
Staten Island, NY 10314

Dr. Mark H. Follansbee—Syracuse Research Corporation, 8191 Cedar St. Akron, NY 14001

Dr. William K. Boyes—National Health and Environmental Effects Research Laboratory,
U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

Dr. Philip J. Bushnell—National Health and Environmental Effects Research Laboratory,
U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

Ms. Beth Hassett-Sipple—Office of Air Quality Planning and Standards, U.S. Environmental
Protection Agency, Research Triangle Park, NC 27711

Dr. Zachary Pekar—Office of Air Quality Planning and Standards, U.S. Environmental
Protection Agency, Research Triangle Park, NC 27711

Dr. John Vandenberg—National Center for Environmental Assessment (B243-01),
U.S. Environmental Protection Agency, Research Triangle Park, NC 27711
CHAPTER 6 - EPIDEMIOLOGICAL STUDIES OF HUMAN HEALTH EFFECTS
ASSOCIATED WITH LEAD EXPOSURE
Chapter Managers/Editors

Dr. Jee Young Kim—National Center for Environmental Assessment (B243-01),
U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

Dr. Dennis Kotchmar—National Center for Environmental Assessment (B243-01),
U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

Dr. David Svendsgaard—National Center for Environmental Assessment (B243-01),
U.S. Environmental Protection Agency, Research Triangle Park, NC 27711
I-xliii
-------
Authors, Contributors, and Reviewers
(cont'd)
Principal Authors
Dr. David Bellinger—Children's Hospital, Farley Basement, Box 127, 300 Longwood Avenue,
Boston, MA 02115 (Section 6.10)

Dr. Margit Bleecker—Center for Occupational and Environmental Neurology, 2 Hamill Road,
Suite 225, Baltimore, MD 21210 (Section 6.3)

Dr. Gary Diamond— Syracuse Research Corporation, 8191 Cedar Street.
Akron, NY 14001 (Section 6.8, 6.9)

Dr. Kim Dietrich—University of Cincinnati College of Medicine, 3223 Eden Avenue,
Kettering Laboratory, Room G31, Cincinnati, OH 45267 (Section 6.2)

Dr. Pam Factor-Litvak—Columbia University Mailman School of Public Health, 722 West
168th Street, Room 1614, New York, NY 10032 (Section 6.6)

Dr. Vic Hasselblad—Duke University Medical Center, Durham, NC 27713 (Section 6.10)

Dr. Stephen J. Rothenberg—CINVESTAV-IPN, Merida, Yucatan, Mexico & National Institute
of Public Health, Cuernavaca, Morelos, Mexico (Section 6.5)

Dr. Neal Simonsen—Louisiana State University Health Sciences Center, School of Public Health
& Stanley S Scott Cancer Center, 1600 Canal Street, Suite 800, New Orleans, LA 70112
(Section 6.7)

Dr. Kyle Steenland—Rollins School of Public Health, Emory University, 1518 Clifton Road,
Room 268, Atlanta, GA 30322 (Section 6.7)

Dr. Virginia Weaver—Johns Hopkins Bloomberg School of Public Health, 615 North Wolfe
Street, Room 7041, Baltimore, MD 21205 (Section 6.4)

Contributors and Reviewers

Dr. J. Michael Davis—National Center for Environmental Assessment (B243-01),
U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

Dr. Lester D. Grant—National Center for Environmental Assessment (B243-01),
U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

Dr. Kazuhiko Ito—Nelson Institute of Environmental Medicine, New York University School of
Medicine, Tuxedo, NY 10987
I-xliv
-------
Authors, Contributors, and Reviewers
(cont'd)
Contributors and Reviewers
(cont'd)

Dr. Kathryn Mahaffey—Office of Prevention, Pesticides and Toxic Substances,
U.S. Environmental Protection Agency, Washington, DC 20460

Dr. Deirdra Murphy—Office of Air Quality Planning and Standards, U.S. Environmental
Protection Agency, Research Triangle Park, NC 27711

Dr. Zachary Pekar—Office of Air Quality Planning and Standards, U.S. Environmental
Protection Agency, Research Triangle Park, NC 27711

Ms. Beth Hassett-Sipple—Office of Air Quality Planning and Standards, U.S. Environmental
Protection Agency, Research Triangle Park, NC 27711

Dr. John Vandenberg—National Center for Environmental Assessment (B243-01),
U.S. Environmental Protection Agency, Research Triangle Park, NC 27711
CHAPTER 7- ENVIRONMENTAL EFFECTS OF LEAD
Chapter Manager/Editor

Dr. Timothy Lewis—National Center for Environmental Assessment (B243-01),
U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

Principle Authors

Dr. Ruth Hull—Cantox Environmental Inc., 1900 Minnesota Court, Suite 130, Mississauga,
Ontario, L5N 3C9 Canada (Section 8.1)

Dr. James Kaste—Department of Earth Sciences, Dartmouth College, 352 Main Street, Hanover,
NH03755 (Section 8.1)

Dr. John Drexler—Department of Geological Sciences, University of Colorado, 1216 Gillespie
Drive, Boulder, CO 80305 (Section 8.1)

Dr. Chris Johnson—Department of Civil and Environmental Engineering, Syracuse University,
365 Link Hall, Syracuse, NY 13244 (Section 8.1)
I-xlv
-------
Authors, Contributors, and Reviewers
(cont'd)
Principle Authors
(cont'd)

Dr. William Stubblefield—Parametrix, Inc. 33972 Texas St. SW, Albany, OR 97321
(Section 8.2)

Dr. Dwayne Moore—Cantox Environmental, Inc., 1550ALaperriere Avenue, Suite 103,
Ottawa, Ontario, K1Z 7T2 Canada (Section 8.2)

Dr. David Mayfield—Parametrix, Inc., 411 108th Ave NE, Suite 1800, Bellevue, WA 98004
(Section 8.2)

Dr. Barbara Southworth—Menzie-Cura & Associates, Inc., 8 Winchester Place, Suite 202,
Winchester, MA 01890 (Section 8.3)

Dr. Katherine Von Stackleberg—Menzie-Cura & Associates, Inc., 8 Winchester Place, Suite
202, Winchester, MA 01890 (Section 8.3)

Contributors and Reviewers

Dr. Jerome Nriagu—Department of Environmental Health Sciences, 109 South Observatory,
University of Michigan, Ann Arbor, MI 48109

Dr. Judith Weis—Department of Biology, Rutgers University, Newark, NJ 07102

Dr. Sharon Harper—National Exposure Research Laboratory (D205-05), U.S. Environmental
Protection Agency, Research Triangle Park, NC 27711

Dr. Karen Bradham—National Research Exposure Laboratory (D205-05), U.S. Environmental
Protection Agency, Research Triangle Park, NC 27711

Dr. Ginger Tennant—Office of Air Quality Planning and Standards, U.S. Environmental
Protection Agency, Research Triangle Park, NC 27711

Ms. Gail Lacey—Office of Air Quality Planning and Standards, U.S. Environmental Protection
Agency, Research Triangle Park, NC 27711

Dr. John Vandenberg—National Center for Environmental Assessment (B243-01),
U.S. Environmental Protection Agency, Research Triangle Park, NC 27711
I-xlvi
-------
Authors, Contributors, and Reviewers
(cont'd)
CHAPTER 8 - INTEGRATIVE SYNTHESIS: INTEGRATIVE SYNTHESIS:
LEAD EXPOSURE AND HEALTH EFFECTS
Chapter Manager/Editor

Dr. Lester D. Grant—National Center for Environmental Assessment (B243-01),
U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

Principal Authors

Dr. Lester D. Grant—National Center for Environmental Assessment (B243-01),
U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

Dr. Lori White—National Center for Environmental Assessment (B243-01),
U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

Dr. Jee Young Kim—National Center for Environmental Assessment (B243-01),
U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

Dr. Dennis J. Kotchmar—National Center for Environmental Assessment (B243-01),
U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

Dr. Srikanth S. Nadadur—National Center for Environmental Assessment (B243-01),
U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

Dr. Mary Ross— National Center for Environmental Assessment (B243-01),
U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

Dr. David Bellinger—Children's Hospital, Farley Basement,
Box 127, 300 Longwood Avenue, Boston, MA 02115

Dr. Vic Hasselblad—Duke University Medical Center, Durham NC 27713

Dr. John Rosen—Division of Environmental Sciences, The Children's Hospital at Montefiore,
The Albert Einstein College of Medicine, 111 E. 210th St., Room 401, Bronx, NY 10467

Contributors and Reviewers

Dr. James Brown—National Center for Environmental Assessment (B243-01),
U.S. Environmental Protection Agency, Research Triangle Park, NC 27711
I-xlvii
-------
Authors, Contributors, and Reviewers
(cont'd)
Contributors and Reviewers
(cont'd)

Dr. Brooke L. Hemming—National Center for Environmental Assessment (B243-01),
U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

Dr. Anuradha Mudipalli—National Center for Environmental Assessment (B243-01),
U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

Dr. David Svendsgaard—National Center for Environmental Assessment (B243-01),
U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

Dr. Deirdra Murphy—Office of Air Quality Planning and Standards, U.S. Environmental
Protection Agency, Research Triangle Park, NC 27711

Dr. John Vandenberg—National Center for Environmental Assessment (B243-01),
U.S. Environmental Protection Agency, Research Triangle Park, NC 27711
EXECUTIVE SUMMARY
Chapter Manager/Editor

Dr. Lester D. Grant—National Center for Environmental Assessment (B243-01),
U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

Principal Authors

Dr. Lester D. Grant—National Center for Environmental Assessment (B243-01),
U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

Dr. Lori White—National Center for Environmental Assessment (B243-01),
U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

Dr. Timothy Lewis—National Center for Environmental Assessment (B243-01),
U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

Contributors and Reviewers

Dr. James Brown—National Center for Environmental Assessment (B243-01),
U.S. Environmental Protection Agency, Research Triangle Park, NC 27711
I-xlviii
-------
Authors, Contributors, and Reviewers
(cont'd)
Contributors and Reviewers
(cont'd)

Dr. Jee Young Kim—National Center for Environmental Assessment (B243-01),
U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

Dr. Dennis J. Kotchmar—National Center for Environmental Assessment (B243-01),
U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

Dr. Anuradha Mudipalli—National Center for Environmental Assessment (B243-01),
U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

Dr. Deirdra Murphy—Office of Air Quality Planning and Standards, U.S. Environmental
Protection Agency, Research Triangle Park, NC 27711

Dr. Srikanth S. Nadadur—National Center for Environmental Assessment (B243-01),
U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

Dr. Mary Ross— National Center for Environmental Assessment (B243-01),
U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

Dr. John Vandenberg—National Center for Environmental Assessment (B243-01),
U.S. Environmental Protection Agency, Research Triangle Park, NC 27711
I-xlix
-------
U.S. Environmental Protection Agency Project Team
for Development of Air Quality Criteria for Lead
Executive Direction

Dr. Lester D. Grant (Director)—National Center for Environmental Assessment-RTF Division,
(B243-01), U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

Scientific Staff

Dr. Lori White (Lead Team Leader)—National Center for Environmental Assessment
(B243-01), U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

Dr. James S. Brown—National Center for Environmental Assessment (B243-01),
U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

Dr. Robert Elias—National Center for Environmental Assessment (B243-01),
U.S. Environmental Protection Agency, Research Triangle Park, NC 27711 (Retired)

Dr. Brooke Hemming—National Center for Environmental Assessment (B243-01), U.S.
Environmental Protection Agency, Research Triangle Park, NC 27711

Dr. Jee Young Kim—National Center for Environmental Assessment (B243-01),
U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

Dr. Dennis Kotchmar—National Center for Environmental Assessment (B243-01),
U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

Dr. Timothy Lewis—National Center for Environmental Assessment (B243-01),
U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

Dr. Anuradha Muldipalli—National Center for Environmental Assessment (B243-01),
U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

Dr. Srikanth Nadadur—National Center for Environmental Assessment (B243-01),
U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

Dr. Paul Reinhart—National Center for Environmental Assessment (B243-01),
U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

Dr. Mary Ross—National Center for Environmental Assessment (B243-01),
U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

Dr. David Svendsgaard—National Center for Environmental Assessment (B243-01),
U.S. Environmental Protection Agency, Research Triangle Park, NC 27711
1-1
-------
U.S. Environmental Protection Agency Project Team
for Development of Air Quality Criteria for Lead
(cont'd)
Technical Support Staff

Mr. Douglas B. Fennell—Technical Information Specialist, National Center for Environmental
Assessment (B243-01), U.S. Environmental Protection Agency, Research Triangle Park, NC
27711 (Retired)

Ms. Emily R. Lee—Management Analyst, National Center for Environmental Assessment
(B243-01), U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

Ms. Diane H. Ray—Program Specialist, National Center for Environmental Assessment
(B243-01), U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

Ms. Donna Wicker—Administrative Officer, National Center for Environmental Assessment
(B243-01), U.S. Environmental Protection Agency, Research Triangle Park, NC 27711 (Retired)

Mr. Richard Wilson—Clerk, National Center for Environmental Assessment (B243-01),
U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

Document Production Staff

Ms. Carolyn T. Perry—Task Order Manager, Computer Sciences Corporation, 2803 Slater Road,
Suite 220, Morrisville, NC 27560

Mr. John A. Bennett—Technical Information Specialist, Library Associates of Maryland,
11820 Parklawn Drive, Suite 400, Rockville, MD 20852

Ms. Samantha Dycus—Publication/Graphics Specialist, TekSystems, 1201 Edwards Mill Road,
Suite 201, Raleigh, NC 27607

Mr. William Ellis—Records Management Technician, InfoPro, Inc., 8200 Greensboro Drive,
Suite 1450, McLean, VA 22102

Ms. Sandra L. Hughey—Technical Information Specialist, Library Associates of Maryland,
11820 Parklawn Drive, Suite 400, Rockville, MD 20852

Dr. Barbara Liljequist—Technical Editor, Computer Sciences Corporation, 2803 Slater Road,
Suite 220, Morrisville, NC 27560

Ms. Michelle Partridge-Doerr—Publication/Graphics Specialist, TekSystems, 1201 Edwards
Mill Road, Suite 201, Raleigh, NC 27607

Mr. Carlton Witherspoon—Graphic Artist, Computer Sciences Corporation, 2803 Slater Road,
Suite 220, Morrisville, NC 27560
I-li
-------
U.S. Environmental Protection Agency
Science Advisory Board (SAB) Staff Office
Clean Air Scientific Advisory Committee (CASAC)
Chair
Dr. Rogene Henderson*—Scientist Emeritus, Lovelace Respiratory Research Institute,
Albuquerque, NM

Members

Dr. Joshua Cohen—Faculty, Center for the Evaluation of Value and Risk, Institute for Clinical
Research and Health Policy Studies, Tufts New England Medical Center, Boston, MA

Dr. Deborah Cory-Slechta—Director, University of Medicine and Dentistry of New Jersey and
Rutgers State University, Piscataway, NJ

Dr. Ellis Cowling*—University Distinguished Professor-at-Large, North Carolina State
University, Colleges of Natural Resources and Agriculture and Life Sciences, North Carolina
State University, Raleigh, NC

Dr. James D. Crapo [M.D.]*—Professor, Department of Medicine, National Jewish Medical and
Research Center, Denver, CO

Dr. Bruce Fowler—Assistant Director for Science, Division of Toxicology and Environmental
Medicine, Office of the Director, Agency for Toxic Substances and Disease Registry, U.S.
Centers for Disease Control and Prevention (ATSDR/CDC), Chamblee, GA

Dr. Andrew Friedland—Professor and Chair, Environmental Studies Program, Dartmouth
College, Hanover, NH

Dr. Robert Goyer [M.D.]—Emeritus Professor of Pathology, Faculty of Medicine, University of
Western Ontario (Canada), Chapel Hill, NC

Mr. Sean Hays—President, Summit Toxicology, Allenspark, CO

Dr. Bruce Lanphear [M.D.]—Sloan Professor of Children's Environmental Health, and the
Director of the Cincinnati Children's Environmental Health Center at Cincinnati Children's
Hospital Medical Center and the University of Cincinnati, Cincinnati, OH

Dr. Samuel Luoma—Senior Research Hydrologist, U.S. Geological Survey (USGS),
Menlo Park, CA
I-lii
-------
U.S. Environmental Protection Agency
Science Advisory Board (SAB) Staff Office
Clean Air Scientific Advisory Committee (CASAC)
(cont'd)

Members
(cont'd)

Dr. Frederick J. Miller*—Consultant, Gary, NC

Dr. Paul Mushak—Principal, PB Associates, and Visiting Professor, Albert Einstein College of
Medicine (New York, NY), Durham, NC

Dr. Michael Newman—Professor of Marine Science, School of Marine Sciences, Virginia
Institute of Marine Science, College of William & Mary, Gloucester Point, VA

Mr. Richard L. Poirot*—Environmental Analyst, Air Pollution Control Division, Department of
Environmental Conservation, Vermont Agency of Natural Resources, Waterbury, VT

Dr. Michael Rabinowitz—Geochemist, Marine Biological Laboratory, Woods Hole, MA

Dr. Joel Schwartz—Professor, Environmental Health, Harvard University School of Public
Health, Boston, MA

Dr. Frank Speizer [M.D.]*—Edward Kass Professor of Medicine, Channing Laboratory, Harvard
Medical School, Boston, MA

Dr. Ian von Lindern—Senior Scientist, TerraGraphics Environmental Engineering, Inc.,
Moscow, ID

Dr. Barbara Zielinska*—Research Professor, Division of Atmospheric Science, Desert Research
Institute, Reno, NV
SCIENCE ADVISORY BOARD STAFF

Mr. Fred Butterfield—CASAC Designated Federal Officer, 1200 Pennsylvania Avenue, N.W.,
Washington, DC, 20460, Phone: 202-343-9994, Fax: 202-233-0643 (butterfield.fred@epa.gov)
(Physical/Courier/FedEx Address: Fred A. Butterfield, III, EPA Science Advisory Board Staff
Office (Mail Code 1400F), Woodies Building, 1025 F Street, N.W., Room 3604, Washington,
DC 20004, Telephone: 202-343-9994)
*Members of the statutory Clean Air Scientific Advisory Committee (CASAC) appointed by the
U.S. EPA Administrator
I-liii
-------
Abbreviations and Acronyms
aFGF
AA
AAL
AALM
AAS
ACBP
ACE
AChE
ACSL
ADCC
ADHD
ADP
AF
A horizon
AHR
ALA
ALAD
ALAS
ALAU
ALM
ALS
ALT
AMD
AMP
ANF
ANOVA
AP
AP-1
APE
ApoE
APP
AQCD
a-fibroblast growth factor
arachidonic acid; atomic absorption
active avoidance learning
All Ages Lead Model
atomic absorption spectroscopy
Achenbach Child Behavior Profile
angiotensin converting enzyme
acetylcholinesterase
Advanced Continuous Simulation Language
antibody-dependent cellular cytotoxicity
attention deficit/hyperactivity disorder
adenosine diphosphate
absorption fraction
uppermost layer of soil (litter and humus)
aryl hydrocarbon receptor
5-aminolevulinic acid; 5-aminolevulinic acid
5-aminolevulinic acid dehydratase
aminolevulinic acid synthase
5-aminolevulinic acid dehydratase
Adult Lead Methodology
amyotrophic lateral sclerosis
alanine aminotransferase; alanine transferase
activity mean diameter
adenosine monophosphate
atrial natriuretic factor
analysis of variance
alkaline phosphatase
activator protein-1
apurinic endonuclease
apolipoprotein E
amyloid precursor protein
Air Quality Criteria Document
I-liv
-------
ASA
AST
ASV
ATP
ATP1A2
ATPase
ATSDR
ATV
AVS
AWQC
P
PFGF
6-p-OH-cortisol
BAEP
BBB
Bcell
BCF
BDNF
BLL
BLM
BMDM
BMI
BMP
BRHS
BSID
BTQ
BUN
BW
CA
45Ca, 47Ca
CA1
CAS
CAA
Ca-ATPase
arylsulfatase
aspartate aminotransferase
anode stripping voltammetry
adenosine triphosphate
sodium-potassium adenosine triphosphase a2
adenosine triphosphate synthase
Agency for Toxic Substances and Disease Research
all-terrain vehicle
acid volatile sulfide
ambient water quality criteria
beta-coefficient; slope of an equation
P-fibroblast growth factor
6-p-hydroxycortisol
brainstem auditory-evoked potentials
blood-brain barrier
B lymphocyte
bioconcentration factor
brain-derived neurotrophic factor
blood lead level
biotic ligand model
bone marrow-derived macrophages
body mass index
bone morphogenic protein
British Regional Heart Study
Bayley Scales of Infant Development
Boston Teacher Questionnaire
blood urea nitrogen
body weight
chromosomal aberration
calcium-45 and -47 radionuclides
cornu ammonis 1 region of hippocampus
cornu ammonis 3 region of hippocampus
Clean Air Act
calcium-dependent adenosine triphosphatase
I-lv
-------
43CaCl2
CaCO3
CaEDTA
CAL
CAMKII
cAMP
CaNa2 EDTA
CANTAB
CAP
Caio(P04)6(OH)2
CASAC
CBCL
CCE
Cd
CD
CDC
CERR
CESD, CES-D
cGMP

CI
CKD
CLRTAP
CMI
CNS
C02
ConA
COX-2
CP
CPT
CRAC
CREB
CRT
CSF
calcium-43 radionuclide-labeled calcium chloride
calcium carbonate
calcium disodium ethylenediaminetetraacetic acid
calcitonin
calcium/calmodulin-dependent protein kinase
cyclic adenosinemonophosphate
calcium disodium ethylenediaminetetraacetic acid
Cambridge Neuropsychological Testing Automated Battery
criteria air pollutant
hydroxyapatite
Clean Air Scientific Advisory Committee
Achenbach Child Behavior Checklist
Coordination Center for Effects
cadmium
Sprague-Dawley CD (rat)
Centers for Disease Control and Prevention
Consolidated Emissions Reporting Rule
Center for Epidemiologic Studies Depression (scale)
cyclic guanosine-3',5'-monophosphate; cyclic
guanylylmonophosphate
confidence interval
chronic kidney disease
Convention on Long-range Transboundary of Air Pollution
cell-mediated immunity
central nervous system
carbon dioxide
concanavalin A
cyclooxygenase-2
coproporphyrin
current perception threshold
calcium release activated calcium reflux
cyclic-AMP response element binding protein
chronic renal insufficiency
cerebrospinal fluid
I-lvi
-------
CSF-1
CTL
CuZnSOD
CWA
CYP
DA
DET
DPS
dfs
DOT
DIAL
DMEM
DMFS
DMSA
DMTU
DNA
DNTC
DOC
DOM
DOS
DPH
DRL
DSA
DTC
DTK
E
E2
EBE
EC
ECF
Eco-SSL
EDRF
EDTA
colony-stimulating factor-1
cytotoxic T lymphocyte
copper and zinc-dependent superoxide dismutase
Clean Water Act
cytochrome (e.g., CYP1 A, CYP-2A6, CYP3A4, CYP450)
dopamine; dopaminergic
diffusive equilibrium thin films
decayed or filled surfaces, permanent teeth
covariate-adjusted number of caries
diffusive gradient thin films
dialkyllead
Dulbecco's Modified Eagle Medium
decayed, missing, or filled surfaces, permanent teeth
2,3-dimercaptosuccinic acid; dimethyl succinic acid
dimethyl thio urea
deoxyribonucleic acid
diffuse neurofibrillary tangles with calcification
dissolved organic carbon
dissolved organic matter
Disk Operating System
1, 6-diphenyl-l,3,5-hexatriene
differential reinforcement of low rate (schedule)
delayed spatial alternation
dithiocarbamate
delayed type hypersensitivity
embryonic day; epinephrine
estradiol
early biological effect
coronary endothelial (cells)
effect concentration for 50% of test population
extracellular fluid
ecological soil screening level
endothelium-derived relaxing factor
ethylenediaminetetraacetic acid
I-lvii
-------
EEDQ
EEG
EOF
EGTA
eNOS
EOD
EP
EPA
EPMA
EPSP
EqP
ERG
ERL
ERM
EROD
ESP
ESRD
ET
ET-AAS
EXAFS
EXANES
F344
FA
FCS
FDA
FEF
FEVi
FGF
FI
FMLP
fMRI
foe
FPLC
FR
7V-ethoxycarbonyl-2-ethoxy-1,2-dihydroquinone
el ectroencephal ogram
epidermal growth factor
ethyleneglycoltetraacetic acid
endothelial nitric oxide synthase
explosive ordnance disposal
erythrocyte protoporphyrin
U.S. Environmental Protection Agency
electron probe microanalysis
excitatory postsynaptic potential
equilibrium partitioning (theory)
electroretinogram
effects range - low
effects range - median
ethoxyresorufin-0-deethylase
electrostatic precipitator
end-stage renal disease
endothelein; essential tremor
electrothermal atomic absorption spectroscopy
extended X-ray absorption fine structure
extended X-ray absorption near edge spectroscopy
Fischer 344 (rat)
fatty acid
fetal calf serum
Food and Drug Administration
forced expiratory flow
forced expiratory volume in one second
fibroblast growth factor (e.g., PFGF, aFGF)
fixed-interval (operant conditioning)
7V-formy 1 -L-methi ony 1 -L-l eucy 1 -L-pheny 1 al anine
functional magnetic resonance imaging
fraction organic carbon
fast protein liquid chromatography
Federal Register; fixed-ratio operant conditioning
I-lviii
-------
FSH
FT3
FT4
FVC
Y-GT
GABA
GAG
GCI
GD
GDP
GEE
GFAAS
GFAP
GFR
GH
GI
GL
GLU
GM
GMP
GnRH
goc
G6PD
GPEI
G-R
GRP78
GSD
GSD;
GSH
GSHPx
GSSG
GST
GTP
GvH
follicle stimulating hormone
free triiodothyronine
free thyroxine
forced vital capacity
y-glutamyl transferase
gamma aminobutyric acid
glycosaminoglycan
General Cognitive Index
gestational day
guanosine diphosphate
generalized estimating equations
graphite furnace atomic absorption spectroscopy
glial fibrillary acidic protein
glomerular filtration rate
growth hormone
gastrointestinal
gestation and lactation
glutamate
geometric mean
guanosine monophosphate
gonadotropin releasing hormone
grams organic carbon
glucose-6-phosphate dehydrogenase
glutathione S-transferase P enhancer element
Graham-Rosenblith Behavioral Examination for Newborns
glucose-regulated protein 78
geometric standard deviation
individual geometric standard deviation
glutathione; reduced glutathione
glutathione peroxidase
oxidized glutathione
glutathione transferase; glutathione S-transferase
guanosine triphosphate
graft versus host (reaction)
I-lix
-------
H+
HAP
Hb
HBEF
H2C03
Hct
HDL
HFE
HFF
HH
HHANES
HHC
HLA
HNO3
H202
HOC1
HOME
HOS-TE-85
HPG
HPLC
H3P04
HPRT
HSAB
H2SO4
HSPG
HTN
HUD
HY-SPLIT
IARC
IBL
ICD
ICP
ICP-AES
ICP-MS
acidity
hazardous air pollutant
hemoglobin
Hubbard Brook Experimental Forest
carbonic acid
hematocrit
high-density lipoprotein (cholesterol)
hemochromatosis gene
human foreskin fibroblasts
hydroxylamine hydrochloride
Hispanic Health and Nutrition Examination Survey
hereditary hemochromatosis
human leukocyte antigen
nitric acid
hydrogen peroxide
hypochlorous acid
Home Observation for Measurement of Environment
human osteosarcoma cells
hypothalamic-pituitary-gonadal (axi s)
high-pressure liquid chromatography
phosphoric acid
hypoxanthine guanine phosphoribosyl transferase
Hard-Soft Acid-Base (model)
sulfuric acid
heparan sulfate proteoglycan
hypertension
U.S. Department of Housing and Urban Development
hybrid single-particle Lagrangian integrated trajectory (model)
International Agency for Research on Cancer
integrated blood lead index
International Classification of Diseases
inductively coupled plasma
inductively coupled plasma atomic emission spectroscopy
inductively coupled plasma mass spectrometry
I-lx
-------
ICRP
IDMS
IEC
IEUBK
IFN
Ig
IGFi
IL
IMPROVE

iNOS
i.p., IP
IQ
IRT
ISCST
IT
i.v., IV
KABC
KID
KLH
K-pNPPase
KTEA
K-XRF
L
LAA ICP-MS
LCso
LDH
LDL
L-dopa
LE
LH
LISREL
LMW
LNAME, L-NAME
International Commission on Radiological Protection
isotope dilution mass spectrometry
intestinal epithelial cells
Integrated Exposure Uptake Biokinetic (model)
interferon (e.g., IFN-y)
immunoglobulin (e.g., IgA, IgE, IgG, IgM)
insulin-like growth factor 1
interleukin (e.g., IL-1, IL-lp, IL-4, IL-6, IL-12)
Interagency Monitoring of Protected Visual Environments
(network)
inducible nitric oxide synthase
intraperitoneal
intelligence quotient
interresponse time
Industrial Source Complex Short Term (model)
intrathecal
intravenous
Kaufman Assessment Battery for Children
Kent Infant Development Scale
keyhole limpet hemocyanin
potassium-stimulated p-nitrophenylphosphatase
Kaufman Test of Educational Achievement
K-shell X-ray fluorescence
lactation
laser ablation inductively coupled plasma mass spectrometry
lethal concentration (at which 50% of exposed animals die)
lactate dehydrogenase
low-density lipoprotein (cholesterol)
3,4-dihydroxyphenylalanine (precursor of dopamine)
Long Evans (rat)
luteinizing hormone
linear structural relationships (model)
low molecular weight
L-7V°-nitroarginine methyl ester
I-lxi
-------
LOAEL
LOWES S
LPO
LPS
LT50
LTD
LTP
LVH
liPIXE
MAO
MAPK
MCH
MCHC
MCV
MDA
MDA-TBA
MDI
MDRD
meso-DMSA
Mg-ATPase
MHC
MK-801
MLR
MMAD
MMSE
MN
Mn-SOD
MRFIT
MRI
mRNA
MRS
MSV
MT
MVV
lowest-observed adverse effect level
locally weighted scatter plot smoother
lipid peroxide; lipid peroxidation
lipopolysaccharide
time to reach 50% mortality
long-term depression
long-term potentiation
left ventricular hypertrophy
microfocused particle induced X-ray emission
monoaminoxidase
mitogen-activated protein kinase
mean corpuscular hemoglobin
mean corpuscular hemoglobin concentration
mean corpuscular volume
malondialdehyde
malondialdehyde-thiobarbituric acid
Mental Development Index
Modification of Diet in Renal Disease (study)
m-2,3-dimercaptosuccinic acid
magnesium-dependent adenosine triphosphatase
major histocompatibility complex
NMD A receptor antagonist
mixed lymphocyte response
mass median aerodynamic diameters
Mini-Mental State Examination
micronuclei formation
manganese-dependent superoxide dismutase
Multiple Risk Factor Intervention Trial
magnetic resonance imaging
messenger ribonucleic acid
magnetic resonance spectroscopy
Moloney sarcoma virus
metallothionein
maximum voluntary ventilation
I-lxii
-------
N,n
NA, N/A
NAA
NAAQS
NAC
NAD
NADH

NADP
NAD(P)H, NADPH
NADS
NAG
Na-K-ATPase
NART
NAS
NASCAR
NAT
NAWQA
NBAS
NCEA-RTP

ND
NE
NEI
NEPSY
NES
NF-KB
NHANES
NHEXAS
NIOSH
NIST
NK
NMDA
NMDAR
number of observations
not available
TV-acetylaspartate; neutron activation analysis
National Ambient Air Quality Standards
TV-acetyl cysteine; nucleus accumbens
nicotinamide adenine nucleotide
reduced nicotinamide adenine dinucleotide; nicotinamide adenine
dinucleotide dehydrogenase
nicotinamide adenine dinucleotide phosphate
reduced nicotinamide adenine dinucleotide phosphate
nicotinamide adenine dinucleotide synthase
7V-acetyl-p-D-glucosaminidase
sodium-potassium-dependent adenosine triphosphatase
North American Reading Test
Normative Aging Study
National Association for Stock Car Automobile Racing
TV-acetyltransferases
National Water-Quality Assessment
Brazelton Neonatal Behavioral Assessment Scale
National Center for Experimental Assessment Division in
Research Triangle Park, NC
non-detectable; not detected; not determined; not done
norepinephrine
National Emissions Inventory
Developmental Neuropsychological Assessment
Neurobehavioral Evaluation System
nuclear transcription factor-KB
National Health and Nutrition Examination Survey
National Human Exposure Assessment Survey
National Institute for Occupational Safety and Health
National Institute for Standards and Technology
natural killer
7V-methyl-D-aspartate
7V-methyl-D-aspartate receptor
I-lxiii
-------
NO
NO2
N03
NOD
NOEC
NOM
NOS
NOX
NPL
NR
NRC
NTP
NTR
02
OAQPS
OAR
OC
OH
1,25-OH-D, 1,25-OHD3
25-OH-D, 25-OH D3
O horizon
ONOO
OR
ORD
OS
OSHA
P
Pio

PAD
PAH
PAI-1
Pb
203
Pb
nitric oxide
nitrogen dioxide
nitrate
autoimmune diabetes prone strain of mice
no-observed-effect concentration
natural organic matter
nitric oxide synthase; not otherwise specified
nitrogen oxide metabolites
National Priorities List
not reported
National Research Council
National Toxicology Program
neurotrophin receptor
superoxide ion
Office of Air Quality Planning and Standards
Office of Air and Radiation
organic carbon
hydroxyl
1,25-dihydroxyvitamin D
25-hydroxyvitamin D
forest floor
peroxynitrate ion
odds ratio
Office of Research and Development
oxidative stress
Occupational Safety and Health Administration
probability value
probability for the occurrence of a blood lead concentration
exceeding 10 |ig/dL
peripheral arterial disease
polycyclic aromatic hydrocarbon
plasminogen activator inhibitor-1
lead
lead-203 radionuclide
I-lxiv
-------
20
4pb, 206pb, 207pb, 208pb
PbB
PbBPs
PbCl2
PbCO3
PbD
PBG-S
PbH
Pb(N03)2
PbO
Pb(OH)2
PbS
PbSO4
PC12
PCV
PDE
PDI
PEC
PFCs
PG
PHA
Pi
PIXE
PKA
PKC
PM
PM2.5
PMN
PMNL
P5N
PND
POMS
stable isotopes of lead-204, -206, -207, -208 respectively
lead-210 radionuclide
blood lead; blood lead concentration
lead binding proteins
lead choride
lead carbonate
interior dust lead concentration
porphobilinogen synthase
hand lead concentration
lead nitrate
lead oxide
lead hydroxide
galena
lead sulfate
pheochromocytoma cell
packed cell volume
phosphodiesterase
Psychomotor Index
probably effect concentration
plaque forming cells
prostaglandin (e.g., PGE2 ,PGF2)
phytohemagglutinin A
inorganic phosphorus
particle induced X-ray emission
protein kinase A
protein kinase C
particulate matter
combination of coarse and fine particulate matter
fine particulate matter
polymorphonuclear leukocyte
polymorphonuclear leukocyte
pyrimidine 5'-nucleotidase
postnatal day
Profile of Mood States
I-lxv
-------
ppb
PPD
ppm
PPVT-R
PRL
PTH
PTHrP
PUFA
PVC
PWM
R2
r
r2
RAAS
RAS
RBA
RBC
RBP
222Rn
ROI
ROS
ROS 17.2.8
RR
ZSEM

SAB
S-BIQ
SBIS-4
s.c., SC
SCAN
SCE
SD
SDS-PAGE
SE
parts per billion
purified protein derivative
parts per million
Peabody Picture Vocabulary Test-Revised
prolactin
parathyroid hormone
parathyroid hormone-related protein
polyunsaturated fatty acid
polyvinyl chloride
pokeweed mitogen
multiple correlation coefficient
Pearson correlation coefficient
correlation coefficient
renin-angiotensin-aldosterone system
renin-angiotensin system
relative bioavailablity
red blood cell; erythrocyte
retinol binding protein
most stable isotope of radon
reactive oxygen intermediate
reactive oxygen species
rat osteosarcoma cell line
relative risk
sum of the molar concentrations of simultaneously
extracted metal
Science Advisory Board
Standford-Binet Intelligence Quotient
Stanford-Binet Intelligence Scale-4th Edition
subcutaneous
test of central auditory processing
sister chromatid exchange
standard deviation; Spraque-Dawley (rat)
sodium dodecyl sulfate-polyacrylamide gel electrophoresis
standard error; Staphylococcus aureus enterotoxin
I-lxvi
-------
SEM
SES
sGC
SHBG
SIMS
SIR
SLE
SMR
SNAP
SO2
SOD
SOILCHEM
SRA
SRBC
SRC
SRD
SRE
SRIXE
SULT
T3
T4
T&E
TB
TEA
TEARS
Tc
Tcell
TEC
TEL
TES
TF
TGF
TH
232
Th
simultaneously extracted metal; standard error of the mean
socioeconomic status
soluble guanylate cyclase
sex hormone binding globulin
secondary ion mass spectrometry
standardized incidence ratio
systemic lupus erythmatosus
standardized mortality ratio
Schneider Neonatal Assessment for Primates
sulfur dioxide
superoxide dismutase
chemical species equilibrium model
Self Reported Antisocial Behavior scale
sheep red blood cell
Syracuse Research Corporation
Self Report of Delinquent Behavior
sterol regulatory element
synchrotron radiation induced X-ray emission
sulfotransferases
triiodothyronine
thyroxine
threatened and endangered (species)
tuberculosis
thiobarbituric acid
thiobarbituric acid-reactive species
cytotoxic T lymphocyte
T lymphocyte
threshold effect concentration
tetraethyllead; triethyl lead chloride
testosterone
transferrin
transforming growth factor (e.g., TGF-a ,TGF-P, TGF-P1)
T-helper lymphocyte
stable isotope of thorium-232
I-lxvii
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ThO
Thl
Th2
THC
TIMS
TLC
TM
TML
TNF
tPA
TPALL
TPBS
TPY
TRH
TRI
TriAL
Trk
TSH
TSP
TSS
TT3
TT4
TTR
TWA
TX
235U, 238U
UC1P
UDP
UGT
UNECE
USGS
UV
VC
VCS
precursor T lymphocyte
T-derived lymphocyte helper 1
T-derived lymphocyte helper 2
CD4,CD8-positive T lymphocytes
thermal ionization mass spectrometry
Treatment of Lead-exposed Children (study)
T-memory lymphocyte
tetramethyllead
tumor necrosis factor (e.g., TNF-a, TNF-P1)
plasminogen activator
transfer rate from diffusible plasma to all destinations
Total Problem Behavior Score
tons per year
thyroid releasing hormone
Toxics Release Inventory
trialkyllead
tyrosine kinase receptor
thyroid stimulating hormone
total suspended particulates
total suspended solids
total triiodothyronine
serum total thyroxine
transthyretin
time-weighted average
tromboxane (e.g., TXB2)
uranium-234 and -238 radionuclides
plasma-to-urine clearance
uridine diphosphate
uridine diphosphate-glucuronyl transferases
United Nations Economic Commission for Europe
United States Geological Survey
ultraviolet
vital capacity
vinyl chloride stabilizer
I-lxviii
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Vd
VDR
VEP
VI
VLDL
VMI
VP
VSMC
WHO
WIC
WISC-III
WISC-R
WPPSI
WRAT-R
w/v
XAS
XPS
X-rays
XRD
XRF
ZPP
deposition velocity
vitamin D receptor
visual-evoked potential
variable-interval
very low density lipoprotein (cholesterol)
visual motor integration
plasma volume
vascular smooth muscle cells
World Health Organization
Women, Infants, and Children (program)
Wechsler Intelligence Scale for Children-Ill
Wechsler Intelligence Scale for Children-Revised
Wechsler Preschool and Primary Scale of Intelligence
Wide Range Achievement Test-Revised
weight per volume
X-ray absorption spectroscopy
X-ray photoelectron spectroscopy
synchrotron radiation
X-ray diffraction
X-ray fluorescence
zinc protoporphyrin
I-lxix
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EXECUTIVE SUMMARY
E.I INTRODUCTION

This document critically assesses the latest scientific information concerning health and

welfare effects associated with the presence of various concentrations of lead (Pb) in ambient air,

as pertinent to providing updated scientific bases for EPA's periodic review of the National

Ambient Air Quality Standards for Lead (Pb NAAQS). As such, this document builds upon

previous assessments published by the U.S. Environmental Protection Agency (EPA), including:

(a) the 1977 EPA document, Air Quality Criteria for Lead; (b) an updated revision of that Lead

Air Quality Criteria Document and an accompanying Addendum published in 1986 (1986 Lead
AQCD/Addendum); and (c) an associated 1990 Supplement to that 1986 AQCD/Addendum.

This document focuses on evaluation and integration of information relevant to Pb NAAQS

criteria development that has become available mainly since that covered by the 1986 and 1990

criteria assessments.

E.I.I Clean Air Act Legal Requirements

As discussed in Chapter 1 of this revised Lead AQCD, Sections 108 and 109 of the Clean

Air Act (CAA) govern establishment, review, and revision of U.S. National Ambient Air Quality

Standards (NAAQS):

• Section 108 directs the U.S. Environmental Protection Agency (EPA) Administrator to list
ubiquitous (widespread) air pollutants that may reasonably be anticipated to endanger public
health or welfare and to issue air quality criteria for them. The air quality criteria are to
reflect the latest scientific information useful in indicating the kind and extent of all
exposure-related effects on public health and welfare expected from the presence of the
pollutant in the ambient air.

• Section 109 directs the EPA Administrator to set and periodically revise, as appropriate, two
types of NAAQS: (a) primary NAAQS to protect against adverse health effects of listed
criteria pollutants among sensitive population groups, with an adequate margin of safety, and
(b) secondary NAAQS to protect against welfare effects (e.g., impacts on vegetation, crops,
ecosystems, visibility, climate, man-made materials, etc.). Section 109 also requires peer
review of the NAAQS and their underlying scientific bases by the Clean Air Scientific
Advisory Committee (CAS AC), a committee of independent non-EPA experts.
E-l
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E.1.2 Chronology of Lead NAAQS Revisions

• In 1971, U.S. EPA promulgated national ambient air standards for several major "criteria"
pollutants (see Federal Register, 1971) that did not include Pb at that time. Later, on October
5, 1978, the EPA promulgated primary and secondary NAAQS for Pb, as announced in the
Federal Register (1979). The primary and the secondary NAAQSs are the same: 1.5 |ig/m3
as a calendar quarterly average (maximum arithmetic mean averaged over 90 days). The
standards were based on EPA's 1977 document, Air Quality Criteria for Lead.

• In 1986, the EPA published a revised Lead AQCD that assessed newly available scientific
information published through December 1985. That 1986 document was mainly concerned
with Pb health and welfare effects, but other scientific data were also discussed to provide a
better understanding of the pollutant in the environment. Thus, the Lead AQCD included
chapters that discussed the atmospheric chemistry and physics of the pollutant; analytical
approaches; environmental concentrations; human exposure and dosimetry; physiological,
lexicological, clinical, epidemiological aspects of Pb health effects; and Pb effects on
ecosystems. An Addendum to the 1986 Lead AQCD was also published concurrently.

• Later, a supplement to the 1986 Lead AQCD/Addendum was published in 1990. That 1990
Supplement evaluated still newer information emerging in the published literature concerning
(a) Pb effects on blood pressure and other cardiovascular endpoints and (b) the effect of Pb
exposure during pregnancy and/or during the early postnatal period on birth outcomes and/or
on the neonatal physical and neuropsychological development of infants and children.

• Evaluations contained in the 1986 Lead AQCD/Addendum and 1990 Supplement provided
scientific inputs to support decision making regarding periodic review and, as appropriate,
revision of the Pb NAAQS, and they were drawn upon by EPA's Office of Air Quality
Planning and Standards (OAQPS) in preparation of a 1990 OAQPS Lead Staff Paper. After
consideration of evaluations contained in these documents, EPA chose not to propose
revision of the Pb NAAQS. At that time, as part of implementing a broad 1991 U.S. EPA
Strategy for Reducing Lead Exposure, the Agency focused on primarily regulatory and
remedial clean up efforts to reduce Pb exposure from a variety of non-air sources and media
judged to pose more extensive public health risks to U.S. populations than remaining air
emissions, as well as taking other actions to reduce Pb emissions to air. By 1990, annual
average ambient air Pb levels had dropped in U.S. urban areas to about 0.15-0.25 |ig/m3 due
to the phasedown of leaded gasoline.

• This revised Lead AQCD, prepared by EPA's National Center for Environmental
Assessment (NCEA), provides scientific bases to support Clean Air Act-mandated periodic
review of the Pb NAAQS. The document assesses the latest available scientific information
(published mainly through December 2005) judged to be useful in deriving criteria as
scientific bases for decisions on possible revision of the current Pb NAAQS.

• A separate EPA Lead Staff Paper, prepared by OAQPS in EPA's Office of Air and Radiation
(OAR), draws upon key findings/conclusions from this document and, together with other
analyses, develops and presents options for consideration by the EPA Administrator in regard
to review, and possible revision, of the Pb NAAQS.
E-2
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E.1.3 Document Organization and Structure

Volume I of this document consists of the present Executive Summary and eight main

chapters of this revised Lead AQCD. Those main chapters focus primarily on interpretative

evaluation of key information, whereas more detailed descriptive summarization of pertinent

studies and/or supporting analyses are provided in accompanying annexes. Volume II contains

(a) the annexes for Chapters 4, 5, and 6 (which assess dosimetric, toxicologic, and epidemiologic
evidence regarding Pb health effects) and (b) the annex for Chapter 7 (which assesses other

information on Pb ecological effects). Topics covered in the main chapters of the present AQCD

are as follows:

• This Executive Summary summarizes key findings and conclusions from Chapters 1 through
8 of this revised Lead AQCD, as they pertain to background information on Pb-related
atmospheric science and air quality, human exposure aspects, dosimetric considerations,
health effect issues, and environmental effect issues.

• Chapter 1 provides a general introduction, including an overview of legal requirements,
the chronology of past revisions of Pb-related NAAQS, and orientation to the structure of
this document.

• Chapters 2 and 3 provide background information on chemistry /physics of Pb, atmospheric
transport and fate, air quality, and multimedia exposure aspects to help to place the ensuing
discussions of Pb health and welfare effects into perspective.

• Chapters 4 through 6 assess dosimetry aspects, toxicologic (laboratory animal) studies, and
epidemiologic (observational) studies of Pb health effects.

• Chapter 7 assesses information concerning environmental effects of Pb on terrestrial and
aquatic ecosystems.

• Chapter 8 then provides an integrative synthesis of key findings and conclusions derived
from the preceding chapters with regard to ambient Pb concentrations, human exposures,
dosimetry, health effects of importance for primary Pb NAAQS decisions, and ecosystem
effects pertinent to secondary Pb NAAQS decisions.
E.2 LEAD SOURCES, EMISSIONS, AND CONCENTRATIONS AND
HUMAN MULTIMEDIA EXPOSURE PATHWAYS

Lead has been observed in measurable quantities in nearly every environmental medium
all over the world. Human exposure to Pb occurs through several routes, as shown in Figure E-l.
E-2
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SURFACE AND
GROUND WATER
DRINKING
WATER
Figure E-l. Simplified diagram of environmental pathways contributing to multimedia Pb
exposure of human populations.
That figure provides a simplified diagram of various routes of exposure through different
environmental media, with a main focus on the ambient air. The multimedia aspects of Pb
exposure can be seen in that Pb emissions to the air contribute to Pb concentrations in water, soil,
and dusts; Pb in soil and dust also can make important contributions to Pb concentrations in
ambient air. The relative contributions of Pb from different media and different sources on
human exposure depend on factors such as the proximity of major sources to the residence and
workplace of the individual, the condition of the residence (especially the presence and condition
of Pb-based paint) and whether the residence is in an urban, suburban, or rural location. This
section briefly summarizes available evidence concerning multimedia Pb sources and exposure
pathways, with main emphasis on pathways involving airborne Pb components.
E-4
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E.2.1 Ambient Lead Sources, Emissions, and Concentrations in the
United States

• Overall, current U.S. ambient Pb concentrations are generally well below the Pb NAAQS
level, except for locations influenced by local sources. During 2000 to 2004, on average,
quarterly mean Pb concentrations at Federal Reference Method monitors ranged from 0.10 to
0.22 |ig/m3 (including point source-related monitors). In the same time period, only one to
five locations from among -200 U.S. sites measured quarterly maximum Pb levels that
exceeded the NAAQS level (1.5 |ig/m3, quarterly max average) in any given year.

• Historically, mobile sources were a major source of Pb emissions, due to the use of leaded
gasoline. The United States initiated the phasedown of gasoline Pb additives in the late
1970s and intensified the phase-out of Pb additives in 1990. Accordingly, airborne Pb
concentrations have fallen dramatically nationwide, decreasing an average of 94% between
1983 and 2002. This is considered one of the great public and environmental health
successes. Remaining mobile source-related emissions of Pb include brake wear,
resuspended road dust, and emissions from vehicles that continue to use leaded gasoline
(e.g., some types of aircraft and race cars).

• Currently, the major stationary sources of Pb are in the industrial sector, including iron and
steel foundries; combustion sources, e.g., energy generation through coal and fuel oil
combustion, or wood combustion and hazardous or solid waste incineration; primary and
secondary Pb smelters; Pb-alloy production facilities; smelters for other metals, such as
copper or nickel; Pb-acid battery plants; and Pb mining and/or processing.

• The resuspension of soil-bound Pb particles and contaminated road dust is a significant
source of airborne Pb. In general, the main source of resuspension is wind and vehicular
traffic, although resuspension through other mechanical processes such as construction,
pedestrian traffic, agricultural operations, and even raindrop impaction is possible. Elevated
Pb levels are found in soil near stationary Pb sources and roadways that were heavily
trafficked prior to gasoline-Pb phasedown; and soil Pb can also be elevated near hazardous
waste cleanup sites.

• Lead can be transported in the atmosphere through mechanisms including deposition and
resuspension of Pb-containing particles. Dry deposition is the process by which pollutants
are removed from the atmosphere in the absence of precipitation. The size of depositing
particles is arguably the most important factor affecting dry deposition rates. Wet deposition
is the process by which airborne pollutants are scavenged by precipitation and removed from
the atmosphere. The size of particles can also influence wet deposition rates, with large
particles being scavenged more efficiently and, hence, tending to be removed closer to their
source of emission than small particles.
E.2.2 Multimedia Lead Exposure Pathways

• Exposure to Pb occurs through a number of routes. In addition to exposure to Pb in the air,
other major environmental routes for exposure to Pb include: Pb in drinking water;
E-5
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Pb-contaminated food; Pb in house dust; and Pb-based paint in older homes. Also, other Pb
exposure sources vary in their prevalence and potential risk, such as calcium supplements,
Pb-based glazes, and certain kinds of miniblinds, hair dye, and other consumer products.

• Lead-based paint exposure has long been one of the most common causes of clinical Pb
toxicity. Lead-based paint was the dominant form of house paint for many decades, and a
significant percentage of homes still contain Pb-based paint on some surfaces. Lead from
deteriorating paint can be incorporated into exterior residential soils and/or house dust.
The associated Pb exposure is often due to ingestion from hand-to-mouth activities and pica,
which are common in children. Inhalation Pb exposure of adults and children can also be
increased markedly during renovation or demolition projects.

• Although marked reductions of Pb in U.S. market basket food supplies have occurred during
the past several decades, Pb-contaminated food still can be a major route of Pb exposure for
some individuals. It was estimated that in 1990, North Americans ingested -50 jig of Pb
each day through food, beverages, and dust; with -30 to 50% of this amount via food and
beverages. With the elimination of Pb solder in U.S. canned food, food-Pb intake has fallen
dramatically in the United States. Recent studies indicate that dietary Pb intake in the United
States ranges from 2 to 10 ng/day. Some imported canned goods, especially from countries
where Pb-soldered cans are still not banned, can be a source of notable dietary-Pb intake for
some U.S. population groups, as can Pb-glazed storage pottery.

• Lead in drinking water results primarily from corrosion of Pb pipes, Pb-based solder, or brass
or bronze fixtures within a residence; very little Pb in drinking water comes from utility
supplies. Lead in drinking water, although generally found at low concentrations in the
United States, has been linked to elevated blood Pb concentrations in some U.S. locations.
In particular, the increasing use of chloramine in municipal water distribution systems, in
place of chlorination as a disinfection process, has led to notable elevations of Pb in tap water
in some U.S. communities in recent years.

• Given the large amount of time people spend indoors, exposure to Pb in dusts and indoor air
can be significant. For children, dust ingested via hand-to-mouth activity is often a more
important source of Pb exposure than inhalation. Dust can be resuspended through
household activities, thereby posing an inhalation risk as well. House dust Pb can derive
both from Pb-based paint and from other sources outside the home. The latter include
Pb-contaminated airborne particles from currently operating industrial facilities or
resuspended soil particles contaminated by deposition of airborne Pb from past emissions.

• In the United States, decreases in mobile sources of Pb, resulting from the phasedown of
gasoline Pb additives, created a 98% decline in emissions from 1970 to 2003. NHANES data
show a consequent parallel decline in blood-Pb levels in children aged 1 to 5 years from a
geometric mean of-15 ng/dL in 1976-1980 to -1-2 ng/dL in the 2000-2004 period.
E-6
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E.3. TOXICOKINETICS AND MEASUREMENT/MODELING OF
HUMAN EXPOSURE IMPACTS ON TISSUE DISTRIBUTION
OF LEAD

At the time of the 1986 Lead AQCD, it was noted that external Pb exposures via various

routes (inhalation, ingestion, dermal) were reflected by increased blood-Pb concentrations, which

served as a key biomarker of Pb exposure and index by which to judge risk of Pb-induced health

effects. It was also recognized (a) that Pb distributed to and accumulated in several bone

compartments and (b) that bone Pb might as a source of long-term internal exposure. Important

findings from newly available studies include the following:

• Blood Pb is found primarily (-99%) in red blood cells. It has been suggested that the small
fraction of Pb in plasma (<0.3%) may be the more biologically labile and lexicologically
active fraction of the circulating Pb. The relationship between Pb intake and blood Pb
concentration is curvilinear; i.e., the increment in blood Pb concentration per unit of Pb
intake decreases with increasing blood Pb concentration.

• New studies investigating the kinetics of Pb in bone have demonstrated that bone Pb serves
as a blood Pb source years after exposure and as a source of fetal Pb exposure during
pregnancy.

• Whereas bone Pb accounts for -70% of the body burden in children, more than 90% of the
total body burden of Pb is found in the bones in human adults. Lead accumulation is thought
to occur predominantly in trabecular bone during childhood and in both cortical and
trabecular bone in adulthood.

• A key issue of much importance in carrying out risk assessments that estimate the potential
likelihood of Pb-induced health effects is the estimation of external Pb-exposure impacts on
internal Pb tissue concentrations. This includes the estimation of typical Pb-exposure
impacts on internal distribution of Pb to blood and bone (as key biomarkers of Pb exposure),
as well as to other "soft tissue" target organs (e.g., brain, kidney, etc.).

• Earlier criteria assessments in the 1977 and 1986 Lead AQCDs extensively discussed the
available slope factor and/or other regression models of external Pb exposure impacts on
blood Pb concentration in human adults and children. Further refinements in regression
modeling of Pb impacts on blood or bone Pb are discussed in Chapter 4 of this document.

• The older slope factor analyses discussed in the 1977 and 1986 Lead AQCDs noted that at
relatively low air-Pb concentrations (<2 |ig/m3), pediatric blood-Pb levels generally increase
by -2 |ig/dL per each 1 |ig/m3 increment in air-Pb concentration.

• Several new empirical analyses have shown that a child's blood Pb is strongly associated
with interior dust Pb loading and its influence on hand Pb. Both exterior soil and paint Pb
contribute to interior dust Pb levels. Not all ingested Pb is absorbed to the same extent.
Factors such as an individual's age and diet, as well as chemical and physical properties of
E-7
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Pb compounds and the media in which they occur, affect absorption, e.g., absorption is
increased by fasting and dietary deficiencies in either iron or calcium. It has been estimated
that for every 1000 ppm increase in soil-Pb concentration, pediatric blood-Pb levels generally
increase by ~1 to 5 |ig/dL in infants and children <6 years old. However, intake of soil-Pb
with low bioaccessibility or bioavailability characteristics can yield distinctly lower-than-
typical blood-Pb increments.

• Information on Pb biokinetics, bone mineral metabolism, and Pb exposures has led to
refinements and expansions of pharmacokinetic models. Three pharmacokinetic models are
currently being used or are being considered for broad application in Pb risk assessment:
(1) the Integrated Exposure Uptake BioKinetic (TEUBK) model for Pb in children developed
by EPA (U.S. Environmental Protection Agency, 1994a,b; White et al., 1998); (2) the
Leggett model, which simulates Pb kinetics from birth through adulthood (Leggett, 1993);
and (3) the O'Flaherty model, which simulates Pb kinetics from birth through adulthood
(O'Flaherty, 1993, 1995).

• These models have been individually evaluated, to varying degrees, against empirical
physiological data on animals and humans and data on blood Pb concentrations in individuals
and/or populations (U.S. Environmental Protection Agency, 1994a,b; Leggett, 1993;
O'Flaherty, 1993). In evaluating models for use in risk assessment, exposure data collected
at hazardous waste sites have been used as inputs to some model simulations (Bowers and
Mattuck, 2001; Hogan et al., 1998). The exposure module in the IEUBK model makes this
type of evaluation feasible.

• Exposure-biokinetics models illustrate exposure-blood-body burden relationships and
provide a means for making predictions about these relationships that can be experimentally
or epidemiologically tested. The EPA IEUBK model for Pb has gained widespread use for
risk assessment purposes in the United States and is currently clearly the model of choice in
evaluating multimedia Pb exposure impacts on blood Pb levels and distribution of Pb to bone
and other tissues in young children <7 years old.

• The EPA All Ages Lead Model (AALM), now under development, aims to extend beyond
IEUBK capabilities to model external Pb exposure impacts (including over many years) on
internal Pb distribution not only in young children, but also in older children, adolescents,
young adults, and other adults well into older years. The AALM essentially uses adaptations
of IEUBK exposure module features, coupled with adaptations of IEUBK biokinetics
components (for young children) and of Leggett model biokinetics components (for older
children and adults). However, the AALM has not yet undergone sufficient development and
validation for its use yet beyond research and validation purposes.
E.4 HEALTH EFFECTS ASSOCIATED WITH LEAD EXPOSURE

Both epidemiologic and toxicologic studies have shown that environmentally relevant
levels of Pb affect many different organ systems. Research completed since the 1986
E-S
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AQCD/Addendum and 1990 Supplement indicates that Pb effects occur at blood-Pb even lower

than those previously reported for many endpoints. Remarkable progress has been made since

the mid-1980s in understanding the Pb effects on health. Recent studies have focused on details

of the associations, including the shapes of concentration-response relationships, especially at

levels well within the range of general population exposures, and on those biological and/or

socioenvironmental factors that either increase or decrease an individual's risk. Key findings

and conclusions regarding important outcomes of newly available toxicologic and epidemiologic

studies of Pb health effects are highlighted below.

Neurotoxic Effects of Pb Exposure

• Neurobehavioral effects of Pb-exposure early in development (during fetal, neonatal, and
later postnatal periods) in young infants and children (<7 years old) have been observed with
remarkable consistency across numerous studies involving varying study designs, different
developmental assessment protocols, and diverse populations. Negative Pb impacts on
neurocognitive ability and other neurobehavioral outcomes are robust in most recent studies
even after adjustment for numerous potentially confounding factors (including quality of care
giving, parental intelligence, and socioeconomic status). These effects generally appear to
persist into adolescence and young adulthood.

• The overall weight of the available evidence provides clear substantiation of neurocognitive
decrements being associated in young children with blood-Pb concentrations in the range of
5-10 |ig/dL, and possibly somewhat lower. Some newly available analyses appear to show
Pb effects on the intellectual attainment of preschool and school age children at population
mean concurrent blood-Pb levels ranging down to as low as 2 to 8 |ig/dL. A decline of
6.2 points in full scale IQ for an increase in concurrent blood Pb levels from 1 to 10 |ig/dL
has been estimated, based on a pooled analysis of results derived from seven well-conducted
prospective epidemiologic studies.

• In the limited literature examining the effects of environmental Pb exposure on adults, mixed
evidence exists regarding associations between Pb and neurocognitive performance.
No associations were observed between cognitive performance and blood Pb levels;
however, significant associations were observed in relation to bone Pb concentrations,
suggesting that long-term cumulative Pb exposure may contribute to neurocognitive deficits
in adults.

• Animal toxicology data indicate that developmental Pb exposures creating steady-state
blood-Pb concentrations of-10 |ig/dL result in behavioral impairments that persist into
adulthood in rats and monkeys. No evident threshold has yet been found; and Pb-induced
deficits, for the most part, have been found to be very persistent, even with various chelation
treatments. However, experimental studies indicate that environmental enrichment during
development can partially mitigate the effects of Pb on cognitive function. In rats,
neurobehavioral deficits that persisted well into adulthood were observed with prenatal,
E-9
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preweaning, and postweaning Pb exposure. In monkeys, such neurobehavioral deficits were
observed both with in utero-only exposure and with early postnatal-only exposure when peak
blood-Pb levels did not exceed 15 |ig/dL and steady-state levels were -11 |ig/dL.

• Learning impairment has been observed in animal studies at blood levels as low as 10 |ig/dL,
with higher level learning showing greater impairment than simple learning tasks. The
mechanisms associated with these deficits include: response perseveration; insensitivity to
changes in reinforcement density or contingencies; deficits in attention; reduced ability to
inhibit inappropriate responding; impulsivity; and distractibility.

• Lead affects reactivity to the environment and social behavior in both rodents and nonhuman
primates at blood Pb levels of 15 to 40 |ig/dL. Rodent studies also show that Pb exposure
potentiates the effects of stress in females.

• Auditory function has also been shown to be impaired at blood Pb levels of 33 |ig/dL, while
visual functions are affected at 19 |ig/dL.

• Neurotoxicological studies in animals clearly demonstrated that Pb mimics calcium and
affects neurotransmission and synaptic plasticity.

• Epidemiologic studies have identified genetic polymorphisms of two genes that may alter
susceptibility to the neurodevelopmental consequences of Pb exposure in children. Variant
alleles of the ALAD gene are associated with differences in absorption, retention, and
toxicokinetics of Pb. Polymorphisms of the vitamin D receptor gene have been shown to
affect the rate of resorption and excretion of Pb over time. These studies are only suggestive,
and parallel animal studies have not been completed.
Cardiovascular Effects of Lead

• Epidemiologic studies have consistently demonstrated associations between Pb exposure and
enhanced risk of deleterious cardiovascular outcomes, including increased blood pressure
and incidence of hypertension. A meta-analysis of numerous studies estimates that a
doubling of blood-Pb level (e.g., from 5 to 10 |ig/dL) is associated with -1.0 mm Hg increase
in systolic blood pressure and -0.6 mm Hg increase in diastolic pressure. Studies have also
found that cumulative past Pb exposure (e.g., bone Pb) may be as important, if not more, than
present Pb exposure in assessing cardiovascular effects. The evidence for an association of
Pb with cardiovascular morbidity and mortality is limited but supportive.

• Experimental toxicology studies have confirmed Pb effects on cardiovascular functions.
Most have shown that exposures creating blood-Pb levels of-20 to 30 |ig/dL for long
periods result in arterial hypertension that persists long after cessation of Pb exposure in
genetically normal animals. One study reported blood pressure increases at blood-Pb levels
as low as 2 |ig/dL in rats. A number of in vivo and in vitro studies provide compelling
evidence for the role of oxidative stress in the pathogenesis of Pb-induced hypertension.
However, experimental investigations of cardiovascular effects of Pb in animals are unclear
as to why low, but not high, levels of Pb exposure cause hypertension in experimental
animals.
E-10
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Renal Effects of Lead

• In the general population, both circulating and cumulative Pb was found to be associated
with longitudinal decline in renal function. Effects on creatine clearance have been reported
in human adult hypertensives to be associated with general population mean blood-Pb levels
of only 4.2 |ig/dL. The public health significance of such effects is not clear, however, in
view of more serious signs of kidney dysfunction being seen in occupationally exposed
workers only at much higher blood-Pb levels (>30-40 |ig/dL).

• Experimental studies using laboratory animals demonstrated that the initial accumulation of
absorbed Pb occurs primarily in the kidneys. This takes place mainly through glomerular
filtration and subsequent reabsorption, and, to a small extent, through direct absorption from
the blood. Both low dose Pb-treated animals and high dose Pb-treated animals showed a
"hyperfiltration" phenomenon during the first 3 months of Pb exposure. Investigations into
biochemical alterations in Pb-induced renal toxicity suggested a role for oxidative stress and
involvement of NO, with a significant increase in nitrotyrosine and substantial fall in urinary
excretion of NOX.

• Iron deficiency increases intestinal absorption of Pb and the Pb content of soft tissues and
bone. Aluminum decreases kidney Pb content and serum creatinine in Pb-intoxicated
animals. Age also has an effect on Pb retention. There is higher Pb retention at a very young
age and lower bone and kidney Pb at old age, attributed in part to increased bone resorption
and decreased bone accretion and, also, kidney Pb.
Effects of Lead on the Immune System

• Findings from recent epidemiologic studies suggest that Pb exposure may be associated with
effects on cellular and humoral immunity. These include changes in serum immunoglobulin
levels. Studies of biomarkers of humoral immunity in children have consistently found
significant associations between increasing blood-Pb concentrations and serum IgE levels at
blood-Pb levels <10 jig/dL.

• Toxicologic studies have shown that Pb targets immune cells, causing suppression of delayed
type hypersensitivity response, elevation of IgE, and modulation of macrophages into a
hyper-inflammatory phenotype. These types of changes can cause increased risk of atopy,
asthma, and some forms of autoimmunity and reduced resistance to some infectious diseases.
Lead exposure of embryos resulting in blood-Pb levels <10 |ig/dL can produce persistent
later-life immunotoxicity.
Effects of Lead on Heme Synthesis

• Lead exposure has been associated with disruption of heme synthesis in both children and
adults. A 10% probability of anemia (hematocrit <35%) is estimated to be associated with a
blood-Pb level of-20 |ig/dL at age 1 year. Increases in blood Pb concentration of about
20-30 |ig/dL are sufficient to halve erythrocyte ALAD activity and sufficiently inhibit
ferrochelatase to double erythrocyte protoporphyrin levels.
E-ll
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• Toxicological studies demonstrated that Pb intoxication interferes with red blood cell (RBC)
survival and alters RBC mobility. Hematological parameters, such as mean corpuscular
volume, mean corpuscular hemoglobin, and mean corpuscular hemoglobin concentration, are
also significantly decreased upon exposure to Pb. These effects are due to internalization of
Pb by RBC. The transport of Pb across the RBC membrane is energy-independent and
carrier-mediated; and the uptake of Pb appears to be mediated by an anion exchanger through
a vanadate-sensitive pathway.

• Erythrocyte ALAD activity ratio (ratio of activated/non activated enzyme activity) has been
shown to be a sensitive, dose-responsive measure of Pb exposure, regardless of the mode of
administration of Pb. Competitive enzyme kinetic analyses in RBCs from both humans and
Cynomolgus monkeys indicated similar inhibition profiles by Pb.
Effects of Lead on Bones and Teeth

• Experimental studies in animals demonstrate that Pb substitutes for calcium and is readily
taken up and stored in the bone and teeth of animals, potentially allowing bone cell function
to be compromised both directly and indirectly by exposure.

• Relatively short term exposure of mature animals to Pb does not result in significant growth
suppression. However, chronic Pb exposure during times of inadequate nutrition has been
shown to adversely influence bone growth, including decreased bone density, decreased
trabecular bone, and growth plates.

• Exposure of developing animals to Pb during gestation and the immediate postnatal period
has clearly been shown to significantly depress early bone growth in a dose-dependent
fashion, though this effect is not manifest below a certain threshold.

• Systemically, Pb has been shown to disrupt mineralization of bone during growth, to alter
calcium binding proteins, and to increase calcium and phosphorus concentration in the blood
stream, in addition to potentially altering bone cell differentiation and function by altering
plasma levels of growth hormone and calciotropic hormones such as vitamin D3 [1,25-
(OH2)D3.

• Periods of extensive bone remodeling, such as occur during weight loss, advanced age,
altered metabolic state, and pregnancy and lactation are all associated with mobilization
of Pb stores from bone of animals.

• Numerous epidemiologic studies and, separately, animal studies (both post-eruptive Pb
exposure and pre- and perinatal Pb exposure studies) suggest that Pb is a caries-promoting
element. However, whether Pb incorporation into the enamel surface compromises the
integrity and resistance of the surface to dissolution, and ultimately increases risk of dental
decay, is unclear.

• Increased risk of dental caries has been associated with Pb exposure in children and adults.
Lead effects on caries were observed in populations whose mean blood-Pb levels were less
than 10 |ig/dL.
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Reproductive and Developmental Effects of Lead

• Epidemiologic evidence suggests small associations between Pb exposure and male
reproductive outcomes, including perturbed semen quality and increased time to pregnancy.
There are no adequate epidemiologic data to evaluate associations between Pb exposure and
female fertility. Most studies have yielded no associations, or weak associations, of Pb
exposure with thyroid hormone status and male reproductive endocrine status in highly
exposed occupational populations.

• New toxicologic studies support earlier conclusions, presented in the 1986 Lead AQCD, that
(a) Pb can produce both temporary and persisting effects on male and female reproductive
function and development and (b) Pb disrupts endocrine function at multiple points along the
hypothalamic-pituitary-gonadal axis. Although there is evidence for a common mode of
action, consistent effects on circulating testosterone levels are not always observed in
Pb-exposed animals. Inconsistencies in reports of circulating testosterone levels complicate
derivation of a dose-response relationship for this endpoint.

• Lead-induced testicular damage (ultrastructural changes in testes of monkeys at blood-Pb
>35 to 40 |ig/dL) and altered female sex hormone release, imprinting during early
development, and altered female fertility all suggest Pb-induced reproductive effects.
However, Pb exposure does not generally produce total sterility. Pre- and postnatal exposure
to Pb has been demonstrated to result in fetal mortality and produce a variety of sublethal
effects in the offspring. Many of the Pb-induced sublethal developmental effects occur at
maternal blood-Pb levels that do not result in clinical (overt) toxicity in the mothers.
Teratogenic effects resulting from Pb exposure reported in a few studies appear to be
confounded by maternal toxicity.
Lead Effects on Other Organ Systems

• Lead impacts the hypothalamic-pituitary-adrenal axis, elevating corticosterone levels and
altering stress responsivity. This may be a potential mechanism contributing to Pb-induced
hypertension, with further possible roles in the etiology of diabetes, obesity and other
disorders.

• Studies of hepatic enzyme levels in serum suggest that liver injury may be present in Pb
workers; however, associations specifically with Pb exposures are not evident. Children
exposed to relatively high levels of Pb (blood Pb >30 |ig/dL) exhibit depressed levels of
circulating 1,25-dihydroxy vitamin D (1,25-OH-D). However, associations between serum
vitamin D status and blood Pb were not evident in a study of calcium-replete children who
had average lifetime blood-Pb concentrations <25 |ig/dL.

• Field studies that evaluated hepatic enzyme levels in serum suggest that liver injury may be
present in Pb workers; however, associations specifically with Pb exposures have not been
well established.
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• Simultaneous induction of the activities of phase II drug metabolizing enzymes and
decreased phase I enzymes with a single exposure to Pb nitrate in rat liver suggest that Pb is
capable of causing biochemical phenotype similar to hepatic nodules.

• Newer studies examined the induction of GST-P at both transcript onal and translational
levels using in vitro systems and indicated a role for Pb-nitrate and Pb-acetate in the
induction process.

• Lead-induced alterations in cholesterol metabolism appear to be mediated by the induction of
several enzymes related to cholesterol metabolism and the decrease of 7 a-hydroxylase, a
cholesterol catabolizing enzyme. This regulation of cholesterol homeostasis is modulated by
changes in cytokine expression and related signaling.

• Newer experimental evidence suggests that Pb-induced alterations in liver heme metabolism
involve perturbations in ALAD activity, porphyrin metabolism, alterations in Transferrin
gene expression, and associated changes in iron metabolism.

• Gastrointestinal (GI) absorption of Pb is influenced by a variety of factors, including
chemical and physical forms of the element in ingested media, age at intake, and various
nutritional factors. The degeneration of intestinal mucosal epithelium leading to potential
malabsorption and alterations in the jejunal ultrastructure (possibly associated with distortion
of glycocalyx layer) have been reported in the intestine of Pb-exposed rats.

• Nutritional studies that varied Pb, Ca, and vitamin D levels in the diet have demonstrated
competition of Pb with Ca absorption. Supplementation with vitamin D has been reported to
enhance intestinal absorption of Ca and Pb. Physiological amounts of vitamin D, when
administered to vitamin D-deficient rats, resulted in elevated Pb and Ca levels. In the case of
severe Ca deficiency, Pb ingestion results in a marked decrease in serum 1,25 hydroxy
vitamin D.
Genotoxic and Carcinogenic Effects of Lead

• Epidemiologic studies of highly exposed occupational populations suggest a relationship
between Pb and cancers of the lung and the stomach; however the evidence is limited by the
presence of various potential confounders, including metal coexposures (e.g., to arsenic,
cadmium), smoking, and dietary habits. The 2003 NTP and 2004 IARC reviews concluded
that Pb and Pb compounds were probable carcinogens, based on limited evidence in humans
and sufficient evidence in animals. Similarly, Pb and Pb compounds would likely be
classified as likely to be carcinogenic to humans according to the new 2005 EPA Cancer
Assessment Guidelines for Carcinogen Risk Assessment, based on animal data even though
the human data are inadequate.

• Studies of genotoxicity consistently find associations of Pb exposure with DNA damage and
micronuclei formation; however, the associations with the more established indicator of
cancer risk, chromosomal aberrations, are inconsistent.
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• Pb is an animal carcinogen and extends our understanding of mechanisms involved to
include a role for metallothionein. Specifically, the recent data show that metallothionein
may participate in Pb inclusion bodies and, thus, serves to prevent or reduce Pb-induced
tumorigenesis.

• In vitro cell culture studies that evaluated the potential for Pb to transform rodent cells are
inconsistent, and careful study of a time course of exposure is necessary to determine
whether Pb actually induces transformation in cultured rodent cells. There is increased
evidence suggesting that Pb may be co-carcinogenic or promotes the carcinogenicity of other
compounds. Cell culture studies do support a possible epigenetic mechanism or
co-mutagenic effects.
Lead-Binding Proteins

• Proteins depending upon sulfur-containing side chains for maintaining conformity or activity
are vulnerable to inactivation by Pb, due to its strong sulfur-binding affinity.

• The enzyme, ALAD, a 280 kDa protein, is inducible and is the major Pb-binding protein
within the erythrocyte.

• The Pb-binding protein in rat kidney has been identified as a cleavage product of a-2
microglobulin. The low molecular weight Pb-binding proteins in human kidney have been
identified as thymosin 04 (molecular weight 5 kDa) and acyl-CoA binding protein
(molecular weight 9 kDa). In human brain, Pb-binding proteins include thymosin P4 and an
unidentified protein of 23 kDa.

• Animal toxicology studies with metallothionein-null mice demonstrated a possible role for
metallothionein as a renal Pb-binding protein.
E.5 HUMAN POPULATION GROUPS AT SPECIAL RISK AND
POTENTIAL PUBLIC HEALTH IMPACTS

• Children, in general and especially low SES (often including larger proportions of African-
American and Hispanic) children, have been well-documented as being at increased risk for
Pb exposure and Pb-induced adverse health effects. This is due to several factors, including
enhanced exposure to Pb via ingestion of soil-Pb and/or dust-Pb due to normal hand-to-
mouth activity and/or pica.

• Even children with low Pb exposure levels (having blood Pb of 5-10 |ig/dL or, possibly,
somewhat lower) are at notable risk, due to apparent non-linear dose-response relationships
between blood Pb and neurodevelopmental outcomes. It is hypothesized that initial
neurodevelopmental lesions occurring at blood-Pb levels <10 |ig/dL may disrupt different
developmental processes in the nervous system than more severe high level exposures.
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• Adults with idiosyncratic exposures to Pb through occupations, hobbies, make-up use, glazed
pottery, native medicines, and other sources are at risk for Pb toxicity. Certain ethnic and
racial groups are known to have cultural practices that involve ingestion of Pb-containing
substances, e.g., ingestion of foods or beverages stored in Pb-glazed pottery or imported
canned food from countries that allow Pb-soldered cans.

• Cumulative past Pb exposure, measured by bone Pb, may be a better predictor of
cardiovascular effects than current blood-Pb levels. African-Americans are known to have
substantially higher baseline blood pressure than other ethnic groups, so Pb's impact on an
already higher baseline could indicate a greater susceptibility to Pb for this group.

• Effects on adults of low-level Pb exposures also include some renal effects (i.e., altered
creatinine clearance) at blood-Pb levels <5 ug/dL. Lead exposure combined with other risk
factors, such as diabetes, hypertension, or chronic renal insufficiency may result in clinically-
relevant effects in individuals with two or more other risk factors.

• At least two genetic polymorphisms, of the ALAD and the vitamin D receptor gene, have
been suggested to play a role in susceptibility to Pb. In one study, African-American
children were found to have a higher incidence of being homozygous for alleles of the
vitamin D receptor gene thought to contribute to greater Pb blood levels. This work is
preliminary and further studies will be necessary to determine implications of genetic
differences that may make certain populations more susceptible to Pb exposure.

• What was considered "low" for Pb exposure levels in the 1980s is an order of magnitude
higher than the current mean level in the U.S. population, and current average blood-Pb
levels in U.S. populations remain perhaps as much as two orders of magnitude above pre-
industrial "natural" levels in humans. There is no level of Pb exposure that has yet been
identified, with confidence, as being clearly not being associated with possible risk of
deleterious health effects. Some recent studies of Pb neurotoxicity in infants have observed
effects at population average blood-Pb levels of only 1 or 2 |ig/dL; and some cardiovascular,
renal, and immune outcomes have been reported at blood-Pb levels below 5 |ig/dL.

• Public health interventions have resulted in declines, over the last 25 years, of more than
90% in the mean blood-Pb level within all age and gender subgroups of the U.S. population,
substantially decreasing the numbers of individuals at likely risk for Pb-induced toxicities.
Nevertheless, estimates of the magnitude of potential public health impacts of Pb exposure
can be substantial for the U.S. population. For example, in estimating the effect of Pb
exposure on intelligence, it was projected that the fraction of individuals with an IQ >120
would decrease from -9% with no Pb exposure to less than 3% at a blood-Pb level of
10 |ig/dL. Also, the fraction of individuals with an IQ >130 points was estimated as being
likely to decrease from 2.25% to 0.5% with a blood-Pb level change from 0 to 10 |ig/dL.
In addition, an estimate of hypertension-related risk for serious cardiovascular events
(coronary disease, stroke, peripheral artery disease, cardiac failure) indicates that a decrease
in blood Pb from 10 to 5 |ig/dL could result in an annual decrease of 27 events per
100,000 women and 39 events per 100,000 men.
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E.6 ENVIRONMENTAL EFFECTS OF LEAD

Chapter 7 assesses the environmental effects of Pb, including discussion, in particular,

of Pb effects on terrestrial and aquatic ecosystems and the methodological approaches used to

study such effects.

E.6.1 Terrestrial Ecosystems

Methodologies Used in Terrestrial Ecosystem Research

• Electron probe microanalysis (EPMA) techniques provide the greatest information on metal
speciation. Other techniques, such as EXAFS (extended X-ray absorption fine structure) and
EXANES (extended X-ray absorption near edge spectroscopy), show great promise and will
be important in solving key mechanistic questions.

• In situ methodologies have been developed to lower soil-Pb relative bioavailability. These
amendments typically fall within the categories of phosphate, biosolid, and Al/Fe/Mn-oxide
amendments. Some of the drawbacks to soil amendment include phosphate toxicity to plants
and increased arsenic mobility at high soil phosphate concentrations. The use of iron(III)
phosphate seems to mitigate arsenic mobility, however increased concentrations of phosphate
and iron limit their application when drinking water quality is a concern.

Distribution of Atmospherically Delivered Lead in Terrestrial Ecosystems

• Total Pb deposition during the 20th century has been estimated at 1 to 3 g Pb nT2, depending
on elevation and proximity to urban areas. Total contemporary loadings to terrestrial
ecosystems are ~1 to 2 mg nT2 year"1. This is a relatively small annual flux of Pb compared
to the reservoir of-0.5 to 4 g nT2 of gasoline additive-derived Pb already deposited in
surface soils over much of the United States.

• Dry deposition can account for 10% to >90% of total Pb deposition. Because Clean Air Act
Legislation has preferentially reduced Pb associated with fine particles, relative contributions
of dry deposition have changed in the last few decades.

• Although inputs of Pb to ecosystems are currently low, Pb export from watersheds via
groundwater and streams is substantially lower than inputs. Therefore, even at current input
levels, watersheds are accumulating anthropogenic Pb.

• Species of Pb delivered to terrestrial ecosystems can be inferred by emission source.
For example, Pb species emitted from automobile exhaust are dominated by particulate Pb
halides and double salts with ammonium halides (e.g., PbBrCl, PbBrChNFLCl), while Pb
emitted from smelters is dominated by Pb-sulfur species. Halides from automobile exhaust
break down rapidly in the atmosphere, via redox reactions in the presence of atmospheric
acids. Lead phases in the atmosphere, and presumably the compounds delivered to the
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surface of the earth (i.e., to vegetation and soils), are suspected to be in the form of PbSC>4,
PbS, andPbO.

• The importance of humic and fulvic acids and hydrous Mn- and Fe-oxides for scavenging Pb
in soils was discussed in some detail in the 1986 Lead AQCD. The importance of these Pb
binding substrates is reinforced by studies reported in the more contemporary literature.

• The amount of Pb that has leached into mineral soil appears to be on the order of 20 to 50%
of the total anthropogenic Pb deposition.

• The vertical distribution and mobility of atmospheric Pb in soils was poorly documented
prior to 1986. Techniques using radiogenic Pb isotopes have been developed to differentiate
between gasoline-derived Pb and natural, geogenic (native) Pb. These techniques provide
more accurate determinations of the depth-distribution and potential migration velocities for
atmospherically delivered Pb in soils.

• Selective chemical extractions have been used extensively over the past 20 years to quantify
amounts of a particular metal phase in soil or sediment rather than total metal concentration.
However, some problems persist with the selective extraction technique: (a) extractions are
rarely specific to a single phase; and (b) in addition to the nonselectivity of reagents,
significant metal redistribution has been found to occur during sequential chemical
extractions. Thus, although chemical extractions provide some useful information on metal
phases in soil or sediment, the results should be treated as "operationally defined," e.g.,
"H2O2-liberated Pb" rather than "organic Pb."

• Soil solution dissolved organic matter content and pH typically have very strong positive and
negative correlations, respectively, with the concentration of dissolved Pb species.
Effects of Lead on Natural Terrestrial Ecosystems

• Atmospheric Pb pollution has resulted in the accumulation of Pb in terrestrial ecosystems
throughout the world. In the United States, anthropogenically-derived Pb represents a
significant fraction of the total Pb burden in soils, even in sites remote from smelters and
other industrial plants. However, few significant effects of Pb pollution have been observed
at sites that are not near point sources of Pb.

• Evidence from precipitation collection and sediment analyses indicates that atmospheric
deposition of Pb has declined dramatically (>95%) at sites unaffected by point sources of Pb,
and there is little evidence that Pb accumulated in soils at these sites represents a threat to
ground water or surface water supplies.

• The effects of Pb and other chemical emissions on terrestrial ecosystems near smelters and
other industrial sites decrease downwind from the Pb source. Several studies using the soil
burden as an indicator have shown that much of the contamination occurs within a radius of
20 to 50 km around the emission source. Elevated metal concentrations around smelters
have been found to persist despite significant reductions in emissions. The concentrations of
Pb in soils, vegetation, and fauna at these sites can be two to three orders of magnitude
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higher than in reference areas. Assessing the risks specifically associated with Pb is difficult,
because these sites also experience elevated concentrations of other metals and because of
effects related to SC>2 emissions. The confounding effect of other pollutants makes the
assessment of Pb-specific exposure-response relationships impossible at the whole-
ecosystem level.

• In the most extreme cases, near smelter sites, the death of vegetation causes a near-complete
collapse of the detrital food web, creating a terrestrial ecosystem in which energy and
nutrient flows are minimal.

• More commonly, stress in soil microorganisms and detritivores can cause reductions in the
rate of decomposition of detrital organic matter. Although there is little evidence of
significant bioaccumulation of Pb in natural terrestrial ecosystems, reductions in microbial
and detritivorous populations can affect the success of their predators. Thus, at present,
industrial point sources represent the greatest Pb-related threat to the maintenance of
sustainable, healthy, diverse, and high-functioning terrestrial ecosystems in the United States.
Terrestrial Species Response/Mode of Action

• Plants take up Pb via their foliage and through their root systems. Surface deposition of Pb
onto plants may represent a significant contribution to the total Pb in and on the plant, as has
been observed for plants near smelters and along roadsides.

• There are two possible mechanisms (symplastic or apoplastic) by which Pb may enter the
root of a plant. The symplastic route is through the cell membranes of root hairs; this is the
mechanism of uptake for water and nutrients. The apoplastic route is an extracellular route
between epidermal cells into the intercellular spaces of the root cortex. The symplastic route
is considered the primary mechanism of Pb uptake in plants.

• Recent work supports previous conclusions that the form of metal tested, and its speciation in
soil, influence uptake and toxicity to plants and invertebrates. The oxide form of Pb is less
toxic than the chloride or acetate forms, which are less toxic that the nitrate form of Pb.
However, these results must be interpreted with caution, as the counter ion (e.g., the nitrate
ion) may also be contributing to the observed toxicity.

• Lead may be detoxified in plants by deposition in root cell walls, and this may be influenced
by calcium concentrations. Other hypotheses put forward recently include the presence of
sulfur ligands and the sequestration of Pb in old leaves as detoxification mechanisms. Lead
detoxification has not been studied extensively in invertebrates. Glutathione detoxification
enzymes were measured in two species of spider. Lead may be stored in waste nodules in
earthworms or as pyromorphite in the nematode.

• Lead effects on heme synthesis (as measured primarily by ALAD activity and protoporphyrin
concentration) were documented in the 1986 Lead AQCD and continue to be studied.
However, researchers caution that changes in ALAD and other enzyme parameters are not
always related to adverse effects, but may simply indicate exposure. Other effects on plasma
enzymes, which may damage other organs, have been reported. Lead also may cause lipid
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peroxidation, which may be alleviated by vitamin E, although Pb poisoning may still result.
Changes in fatty acid production have been reported, which may influence immune response
and bone formation.

Insectivorous mammals may be more exposed to Pb than herbivores, and higher trophic-level
consumers may be less exposed than lower trophic-level organisms. Nutritionally-deficient
diets (including low calcium) cause increased uptake of Pb and greater toxicity in birds.

Interactions of Pb with other metals are inconsistent, depending on the endpoint measured,
the tissue analyzed, the animal species, and the metal combination.
Exposure/Response of Terrestrial Species

• Recent critical advancements reported in the current Lead AQCD in understanding toxicity
levels relies heavily on the work completed by a multi-stakeholder group, consisting of
federal, state, consulting, industry, and academic participants, led by the EPA to develop
Ecological Soil Screening Levels (Eco-SSLs).

• Eco-SSLs are concentrations of contaminants in soils that would result in little or no
measurable effect on ecological receptors. The Eco-SSLs are intentionally conservative in
order to provide confidence that contaminants which could present an unacceptable risk are
not screened out early in the evaluation process. That is, at or below these levels, adverse
effects are considered unlikely. Due to conservative modeling assumptions (e.g., metal exists
in most toxic form or highly bioavailable form, high food ingestion rate, high soil ingestion
rate) which are common to screening processes, several Eco-SSLs are derived below the
average background soil concentration for a particular contaminant.

• The Eco-SSLs for terrestrial plants, birds, mammals, and soil invertebrates are 120, 11, 56,
and 1700 mg Pb/kg soil, respectively.
E.6.2 Aquatic Ecosystems

Methodologies Used in Aquatic Ecosystem Research

• Many of the terrestrial methods can also be applied to suspended solids and sediments
collected from aquatic ecosystems. Just as in the terrestrial environment, the speciation of Pb
and other trace metals in natural freshwaters and seawater plays a crucial role in determining
their reactivity, mobility, bioavailability, and toxicity. Many of the same speciation
techniques employed for the speciation of Pb in terrestrial ecosystems are applicable in
aquatic ecosystems.

• There is now a better understanding of the potential effects of sampling, sample handling,
and sample preparation on aqueous-phase metal speciation. Thus, a need has arisen for
dynamic analytical techniques that are able to capture a metal's speciation, in-situ and in real
time.
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• With few exceptions, ambient water quality criteria (AWQC) are derived based on data from
aquatic toxicity studies conducted in the laboratory. In general, both acute (short term) and
chronic (long term) AWQCs are developed. Depending on the species, the toxicity studies
considered for developing acute criteria range in length from 48 to 96 hours.

• Acceptable chronic toxicity studies should encompass the full life cycle of the test organism,
although for fish, early life stage or partial life cycle toxicity studies are considered
acceptable. Acceptable endpoints include reproduction, growth and development, and
survival, with the effect levels expressed as the chronic value.

• The biotic ligand model (BLM), which considers the binding of free metal ion to the site of
toxic action and competition between metal species and other ions, has been developed to
predict the toxicity of several metals under a variety of water quality conditions. However,
there are limitations to this tool in deriving AWQC because, currently, limited work has been
conducted in developing chronic BLMs (for any metals, let alone Pb) and the acute BLMs to
date do not account for dietary metal exposures.
Distribution of Lead in Aquatic Ecosystems

• Atmospheric Pb is delivered to aquatic ecosystems primarily through deposition (wet and/or
dry) or through erosional transport of soil particles.

• A significant portion of Pb in the aquatic environment exists in the undissolved form (i.e.,
bound to suspended particulate matter). The ratio of Pb in suspended solids to Pb in filtrate
varies from 4:1 in rural streams to 27:1 in urban streams.

• The oxidation potential of Pb is high in slightly acidic solutions, and Pb2+ binds with high
affinity to sulfur-, oxygen-, and nitrogen-containing ligands. Therefore, speciation of Pb in
the aquatic environment is controlled by many factors (e.g., pH, redox, dissolved organic
carbon, sulfides). The primary form of Pb in aquatic environments is divalent (Pb2+), while
Pb4+ exists only under extreme oxidizing conditions. Labile forms of Pb (e.g., Pb2+, PbOH+,
PbCOs) are a significant portion of the Pb inputs to aquatic systems from atmospheric
washout. Lead is typically present in acidic aquatic environments as PbSC>4, PbCU, ionic Pb,
cationic forms of Pb-hydroxide, and ordinary Pb-hydroxide (Pb(OH)2). In alkaline waters,
common species of Pb include anionic forms of Pb-carbonate (Pb(CC>3)) and Pb(OH)2.

• Lead concentrations in lakes and oceans were generally found to be much lower than those
measured in the lotic waters assessed by NAWQA. In open waters of the North Atlantic the
decline of Pb concentrations has been associated with the phasing out of leaded gasoline in
North America and Western Europe. However, in estuarine systems, it appears that similar
declines following the phase-out of leaded gasoline are not necessarily as rapid.

• Based on a synthesis of NAWQA data from the United States, Pb concentrations in surface
waters, sediments, and fish tissues (whole body) respectively range from: 0.04 to 30 jig/L
(mean = 0.66, median = 0.50, 95th %tile =1.1); 0.5 to 12,000 mg/kg (mean = 120, median =
28, 95th %tile = 200); and 0.08 to 23 mg/kg (mean = 1.03, median = 0.59, 95th %tile = 3.24).
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Effects of Lead on Natural Aquatic Ecosystems

• Lead exposure may adversely affect organisms at different levels of organization, i.e.,
individual organisms, populations, communities, or ecosystems. Generally, however, there is
insufficient information available for single materials in controlled studies to permit
evaluation of specific impacts on higher levels of organization (beyond the individual
organism). Potential effects at the population level or higher are, of necessity, extrapolated
from individual level studies. Available population, community, or ecosystem level studies
are typically conducted at sites that have been contaminated or adversely affected by multiple
stressors (several chemicals alone or combined with physical or biological stressors).
Therefore, the best documented links between Pb and effects on the environment are with
effects on individual organisms.

• Natural systems frequently contain multiple metals, making it difficult to attribute observed
adverse effects to single metals. For example, macroinvertebrate communities have been
widely studied with respect to metals contamination and community composition and species
richness. In these studies, multiple metals were evaluated and correlations between observed
community level effects were ascertained. The results often indicate a correlation between
the presence of one or more metals (or total metals) and the negative effects observed.
While, correlation may imply a relationship between two variables, it does not imply
causation of effects.

• In simulated microcosms or natural systems, environmental exposure to Pb in water and
sediment has been shown to affect energy flow and nutrient cycling and benthic community
structure.

• In field studies, Pb contamination has been shown to significantly alter the aquatic
environment through bioaccumulation and alterations of community structure and function.

• Exposure to Pb in laboratory studies and simulated ecosystems may alter species competitive
behaviors, predator-prey interactions, and contaminant avoidance behaviors. Alteration of
these interactions may have negative effects on species abundance and community structure.

• In natural aquatic ecosystems, Pb is often found coexisting with other metals and other
stressors. Thus, understanding the effects of Pb in natural systems is challenging given that
observed effects may be due to cumulative toxicity from multiple stressors.
Aquatic Species Response/Mode of Action

• Recent research has suggested that due to the low solubility of Pb in water, dietary Pb
(i.e., Pb adsorbed to sediment, paniculate matter, and food) may contribute substantially to
exposure and toxicity in aquatic biota.

• Generally speaking, aquatic organisms exhibit three Pb accumulation strategies:
(1) accumulation of significant Pb concentrations with a low rate of loss, (2) excretion of
Pb roughly in balance with availability of metal in the environment, and (3) weak net
accumulation due to very low metal uptake rate and no significant excretion.
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• Protists and plants produce intracellular polypeptides that form complexes with Pb.
Macrophytes and wetland plants that thrive in Pb-contaminated regions have developed
translocation strategies for tolerance and detoxification.

• Like aquatic plants and protists, aquatic animals detoxify Pb by preventing it from being
metabolically available, though their mechanisms for doing so vary. Invertebrates use
lysosomal-vacuolar systems to sequester and process Pb within glandular cells. They also
accumulate Pb as deposits on and within skeletal tissue, and some can efficiently excrete Pb.
Fish scales and mucous chelate Pb in the water column, and potentially reduce visceral
exposure.

• Numerous studies have reported the effects of Pb exposure on blood chemistry in aquatic
biota. Plasma cholesterol, blood serum protein, albumin, and globulin concentrations were
identified as bioindicators of Pb stress in fish.

• Nutrients affect Pb toxicity in aquatic organisms. Some nutrients seem capable of reducing
toxicity. Exposure to Pb has not been shown to reduce nutrient uptake ability, though it has
been demonstrated that Pb exposure may lead to increased production and loss of organic
material (e.g., mucus and other complex organic ligands).

• Avoidance responses are actions performed to evade a perceived threat. Some aquatic
organisms have been shown to be quite adept at avoiding Pb in aquatic systems, while others
seem incapable of detecting its presence.

• The two most commonly reported Pb-element interactions are between Pb and calcium and
between Pb and zinc. Both calcium and zinc are essential elements in organisms and the
interaction of Pb with these ions can lead to adverse effects both by increased Pb uptake
and by a decrease in Ca and Zn required for normal metabolic functions.
Exposure/Response of Aquatic Species

• The 1986 Lead AQCD reviewed data in the context of sublethal effects of Pb exposure. The
document focused on describing the types and ranges of Pb exposures in ecosystems likely to
adversely impact domestic animals. As such, the 1986 AQCD did not provide a
comprehensive analysis of the effects of Pb to most aquatic primary producers, consumers,
and decomposers.

• Waterborne Pb is highly toxic to aquatic organisms, with toxicity varying with the species
and life stage tested, duration of exposure, form of Pb tested, and water quality
characteristics.

• Among the species tested, aquatic invertebrates, such as amphipods and water fleas, were the
most sensitive to the effects of Pb, with adverse effects being reported at concentrations as
low as 0.45 |ig/L (range: 0.45 to 8000 |ig/L).

• Freshwater fish demonstrated adverse effects at concentrations ranging from 10 to
>5400 |ig/L, depending generally upon water quality parameters.
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Amphibians tend to be relatively Pb tolerant; however, they may exhibit decreased enzyme
activity (e.g., ALAD reduction) and changes in behavior (e.g., hypoxia response behavior).
E.6.3 Critical Loads for Lead in Terrestrial and Aquatic Ecosystems

• Critical loads are defined as threshold deposition rates of air pollutants that current
knowledge indicates will not cause long-term adverse effects to ecosystem structure and
function. A critical load is related to an ecosystem's sensitivity to anthropogenic inputs of a
specific chemical.

• The critical loads approach for sensitive ecosystems from acidification has been in use
throughout Europe for about 20 years. Its application to Pb and other heavy metals in Europe
is more recent. European critical load values for Pb have been developed but are highly
specific to the bedrock geology, soil types, vegetation, and historical deposition trends in
each European country. To date, the critical loads framework has not been used for
regulatory purposes in the United States for any chemical. Considerable research is
necessary before critical load estimates can be formulated for ecosystems extant in the
United States.

• Speciation strongly influences the toxicity of Pb in soil and water and partitioning between
dissolved and solid phases determines the concentration of Pb in soil drainage water, but it
has not been taken into account in most of the critical load calculations for Pb performed to
date.

• Runoff of Pb from soil may be the major source of Pb into aquatic systems. However, little
attempt has been made to include this source into critical load calculations for aquatic
systems due to the complexity of including this source in the critical load models.

In summary, due to the deposition of Pb from past practices (e.g., leaded gasoline, ore

smelting) and the long residence time of Pb in many aquatic and terrestrial ecosystems, a legacy

of environmental Pb burden exists, over which is superimposed much lower contemporary Pb

loadings. The potential for ecological effects of the combined legacy and contemporary Pb

burden to occur is a function of the bioavailability or bioaccessibility of the Pb, which, in turn, is

highly dependent upon numerous site factors (e.g., soil organic carbon content, pH, water

hardness). Moreover, while the more localized ecosystem impacts observed around smelters are

often striking, these perturbations cannot be attributed solely to Pb. Many other stressors (e.g.,

other heavy metals, oxides of sulfur and nitrogen) can also act singly or in concert with Pb to

cause such notable environmental impacts.
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1. INTRODUCTION
The present document critically assesses the latest scientific information concerning
health and welfare effects associated with the presence of various concentrations of lead (Pb)
in ambient air, as pertinent to providing updated scientific bases for EPA's current periodic
review of the National Ambient Air Quality Standards for Lead (Pb NAAQS). As such, this
document builds upon previous assessments published by the U.S. Environmental Protection
Agency (EPA), including: (a) the document, Air Quality Criteria for Lead(U.S. Environmental
Protection Agency, 1977); (b) an updated revision of that Lead Air Quality Criteria Document
(Lead AQCD) and an accompanying Addendum published in 1986 (U.S. Environmental
Protection Agency, 1986a, b); as well as (c) an associated 1990 Supplement (U.S. Environmental
Protection Agency, 1990). This document focuses on evaluation and integration of information
relevant to Pb NAAQS criteria development that has become available mainly since that covered
by the 1986 and 1990 criteria assessments.
This introductory chapter (Chapter 1) of the revised Lead AQCD presents:
(a) background information on pertinent Clean Air Act legislative requirements, the criteria
and NAAQS review process, and the history of previous Lead criteria reviews; (b) an overview
of the current Lead criteria review process, associated key milestones, and schedule; and
(c) an orientation to the general organizational structure and content of this revised Lead AQCD.
1.1 LEGAL AND HISTORICAL BACKGROUND
1.1.1 Legislative Requirements
Two sections of the Clean Air Act (CAA) govern the establishment, review, and revision
of NAAQS. Section 108 (42 U.S.C. 7408) directs the Administrator of the U.S. Environmental
Protection Agency (EPA) to identify ambient air pollutants that may be reasonably anticipated to
endanger public health or welfare and to issue air quality criteria for them (U.S. Code, 2003a).
These air quality criteria are to reflect the latest scientific information useful in indicating the
kind and extent of all identifiable effects on public health or welfare that may be expected from
the presence of a given pollutant in ambient air.
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Section 109(a) of the CAA (42 U.S.C. 7409) directs the Administrator of EPA to propose
and promulgate primary and secondary NAAQS for pollutants identified under Section 108 (U.S.
Code, 2003b). Section 109(b)(l) defines a primary standard as one that, in the judgment of the
EPA Administrator, is requisite to protect the public health (see inset below) based on the criteria
and allowing for an adequate margin of safety. The secondary standard, as defined in Section
109(b)(2), must specify a level of air quality that, in the judgment of the Administrator, is
requisite to protect the public welfare (see inset below) from any known or anticipated adverse
effects associated with the presence of the pollutant in ambient air, based on the criteria.
EXAMPLES OF PUBLIC
HEALTH EFFECTS
Effects on the health of the general
population, or identifiable groups within the
population, who are exposed to pollutants in
ambient air
Effects on mortality
Effects on morbidity
Effects on other health conditions including
indicators of:
• pre-morbid processes,
• risk factors, and
• disease
EXAMPLES OF PUBLIC
WELFARE EFFECTS
Effects on personal comfort and well-being
Effects on economic values
Deterioration of property
Hazards to transportation
Effects on the environment, including:
animals
climate
crops
materials
soils
vegetation
visibility
water
weather
wildlife
Section 109(d) of the CAA (42 U.S.C. 7409) requires periodic review and, as appropriate,
revision of existing criteria and standards (U.S. Code, 2003b). If, in the EPA Administrator's
judgment, the Agency's review and revision of criteria make appropriate the proposal of new or
revised standards, such standards are to be revised and promulgated in accordance with Section
109(b). Alternatively, the Administrator may find that revision of the standards is inappropriate
and conclude the review by leaving the existing standards unchanged. Section 109(d)(2) of the
1977 CAA Amendments also requires that an independent scientific review committee be
established to advise the EPA Administrator on NAAQS matters, including the scientific
soundness of criteria (scientific bases) supporting NAAQS decisions. This role is fulfilled by the
Clean Air Scientific Advisory Committee (CASAC), which is administratively supported by U.S.
EPA's Science Advisory Board (SAB).
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1.1.2 Criteria and NAAQS Review Process
Periodic reviews by EPA of criteria and NAAQS for a given criteria air pollutant progress
through a number of steps, starting with the preparation of an air quality criteria document
(AQCD) by the National Center for Environmental Assessment Division in Research Triangle
Park, NC (NCEA-RTP), a unit within EPA's Office of Research and Development (ORD). The
AQCD provides a critical assessment of the latest available scientific information upon which
the NAAQS are to be based. Drawing upon the AQCD, the Office of Air Quality Planning and
Standards (OAQPS), a unit within EPA's Office of Air and Radiation (OAR), prepares a Staff
Paper that (a) evaluates policy implications of key studies and scientific information assessed in
the AQCD; (b) presents relevant exposure and risk analyses; and (c) also presents EPA staff
conclusions and recommendations for standard-setting options for the EPA Administrator to
consider. The Staff Paper is intended to help "bridge the gap" between the scientific assessment
contained in the AQCD and the judgments required of the Administrator in determining whether
it is appropriate to retain or to revise the subject NAAQS.
Iterative drafts of both the AQCD and the Staff Paper (as well as other analyses, such as
associated exposure and/or risk assessments supporting the Staff Paper) are made available for
public comment and CAS AC review. Final versions of the AQCD and Staff Paper incorporate
changes in response to CASAC review and public comment. Based on the information in these
documents, the EPA Administrator proposes decisions on whether to retain or revise the subject
NAAQS, taking into account public comments and CASAC advice and recommendations. The
Administrator's proposed decisions are published in the Federal Register, with a preamble that
delineates the rationale for the decisions and solicits public comment. After considering
comments received on the proposed decisions, the Administrator makes a final decision that is
promulgated via & Federal Register notice that addresses significant comments received on the
proposal.
Promulgated NAAQS decisions involve consideration of the four basic elements of a
standard: indicator, averaging time, form, and level. The indicator defines the pollutant to be
measured in the ambient air for the purpose of determining compliance with the standard.
The averaging time defines the time period over which air quality measurements are to be
obtained and averaged, considering evidence of effects associated with various time periods of
exposure. The form of a standard defines the air quality statistic that is to be compared to the
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level of the standard (i.e., an ambient concentration of the indicator pollutant) in determining
whether an area attains the standard. The form of the standard specifies the air quality
measurements that are to be used for compliance purposes (e.g., the 98th percentile of an annual
distribution of daily concentrations; the annual arithmetic average), the monitors from which the
measurements are to be obtained (e.g., one or more population-oriented monitors in an area), and
whether the statistic is to be averaged across multiple years. These basic elements of a standard
are the primary focus of the staff conclusions, and recommendations posed in the Staff Paper are
explicitly specified in the ensuing NAAQS rulemaking, building upon the policy-relevant
scientific information assessed in the AQCD and on policy analyses contained in the Staff Paper.
These four elements, taken together, determine the degree of public health and welfare protection
afforded by the NAAQS.

1.1.3 Regulatory Chronology
In 1971, U.S. EPA promulgated national ambient air standards for several major "criteria"
pollutants (see Federal Register, 1971), but did not include Pb among them at that time. Later,
on October 5, 1978, the EPA promulgated primary and secondary Pb NAAQS under Section 109
of the CAA (43 FR 46258), as announced in the Federal Register (1979). Identical primary and
the secondary Pb standards were established at the time: 1.5 |ig/m3 as a quarterly average
(maximum arithmetic mean averaged over a calendar quarter). Those standards were based on
scientific assessments in EPA's original Air Quality Criteria for Lead (U.S. Environmental
Protection Agency, 1977) or "1977 Lead AQCD."
In 1986, the EPA published a revised Lead AQCD (U.S. Environmental Protection
Agency, 1986a). The 1986 Lead AQCD assessed newly available scientific information on the
health and welfare effects associated with exposure to various concentrations of Pb in ambient
air, based on literature published through 1985. That 1986 document was principally concerned
with the health and welfare effects of Pb, but other scientific data were also discussed in order to
provide a better understanding of the pollutant in the environment. Thus, the 1986 document
included chapters that discussed the atmospheric chemistry and physics of the pollutant;
analytical approaches; environmental concentrations; human exposure and dosimetry;
physiological, lexicological, clinical, and epidemiological aspects of Pb health effects; and Pb
effects on ecosystems. An Addendum to the 1986 Lead AQCD was also published along with it
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(U.S. Environmental Protection Agency, 1986b). Subsequently, a Supplement to the 1986 Lead
AQCD/Addendum was published by EPA in 1990 (U.S. Environmental Protection Agency,
1990a). That 1990 Supplement evaluated still newer information emerging in the published
literature concerning (a) Pb effects on blood pressure and other cardiovascular endpoints and
(b) the effects of Pb exposure during pregnancy or during the early postnatal period on birth
outcomes and/or on the neonatal physical and neuropsychological development of affected
infants and children.
The evaluations contained in the 1986 Lead AQCD/Addendum and 1990 Supplement
provided scientific inputs to support decision-making regarding CAA-mandated periodic review
and, as appropriate, revision of the Pb NAAQS; and they were drawn upon by EPA's Office of
Air Quality Planning and Standards in preparation of an associated OAQPS Lead Staff Paper
(U.S. Environmental Protection Agency, 1990b). Based on the scientific assessment in the 1986
Lead AQCD/Addendum and the 1990 Supplement, as well as associated exposure/risk analyses,
the 1990 Staff Paper recommended that the EPA Administrator consider a range of standards for
the primary Pb NAAQS of 0.5 to 1.5 |ig/m3 (30-day arithmetic mean). After consideration of
evaluations contained in the above documents, EPA chose not to propose revision of the Pb
NAAQS. As part of implementing a broad, integrated Strategy for Reducing Lead Exposures
(U.S. Environmental Protection Agency, 1991), the Agency focused efforts primarily on
regulatory or remedial clean-up actions aimed at reducing Pb exposures from a variety of non-air
sources judged to pose more extensive public health risks to U.S. populations as well as on other
actions to reduce Pb emissions to air.
Changes in relative contributions of various Pb sources and exposure pathways to human
exposures in the United States, and EPA actions to reduce such exposures, thusly provide very
important background context for this current Pb criteria and NAAQS review. Since 1978, the
amount of Pb emitted into the air nationally has markedly declined. For example, as illustrated
in Chapters 2 and 3 of this document, from 1982 to 2002 Lead emissions into the air decreased
by 93%, and the average air quality concentration of Pb decreased by 94% from 1983 to 2002
(http://www.epa.gov/airtrends/lead2.html). Total Pb emissions into the air decreased from about
220,000 tons in 1970 to less than 4,000 in 1999. This decline is mainly attributable to EPA's
regulatory actions that led to notable reductions in the content of Pb in gasoline (see, for
example, 50 FR 9386), which substantially altered basic patterns of air Pb emissions in the
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United States (http://www.epa.gov/airtrends/lead2.html). Emissions from stationary sources
have also been greatly reduced (http://www.epa.gov/airtrends/lead2.html, Figure 2-11); but,
given the even greater reductions in emissions from transportation sources, industrial processes
(including smelters and battery manufacturers) now constitute a larger percentage of remaining
Pb emissions to the atmosphere (http://www.epa.gov/airtrends/lead2.html, Figure 2-12).
In short, Pb emissions into the atmosphere decreased greatly in the 1980's and 1990's, a trend
that has continued through to the present. Consequently, airborne Pb now represents only a
relatively small component of total exposure to Pb in the United States, such that the principal
sources and pathways for U.S. Pb exposure among the classically-defined most sensitive
population group (young children) involve non-inhalation pathways, e.g., ingestion of Pb from
deteriorating paint, dust, historically contaminated soil, drinking water, and food. While these
downward trends in air Pb exposures nationwide are encouraging, several important sources of
air Pb exposure may still persist in some localities. Lead emissions from specific stationary
sources and/or reentrainment of Pb-contaminated soils (including from past deposition of
airborne Pb) may still have significant impacts on a local level. Recognition of the multimedia
nature of Pb exposure of the general population has been important historically and sorting out
relative contributions to total Pb exposure burdens represents an important input to the current
periodic Pb NAAQS review effort.
Since the 1980's, EPA has played a major, effective role in working to reduce the main
sources of Pb exposure for most children, including deteriorating Pb-based paint, Pb-
contaminated dust, and Pb-contaminated residential soil (http://www.epa.gov/lead/). For
example, EPA has established standards for Pb-based paint hazards and Pb dust cleanup levels in
most pre-1978 housing and child-occupied facilities, and is now developing standards for those
conducting renovation activities that create Pb-based paint hazards and for the management and
disposal of Pb-based debris (http://www.epa.gov/lead/regulation.htm). In addition, EPA has
developed standards for management of Pb in solid and hazardous waste, continues to oversee
the cleanup of Pb contamination at Superfund facilities, and issued regulations to reduce Pb in
drinking water (http://www.epa.gov/lead/sources.htm). Beyond taking specific regulatory
actions, the Agency's Lead Awareness Program also continues to work to protect human health
and the environment against the dangers of Pb by conducting research and designing educational
outreach efforts and materials (http://www.epa.gov/lead/).
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Since the 1980's, EPA has also promulgated regulations under Section 112 of the Clean
Air Act (42 U.S.C. § 7412), to address emissions of Pb components and other toxic pollutants
from both primary Pb smelters and secondary Pb smelters (40 CFR Subparts X and TTT). Under
section 112(d), these emission standards are to require "the maximum degree of reduction in
emissions" that are "achievable." Thus, EPA promulgated section 112(d) standards for
secondary Pb smelters on June 23, 1995 (60 Fed. Reg. 3587) and revised them on June 13, 1997
(62 Fed. Reg. 32209), followed by promulgation of section 112(d) standards for primary Pb
smelters on June 4, 1999 (64 Fed. Reg. 30194).
1.2 CURRENT LEAD CRITERIA AND NAAQS REVIEW
1.2.1 Procedures and Key Milestones for Document Preparation
It is important to emphasize at the outset that development of the present revised Lead
AQCD included substantial external (non-EPA) expert inputs and opportunities for public input
through (a) public workshops involving the general scientific community, (b) several iterative
reviews of successive drafts of this document by CAS AC, and (c) comments from the public on
successive drafts. Such extensive external inputs received via these mechanisms help to ensure
that the review of the Pb NAAQS will be based on critical assessment in this document of the
latest available pertinent science.
The approach used for developing this revised Lead AQCD has built on experience
derived from other recent criteria document preparation efforts. This includes close coordination
between NCEA-RTP and OAQPS staff, as well as with others, throughout the document
preparation and review process. Briefly, the respective responsibilities for production of the
document and meeting key milestones were as follows. An NCEA-RTP Lead Team was
designated as being responsible for creation and implementation of a project plan for developing
the Lead AQCD, taking into account input from individuals in other ORD units, OAQPS, and
other EPA program/policy offices identified as part of the EPA Lead Work Group. The Lead
Team defined critical issues and topics to be addressed by the authors and provided direction in
order to focus on evaluation of those studies most clearly identified as likely being important for
U.S. air standard setting purposes. Criteria document materials were authored in part by
NCEA-RTP Lead Team staff with appropriate expertise in particular areas and by non-EPA
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consultants to EPA who are recognized experts in pertinent specific areas (e.g., Pb biokinetic
modeling, toxicology, and epidemiology).
Key milestones for development of this Lead AQCD are listed in Table 1-1. As a first
step, EPA announced on November 9, 2004 the official initiation of the current periodic review
of Lead air quality criteria.
Table 1-1. Key Milestones and Projected Schedule for Development of Revised Lead Air
Quality Criteria Document (Lead AQCD)
Major Milestones
Target Dates
1. Literature Search
2. Federal Register Call for Information
3. Prepare Draft Lead AQCD Project Work Plan
4. Release Draft Proj ect Plan for Public Comment/CAS AC Review
5. Public Comment Period
6. CAS AC Teleconsultation on Proj ect Work Plan
7. Workshop Drafts of Lead AQCD Chapters
8. Peer Consultative-Review Workshop(s)
9. Release First External Review Draft
10. Public Comment Period
11. CAS AC/SAB Public Review Meeting (First Ext. Rev. Draft)
12. Release Second External Review Draft
13. Public Comment Period
14. CAS AC/SAB Public Review Meeting (Second Ext. Rev. Draft)
15. Release Revised Drafts of Executive Summary and Integrative Synthesis
Chapter for Public Comment and CASAC Review
16. CASAC Review of Executive Summary and Integrative Synthesis
via Public Teleconference
17. Final Lead AQCD
Ongoing
November 9, 2004
Nov-Dec 2004
January 2005
Jan/Feb 2005
March 28, 2005
May/June 2005
July/August 2005
December 1, 2005
Dec 2005-Feb 2006
Feb. 28-Mar 1, 2006
May 2006
May/June 2006
June 28-29,2006

July 31,2006

August 15, 2006
October 1, 20061
1 Court-ordered deadline for EPA to produce a final Lead AQCD in relation to Missouri Coalition for the
Environment v. EPA. Civil Action No. 4:04-CV-00660 (ERW) (E.D. Mo. Sept. 14, 2005). Also, note that
materials contributed by non-EPA authors, at times, were modified by EPA staff in response to internal and/or
external review comments and that EPA is responsible for the ultimate content of this Lead AQCD.
-------
More specifically, under processes established in Sections 108 and 109 of the Clean Air
Act, U.S. EPA began by announcing in the Federal Register (69 FR 64,926) the formal
commencement of the current review process with a call for information (see Federal Register,
2004). In addition, EPA prepared a January 2005 draft Lead AQCD Work Plan that was made
available for public comment and was the subject of teleconsultation with CAS AC on March 28,
2005 as a means by which to communicate the process and timeline for development of a revised
Lead AQCD. Next, expert consultants to NCEA-RTP and NCEA-RTP staff (a) carefully
evaluated pertinent new studies obtained via the call for information and via ongoing literature
searches conducted by NCEA-RTP information retrieval specialists and (b) prepared preliminary
draft chapter materials for inclusion in this revised Lead AQCD. Those preliminary draft
materials then underwent expert peer discussion at public workshops organized and conducted
by NCEA-RTP in July/August, 2005. After consideration of comments received at the
workshops, appropriate revisions were made in the draft materials and incorporated into the First
External Review Draft of the Lead AQCD, which was made available for public comment (for
90 days) and CASAC review at a public meeting on February 28-March 1, 2006. EPA, taking
into account CASAC and public comments, then incorporated revisions into the draft AQCD
before releasing a Second External Review Draft (May 2006) of it for further review by the
public and by CASAC at a public meeting held June 28-29, 2006. Revised drafts of the
Executive Summary and the Integrative Synthesis were next quickly prepared and made
available on August 1, 2006 for public comment and CASAC teleconference on August 15,
2006. Further revisions were then incorporated, in response to the last two public comment and
CASAC reviews, to complete the final version of this revised Lead AQCD for issuance by
October 1, 2006. The final document (dated October 2006) was then published, and its
availability to the public announced in the Federal Register.
Drawing upon evaluations in this Lead AQCD and other Pb exposure/risk analyses, an
associated Lead Staff Paper prepared by EPA's Office of Air Quality Planning and Standards
(OAQPS) assesses policy implications of key information in the Lead AQCD, reports pertinent
exposure and risk analyses, and poses possible options for the EPA Administrator to consider
regarding whether to retain or, if appropriate, revise the Pb NAAQS. Taking into account
CASAC and public comments, the EPA Administrator will consider the options posed in the
Staff Paper; propose decisions regarding possible retention or revision of the primary and/or
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secondary Pb NAAQS via the Federal Register for public comment; and then, after consideration
of comments received, promulgate final Pb NAAQS review decisions.
1.3 ORGANIZATIONAL STRUCTURE AND CONTENT OF
THE DOCUMENT
1.3.1 Ascertainment of Literature and General Document Format
Lists of references published since completion of the 1986 Lead AQCD/Addendum and
1990 Supplement were made available to the authors. The references were mainly selected from
information data base (e.g., Pub Med) searches conducted by EPA. However, additional
references have been added as work has proceeded in creating the present Lead AQCD materials.
As an aid in selecting pertinent new literature, the authors were also provided with a summary of
issues to be addressed in this revised Lead AQCD. Many such issues identified in the course of
previous Lead criteria assessments, through interactions between EPA Lead Team and Lead
Work Group members, and via workshop discussions.
The general format used in this document is to open each new chapter (or main section)
for the updated Lead AQCD with concise summary of key findings/conclusions from previous
Lead criteria assessments, especially the 1986 Lead AQCD/Addendum (U.S. Environmental
Protection Agency, 1986a,b) and 1990 Supplement (U.S. Environmental Protection Agency,
1990). After presentation of such background information, the remainder of each chapter or
section typically provides an updated discussion of newer literature and resulting key
conclusions. In some cases where no new information is available, the summary of key findings
and conclusions from the previous Pb criteria assessment(s) must suffice as the basis for current
key conclusions. Increased emphasis is placed in the main chapters of this revised Lead AQCD
on interpretative evaluation and integration of evidence pertaining to a given topic than was
typical of many previous EPA air quality criteria documents, with more detailed descriptions of
individual studies or other supportive information being provided in a series of accompanying
annexes.
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1.3.2 Organization and Content of the Document
This updated Lead AQCD critically assesses scientific information on health and welfare
effects associated with exposure to the concentrations of Pb in ambient air. The document is not
intended to be a detailed, exhaustive review of the literature. Rather, the cited references reflect
the current state of knowledge regarding important issues pertinent to decisions regarding
possible revision by EPA of the Pb NAAQS. Although emphasis is placed mainly on the
discussion of health and welfare effects data, other scientific information also is evaluated in
order to provide a better understanding of the nature, sources, distribution, and concentrations of
Pb in ambient air, as well as the measurement of human exposure to Pb.
The focus of discussion is on assessment of selected pertinent scientific information
newly published since the last prior assessments of air quality criteria for Pb contained in the
1986 Lead AQCD/Addendum and 1990 Supplement. As noted earlier, key findings and
conclusions from the 1986 Lead AQCD/Addendum and 1990 Supplement are typically first
briefly summarized at the outset of discussion of a given topic, with appropriate reference back
to the previous criteria assessment materials. Typically, important prior studies are more
specifically discussed only if they are open to reinterpretation in light of newer data and/or are
judged to be potentially useful in decisions on revision of the Pb NAAQS. Generally, only
information that has undergone scientific peer review and has been published (or accepted for
publication) in the open literature through December 31, 2005 has been considered in this
revised Lead AQCD. Certain other unpublished analyses (e.g., EPA analyses of recently
available U.S. Lead air quality data) are also considered, depending on the importance of the
subject information and its pertinence to criteria development for Pb NAAQS, as determined in
consultation with CASAC.
This Lead AQCD consists of two volumes. The first volume (Volume I) includes eight
chapters that comprise the main body of the revised Lead AQCD and an Executive Summary for
all chapters. In Volume I, this introductory chapter (Chapter 1): (a) provides brief statements
regarding the purpose of the document; (b) presents information on the legislative background
and regulatory chronology of Pb criteria reviews; and (c) presents an overview of the
organization of the document. Chapter 2 discusses the physics and chemistry of Pb, as well as
sources, emissions, transport and deposition/fate. Chapter 3 next provides information on
environmental concentrations, dispersal patterns, and multimedia exposure pathways. Chapter 4
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focuses on the measurement of concentrations of Pb in biological samples and the modeling of
multimedia exposure impacts on human internal Pb burdens, especially as indexed by blood Pb
or bone Pb concentrations. Then, Chapter 5 discusses toxicologic studies of Pb health effects in
humans, laboratory animals, and in vitro test systems; whereas Chapter 6 assesses Pb-related
epidemiologic (observational) studies of human population groups. Chapter 7 deals with
ecological and other environmental effects of Pb. Lastly, Chapter 8 provides an overall
integrative synthesis of key information drawn from the earlier chapters to delineate: human Pb
exposure pathways and trends; health effect findings and conclusions of most importance for
derivation of primary Pb NAAQS; and key types of welfare effects (in this case, ecosystem)
findings pertinent to the derivation of secondary Pb NAAQS. Volume II of this revised Lead
AQCD includes several annexes containing detailed descriptive materials supporting the more
interpretative evaluations presented in several of the main chapters dealing with Pb-related health
and/or ecological effects.
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REFERENCES

Federal Register. (1971) National primary and secondary ambient air quality standards. F. R. (April 30)
36: 8186-8201.
Federal Register. (1979) National primary and secondary ambient air quality standards: revisions to the National
Ambient Air Quality Standards for lead. F. R. (February 8) 44: 8202-8237.
Federal Register. (2004) Air quality Criteria Document for Lead: Call for Information. F. R. (November 9)
69: 64926-64928.
U.S. Code. (2003a) Clean Air Act, §108, air quality criteria and control techniques.. U. S. C. 42: §7408.
U.S. Code. (2003b) Clean Air Act, §109, national ambient air quality standards. U. S. C. 42: §7409.
U.S. Environmental Protection Agency. (1977) Air quality criteria for lead. Research Triangle Park, NC: Health
Effects Research Laboratory, Criteria and Special Studies Office; EPA report no. EPA/600/8-77-017.
Available fromNTIS, Springfield, VA; PB-280411.
U.S. Environmental Protection Agency. (1986a) Air quality criteria for lead. Research Triangle Park, NC: Office of
Health and Environmental Assessment, Environmental Criteria and Assessment Office; EPA report no.
EPA/600/8-83/028aF-dF. 4v. Available from: NTIS, Springfield, VA; PB87-142378.
U.S. Environmental Protection Agency. (1986b) Lead effects on cardiovascular function, early development, and
stature: an addendum to U.S. EPA Air quality criteria for lead. In: Air quality criteria for lead, v. 1. Research
Triangle Park, NC: Office of Health and Environmental Assessment, Environmental Criteria and Assessment
Office; pp. A1-A67; EPA report no. EPA/600/8-83/028aF. Available from: NTIS, Springfield, VA;
PB87-142378.
U.S. Environmental Protection Agency. (1990a) Summary of selected new information on effects of lead on health
and supplement to 1986 air quality criteria for lead. Research Triangle Park, NC: Office of Health and
Environmental Assessment, Environmental Criteria and Assessment Office; EPA report no. EPA/600/8-89.
Available fromNTIS, Springfield, VA; PB92-235670.
U.S. Environmental Protection Agency. (1990b) Review of the national ambient air quality standards for lead:
assessment of scientific and technical information: OAQPS staff paper. Research Triangle Park, NC:
Office of Air Quality Planning and Standards; report no. EPA-450/2-89/022. Available from: NTIS,
Springfield, VA; PB91-206185.
U.S. Environmental Protection Agency (1991) U.S. EPA Strategy for Reducing Lead Exposure. Available from
U.S. EPA Headquarters Library/Washington, D.C. (Library Code EJBD; Item Call Number: EAP
100/1991.6; OCLC Number 2346675).
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2. CHEMISTRY, SOURCES, AND TRANSPORT
OF LEAD
The purpose of this chapter is to provide background information on the chemical
properties of lead (Pb) that are relevant to its transport within the environment, its transport into
ecosystems and its impact on human health; to discuss the known sources of Pb in the
environment; and to outline the mechanisms by which Pb is transported within the environment.
The chapter does not provide a comprehensive list of all sources of Pb, nor does it provide
emission rates or emission factors for all source categories, since such information is available
for only a limited number of sources. Rather, the chapter provides data on the chemistry,
sources, and transport of Pb where information is available in the literature or via publicly
accessible EPA databases, websites, and reports. Particle size distribution data for Pb are even
scarcer than information on total Pb emissions from sources; but particle size data are presented
where such data are available.
2.1 PHYSICAL AND CHEMICAL PROPERTIES OF LEAD
Properties of Elemental Lead
Elemental Pb possesses an array of useful physical and chemical properties, making it
among the first metals to be extracted and used by humankind. It has a relatively low melting
point (327.51), is a soft, malleable, and ductile metal, a poor electrical conductor, and is easily
cast, rolled and extruded. Although sensitive to environmental acids, after exposure to
environmental sulfuric acid (H^SO/t), metallic Pb becomes impervious to corrosion due to
weathering and submersion in water. This effect is due to the fact that Pb lead sulfate (PbSO/t),
the relatively insoluble precipitate produced by reaction of Pb with H2SO4, forms a protective
barrier against further chemical reactions (Schweitzer, 2003). This aspect of its chemistry made
Pb especially convenient for protective surface coatings (e.g. paint), roofing, containment of
corrosive liquids, and (until the discovery of its adverse health effects), construction of water
supply systems.
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Lead is readily extracted from galena, a widely available lead sulfide mineral form (PbS),
by froth flotation, followed by roasting in the presence of a limited amount of oxygen to form
litharge, one of two forms of lead oxide (PbO). Elemental Pb is then isolated by reducing PbO
by way of heating in the presence of elemental carbon (coke, charcoal) (Greenwood and
Earnshaw, 1984). This and other extraction and recovery processes are discussed in greater
detail later in this chapter.
Lead alloys constitute 60% of Pb used in industry (Prengaman, 2003). The major alloying
elements are antimony, calcium, tin, copper, tellurium, arsenic, and silver. Selenium, sulfur,
bismuth, cadmium, indium, aluminum, and strontium are also sometimes used. Lead alloys are
found primarily in Pb acid batteries, solder, ammunition, and cable sheathing (Prengaman,
2003). Table 2-1 provides a list of Pb alloys in wide use by industry.
Some of the physical properties of elemental Pb are listed in Table 2-2. The most
important of these properties, when evaluating the transport routes for Pb within the atmosphere,
is its boiling point. As indicated, Pb will only exist in the vapor phase at or above 1750°C.
Therefore, at ambient atmospheric temperatures, elemental Pb will deposit to surfaces or exist in
the atmosphere as a component of atmospheric aerosol.

Oxidation States of Lead
Lead is the heaviest congener of carbon and shares many properties with the other
elements found in the same column of the periodic chart (silicon, germanium, and tin).
As Group IV elements, these elements have four valence electrons (2p and 2 s), allowing for
both divalent and tetravalent compounds.
Due to its high atomic number (82), the valence electron orbitals of the Pb atom exist at a
comparatively large distance from its nucleus. As with s and/? orbitals at any quantum level,
electrons in the 6s orbital tend to occupy space near the nucleus with greater probability than
those in the 6p orbital. The strong attraction produced by the large Pb nucleus, combined with
the long distance that the 6s electrons must travel, result in electron accelerations to relativistic
speeds. The Theory of Relativity states that as the velocity of matter approaches the speed of
light, its apparent mass increases. In this instance, the electrons in the Pb 6s orbital experience
an increase in weight, which increases the attractive effect of the positive nuclear charge, which
contracts the diameter of the Pb 6s orbital (Pitzer, 1979). This "relativistic effect" on valence
2-2
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Table 2-1. Lead Alloys and Their Industrial Applications
Pb Alloy
Uses
Pb-Antimony

Pb-Calcium

Pb-Tin

Pb-Copper

Pb-Silver

Pb-Tellurium
Pb-Bismuth

Pb-Cadmium

Pb-Indium

Pb-Strontium
Pb-Lithium
Pb-Antimony-Tin
Pb-Calcium-Aluminum
Pb-Calcium-Tin
Pb-Calcium-Silver
Pb-Antimony-Silver

Pb-Silver-Tin

Pb - Strontium-Tin
Pb -Lithium-Tin
Grids, posts, and connectors for Pb-acid batteries, ammunition, cable sheathing,
anodes, tank linings, pumps, valves, and heating and cooling coils
Automotive, standby power, submarines, and specialty sealed batteries,
electrowinning anodes, cable sheathing, sleeving, specialty boat keels, and Pb
alloy tapes
Soldering for electronics, general purposes, automobile radiators, and heat
exchangers, corrosion resistant coatings on steel and copper, cable sheathing,
fuses, sprinkler system alloys, foundry pattern alloys, molds, dies, punches,
cores, mandrels, replication of human body parts, and filters for tube bonding
Pb sheet, pipe, cable sheathing, wire, fabricated products, tank linings, tubes for
acid-mist precipitators, steam heating pipes for sulfuric acid or chromate plating
baths, and Pb sheathing for roofs
Anodes, high-temperature solders, insoluble anodes in the electrowinning of
zinc and manganese, and soft solders
Pipes, sheets, shielding for nuclear reactors, and cable sheathing
Fuses, sprinkler system alloys, foundry pattern alloys, molds, dies, punches,
cores, mandrels, solders, replication of human body parts, and filters for tube
bonding
Fuses, sprinkler system alloys, foundry pattern alloys, molds, dies, punches,
cores, mandrels, solders, replication of human body parts, and filters for
tube bonding
Fuses, sprinkler system alloys, foundry pattern alloys, molds, dies, punches,
cores, mandrels, solders, replication of human body parts, filters for tube
bonding, and joining metals to glass
Battery grids
Bearings, Pb-acid battery grids
Printing, bearings, solders, slush castings, and specialty castings
Negative battery grids of Pb-acid batteries
Positive grids of Pb-calcium batteries, and Pb anodes for electrowinning
Zinc electrowinning
Anodes used for the production of thin copper foil in electronics, and anodes in
cathodic protection of steel pipes and structures in water
Anodes in cathodic protection of steel pipes and structures in water, and soft
solders
Anodes for copper electrowinning
Pb-acid battery grids
Source: Prengaman (2003).
2-3
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Table 2-2. Physical Properties of Elemental Lead
Physical Property
Atomic number 82
Atomic weight 207.2
Valence electrons [Xe]4f145d106s26p2
Melting point 328 °C
Boiling point 1750 °C
Density 11.34g/cm3
Atomic radius 146 pm
Standard reduction potential - 0.126V
Oxidation numbers +2, +4
lonization Energy 715.6 kJ/mol

Source: Kotz and Purcell (1991).
electrons are proportional to the square of atomic number, and manifests within the Group IV
elements as a distinctly increasing trend in the stability of the divalent state from Si down to Pb.
In the case of Pb, the two 6s electrons behave as if they were chemically inert, leaving only the
two 6p electrons available for bonding or oxidation under ordinary conditions. For this reason,
the relativistic effect is also known as the "inert pair effect." Consequently, Pb(II) is the most
common oxidation state in which Pb is found in the environment (King, 1995; Claudio et al.,
2003).
Lead is distinguished from other elements that are subject to relativistic effects by its
preference for forming tetravalent (Pb(IV) organometallic compounds. However, in fact, it is
only with rare exception that Pb(II) organometallic compounds form (Pelletier, 1995;
Greenwood and Earnshaw, 1984). All simple alkyllead compounds, such as the well-known fuel
additives, tetramethyllead (TML) and tetraethyllead (TEL) are composed of Pb(IV). In contrast,
inorganic Pb(IV) compounds (such as PbO2) are strong oxidants and unstable with respect to
their Pb(II) analogs. There are, overall, more than 200 known organolead compounds
(Harrison, 1985).
In relation to the other Group IV metals, however, Pb forms the least stable and most
reactive organometallic derivatives. This is largely due to the weak bond between Pb and

2-4
-------
carbon, consistent with its large atomic size, and the influence of the relativistic effect on its
valence orbitals. Specifically, the mean bond dissociation energies of the metal-carbon bonds for
Group IV elements are 56.7 kcal/mol for germanium, 46.2 kcal/mol for tin, and 30.8 kcal/mol for
Pb (Shapiro & Frey, 1968). Organolead compounds are thermally unstable and will decompose
to metallic Pb and free radicals at relatively low temperatures (Willemsens and van der Kerk,
1965). For example, TML decomposes at temperatures above 200°C, and TEL decomposes at
temperatures above 1 10°C (King, 1995). In solution, organolead compounds decompose in the
presence of UV radiation (1 hr/254 nm) and sunlight (Gomez Ariza et al., 2000).
Tetralkyllead compounds have atmospheric residence times ranging from a few hours to a
few days (Pelletier, 1995). TML and TEL react with OH in the gas-phase, following pseudo-first
order kinetics, to form a variety of products that include ionic trialkyllead (TriAL), dialkyllead
(DiAL) and metallic Pb. Trialkyllead is slow to react with OH and is quite persistent in the
atmosphere (Hewitt and Harrison, 1986; Harrison and Laxen, 1980).

Lead Oxides, Chalcogenides, and Salts
A rich variety of inorganic Pb compounds and complex salts can be prepared in the
laboratory under conditions of temperature and pressure not usually seen in the environment.
Information on the many possible organic and inorganic Pb compounds can be found in the text
by Greenwood and Earnshaw (1984). Several representative Pb salts and oxides are described in
Tables 2-3 and 2-4. Inorganic Pb compounds that can be found in the environment are the main
focus of this discussion.
As explained earlier, Pb exists preferentially in its +2 oxidation state in the environment.
Under aqueous acidic conditions, Pb readily oxidizes, with a strongly positive electrochemical
potential (E° = 1.355 V) and a large equilibrium constant (K = 10 91 6), to form Pb(II) (Singley,
1994):
2Pb + O2 + 4H+ -» 2Pb2+ + 2H2O (2-1)
Table 2-5 lists the various Pb compounds and salts that are present naturally or are
introduced into the environment by anthropogenic activities. From this list, it is clear that only a
relatively limited number of salts and covalently-bound Pb compounds are of significance in
2-5
-------
Table 2-3. Lead Salts: Names, Formulae, Physical Characteristics, and Uses
Category Compound Name
Formula
Form
Uses
Pb Acetates Anhydrous Pb
Acetate
Basic Pb Acetate
Pb Acetate
Trihydrate

Pb Tetraacetate
Pb Pb Carbonate
Carbonates
Pb(C2H302)2

Pb(C2H3O2)2-
2Pb(OH)2
Pb(C2H302)2-
3H20

Pb(C2H3O2)4
PbC03
Basic Pb Carbonate (Pb(CO3)2)2
Pb(OH)2
Pb Halides Pb Fluoride
Pb Chloride
PbF,
PbCl2
Pb Bromide
Pb Iodide
PbBr2
Pbl,
Pb Silicates Pb Monosilicate 3PbO
2SiO2

Pb Bisilicate PbO
0.03A12O3-
1.95Si02
Tribasic Pb Silicate 3PbO-SiO2

Pb Sulfates Tribasic Pb Sulfate 3PbO
PbSO4
H20
White, crystalline
solid
Heavy, white
powder
White,
monoclinic
crystalline solid
Colorless,
monoclinic
crystalline solid
Colorless,
orthorhombic
crystals
White, hexagonal
crystals
Colorless,
orthorhombic
crystals
White,
orthorhombic
needles
White,
orthorhombic
crystals
Powdery, yellow,
hexagonal
crystals
White, trigonal
crystalline
powder
Pale yellow
powder

Reddish-yellow
powder
Fine, white
powder
Preparing other Pb salts.

Sugar analysis.

Making other Pb compounds, mordant for cotton
dyes, water repellant, processing agent for
cosmetics, perfumes, and toiletries.
Oxidizing agent in organic synthesis, cleaving of
a-hydroxy acids, introducing acetyl groups in
organic molecules.
Catalytic polymerization of formaldehyde,
improving the bonding of polychloroprene to
metals in wire-reinforced hoses, a component of
high-pressure lubricating greases, and a lubricant
for polyvinyl chloride.
Ceramic glazes, a curing agent with peroxides to
form polyethylene wire insulation, a
color-changing component of temperature-
sensitive inks, a component of lubricating grease,
and a component of weighted nylon-reinforced
fish nets made of polyvinyl chloride fibers.
Glass sealing disks for IR sensors, wear-resistant
automotive shock absorbers, electrolytic
deposition of Pb, flux for brazing of aluminum
and its alloys, optical glass fibers for IR
transmission, and thin film batteries.
Artist's pigment, precursor of organolead
compounds, seawater-activated batteries,
expanding polymer mortar, flux for soldering cast
iron and cast brass, sound-insulating rubber
sealants, corrosion inhibitor for galvanized steel,
and infrared-transmitting glasses for CO2 lasers.
Filler for flame-resistant polypropylene, glass
optical waveguides for infrared thermometers and
catalysts for producing polyesters.
Aerosols for cloud seeding, making high-contrast
photographic images of laser radiation, high
capacity cathodes in lithium batteries, and
low-temperature thermographic copying
materials.
Formulating Pb-bearing glazes for ceramics,
source of PbO in glass manufacturing.

Ceramic glazes.
Glass and frit production.

Providing long-term heat stability to PVC,
electrical insulation, activation for
azodicarbonamide blowing agents for vinyl foam.
Source: Carr(2003).
2-6
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Table 2-4. Lead Oxides: Names, Formulae, Physical Characteristics, and Uses
Name
Formula
Form
Uses
Pb Monoxide
PbO
Pb Dioxide
PbO9
Pb Sesquioxide
RedPb
Reddish below 489°C,
yellow at high
temperatures
Brownish-black
crystalline powder of
fine flakes
Pb2O3 Amorphous,
orange-yellow powder

Pb3O4 Brilliant orange-red
pigment
Pastes for the grids of Pb-acid batteries, optical,
electrical, and electronic glasses, glazes for fine
tableware, vulcanizing agent for rubber, Pb soaps
used in driers as varnishes, high-temperature
lubricants, neutralizing agent in organic synthesis,
heat stabilizer in plastics, and starting material in
the production of pigments.

Active material of the positive plates in Pb-acid
batteries, oxidizing agent in the manufacture of
chemicals, dyes, matches, pyrotechnics, and liquid
polysulfide polymers, antifriction agent for plastic
sliding bearings, ballistic modifiers in high-energy
propellants, electrodes for seawater electrolysis,
filters for desulfurization of waste gases,
vulcanizing agents for butyl-rubber
puncture-sealing layers inside tires.

Ballistic modifier for high-energy propellants,
cathode material in lithium batteries, additive to
increase the shattering force of explosives.

Pigment in anticorrosion paints for steel surfaces,
Pb oxide pastes for tubular Pb-acid batteries,
ballistic modifiers for high-energy propellants,
ceramic glazes for porcelain, lubricants for hot
pressing metals, radiation-shielding foam coatings
in clinical x-ray exposures, and rubber adhesives
for roadway joints.
Source: Carr(2003).
the environment, i.e., sulfates (PbSO4), chlorides (PbCl2), carbonates (PbCO3, Pb(HCO3) 2),

hydroxides (Pb(OH) 2), nitrates (Pb(NO3) 2), phosphates (PbPO4, Pb(HPO4) 2), silicates, oxides

(PbO, Pb3O4), and PbS. With the exception of the covalently-bound sulfide and oxide, these

compounds are derived from acids (or the related anions) that are common in the environment,

such as sulfuric acid (H2SO4), nitric acid (HNO3)~, phosphoric acid (H3PO4), and carbonic acid

(H2CO3), an acid that forms when CO2 dissolves in water. Lead salts, once formed, tend to be

only slightly soluble in neutral solutions, but are quite soluble in the presence of acid (Weast

etal., 1988).
2-7
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Table 2-5. Lead Compounds Observed in the Environment
Location
Observed Pb Compounds
Minerals
Smelting Aerosols

Coal Combustion Aerosols

Coal Combustion Flue Gases

Wood Combustion

Waste Incineration Aerosols

Soils Near Mining Operations
Motor vehicle exhaust (combustion of leaded fuel)
Roadside dust

Other mobile sources:
Brake wear, wheel weights
NASCAR vehicle emissions
Aircraft engine wear
Lawn mowers
PbS (Galena)
PbO (Litharge, Massicot)
Pb3O4 (Minium or "Red Lead")
PbCO3 (Cemssite)
PbSO4 (Anglesite)

Pb°, PbS
PbSO4,PbO,PbSO4.PbO
PbCO3
Pb silicates

PbS
PbSe

Pb°, PbO, PbO2 (Above 1150K)
PbCl2 (Low rank coals, above 1150K)
PbSO4 (Below 1150 K)

PbCO3
PbCl2
PbO

PbCO3
PbSO4
[PbFe6(S04)4(OH)12]
[Pb5(P04)3Cl]
[Pb4S04(C03)2(OH)3]
PbS-Bi2S3
Pb oxides, silicates

PbBrCl
PbBrCl-2NH4Cl
PbBrCl-NH4Cl

PbSO4, Pb°, PbSO4 (NH4)SO4, Pb3O4,
PbO-PbSO4 and 2PbCO3-Pb(OH)2,PbSO4

Pb°
Pb halides
Pb°
Pb halides (Battery leakage)
"Source: Biggins and Harrison (1979,1980).
Lead Coordination Chemistry, and Its Role in Biochemistry

The formation of coordinate covalent complexes represents a different class of chemical

interaction from the formation of simple covalent compounds and salts. "Coordinate covalent"

bonds form when anions or neutral molecules interact with metal ions in solution that are capable
2-8
-------
of donating both of the electrons required to form a covalent bond. These molecules (or anions)
are called "ligands" or "electron donors." Ligands possess a filled valence orbital with a
geometry that allows it to overlap to a substantial degree with an empty orbital associated with
the metal ion. In the case of Pb, its large atomic size is associated with several out-lying empty
atomic orbitals leading to a tendency to form a large number of coordinate covalent bonds
(Claudio et al., 2003). This is suggested by the coordination number (9) of PbCb, in its
crystalline form, which is able to share electrons with 9 adjacent chloride ions (CF) (Douglas
etal., 1983).
Molecules capable of serving as ligands for metal ions in solution take many forms.
"Monodentate" ligands are molecules capable of providing 2 electrons to form a single
coordinate bond, such as water (H2O), ammonia (NH3); "multidentate" ligands can participate in
more than one coordinate bond. A common term for the binding of a metal ion by a multidentate
ligand is "chelation." The chelating agent, ethylenediaminetetraacetic acid (EDTA), is a well
known hexadentate ligand, containing 6 functional groups capable of forming 6-coordinate
bonds with metal ions in aqueous solution. Proteins, particularly the active sites of enzymes,
contain functional groups—usually associated with amino acid side chains—that can serve as
ligands for metal ions. In fact, the zinc finger proteins must form coordinate complexes with
Zn2+ ions to stabilize their active conformations (Claudio et al., 2003).
Several types of equilibrium constants for ligand-metal interactions can be derived,
depending on the property of interest. One formulation, the "binding constant (Kb)," between the
free metal ion and ligands in- solution with the ligand-metal complex, is derived below for a
negatively charged ligand:
[ML n~1]
Kh = binding constant = - (2-2)
[MLJ+HL-]
Where:
[ML-1] .
bl
_ [ML2-2]
b2
Etc.
2-9
-------
Kbi provides a measure of the stability of a solution of the free metal ion, Mn+ and an
individual ligand, L, compared to the complex of MLn+. Alternatively, KM gives an indication of
the strength of the interaction between Mn+ and L. Thus, Kb2 indicates the strength of the
interaction between the MLn+ complex and an additional ligand, L. Subsequent additions of
ligands to the complex are described following the same convention. Binding constants are
useful, in particular, for evaluating the strength of interactions between metals and small
(monodentate) ligands. The form typically used to evaluate binding between metals and proteins
is the "dissociation" constant, Kd. The example given here is for a neutral ligand:
n+ -i T—r o -i
Kd = dissociation constant = - - — — — — - (2-3)
[MLxn+]
Where:
Kdl=-
Kd2 =
[MLn+] '
[MLn+][L°]
m /rr n+ -i
[ML2 ]
Etc.

Kd is the inverse of Kb, in that it refers to stability of the existing complex between ligand
and metal, versus the free metal ion and the free ligand. Kii is a measure of the strength of the
bond between an individual ligand and the metal-ligand complex. Kd2 indicates the strength of
the interaction as the second ligand is, subsequently, removed. A variety of quantitative,
analytical methods are available for measuring the binding and dissociation constants for specific
combinations of metals and ligands.
A simple, qualitative model commonly used for discussing the relative strength of
coordinate covalent bonding between different metals and ligands is the Pearson's Hard-Soft
Acid-Base (HSAB) model (Douglas et al., 1983). Heavier metals, such as Pb, which have more
electrons and more spatially diffuse valence orbitals, are described as "soft" (Lewis) acids.
Lighter metals, with fewer electrons and more closely-spaced valence orbitals, are described as
"hard" (Lewis) acids. These metals tend to preferentially bond with ligands having similar
electronic properties. Hard acids tend, for example, to prefer oxygen-based ligands, i.e.,
2-10
-------
"hard bases," and soft acids prefer ligands based on larger atoms, such as sulfur and selenium,
i.e., "soft bases."
The HSAB concept is useful for understanding the behavior of Pb in the biological
context. Lead forms coordinate covalent bonds with ligand atoms with effectiveness that
declines with atomic size. For example, Pb forms especially stable bonds with sulfur and
sulfur-containing compounds, but somewhat less so with carboxylic acids (O-based ligands) and
imidazoles (N-based ligands) (Claudio et al., 2003).
In biological systems, Pb competes very effectively with native or homeostatic metal ions
for binding with the sulphahydryl, carboxyl and imidazole side-chains comprising enzyme active
sites. This competition leads to inhibition of enzyme activity, as well as the replacement of
calcium in bone and, ultimately, to a substantial list of negative human health effects. The
relative strength of these different interactions appears to be reasonably well-predicted by the
HSAB model.
By far, the most effective biological ligands for Pb are amino acid side-chains containing
sulfur and selenium. Smaller electron donors (hard bases), such as carboxylic acids that bind Pb
via electrons associated with oxygen, form weaker bonds. These complexes are generally more
labile, i.e., bonds form and break rapidly, thus allowing more effective competition at protein
binding sites amongst metals available in solution. Simple ligands that are examples of this case
are the amine functional group, -NH, and the thiol functional group, -SH. The amine group has a
Pb binding constant on the order of 100, whereas the thiol group binding constant is on the order
of 107. Example proteins in this instance are carboxypeptidase A, a zinc-binding protein, with
carboxylate and histidine side-chains, and the four cysteine zinc finger consensus peptide,
CP-CCC. Carboxypeptidase A has a Pb dissociation constant of approximately 10"4 M, versus
that of the zinc finger protein, which is 3.9 x 10"14 M. Claudio et al. (2003) concluded, on the
basis of these values, that carboxypeptidase A is unlikely to be a protein associated with Pb
poisoning, while cysteine-rich proteins, including the zinc enzyme, d-aminolevulinic acid
dehydratase (ALAD), the second enzyme in the heme biosynthetic pathway, are more likely
targets. ALAD active site, with its Cys3 active site, is known to be inhibited at femtomolar
(10"15 M) concentrations of Pb in vitro.
Figure 2-1 illustrates the wide array of possible inhibitory interactions between Pb2+ and
proteins responsible for transduction at nerve synapses. Targets for Pb2+ interference at the
2-11
-------
Figure 2-1. Multiple possible molecular targets for interference by Pb + ion at
nerve synapses.
Source: Nihei and Guilarte (2002).
2+
presynaptic terminal include synaptic vesicles, ionotropic receptors, Ca and other channel
proteins, and kinase proteins. At the postsynaptic interface, ionotropic proteins, dopamine
receptors, protein kinase-C isoenzymes and ion channel proteins are amongst the proteins subject
to interference by Pb2+ (Nihei and Guliarte, 2002).
Additional information concerning the physical aspects of Pb coordination chemistry and
its role in biological systems can be obtained from the substantial review by Claudio et al.
(2003). An extensive discussion of the neuro- and other toxic effects associated with exposure
to Pb can be found in Chapter 5 of this document.
2-12
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2.2 SOURCES OF LEAD
In this section, information is summarized with regard to a number of major sources of
Pb, categorized as natural sources, stationary point sources, and mobile sources. In addition to
these categories, fugitive emissions such as resuspension of Pb in soil and dust can be important.
Resuspension is considered a transport route and is discussed in Section 2.3.

2.2.1 Natural Sources
Common natural sources of airborne Pb include volcanoes, sea-salt spray, biogenic
sources, wild forest fires, and wind-borne soil particles in rural areas with background soil
concentrations. Natural sources combined contribute an estimated 19,000 metric tons of Pb to
the air each year (Nriagu and Pacyna, 1988). However, there is significant variability in the Pb
emissions from volcanoes and forest fires and considerable uncertainty in biogenic and sea-salt
emissions of Pb (Nriagu, 1989). Table 2-6 shows the median value and the range of annual
emissions worldwide for natural sources of airborne Pb.
Table 2-6. Worldwide Annual Emissions of Lead from Natural Sources
Amount Emitted: Range
Source (thousands of metric tons/yr)
Wind-borne soil particles
Sea salt Spray
Volcanoes
Wild Forest Fires
Biogenic, continental particulates
Biogenic, continental volatiles
Biogenic marine sources
Total
0.3-7.5
0.02-2.8
0.54-6.0
0.06-3.8
0.02-2.5
0.01-0.038
0.02-0.45
0.97-23
Amount Emitted: Median
(thousands of metric tons/yr)
3.9
1.4
3.3
1.9
1.3
0.20
0.24
12
Source: Nriagu (1989).
2-13
-------
Natural Pb emissions worldwide are somewhat greater than the estimated 1600 metric
tons/year of Pb emitted from anthropogenic stationary and mobile sources in the United States in
the year 2002 (U.S. Environmental Protection Agency, 2003). However, many countries around
the world have much greater Pb emissions than the United States from stationary and mobile
sources, including several countries that still use leaded gasoline. Furthermore, the EPA estimate
does not account for emissions of Pb in resuspended soil. Harris and Davidson (2005) estimate
that stationary and mobile source emissions account for only about 10% of the total Pb emissions
in the South Coast Air Basin of California; the remaining 90% of the emissions are from
resuspended soil. The soil contains elevated Pb levels because of the many decades of leaded
gasoline use. Therefore, on a worldwide basis, anthropogenic emissions of Pb are expected to be
much greater than natural emissions.
There are four stable isotopes of Pb: 204Pb, 206Pb, 207Pb, and 208Pb. The last three of these
isotopes are produced by decay of 238U, 235U, and 232Th respectively. The concentrations of
natural vs. anthropogenically derived Pb in environmental media are often determined through
isotopic ratios. Most minable Pb ores exhibit ratios of 206Pb/207Pb between 0.92 and 1.20
(Erel et al., 1997). Rock released or "natural" Pb, however, generally exhibits a higher
Deep soil samples converge to ratios of 206Pb/207Pb ~ 1.21 and 208Pb/206Pb ~ 2.05, which
are considerably different from the natural ratios found in adjacent bedrock (Erel et al., 1997).
For more information on isotopic ratios of Pb and their uses as environmental tracers, see
Chapter 7 of this document.
Natural aerosol Pb tends to have large particle sizes (Reuer and Weiss, 2002). As a result,
it deposits rapidly and has an atmospheric residence time of a few hours to -10 days (Reuer and
Weiss, 2002). The average downward flux is estimated as 0.012 mg nT2 yr"1 for natural Pb in all
forms (Hindi er et al., 1999).
Lead concentrations in air and soil have most likely been elevated by anthropogenic
activities at least since the rise of the Greek and Roman societies, both of which extensively used
Pb. The natural background (i.e., pre-Greek and Roman eras) concentration of Pb in soil is
<0. 1 ppm (Bindler et al., 1999; Erel et al., 1997). This is significantly higher than the adjacent
bedrock but is approximately equal to concentrations found in bedrock residues such as quartz
and clay (Erel et al., 1997). An estimated 3.1 x 1014 metric tons of Pb are dispersed within the
2-14
-------
continental crust (Reuer and Weiss, 2002). Of this, -9.3 x 107 metric tons of Pb are found in Pb
ores. Table 2-7 lists the naturally occurring concentrations of Pb in bedrocks, ocean crusts, and
continental crusts. Spatially, background levels of Pb vary considerably.

Table 2-7. Naturally Occurring Lead Concentrations in Major Rock Types
Lithology Natural Pb Concentration (ppm)
Continental Crust 15.0
Oceanic Crust 0.9
Basalts, Gabbros 3.5
Limestones 5.0
Granulites 9.8
Greywackes 14.0
Gneisses, Mica Schists 22.0
Shales 22.0
Granites 32.0

Source: Reuer and Weiss (2002).
Natural Pb in surface water is derived from four different sources: biogenic material,
aeolian particles, fluvial particles, and erosion (Ritson et al., 1994). As discussed in Chapter 7
(Section 7.2.2), Pb is generally present in aquatic systems as Pb salts, such as PbSCu, PbCU,
Pb(OH)2, etc. About 90% of natural Pb in the open ocean is in the dissolved phase (Reuer and
Weiss, 2002). Organic ligands are complexed with 50 to 70% of this Pb, with the balance of Pb
in open ocean waters being found in inorganic compounds (Reuer and Weiss, 2002). Biological
particles in the open ocean scavenge a significant portion of the Pb complexes, which have an
estimated two-year residence time in the surface waters (Reuer and Weiss, 2002).
A naturally occurring radioactive isotope of Pb, 210Pb, is commonly used as a tracer to
determine how particles are transported through the environment. The source of 210Pb is the
238U decay series. In this process, gaseous 222Rn is produced, which escapes from the soil and
enters the atmosphere. As radon decays into 210Pb, the particulate Pb deposits onto soils and
2-15
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surface waters all over the world. The surfaces of all soils have been exposed to atmospherically
derived 210Pb particles (Kaste et al., 2003).
Particles of 210Pb tend to be submicron, with an average size of 0.53 jim AMD (Winkler
et al., 1998). The mean residence time for 210Pb in the air is -4 to 5 days but has been estimated
to be as long as 8 days with some seasonal variability (Winkler et al., 1998). The downward flux
has been estimated as 136 Bq nT2 yr"1 for 210Pb (Joshi et al., 1991).
Atmospheric deposition is likely the largest source of 210Pb to water bodies. Leaching
of Pb naturally contained in host rock is a very small source to water (Toner et al., 2003).
In surface waters, 210Pb is primarily in particulate form, whereas dissolved Pb is transported
more readily (Joshi et al., 1991). Dissolved 210Pb is scavenged by suspended matter (Carvalho,
1997). The residence time of dissolved 210Pb is -30 days, although partial re-dissolution from
bottom sediments probably occurs (Carvalho, 1997). One estimate found that -56% of
atmospherically derived 210Pb in lakes of the Canadian Shield was retained in the sediment
(Joshi etal., 1991).
Many authors have measured concentrations of 210Pb in plants (including foodstuffs) and
animals (including humans). Holtzman (1978) summarized these measurements. Concentrations
in United States vegetation range between 30 pCi/kg and 70,000 pCi/kg for wheat and lichens
respectively. The estimated human consumption of 210Pb from vegetation averages 1.4 pCi/day
in the United States. Overall, the concentrations of 210Pb in animals vary significantly depending
on the type of tissue or organ measured. However, concentrations are generally higher in
animals with higher rates of Pb intake.

2.2.2 Lead Emissions in the United States
Currently, the major use of Pb in the United States is in Pb-acid batteries, for which the
demand is increasing (Socolow and Thomas, 1997). Other major uses are for glass, paints,
pigments, and ammunition. United States consumption of Pb by industry is shown in Figure 2-2.
The consumption reached -1.4 million metric tons per year in the mid 1990s (Socolow and
Thomas, 1997). Approximately 910,000 metric tons of this was secondary production,
indicating high rates of Pb recycling.
2-16
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H Other
[gj Bearing Metals
13 Brass and Bronze
1HO Casting Metal
• Type metal
H Pipe and Sheet Lead
Q Caulking Lead
Q Cable Coverings
M Solder
§3 Ammunition
Q Pigments and paints
HJ Gasoline additives
[1 Batteries
1968 1972 1976 1980 1984 1989 1993 1997 2001 2003
Year
Figure 2-2. Annual lead production and use in the United States (1968 - 2003).
Source: U.S. Bureau of Mines (1968-1995) and USGS (1996-2003).
Table 2-8 presents estimates of annual emissions rates for Pb in the U.S. from 1990 and
2002 (20 tons per year or more in 2002). These values were extracted from the 1990 Emissions
Inventory of Forty Potential Section 112(k) Pollutants and the 2002 National Emissions
Inventory (NEI) (U.S. Environmental Protection Agency, 1999, 2006a). For criteria air
pollutants, sources are required to report emissions either annually or every three years,
depending on the size category of the sources. The NEI contains estimates of facility-specific air
pollutant emissions and the source-specific factors necessary for modeling such as location and
facility characteristics (e.g., stack height, exit velocity, temperature). Complete source category
coverage is needed, and the NEI contains estimates of emissions from stationary point and
nonpoint (sources such as residential heating that are inventoried at the county level) and mobile
source categories. The NEI contains individual stack and fugitive emissions estimates at
individual geocoordinates for point sources. County level estimates are provided in the NEI
for nonpoint and mobile sources. Further information on the NEI is available at:
http://www.epa.gov/ttn/chief/net/index.html. Note that all emission rates are estimated based
on periodic reporting by sources of their activities, production, and fuel consumption, multiplied
2-17
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Table 2-8. Annual Air Emission Rates for U.S. Lead Sources (20 Tons per Year or
Greater) for 1990 and 2002, Ordered by Emissions Levels in 2002. These Data were
Extracted from the 1990 Emissions Inventory of Forty Potential Section 112(k)
Pollutants and the 2002 National Emissions Inventory (NEI) (U.S. EPA. 2006a).
1990 112(k) Total
Source Category Name Emissions (TPY) a
ALL CATEGORIES
Industrial/Commercial/ Institutional
Boilers & Process Heaters
Utility Boilers: Coal
Mobile sources
Iron and Steel Foundries
Hazardous Waste Incineration (incl onsite)
Primary Pb Smelting
Electric Services
National Security
Municipal Waste Combustors: Small
& Large
Integrated Iron & Steel Manufacturing (b)
Pressed and Blown Glass and Glassware
Manufacturing
Secondary Nonferrous Metals
Pb and Zinc Ores
Pb Acid Battery Manufacturing
Stainless and Nonstainless Steel
Manufacturing (EAF)
Primary Copper Smelting
Portland Cement Manufacturing
Primary Metal Products Manufacturing
3270.0
33.7
72.0
1198.0
13.5
96.7
220.0
c
d
80.4
23.5
52.4
55.8
c
54.8
84.0
152.0
e
2.0
2002 Total
Emissions (TPY)a
1435.0
247.0
165.4
142.8
110.0
69.7
58.9
53.2
34.1
32.8
32.7
30.4
28.3
25.6
24.9
23.4
21.7
21.4
20.5
1990%
emissions

1.03
2.20
36.64
0.41
2.96
6.73
—
—
2.46
0.72
1.60
1.71
—
1.68
2.57
4.65
—
0.06
2002%
emissions

17.21
11.53
9.95
7.67
4.86
4.10
3.71
2.38
2.29
2.28
2.12
1.97
1.78
1.73
1.63
1.51
1.49
1.43
a) Categories with emissions >20 TPY in 1990 112(k) or 2002 NEI.
b) Listed as Blast Furnaces and Steel Mills in 1990 112(k).
c) Source category was not included in the 1990 112(k) inventory.
d) Source category was included in the 1990 112(k) inventory but no Pb emissions were identified.
e) Source category was included in 1990 112(k) inventory but no Pb emissions were quantified.
2-18
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by sector-wide average emission factors per unit of activity, production, or fuel consumed.
There are numerous uncertainties in the estimation of emissions inventory data. Emissions
measurements on a per facility basis are scarce. EPA's AP-42 document includes emission
factors for many different processes and operations The uncertainties underlying emissions
estimates are greater for some sources than others, and the AP-42 document includes emission
factor ratings from A (excellent) through E (poor) based on the quality of the testing or
measurement data and how well the factor represents the emission source. Further information
on emission factors is available at: http://www.epa.gov/ttn/chief/ap42/.
Regardless of these uncertainties, the long-term trend in U.S. Pb emissions is very clear.
Total Pb emissions to the air declined by about half between 1990 and 2002, from -3300 tons
per year to -1400 tons per year, largely due to the decline in emissions from mobile sources.
In the summary data for 1990, mobile sources were by far the largest source of Pb emissions to
the air. In 2002 NEI data, mobile sources remain an important source of airborne Pb emissions
(from specific types of mobile sources discussed further in Section 2.2.4), but industrial sources
and combustion sources have become the major sources of U.S. Pb emissions. As shown in
Table 2-8, the largest U.S. emissions sources for airborne Pb in 2002 are industrial sources, such
as industrial/commercial/institutional boilers or process heaters and utility boilers.
Another source of information on U.S. Pb emissions to air is the EPA's Toxics Release
Inventory (TRI) (U.S. Environmental Protection Agency, 2006b). Reported emissions to the air
in 2004, including fugitive and point source emissions from U.S. facilities, totaled over 1 million
pounds, or -500 tons. This is consistent with the emissions estimates reported for 2002 in the
NEI (1400 tons), in that the TRI includes emissions data reported by facilities, whereas the NEI
includes estimated emissions from a broader group of sources. For more information about the
TRI, see: http://www.epa.gov/ebtpages/emerreportingtoxicsreleaseinventorytri.html.
Information is also available from the peer-reviewed literature from studies done at
specific sites. These data are summarized in the following discussions. Section 2.2.3 focuses on
data available for numerous stationary source categories, and Section 2.2.4 presents data on Pb
emissions from mobile sources.
2-19
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2.2.3 Stationary Sources
As observed in the preceding section, emissions estimates and measurements on a per
facility basis are scarce. The AP-42 document of the EPA includes emission factors for many
different processes and operations. For Pb-processing facilities, these emission factors are
usually expressed as grams of Pb emitted per kg of Pb processed (U.S. Environmental Protection
Agency, 2005). In general, AP-42 data are not listed in the following sections except in the
absence of newer, more robust or peer-reviewed data on process emissions. Although AP-42
emission factors can provide a first order estimate, they are limited in that they are often derived
from one individual source and do not reflect the variability between sources (U.S.
Environmental Protection Agency, 2006c). Also, in many cases, AP-42 emission factors do not
account for process parameters. In some cases, AP-42 may complement the data listed below,
and the reader is referred there for emission factors not given in this chapter. Up-to-date,
accurate emissions estimates are critical as inputs to models predicting airborne concentrations,
and more research in this area is needed.

Primary and Secondary Lead Smelters
Primary Pb smelting is the process by which elemental Pb is recovered from Pb ore.
Lead ore is primarily in the form of galena (PbS) but can also occur as plattnerite (PbO2),
cerussite (PbCOs), and anglesite (PbSO4) (Reuer and Weiss, 2002). Producing elemental Pb
from ore involves three processes - sintering, reduction, and refining - each with its own
characteristic emissions. Primary Pb production in the United States emitted about 58.9 metric
tons of Pb in 2002, -4% of total anthropogenic Pb emissions in the United States (U.S.
Environmental Protection Agency, 2006a). Currently, there is only one remaining primary Pb
smelting facility still operating in the United States. The Pb emissions from this facility,
Doe Run's Herculaneum, Mo. facility, were 25 tons in 2005.
Secondary Pb smelters reclaim scrap Pb. Both the principal input to and the principal
major product market of secondary smelters are Pb-acid batteries. Secondary Pb production
contributed 82% of total Pb production in 2003 (USGS, 2003). Secondary Pb production in the
United States emitted about 4.3 metric tons of Pb in 2002, <1.0% of total anthropogenic Pb
emissions in the United States (U.S. Environmental Protection Agency, 2006a). Although
recycling of Pb-acid batteries with minimal emissions may be possible (Socolow and Thomas,
2-20
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1997) secondary smelters and battery recycling facilities are still one of the most significant
stationary sources of airborne Pb emissions.
The quantity of Pb emitted from a given facility is highly variable and depends on facility
processes and meteorological conditions such as wind speed and ambient temperature.
Emissions estimates are typically performed through direct measurements, mass balances,
process models, inverse inferences, or emissions factors (Frey and Small, 2003).
Emissions from smelters have been measured in several cases. A survey of approximately
50 European Pb smelters had mean emission factors of 0.1 grams and 0.05 grams of Pb emitted
per kg of Pb processed for primary and secondary Pb smelters respectively (Baldasano et al.,
1997). Measurements of emissions from the blast furnace of a primary smelter were between
1.2 and 3.8 kg Pb/hr (Bennett and Knapp, 1989). The acid-sinter at the same plant emitted
between 0.4 and 8.5 kg Pb/hr (Bennett and Knapp, 1989). Emissions occur during every stage of
the overall smelting process. Because the process emissions mostly are controlled to conserve
raw materials, the largest source of emissions is likely to be fugitive dust from the transport,
grinding, and storage of battery scrap (Kimbrough and Suffet, 1995).
Much work has been done to determine the species of Pb emitted from the various
smelting processes. The fraction of Pb in particulate matter (PM) emissions varies significantly
between processes and depends on the type of furnace used. However, Pb is often the dominant
element in smelter emissions. Lead can be emitted either in PM or in fumes. Lead fume
emissions are particularly high if Pb blast furnace bullion is transferred in an open ladle (Wang
and Morris, 1995). Major components of particulate Pb emissions are PbS, PbSC>4, PbSCVPbO,
and elemental Pb; and minor species are PbCOs, PbO, Pb silicates, and PbO litharge (Batonneau
et al., 2004; Harrison and Williams, 1983; Ohmsen, 2001; Sobanska et al., 1999; Rieuwarts and
Farago, 1995).
The distribution of particle sizes varies depending on temperature, process, and the
conditions of each facility. Ohmsen (2001) found that Pb emissions from a blast furnace tend to
be less than 1 jim in size and have a smaller diameter than particulate emissions from either the
sintering process or storage areas. Higher temperatures (>600 °C) in the blast furnace tend to
produce emissions with finer particle sizes. Dusts from the raw materials area tend to fall
between 10 and 100 jim, whereas dusts from the refinery tend to fall between 1 and 30 jim
(Ohmsen, 2001). Sobanska et al. (1999) found that just 15% of dust particles by mass emitted
2-21
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from a "water jacket" furnace were <10 jim in diameter, and the remaining 85% fell between
10 and 100 jim. Measurements of Harrison et al. (1981) at a primary smelter found particles
derived from combustion processes to be typically between 0.1 and 2 jam, but also showed that
these particles could agglomerate to more than 10 jim if confined to ventilation ducts. Reported
sizes from primary smelting processes are shown in Table 2-9.
Table 2-9. Mass-Median Aerodynamic Diameters for Particles from Various Processes
at Primary Lead Smelters
Primary Smelter
Process
Raw Materials
Sinter
Blast Furnace
Copper Drosser
Refinery

Harrison et al.
(1981)
—
5.1 um
3.4 um
9.4 um
—
Average Particle Size
Ohmsen
(2001)
40 um (range = 10-100 um)
range = 10-300 um
90% of particles were <1 um
range = 10-300 um
range = -1-100 um,
mostly <20 um

Bennett and Knapp
(1989)
—
0.91 um, 80% of
particles <10 um
1.1 um, 88% of
particles <10 um
—
—
Note: Where there were multiple data points, geometric means were used. Data for Harrison et al. (1981) were
occasionally given as >11 um. These values were replaced with 11 um before calculating the geometric mean.
Thus, these values represent a lower limit.
Source: Harrison et al. (1981), Ohmsen (2001), Bennett and Knapp (1989).
Lead concentrations in stack outlets have been measured in several cases. Measurements
taken at the stack of a blast furnace in a primary smelter ranged between 3.7 and 7.3 mg/m3
(Bennett and Knapp, 1989). Stack concentrations at the sinter plant of the same facility ranged
between 4.5 and 71.0 mg/m3 (Bennett and Knapp, 1989). Two stacks on a blast furnace at a
secondary smelting facility had Pb concentrations of 0.002 and 0.0137 mg/m3 (Sturges and
Harrison, 1986). The average values of-50 European smelters were 2 mg/m3 for both primary
and secondary smelters (Baldasano, et al., 1997).
The ambient air concentrations in the immediate vicinity of smelters tend to be elevated to
varying degrees depending on facility operations and meteorological conditions. In the UK, an
2-22
-------
increase of 1.5 |ig/m3 in the local ambient air was attributed to the emissions of a single
secondary Pb smelter (Sturges and Harrison, 1986). Fenceline measurements at two secondary
smelters located in California ranged between 0.85 and 4.0 |ig/m3 (Kimbrough and Suffet, 1995).
Air-Pb concentration data measured at 50 m, 500 m, and 800 m from the plant were slightly
lower but generally on the same order of magnitude as the fenceline values. Ambient air
concentrations measured at 12 sites within several hundred meters of three secondary Pb
smelters in Manitoba were elevated (Tsai, 1987). The geometric means of these samples, which
were taken over three month time spans, ranged between 0.11 and 1.69 |ig/m3. Also, the area
was shown to be much less likely to meet the Manitoba guideline of <5 |ig/m3 for a 24-hour
average when the smelters were operating than when they were not.

NonleadMetallurgical Processes
Emissions of Pb from nonlead smelters can be significant. Emissions from smelters,
metal works, and metal refineries depend on the type of equipment used to process the metals,
the concentrations of Pb in the initial material (ore, recycled material, or alloy), the type and
effectiveness of pollution controls at the facility, and the temperature of operations (Pacyna,
1986). Little work has been done to speciate Pb emissions from metallurgical facilities, although
Pb emissions from a primary copper-nickel smelter are primarily in the form of PbO (Barcan,
2002). Emissions from non-Pb metallurgical processes are summarized in Table 2-10.

Ore Mining and Processing
Lead mining occurs in 47 countries, although primary Pb production is on the decline
(Dudka and Adriano, 1997). World mine production of Pb is -2.8 million metric tons per year
(Wernick and Themelis, 1998). The reserve base of Pb is estimated to be about 120 million
metric tons, which will sustain current rates of mine production for 43 years (Wernick and
Themelis, 1998).
Mines can be a significant source of metal emissions to the atmosphere. Lead and zinc
ores, which are often mined together, frequently contain high concentrations of cadmium and
arsenic (Pacyna, 1986). An emission factor for Pb mines has been reported as 0.91 grams of Pb
emitted to the air per kg of Pb mined (Pacyna, 1986).
2-23
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Table 2-10. Emissions of Lead from Nonlead Metallurgical Processes
to
to
Metallurgical Plant Lead Emissions
Particle Sizes
MMAD = Mass median
aerodynamic diameter
Location
Source
Aluminum (secondary)

Aluminum (secondary)
Aluminum (secondary)
burning drying

Aluminum (secondary)
reverberatory furnace

Antimony

Brass/Bronze refinery
Brass/Bronze refinery -
blast furnace

Brass/Bronze refinery -
crucible furnace

Brass/Bronze refinery -
cupola furnace

Brass/Bronze refinery -
reverberatory furnace

Brass/Bronze refinery -
rotary furnace
0.81±0.014% of PM emissions

0.098±0.031% of PM emissions

1.01x10-3-3.52x10-3 kg/mt produced
(venturi scrubber)

3.38x10-6-7.40x10-6 kg/mt produced
(baghouse)

1.05x10-2-1.13x10-2 kg/mt produced
(multiple cyclones)

5.0xlO-4-l.lxlO-3 kg/mt processed
(baghouse)

0.17±0.04% of PM emissions

0.11±0.02% of PM emissions

0.01-1% of PM emissions

16 g/ton produced

10 g/ton produced

65 g/ton produced

60 g/ton produced

60 g/ton produced
Fine (<2.5 um)

Coarse (2.5-10 um)
n.a.

n.a.

Fine (<2.5 um)

Coarse (2.5-10 um)

n.a.

n.a.
Philadelphia, USA

Philadelphia, USA
U.S.

U.S.

Philadelphia, USA

n.a.

n.a.
Olmezetal. (1988)

Olmezetal. (1988)
U.S. EPA (1998)

U.S. EPA (1998)

Olmezetal. (1988)

Lee and Von Lehmden
(1973)

Pacyna (1986)

Pacyna (1986)
-------
Table 2-10 (cont'd). Emissions of Lead from Nonlead Metallurgical Processes
Metallurgical Plant
Lead Emissions
Particle Sizes
MMAD = Mass median
aerodynamic diameter
Location
Source
to
to
Brass/Bronze production

Copper-Nickel

Copper-Nickel (primary)

Copper (primary) smelter

Copper (primary) converter

Copper (secondary)
reverberatory furnace

Copper (secondary) smelter

Copper Smelter - furnace

Copper Smelter - sinter

Copper Smelter (secondary)

Iron Ore Recovery and
Ni refinery
25 kg/mt produced
(high-leaded alloys)

6.6 kg/mt produced
(red and yellow Pb alloys)

2.5 kg/mt produced (other alloys)

184mt/yr, 21 kg/hr

13.4 mt/year

0.6-1.4% of PM emissions

2.3-3.6 kg/ton produced

3.1 kg/ton produced

3.0x 10~2 kg/ton produced

0.27 kg/ton produced

2.5 - 25 kg/mt produced

5.0*10~4 kg/mt processed

0.24-0.52 kg/hr

below detection

54-214 g/ton produced

6 mt/year
n.a.

1.2 urn MMAD

0.9 urn MMAD

n.a.

0.87 urn MMAD

<0.10 urn MMAD

n.a.

Coarse (2.5-10 urn)
U.S.

Copper Cliff, Ontario

Falconbridge, Ontario

Monchegorsk, Russia

Poland

n.a.

U.S.

n.a.

n.a.
U.S. EPA (1998)

Chan and Lusis (1986)

Barcan (2002)

Pacyna (1986)

U.S. EPA (1998)

Bennett and Knapp (1989)

Pacyna (1986)
Copper Cliff, Ontario Chan and Lusis (1986)
-------
Table 2-10 (cont'd). Emissions of Lead from Nonlead Metallurgical Processes
to
to
Metallurgical Plant

Iron foundry cupola
Iron foundry -reverberatory
furnace
Iron foundry - electric
induction furnace
Iron foundry - casting
Iron and Steel foundry

Steel works - electric-arc
furnace
Zinc-Cadmium (primary)
Zinc Smelter - furnace
Zinc Smelter - sinter
Particle Sizes
MMAD = Mass median
Lead Emissions aerodynamic diameter
0.05-1.10 kg/mt produced
(no control device)
7.80 xlO"4 kg/mt processed
(afterburner, venturi scrubber)
6.95 x 10-4-2.23 x 10-3 kg/mt n.a.
produced (baghouse)
6.00xlO"3-7.00xlO"2kg/mt produced n.a.
(no control device)
4.45xlO"3-5.00xlO"2 kg/mt produced n.a.
(no control device)
2.40 x 10"3 kg/mt processed n.a.
(afterburner, venturi scrubber)
0. 01-0.1% of PM emissions n.a.

4.1-16.3 g/ton produced n. a.
1 .2-25 kg/ton produced n.a.
0.86-1.5 kg/hr 1.8-2.2 urn MMAD
3.6-6.0 kg/hr 0.9-2.1 urn MMAD
Location Source

U.S. U.S. EPA (1998)
U.S. U.S. EPA (1998)
U.S. U.S. EPA (1998)
U.S. U.S. EPA (1998)
n.a. Lee and Von Lehmden
(1973)
n.a. Pacyna (1986)
n.a. Pacyna (1986)
n.a. Bennett and Knapp (1989)
n.a. Bennett and Knapp (1989)
Source: Olmez et al. (1988), Lee and Von Lehmden (1973), Pacyna (1986), Chan and Lusis (1986), Barcan (2002), Bennett and Knapp (1989).
-------
Since Pb is mined in the form of galena (PbS), emissions from Pb mines tend also to be in
the form of galena (Dudka and Adriano, 1997). However, other species have been detected.
In mine spoils, Pb is typically galena and secondary alternation products such as plumbojarosite
[PbFee(SO4)4(OH)i2] (Rieuwerts and Farago, 1995). Other Pb forms detected in the vicinity of
mines are: pyromorphite [Pbs^O^Cl], which has a low bioavailability; PbCOs, which is
formed from the weathering of galena in the soil; leadhillite [Pb4SO4(CO3)2(OH)2]; PbSrE^Ss;
Pb oxides; Pb silicates; and PbSC>4 (Rieuwerts and Farago, 1995).
Although mining can be considered a point source to air, mine wastes can have a major
widespread effect on soil and water (Riewerts and Farago, 1995). Mines produce four different
types of large-volume waste: mine waste, which consists of overburden and barren rocks,
tailings, dump heap leachate, and mine water (Dudka and Adriano, 1997). Tailings especially
are major sources of metal contamination to soil and water (Bridge, 2004). Acid mine drainage
can contain highly elevated levels of Pb, >3000 |ig/L, and can contaminate vast areas (Bridge,
2004; Kurkjian et al., 2004). Soil contamination from both active and abandoned mines can be a
significant source of airborne Pb from fugitive or wind blown matter. Resuspension of
contaminated soil is addressed later in this chapter, and soil Pb concentrations near mines are
discussed in Chapter 3.
Mining of materials other than Pb can also release Pb to the atmosphere. Zinc-copper
ores, for example, contain Pb in the range of 100-100,000 ppm (Lee and Von Lehmden, 1973),
and about 6.1% of all Pb in the United States is extracted from "zinc mines" (Dudka and
Adriano, 1997).
In an underground gold mine, high Pb-particulate concentrations were associated with
blasting (Annegarn et al., 1988). These particles were primarily Pb oxides and submicron in
size. A source apportionment analysis on airborne PM in an underground gold mine found that
the significant sources of Pb were rock dust and diesel exhaust (McDonald et al., 2003).
Concentrations of airborne Pb inside the mine were measured at 0.21 |ig/m3.

Stationary External Combustion: Coal Combustion
Coal combustion can be a significant local source of Pb emissions as well as a
considerable regional source of airborne Pb. Coal is commonly burned as a fuel for utilities,
industries, and commercial and institutional facilities. Coal utility boilers accounted for 165 tons
2-27
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of Pb emissions in 2002 (10% of total emissions) (U.S. Environmental Protection Agency,
2006a). Coal is pulverized, fluidized, or gasified before combustion. Generally, Pb impurities
will volatilize early in the combustion process, although the precise rate of vaporization depends
on the distribution of Pb particles in the coal and the particle sizes (Lockwood and Yousif, 2000).
As Pb vapors cool, they condense, either forming individual particles or condensing on the
surface of ash particles (Lockwood and Yousif, 2000; Furimsky, 2000; Clarke, 1993; Pacyna,
1986). A high surface area to volume ratio makes fine ash particles better candidates for surface
sorption than coarse particles. Additionally, recondensed Pb particles tend to be fine, with an
average size of 0.2 jim (Lockwood and Yousif, 2000). The fine fraction of PM from coal
combustion has an enrichment factor of-22 (Lockwood and Yousif, 2000).
The primary contributor of Pb emissions from coal combustion is the Pb content of the
coal itself. Lead is present in all coal samples in varying amounts, depending on the location of
the coalfield and even the location of the coal sample within a coalfield. Generally, Pb is present
in trace amounts in the form of PbS, but it can also be present as pyrite and PbSe (Lockwood and
Yousif, 2000; Mukherjee and Srivastava, 2005). The type of the coal - either bituminous,
subbituminous, or lignite - does not seem to correlate with the quantity of trace elements
(Mukherjee and Srivastava, 2005). The age of the coal also does not seem to impact the Pb
concentration (Ghosh et al., 1987). The most important factors contributing to Pb content of
uncombusted coal seem to be local environmental conditions at the time the coal formed and
relative proportions of organic and inorganic matter (Pacyna, 1986; Ghosh et al., 1987).
Globally, Pb concentrations in coal range between 2 and 80 ppm (Mukherjee and Srivastava,
2005). Table 2-11 lists the range of Pb concentrations measured in four different coal
components.

Table 2-11. The Range of Lead Concentrations in Coal Lithotypes
Coal Lithotype Range of Lead Concentrations (ppm)
Vitrain 0.30-16.17
Clarain 4.84-17.55
Durain 4.10-11.76
Fusain 3.64-15.60
Source: Ghosh etal. (1987).
2-28
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Coal is often combined with limestone to attenuate sulfur dioxide emissions. However,
limestone can contain trace elements and has been shown to increase emissions of Pb by four to
six times in a fluidized bed system compared to tests performed without a limestone addition
(Clarke, 1993). Other measurements performed on a fluidized bed system found that increasing
limestone increased particulate emissions of Pb, but decreased gaseous emissions of Pb. The
overall emissions of Pb (gaseous + particulate) remained relatively constant (Furimsky, 2000).
Limestone had a negligible effect on pressurized fluidized bed systems, although Pb emissions
from gasification systems may increase with limestone additions (Clarke, 1993).
Emissions from coal combustion depend a great deal on the process conditions at a given
facility. In addition to the type of boiler, conditions such as temperature, heating rate, exposure
time at elevated temperatures, and whether the environment is oxidizing or reducing can affect
emissions (Pacyna, 1986). For Pb, changes in the temperature affect the size of particles, the
amount of Pb in the vaporized fraction, and the species of the emissions. At combustion
temperatures of 1800 K, about 0.1% of the total ash produced was vaporized (Lockwood and
Yousif, 2000). At 2800 K, the vaporized fraction of the ash was increased to 20%. The ratio of
air to coal during combustion can also have a major effect on emissions (Furimsky, 2000). In a
fluidized bed system, increasing the air to coal ratio from 1.0 to 1.10 decreased the gas to solid
ratio for Pb emissions from 1.5 to 0.18 (Furimsky, 2000).
Uncontrolled combustion of coal can also occur - usually as natural, in-ground coal fires
- and such combustion can emit Pb (Finkelman, 2004). Although such fires may, at times, be
locally importance, they are not discussed in detail here.
Controlled combustion is the norm for industries and utilities. The major pollution
control systems are electrostatic precipitators (ESP), wet scrubbers, and baghouses. In general,
pollution control systems are most effective at removing large particles and are least effective at
removing submicron particles. ESPs are highly efficient and can remove particles with >99.9%
efficiency depending on particle size, ash resistivity, flue gas temperature, and moisture content
(Clarke, 1993). ESPs are used at more than 90% of coal-fired utility boilers in the United States
(Senior et al., 2000). Particles that escape EPSs typically range from 0.1 to 1.0 jim in diameter
(Senior et al., 2000). Wet scrubbers are also more than 99% efficient, with the majority of
particles that escape being <2 jim in size (Pacyna, 1986). Wet scrubbers are used less commonly
than ESPs and baghouses (Senior et al., 2000). Baghouses or fabric filters are frequently used by
2-29
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coal-fired utilities. As with ESPs and wet scrubbers, the collection efficiency of baghouses is a
function of particle size (Senior et al., 2000). Baghouses are >99% effective with mass
emissions averaging <20 mg/m3 (Clarke, 1993).
Little published information is currently available regarding the actual quantity of Pb
emitted from coal-fired boilers. The EPA AP-42 program publishes emission factors for typical
coal-fired boilers, although using process data specific to a given facility is likely to be more
accurate. Clarke (1993) reported emissions from fluidized beds. Of the processes tested, the Pb
emissions were highest from a 0.5 m bed with a limestone sorbent, second highest with a 1.0 m
bed without a limestone sorbent, and lowest with a 0.5 m bed without a limestone sorbent
(Clarke, 1993). Reducing the depth of the fluidized bed by 50% decreased the emissions of trace
elements by ~5 to 50%, probably because deeper beds undergo attrition of ash (Clarke, 1993).
Olmez et al. (1988) reported on Pb mass fractions of PM in a stack of a coal-fired power plant.
For fine particles, Pb constituted 0.04 ± 0.004%, whereas for coarse particles, Pb constituted
0.03 ± 0.002%. Coal combustion products that underwent long-range transport from the
coal-fired power plants of the Midwest contributed an estimated 0.05 |ig/m3 to the ambient air in
Boston (Thurston and Spengler, 1985). Table 2-12 lists the emission factors for three different
types of coal, in three different types of power plants.
Table 2-12. Lead Emission Factors for Coal Combustion in Three Different Furnaces
Rank
Bituminous with
control device
Bituminous without
control device
Subbituminous with
control device
Lignite with control
device
Pulverized coal
Anthracite
Cyclone Furnace
8.5xl(T14kg/J
2.10xl(T4kg/mt
2.18xl(T13kg/J
1.03xl(T13kg/J
1.44xl(T14kg/J
5071b/1012Btu

Stoker Furnace
128xl(T13kg/J
2.18xl(T13kg/J
1.56xl(T13kg/J
2.17xl(T14kg/J

4.45xl(T3kg/mt
Pulverized Furnace Source
5.5xl(T14kg/J Pacyna(1986)
2.10xlCT4kg/mt U.S. EPA (1998)
2.18xl(T13kg/J U.S. EPA (1998)
6.2xl(T14kg/J Pacyna(1986)
92x KT13 kg/J Pacyna (1986)
U.S. EPA (2005)
U.S. EPA (1998)
Source: Pacyna (1986), U.S. Environmental Protection Agency (1998, 2005).
2-30
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The species of Pb emitted from coal depend on process conditions. PbSC>4 was found to
be the dominant Pb compound in flue gas up to 1150 K (Lockwood and Yousif, 2000).
Above this temperature, elemental Pb and PbO, both in the vapor phase, dominate. As the
temperature increases, the equilibrium shifts toward elemental Pb (Lockwood and Yousif, 2000).
In pulverized coal combustion at 1800K, the Pb species found in the gas phase were PbO,
elemental Pb, PbCl, and PbCb (Furimksy, 2000). The solid phase was comprised of PbO,
PbO»SiO2, elemental Pb, and PbO2 (Furimksy, 2000). As the flue gas cools, the Pb composition
changes. PbCb increases and is the main constituent of the gas phase before condensation
occurs at 900K. If low rank low chlorine coal is used, then PbO and elemental Pb will dominate
the gas phase. At 1500K, PbSO4 dominates the particulate phase; at 1800K PbO2 was the
predominant Pb compound in the particulate phase (Furimksy, 2000).
Lead emissions from coal combustion in industrial, commercial, and residential boilers
are similar to the values listed above for utility boilers. Table 2-13 lists emission factors for coal
combustion.
Table 2-13. Lead Emissions from Industrial, Commercial, and
Residential Coal Combustion
Coal-fired unit Emission factor (g/metric ton)
Industrial cyclone boiler 1.2
Industrial stoker boiler 7.7
Industrial pulverized coal boiler 4.5
Commercial/Residential boiler (stoker or hand-fired) 2.7

Note: Data for industrial boilers, assuming 10% ash fraction and 85% efficient control devices.
Source: Pacyna (1986).
Stationary External Combustion: Fuel Oil Combustion
Fuel oil combustion constitutes 15% of fossil fuel energy production in the United States.
(U.S. Environmental Protection Agency, 1998). As with coal, fuel oil is used to generate energy
for utilities, industries, and commercial and residential boilers. The discussion below focuses on
electric power utilities, which are the largest users of fuel oil.
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Fuel oil is generally combusted in tangentially-fired or wall-fired boilers. Emissions of
Pb from oil combustion depend on the process conditions, the amount of Pb in the oil, and the
amount of sulfur in the oil (Pacyna, 1986) (see Table 2-14).
Table 2-14. Lead Emission Factors for Oil-Fired Utility Boilers
Boiler Type
Emission Factor
Control Device
Residual oil-fired boiler, No. 6 oil, normal firing
Residual oil-fired boiler, No. 6 oil, normal firing

Residual oil-fired boiler, No. 6 oil, tangential firing
Residual oil-fired boiler, No. 5 oil, normal firing
Distillate oil grades 1 and 2
Oil-fired utility boiler
Oil-fired utility boiler
4.33xl(T15kg/J
9.35xl(T15kg/J
5.43xl(T15-1.22xl(r14kg/J
4.33xl(T15kg/J
6.89xl(T15kg/J
3.84xl(T15kg/J
2.61b/trillionBtu
9.01b/trillionBtu
None
Flue gas recirculation

None
None
None
PM control
PM/SO2 control
Source: U.S. Environmental Protection Agency (1998).
The Pb concentration in the oil is the most important factor for determining eventual Pb
emissions from combustion. Lead concentration in crude oil range between 0.001 to 0.31 ppm
(Pacyna, 1986). The heavy fractions of crude oil tend to possess higher metal concentrations,
trending to larger metal concentrations with increasing weight. Refining oil removes about 10%
of metals (Pacyna, 1986).
As with coal, process conditions and the presence of pollution control devices greatly
affect the rate and characteristics of emissions from fuel oil combustion. Emissions from
oil-fired boilers depend on the efficiency of combustion and how much deposited material has
built up in the boiler (Pacyna, 1986). Additionally, poor mixing, low flame temperatures, and a
short residence time in the combustion zone cause overall particulate emissions to be greater and
individual particle sizes to be larger (Pacyna, 1986). Oil, which is typically atomized prior to
2-32
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combustion, will emit larger particles and have a higher particulate loading when atomization is
done at low pressures. Conversely, high pressure atomization leads to smaller particles and
lower particulate loadings (Pacyna, 1986). In general, about 90% of PM mass is <2.5 jim in
diameter (Olmez et al., 1988).
Emission factors published in the literature are limited. An average emission factor for
European oil-fired power plants was reported as 126 jig Pb/MJ for oil containing 1% sulfur
(Pacyna, 1986). Lead emissions are higher for oils with greater sulfur contents. Olmez et al.
(1988) report Pb mass fractions for two oil-fired power plants in Philadelphia. Lead was found
to be 1.0% ± 0.2% and 1.8% ± 0.6% in the fine PM fraction in these two plants, respectively, and
0.48% ± 0.2% and 3% ± 0.4% in the coarse fraction. Lead in PM at the Philadelphia plants was
enriched by more than a factor of 1000 compared to the Pb concentration in the fuel oil. Lead in
PM for seven other oil-fired power plants was enriched by more than a factor of 100 (Olmez
et al., 1988). A plant in Boston increased the ambient concentration of fine Pb aerosols by an
estimated 0.05 |ig/m3 and the ambient concentration of coarse Pb aerosols by 0.003 |ig/m3
(Thurston and Spengler, 1985).
The combustion of used oil is also common. About 75% of used oil, which is generated
in the transportation, construction, and industrial sectors, is burned as fuel oil (Boughton and
Horvath, 2004). The Pb concentration of used oils is markedly higher than that of low-sulfur
crude-based heavy fuel oils (Boughton and Horvath, 2004). Emissions from used oil combustion
are estimated at -30 mg of Pb from the combustion of 1 L of used oil. This is 50 to 100 times
higher than emissions from crude-derived fuel oils.
Emission rates for industrial boilers are similar to those of utility boilers. Industrial
oil-fired boilers are not usually equipped with pollution control devices. Approximately 6.4 g of
Pb are emitted per 1000 L of fuel oil burned with a sulfur content of 1% (Pacyna, 1986).
Commercial and residential boilers, which are also not typically equipped with pollution control
devices, have emissions of-3.3 g of Pb emitted per 1000 L of fuel oil (Pacyna, 1986).

Stationary External Combustion: Wood Combustion
Wood-fired boilers are used almost exclusively by industries that produce wood or wood
products. These include pulp and paper mills, lumber production facilities, and furniture
manufacturers (U.S. Environmental Protection Agency, 1998). The materials used as fuel may
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include bark, slabs, logs, cuttings, shavings, pellets, and sawdust. During combustion, elemental
pollutants such as Pb are converted to their oxide forms. These are hydrated and later carbonated
under atmospheric conditions (Demirbas, 2003a).
As with coal and oil, the largest factor affecting emissions from wood combustion is the
concentration of Pb in the fuel. Lead concentrations tend to be very low for virgin wood. The
median Pb concentration in 24 pine and spruce samples was 0.069 ppm (Krook et al., 2004).
The concentrations of Pb in spruce, beech, oak, pine, and ailanthus are listed in Table 2-15.
Table 2-15. Lead Concentrations in Biomass, Char, and Ash Samples
from Spruce, Beech, Oak, Pine, and of Ailanthus Trees
Wood
Spruce trunk wood
Beech trunk wood
Oak trunk wood
Pine trunk wood
Ailanthus trunk wood
Spruce bark
Beech bark
Oak bark
Pine bark
Ailanthus bark
Biomass (ppm)
0.32a
0.36a
0.27a
n.a.
n.a.
0.38a
0.43a
0.31a
n.a.
n.a.
Char (ppm)
2.5a
2.6a
2.1a
n.a.
n.a.
3.r
3.3a
2.5a
n.a.
n.a.
Ash (ppm)
33.2a'b
35.0a'b
28.4a'b
34.9b
32.7b
5.2a, 36.2b
3.8a,40.8b
4.0a, 34.0b
38.7b
35.7b
"Source: Demirbas (2003a).
b Source: Demirbas (2003b).
Waste wood recovered from construction and demolition sites is increasingly used as fuel.
Although most of this wood is untreated, some can have elevated levels of metals from surface
treatment of the wood or industrial preservatives (Krook et al., 2004). In addition, waste wood
commonly contains contaminants such as metal pieces, concrete, stone, gravel, glass, and soil,
which may increase metal emissions during combustion. Lead has been measured in waste wood
at levels -40 times higher than levels found in virgin wood. The median concentration of Pb in
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recovered waste wood in Sweden was 33 ppm (Krook et al., 2004), whereas Pb in recovered
waste wood from Germany and the Netherlands had a median value of 110 ppm.
Emissions of metals from wood are affected by process conditions. Good air-fuel mixing
and high furnace temperatures keep emissions low (Demirbas, 2003a). Additionally, emissions
depend on whether or not the wood was combined with other fuels, the feed rate, the physical
state of the wood, the stack temperature, the geometry of the boiler, which can act as an inertial
particulate collector, the draft setting, and the amount of moisture in the fuel (Demirbas, 2003a;
Pels et al., 1990; Pacyna, 1986).
Pollution control devices may be present with large-scale wood-fired boilers. These can
greatly reduce PM emissions. However, in a wood-burner installation in Ontario, a cyclone was
found to have an efficiency of just 53% for total PM mass (Pels et al., 1990). For particles
<2 jim in diameter, the PM concentrations downstream of the cyclone were actually greater than
those upstream, probably indicating that larger particles were breaking apart during passage
through the cyclone. The emissions of Pb from wood combustion are highly variable. The
emission factor for wet fuel at a large-scale wood burner was 0.0006 g Pb/kg fuel (Pels et al.,
1990). For dry fuel, emission factors were in the range <0.00035 to 0.0014 g Pb/kg fuel burned,
with an average of 0.00056 g Pb/kg fuel (Pels et al., 1990). Emissions from a wood stove and a
fireplace were estimated as 0.007 g Pb and 0.0047 g Pb per kg of wood burned, respectively
(Pacyna, 1986). The recently updated AP-42 emission factor for Pb from wood residue
combustion in boilers is 4.8^10-5 Ib/MMBtu (U.S. Environmental Protection Agency, 2005).
Lead emissions from combustion of waste wood are higher than emissions from
combustion of virgin wood. Although emission factors are not available, the concentration of Pb
in ash from waste wood combustion is elevated above that from the combustion of virgin wood
(Krook et al., 2004).
Data on particle sizes and species of emitted aerosols from wood combustion are not
readily available.

Stationary Combustion Sources: Solid Waste Incineration
The amount of municipal waste incinerated (-15% of waste) has remained stable over the
past decade. In earlier years, municipal waste incineration was an important source of Pb
emissions; and, locally it is still a concern in some places (Walsh et al., 2001). In New York
2-35
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City in the late 1960s, for example, Pb emissions from refuse incineration were between 602 and
827 tons per year, which was about 40 to 50% of Pb emissions from cars that totaled -1752
metric tons in the same area (Walsh et al., 2001). Due to new requirements for the use of
Maximum Achievable Control Technology (MACT), combustion is estimated to be <10 tons per
year nationally.
Incinerator residue is partitioned into bottom ash, fly ash, and flue gas. Here the focus is
on Pb in flue gas, due to its importance in increasing airborne Pb concentrations (Chang et al.,
1999). Lead in incinerator effluents is derived primarily from the noncombustible materials that
end up in refuse (Pacyna, 1986). Under MACT requirements, the most common air pollution
control technology used at municipal waste incinerators is a spray dryer-fabric filter scrubbing
system, enhanced with activated carbon injection.
Factors that affect the quantity of Pb emitted from incinerators include combustion
temperature, the amount of Pb in the refuse, process conditions, moisture content, the addition of
reactive species such as calcium, magnesium, and aluminum, and the addition of sorbents. Of all
these factors, temperature seems to have the greatest impact on metal volatility (Chen and Yang,
1998). Metal volatilization is fast during the initial stages of combustion but levels off after
about 15 minutes (Ho et al., 1993; Chen and Yang, 1998). When plastics only were burned, Pb
volatility was at 18% at 600 °C, 61% at 800 °C, and 91% at 1000 °C (Chen and Yang, 1998).
Figure 2-3 shows the percent volatility for Pb at four different combustion temperatures over
25 minutes of combustion time. Chang et al. (1999) derived the following relationship for Pb
emissions from a fixed bed refuse incinerator in Taiwan:
lnE(wt%) = -3.083J11-"3' + 3.659 (2-4)

where E is the weight percent of Pb in particulate emissions, and T is the combustion
temperature in Kelvin.
The amount of Pb emitted is dependent on the quantity of Pb in refuse. Typical sources of
Pb include paper, inks, cans and other metal scrap, and plastics. For U.S. municipal solid waste,
Pb concentrations vary between 110 and 1500 ppm, with an average of about 330 ppm (Durlak
et al., 1997). Because other countries have very different waste compositions, Pb concentrations
elsewhere can vary greatly.
2-36
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100
90 -

80 -
A
70 -

60 J
50 -

40 -
30 -

20 -

10 -
0
Symbol
•e-
*
4
jf.
Temp.
600'C
800°C
902 ratio from 4:1 to 1:4 increased Pb volatility. An increase in the gas velocity can also
increase Pb emissions, although this is a relatively minor effect (Chang et al., 1999; Chen and
Yang, 1998).
The moisture content in an incinerator can affect the behavior of Pb. At a typical
temperature of 950 °C, decreasing the moisture level from 37% to 5% increased Pb in the fly ash
from 54% to 58% (Durlak et al., 1997). Similarly, decreasing the relative humidity from 60% to
40% at 900 °C increased the Pb volatility from 67% to 76%, respectively (Chen and Yang,
2-37
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1998). In addition to these direct effects, moisture can indirectly affect emissions by altering the
combustion temperature (Durlak et al., 1997).
Additives can reduce metal emissions from incinerators. Additives such as calcium,
magnesium, and aluminum react with metals and bind them. This has been shown to reduce the
formation of metal particles. Adding A1(NC>3)3, for example, reduced quantities of PbCl2 emitted
(Ho et al., 1993). The addition of Ca(OH)2 did not affect volatility at 600 °C (lower limit for
combustion temperature) or at 1000 °C (upper limit for combustion temperature) (Chen and
Yang, 1998). However, Ca(OH)2 did appreciably limit Pb emissions at intermediate
temperatures.
Sorbents can also reduce metal emissions. Sorbents function by binding metal vapors
through heterogeneous chemical absorption and/or condensation before vaporized metals are
able to form particles (Ho et al., 1993). In a fluidized bed incinerator, the efficiency of metal
capture with sorbents varied between 4.9% and 94.5% (Ho et al., 1993), depending on
temperature. Low efficiencies were observed at high and low temperatures, whereas optimal
efficiency was observed in the intermediate range of-600 to 800 C. Limestone was shown to be
a more effective sorbent than sand.
Emissions from refuse incinerators have been reported as 0.018 g of Pb emitted per kg of
refuse, assuming a control device with 85% efficiency (Pacyna, 1986). A source apportionment
study showed that refuse incineration increased the ambient concentration of Pb by an estimated
0.008 |ig/m3 (Thurston and Spengler, 1985). This was observed after incinerators had been
banned in the area, probably indicating prohibited residential refuse combustion. Lead in PM
emissions has been reported to be between 6.9% and 8.9%, with an average of 8.1% (Pacyna,
1986). Three U.S. incinerators had emissions in which Pb constituted 8.2 ± 1.6% of the PM
(Olmezetal., 1988).
Chlorine plays a critical role in determining the speciation of Pb emissions. Lead in the
incineration system exists primarily as chlorine species (either PbCl or PbCl2) (Durlak et al.,
1997). However an increase in moisture content decreases the levels of free chlorine, which has
the subsequent effect of shifting Pb from gaseous PbCl2 to PbO in particulate form. PbCl2(g) is
completely volatilized at 430 °C (Chen and Yang, 1998; Chang et al., 1999). Above 800 C
PbCl2 slowly decomposes and PbO(g) and PbCl(g) are present in greater concentrations.
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The presence of sodium can also affect speciation. Sodium has a greater affinity for
binding with chlorine than Pb (Durlak et al., 1997). Thus, increasing the sodium content
effectively shifts the dominant Pb compound from PbCl2 to PbO. Decreasing the sodium content
from 6560 ppm to 4500 ppm (the average value observed in municipal solid waste) was
responsible for increasing Pb in the fly ash from 35% to 60% at average moisture levels (Durlak
et al., 1997). High sodium concentrations attenuate the influence of moisture on Pb emissions.
Lead emissions tend to concentrate in the submicron size range (Chang et al., 1999;
Olmez et al., 1988). Lead in the fine fraction was enriched by a factor of more than 105 at
several U.S. incinerators compared with the concentration of Pb in the solid waste (Olmez et al.,
1988). Lead in the coarse fraction was enriched by a factor of more than 1000.

Stationary Combustion Sources: Sewage Sludge Combustion
Sewage sludge incinerators exist at approximately 200 sites in the United States (U.S.
Environmental Protection Agency, 1998). Lead can enter the sewage waste stream through car
washes, galvanized material, pipe erosion, pigments, food, processed chemicals, and roofs
(Krook et al., 2004). As in other combustion processes, Pb impurities vaporize during
incineration and then condense.
The Pb content of dry sludge from sewage sludge incinerators varies between 80 and
26,000 ppm, with an average of 1,940 ppm (Pacyna, 1986). The emission factor for Pb from
sewage sludge incineration was 140 g/ton. Sewage sludge cake taken from an industrial
wastewater treatment plant in Taiwan had Pb levels of 1,500 ppm (Chang et al., 1999). Prior to
combustion, Pb is either bound to organic matter in sludge or is present as a carbonate
(Lockwood and Yousif, 2000).
In sewage sludge incinerators, higher temperatures are associated with higher Pb
emissions (Pacyna, 1986). Additionally, sewage sludge incinerators tend to be equipped with
venturi scrubbers with efficiencies of 90 to 99% (Pacyna, 1986). Other pollution control devices
are less common.
Sorbents can be effective pollution controls. Kaolinite, in particular, was shown to reduce
Pb emissions significantly (Lockwood and Yousif, 2000).
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Emissions have been estimated as 0.14 g Pb emitted per kg of sewage sludge combusted
(Pacyna, 1986). The fine fraction of particulate emissions in an experimental setup was enriched
with Pb by a factor of 2.5 (Lockwood and Yousif, 2000).

Stationary Combustion Sources: Scrap Tire Combustion
Waste tires are increasingly used as a fuel, although uncontrolled burning as a result of
accidents or illegal activity is common (U.S. Environmental Protection Agency, 1998). One
analysis showed that uncontrolled combustion resulted in Pb emissions on the order of 0.47 mg
Pb/kg tire for tires that had been cut into four to six pieces (Lemieux and Ryan, 1993).
Emissions were lower for shredded tires, at 0.10 mg Pb/kg tire, probably because of greater
oxygen transport between tire pieces. Another analysis detected trace amounts of Pb in the
smoke from the combustion of tire bodies but did not detect Pb emissions when the tread was
burned (Wagner and Caraballo, 1997).

Lead-acid Battery Manufacturing
Lead-acid batteries constituted 84% of Pb consumed in 2003, as shown in Figure 2-4
(USGS, 2003). Lead-acid batteries are manufactured from Pb alloy ingots and Pb oxide. Lead
alloy ingots are produced by smelters, the emissions of which are characterized earlier in this
chapter. Lead oxide is either produced on-site or is outsourced (U.S. Environmental Protection
Agency, 1998). In 1975, at one facility in Pennsylvania, ambient Pb concentrations ranged from
4.1 to 5.2 |ig/m3 at several sites near the property. Lead acid battery manufacturing contributed
1.7% of Pb emissions in the U.S., or -25 tons, in the 2002 NEI (U.S. Environmental Protection
Agency, 2006a).
Lead-acid battery manufacture consists of the following processes: grid casting or
stamping, paste mixing, plate stacking, group assembly, and battery assembly into the battery
case (U.S. Environmental Protection Agency, 1998). Each process has its own characteristic
emissions of Pb. Emissions from Pb oxide manufacture tend also to be in the form of Pb oxides.
These emissions are typically controlled via a baghouse. The sites of other processes are usually
equipped with baghouses or impingement wet scrubbers (U.S. Environmental Protection
Agency, 1998).
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Cement Manufacturing
The manufacture of Portland cement emits relatively low quantities of Pb. In the 2002
NEI, Portland cement manufacturing resulted in -21 tons of emissions, or about 1.5% of total
U.S. Pb emissions (U.S. Environmental Protection Agency, 2006a). Trace amounts of Pb are
present in the raw materials of calcium, silicon, aluminum, and iron (U.S. Environmental
Protection Agency, 1998). As the raw materials are thermo-treated, most of the Pb is trapped in
the resulting clinker, although some is released as PM (U.S. Environmental Protection Agency,
1998). Additionally, emissions result from the combustion of the coal, natural gas, or waste tires
used to fire the kiln (Pacyna, 1986; U.S. Environmental Protection Agency, 1998).
Emissions are reduced significantly through the use of pollution control devices. ESPs
and baghouses are both common although baghouses tend to be more effective. Lead is present
in the emitted PM in the range of 100 to 1000 ppm (Lee and Von Lehmden, 1973). Emission
factors for cement production are listed in Table 2-16.
Table 2-16. Emission Factors for Processes Used in Cement Manufacture by
Control Device
Pollution Control Device
Process
Dry Process (total)
Kiln/cooler
Dryer/grinder
Wet Process (total)
Kiln/cooler
Dryer/grinder
Multi-cyclones
16.0
12.0
4.0
12.0
10.0
2.0
ESP
4.0
3.0
1.0
3.0
2.5
0.5
Baghouse
0.16
0.12
0.04
0.12
0.10
0.02
Note: Units are g Pb/metric ton cement.
Source: Pacyna (1986).
Glass Manufacturing
The production of leaded glass emits significant quantities of Pb. Its uses primarily
include Pb crystal, cathode ray tubes for televisions, and optical glasses such as binoculars,
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microscopes, and telescopes (U.S. Environmental Protection Agency, 1998). Leaded glass is
composed of silica sand and Pb oxide. Lead oxide concentrations in U.S.-produced leaded glass
typically range between 12% and 60% but can be as high as 92% (U.S. Environmental Protection
Agency, 1998).
The basic process of glass manufacturing includes blending the raw materials, melting,
and forming and finishing. Lead emissions can occur during all of these processes. During
blending, forming, and finishing, Pb is emitted as part of fugitive dust emissions in minor
quantities (Shapilova and Alimova, 2000; U.S. Environmental Protection Agency, 1998).
The major source of Pb emissions derives from the melting process. Emissions from
melting depend mostly on the amount of Pb oxide in the raw material (Shapilova and Alimova,
2000; U.S. Environmental Protection Agency, 1998). Other factors are the type and efficiency of
the furnace, the waste-gas volume, the smoke-flue length, and the efficiency of pollution control
devices (Shapilova and Alimova, 2000). Electric furnaces emit significantly less Pb than
gas-flame furnaces. One analysis found that the rate of Pb emissions from a gas-flame
regenerative furnace was more than seven times higher than the rate of emissions from a deep
tank electric furnace (Shapilova and Alimova, 2000). An electrostatic precipitator is the most
efficient pollution control device for glass manufacturing operations, and ESPs are between 80%
and 90% effective. Wet scrubbers are relatively ineffective. Rates of Pb emissions from several
types of furnaces are listed in Table 2-17.

Table 2-17. Rate of Lead Compound Emissions from Glass-Melting Furnaces
Equipment
Electric tank furnace with gas-heated
working zonea
Electric tank furnace with gas-heated
working zone
Gas-flame potter furnace3
Slag-lining electric furnace with
gas-heated working zone
Product
Glass with 1 6% PbO
Glass with 1 6% PbO
Glass with 16% PbO
Glass with 64.5% PbO
Lead Compound
Emissions g/sec)
0.134
0.002
0.004
0.004
a Fitted with a "cassette pulse filter" designed specifically to capture particulate emissions from small-sized,
glass-melting furnaces.
Source: Shapilova and Alimova (2000).

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Ammunition Production and Shooting Ranges
In 2003, 48,800 metric tons of Pb were consumed in the United States for the production
of ammunition (USGS, 2003). Additionally, some Pb is used to produce Pb azide or Pb
styphnate, which is a detonating agent. Small arms manufacturing plants are likely emitters of
Pb although the actual quantity is unknown.
Shooting ranges, both outdoor and indoor, may have a local impact on airborne Pb
concentrations. Lead is emitted both from cast Pb bullets and Pb-based primers (Gulson et al.,
2002). The propellants contain <2 ppm Pb and seem to have a negligible effect on air Pb
concentrations. A 97% reduction in the air Pb concentrations was observed when Cu-jacketed
bullets replaced cast Pb bullets (Gulson et al., 2002). In comparing the Pb exposure of
personnel, there seems to be little difference between indoor and outdoor firing ranges (Gulson
et al., 2002). One study found that soil Pb concentrations at an outdoor firing range were
elevated by up to 2600 times background concentrations, indicating significant atmospheric
deposition (DeShields et al., 1998).
An additional source of Pb emissions may be explosive ordnance disposal (EOD) (U.S.
Environmental Protection Agency, 1998). Emissions from EOD are either from the combustion
or detonation of the propellant and primer material or from nonenergetic wastes such as
containers and other wastes associated with the propellant (U.S. Environmental Protection
Agency, 1998).

Demolition
A study of Pb dust-fall during the demolition and debris removal of urban row houses
found that Pb was released in very large quantities (Farfel et al., 2003). Many of the row houses
demolished at three sites in Baltimore, MD contained Pb-based paint in addition to being near
sites with elevated levels of Pb in street dust (-700 ppm), sidewalk dust (-2000 ppm), and
residential entryway mat dust (-750 ppm). The results of the study showed that dust fall within
10m of the demolition sites was much higher than baseline measurements and was highly
enriched with Pb (Farfel et al., 2003). The geometric mean Pb dust fall rate increased to 410 jig
Pb/m2/hr during demolition and to 61 jig Pb/m2/hr during debris removal. The baseline rate was
just 10 jig Pb/m2/hr. The Pb concentration in dust fall was 2600 ppm during demolition,
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1500 ppm during debris removal, and 950 ppm at baseline (Farfel et al., 2003). Thus, demolition
and debris removal can be a major source of airborne Pb under some conditions.

Other Stationary Sources of Lead Emissions
There are additional stationary sources of Pb emissions that have not been mentioned
above. Each of these sources are relatively small, but may be an important local source.
Previously unmentioned Pb sources include: medical waste incineration, hazardous waste
incineration, drum and barrel reclamation, crematories, pulp and paper mills, pigment
production, Pb cable coating production, frit manufacturing, ceramics and glaze production, type
metal production, pipe and sheet Pb production, abrasive grain processing, solder manufacturing,
electroplating, resin stabilizer production, asphalt concrete production, paint application, and
rubber production.

2.2.4 Mobile Sources
Automotive Sources of Lead Emissions
Lead is used to manufacture many components in on-road vehicles including the battery,
bearings, paint primers, corrosion-resistant gas tanks, and some plastic and ceramic electrical
components (U.S. Environmental Protection Agency, 1998). The major sources of Pb
emissions—fuel combustion and vehicle wear—are considered below.

Emissions from Combustion of Unleaded Gasoline
Although its phaseout began in 1975, some Pb was still added to gasoline in the United
States as an anti-knock additive at the time of the 1986 Lead AQCD. The ban on Pb additives to
most U.S. motor vehicle gasoline did not fully take effect until 1995, with Pb additives
continuing to be added to piston-engine aircraft fuel and gasoline for some types of race cars.
Since the phaseout of Pb additives from motor vehicle fuel, U.S. airborne-Pb concentrations have
fallen dramatically nationwide. This is considered one of the greatest public and environmental
health successes (Nriagu, 1990). For example, airborne concentrations in the United States fell
an average of 94% from 1983 to 2002 and 57% from 1993 to 2002 (U.S. Environmental
Protection Agency, 2003).
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Most countries have made a similar move away from leaded fuel, but a few continue the
practice of adding tetraethyl Pb to automotive gasoline, particularly most countries in Africa and
some in Asia. Worldwide Pb consumption for gasoline peaked in the 1970s at just under
400,000 metric tons, but by 1993, this value fell to about 70,000 metric tons (Socolow and
Thomas, 1997). Leaded gasoline was the largest source of air emissions throughout the 1970s
and 1980s (Socolow and Thomas, 1997). In Pakistan, a country that continues to use leaded fuel,
the airborne Pb concentrations in the urban center of Karachi range between 2.0 and 19 |ig/m3
(Parekh et al., 2002). This is 2 to 3 orders of magnitude higher than typical urban concentrations
in the United States.
In the absence of tetraethyl Pb additives, Pb is emitted from automobiles as a trace
element in PM. Metals enter the vehicle in trace amounts, naturally occurring in gasoline and
also a component of lubricating oil. The amount of PM that is emitted from the car depends on
a number of variables including the ambient temperature, the cruising speed, the amount of
stop-and-go activity, the type of catalyst, the fuel quality, the phase of driving, and the age, size,
maintenance level, and engine type of the vehicle. EPA promulgated PM standards in 2000 for
light-duty vehicles that cap PM emissions at 0.01 g/mile.
The amount of Pb that naturally occurs in gasoline is -0.00005 g/L (Harris and Davidson,
2005). An estimated 30 to 40% of this Pb deposits in the engine and exhaust system; the balance
is emitted (Huntzicker et al., 1975; Loranger and Zayed, 1994). The deposition of Pb in the
engine and exhaust system varies greatly under different driving conditions, and Pb that has been
retained here is re-entrained into the exhaust under certain transient operating conditions, such as
heavy acceleration. Due to this complex behavior, inventory estimates of Pb from motor
vehicles (e.g., those presented in Table 2-8) are derived from Pb concentrations in fuel and fuel
consumption estimates and not emissions testing data. Data from engine tests of light-duty and
heavy-duty vehicles are discussed below. These data are helpful in understanding the size
fraction for the emitted Pb, but the values reported highlight the variability in emissions and
large uncertainties in measurements made. Lead is also present in lubricating oil, and
concentrations can vary widely. However, since EPA does not regulate Pb in lubricating oil, no
systematic data have been collected or are now available on Pb levels in lubricating oil.
Particulate matter emissions have been shown to be higher in older vehicles than in newer
vehicles (Gillies et al., 2001; Cadle et al., 1999). Gillies et al. (2001) compared emission factors
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from several studies, and found that emission factors from car models between the years 1964
and 1983 had emission factors for PM that were about an order of magnitude higher than models
from the 1990s. This was true even of catalyst-equipped vehicles. Similarly, Cadle et al. (1999),
using unleaded gasoline, tested 195 cars with model years between 1971 and 1996. Their results,
which are listed in Table 2-18, show an increase in Pb emission rates with automobile age.
Table 2-18. Lead Emission Factors During Summer Versus Winter for
Automobiles with Model Years Between 1971 and 1996

Vehicle Category
1991-1996
1986-1990
1981-1985
1971-1980
Smokers
Diesel
Emission
Summer
0.003
0.027
0.006
0.043
0.035
0.15
Factors in mg/mile
Winter
0.019
0.019
0.103
0.222
0.282
0.142
Note: "Diesel" denotes diesel automobiles, "Smokers" denotes automobiles with visible emissions.
Source: Cadle etal. (1999).
Vehicles that have visible tailpipe emissions are known as "smokers." The emissions of
almost all pollutants are elevated from smoking vehicles compared to their non-smoking
counterparts. Emission rates of Pb from smokers are an order of magnitude higher than typical
cars manufactured in the 1990s, as shown in Table 2-18. Interestingly, another study found that
smokers and other high-emitting vehicles emitted more Pb after undergoing repair than before
(Cadle et al., 1997). The emission rate of Pb before repair had an average value of 0.03 mg/mi
with a standard deviation of 0.05 mg/mi. After repair, the emission rate for Pb increased to
0.16 mg/mi with a standard deviation of 0.5 mg/mi. The authors explain this surprising result by
suggesting that either changes in combustion conditions caused elemental deposits from the
engine and exhaust system to be released, or PM deposited during repair and testing was not
removed before emissions testing (Cadle et al., 1997).
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Table 2-18 also shows the effect of the ambient temperature on emission rates of Pb.
Emissions tend to be higher during cold months than during warm months (Cadle et al., 1999).
The rate of emissions is largely dependent on the phase of driving. The Federal Test
Procedure analyzes three phases: cold start, hot stabilized, and hot start, the results of which are
shown in Table 2-19. Driving cycles that are not included are the highway fuel economy test,
and a high speed, high load cycle known as US06 (Cadle et al., 1999). Emissions were
significantly higher during cold start than during the hot stabilized and hot start phases.
Table 2-19. Lead Emission Factors for Different Driving Phases for Automobiles with
Model Years Between 1971 and 1996
Vehicle Category
1991-1996
1986-1990
1981-1985
1971-1980
Smokers
Diesel

Cold Start
0.005
0.041
0.016
0.112
0.116
0.190
Summer Emission Factors in mg/mile
Hot Stabilized
0.002
0.020
0.002
0.015
0.010
0.048

Hot Start
0.002
0.031
0.006
0.044
0.031
0.313
Source: Cadle etal. (1999).
Despite the large variability in Pb emissions, several studies describe average on-road
emission factors for a typical fleet. Sternbeck et al. (2002) measured metal concentrations
in two tunnels in Gothenburg, Sweden. The emission factors subsequently derived were
0.036 ± 0.0077 mg/km per vehicle and 0.035 ± 0.014 mg/km per vehicle for the two tunnels,
respectively. Another tunnel study was performed on a fleet comprised of 97.4% light-duty
vehicles and 2.6% heavy-duty vehicles in the Sepulveda Tunnel in California (Gillies et al.,
2001). The emission factors for Pb were 0.08 mg/km per vehicle and 0.03 mg/km per vehicle in
the PMio and PM2.5 fractions respectively. Lough et al. (2005) analyzed emissions from on-road
vehicles in two tunnels in Milwaukee, Wisconsin. Trucks constituted between 1.5% and 9.4% of
the vehicles, with the balance comprised of passenger cars. Lead emission rates were on the
order of 0.01 mg/km per vehicle and 0.1 mg/km per vehicle in the summer and winter
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respectively. Cadle et al. (1999) analyzed 195 in-use, light-duty vehicles using two
dynamometers. Their results are shown in Tables 2-18 and 2-19. A test on noncatalyst-
equipped, light-duty vehicles found that Pb constituted about 0.03% of the fine particle mass
emitted from these vehicles (Kleeman et al., 2000).
Vehicle-derived Pb seems to have a bimodal size distribution. The submicron mode is
likely the product of combustion or high temperatures, and therefore probably came from the
tailpipe (Lough et al., 2005; Harrison et al., 2003; Abu-Allaban et al., 2003). The coarse mode,
with an approximate size range of 1.0 to 18 jim in diameter, is likely a product of physical
processes such as road dust resuspension and tire or brake wear (Lough et al., 2005;
Abu-Allaban et al., 2003). More than 80% of the airborne Pb particles near a roadway were
-------
Table 2-20. Lead Concentration in Particulate Matter Emissions and Lead Emissions
Factors for Buses and Trucks Fueled with Diesel No. 2 and Jet A Fuel
Concentration of
Fuel and Vehicle Type Pb in PM (%)
Truck, Diesel No. 2
Truck and Bus, Diesel
No. 2
Truck and Bus, Jet A
Bus, Jet A and Diesel No. 2
with paniculate trap
Bus, Jet A with paniculate
trap
Phoenix PM10 study
0.0007
0.0006
0.0010
0.0009
0.0028
0.0147
Uncertainty
(%)
0.0028
0.0025
0.0055
0.0052
0.0132
0.0294
Emission Factor
(mg/km)
0.0053
0.0045
0.0050
0.0016
0.0018
n.a.
Uncertainty
(mg/km)
0.0187
0.0188
0.0214
0.0100
0.0085
n.a.
The results of Chow et al. (1991) on heavy-duty paniculate emissions in Phoenix are listed in the last row
for comparison.
Source: Lowenthal et al. (1994).
derived in other studies. Estimates of Pb emissions from brake pads in Sweden were just under
200 jig/km per vehicle (Sternbeck et al., 2002). This is an order of magnitude higher than the
tailpipe emissions measured by Cadle et al. (1999).
Up to 35% of brake pad mass loss is emitted as airborne PM (Garg et al., 2000). One
study that analyzed particulate emissions from seven different brake pad formulations found that
only one type of brake pad described as "potassium titanate, aramid, and copper fiber" emitted
PM with a measurable Pb fraction (Garg et al., 2000).
A joint study in Reno, NV and Durham/Research Triangle Park, NC found that the
dominant contributors to roadside PM were resuspended road dust and tailpipe emissions
(Abu-Allaban et al., 2003). However, brake wear was a significant source of PM in places where
strong braking occurred, such as at freeway exits (Abu-Allaban et al., 2003).
Particulate matter emissions from brake pads were primarily in the fine fraction.
Eighty-six percent and 63% of airborne PM was smaller than 10 and 2.5 |im, respectively (Garg
et al., 2000). It is expected that Pb particles from mechanical processes such as brake wear
would be in the coarse fraction. However, smaller particles may be observed if Pb is vaporized
from hot brake surfaces (Harrison et al., 2003; Lough et al., 2005).
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Lead weights used to balance vehicle wheels may be an additional source of elevated
roadside Pb concentrations. Deposition of Pb from wheel weights at one intersection in
Albuquerque, NM was estimated to be between 50 and 70 kg/km per year (Root, 2000). Wheel
weights are 95% Pb, 5% antimony, and typically weigh between 7 and 113 grams. These wheel
weights can become dislodged during quick stops. Although deposited pieces of wheel weights
are quite large, Pb is very malleable and can be worn away into respirable particles by being run
over by vehicles (Root, 2000).

Emissions from Racing Vehicles
Vehicles used in racing (including cars, trucks, and boats) are not regulated by the EPA
according to the Clean Air Act and can therefore use alkyl-lead additives to boost octane. Data
on Pb levels in racing fuel and rates of Pb emissions are scarce. The U.S. Department of Energy
stopped tracking information on the production of leaded gasoline for non-aviation use in 1990
(U.S. Environmental Protection Agency, 2002). However, the National Motor Sports Council
reports that -100,000 gallons of leaded gasoline were used by National Association for Stock
Car Automobile Racing (NASCAR) vehicles in 1998 (U.S. Environmental Protection Agency,
2002).
As was the case with on-road emissions during the time of universal leaded gasoline use,
the combustion of racing fuel likely elevates airborne Pb concentrations in the nearby area. This
may pose some health risk for some subpopulations, such as residents living in the vicinity of
racetracks, fuel attendants, racing crew and staff, and spectators. However, EPA has formed a
voluntary partnership with NASCAR with the goal of permanently removing alkyl-Pb from
racing fuels used in the Busch, Winston Cup, and Craftsman Truck Series (U.S. Environmental
Protection Agency, 2002). In January of 2006, NASCAR agreed to switch to unleaded fuel in its
racecars and trucks beginning in 2008.
In addition to racing vehicles and piston engine aircraft, legally permitted uses of leaded
fuel include construction machinery, agricultural equipment, logging equipment, industrial and
light commercial equipment, airport service equipment, lawn and garden equipment, and
recreation equipment, including boats, ATVs, jet skis, snowmobiles, etc. (U.S. Environmental
Protection Agency, 2000). Given the relative unavailability of leaded fuel, it is unlikely that it is
commonly used for any of these purposes other than racing vehicles.
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Emissions from the combustion of leaded fuel are generally in the form of submicron
particles of inorganic Pb halides.

Aircraft
Piston-engine aircraft use leaded fuel. Aviation fuel, or "avgas," contains 0.1 to 1.0 g of
tetraethyl Pb additives per liter. About 32.7% of general aviation aircraft use avgas, and the
remainder use jet fuel, which does not contain Pb additives (U.S. FAA, 1996). The overall
fraction of aviation fuel containing Pb additives is unknown.
In the South Coast Air Basin of California, emissions of Pb from general aviation aircraft
have been estimated at 634 ±110 kg/year (Harris and Davidson, 2005). This corresponds to
0.54 grams of Pb released per flight. Approximately 267 kg of the total was emitted below the
mixing height in 2001, which could be a local source of Pb exposure.
Commercial jet aircraft do not use leaded fuel. However, they are also likely sources of
some Pb emissions. In-flight sampling of contrails from a DC-8 and a 757 showed that metals
constituted more than 11% and 5.2% of PM, respectively (Twohy and Gandrud, 1998). This is a
lower limit for the fraction of metals in emissions since almost half of the particles in contrails
are from the ambient air (Twohy and Gandrud, 1998). No known estimates have been made of
the quantity of Pb in commercial aircraft emissions. However, the dominant metals seem to be
Fe, Cr, and Ni (Karcher, 1999). These are the primary components of stainless steel and indicate
that aircraft engine erosion may be a significant source of metal emissions (Karcher, 1999).
Metal particles in contrails have two modes. One is submicron with an average diameter
of about 0.36 jim (Karcher, 1999; Twohy and Gandrud, 1998), whereas the other larger mode is
~1 |im in diameter and has a morphology that suggests mechanical generation (Karcher, 1999).

Lawn-Care Equipment
A life cycle assessment used to compare gasoline-, electricity-, and battery-powered lawn
mowers found that electricity-powered mowers had the fewest overall emissions over their
lifetime (Sivaraman and Lindner, 2004). Battery-powered mowers are fitted with a lead-acid
battery. The total amount of Pb released to the environment a typical Pb-acid battery over its
lifetime averages -0.052 kg Pb, which includes consideration of raw material extraction and
refining, energy production, Pb mining and refining, battery manufacture, and battery recycling
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(Sivaraman and Lindner, 2004). Electricity-powered lawn mowers presumably emit less PM and
Pb than gasoline-powered mowers. This is a reasonable assumption, since utility generation
plants tend to be fitted with pollution control devices and internal combustion engines of
gasoline-powered mowers do not.

Other Sources of Lead Emissions
Lead emissions are associated with the combustion of any fossil fuel. Thus, any of the
following may be additional types of mobile Pb emission sources not separately addressed
above: construction equipment, off-road recreational vehicles, generators, marine vessels,
railroad locomotives, agricultural equipment, logging equipment. However, detailed data on Pb
emissions from such sources are not readily available.
Additionally, the resuspension of Pb-contaminated soil and dust is a major source of
airborne Pb. Since fugitive dust emissions are not considered a primary source of airborne Pb a
discussion of resuspended soil particles is omitted here and covered in Section 2.3.3 as a mode of
transport for Pb particles through the environment.
2.3 TRANSPORT WITHIN THE ENVIRONMENT
2.3.1 Atmospheric Transport of Lead Particles
Atmospheric Dispersion
The atmosphere is the major environmental transport pathway for anthropogenic Pb
(Reuer and Weiss, 2002). Airborne Pb tends to be mainly in the form of submicron aerosols
(Davidson and Rabinowitz, 1992; Davidson and Osborn, 1986; Harrison, 1986; Lin et al., 1993).
The mass median diameter averaged for several studies is 0.55 jim (Milford and Davidson,
1985). A study performed in 1991, after leaded gasoline was no longer the predominant source
of Pb in the atmosphere, showed a bimodal distribution for Pb particles, with the larger peak in
the fine fraction (Lin et al., 1993). The mass median diameter for Pb samples was 0.38 ±
0.06 |im in the fine fraction and 8.3 ± 0.6 jim in the coarse fraction. Since small particles are
much slower to deposit than larger particles, Pb can be transported great distances in the
atmosphere. Detectable quantities of Pb have been found even in the most remote places on
earth. Because much of the airborne Pb is generally associated with fine particles, atmospheric
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dispersion models used for gaseous pollutants can be applied to estimate atmospheric flows of Pb
under certain conditions. Use of such dispersion models is more accurate for submicron Pb
emitted from stacks than it is for larger particles resulting from fugitive emissions, such as
resuspended soil particles.
Airborne concentrations of species emitted from a point source are frequently described
by a Gaussian distribution. This simple description holds true only when turbulence is stationary
and homogeneous. However, the Gaussian model can be modified to account for more complex
atmospheric conditions. For a thorough discussion of assorted Gaussian plume models and
parameters, see Seinfeld and Pandis (1998). Gaussian models are, in general, reasonably
accurate for small-scale work, i.e., within -100 km of the source.
The rate and direction of dispersion are dependent both on pollutant characteristics and
meteorological conditions. Important meteorological factors include windspeed, surface
roughness, inversion frequency, inversion duration, and temperature.
A Gaussian dispersion model (EMITEA-AIR) was applied to theoretical primary and
secondary Pb smelters in Europe (Baldasano et al., 1997). This model accounts for plume rise,
as well as interactions between the plume and terrain. Two sites were modeled. Conditions in
Copenhagen, Denmark included flat terrain, dominant strong winds, neutral or stable turbulence,
and an annual mean temperature of 10°C. Conditions in Catalunya, Spain had a complex terrain,
weak winds, unstable turbulence, and an annual mean temperature of 15°C.
The results of these modeling efforts showed that airborne Pb concentrations were both
lower and more symmetric surrounding the Copenhagen site than surrounding the Catalunya site
(Baldasano et al., 1997). Concentrations at the Copenhagen site had a maximum value of
0.004 |ig/m3. Concentrations at the Catalunya site ranged between 0.065 and 0.3 |ig/m3. The
prevalence of calm winds and the complex terrain were the most important factors contributing
to high Pb concentrations surrounding the Catalunya smelter.
Modeling efforts for an abandoned battery recycling facility using the EPA Industrial
Source Complex Short Term (ISCST) model, based on Gaussian equations, showed good
agreement with measured airborne Pb concentrations (Small et al., 1995). Model predictions at
three sites at distances between 240 and 310m from the stack were between 3.8 and 4.4 |ig/m3,
whereas measured Pb concentrations taken when the plant was in full operation had averages
between 4.1 and 5.2 |ig/m3.
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For long-range transport modeling, Lagrangian trajectory or Eulerian grid models are
commonly employed. These models determine how a parcel of air moves relative to the moving
fluid and a fixed coordinate system, respectively. Two Lagrangian experiments were performed
in the Azores in the northern Atlantic (Veron and Church, 1997). Retrospective air mass
trajectories based on the hybrid single-particle Lagrangian integrated trajectory (HY-SPLIT)
model found that air masses enriched with Pb had been over continental regions ten days prior to
testing. This is consistent with current understanding that most Pb emissions are from sources on
continents, not from oceanic sources. Airborne Pb at this remote location was transported from
several different countries (Veron and Church, 1997).
Similarly, backward air mass trajectories estimated for Greenland showed that the highest
air concentrations of metals were in air parcels that had been over continental regions five days
earlier (Davidson et al., 1993). The model used in this study employed a constant acceleration
formulation of the trajectory equations and encompassed air parcel movements affected by
terrain and meteorology. The air masses with the highest metal concentrations were traced back
to polluted regions, including the Arctic Basin, eastern North America, and Western Europe
(Davidson et al., 1993).
A numerical model that combined weather system modeling with three-dimensional
Lagrangian transport and diffusion modeling was used to determine the foreign contributions of
Pb to airborne concentrations in Israel (Erel et al., 2002). These predictions, in conjunction with
isotopic measurements, indicated that Israel received significant amounts of Pb from Egypt,
North Africa, the United Arab Emirates, Jordan, Turkey, and Eastern Europe (Erel et al., 2002).

Historical Records of Atmospheric Lead Transport and Deposition
An important field of research involves analyzing natural records of Pb deposited from the
atmosphere. Lead concentrations are measured in media such as soil, sediments, ocean water,
peat bogs, plants, snowpacks, or ice cores. Based on concentrations, ratios to other pollutants, or
isotopic compositions, an airborne concentration is back calculated and, in some cases, the major
emitters can be identified. Sediments can provide records dating back several million years, peat
bogs can reach back to the late glacial period (-15000 years ago), corals and trees can record up
to several hundred years, and lichens and mosses can provide recent deposition data (Weiss
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et al., 1999). Additionally, some applications can yield data showing variation with seasons or
climate. These methods have been used to monitor both short and long-range transport.
Many studies have shown a pattern of sediment Pb concentrations increasing to reach
peak concentrations in layers representing deposition during the 1970's followed by marked
declines in more recent years. For example, Figure 2-4 presents data on Pb concentrations in
sediment samples from 12 lakes in the Great Lakes area (Yohn et al., 2004). Other such studies
have been conducted in the Okefenokee Swamp in Georgia (Jackson et al., 2004), Lake Clair in
Quebec (Ndzangou et al., 2005), in two ponds near a Superfund site in Massachusetts (Norton
et al., 2004), lake sediments near Sudbury, Ontario (Belzile et al., 2004), several Canadian shield
lakes (Gallon et al., 2004; Gallon et al., 2006), and several lakes in Scotland (Eades et al., 2002).
A similar pattern has been seen in peat cores from three Southern Ontario bogs (Givelet et al.,
2003) and two peat bogs in Spain (Cortizas et al., 2002), and peat deposits in Switzerland
(Shotyk, 2002). Nieminen et al. (2002) show a pattern of declining Pb concentrations with
greater height in peat cores at a "background" site, but markedly higher Pb concentrations in peat
cores collected near a Cu/Ni smelter in Southwest Finland. For a comprehensive look at natural
historical records, the reader is referred to review articles by Weiss et al. (1999), Boutron et al.
(1994), and Garty (2001).

2.3.2 Deposition of Airborne Particles
Deposition (both dry and wet) is the major removal mechanism for atmospheric pollutants
and atmospheric deposition can be a major source of Pb into lakes (Balistrieri et al., 1995). Here
the main focus is on deposition data published specifically for Pb aerosols, although the literature
on particle deposition is extensive.

Dry Deposition
Dry deposition is the process by which pollutants are removed from the atmosphere in the
absence of precipitation. The downward flux, -F, is characterized by:

-F = VdC (2-5)
where C is the airborne concentration in |ig/m3 and Vd is the deposition velocity in m/second.
The deposition velocity is an empirical quantity defined by Equation 2-5 as the ratio of F to C

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1995
Littlefield*
•Elk
Higgins
Mullett
Crystal M
Gratiot
Cadillac
Paw Paw
Cass
•Gull
• Crystal B
Whilmore
1815
0.00 0.20 0.40 0.60 0.80 1.00 1.20
Lead sediment concentrations normalized to peak
Figure 2-4. Lead concentrations in sediment samples in 12 Michigan lakes. The
concentrations are normalized by the peak Pb concentration in each lake;
peak Pb concentrations ranged from approximately 50 to 300 mg/kg.
Source: Yohn et al. (2004).
with units of m/s. It should be noted that both the airborne concentration and the deposition
velocity are dependent on vertical height.
The physical factors governing dry deposition are often described in a manner analogous
to electronic resistances (Davidson and Wu, 1990). The parameters of aerodynamic resistance,
boundary layer resistance, and surface resistance run in parallel with sedimentation resistance or
gravity. The relative importance of each of these resistances varies with particle size and
meteorological conditions (Wu et al., 1992a).
The size of depositing particles is arguably the most important factor affecting deposition
rates. For very small particles, Brownian motion is the dominant mechanism that transports
particles through the viscous sublayer that borders surfaces (Nicholson, 1988a). For large
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particles, sedimentation is the most important process governing particle deposition.
For intermediate particles, impaction and interception largely determine deposition rates.
The deposition velocity has the most uncertainty for these intermediate sized particles
(Nicholson, 1988a). Although most of the airborne Pb mass was associated with submicron
particles, only about 0.5% of the Pb particle mass undergoing dry deposition in Chicago was
<2.5 |im in diameter (Lin et al., 1993). Additionally, more than 90% of Pb particle mass that
undergoes dry deposition is in an insoluble chemical form (Gatz and Chu, 1986).
Deposition velocities for Pb are in the range of 0.05 to 1.3 cm/s. Table 2-21 is a
compilation of data from the literature. Figure 2-5 shows the variation of deposition velocity
for Pb as a function of particle size. Dry deposition flux values have been reported for several
studies. An estimated flux of 370 to 1000 |ig/m2-year (-0.37-1.0 mg/m2-year) was based on data
collected from two sites on Chesapeake Bay during 1990-1991 (Wu et al., 1994). The mean air
Pb concentration associated with PM2.5 over an estuary in the New York-New Jersey Harbor
Bight area was 4.9 ng/m3, and dry deposition flux associated with PM2.5 was estimated to be
0.15-0.76 mg/m2-year (Gao et al., 2002). Dry deposition flux at several sites near the Lake
Michigan ranged from 0.023 to 0.038 mg/m2-day (8.4-14 mg/m2-year) using data collected in
1993-1995 (Yi et al., 2001). In seven urban sites across the metropolitan Detroit area, estimates
of dry deposition flux rates decreased by about an order of magnitude from 1982 to 1991, from
nearly 10 to below 1 g/km2-day (nearly 4 to below 0.4 mg/m2-year) (Pirrone et al., 1995). Dry
deposition flux of 7.0 |ig/m2-day (2.6 mg/m2-year) was estimated over the Ligurian Sea (Migon
et al., 1997). Total deposition flux of Pb (in fine and coarse fraction particles) into Lake Erie
was 43.3 ng/m2-hour in April 1992 (Keeler and Pirrone, 1996). In an industrial area of northern
France, dry deposition of Pb was estimated to be 40 to 80 |ig/m2-hour (Franssens et al., 2004).
Based on Pb in both fine and coarse particles, dry deposition was estimated to be 920 |ig/m2-year
(0.920 mg/m2-year) in Lake Superior, 950 |ig/m2-year (0.950 mg/m2-year) in Lake Michigan,
and 780 |ig/m2-year (0.780 mg/m2-year) in Lake Erie (Sweet et al., 1998).

Wet Deposition
Wet deposition is the process by which airborne pollutants are scavenged by precipitation
and removed from the atmosphere. The flux of a depositing species can be defined through the
following equation:
2-57
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Table 2-21. Dry Deposition Velocities for Lead Particles
Vd (cm/s)
0.26
0.56
0.06 ±0.02
0.06 ±0.02
0.09 ±0.03
0.26
0.14±0.13
0.15 ±0.07
0.41
0.43
^ 0.19
o° 0.33 ±0.03
0.31 ±0.02
0.37 ±0.04
0.31 ±0.02
0.28 ±0.05
0.34 ±0.05
0.9 ±0.3
1.3±0.5
0.05
0.005
0.06 ±0.01
0.46
0.06
0.13
0.27-0.74
MMAD
(jim)
all
all
all
all
all
all
10% >4
0.87
0.68
0.75
0.70

0.79
0.79
0.5
0.5
0.5

0.3
82%<1

Surface
water
orchard grass

all
teflon plates
teflon plates
water
water
land
alfalfa + oil
alfalfa + oil
grass + oil
grass + oil
soil
soil
beech canopy
spruce canopy
polyethylene petri dish
oak
polyethylene petri dish
filter paper
bucket
water
water
Other

model of Rojas et al. (1993)
model of Slinn and Slinn (1 980)
model of Williams (1982)
mass balance model

stable conditions
unstable conditions
stable conditions
unstable conditions
stable conditions
unstable conditions
throughfall
throughfall

foliar extraction

aerometric mass balance

Reference
Davidson and Rabinowitz (1992)
Davidson and Rabinowitz (1992)
Rojas etal. (1993)
Rojas etal. (1993)
Rojas etal. (1993)
Friedlander etal. (1986)
Davidson and Wu( 1990)
Davidson et al. (1985)
Davidson and Wu (1990)
Davidson and Wu( 1990)
Davidson and Wu ( 1 990)
El-Shobokshy (1985)
El-Shobokshy (1985)
El-Shobokshy (1985)
El-Shobokshy (1985)
El-Shobokshy (1985)
El-Shobokshy (1985)
Davidson and Wu ( 1 990)
Davidson and Wu (1990)
Davidson and Wu( 1990)
Davidson and Wu( 1990)
Davidson and Wu (1990)
Davidson and Wu( 1990)
Davidson and Wu( 1990)
Davidson and Wu( 1990)
Wu etal. (1994)

taken from Dedeurwaerder et al. (1983)
taken from Dedeurwaerder et al. (1983)
taken from Dedeurwaerder et al. (1983)

taken from Hofken et al. (1983)
taken from Hofken et al. (1983)
taken from Lindberg and Harriss (1981)
taken from Lindberg and Harriss (1981)
taken from Lindberg and Harriss (1981)
taken from Pattenden et al. (1982)
taken from Rohbock (1982)
taken from Sievering et al. (1979)

Source: Davidson and Rabinowitz (1992), Rojas et al. (1993), Friedlander et al. (1986), Davidson and Wu (1990), Davidson et al. (1985), and El-Shobokshy (1985), Wu et al. (1994).
-------
Deposition Velocity
ilUUU
1/j 1000
E
u
>*
o 0.100
o
0)
c
5 0.010
v>
0
Q.
Q)
Q
n nn-i
I
a
-s
-\
i
. *
- '
„
: .
_
r
-
~

f "'*
/

k
\
X
^

^
\
1 ^

^
-p /
/
1 i
/ v
/
^ Density = 6 g cm"3
*~ ~™
/

i (MINI I I I I I i II 1 1 1 1 1 1 1
0.01 0.10 1.00 8.00
Mean Stokes Diameter ((jm)
Figure 2-5. The deposition velocity plotted against the geometric mean Stokes diameter
for particles with a density of 6 g/cm~3 (i.e., lead). Error bars are shown and
the arrow indicates a negative value for the lower limit of uncertainty.

Source: Reprinted from Main and Friedlander (1990).
F = VPCP
(2-6)
where Vp is the rate of precipitation in cm/s and Cp is the concentration of the chemical species
in the precipitation in |ig/L (Miller and Friedland, 1994).
The size of particles can influence wet deposition rates. Large particles are scavenged
more efficiently. Lead, which is found primarily in the submicron size range, does not undergo
wet deposition as easily as many of the crustal elements (Davidson and Rabinowitz, 1992).
Conko et al. (2004) note a seasonal trend in wet deposition rates for Pb in Reston, VA.
The annual deposition rate was 440 |ig/m2-year. The highest concentrations were observed in
the summer months, which the authors attribute to increased emissions from electric power
plants. Wet depositional flux of Pb in 1991-1996 was found to be 51 ng/cm2-year
(510 |ig/m2-year) at a site along the Chesapeake Bay, 39 ng/cm2-year (390 |ig/m2-year) at a site
at the mouth of Delaware Bay (Kim et al., 2000), 3.1 |imol/m2-year (642 |ig/m2-year) in two
2-59
-------
streams in western Maryland in 1997-1998 (Lawson and Mason, 2001), and ranged from -300 to
600 |ig/m2-year at four sites in North-central Maryland (Scudlark et al., 2005). Concentrations
of Pb in rainwater in Massif Central in France ranged from 1.30 to 465 |ig/L in 1994-1995 with a
mean of 50.2 |ig/L and an estimated annual flux of Pb from rainwater of 15.4 mg/m2-year (Roy
and Negrel, 2001). In Lake Superior, Lake Michigan, and Lake Erie, respectively, Sweet et al.
(1998) estimated wet deposition rates of 550 |ig/m2-year, 640 |ig/m2-year, and 1000 |ig/m2-year.
Precipitation activity has been linked to variability in wet deposition rates. Intense rain
showers had lower Pb concentrations than slow, even rainfalls (Chow, 1978). Thunderstorms
typically did not have detectable quantities of Pb but occasionally produced very high levels.
The Pb concentration in rainfall does not appear to be correlated to the amount of time between
rainfalls, but meteorological conditions such as a thermal inversion preceding a rainfall may
affect the Pb content (Chow, 1978).
Lead in rainwater includes both dissolved and particulate material. Approximately 83%
of Pb in wet deposition samples was in a soluble form, compared to less than 10% in dry
deposition samples (Gatz and Chu, 1986).
Typical Pb concentrations in precipitation are listed in Table 2-22. The table shows a
pronounced downward trend with time presumably due to the phaseout of leaded fuel. A trend
of reduced Pb concentrations in precipitation was shown in data from Lewes, DE, where average
concentrations declined ~3 |ig/L in 1982 to <1.0 |ig/L in 1989 (Scudlark et al., 1994).

Bulk Deposition
Bulk deposition is the rate of dry and wet deposition combined. It is typically sampled in
open buckets or other open containers. This is often used to estimate the overall rate of
atmospheric input to soil, surface water, or other terrestrial media. However, it is understood
that dry deposition onto surrogate surfaces may differ greatly from dry deposition onto natural
surfaces. Gelinas and Schmit (1998) reported a bulk (wet and dry) deposition rate of
2.61 mg/m2-year in an agricultural area near Montreal in 1993-1995. In this study, the maximum
fluxes were observed in the spring and fall for PM (due to agricultural tilling practices) and
lowest rates were seen in winter and summer. The authors suspect that this is due to decreased
traffic in the summer months. The average deposition rates (wet + dry) estimated for Pb in the
Massachusetts Bay in 1992-1993 were 2700 |ig/m2-year (Golomb et al., 1997). Total (dry and
2-60
-------
Table 2-22. Lead Concentrations in Rainwater in the United States
Dates of
Testing
1966-1967
1971-1972
1975-1976
1977-1978
1982
pre-1982
pre-1982
pre-1982
1982
1988-1989
1989
1993-1994
1993-1994
1993-1994
1998
1998
1998
Precipitation
concentration
(Mg/L)
32.7
31.2
25.2
15.6
17.0
44
5.4-147
12
0.59-64
0.09
0.02-0.41
~3
1.9
~1
0.7 ±0.4
0.9 ±0.6
1.1 ±0.8
0.47 ±0.55
0.76
0.54
Cloudwater
concentration
Oig/L)
98.1
93.6
75.6
46.8
51
n.a.
n.a.
n.a.
n.a.
5.4
n.a.
n.a.
n.a.
n.a.
n.a.
0.58
5.45
Location
Northeastern US
Northeastern US
Northeastern US
Northeastern US
Northeastern US
Urban areas
Rural areas
Remote areas
Lewes, DE
Northeastern US
Lewes, DE
Lake Superior
Lake Michigan
Lake Erie
Reston, VA
Mt. Mansfield, VT
Mt. Mansfield, VT
Source
Lazrusetal. (1970)a
Schlesinger and Reiners (1974)a
Smith and Siccama (1981)a
Smith and Siccama (1981)a
Scherbatskoy and Bliss (1984)a
Galloway etal. (1982)b
Galloway etal. (1982)b
Galloway etal. (1982)b
Scudlark etal. (1994)
Miller and Friedland (1991)a
Scudlark etal. (1994)
Sweet etal. (1998)
Sweet etal. (1998)
Sweet etal. (1998)
Conko et al. (2004)
Lawson et al. (2003)
Malcolm et al. (2003)
Source: a Cited in Miller and Friedland (1994), b cited in Davidson and Rabinowitz (1992), and Conko et al. (2004).
wet) deposition flux to the Rouge River watershed was estimated to be 3167 g/km-year
(3167 |ig/m2-year) from 1982 to 1992; the deposition flux rate declined sharply across this time
period (Pirrone and Keeler, 1996). In an area influenced by many stationary sources as well as
motor vehicle emissions in Paris, France, deposition flux was 39 mg/m2-year in 1994-1995,
9.5 mg/m2-year in 1999-2000, and slightly lower in 2001-2002 (Azimi et al., 2005).
The ratio of dry to wet deposition has been estimated to be 1.5, 0.4, and 0.25 in marine,
rural, and urban areas respectively (Galloway et al., 1982). The estimated ratio of dry deposition
to wet deposition ranged between 0.1 and 0.5 in arctic regions (Davidson and Rabinowitz, 1992).
2-61
-------
Lindberg et al. (1982) reported a ratio of 0.8 for wet to dry deposition in the Tennessee Valley.
In a literature survey, Hicks (1986) found that this ratio varied between 0.4 and 1.8.
Deposition can be influenced by the extent of vegetation cover. Scudlark et al. (2005)
reported that dry deposition comprises <50% of total atmospheric Pb deposition in the
Chesapeake Bay area, with the greatest contribution from wet deposition in winter during
reduced leaf cover.

2.3.3 Resuspension of Lead-Containing Soil and Dust Particles
The resuspension of soil-bound Pb particles and contaminated road dust can be a
significant source of airborne Pb in areas near major sources of Pb emissions. Resuspension by
wind and vehicular traffic is emphasized here, although resuspension through other mechanical
processes (such as pedestrian traffic, agricultural operations, construction, and even raindrop
impaction) is possible. In general, mechanical stresses are more effective at resuspending
particles than wind (Sehmel, 1980; Nicholson, 1988b).
Rapid calculations of ambient, respirable concentrations of Pb from resuspension can be
performed through the use of fugitive dust emission factors. The emission rate of a pollutant as
PMio can be estimated through the following equation (Cowherd et al., 1985):

RIO = a Eio A (2-7)

where RIO is the emission rate of a contaminant as PMio (units of mass/time), a is the fraction of
contaminant in the PMio size range (mass/mass), EIO is the PMio emission factor (mass/source
extent), and A is the source extent (in source dependent units, which are typically area but can
be volume).
Emission factors for fugitive dust depend on whether the predominant force of
resuspension is traffic or other mechanical disturbance, or wind. Emission factors are not
recommended for detailed calculations but can provide order of magnitude assessments with
minimal effort. Condition specific equations for fugitive dust emissions are provided by
Cowherd et al. (1985) and AP-42 (U.S. Environmental Protection Agency, 2005).
Understanding the physics of resuspension from natural winds requires analyzing the
wind stresses on individual particles, including frictional drag, form drag, gravitation, and the
2-62
-------
Bernoulli effect (Sehmel, 1980). Although this analysis can be accurate on a small scale,
predicting resuspension on a large scale generally focuses on empirical data for continual soil
movement due to three processes: saltation, surface creep, and suspension (Sehmel, 1980;
Nicholson, 1988b). Saltation is the process by which particles in the 100 to 500 |im size range
bounce or jump close to the surface. The low angle at which these particles strike the surface
can transfer momentum to smaller particles allowing them to be suspended into the atmosphere
(Sehmel, 1980; Nicholson, 1988b). Depending on soil conditions, saltation can be responsible
for moving 50 to 75% of surface particles. Surface creep is the rolling or sliding motion of
particles, which is induced by wind stress or momentum exchanged from other moving particles.
This generally applies to large particles 500 to 1000 jim in diameter and moves 5 to 25% of soil
by weight (Sehmel, 1980; Nicholson, 1988b). Suspension is the process that actually ejects
particles into the air. This affects particles smaller than 100 |im in diameter and moves 3 to 40%
of soil by weight (Sehmel, 1980; Nicholson, 1988b).
Resuspension is often defined in terms of a resuspension factor, K, with units of m"1, or a
resuspension rate (A), with units of sec"1. The resuspension factor was used in early research on
reentrainment and is defined by:

where Cair is the airborne concentration of a chemical species and Csoii is the surface soil
concentration of the same species. K has significant limitations, in that it is dependent both on
the height at which Ca;r is measured and the depth to which Cso;i is measured. This factor also
assumes that all airborne material is a direct result of resuspended soil-bound material, which is
not the case in most situations (Sehmel, 1980; Nicholson, 1988b). Additionally, K cannot be
used if soil concentrations are not uniform across the area of interest (Nicholson, 1988b).
The resuspension rate, A, is the fraction of a surface contaminant that is released per time
and is defined by:
A= (2-9)
^
2-63
-------
where R is the upward resuspension flux, and A has units of s"1. Although A is also dependent
on the depth to which soil concentrations are measured, the resuspension rate has a number of
advantages over K. Most notably, it can be applied to non-uniform areas of soil contamination,
and it allows for other sources of airborne contaminants. It cannot be determined experimentally
and is usually deduced by fitting results to a numerical model of airborne dispersion and
deposition for the pollutant of interest (Nicholson, 1988b). Resuspension rates are dependent on
many factors, including wind speed, soil moisture, particle sizes, the presence of saltating
particles, and the presence of vegetation. Typical values for A can cover 9 orders of magnitude
in the range of 10'12-10'4 s'1 (Sehmel, 1980; Nicholson, 1988b).
Nicholson (1993) notes that A increases with increasing particle diameter because larger
particles protrude faster into the turbulent air stream and the drag force increases more quickly
than adhesive forces. Furthermore, in a laboratory resuspension chamber, the yields of
resuspended matter decreased approximately linearly with increases in the geometric mean
particle sizes of the bulk soil (Young et al., 2002). Lead is associated with the smaller size
ranges in the distribution of soil particles (Van Borm et al., 1988). Young et al. (2002) suggest
that this is because the higher specific surface area of small particles means that there are higher
contents of organic matter or Fe/Al oxides that serve as Pb binding sites.
Saltation is a particularly important factor in determining resuspension rates. Saltation
moves large quantities of soil particles and is highly efficient at ejecting particles into the
airstream. Saltating particles rotate between 200 and 1000 revolutions/second and are ejected
almost vertically (Sehmel, 1980). Saltating particles strike the surface at very small angles -
almost horizontally - and cause an avalanching effect. In the absence of saltation, very little
resuspension would occur at all (Sehmel, 1980; Nicholson, 1993). Because resuspension is
driven by saltation and not the direct pick-up by wind, the size distribution of resuspended
particles does not change with windspeed (Young et al., 2002).
Vehicular resuspension is the result either of shearing stress of the tires or turbulence
generated by the passing vehicle (Nicholson, 1988b; Nicholson et al., 1989). This process can be
particularly important, since the most contaminated roadways tend to have the most traffic.
As with wind resuspension, a number of factors can affect the rate of resuspension from
vehicular motion. These factors include vehicle size, vehicle speed, moisture, and particle size.
2-64
-------
Lead in street dust appears to have a bimodal distribution. The fine mode is likely from
vapor phase condensation from combustion engines, whereas the coarse mode is likely from
either vehicle wear or significant coagulation of smaller particles. Al-Chalabi and Hawker
(1997) observed that in roadways with significant resuspension, Pb concentrations were lower,
indicating either dispersion from the source or the scavenging of smaller Pb particles by coarser
particles. Abu-Allaban et al. (2003) similarly observed that Pb in road dust tended to be in the
coarse mode. Measurements performed in tunnel tests indicated that < 17% of Pb in PMio was
smaller than 2.5 jim (Lough et al., 2005).
Resuspension may occur as a series of events. Short episodes of high windspeeds, dry
conditions, and other factors conducive to resuspension may dominate annual averages of
upward flux (Nicholson, 1988b, 1993).
Lead concentrations in suspended soil and dust vary significantly. In suspended soils
sampled near industrial emitters of Pb, PMio-bound Pb varied between 0.012 and 1.2 mg Pb/ kg
of bulk soil (Young et al., 2002). Tsai and Wu (1995) measured Pb in airborne particles that was
30 times higher than Pb in road dust. This enrichment factor was much higher than for other
pollutants, which may indicate that Pb is more easily resuspended than other contaminants.
Fractions of Pb observed in suspended dusts and soils are listed in Table 2-23.
The contribution of resuspended soil and dust to the airborne burden may be significant,
particularly from highly contaminated sites. A source apportionment study in Boston indicated
that soil resuspension increased the airborne concentration of Pb by as much as 0.02 |ig/m3 in the
fine mode (Thurston and Spengler, 1985). Isotopic measurements in Yerevan, Armenia credited
resuspension of contaminated soil with 75% of the atmospheric Pb in 1998 (Kurkjian et al.,
2002). Calculations based on road dust emissions and Pb weight fractions indicate that
resuspension was responsible for -40% of overall Pb emissions to the South Coast Air Basin of
California in 1989 (Lankey et al., 1998). Resuspension estimates based on modeling efforts
for the same area suggest that resuspension contributed -90% of overall Pb emissions in 2001
(Harris and Davidson, 2005). Figures 2-6 and 2-7 illustrate how air and soil concentrations
may be affected by long-term resuspension.
2-65
-------
Table 2-23. Observed Percentages of Lead in Resuspended Particulate Matter
Source
Paved road dust

Paved road dust

Unpaved road
dust
Unpaved road
dust
Unpaved road
dust
Unpaved road
dust
Agricultural soil
Agricultural soil
Agricultural soil
Agricultural soil
Agricultural soil
Agricultural soil
Playa dust

Sand and gravel
storage
Construction site

Location
Urban
San Joaquin Valley
Urban
Fresno, CA
Urban
Reno and Sparks, NV
Rural
San Joaquin Valley
Rural
San Joaquin Valley
Rural
Bakersfield, CA
Residential
San Joaquin Valley
Staging area
San Joaquin Valley
San Joaquin Valley
San Joaquin Valley
San Joaquin Valley
San Joaquin Valley
San Joaquin Valley
Stockton, CA
Rural
Reno and Sparks, NV
Visalia, CA

Urban
Reno and Sparks, NV
Pb fraction of
PM10 mass (%)
0.0161 ±0.0031

0.3 ±0.03

0.0057 ±0.0028

0.0058 ±0.0073

0.01

0.0203 ±0.0133

0.0043 ± 0.0008

0.0063 ±0.0059
0.0031 ±0.0025
0.0062 ± 0.0034
0.0024 ± 0.0082
0.003 ± 0.0025
0.01

0.02

Pb fraction of
PM2.5 mass (%) Reference
Chow etal. (2003)

0.4 Chow etal. (1994)

l.E-02 Gillies etal. (1999)

Chow etal. (2003)

Chow etal. (1994)

Chow etal. (2003)

Chow etal. (2003)
Chow etal. (2003)
Chow etal. (2003)
Chow etal. (2003)
Chow etal. (2003)
Chow etal. (1994)
l.E-03 Gillies etal. (1999)

Chow etal. (1994)

l.E-03 Gillies etal. (1999)

Source: Chow et al. (1994, 2003) and Gillies et al. (1999).
2.3.4 Runoff from Impervious Surfaces
The runoff of water from impervious surfaces may be a significant transport route for Pb
from urban areas to soil, waterways, and catchment basins. As water runs off roadways and
buildings, it can become laden with dissolved and suspended matter. Dust on roadways contains
2-66
-------
0,010
c
0
"•5

S
*j
c
o
o
c
0
o

"6
(0
0,000 \
1970 1975
1980 1985

Year
1990 1995 2000
Figure 2-6. Modeled soil concentrations of lead in the South Coast Air Basin of California

based on four resuspension rates.

Source: Reprinted from Harris and Davidson (2005).
Symbol
Calc. of Airborne Pb Cone.
A=1x10-10/s

A=1x10-11/s

Measured
1975
1980 1985

Year
1990 1995 2000
Figure 2-7. Modeled and measured airborne concentrations of lead in the South Coast Air

Basin of California based on two resuspension rates.

Source: Reprinted from Harris and Davidson (2005).
2-67
-------
a significant fraction of Pb due to vehicle wear, vehicle emissions, road wear, fluid leakage, and
atmospheric deposition. Lead in road dust is discussed in further detail in Sections 2.3.3 and 3.2
of this document. Additionally, Pb-containing paints, gutters, roofing materials, and other
housing materials may leach with rainfall.
Urban catchments in Fresno, CA had highly elevated soil-Pb concentrations indicative of
high Pb concentrations in runoff waters (Nightengale, 1987). Basins in use since 1962, 1965,
and 1969 had surface soil-Pb concentrations of 570, 670, and 1400 ppm, respectively. Nearby
control soils had surface-Pb concentrations between 8.3 and 107 ppm.
Urban runoff released into a stream in State College, PA caused significant spikes in Pb
concentrations (Lieb and Carline, 2000). Concentrations upstream of the release point were
1.5 ng/L. Downstream concentrations were 1.8 |ig/L when there was no precipitation, and but
averaged 14.6 |ig/L during storm events.
The amount of Pb is removed from roadways and buildings by rainwater depends in part
on the intensity of the storm. Experiments performed by Davis and Burns (1999) indicated that
high intensity storms washed away significantly more exterior house paint than low-intensity
storms. Other experiments showed that the amount of Pb in roadway runoff increased
significantly with length of dry period prior to a rain event (Hewitt and Rashed, 1992).
Lead in runoff water primarily occurs in the particulate form, with only a relatively small
fraction in the dissolved form (Hewitt and Rashed, 1992; Davis and Burns, 1999; Roger et al.,
1998). Between 69% and 93% of Pb washed from painted structures was reported to be in
parti culate form (Davis and Burns, 1999). More than 90% of Pb in highway runoff from a rural
highway in the UK was in the particulate phase (Hewitt and Rashed, 1992). Roger et al. (1998)
found that Pb particles in a motorway catchment in France were typically <50 jim in diameter.
Also, samples taken from road water samples in France showed that most Pb was in an
inorganic, non-bioavailable form (Flores-Rodriguez et al., 1994).
Amounts of Pb from roadways vary by region, rainfall intensity, maximum inflow,
rainfall duration, and antecedent dry weather periods (Shinya et al., 2000). Based on
measurements taken near a roadway in France, Pb concentrations in runoff water ranged between
0.46 and 4.57 g Pb/kg of suspended PM (Roger et al., 1998). Another study of French roadways
found an average Pb content of 2.36 g Pb/kg of dried material (Flores-Rodriguez et al., 1994).
Thirteen storm events studied at a heavily trafficked, rural highway in England showed mean Pb
2-68
-------
contents of 181 |ig/L (Hewitt and Rashed, 1992). Of this total, 16.2 ± 6.9 |ig/L was in the
dissolved phase and 165 ± 101 jig/L in the particulate phase. An additional 0.36 jig/L was in an
organic form. Mean Pb concentrations during four rain events studied near a roadway in Japan
ranged between 17 and 39 |ig/L (Shinya et al., 2000). The initial concentrations were higher,
ranging from 130 to 567 |ig/L. This indicates the presence of a first flush effect, in which much
of the Pb contamination is removed within the initial period of rainfall. Hewitt and Rashed
(1992) evident in a similar downward trend in Pb concentrations with time. However, no first
flush phenomenon was evident in a study by Taebi and Droste (2004), which evaluated
combined urban runoff transported to a mixed residential and commercial urban catchment in
Iran. The Pb concentrations for each of 10 major rainfall events ranged between 0.018 and
0.558 |ig/L. The arithmetic mean for all 10 events was 0.278 |ig/L.
Studies of runoff from building materials showed high Pb concentrations from painted
wood and painted brick, particularly if the paint is more than 10 years old (Davis and Burns,
1999; Davis et al., 2001). The maximum Pb concentrations were 1,900 |ig/L and 28,000 |ig/L
associated with painted exterior wood and brick surfaces, respectively (Davis and Burns, 1999).
Lead from paint is released into waters in both particulate and dissolved form. Lead
concentrations observed in runoff from building surfaces are listed in Table 2-24.
Table 2-24. Lead Concentrations Observed in Runoff From Building Surfaces
Geometric
Substance Mean (jig/L)
Block (painted)
Brick (painted)
Wood (painted)
0-5 yr. old paint
5-10 yr. old paint
>10yr. old paint
Roofs
Residential roofs
Commercial roofs
Institutional roofs
9.2
22
43
8.0
18
81
6.0

Median
(Mg/L)
8.0
16
49
8.1
14
88
5.2
2
12
64
Mean
Hg/L)
38
580
170
27
120
810
38
1.5
62
64
Range (jig/L)
<2-590
<2-28000
<2-1900
<2-370
<2-2600
<2-28000
<2-590

Reference
Davis and Burns (1999)
Davis and Burns (1999)
Davis and Burns (1999)
Davis and Burns (1999)
Davis and Burns (1999)
Davis and Burns (1999)
Davis and Burns (1999)
Davis etal. (2001)
Davis etal. (2001)
Davis etal. (2001)
Source: Davis and Burns (1999) and Davis et al. (2001).
2-69
-------
Matthes et al. (2002) studied runoff from lead sheet to simulate lead in gutters, roofs,
piping, siding, and sculptures. Typical concentrations in runoff ranged between 700 and
3700 mg/L. This was attributed to the solubility of cerrusite (Pb carbonate) and hydrocerrusite
(Pb hydroxy carbonate), which form on the surface of air-exposed Pb. Lead corrosion (cerrusite
and hydrocerrusite) dissolution rates from Pb sheets were measured at 14.3 to 19.6 millimoles of
lead/m2 per year (Matthes et al., 2002).
The amount of Pb removed by runoff events varies. Hewitt and Rashed (1992) estimate
that -8% of Pb and 5% of organic Pb emitted from vehicles is removed by highway drainage
waters. Shinya et al. (2000) estimate that total Pb loads for a roadway in Japan prior to four
storm events ranged between 0.053 and 0.771 mg Pb/m2. These storm events removed half of
the load in 0.07 to 3.18 hours after the start of the rainfall event.
Davis et al. (2001) estimate the total annual loading of Pb from all sources to be between
0.069 and 0.18 kg Pb/ha. They estimate that 80 to 90% of this is derived from runoff from
buildings.

2.3.5 Leaching of Soil Lead
Although Pb is generally relatively immobile and has a long retention in most soils, soil
Pb has some capacity to leach through the soil column, potentially contaminating ground water.
Lead sorbs strongly to constituents of the soil matrix and is only weakly soluble in pore water, so
the leaching of Pb is a much slower process than the leaching of many other contaminants
(Marcos et al., 2002; Zhang and Xu, 2003; Unlii, 1998; Pang et al., 2002). The sorbing capacity
of the soil and the solubility of the contaminants can be affected by the hydraulic conductivity of
the soil, the composition of the soil solution, the content of the soil organic matter, the content of
the soil clay minerals, soil pH, microbial activity, preferential flow through plant root channels
and animal holes, and geochemical reactions (Rhue et al., 1992; Elzahabi and Yong, 2001). The
experiments of Erel et al. (1997) on soil columns indicate that anthropogenic Pb is more readily
available for leaching than Pb that naturally occurs in the soil.
Lead can bind to many different surfaces in the heterogeneous soil matrix. This
adsorption greatly affects mobility and is dependent on the characteristics of the soil and Pb
compounds. Lead is partitioned between the soil water solution, precipitated forms, secondary
Fe or Mn oxides, carbonates, organic matter, sulfides, or the surfaces of clay, humus, or silicate
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particles (Badawy et al., 2002; Venditti et al., 2000; Cajuste et al., 2000; Erel and Patterson,
1994). The most labile fraction of Pb is adsorbed to the surfaces of colloid soil particles, which
may include organic matter, clay, oxides, or carbonates (Erel et al., 1997). Lead leached from a
limestone soil during a sequential fractionation procedure was exclusively in the iron/manganese
oxide form (Hee, 1994). A study of industrially contaminated soils found that -50% to 60% of
the Pb was not susceptible to leaching during any phase of a sequential fractionation procedure
(Cajuste et al., 2000). The remaining Pb was found primarily in the carbonate and Fe-Mn oxide
fractions, with sizeable amounts in the organic and exchangeable phases. None of the Pb was
water soluble. Maskall and Thornton (1998) also observed a high fraction of Pb in the carbonate
form in highly contaminated soil. The unusual presence of carbonate-bound Pb is probably due
to the formation of cerrusite (PbCO3) in soils contaminated with calcareous slag wastes (Maskall
and Thornton, 1998). Lead migration in this contaminated soil was associated with Fe-Mn
oxides. A third contaminated site was tested by Jing et al. (2004). These soils showed 57% of
Pb in the Fe-Mn oxide form, 29% in the carbonate form, and just 5% in the residual, soil-bound
form.
High chlorine content in soil has been shown to increase Pb leaching (Urdu, 1998).
Chloride complexation with Pb enhances lead solubility.
The pore-water velocity is inversely proportional to sorption rates. At low flow, the
longer retention times Pb to more complete sorption of Pb to soil particles (Pang et al., 2002).
In laboratory experiments on soil columns, transport of Pb was enhanced by the
introduction of soil colloid suspensions (Karathanasis, 2000). Colloids increased transport of not
only colloid-bound Pb but also dissolved Pb. Colloid transport was enhanced by increasing the
colloid surface charge, increasing the pH, increasing the amount of organic carbon, increasing
the soil macroposity, decreasing the colloid size, and decreasing the Al, Fe, and quartz contents
(Karathanasis, 2000). Colloid binding and co-transport of Pb are important mechanisms for Pb
migration, but colloids also enhance the flow of Pb through physical blockage from exchange
sites, competitive sorption, and organic complexation (Karathanasis, 2000). Denaix et al. (2001)
observed that most of the Pb-transporting colloids in an acidic, loamy soil were biological in
nature. The Pb concentration in the colloid fraction was not correlated with pH, colloidal
organic carbon contents, or colloidal silicon concentrations (Denaix et al., 2001). Approximately
50% of the total Pb transfer in these experiments was attributed to colloidal transfer.
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At low pH, metal species bound to carbonates, hydroxides, and other soil matrix
components are more likely to dissolve into solution (Maskall and Thornton, 1998; Elzahabi and
Yong, 2001; Badawy et al., 2002). This increases the rate of Pb migration through the soil. The
experiments of Jing et al. (2004), which followed eight different leaching protocols, suggest that
pH is the primary factor in determining the concentration of Pb in leached solution. At pH >12,
Pb forms soluble hydroxide anion complexes and leaches out of the soil column. At pH between
6 and 12, Pb leachibility is low due to adsorption and precipitation. At pH <6 free Pb ions leach
into the pore water and are removed from the soil columns. Rhue et al. (1992) observed that
organic Pb species (Me2Pb2+ and Et2Pb2+) were best absorbed at pH 6.2 and 7.2, respectively.
Sorption decreased at pH <5 and >8.2 (Rhue et al., 1992).
A partition coefficient, Kd, is often used to describe the susceptibility of Pb to leaching.
This value is used to compare the fractionation of a contaminant between liquid and solid forms.
Kd is defined by the following equation:

Kd = S/C' (2-10)

where S is the total concentration of Pb adsorbed in the solid phase and C' is the concentration of
Pb in pore water solution (Elzahabi and Yong, 2001). Kd increases with increasing pH (up to
7.0) and increasing distance from the leachate source (Elzahabi and Yong, 2001; Sheppard and
Sheppard, 1991). Kd decreases with an increase in the influent heavy metal concentration and
the degree of saturation (Elzahabi and Yong, 2001). The highest value of Kd appears to be near
the source of Pb contamination. Values of Kd in the literature cover many orders of magnitude
between 1.20 L/kg and "infinity" (when no Pb can be detected in pore water). These values are
listed in Table 2-25. For more information on Pb solid-solution partitioning see Chapter 7.
The rate of migration through the soil has been estimated in many different studies.
Using Pb isotopes, Erel et al. (1997) estimate the rate of Pb migration to be 0.5 cm/year in soils
collected from rural locations in Israel. Sheppard and Sheppard (1991) measured the rate of flow
through spiked soils, which were highly acidic and had a low organic matter content. These
soils, which were especially susceptible to leaching, exhibited migration rates of 0.3 cm/day
during the first year of experiments. The migration rate appeared to slow down in subsequent
years. Cores taken at smelting sites used during the Roman era, medieval times, and the
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Table 2-25. Soil/Water Partition Coefficients for Several Different Soils and Conditions
Kd (L/kg)
12.68
3.23
1.20
1.36
-6000
-3000
-5000
20
9000
92.99
14.25
125.58
95.51
1330 ±200
Beginning Soil
pH Water Content (%) Soil Type
4.0
4.0
3.5
3.5
n.a.
n.a.
n.a.
4.9
4.8
4.45
4.45
5.01
5.01
3.0-4.0
26.69
28.20
26.29
26.32
n.a.
n.a.
n.a.
n.a.
n.a.
n.a.
n.a.
n.a.
n.a.
n.a.
Illitic (spiked)
Illitic (spiked)
Illitic (spiked)
Illitic (spiked)
Brown pseudopodzolic
Rendzina
Gley podzolic
Acidic
(low-organic-matter
sand)
Sphagnum peat
Mining site
Mining site
Mining site
Mining site
Acidic (high-
organic-matter peat)
Reference
Elzahabi and Yong (2001)
Elzahabi and Yong (2001)
Elzahabi and Yong (2001)
Elzahabi and Yong (2001)
Alumaa et al. (2002)
Alumaa et al. (2002)
Alumaa et al. (2002)
Sheppard and Sheppard (1991)
Sheppard and Sheppard (1991)
Merrington and Alloway (1994)
Merrington and Alloway (1994)
Merrington and Alloway (1994)
Merrington and Alloway (1994)
Deiss et al. (2004)
Source: Elzahabi and Yong (2001), Alumaa et al. (2002), Sheppard and Sheppard (1991), Merrington and
Alloway (1994), and Deiss et al. (2004).
18th century underwent sequential extraction (Maskall and Thornton, 1998). The estimated Pb
migration rates at the Roman, medieval, and 18th century sites were 0.07 to 0.54 cm/year,
0.31 to 1.44 cm/year, and 0.11 to 1.48 cm/year, respectively. Using Pb measurements from soil
cores in agricultural fields, Zhang (2003) estimated a maximum deposition flux of approximately
27 (+14) |ig/cm2-year (270 mg/m2-year) to have occurred around 1940 (attributed by the author
to residential coal burning). Measurements made for more recent time periods were lower, but
accompanied by such variability as to be indistinguishable from zero via the study methodology
in more recent years. Over the 50-year period from 1950-2000, the authors estimated a loss from
surface soils (to lower horizon soils) of more than 50% of atmospherically-derived Pb.
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Mass balance calculations of Miller and Friedland (1994) suggest that the response time
(or soil horizon flushing time) for Pb is 17 and 77 years in the organic horizons of northern
hardwood and spruce-fir forests, respectively. Similar calculations by Kaste et al. (2003) at the
same site predicted that anthropogenic Pb will take -60 and -150 years to be transported through
the organic horizon in the deciduous and spruce-fir forests, respectively. The difference in
response times for the two forests may be due to differences in the litter depth and/or in the rate
of litter decomposition.
Soil tested from a car battery salvage facility showed a significantly greater Pb
concentration in the leached solution than in a reference soil (Jensen et al., 2000). Lead
concentrations in the leached solution went as high as 8000 |ig/L. Other industrially
contaminated soils did not show such high rates of leaching, but these other soils had nearly
neutral pHs.
Isotopic ratios in soil cores in the Sierra Nevada, California showed that 21% of Pb at a
depth of 30 cm had anthropogenic origins and had migrated from the surface (Erel and Patterson,
1994). The remaining 79% of Pb at this depth was naturally occurring.
Physical mixing of soils through animal activity may also increase the rate of Pb
migration. Mace et al. (1997) observed a significant decrease in Pb transport time through soil
as a result of rodent activity in a southern California location.
Vilomet et al. (2003) used isotopes to trace the leaching of Pb from a landfill into
groundwater in France. The active landfill has been in use since 1900 and has no bottom liner.
Detectable quantities of leached Pb were seen as far as 4600 m downgradient (Vilomet et al.,
2003).

2.3.6 Transport in Aquatic Systems
Chemical, biological, and mechanical processes govern the cycling of Pb in aquatic
environments. The main focus here is on the exchange between sediment and surface water,
which is affected by many different factors including salinity, the formation of organic
complexes, redox conditions, and pH (Arakel and Hongjun, 1992).
Lead enters surface waters from a number of sources. Atmospheric deposition is the
largest source, but urban runoff and industrial discharge are also significant (Peltier et al., 2003;
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Hagner, 2002; Perkins et al., 2000). As expected, concentrations in surface waters are highest
near sources of pollution.
The dispersal of Pb in waterways is relatively quick. If Pb is emitted into waterways as a
point source, water concentrations decrease rapidly downstream of the source (Rhoads and
Cahill, 1999; Hagner, 2002; Kurkjian et al., 2004; Peltier et al., 2003). Lead is removed from the
water column through flushing, evaporation, or sedimentation (Schell and Barnes, 1986).
Kurkjian et al. (2004) noted that first order approximations of concentrations of non-conservative
pollutants (such as Pb) can be made by using the exponential decay curve:
C = C0ekx (2-11)
where C is the pollutant concentration, C0 is the concentration at the source, x is the downstream
distance from the source, and k is the decay rate in km1. For the Debed River in Armenia,
Kurkjian et al. (2004) found that a decay rate of 0.57 km1 provided the best fit to measured
Pb concentrations.
Metals in waterways are transported primarily as soluble chelates and ions, constituents of
PM, or by adsorption onto suspended organic or inorganic colloids (Arakel and Hongjun, 1992).
The last two are the most important for Pb. The predominant chemical forms of Pb that interact
with aqueous ecosystems are PbO and PbCOs (Schell and Barnes, 1986). Lead is adsorbed on
colloids that are typically secondary clay minerals, Fe-Mn oxides or hydroxides, or organic
compounds (Arakel and Hongjun, 1992). The concentration of Pb appears to increase with
increasing salinity (Arakel and Hongjun, 1992).
Schell and Barnes (1986) describe water columns as "transient reservoirs" for pollutants.
They found mean residence times for Pb in two lakes and a reservoir to be between 77 and
250 days, although it should be noted that residence times tend to be shorter in turbulent
waterways. Lead concentrations in water are attenuated by the presence of A1(OH)3
precipitation, which is responsible for an estimated 54% of total Pb loss, and by the adsorption of
Pb onto other particles which settle out of the water column, which makes up the other 46% of
Pb loss (Kurkjian et al., 2004). Schell and Barnes (1986) measured sedimentation rates for
anthropogenic Pb, which ranged between 0.036 g cm2 a1 and 0.064 g cm2 a1.
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Concentrations of Pb in sediments roughly follow the concentrations of Pb in overlying
water (Kurkjian et al., 2004; Rhoads and Cahill, 1999). Thus, Pb concentrations in sediment are
highest near sources and decrease downstream.
Lead preferentially sorbs onto small particles rather than large particles. Small grain sizes
and the larger surface area per unit weight Pb to greater potential for adsorption (Rhoads and
Cahill, 1999). Concentrations of metals increase approximately logarithmically with decreasing
particle size.
Organic matter in sediment has a high capacity to accumulate trace elements. High humic
levels may Pb to greater Pb contamination in sediments (Rhoads and Cahill, 1999; Kiratli and
Ergin, 1996).
Sulfides are another potential source of Pb adsorption. This is especially true under
anoxic conditions (Kiratli and Ergin, 1996; Perkins et al., 2000). An increase in the amount of
sulfide in pore water was shown to decrease the dissolved concentration of Pb (Peltier et al.,
2003).
Lead in sediment can also be sequestered on iron or manganese oxides (Peltier et al.,
2003; Gallon et al., 2004; Schintu et al., 1991). These forms may make Pb more susceptible to
recycling into the overlying water column (Schintu et al., 1991).
Lead appears to be relatively stable in sediment. It has a very long residence time, and
many studies suggest that Pb is not mobile in the sediment. However, many other studies
suggest that Pb-containing particles can be remobilized into the water column (Ritson et al.,
1999; Steding et al., 2000; Hlavay et al., 2001; Kurkjian et al., 2004; Peltier et al., 2003; Gallon
et al., 2004). For example, Steding et al. (2000) observe that isotopic concentrations of Pb in the
San Francisco Bay match those of leaded gasoline from the 1960s and 1970s, suggesting that
recontamination by sediment may be a significant source of Pb to overlying waters. Ritson et al.
(1999) similarly observed that there was a negligible reduction in Pb concentrations in the
San Francisco Bay despite the closing of a nearby Pb smelter, the implementation of municipal
effluent controls, and the elimination of Pb additives to gasoline. That concentrations have
remained high may suggest recycling of sediment Pb. Similarly, in a study of water Pb
concentrations in the North Sea, Pb concentrations did not decrease significantly with the
elimination of major sources (Hagner, 2002). This also may indicate continued high rates of
atmospheric deposition or cycling of Pb stored temporarily in sediment.
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Modeling efforts of Gallon et al. (2004) indicate that processes that resuspend sediment
(such as diffusion, bioturbation, and bioirrigation) are small compared to sedimentation of
colloidal particles. Kurkjian et al. (2004) suggest a correction factor for equation (2-11) to
account for the contribution of Pb from sediment.
C = C0e(-kx) + Is (2-12)

where Is is the amount of Pb resuspended into the water column. Depending on the region of the
river under discussion, the authors extrapolated Is values in the range of 1.3-2.8 jig Pb/L.
2.3.7 Plant Uptake
Plants that take up Pb can be a source of Pb exposure for wildlife, livestock, and humans
that consume contaminated plants. More thorough discussion of soil Pb extraction by plants and
subsequent effects on ecosystem health can be found in Chapter 7 of this document.
Plants grown in soils contaminated by mine spoils (e.g., Cobb et al., 2000), smelting
operations (e.g., Barcan et al., 1998), sludge amendments (e.g., Dudka and Miller, 1999),
contaminated irrigation water (e.g., Al-Subu et al., 2003), or Pb-containing agrochemicals (e.g.,
Azimi et al., 2004) have higher than natural concentrations of Pb. In general, higher Pb
concentrations in soils typically result in increased Pb levels in plants.
Although the transfer of soil Pb to plants and direct stomatal uptake of atmospherically
deposited Pb are generally small, all plants can accumulate Pb to some degree (Finster et al.,
2004). The rate of uptake is affected by plant species, soil conditions, and Pb species.
Of all the factors affecting root uptake, pH is believed to have the strongest effect (Dudka
and Miller, 1999). Acidic soils are more likely to have Pb in solution and therefore available for
absorption. This is sometimes attenuated by liming.
Most Pb in plants is stored in roots, and very little is stored in fruits (e.g., Finster et al.,
2004; Cobb et al., 2000). Of 33 edible plants grown in urban gardens, roots had a median Pb
concentration that was 12% of the soil-Pb concentration (Finster et al., 2004). Shoot Pb, when
detectable, was just 27% of root Pb. Root vegetables seem the most prone to Pb uptake,
followed by leafy vegetables (Dudka and Miller, 1999; Finster et al., 2004). Fruits and grains do
not seem as susceptible to Pb contamination.
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Metals that are applied to soil as salts (usually as sulfate, chloride, or nitrate salt) are
accumulated more readily than the same quantity of metal added via sewage sludge, flue dust, or
fly ash (Dudka and Miller, 1999). This is likely because metal salts Pb to the formation of metal
chloride complexes and ion pairs, which can increase metal diffusion and subsequent root
uptake.

2.3.8 Routes of Exposure for Livestock and Wildlife
There are many routes of Pb exposure for livestock and wildlife, including food ingestion,
drinking water, and inhalation for terrestrial organisms. For aquatic organisms, the main routes
of Pb exposure are food ingestion and water intake. A few representative studies which have
analyzed routes of Pb exposure for livestock and wildlife are summarized here. For a discussion
of health effects, toxicity, and Pb concentrations in animal tissue, see Chapter 7.
Lead concentration of plants ingested by animals is primarily a result of atmospheric
deposition of Pb particles onto plant surfaces rather than the uptake of soil Pb through plant roots
(Steinnes, 2001; Palacios et al., 2002; Dudka and Miller, 1999). The uptake of Pb by the lowest
trophic levels - invertebrates, phytoplankton, and krill for example - are some of the most
important avenues for introducing Pb into food chains (Pilgrim and Hughes, 1994;
Sanchez-Hernandez, 2000; Hagner, 2002).
Some of the highest levels of Pb exposure in animals occur near major sources like
smelters. In two studies of horses living near smelters, the estimated ingestion rate was in the
range of 2.4 to 99.5 mg Pb/kg body weight per day (Palacios et al., 2002) and 6.0 mg Pb/kg body
weight per day (Liu, 2003). Both exposure rates were well above the estimated fatal dose for
horses. Sheep grazing near smelters were similarly poisoned (Liu, 2003; Pilgrim and Hughes,
1994). Installation of pollution controls at a Pb smelter in Slovenia greatly reduced the amount
of Pb in nearby vegetation and the blood-Pb levels of cows grazing on this vegetation (Zadnik,
2004). However, Pb concentrations in topsoil at this site have not noticeably declined in the
20 years since the pollution controls were implemented.
The amount of Pb entering the food chain depends highly on the species of the animal, the
species of their food, and where the organisms live. A study of sheep living in the southernmost
part of Norway (the most polluted part of the country), showed a strong correlation between liver
Pb concentrations and moss concentrations (Steinnes, 2001). The sheep fed almost exclusively
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on a grass that easily picks up atmospherically deposited Pb. Correspondingly high Pb levels
were also observed in hare and black grouse in this region. Similarly, a study of Pb
concentrations in raccoon tissues showed much higher Pb levels in urban raccoons than in rural
raccoons (Khan et al., 1995). This may be because urban raccoons are exposed to higher air
concentrations, ingest human refuse, or frequently visit storm sewers. In general, ruminant
animals appear to be more resistant to Pb ingestion than monogastric animals (Humphreys,
1991).
Acute Pb poisoning observed in Laysan albatross (Phoebastria immutabilis) chicks was
traced to the direct ingestion of paint chips by using isotopic analysis (Finkelstein et al., 2003).
Blood-Pb concentrations in P. immutabilis at the Midway Island National Wildlife Refuge had a
geometric mean of 190 |ig/dL. P. immutabilis chicks at a reference site had blood-Pb levels of
4.5 |ig/dL. Lead levels are somewhat elevated even in Antarctic animals (Sanchez-Hernandez,
2000). Antarctic food systems are supported by krill (Euphausia superba), which is the primary
food source for organisms in higher trophic levels. Lead concentrations measured in E. superba
were in the range of 0.17-12.0 ppm by dry weight. This is probably elevated above natural levels
due to anthropogenic Pb inputs (Sanchez-Hernandez, 2000).
Lead contamination in mammals and fish livers was shown to be higher in highly polluted
coastal zones than in the open sea (Hagner, 2002). In foraminifers, which are meiobenthic
organisms, high sediment Pb concentrations corresponded to high tissue Pb concentrations.
Sediment concentrations were 10 to 20 times higher than foraminifer concentrations. Fish take
in Pb either in their food or in water through their gills. The relative importance of these two
mechanisms depends largely on the fish species. A literature survey suggests that there has been
no observable decrease in fish muscle and liver concentrations of Pb over the past twenty years
in marine or freshwater environments (Hagner, 2002). Lead concentrations in the harbor
porpoise (Phocoenaphocoend) appear to increase with the age of the animal, but this was not
evident for the common seal (Phoca vitulina) (Hagner, 2002). Shrimp (Palaemonetes varians)
were shown to absorb 4 to 8% of the Pb content of its prey (Boisson et al., 2003). Between 52%
and 57% of the Pb accumulated from food was irreversibly retained in P. varians tissue. Just 2%
of dissolved Pb accumulated from water was retained in tissue (Boisson et al., 2003).
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2.4 METHODS FOR MEASURING ENVIRONMENTAL LEAD
The previous 1986 Lead AQCD (U.S. Environmental Protection Agency, 1986) contained
a detailed review of sampling and analytical methods for measuring Pb in environmental media.
Included in that document were discussions of site selection criteria, sampling methods, sample
preparation, and analysis techniques. Furthermore, the document included discussion of
sampling of Pb emissions from mobile and stationary sources. In this section, only a very brief
summary is provided for approaches for sampling and analysis of Pb in environmental media.
For a more comprehensive discussion, the reader is referred to the 1986 Lead AQCD.
Emissions can be estimated from measurements at sources using grab samples, periodic
samples, or continuous monitoring. Determining the rate of emissions requires knowing both the
fluid flow rate and the concentration of Pb in the fluid, usually air or water. Thus, it is much
easier to measure emissions from stacks than it is to measure fugitive, diffuse, or nonpoint
emissions (Frey and Small, 2003). Much of the recent improvement in the measurement of Pb
emissions from sources is due to better sampling and analytical equipment. For example, better
dilution tunnels can provide reliable samples from in-stack sampling, and improved analytical
methods (such as inductively coupled plasma mass spectrometry) permit determination of Pb at
much lower levels than in years past. This means that it is possible to obtain data from short
sampling runs, permitting better time resolution.
Wet deposition can be collected using precipitation buckets that seal tightly immediately
before and after rain. Dry deposition on land can be sampled using surrogate surfaces, such as
Teflon plates (Davidson et al., 1985; Davidson and Wu, 1990); or it can be done alternatively by
leaf-washing (Lindberg and Lovett, 1985) or by sampling throughfall precipitation that washes
previously deposited Pb off the vegetation and onto the forest floor (Wu et al., 1992b). Dry
deposition onto bodies of water is more difficult to estimate, usually requiring airborne
concentrations used in conjunction with deposition velocity estimates (Zufall and Davidson,
1997). Subsequent analysis of all of these samples can be performed by atomic absorption
spectrometry, neutron activation analysis, x-ray fluorescence, or proton-induced x-ray emission
(Koutrakis and Sioutas, 1996), or by inductively-coupled plasma mass spectrometry (ICP-MS)
(U.S. Environmental Protection Agency, 1991).
Recently developed single-particle instruments can identify which particles contain Pb,
and what other elements are present in the same particle. Such instrumentation can also provide
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information on the size of the particle (Pekney et al., 2006; Silva and Prather, 1997). Although
such instruments are not able to determine the precise mass of Pb in each particle, they can
provide valuable data on the characteristics of particles that contain Pb from individual sources
or source categories. Such "fingerprinting" methods can be used to identify sources of
Pb-containing particles in the environment.
2.5 SUMMARY
This chapter discusses sources of airborne Pb that generally fall in three categories:
natural sources, stationary sources, and mobile sources.
Nationwide, U.S. airborne Pb emissions fell 98% between 1970 and 2003 (U.S.
Environmental Protection Agency, 2003). The elimination of alkyllead additives to automotive
gasoline was principally responsible for the drop, although air-Pb emissions fell by 5% from
1993 to 2002 after the total phaseout of leaded fuel (U.S. Environmental Protection Agency,
2003). Figure 2-8 shows the decline in estimated U.S. Pb emissions from 1982 through 2002.
80,000
60,000
40,000
20,000
in 1985, EPA refined its methods for estimating emissions
Transportation
Industrial Processes
'uel Combustion
1982-02: 93% decrease
1993-02: 5% decrease
85
92 93 94 95 96 97 98 99 00 01 02
Year
Figure 2-8. Trends in U.S. air lead emissions, 1982-2002.
Source: U.S. Environmental Protection Agency (2003).
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For most of the past 50 to 60 years, the primary use of Pb was as additives for gasoline.
Leaded gasoline use peaked in the 1970s, and worldwide consumption has declined since
(Nriagu, 1990). The largest source of air-Pb emissions was leaded gasoline throughout the 1970s
and 1980s. In 1980, on-road vehicles were responsible for -80% of air-Pb emissions, whereas in
2002, on-road vehicles contributed less than half of a percent (U.S. Environmental Protection
Agency, 2003). In every case where the U.S. Pb NAAQS has been exceeded since 2002,
stationary point sources were responsible (www.epa.gov/air/oaqps/greenbk/inte.html).
Pirrone and Keeler (1996) estimated that motor vehicles were the source of 85% of total
Pb deposition to the Rouge River watershed in 1982, but only 2.1% of Pb deposition flux in
1992; in 1992, the major sources of Pb deposition flux were estimated to be steel (36%),
municipal solid waste combustion (29%), sewage sludge incineration (21%) and coal combustion
(10.4%). Source apportionment analyses in PM2.5 samples collected in 1998-1999 in the
New York - New Jersey Harbor Bight area indicated that fossil fuel combustion, oil combustion,
metal processing industry, and waste incineration were the major sources of Pb in fine particles
(Gao et al., 2002).
Stationary sources emitted an estimated 1,662,000 kilograms nationwide in 2000 (U.S.
Environmental Protection Agency, 2003). The largest emitters are now in the industrial sector,
which includes iron and steel foundries, smelters, combustion of hazardous and solid waste, and
others (Harris and Davidson, 2005). These emissions are not confined to the air 90 facilities
nationwide also generate 90% of the Pb-containing solid hazardous waste (Chadha et al., 1998).
Nationwide air emissions in 2000 were estimated as 1,600 tons in total emissions of Pb, with
contributions of 288 tons from iron and steel foundries, 234 tons from industrial boilers/heaters,
165 tons from coal-fired utility boilers, and 143 tons from mobile sources (see Table 2-8).
It should be noted that Pb emission inventories have significant omissions and
discrepancies (Harris et al., 2006; Chadha et al., 1998). An analysis of four emission inventories
for Pb in southern California showed that major Pb emitters were missed by all four databases,
and that the databases were not consistent with one another nor updated regularly (Harris et al.,
2006). Thus, the data noted above are probably a lower limit for Pb emissions. Efforts to
develop accurate databases of Pb emissions are needed.
Natural processes contribute only a relatively small amount to the overall load of Pb in the
environment. Nriagu and Pacyna (1988) estimate mean global natural emissions are at least an
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order of magnitude smaller than anthropogenic emissions. Natural sources include volcanoes,
seasalt spray, wild forest fires, wind-borne soil particles, and biogenic processes (Nriagu, 1989).
Air is the major transport route for Pb emissions. Deposition of airborne pollutants to
surfaces has been observed in the most remote places on Earth, including the Arctic and
Antarctic. Mass balance calculations performed on an agricultural plot in France indicate that
atmospheric deposition is the dominant source of Pb to soil even when Pb-containing fertilizer is
applied (Azimi et al., 2004). However, on a local scale, solid waste disposal or mine tailings
may be the predominant source of soil Pb.
A rigorous comparison of resuspension, leaching, and plant uptake "removal" rates for
soil Pb has not been undertaken. Resuspension of Pb-containing particles is likely the dominant
removal mechanism from surface soil when soil pH is high. Leaching may dominate when soil
pH is low. Leaching of Pb through soil occurs more rapidly than uptake to pea or wheat crops
(Azimi et al., 2004). More research is needed to compare removal rates for other plants with soil
Pb migration and resuspension rates.
Surface waters are contaminated through several routes. On a global scale, sediment
resuspension and wet and dry deposition are the predominant contributors to Pb concentrations
in surface water. On a local scale, industrial effluent and urban runoff may dominate.
The major routes of Pb transport into the food chain appear to be the ingestion of
contaminated plants, ingestion of contaminated water, and inhalation of contaminated air.
Research into the relative importance of each of these transport routes is needed.
Measurements conducted in essentially any ecosystem worldwide show some level of Pb
contamination. Anthropogenic Pb reaches these ecosystems through many possible transport
routes, such as those shown in Figure 2-9.
2-83
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Paints (weathering,
burning, sanding)
Oil & Coal
combustion
Smelting, refining

Cement production
Mining
Incineration
Sewage sludge
Municipal waste
ATMOSPHERE
Dry
Wet deposition
deposition
Leaching from
waste disposal
Battery
Manufacturing/
recycling

Aircraft
Lead pipes
Storage tanks
Sewage effluent
Industrial effluent
Solid waste
disposal
Sewage sludge
Inorganic phosphatic
fertilizer
Paint
Figure 2-9. Transport pathways for lead in the environment.

Source: Modified from Zabel (1993).
2-84
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3. ROUTES OF HUMAN EXPOSURE TO LEAD AND
OBSERVED ENVIRONMENTAL CONCENTRATIONS
Introduction
This chapter assesses information regarding human exposures to lead (Pb) via various
media and routes of exposure. Lead has been observed in measurable quantities in nearly every
environmental medium all over the world. As summarized in the 1986 Lead Air Quality Criteria
Document (1986 Lead AQCD), human exposure to Pb occurs through various routes, as shown
in Figure 3-1. That figure is a simplified diagram of multimedia routes of exposure via various
environmental media, with a focus on the ambient air. The multimedia aspects of Pb exposure
can be seen in that Pb emissions to the air contribute to Pb concentrations in water, soil and
dusts; and Pb in soil and dust also can make important contributions to Pb concentrations in
ambient air. The relative contributions of Pb from different media and different sources to
human Pb exposure depend on factors such as the proximity of major sources to the residence
and workplace of the individual, the condition of the residence (especially the presence and
condition of any Pb-based paint present) and whether the residence is in an urban, suburban or
rural location.
In general, Pb exposure in the United States has fallen with the elimination of leaded
gasoline, Pb-based paint and Pb solder in cans. As airborne concentrations of Pb have fallen in
the United States, a corresponding decrease in blood-Pb levels of the U.S. population has
occurred. In a meta-analysis of 19 studies from six continents, a strong linear correlation was
observed between blood-Pb levels and gasoline Pb levels (Thomas et al., 1999). As gasoline Pb
was reduced to zero in the study countries, airborne Pb concentrations declined and converged to
less than 0.2 |ig/m3, and blood Pb levels also declined, converging to a median of 3 |ig/dL.
However, the potential for high Pb exposures remains, particularly in areas near major Pb
sources or with exposures to Pb-based paint or high Pb levels in drinking water.
This chapter discusses the evidence related to current concentrations of Pb in different
media and human Pb exposure contributions from various media. The chapter first focuses on
exposure to Pb from ambient air, briefly summarizing information available from various U.S.
3-1
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CRUSTAL
WEATHERING
INDUSTRIAL
EMISSIONS
SURFACE AND
GROUND WATER
DRINKING
WATER
Figure 3-1. Principal pathways of lead from the environment to humans. Heavy arrows
are those pathways discussed in greatest detail in this chapter.
Source: Modified from 1986 Lead AQCD (U.S. Environmental Protection Agency, 1986).
ambient air monitoring networks. The chapter also assesses information related to Pb exposure
from soil or dust, drinking water, dietary intake of foods or beverages, Pb-based paint, and from
various other sources. In each section, Pb concentrations reported for these various media are
discussed, along with available evidence on the contribution of Pb in multiple media to human
Pb exposures. As discussed in more detail in Chapters 4 and 6, concentrations of Pb in blood
and bone are the most common indices used as biological indicators or biomarkers of human
Pb exposure.
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3.1 EXPOSURE: AIR
Widespread emissions from anthropogenic sources (described in Chapter 2) have
contributed to elevated environmental Pb concentrations. In fact, airborne Pb concentrations in
many places throughout the world increased several orders of magnitude during the past seventy
years, largely due to the use of leaded gasoline additives (Miller and Friedland, 1994). The
lowest Pb concentrations measured are at the South Pole, where an average concentration of
0.076 ng/m3 was recorded (Maenhaut et al., 1979). Even at this remote location, it is likely that
the airborne Pb levels have exceeded historical background levels. This is evidenced by Pb
concentrations in Arctic ice sheets that increased from <1 ng/kg in 800 BC to 200 ng/kg in the
1960's (Murozumi et al., 1969).
Airborne concentrations of Pb in the United States have fallen dramatically over the last
30 years due largely to the phase out of leaded gasoline additives. Figure 3-2 shows the sharply
declining trend in overall U.S. airborne Pb concentrations since 1983. Major declines over
several orders of magnitude have been observed not only in urban areas, but also in rural regions
and remote locations. Data taken at rural sites throughout the United States since 1979 showed a
similar decline (Eldred and Cahill, 1994).
The United States has not been the only country to experience a significant drop in
airborne Pb concentrations. In the early 1980's, 5% of Europe's urban population was exposed
to air Pb concentrations above the World Health Organization's (WHO) recommended limit of
0.5 |ig/m3 for an annual average (Fenger, 1999; WHO, 2000). By the late 1980s, this value had
fallen, and very few locations reported concentrations above 0.5 |ig/m3. These areas were
primarily near large, uncontrolled metal industries (Fenger, 1999). Notable decreases in airborne
Pb have even been seen in remote locations. For example, measurements made in Bermuda
between 1993 and 1994 showed that, despite its remote location, airborne-Pb concentrations had
fallen by an order of magnitude since the 1970s and by a factor of four since the 1980s (Huang,
1996). Similarly, measurements taken at the South Pole were routinely below the detection limit
in 2000-2001, which indicates a significant improvement in Antarctic air quality since the 1970s
(Arimoto et al., 2004).
3-3
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c
o
V-»
2
-H^
c
0)
u
c
o
o
1.6
1.4
1.2
1.0
0.8
0.6
0.4
0.2
0.0
NAAQS
42 Sites
- 90% of sites have concentrations below this line
1983-02: 94% decrease
1993-02: 57% decrease
10% of sites have concentrations below this line
83 84 85 86 87 88 89 90 91 92 93 94 95 96 97 98 99 00 01 02
Year
Figure 3-2. Airborne Pb concentrations measured at FRM sites, averaged across the
United States for the years 1983 through 2002. The data are plotted in terms
of maximum arithmetic mean averaged over a calendar quarter and are
shown in relation to the current NAAQS of 1.5 ug/m3 (quarterly average).
Source: http://www.epa.gov/airtrends/lead.html.
3.1.1 Routine Monitoring of Lead in U.S. Ambient Air
Ambient air Pb concentrations are measured by four monitoring networks in the United
States, all funded in whole or in part by EPA. For compliance with the current Pb NAAQS,
quarterly average airborne concentrations of Pb are not to exceed 1.5 |ig/m3. Between
September 2001 and September 2002, there were just four areas in the United States not in
attainment of this standard: Liberty-Acadia, MO; Herculaneum, MO; East Helena, MT; and
Lame Deer, MT (U.S. Environmental Protection Agency, 2003). In 2004, there were only two
areas out of attainment (www.epa.gov/air/oaqps/greenbk/inte.html).

NAAQS Compliance Monitoring Sites - Federal Reference Method
This network is comprised of official state/local Pb monitoring stations which measure Pb
in total suspended particulate matter (TSP), i.e., particles up to about 30 microns, for the purpose
of determining compliance with the Pb NAAQS. These stations use samplers and laboratory
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analysis methods which have either Federal Reference Method (FRM) or Federal Equivalence
Method (FEM) status. The FRM and FEM method descriptions can be found in 40 CFR part 50,
Appendix G. Sampling is conducted for 24-hour periods, with a typical sampling schedule of
1 in 6 days. About 250 sampling sites operated during 2005. These sites provide a total Pb
measurement and are intended to be used for determining compliance with the Pb NAAQS. The
locations of these sites are shown in Figure 3-3. The state/local agencies which operate these
sites report the data to EPA's Air Quality System where they are accessible via several web-
based tools. Many of the stations in this network have been in operation since the 1970s. EPA's
series of annual air quality trends reports have used data from this network to quantify trends in
ambient air Pb concentrations. The most recent Trends Report for Lead can be found at
http://www.epa.gov/airtrends/lead.html.
>
° .Active Sites
• Inactive Sites
Figure 3-3. United States Lead TSP monitoring sites from 2000-2006. Monitor
information and data available at: http://www.epa.gov/air/data.
3-5
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Max quarterly average concentrations of Pb measured at the FRM monitors in 2000 to
2004 are, on average, quite low, with the composite mean ranging from 0.03 to 0.05 |ig/m3
(excluding point source-related monitors) and 0.10 to 0.22 (including point source-related
monitors). When data from point source-oriented monitors are included, in any given year
during 2000 to 2004 only one to five U.S. locations (from among -200 sites) had measured max
quarterly average Pb levels that exceeded the NAAQS level (1.5 |ig/m3, max quarterly average).

PM2.s Speciation Trends Network
This is a U.S. network of about 200 PM2 5 speciation sites. This network consists of
54 long-term trends sites, commonly referred to as the Speciation Trends Network (STN), and
about 150 supplemental sites. Nearly all of these state/local sites are in urban areas, often at the
location of highest known PM2.5 concentrations. Sites in this network determine the Pb
concentrations in PM2 5 samples and, as such, do not measure Pb in the size fraction >2.5 jim in
diameter. Lead is quantified via the XRF method. The standard operating procedure for metals
by XRF is available at: http://www.epa.gov/ttnamtil/files/ambient/pm25/spec/xrfsop.pdf
Data are managed through the Air Quality System. These sites generally began operation
around 2000.
The locations of these sites are shown in Figure 3-4a; and Figure 3-4b shows the average
maximum quarterly mean concentrations of Pb observed at those sites that were at or above
0.005 |ig/m3 for 2002-2005. In these data, the highest max quarterly average Pb concentration
reported was 0.168 |ig/m3, and the composite average of max quarterly average concentrations
was 0.0080 |ig/m3. As is shown in Figure 3-4b, the max quarterly average Pb concentration
exceeded 0.1 |ig/m3 at only one location during this 5-year period.

IMPROVE Network - PM25 Speciation
In the Interagency Monitoring of Protected Visual Environments (IMPROVE) network,
PM2 5 monitors are placed in "Class 1" areas (including National Parks and wilderness areas) and
are mostly in rural locations. This network is administered by the National Park Service, largely
with funding by EPA, on behalf of state air agencies that use the data to track trends in rural
visibility. Lead in the PM2.5 is again quantified via the XRF method. Data are managed and
made accessible mainly through the IMPROVE website, but also are available via the Air
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o Trends
: Supplemental
• Tribal
Figure 3-4a. Locations monitored by the Speciation Trends Network (STN).
• > 030 pg/m3 (2 monitor-methods)
• .010-.030|jg/m3(30)
o .005 -.010 ^g/m3(73)
. <.005Mg/m3(167)
Figure 3-4b. The average maximum quarterly mean Pb concentrations observed in
by the STN.
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Quality System. The oldest of these sites began operation in 1988, while many others began in
the mid 1990s. The locations of these sites are shown in Figure 3-5a. There are 110 formally
designated "IMPROVE" sites located in or near national parks and other Class I visibility areas,
virtually all of these being rural. Approximately 80 additional sites at various urban and rural
locations, requested and funded by various parties, are also informally treated as part of the
network. Samplers are operated on the same 1 in 3 day schedule as the STN by several different
federal, state, and tribal host agencies (see: http://vista.cira.colostate.edu/IMPROVE/).
Figure 3-5a. The Interagency Monitoring of Protected Visual Environments
(IMPROVE) network of PM2.s monitors. Monitor site information
available at:
http://vista.cira.colostate.edu/IMPROVE/overview/IMPROVEProgram file
s.htm.
Figure 3-5b shows IMPROVE sites that detected ambient air Pb concentrations in PM2.5
at or above 0.0008 |ig/m3 between 2000 and 2005. In the data from the IMPROVE network, the
highest max quarterly average Pb concentration reported was 0.008 |ig/m3, and the composite
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Figure 3-5b.
• > .030 M9/m3 (2 monitor-methods)
• .010-.030Mg/m3(30)
o .005-.010pg/m3(73)
» <.005|jg/m3(167)

IMPROVE sites with Pb PM2.5 concentrations at or above 0.0008 ug/m3
between 2000 and 2005. Air quality data available at:
http://vista.cira.colostate.edu/views/web/General/Data.aspx.
average of max quarterly average concentrations was 0.002 |ig/m . These levels are
considerably lower than those obtained in the PM2.5 speciation monitoring network, reflecting
the fact that speciation monitors are generally located in urban areas while the IMPROVE sites
are in national parks and wilderness areas. Recent studies have also reported that concentrations
of airborne Pb are sometimes several orders of magnitude higher in urban areas compared to
remote regions (Schroeder et al., 1987; Malm and Sisler, 2000). Rural areas tend to have Pb
concentrations falling somewhere between those of urban and remote areas. Thus, urban
populations are typically exposed to distinctly higher levels of airborne Pb than rural or
remote residents.
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National Air Toxics Trends Stations — PMio speciation
The National Air Toxics Trends Stations (NATTS) network of 24 sites monitors mostly
urban, but some rural, areas. These sites are also operated by 22 state or local host agencies.
All collect particulate matter as PMio for toxic metals analysis and, as such, do not measure Pb
in the size fraction >10 jim in diameter. Lead in the collected sample is quantified via the
ICP/MS method. The standard operating procedure for metals by ICP/MS is available at:
http://www.epa.gov/ttn/amtic/airtox.html. These NATTS sites are relatively new, with 2004
being the first year in which all were operating. The Air Quality System can be accessed at
http://www.epa.gov/ttn/airs/airsaqs/ (see Figure 3-6a for the locations of the NATTS
monitoring sites).
Figure 3-6a. The National Air Toxics Trends Stations (NATTS) network. Monitor site
information available at: http://www.epa.gov/ttn/amtic/airtoxpg.html.
Figure 3-6b shows the arithmetic mean of the maximum quarterly average Pb
concentrations in PMio observed at the NATTS sites during 2002 through 2006. In the data
from PMio monitors in the NATTS network, the highest max quarterly average Pb concentration
observed was 0.039 |ig/m3 during 2002 to 2005, and the composite mean of the max
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Figure 3-6b.
• > .012 (jg/m3 (2 monitor-methods)
• .006-.012 M9/m3(9)
o .003-.006 tjg/m3{11)
. s.003 |jg/m3(4)
Arithmetic mean of maximum quarterly average Pb concentrations
measured in PMio at NATTS network sites during 2002 through 2005. Air
quality data available at: http://www.epa.gov/ttn/amtic/airtoxpg.html.
quarterly average concentrations was 0.012 |ig/m3. Data are managed through the Air Quality
System.
In addition to these four networks, various organizations have operated other sampling
sites yielding data on ambient air concentrations of Pb, often for limited periods and/or for
primary purposes other than quantification of Pb itself. Most of these data are accessible via the
Air Quality System. In an effort to gather as much air toxics data (including Pb) into one
database, the EPA and STAPPA/ALAPCO created the Air Toxics Data Archive. The Air Toxics
Data Archive can be accessed at: http://vista.cira.colostate.edu/atda/.

Pb Concentrations in Different PM Size Classes
Airborne Pb concentrations are measured in three PM size fractions, as discussed above -
TSP, PMio and PM2.s - by the various monitoring networks. There are not many sites where Pb
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measurements are made in different PM size fractions at the same location (and where Pb values
exceed the minimum detection limit). From among U.S. monitoring sites, two obtain Pb
concentrations in both TSP and PMi0, 16 have Pb data available from both TSP and PM2.5, and
13 for both PMio and PM2.5.
In a combined analysis of data from all collocated monitoring sites, there is typically a
good correlation between Pb measurements in TSP and PMio (r = 0.73 at the one site with
10+ paired observations) and between Pb measurements in PMio and PM2 5 (r = 0.69 for 11 sites
with 10+ paired observations). However, the correlation between Pb measurements in TSP and
PM2 5 is generally not as high (r = 0.46 for 13 sites with 10+ paired observations). There is
substantial variability in the correlation between Pb concentrations in TSP and PM2.5 samples at
different sites. For those sites with at least 10 paired observations, the correlation coefficients
range from -0.25 to +0.95.
Table 3-1 summarizes data for Pb concentrations determined from several particle size
fractions in the monitoring networks discussed above. Focusing on the Pb concentrations
reported from TSP, PMio and PM2 5 samples in urban areas (i.e., not from the IMPROVE
network), it can be seen that the mean and median values are not markedly different, though in
general PM2.5 mass is about 50% of the mass of PMio, which is then about 50% of the mass of
TSP depending on the given area.
Table 3-1. Descriptive Statistics for Lead Measurements (in ug/m3) from Monitors*
Using Different Size Fractions of PM for Recent Years**
Particle size (network)
TSP*** (FRM, n~200)
TSP*** (FRM, n~200) excluding
point source sites
PM10 (NATTS, n=26)
PM2 5 (Speciation, n=272)
PM25 (IMPROVE, n=167)
Minimum
0.00
0.00
0.0027
0.002
0.0005
Mean
0.01-0.22
0.03-0.05
0.0116
0.008
0.0016
Median
0.02-0.04
0.01-0.02
0.0101
0.005
0.0013
Maximum
1.92-9.13
0.26-1.75
0.039
0.168
0.0065
* Excluding monitors representative of point source emissions
* * 2000-2004 for data from IMPROVE and TSP; 2002-2005 for data from the PM2 5 speciation network
and NATTS.
*** Data for TSP presented as range of values for each year
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Several research studies have reported Pb concentrations in different PM size fractions.
For example, in a rural area in the Southeastern United States, Goforth and Christoforou (2006)
reported average Pb concentrations of 6.11 ng/m3 in PM2 5 and 15.04 ng/m3 in TSP samples.
The average total mass of PM2.5 and TSP were, respectively, 9.5 |ig/m3 and 19.1 |ig/m3; thus,
Pb constituted a similar very small proportion of particles in each size fraction. In another study,
Singh et al. (2002) reported concentrations of metals in several PM size classes from two areas in
the Los Angeles Basin. In Downey, a site where refineries and traffic contribute heavily to
particle concentrations, Pb was proportionally greater in the fine and ultrafme fractions of PMi0.
In Riverside, which is considered a receptor site for particles transported from the Los Angeles
basin and also has agricultural sources, Pb was proportionally greater in the coarse fraction of
PMio. In Boston, MA, Pb concentrations of 326 ng/m3 and 75.6 ng/m3 were reported from PM25
and PMio-2.5, respectively (Thurston and Spengler, 1985). While there is clearly variation
between sites, these findings generally suggest that Pb is somewhat more likely to be found in
fine fraction particles than in larger particle sizes.
Lead is measured in three PM size fractions in only a few locations in the United States.
Table 3-2 shows Pb concentrations from three such areas: Wayne Co., MI (Detroit); Multnomah
Co., OR (Portland); and St. Louis City, MO.
Table 3-2. Maximum Quarterly Mean and Overall Average Quarterly Mean Lead
Measurements (in ug/m3) from U.S. Monitors using Different Size Fractions of PM for
Recent Years

Location
Wayne Co., MI
Multnomah Co., OR
St. Louis City, MO
Pbin
Maximum
0.041
0.015
0.216
TSP
Average
0.0313
0.0123
0.0216
Pb in PM10
Maximum
0.039
0.0273
0.0153
Average
0.0283
0.0188
0.0143
Pb in PM2 5
Maximum
0.0182
0.0178
0.0136
Average
0.0165
0.0118
0.0136
Note that although Pb was measured in TSP, PM10, and PM25 at the same site in each of the above three locations,
different PM monitoring methods were used for the different PM size fractions at a given size, contributing to
apparent anomaly of PM10 Pb value being higher than TSP Pb value for the Oregon site.
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Temporal and Spatial Variation in Pb Concentrations
Some seasonal variability is common for air Pb concentrations. However, the extent to
which seasonal variability is present depends on precipitation trends, changes in wind direction,
and mixing height variability for a given area. For example, a relative maximum was observed
in the winter in the Arctic because of the lack of precipitation during winter months (Heidam,
1986), whereas a relative maximum was observed in the summer in Bermuda when winds come
predominantly from Africa and Europe (Huang, 1996). Chiaradia and Cupelin (2000) observed
no seasonality in Pb concentrations in Geneva, Switzerland. Lead measurements taken at a
number of U.S. and French cities suggest some seasonal variation, based on seasonal differences
in mixing height (Del Delumyea and Kalivretenos, 1987).
Measurements made in Riverside, CA show diurnal trends (Singh et al., 2002). Lead
concentrations are high in the morning (6 to 10 a.m.) and the late afternoon (4 to 8 p.m.). This is
most probably indicative of heavy traffic, despite the use of unleaded gasoline, a depressed
atmospheric mixing height in the morning, and advection from Los Angeles traffic. Lead
concentrations in Riverside are significantly lower during midday (10 a.m. to 4 p.m.) and night
(8 p.m. to 6 a.m.).
Concentrations of Pb are dependent on height. This is particularly true if Pb is emitted at
street level from traffic. Measurements performed at roadsides in Hong Kong in 1997 showed
much higher Pb concentrations at breathing level than at rooftop level (Chan et al., 2000).
Similarly, Pb concentrations measured at four elevations in Berne, Switzerland showed a
pronounced decrease with height (Galli and Nyffeler, 1987). Some leaded gasoline was still
used in Hong Kong and Switzerland during these two studies. Also, measurements made in an
urban street canyon in Lahti, Finland showed that Pb concentrations declined by a factor of five
from street level (1.5m) to rooftop level (25m) (Vakeva et al., 1999).

3.1.2 Observed Concentrations - Indoor Air
Concentrations of Pb can be elevated indoors. Lead in indoor air is directly related to Pb
in housedust, which poses both an inhalation and an ingestion risk and is discussed in more detail
in Section 3.2. Strong correlations have been observed in a Boston study between indoor air,
floor dust, and soil Pb concentrations (Rabinowitz et al., 1985a). In the National Human
Exposure Assessment Survey (NHEXAS) study of six Midwestern states, Pb concentrations in
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personal air were significantly higher than either indoor or outdoor concentrations of air Pb
(Clayton et al., 1999). The predominant sources of indoor air Pb are thought to be outdoor air
and degraded Pb-based paint.
Lead concentrations tend to be somewhat elevated in houses of smokers. In a nationwide
U.S. study, blood-Pb levels were 38% higher in children who exhibited high cotinine levels,
which reflect high secondhand smoke exposure (Mannino et al., 2003). Lead is present both in
tobacco and in tobacco smoke, although Pb concentrations in tobacco have fallen in parallel with
decreases in airborne Pb concentrations (Mannino et al., 2003).
Another source of Pb in residential air is metal-cored candlewicks. The U.S. Consumer
Product Safety Commission banned the use of metal-cored candlewicks that contain more than
0.06% Pb as of October 15, 2003 (USGS, 2003). However, prior to this time, Pb emissions from
metal-core wicks were measured in the range of 0.5 to 66 jig/hour according to one study
(Nriagu and Kim, 2000) and 100 to 1700 jig/hour according to another study (Wassan et al.,
2002). In homes where such candles were burned, airborne Pb concentrations could have been
well above ambient air levels.

3.1.3 Observed Concentrations - Occupational
Lead concentrations inside work places can also be elevated. Thus, inhalation of Pb
during work hours is an additional route of exposure for some subpopulations.
Feng and Barratt (1994) measured Pb concentrations in two office buildings in the United
Kingdom (UK). In general, concentrations in the UK office buildings were higher than those in
nearby houses. Office dust Pb was concentrated in the organic and residual fractions, unlike
house dust which was bound to carbonate and Fe-Mn oxides. This indicates that offices and
houses may have different Pb sources. Office building Pb also tends to be in the coarse mode,
unlike house dust Pb that predominantly occurs in fine particles (Feng and Barratt, 1994).
As expected, Pb concentrations tend to be highly elevated within manufacturing facilities
for Pb-based products (Rieuwerts et al., 1999; Harrison et al., 1981; Tsai et al., 1997). Thus,
occupational exposure can represent a major Pb exposure route for employees working in such
facilities. For example, measurements taken in a battery manufacturing plant in the Czech
Republic found Pb concentrations in floor dust to be 47,700 ppm outside of the assembly plant,
39,200 ppm inside the assembly plant, and 73,700 ppm in the battery grid storage area
3-15
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(Rieuwerts et al., 1999). In another study in Taiwan, airborne Pb concentrations in a battery
manufacturing plant, a metallic film capacitor plant, and a Pb powder plant were 140 ±
112 |ig/m3, 281 ± 114 |ig/m3, and 485 ± 245 |ig/m3, respectively (Tsai et al., 1997). Work sites
that use mechanical actions such as abrasion, friction, and cutting typically generate large
particles. However, work sites that use high temperature operations generate small, respirable
particles. At the three sites listed above, particle sizes were predominantly >10 jam in diameter
(Tsai et al., 1997). A Pb-Zn smelter in the UK similarly showed much larger Pb particle sizes
inside the facility than outside of the facility (Harrison et al., 1981). This may be because
concentrations are high enough indoors to coagulate. Floor dusts (<60 jim) taken from each
process site in the overall smelting process contained the same Pb species as the aerosols emitted
from each process, which are discussed in Section 2.2.
Residential renovation and paint removal can also be major sources of Pb exposure for
both workers and residents. Dry sanding, abrasive blasting, and burning, welding, or heating
surfaces covered with Pb-based paint typically generate highly dangerous airborne Pb levels
(Jacobs, 1998). Geometric mean and maximum air Pb concentrations observed during each of
these processes (as reported by Jacobs, 1998) are listed in Table 3-3. Daniels et al. (2001)
measured airborne Pb concentrations during exterior paint removal from residences via wet
abrasive blasting technology. The eight-hour, time-weighted average (TWA) air exposures
measured via personal monitors ranged between 55.1 and 81.5 |ig/m3. Area air Pb
concentrations were between 20.5 and 26.9 |ig/m3.
Lead-based paints were the predominant coating for U.S. highway bridges for many years.
Paint removal during bridge renovation projects has also been cited as a major source of Pb
exposure for workers. As with residential renovation, Pb concentrations during industrial paint
removal depend largely on the technology used. Generally, abrasive blasting techniques are
used, which breaks Pb coatings into small particles that can be inhaled or ingested if hands are
not washed prior to eating or smoking (Chute and Mostaghim, 1991). Vacuum blasting may
reduce occupational exposures. Personal monitors worn during vacuum blasting on a bridge
registered air Pb concentrations between 27 and 76 |ig/m3, with a geometric mean of 55 |ig/m3
(Mickelson and Johnston, 1995). Concentrations measured eleven meters from the removal
processes fell to 0.1 and 2 |ig/m3 over an eight-hour TWA.
3-16
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Table 3-3. Airborne Lead Concentrations in Areas Surrounding Residential
Lead-Based Paint Abatement Activities
Abatement Technique
Preparation (e.g., carpet removal)
Abrasion
Chemical stripping
Encapsulation
Heat gun
Component replacement
Cleaning
Geometric Mean
(ug/m3)
2
8
3
2
7
3
2
Maximum Exposure
(ug/m3)
206
403
476
72
915
121
590
Source: Jacobs (1998).
Certain types of mining operations can also result in occupational Pb exposure. For
example, air-Pb concentrations measured in underground gold mines were somewhat elevated,
but comparable, to ambient Pb concentrations due to adequate air exchange (Annegarn et al.,
1988). Air Pb concentrations ranged between 1.4 |ig/m3 and 800 |ig/m3 and were highly
dependent on the process being undertaken (Annegarn et al., 1988). However, another source
apportionment study in a Nevada gold mine measured Pb concentrations that averaged
0.21 |ig/m3 (McDonald et al., 2003).
Children of Pb workers are also at increased risk for exposure. In a meta-analysis of take-
home Pb exposure, the geometric mean blood Pb level for children was 9.3 |ig/dL (Roscoe et al.,
1999). This was significantly higher than the geometric mean of 3.6 |ig/dL for children overall.
Similarly, 52% of children of Pb workers had blood Pb levels at or above 10 |ig/dL, compared
with just 8.9% of children nationwide (Roscoe et al., 1999). Having a parent in an automobile
body or maintenance occupation also appears to contribute to increased blood Pb levels in
children (Murgueytio et al., 1998a).
3-17
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3.2 EXPOSURE: SOIL AND DUST
Contaminated soil can be a potential source of Pb exposure for humans. Soil Pb can be
directly ingested through hand-to-mouth behavior common in children, indirectly ingested
through contaminated food, or inhaled when breathing air containing resuspended soil particles.
Soil ingestion, as reported by parents, peaks during the second year of life and diminishes
thereafter (Lanphear et al., 2002). Soil Pb concentrations measured in urban, residential, and
industrial areas are discussed here. Soil-Pb concentrations in areas not influenced by Pb sources
have been estimated to be in the range of 1 to 200 ppm, with an average of 15 ppm (Zimdahl and
Skogerboe, 1977). In a U.S survey, Pb concentrations in agricultural soils were found to range
from <1 to 135 mg/kg, with a mean of 12.3 mg/kg (Holmgren et al., 1993). In a review of
studies on soil Pb contamination, Markus and McBratney (2001) reported a broad range of
soil-Pb concentrations. It should be noted that soil-Pb measurements are difficult to compare,
given the variety of extraction techniques and depths of soil cores analyzed in each study.
Lead in soil is derived mainly from atmospheric deposition, both from local sources and
long-range transport (Erel et al., 1997; Markus and McBratney, 2001; Sheets et al., 2001).
In general, soil in urban and residential areas is contaminated primarily via atmospheric
deposition, direct application of agricultural chemicals, and natural mineral weathering of parent
rock (Paces, 1998). At a local level, soil Pb contamination can be derived from agricultural and
food wastes, animal wastes and manure, logging and other wood-cutting activities, urban refuse,
municipal sewage sludge, miscellaneous organic wastes (including excreta, solid wastes from
metal manufacturing, coal fly ash and bottom fly ash, peat for agricultural and fuel uses),
wastage of commercial products, mine tailings, and smelter slags and wastes (Nriagu and
Pacyna, 1988). Flaking and peeling of Pb-based paint can also be a significant source of soil Pb
near old structures (Small et al., 1995; Finkelstein et al., 2003).
The retention time for Pb in the soil is much longer than it is in the air, such that the time
in which a change in emissions is reflected in soil Pb concentrations is much longer than for air
Pb concentrations. The only "removal" mechanisms for soil Pb are resuspension, mechanical
mixing from tilling, landscaping and animals, and leaching, the last of which is known to be a
slow process (see Chapter 2 of this document for details). Retention of Pb in soil is influenced
by soil characteristics, including the organic matter content (Schwab et al., 2005; Vega et al.,
2006). Modeling efforts by Harris and Davidson (2005) in southern California predict that
3-18
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reduction of surface soil Pb concentration to the steady state concentrations associated with
current Pb emissions will not be achieved for more than a hundred years, assuming emission
rates stay constant. Miller and Friedland (1994), employing a dynamic analysis, estimated a time
to steady state for Pb concentrations in the organic soil horizon of two forest types in the
northeastern United States based on estimates of 1989 atmospheric deposition. These times to
steady state were estimated to be approximately equal to five response times (5 times the
estimate of soil flushing time) or 90 years or 400 years for a northern hardwood forest and a
subalpine spruce-fir forest, respectively. The mean response times for Pb in these soils were
estimated as 17 and 77 years, respectively. A later study in the same region estimated the
response times as 60 years and 150 years for the two forests, respectively (Kaste et al., 2003).

3.2.1 Concentrations of Soil Lead in Urban Areas
The concentration of soil Pb varies significantly throughout urban areas, depending on
proximity to stationary sources and roadways and on wind speed and direction. In urban and
industrial areas, mean soil-Pb concentrations have been found to range from 23 to 1275 mg/kg,
with peak concentrations of over 10,000 mg/kg (Markus and McBratney, 2001).
The major sources of Pb in urban soils are (a) automotive traffic from the days of leaded
gasoline (Sheets et al., 2001; Mielke, 1993; Sutherland, 2000) and (b) deteriorating exterior
Pb-based paint. Soil concentrations decrease both with depth and distance from roadways.
In one study of 831 homes in the United States, 7% of housing units were found to have soil-Pb
levels exceeding 1200 ppm, the U.S.EPA/HUD standard for soil Pb concentration outside of play
areas (Jacobs et al., 2002). Soil-Pb concentrations were related to the presence of deteriorating
exterior Pb-based paint; 24% of housing units that had deteriorated, exterior, Pb-based paint had
bare soil-Pb concentrations in excess of 1200 ppm, whereas just 4% of homes without
deteriorating paint had soil-Pb levels greater than 1200 ppm (Jacobs et al., 2002). Soil-Pb
concentration was also generally higher in homes constructed in earlier years (e.g., before 1940)
than more recently constructed homes. Another review of studies conducted in several urban
areas concluded that automotive Pb was a more important influence than the presence of Pb-
based paint under some conditions (Mielke, 1993).
Extensive studies in Baltimore, New Orleans, and cities throughout Minnesota found the
highest soil-Pb concentrations in the central sections of each city, where traffic and population
3-19
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density are greatest (Mielke, 1991, 1993). The lowest concentrations were found in the outskirts
of these cities and in smaller cities. In all of these studies, the age of housing did not seem to be
a major factor, which suggests that the impacts of Pb-based paint may be dominated by historic
emissions of leaded gasoline additives. However, the fact that the highest concentrations are
typically found in the inner city (generally disproportionately populated by minorities and the
poor) suggests that these groups are likely most at risk for Pb exposure from contaminated soil.
Some of the highest soil-Pb concentrations are found near major roadways. Surface
soil-Pb concentrations measured near a major freeway in Cincinnati, OH, fell between 59 and
1980 ppm (Turer et al., 2001). These concentrations dropped off dramatically with depth.
An estimated 40% of Pb from automobile exhaust is retained in the nearby soil (Turer et al.,
2001).
Measurements by Erel et al. (1997) in Israel show that soil-Pb concentrations decrease
more rapidly with depth near roadways than far from roadways. In a soil profile extracted near a
local road, Pb concentrations fell by a factor of 42 between the surface and 30 to 36 cm from the
surface. However, far from the roadway, Pb concentrations decreased by about a factor of
3 between the surface and 30 to 36 cm below the surface.
Several authors making measurements during the days of leaded gasoline usage reported
elevated Pb concentrations in soil that decreased with distance from roadways. For example,
Pierson and Brachaczek (1976) reported soil-Pb levels that decreased from >1000 ppm adjacent
to the road down to less than 200 ppm at 12.5 m from the roadway edge. These concentrations
have likely stayed high despite the elimination of leaded gasoline use. Harris and Davidson
(2005) have also shown through use of a mass balance model that elevated Pb concentrations in
soil are likely to remain high for hundreds of years; and this is consistent with other studies
showing similarly long residence times for Pb in soil (e.g., Dudka and Adriano, 1997).
Several studies have assessed the impact of soil Pb concentrations on blood-Pb levels.
Without accounting for other sources of Pb intake, Duggan and Inskip (1985) estimated that, for
every 1000 ppm increase in soil-Pb concentration, children's blood-Pb levels increase 5 |ig/dL.
Aschengrau et al. (1994) reported decreases in blood-Pb levels of 1.12 to 1.35 |ig/dL per
1000 ppm reductions in soil Pb concentrations during a randomized control trial. The results of a
pooled analysis of 12 studies showed an average 3.8 |ig/dL increase in blood-Pb levels per
1000 ppm increase in soil-Pb levels (Lanphear et al., 1998). Soil abatement at a Superfund site
3-20
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resulted in a mean 3.5 |ig/dL decrease in blood-Pb levels for 6 to 36 month old children
(Lanphear et al., 2003). A smaller reduction in blood-Pb levels was observed for 36 to 72 month
old children because of age differences, Pb intake from other sources, and mouthing behaviors.
Murgueytio et al. (1998b) observed a 2.8 |ig/dL increase in blood-Pb levels with increases in soil
concentrations of 1000 ppm. Accounting for age differences and, therefore, the redistribution of
bone-Pb stores, (Gwiazda et al., 2005) reconciles many of the apparent differences between the
results of the Lanphear et al. (1998, 2003), Murgueytio et al. (1998b), and Aschengrau et al.
(1994) studies.

3.2.2 Soil-Lead Concentrations Near Stationary Sources
Concentrations Near Lead Smelters
Lead in soil is highly elevated near sources of Pb emissions. In particular, areas around
stationary facilities, such as smelters and battery disposal sites, can have very high levels of
soil Pb.
Major smelter deposits exist primarily within a 0.5 km radius of the stack (Chatterjee and
Banerjee, 1999; Rieuwerts et al., 1999), although some studies have found elevated soil-Pb
concentrations as far away as 30 km (Liu, 2003). Franssens et al. (2004) used isotopic
measurements to show that between 50% and 80% of dry depositing Pb within a 3 to 4 km radius
of a Pb-zinc smelter had an industrial origin.
Soil-Pb concentrations decrease dramatically with distance from the source, and they
depend greatly on windspeed and direction (Kimbrough and Suffet, 1995; Palacios et al., 2002;
Suchara and Sucharova, 2004). Godin et al. (1985) measured soil-Pb concentrations that were
almost proportional to the inverse of the distance from the source and the square root of the wind
direction frequency. Suchara and Sucharova (2004) estimated an exponential decrease in soil-Pb
concentration with distance from a Pb smelter in the Czech Republic. Data collected within a
14 km radius showed an exponential decrease in soil-Pb concentration with distance from the
source. Exponential decreases in soil-Pb concentrations have been suggested elsewhere, as well
(e.g., Chatterjee and Banerjee, 1999; Rieuwerts et al., 1999). Results of Chatterjee and Banerjee
(1999) indicate that Pb concentrations remain relatively constant within about 250 meters of the
source and decrease with distance after this. Examples of data showing decreases in soil-Pb
concentration with distance from major sources are shown in Table 3-4.
3-21
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Table 3-4. Concentrations of Soil Lead with Distance from Lead Smelters
Distance from Smelter
(m)
Fenceline
20
30
40
100
123 - 256
250
400
500
700
1500
3000
5000
10000
20000
30000
Concentration
(ppm, dry weight)
2300a'4, 46700 ± 2100a'5, 12650b'6
5657dJ
3937dJ
3253d'1
783d'1'312.8±98.7e'2'1800a'4
636 ± 522C'8
229dJ, 20200 ±1100a'5
127dJ
400 ± 20a'5
792ej
519C'3
242C'3
216.7 ±87.6e'2, 137C'3
110.3±76.4e'2
57.4 ± 24.9e'2
32.9±21.4e'2
Note: In cases where multiple transects were sampled, only the downwind transects are shown.
Values are given as mean ± standard deviation.

"Depth sampled was not defined ^alacios et al. (2002)
bSample depth was 0-5 cm 2Liu (2003)
"Sample depth was 0-10 cm 3Godin et al. (1985)
dSample depth was 0-15 cm 4Kimbrough and Suffet (1995)
eSample depth was 0-30 cm 5Chatterjee and Banerjee (1999)
6Rieuwerts et al. (1999)
7Venditti et al. (2000)
8Young et al. (2002)
As in the case for urban soils, Pb concentrations decrease significantly with depth near

industrial sites. As an example, Table 3-5 lists a Pb concentration profile measured near a Pb

smelter in northern France.
3-22
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Table 3-5. Soil Lead Concentration Profile Measured Near a
Lead Smelter in Northern France
Depth (cm)
0-6
6-9
9-36
36-50
50-70
70-85
85-120
120-165
Soil Horizon
Oi
Oa
Ag
ABg
BAg
Bg
HC2g
HC3g
Soil Cone, (ppm)
2340
4480
383
21.7
18.2
17.1
12.4
10.2
Source: Denaixetal. (2001).
The species of metals found near smelters vary depending on soil conditions. One study
observed Pb in topsoil that was either in the form of Pb5(PO4)3Cl or Pb(II) compounds that were
adsorbed onto Fe(II) oxides or associated with clay particles (Batonneau et al., 2004). Other
measurements at a site contaminated with automotive battery wastes showed Pb species in the
soil to be Pb(CO)3, Pb(CO3)2, Pb(OH)2, PbO, and PbSO4 (Pichtel et al., 2000). Other studies
have shown Pb contamination bonded to bacteria (Denaix et al., 2001), carbonate (Maskall and
Thornton, 1998; Pichtel et al., 2000; Venditti et al., 2000), sulfide phases (Pichtel et al., 2000;
Venditti et al., 2000), organic phases (Pichtel et al., 2000; Venditti et al., 2000) and Fe-Mn
oxides (Venditti et al., 2000). The prevalence of carbonate forms in Pb-contaminated soil is due
to coinciding contamination with calcareous slag wastes (Maskall and Thornton, 1998).
Soil-Pb concentrations do not appear to have noticeably decreased in areas surrounding
smelters despite the implementation of pollution controls. A smelter in Slovenia was fitted with
protective filters in 1978 (Zadnik, 2004). Since that time, Pb concentrations have fallen
dramatically in hay samples and cow blood within 10 km of the smelter; however, soil-Pb
concentrations in areas around the smelter did not decrease between 1978 and 2003 (Zadnik,
2004). Similarly, a Pb-zinc smelter in British Columbia, Canada was replaced by a new smelting
facility in 1997 (Hilts, 2003). Airborne Pb concentrations fell by nearly 75%, and Pb levels fell
by 50% in outdoor dustfall, street dust, and indoor dustfall fell by 50%. However, no statistically
3-23
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significant decline was observed in soil-Pb concentrations nor in Pb concentrations in carpeting
inside nearby residences (Hilts, 2003). Soil-Pb concentrations at five U.S. factory sites, which
had closed decades ago, were elevated as well (Rabinowitz, 2005). Also, many sites where
smelters had previously operated but are currently unrecognized as such (Eckel et al., 2001)
may represent a previously unidentified Pb exposure risk for nearby populations.

Concentrations Near Mines
Concentrations of Pb are highly elevated near mines as well. Lead and zinc mines, in
particular, have large deposits of Pb in nearby soil, but mines used for extracting other metals
can also have Pb-contaminated soil nearby. Mine sites are contaminated by the disposal of mine
tailings, acid mine drainage, and atmospheric deposition of airborne emissions (Dudka and
Adriano, 1997). Mines in the Midwestern United States produced an estimated 480 Tg of Pb
tailings and 50 Tg of Pb mine wastes between 1910 and 1981 (Dudka and Adriano, 1997).
Concerns about the impact of mine tailing piles from Pb mines are being addressed under the
Comprehensive Environmental Response, Compensation, and Liability Act (CERCLA). Several
Midwestern U.S. states have been allowing the Pb mine waste piles to be used for roadway
resurfacing after washing and encapsulation of the waste into asphalt. However, if mine
overburden is used for surfacing of unpaved roads, parking lots, highway construction, etc.,
without first undergoing washing and encapsulation, it can potentially increase the risk for
human Pb exposure.
Lead is widely dispersed in areas surrounding mining sites (Dudka and Adriano, 1997;
Rieuwerts and Farago, 1995). Thus, it is not easy to determine a relationship between distance
and soil concentration, as is the case for smelting emissions. However, a study of an abandoned
Pb-zinc mine in Tyndrum, Scotland located near a river showed that fluvial transport had carried
Pb contamination at least as far as 6.5 km, although contamination is suspected as far as 25 km
downstream (MacKenzie and Pulford, 2002). Examples of soil-Pb concentrations measured near
mining sites are shown in Table 3-6.
Lead is found in many different forms near mining sites. It is commonly found in its
mineral form of galena (Rieuwerts and Farago, 1995; Dudka and Adriano, 1997). However,
in mine spoils, Pb is also found as plumbojarosite [PbFe6(SO4)4(OH)i2], pyromorphite
3-24
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Table 3-6. Soil Concentrations Measured Near Mining Sites
Location
Wales, UK
Halkyn, UK
Shipham, UK
Shipham, UK
Derbys, UK
Winster, UK
Leadville, US
Derbys, UK
Shipham, UK
Pribram, Czech
Republic
Tyndrum,
Scotland
Goldenville,
Canada
Sao Domingos,
Portugal
Jasper County,
Missouri, U.S.
Dubuque, Iowa,
U.S.
Type of
Mine
Pb
Pb-Zn
Zn,Pb
Zn,Pb
Pb
Pb
Pb
Pb
Zn,Pb
Pb
Pb-Zn
Au
Cu
Pb
Zn,Pb
Main Period
of Operation
historic,
not specified
1845-1938
1700-1850
1650-1850
18th and
19th cent.
Up to end of
18th cent.
1860s-1960s
18th and
19th cent.
18th and
19th cent.
18th-20thcent.
Up to 1862
1869-1927
Pre-Roman-
Roman times
1850-1957
19th century
Depth
(cm)
0-15
0-15
0-15
0-5
0-5
0-5
n.a.
0-15
0-15
0-5
n.a.
n.a.
0-30
n.a.
0-20
Mean cone.
(ppm)
1159
1127
7900
2002
5610
7140
1110
1800
7360 (max)
1451
13000
70-120
2694
574 ±691
791
Reference
Gallacher et al. (1984) (taken
from Rieuwerts and Farago,
1995)
Davies et al. (1985) (taken from
Rieuwerts and Farago, 1995)
Mattigod et al. (1986) (taken
from Rieuwerts and Farago,
1995)
Thornton (1988) (taken from
Rieuwerts and Farago, 1995)
Thornton et al. (1990) (taken
from Rieuwerts and Farago,
1995
Cotter-Howells and Thornton
(1991) (taken from Rieuwerts
and Farago, 1995)
Cook et al. (1993) (taken from
Rieuwerts and Farago, 1995
Li and Thornton (1993) (taken
from Rieuwerts and Farago,
1995)
Li and Thornton (1993) (taken
from Rieuwerts and Farago,
1995)
Rieuwerts and Farago (1996)
(taken from Rieuwerts and
Farago, 1995)
MacKenzie and Pulford (2002)
Wong et al. (2002)
Freitas et al. (2004)
Murgueytio et al. (1998b)
Mbila and Thompson (2004)
3-25
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[Pb5(PO4)3Cl], Pb carbonate [PbCO3], leadhillite [Pb4SO4(CO3)2(OH)2], PbS'Bi2S3, Pb oxides,
Pb silicates, and Pb sulfate [PbSO4] (Rieuwerts and Farago, 1995; Mbila and Thompson, 2004).
Lead tends to be more heavily concentrated in smaller soil grain sizes than in larger grain
sizes (MacKenzie and Pulford, 2002). Results of one study are listed in Table 3-7. Young et al.
(2002) observed that the Pb concentration was much higher in the <38 jim size range than in the
300 |im to 2 mm size range in contaminated soils. This is likely due to the higher specific
surface area of smaller soil particles and the fact that Pb tends to bond with organic matter and
Fe/Al oxides, which can also concentrate in smaller size particles (Young et al., 2002). Also,
Rieuwerts and Farago (1995) note that soil-Pb particles are typically larger in mining areas than
in smelting areas.
Table 3-7. Concentrations of Lead in Soils Grouped by Soil Grain Size
Size Fraction
>180(im
53-180 (im
<53 (im
Pb cone, of main
mine waste
0.91%
1.5%
4.5%
Pb cone, of processing
site waste
17%
14%
18%
Source: MacKenzie and Pulford (2002).
Lead concentrations in peat have also been shown to decrease with depth. Figure 3-7
illustrates two peat profiles sampled near an abandoned Pb mine.
Blood-Pb levels are typically elevated for people living near Pb mines. Soil collected
at residences near the Tar Creek Superfund Site, which is a Pb mining area in northeastern
Oklahoma, showed wind-dispersed mine wastes (Lynch et al., 2000). More than 20% of soils
exceeded the EPA action level of 500 ppm; and children's blood-Pb levels tended to be higher
when compared to children living outside the Superfund towns. In this same area, Malcoe et al.
(2002) showed that blood-Pb levels were highest among African-American, Mexican-American,
and poor children. Blood-Pb levels were most commonly correlated with mean floor dust Pb
3-26
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(i) Pb concentration (ing
1000 4000 6000 1000 21 Off,
(b)
chronology

1966

1922

ill!

1876

1IS2

1I2J
10 -
12 -
48
I
1.12 1.14 1.16 l.tl
>wpb/i«pb itom riito
Pb concenirition (mg kg'1)
100 2M
. I.I.
11)2
1172
1.11 1.14 l.li i.li
10W°'Pb Horn ntio
Figure 3-7. The changes in lead concentration with depth in two peat cores. Core A was
taken at a location adjacent to the ore processing area of the abandoned lead
mine in Tyndrum, Scotland. Core B was taken 0.5 km from the main mine
waste dump at the same site.
Source: MacKenzie and Pulford (2002).
loading and with soil Pb, especially front yard soil (Malcoe et al., 2002). At the Jasper County
Superfund Site in southwestern Missouri, homes had significantly higher soil and dust Pb levels
and significantly higher blood Pb levels than areas outside of the Superfund site (Murgueytio
et al., 1998b). A strong statistical relationship observed there between blood-Pb levels and dust-,
soil-, and paint- Pb concentrations.

3.2.3 Observed Concentrations - House Dust
Given the large amount of time people spend indoors, exposure to Pb in dusts and indoor
air can be significant. For children, dust ingested via hand-to-mouth activity can be a more
important source of Pb exposure than inhalation (Adgate et al., 1998; Oliver et al., 1999).
3-27
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However, dust can be resuspended through household activities (e.g., Ferro et al., 2004), thereby
posing an inhalation risk as well. The particle size of "dust" is not well defined, although 50 jim
or 75 |im in diameter is sometimes given as an upper limit. In a study performed in the UK, Pb
in housedust tended to be bound to the carbonate or Fe-Mn oxides (Feng and Barratt, 1994).
Lead can be measured as a concentration in house dust (in |ig/g) or as dust Pb loading per
surface area (jig/ft2). One survey reported that dust-Pb loading was a more important predictor
of children's blood-Pb concentrations than dust-Pb concentration (Lanphear et al., 1995).
Lead in housedust can derive from a number of different sources. Lead appears both to
come from sources outside the home (Jones et al., 2000; Adgate et al., 1998) and from Pb-based
paint (Hunt et al., 1993; Lanphear et al., 1996). A chemical mass balance study in Jersey City,
NJ observed that crustal sources contributed almost half of the Pb in residences, Pb-based paint
contributed about a third, and deposition of airborne Pb contributed the remainder (Adgate et al.,
1998). Residential concentrations measured at the Bunker Hill Superfund Site in northern Idaho
indicate that the Pb concentration in houses depends primarily on the neighborhood soil-Pb
concentration (Von Lindern et al., 2003a, 2003b). However, factors such as household hygiene,
the number of adults living in the house, and the number of hours children spend playing outside
were also shown to affect Pb concentrations. Using a classification scheme, Hunt et al. (1993)
identified sources of Pb in housedust in London for various particle size ranges. In the 64 to
1000 |im size range, the predominant source of Pb was Pb-based paint. However, in the <64 jim
size bin, paint, road dust, and garden soil were significant contributors. Lead deposition
measured on an interior plate near an open window, an unsheltered exterior plate, and a sheltered
exterior plate in New York City were 4.8, 14.2, and 32.3 |ig/(ft2 week) or -52, 153, and
348 |ig/(m2 week), respectively (Caravanos et al., 2006). Data from a control (interior plate,
closed window) showed deposition that was primarily from exterior, environmental sources
as well.
Living near a smelter or a mine contributes significantly to the Pb load in residences
(Rieuwerts and Farago, 1995; Rieuwerts et al., 1999; Sterling et al., 1998). Homes of mine and
smelter employees tend to have Pb levels elevated above those of nearby houses, indicating that
Pb can be transported into homes via workers (Rieuwerts et al., 1999). In a U.S., study, mining
wastes, paint, and soil were all shown to contribute to housedust Pb (Sterling et al., 1998). Soil
3-28
-------
and mining wastes accounted for more than 50% of Pb in housedust. Lead-based paint
contributed 16-23% of Pb in housedust in a mining community (Sterling et al., 1998).
Renovation, and especially old paint removal, can greatly increase Pb levels inside the
home (Laxen et al., 1987; Jacobs, 1998; Mielke et al., 2001). Removal of exterior paint via
power sanding released an estimated 7.4 kg of Pb as dust, causing Pb levels inside one house to
be well above safe levels (Mielke et al., 2001). Remaining in a residence during the deleading
procedure can be acutely dangerous (Rey-Alvarez and Menke-Hargrave, 1987). Deleading by
dry-scraping and sanding has been shown to raise children's blood-Pb levels during the process,
but deleading by covering or replacing painted surfaces decreased children's blood-Pb levels
during the abatement process (Amitai et al., 1991). Excessive Pb exposure can occur even after
Pb abatement. In one prospective controlled study, an average blood-Pb increase of 6.5 |ig/dL
was observed among children whose homes had undergone Pb-based paint abatement
(Aschengrau et al., 1997). Clark et al. (2004) found that despite adherence to U.S. Department
of Housing and Urban Development (HUD) post-abatement standards, six-month-old children
who lived in houses that had recently undergone Pb abatement were eleven times more likely to
have blood-Pb levels increase by 5 |ig/dL or more compared to a control group. These studies
suggest that existing clearance standards may not be adequate to fully protect children from Pb
exposure following abatement or other Pb hazard controls (Clark et al., 2004).
Examples of Pb concentrations and dust Pb loading levels measured in house dust, school
dust, and nursing home dust are shown in Table 3-8. It should be noted that dust-Pb loadings
may be better predictors of blood-Pb levels than dust concentrations (Lanphear et al., 1995,
1998). Standards for residential Pb loadings of housedust were set by EPA in 2001 to be
40 |ig/ft2 (430 |ig/m2) for floors and 250 |ig/ft2 (2690 |ig/m2) for windowsills.
An additional concern is attic dust-Pb or dust found in roof cavities. Significant deposits
of atmospheric Pb can build up in these spaces. This dust can seep into living spaces through
ceiling decorative artwork, cracks between the wall and ceiling, electric light fittings, wall vents,
or exhaust, roof, and ceiling fans (Davis and Gulson, 2005). Additionally, renovations, housing
additions, ceiling collapses, and storm damage can produce large plumes of attic dust (Davis and
Gulson, 2005).
Studies comparing Pb concentrations in attic dust with house age showed a high
correlation between attic dust-Pb levels and ambient air-Pb concentration data measured
3-29
-------
Table 3-8. Examples of Lead Concentrations and Dust Lead Loadings in Indoor Dust
Concentration or Loading of
Lead (ppm unless otherwise
indicated) Location Surface Reference
oo
o
503 (mean)
308 (median)
43-13,600
9 (geometric mean)
1.5-48.9
117-362
1598

3025-4140
1283
114-185
1984
348
340
786
1870
1560
726
435
857±91ppminPM60
1133±119ppminPM10
975 ppm in PM53
481 ppminPM25o
Edinburgh Scotland Floor dust
Edinburgh Scotland Floor dust
Edinburgh Scotland Floor dust
Various parts of Denmark Floor dust
Various parts of Denmark Floor dust
UK Floor dust
Helena and Silver Valleys, US
(near 2 Pb smelters)
Trail, B.C. Canada (near Pb smelter)
Illinois (near Pb smelter)
Landskrona, Sweden (near Pb smelter) Floor dust
Pribram, Czech Republic (near Pb smelter) Floor dust
Wales, UK (near a mining site) Floor dust
Halkyn, UK (near a mining site) Floor dust
Shipham, UK (near a mining site) Floor dust
Derbys, UK (near a mining site) Floor dust
Winster, UK (near a mining site) Floor dust
Leadville, US (near a mining site)
Pribram, Czech Republic (near a mining site) Floor dust
Jersey City, NJ Floor dust
Jersey City, NJ Floor dust
Public school in Port Pirie, Australia Floor dust
Public school in Port Pirie, Australia Floor dust
Laxenetal. (1987)
Laxenetal. (1987)
Laxenetal. (1987)
Jensen (1992)
Jensen (1992)
Feng and Barratt( 1994)
Schilling and Bain (1988)a
Hertzmanetal. (1991)a
Kimbroughetal. (1994)a
Faragoetal. (1999)a
Rieuwerts and Farago (1996)a
Gallacheretal. (1984)a
Daviesetal. (1985)a
Thornton (1988)a
Thornton etal. (1990)a
Cotter-Howells and Thornton (1991)a
Cook etal. (1993)a
Rieuwerts and Farago (1996)a
Adgate et al. (1998)
Adgate et al. (1998)
Oliver etal. (1999)
Oliver etal. (1999)
-------
Table 3-8 (cont). Examples of Lead Concentrations and Dust Lead Loadings in Indoor Dust
Concentration or Loading of Lead
(ppm unless otherwise indicated)
1693-6799 ppm in PM53
1407-4590 ppm in PM250
954
14.4 (ng/m3)
26.8 (ng/m3)
558 ± 544 ppm in TSP
612±518ppminPM10
547 ± 5 12 ppm in PM2 5
5 140 (ng/m2)
18230 (ng/m2)
24,6 (ng/m2)
3 158 (ng/m2)
219.3 (ng/m2)
89.3 (ng/m2)
191.5 (ng/m2)
26.9 (ng/m2)
20.7 (ng/m2)
50.9 (ng/m2)
95.5 (ng/m2)
65.8 (ng/m2)
39.6 (ng/m2)
63.6 (ng/m2)
Location
Houses in Port Pirie, Australia
Houses in Port Pirie, Australia
Households in Midwest, US
Households in Midwest, US
Households in Midwest, US
Nursing homes in Vienna
Nursing homes in Vienna
Nursing homes in Vienna
Households in Midwest, US
Households in Midwest, US
Boston, MA
Cincinnati, OH
Cincinnati, OH
Rochester, NY
Rochester, NY
Butte, MT
Bingham Creek, UT
Leadville, CO
Magna, UT
Sandy, UT
Midvale, UT
Palmerton, PA
Surface
Floor dust
Floor dust
Windowsill dust
Airborne
Airborne, personal air
Airborne
Airborne
Airborne
Surface dust
Windowsill dust
Floor dust
Floor dust
Floor dust
Floor dust
Floor dust
Floor dust
Floor dust
Floor dust
Floor dust
Floor dust
Floor dust
Floor dust
Reference
Oliver etal. (1999)
Oliver etal. (1999)
Clayton etal. (1999)
Clayton etal. (1999)
Clayton etal. (1999)
Komarnicki (2005)
Komarnicki (2005)
Komarnicki (2005)
Clayton etal. (1999)
Clayton etal. (1999)
cited in Lanphear et al. (1998)
cited in Lanphear et al. (1998)
cited in Lanphear et al. (1998)
cited in Lanphear et al. (1998)
cited in Lanphear et al. (1998)
cited in Lanphear et al. (1998)
cited in Lanphear et al. (1998)
cited in Lanphear et al. (1998)
cited in Lanphear et al. (1998)
cited in Lanphear et al. (1998)
cited in Lanphear et al. (1998)
cited in Lanphear et al. (1998)
"Cited in Rieuwerts and Farago (1995).
-------
throughout the lifetime of the house (Chiaradia et al., 1997; Ilacqua et al., 2003). Attic dust may
even serve as a proxy for estimating historic ambient Pb concentrations, although the resolution
on such calculations would be low. Attic dust-Pb concentrations measured in Australia were an
order of magnitude higher in houses near a copper smelter compared to houses far from the
smelter (Chiaradia et al., 1997). However, isotopic analyses showed that alkyl-Pb additives
were the overall dominant source of Pb contamination in attic dust, suggesting that gasoline
emissions had a greater influence than the smelter. The geometric mean concentration of Pb
measured in attics in Sydney was 1660 ppm near industrial sites, 1173 ppm near semi-industrial
sites, 447 ppm in non-industrial sites, and 16 ppm in background crustal materials (Davis and
Gulson, 2005).
Even at low dust-Pb loading levels, Pb in housedust can have an effect on children's
blood-Pb levels. Epidemiological studies show that, at a median floor dust-Pb loading level of
5 jig/ft2 (54 jig/m2), approximately 5% of children have blood-Pb levels >10 |ig/dL (Lanphear
et al., 1998, 2005; Malcoe et al., 2002). At a floor dust-Pb loading of 50 |ig/ft2 (540 |ig/m2),
the percentage of children with blood-Pb levels > 10 |ig/dL rose to 20% (Lanphear et al., 1998)
In another study, children exposed to floor dust-Pb loadings in excess of 25 |ig/ft2 (270 |ig/m2)
were at eight times greater risk of having blood-Pb levels > 10 |ig/dL compared to children
exposed to levels below 2.5 |ig/ft2 (27 |ig/m2) (Lanphear et al., 2005).
Throughout early childhood, floor dust-Pb contamination is a source of exposure. Lead-
contaminated windowsill dust becomes an additional source of Pb intake during the second year
of life when children stand upright. Because of normal mouthing behaviors and increased
mobility, the highest blood-Pb levels are seen in children between 18 and 36 months of age
(Clark et al., 1991). This peak is typically observed after a rapid rise in blood-Pb levels between
6 and 12 months of age.

3.2.4 Concentrations of Lead in Road Dust
Elevated concentrations of Pb in road dust pose an important exposure risk through wind
and traffic resuspension, as outlined in Chapter 2 of this document.
The primary source of Pb in road dust is adjacent soil (De Miguel et al., 1997). However
traffic emissions, the weathering and corrosion of building materials (De Miguel et al., 1997),
and brake pad wear (Garg et al., 2000) are additional sources. Between 60 to 90% of the mass of
3-32
-------
road dust consists of soil particles (Adgate et al., 1998). Soil is still an important reservoir for Pb
emitted from vehicles despite the widespread phase out of leaded gasoline. The concentration of
Pb in road dust is generally elevated. This is particularly true in urban areas. Additionally,
measurements reported in 2003 in the San Joaquin Valley of California show concentrations that
were significantly lower than concentrations measured in the same area in 1987 (Chow et al.,
2003). Examples of road dust Pb data reported in the literature are listed in Table 3-9. Metals in
road dust tend to be associated with small size grains. Measurements of Kuang et al. (2004)
show that metals are concentrated in grains smaller than 0.125 mm in diameter.
De Miguel et al. (1997) observed a steep gradient in road dust concentrations of Pb in the
north-south direction in Oslo, Norway. This indicates that Pb concentrations are much higher in
the highly urbanized areas and lower in the suburban and residential areas. This is consistent
with traffic and building construction, renovation, and weathering of building materials being the
dominant source of Pb to soil and subsequently road dust (De Miguel et al., 1997).
3.3 EXPOSURE: DRINKING WATER
Lead in drinking water primarily results from corrosion from Pb pipes, Pb-based solder,
or brass or bronze fixtures within a residence (Lee et al., 1989; Singley, 1994; Isaac et al., 1997).
Very little Pb in drinking water comes from utility supplies. Experiments of Gulson et al. (1994)
have confirmed this by using isotopic Pb analysis. Tap water analyses for a public school,
apartments, and free standing houses also indicate that the indoor plumbing is a greater source of
Pb in drinking water than the utility, even for residences and schools serviced by Pb-pipe water
mains (Moir et al., 1996). Ratios of influent Pb concentration to tap concentrations in homes in
four municipalities in Massachusetts ranged between 0.17 to 0.69, providing further confirmation
that in-home Pb corrosion dominates the trace quantities of Pb in municipal water supplies (Isaac
et al., 1997). The information in this section addresses Pb concentrations in water intended for
human consumption only. However, such water comes from the natural environment, and
concentrations of Pb found in natural systems are discussed in Chapter 7.
The chemical composition of water distribution pipes is of great importance when
considering how much Pb is leached into drinking water. Copper piping with Pb-based solder
has largely replaced pure Pb piping in the United States. A 1988 survey of 94 U.S. water
3-33
-------
Table 3-9. Examples of Observed Road Dust Lead Concentrations
Cone. Of Lead (ppm)
Location
Land Use
Reference
OJ
oo
180 ± 14
1927 ± 508
536 ±39
57.2 ±27.3
-100
1209 ± 170 (PM25)
1061 ± 155 (PM10)
588 ±688
470 ± 524
151 ±124
161±31
57 ±28
109 ± 74
58 ±73
203 ±133
43 ±8
101 ±88
Oslo, Norway
Madrid, Spain
Calcutta, India
Beijing, China
Reno-Sparks, NV
Hong Kong
Hong Kong
Honolulu, HI
Honolulu, HI
Honolulu, HI
San Joaquin Valley, CA
San Joaquin Valley, CA
San Joaquin Valley, CA
San Joaquin Valley, CA
San Joaquin Valley, CA
San Joaquin Valley, CA
San Joaquin Valley, CA
urban, paved road
urban, paved road
near Pb smelter, paved
urban, paved road
urban, paved road
urban, paved road
urban, paved road
urban, paved road
urban, paved road
urban, paved road
urban, paved road
rural, paved road
composite, paved road
agricultural unpaved road
residential unpaved road
staging area soil
unpaved composite
De Miguel etal., 1997
De Miguel et al., 1997
Chatterjee andBanerjee, 1999
Kuangetal., 2004
Gillies et al., 1999
Ho et al., 2003
Ho et al., 2003
Sutherland et al., 2003
Sutherland et al., 2003
Sutherland et al., 2003
Chow et al., 2003
Chow et al., 2003
Chow et al., 2003
Chow et al., 2003
Chow et al., 2003
Chow et al., 2003
Chow et al., 2003
-------
companies nationwide revealed that copper pipe was present in 73% of homes, galvanized pipe
was present in 13% of homes, a mixture of galvanized and copper was present in 11% of homes,
and plastic pipes were present in 2% of homes (Lee et al., 1989). An analysis of PVC pipes
indicated that some Pb is leached from PVC in measurable amounts (Sadiq et al., 1997). PVC,
which contains -1% Pb, increased the tap water concentration to an average of 0.017 ±
0.038 mg/L, which was a statistically significant increase over the influent concentration of
0.011 ± 0.026 mg/L (Sadiq et al., 1997). Guo (1997) suggested that Pb may be leached from
cement-mortar lined pipes in significant quantities if the cement was made from clinker derived
from combusted, hazardous materials.
In addition to piping, Pb may leach from faucets. Water-Pb measurements performed for
12 faucets of different compositions typically found in homes indicated that new cast-brass
faucets leached more Pb than any of the other designs (Gardels and Sorg, 1989). Water-Pb
levels were below the detection limit from a plastic faucet. In houses with copper piping and
Pb-based solder, brass fixtures may contribute as much as 50% of Pb in drinking water (Lee
etal., 1989).
The primary type of solder used in the United States was 50-50 tin-Pb solder (50% tin,
50% Pb) before the Safe Drinking Water Act amendments of 1986 were enacted (U.S. EPA
2006). Although new or repaired pipes may not use solder containing more than 0.2% Pb, 50-50
solder still exists in many older structures. In comparing Pb leached from 50-50 tin-Pb solder,
95-5 tin-antimony solder, and a liquefied 50-50 tin-Pb formulation that contained a flux, Birden
et al. (1985) showed that the liquefied 50-50 formulation leached the most Pb into drinking
water. The 95-5 tin-antimony solder was the safest with respect to drinking water quality.
Measurements of metals leached from four, nonlead-based solders in copper pipes were made by
Subramanian et al. (1991, 1994). Of the four solders tested (95-5 Sn-Sb, 96-4 Sn-Ag, 94-6 Sn-
Ag, and 95.5-4.0-0.5 Sn-Cu-Ag), all showed that metals (Ag, Cd, Cu, Sb, Sn, and Zn) were
leached in small enough quantities to make these solders safe alternatives to Pb-based solders.
Lead corrosion is essentially an electrochemical process. Electrons may be transferred
from the metal (Pb) to the solution (drinking water), where the major electron acceptors are
dissolved oxygen, hydrogen ions, or disinfectant residuals (Singley, 1994). Alternatively, when
two different metals are in contact, there is a difference in potential, and the difference in
electron demand may increase corrosion (Singley, 1994). In either case, lowering the pH and
3-35
-------
increasing the dissolved oxygen demand are known to increase rates of corrosion. The combined
pH and alkalinity of water are sometimes described as the aggressiveness of the water and is
measured using the Langelier Index. A pH above 8.0 is generally considered safe for Pb
leaching (e.g., Lee et al., 1989; Frey, 1989). Also, the corrosion process occurs faster at high
temperatures than at low temperatures (e.g., Thompson and Sosnin, 1985; Lee et al., 1989).
There are conflicting reports on the effect of chlorine in water. Chlorine, which is
typically used as a disinfectant in municipal supplies, may increase the rate of corrosion by
providing a source of electron acceptors (Singley, 1994). However, measurements of Lee et al.
(1989) show an absence of statistically significant change in Pb levels with increasing
concentration of free chlorine. Laboratory tests of Edwards and Dudi (2004) show that chlorine
reacts with soluble Pb2+ to precipitate a red-brown colored Pb solid. This solid is highly
insoluble, even at a pH of 1.9 for twelve weeks. Thus, chlorine may actually lessen the overall
quantity of Pb in drinking water. Elevated levels of Pb in drinking water in Washington DC in
2000 were traced to a change from chlorine to chloramine disinfectant. The red-brown Pb solid
does not form in the presence of chloramines, and the data suggest that chloramines dramatically
increase the amount of Pb leached from brass (Edwards and Dudi, 2004).
Flouridating water does not seem to affect the solubility or reactivity of Pb compounds
(Urbansky and Schock, 2000).
Corrosion inhibitors are sometimes added to water to inhibit scaling or iron precipitation.
Zinc orthophosphate in the range of 0.4 to 0.6 mg/L is an effective inhibitor for Pb corrosion
(Lee et al., 1989). Results indicate that zinc orthophosphate is more effective at reducing Pb
levels than increasing the pH. Soluble Pb release is reduced by up to 70% with the addition of
orthophosphate (Edwards and McNeill, 2002). Other proposed corrosion inhibitors such as
sodium zinc hexametaphosphate or sodium hexametaphosphate are not effective at reducing Pb
corrosion (Lee et al., 1989). In fact, results of McNeill and Edwards (2004) indicate that
hexametaphosphate increased the levels of soluble Pb in drinking water. Each milligram per liter
of hexametaphosphate increased the Pb content by -1.6 mg/L after a 72 hour stagnation period in
pure Pb pipes (Edwards and McNeill, 2002).
The length of time that drinking water remains in a pipe also affects the water-Pb
concentration. Thus, a first flush phenomenon is generally observed in the morning after water
has stayed in the pipe through the night. An estimated 47% of total leached Pb was observed in
3-36
-------
the first 500 mL of water after prolonged stagnation (Singh and Mavinic, 1991). Gardels and
Sorg (1989) demonstrated that 60 to 75% of total Pb leached appeared in the first 125 mL of
water after prolonged stagnation. For cold water, the peak Pb concentrations occurred in the first
or second 25 mL sample and decreased exponentially with time thereafter. For hot water, the
peak Pb concentrations occurred in the second or third 25 mL sample before decreasing
exponentially (Gardels and Sorg, 1989). In a system where fully flushed water had a Pb content
of 1.7 ng/L, removing just 125 mL of water from the tap every hour kept Pb concentrations
elevated (35 to 52 |ig/L) throughout the day (Gulson et al., 1997). Lytle and Schock (2000)
showed a temporary exponential increase in Pb concentration with stagnation time before the
rate leveled off. After 10 hours of stagnation, -50 to 70% of the maximum Pb concentration had
been achieved, although Pb levels continued to increase even after 90 hours of stagnation. Their
results are shown in Figures 3-8 and 3-9. It should be noted that the shape of the stagnation-
concentration curves was the same for all situations regardless of water quality.
Some examples of Pb concentrations in drinking water are shown in Table 3-10. The Pb
standard for drinking water was set by the U.S. EPA in 1988, with a maximum allowable limit of
5 |ig/L for water entering the distribution system (Frey, 1989). Longitudinal observations
suggest that temporal variation is small for individual households compared to between-home
variation (Clayton et al., 1999).
Lead in drinking water can be either in paniculate or soluble form. Lead can be in the
form of aqueous ions or complexes, particularly when pH is low. Solids are the product of
nonadherent corrosion deposits, eroded pieces of plumbing material, or background inputs from
the distribution system (Lytle et al., 1993). Lead particles are released when pH and alkalinity
are low, and they typically occur in the form of hydrocerrusite scales (McNeill and Edwards,
2004). The main Pb products of corrosion include: CaCO3; PbCO3; Pb3(CO3)2(OH)2;
PblO(CO3)6(OH)6O; Pb5(PO4)3OH; and PbO (Lytle et al., 1993; McNeill and Edwards, 2004).
Based on the conditions described above, models to predict drinking water Pb concentrations
have been proposed (e.g., Clement et al., 2000; Van Der Leer et al., 2002). Lead in water,
although it is typically found at low concentrations in the United States, has been linked to
elevated blood-Pb concentrations. In a study of mothers and infants in Glasgow, Scotland tap
water was the main correlate of raised maternal blood-Pb levels (Watt et al., 1996). In a
prospective study, children exposed to water with Pb concentrations greater than 5 ppb had
3-37
-------
1,00
0,90
OJO
I 0.50
TJ
J
0.10
0.10
0.90
CQ -C- Agerf4»-I62dl¥i
15
T}
2 S
0 10 20 3fl 40 50 W 70 §0 90
Stagnation time I'h)
Impact of stagnation time on lead and dissolved oxygen concentration in lead pipe
(13 mm diameter) exposed to softened water in Study A.
Figure 3-8. The change in lead concentration versus stagnation time. (Reprinted from
Lytle and Schock, 2000).
blood-Pb levels -1.0 |ig/dL higher than children with water-Pb levels less than 5 ppb (Lanphear
et al., 2002). Thus, water may not be a trivial source of Pb under some conditions. The 1991
EPA Lead and Copper Rule requires that public water utilities conduct monitoring of Pb from
customer taps - generally every six months, annually, or triennially, depending on the levels of
Pb levels observed in drinking water. Less frequent monitoring is required if levels are low.
The rule established a tap water limit ("action level") of 0.015 mg/L (15 ppb) for Pb, based
on the 90th percentile concentration, above which corrective action is required (see
http://www.epa.gov/safewater/lcrmr/implement.html). The Safe Drinking Water Information
3-38
-------
LOO
0,90
OJO
OJO
§
•o
J
OJO
0.20
0.10
o.oo H
Aged 161-184 days
Aged 455-462 days
2 9
yft
5
10 20 .10 40 50 SO
Stagnation unw (hi
80 90
Impact of stagnation time on lead and dissolved oxygen concentration in lead pipe (13 mm diameter)
exposed to non-softened water in Study A.

Figure 3-9. Change in lead concentration versus stagnation time. (Reprinted from Lytle
and Schock, 2000).
System/Federal Version (SDWIS/FED) maintains a database to which public water utilities are
required to submit monitoring data. States have been required to report to EPA the 90th
percentile Pb concentrations reported by water systems serving more than 3,300 people. The
data available up through 2005 show that about 96% of the utilities that monitored and reported
90th percentile results are below the action level (see http://www.epa.gov/safewater/lcrmr/
lead_data.html). For illustrative purposes, Table 3-10 shows 90th percentile drinking water Pb
concentrations for a selection of large U.S. cities reported in 1992, 1993, and in more recent
years. These are examples of high tap water concentrations that exceed the action level for Pb in
tap water and are thusly notably higher than the mean Pb concentrations reported in Table 3-11.
3-39
-------
Table 3-10. Examples of Tap Water Concentrations of Lead
Water Concentration
(Hg/L)
20
13
0.70
0.32
16
8
3
2
6
5
17
7.7
25.0
15.3
11.6
3.92
0.84
Location
Vancouver, Canada
Vancouver, Canada
Arizona
Mexico/US border
Halifax, Canada
Halifax, Canada
Halifax, Canada
Halifax, Canada
Halifax, Canada
Halifax, Canada
Dharan, Saudi Arabia
Clinton, MA
Gardner, MA
Fall River, MA
New Bedford, MA
Midwest, US
Midwest, US
Residence Type
Apartments
Houses
Residences
Residences
Houses
Houses
Apartments
Apartments
Public School
Public School
Community sites
Residences
Residences
Residences
Residences
Residences
Residences
Description
copper or plastic pipes
copper or plastic pipes
—
—
standing water
running water
standing water
running water
standing water
running water
PVC pipes
standing water
standing water
standing water
standing water
standing water
flushed water
Reference
Singh and Manivic (1991)
Singh and Manivic (1991)
Sofuoglu et al. (2003)
Sofuoglu et al. (2003)
Moiretal. (1996)
Moiretal. (1996)
Moiretal. (1996)
Moiretal. (1996)
Moiretal. (1996)
Moiretal. (1996)
Sadiqetal. (1997)
Isaac etal. (1997)
Isaac etal. (1997)
Isaac etal. (1997)
Isaac etal. (1997)
Clayton etal. (1999);
Thomas etal. (1999)
Clayton etal. (1999);
Thomas etal. (1999)
-------
Table 3-11. 90* Percentile Tap Water Lead Concentrations for a Selection of U.S. Cities
Exceeding the EPA Pb Action Level
State
AZ
DC
FL
IA
IL
MI
MN
MN
NJ
NY
NY
OH
OR
PA
SC
TX
VA
WA
Water system
Phoenix Municipal
Washington Aqueduct
Miami Beach, City of
Cedar Rapids
Chicago
Detroit
Minneapolis
St. Paul
Bayonne Water Dept.
Syracuse
Yonkers
Columbus
Portland
Philadelphia Water Dept.
Columbia, City of
Galveston
City of Richmond
Tacoma
90th %ile
(ppb)
1993
19
18
27
80
10
21
19
54
18
50
68
15
41
322
40
18
16
32
90th %ile
(ppb)
1992
11
39
4
42
20
15
32
28
25
40
110
16
53
15
114
6
25
17
90th %ile
(ppb)
Recent
1
63
8
6
7
12
6
11
18
25
18
1
8
13
6
2
4
12
Recent Monitoring
Period
1/1/2003
7/1/2003
1/1/2001
1/1/2003
1/1/1999
1/1/2002
1/1/2002
1/1/2003
7/1/2001
1/1/2003
1/1/2003
1/1/2002
7/1/2003
1/1/2002
1/1/2002
1/1/2000
1/1/2000
1/1/2001
-12/31/2003
-12/31/2003
-12/31/2001
-12/31/2003
-12/31/2001
-12/31/2002
-12/31/2002
-12/31/2003
-12/31/2001
- 6/30/2003
- 6/30/2003
-12/31/2002
- 6/30/2006
-12/31/2002
-12/31/2002
-12/31/2002
-12/31/2002
-12/31/2003
3.4 EXPOSURE: DIETARY INTAKE
Although notable reductions of Pb in U.S. market basket food supplies have occurred
during the past several decades, Pb exposure via consumption of food and beverages can still be
a major route of exposure for some groups. As is true for Pb exposure via inhalation, Pb
exposure via ingestion has also decreased in the U.S. population. In general, food Pb
concentrations have decreased as a direct result of the decrease in airborne emissions of Pb from
automotive gasoline, as well as the reduction in the use of Pb solder in cans. In a detailed study
of Pb ingestion in food, Flegal et al. (1990) showed that North Americans ingested an estimated
3-41
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50 jig of Pb each day through food, beverages, and dust; and -30 to 50% of this amount was
from food and beverages. In 1987, the global average daily intake of Pb was estimated to be
about 80 jig/day from food and 40 ng/day from drinking water, according to estimates made by
the UN Environment Program (Juberg et al., 1997).
More recent data from the Food and Drug Administration's Total Diet Studies show that
estimates of daily Pb intake from food dropped substantially between 1982-1984 and 1994-1996
(Egan et al., 2002). In the 1994-1996 Total Diet Study, 74% of samples were found to be below
the detection limit for Pb, and daily intake values were presented as ranges reflecting different
methods being used to account for measurements below the limit of detection. Infants aged 6-11
months had the lowest estimated dietary Pb intakes; in 1994-1996, intake estimates ranged from
0.8 to 5.7 |ig/day (Egan et al., 2002), while in 1982-1984, the average estimated intake was
16.7 |ig/day (Gunderson et al., 1988). Similarly, children aged 2 years had estimated dietary Pb
intakes in the range of 2.4 to 10.1 ng/day in 1994-1996 (Egan et al., 2002), and Pb intake from
the diet was estimated to be larger in 1982-1984, with a mean value of 23.0 |ig/day (Gunderson
et al., 1988). For older children and adults, dietary Pb intakes in 1994-1996 were in the range of
4 to 19 jig/day (Egan et al., 2002); average dietary intake values in 1982-1984 for adults and
children older than 13 years ranged from 28.7 to 41.3 |ig/day (Gunderson et al., 1988).
Similar intake levels were observed in a study of children and their mothers in Omaha,
Nebraska, where the estimated rate of Pb ingestion was 1.8 |ig/day, 3.3 |ig/day, 4.1 |ig/day and
7.5 jig/day, for age groups of 0 to 12 months, 13 to 24 months, 2 to 6 years, respectively
(Manton et al., 2005). In this study, the authors observed that much of the Pb in diet was derived
from contamination of foods by household dust (Manton et al., 2005). In one report from the
NHEXAS study in Maryland, the mean intake of Pb in the diet was 7.6 |ig/day (Scanlon et al.,
1999); a subsequent analysis in this study reported a daily dietary intake of 8.14 |ig/day (Ryan
et al., 2001). The accompanying longitudinal study showed that Pb dietary exposures vary little
over time (Scanlon et al., 1999). In the Midwest, Pb concentrations in foods consumed by
children 0 to 6 years old were similar or lower than adults, but on a body weight basis Pb intake
rates were 1.5 to 2.5 times higher for young children (0.26 |ig/kg body weight/day for children
0 to 7 yrs and 0.10 |ig/kg body weight for people overall) (Thomas et al., 1999). Overall, a small
percentage of the population exceeded health-based intake levels set by FAO/WHO (Thomas
et al., 1999). In Australia, women between 20 and 39 years of age ingest between 7.3 and
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9.7 jig/day (Gulson et al., 200la). Infants that are breast-fed take in -0.73 ng/day compared to
1.8 jig/day for formula-fed infants (Gulson et al., 2001a). Australian children ingest
-6.4 jig/day. Overall, recent studies conducted in the United States indicate that daily Pb intake
from the diet ranges from about 1 to 10 jig/day. Some researchers have estimated the
contribution of Pb from sources other than food in the diet. For example, Melnyk et al. (2000)
estimated a daily Pb intake of 8.4 |ig/day in children based on a diet survey, but when they
estimated exposure due to the handling of food by children (including Pb from the floor and
house dust), the daily Pb intake from ingestion was estimated at 29.2 |ig/day.
Potential dietary exposure to Pb can be influenced by a range of factors. Nutrition status
and fasting can affect absorption rates. In a study of adult men, increased dietary vitamin D was
found to decrease Pb concentration in bone, while increased dietary vitamin C intake was
associated with decreased Pb concentrations in blood (Cheng et al., 1998). Fasting conditions
have been shown to increase Pb absorption dramatically (Rabinowitz et al., 1980).
Some recent exposure studies have evaluated the relative importance of diet to other
routes of Pb exposure. In reports from the NHEXAS, Pb concentrations measured in households
throughout the Midwest were significantly higher in solid food compared to beverages and tap
water (Clayton et al., 1999; Thomas et al., 1999). However, beverages appeared to be the
dominant dietary pathway for Pb according to the statistical analysis (Clayton et al., 1999),
possibly indicating greater bodily absorption of Pb from liquid sources (Thomas et al., 1999).
Dietary intakes of Pb were greater than those calculated for intake from home tap water or
inhalation on a |ig/day basis (Thomas et al., 1999). The NHEXAS study in Arizona showed that,
for adults, ingestion was a more important Pb exposure route than inhalation (O'Rourke et al.,
1999). Egeghy et al. (2005) did not find a significant association between blood Pb
concentration and Pb measured in any of the environmental media in the NHEXAS in Maryland;
however, the authors note that the short time frame of environmental sample collection was
likely not long enough to reflect the long half-life of Pb in the body.
It should be noted that concentrations in food can be very low and are frequently below
detection limits. In one study in New Jersey, average Pb concentrations measured in various
solid foods and beverages were 20 |ig/kg and 3.1 |ig/kg, respectively (Melnyk et al., 2000).
Examples of Pb concentrations measured in several foods are shown in Table 3-12. In the Food
and Drug Administration's Total Diet Study, dietary intake of Pb was distributed among grains,
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Table 3-12. Examples of Lead Concentrations in Food Products
Food
Concentration
Location
Description
Reference
Barley, grain
Barley, grain
Potato tubers, peeled
Potato tubers, peeled
Lettuce
Spinach
Potatoes
Wheat
Rice
Sweet corn
Field corn
Carrots
Onions
Tomatoes
Peanuts
Soybeans
Applesauce, canned
Fruit cocktail, canned
Spinach, fresh
Peaches, canned
0.4 ppm
2.0 ppm
0.21 ppm
0.89 ppm
0.19 ppm
0.53 ppm
0.03 ppm
0.02 ppm
0.01 ppm
0.01 ppm
0.01 ppm
0.05 ppm
0.04 ppm
0.03 ppm
0.01 ppm
0.04 ppm
8.5 ug/serving
7.1 ug/serving
2.4 ug/serving
6.0 ug/serving
Uncontaminated soil
Zn-Pb smelter contaminated
Uncontaminated soil
Zn-Pb smelter contaminated
Edible portion, untreated soil
Edible portion, untreated soil
Edible portion, untreated soil
Edible portion, untreated soil
Edible portion, untreated soil
Edible portion, untreated soil
Edible portion, untreated soil
Edible portion, untreated soil
Edible portion, untreated soil
Edible portion, untreated soil
Edible portion, untreated soil
Edible portion, untreated soil
FDA Total Diet Study
FDA Total Diet Study
FDA Total Diet Study
FDA Total Diet Study
Dudka and Miller (1999)
Dudka and Miller (1999)
Dudka and Miller (1999)
Dudka and Miller (1999)
Dudka and Miller (1999)
Dudka and Miller (1999)
Dudka and Miller (1999)
Dudka and Miller (1999)
Dudka and Miller (1999)
Dudka and Miller (1999)
Dudka and Miller (1999)
Dudka and Miller (1999)
Dudka and Miller (1999)
Dudka and Miller (1999)
Dudka and Miller (1999)
Dudka and Miller (1999)
Jubergetal. (1997)
Jubergetal. (1997)
Jubergetal. (1997)
Jubergetal. (1997)
-------
Table 3-12 (cont'd). Examples of Lead Concentrations in Food Products
Food
Concentration
Location
Description
Reference
Pears, canned
Strawberries, fresh
Apple juice, bottled
Wine
Vaccinium vitis-idaea
Vaccinium myrtillus
Rubus chamaemoms
Empetmm hermaphroditum
Leccinum auranticcum
Leccinum sacbrum
Russul vesea
Xerocomus subtomentosus
Suillus luteus
Lactarius trivialis
Lactarius torminosus
Lettuce
Lettuce
Lettuce
Carrots
Carrots
4.9 ng/serving
1 . 1 ug/serving
2.6 ug/serving
7.7 ug/serving
0.4-2.3 ppm
0.7-1.6 ppm
0.3-4.7 ppm
0.3-1.5 ppm
0.8-2.3 ppm
1.1-5.2 ppm
1.1-3.4 ppm
1.3-3.1 ppm
2.0-2.3 ppm
1.1-3.1 ppm
0.6-3. 5 ppm
0.65-1. 3 ppm
0.15-0.46 ppm
0.36 ppm
0.07-0.28 ppm
<0.02-0.09 ppm
Monchegorsk, Russia
Monchegorsk, Russia
Monchegorsk, Russia
Monchegorsk, Russia
Monchegorsk, Russia
Monchegorsk, Russia
Monchegorsk, Russia
Monchegorsk, Russia
Monchegorsk, Russia
Monchegorsk, Russia
Monchegorsk, Russia
Copenhagen, Denmark
Copenhagen, Denmark
Copenhagen, Denmark
Copenhagen, Denmark
Copenhagen, Denmark
FDA Total Diet Study
FDA Total Diet Study
FDA Total Diet Study
FDA Total Diet Study
Berry, near Ni-Cu smelter
Berry, near Ni-Cu smelter
Berry, near Ni-Cu smelter
Berry, near Ni-Cu smelter
Mushroom, near Ni-Cu smelter
Mushroom, near Ni-Cu smelter
Mushroom, near Ni-Cu smelter
Mushroom, near Ni-Cu smelter
Mushroom, near Ni-Cu smelter
Mushroom, near Ni-Cu smelter
Mushroom, near Ni-Cu smelter
Close to lead smelter
Far from lead smelter
Background concentration
Close to lead smelter
Far from lead smelter
Jubergetal. (1997)
Jubergetal. (1997)
Jubergetal. (1997)
Jubergetal. (1997)
Barcanetal. (1998)
Barcanetal. (1998)
Barcanetal. (1998)
Barcanetal. (1998)
Barcanetal. (1998)
Barcanetal. (1998)
Barcanetal. (1998)
Barcanetal. (1998)
Barcanetal. (1998)
Barcanetal. (1998)
Barcanetal. (1998)
Moseholm et al. (1992))
Moseholm et al. (1992))
Moseholm et al. (1992))
Moseholm etal. (1992))
Moseholm etal. (1992))
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Table 3-12 (cont'd). Examples of Lead Concentrations Food Products
Food
Concentration
Location
Description
Reference
Carrots
Potatoes
Potatoes
Potatoes
Kale
Kale
Kale
Wine
Breast milk
Infant formula
Baby food
Brassica juncea
Triticum aestivum L.
Basil
Cabbage
Cilantro
Collard greens
Coriander
Ipasote
Lemon balm
0.02-0.03 ppm
<0.02-0.12ppm
O.02-0.06 ppm
O.02 ppm
1.4-9.3 ppm
0.58-2.4 ppm
0.52-0.72 ppm
65ug/L
0.55 ug/kg
1.6 ug/kg
2.9 ug/kg
298.3 ppm
19.2 ppm
<10 ppm
<10 ppm
49 ppm
12 ppm
39 ppm
14 ppm
20 ppm
Copenhagen, Denmark
Copenhagen, Denmark
Copenhagen, Denmark
Copenhagen, Denmark
Copenhagen, Denmark
Copenhagen, Denmark
Copenhagen, Denmark
France
Australia
Australia
Australia
Taihe, China
Taihe, China
Chicago, IL
Chicago, IL
Chicago, IL
Chicago, IL
Chicago, IL
Chicago, IL
Chicago, IL
Background concentration
Close to lead smelter
Far from lead smelter
Background concentration
Close to lead smelter
Far from lead smelter
Background concentration
Vintage 1990-1995
Indian mustard, near lead smelter
Common wheat, near lead smelter
Edible portion, urban garden
Edible portion, urban garden
Edible portion, urban garden
Edible portion, urban garden
Edible portion, urban garden
Edible portion, urban garden
Edible portion, urban garden
Moseholmetal. (1992))
Moseholmetal. (1992))
Moseholmetal. (1992))
Moseholmetal. (1992))
Moseholmetal. (1992))
Moseholmetal. (1992))
Moseholmetal. (1992))
Medina et al. (2000)
Gulsonetal. (200 Ib)
Gulsonetal. (200 Ib)
Gulsonetal. (200Ib)
Cui et al. (2003)
Cui et al. (2003)
Finster et al. (2004)
Finster et al. (2004)
Finster et al. (2004)
Finster et al. (2004)
Finster et al. (2004)
Finster et al. (2004)
Finster et al. (2004)
-------
Table 3-12 (cont'd). Examples of Lead Concentrations in Food Products
Food
Mint
Mustard greens
Parsley
Red chard
Rhubarb
Sage
Swiss chard
Thyme
Carrot
Onion
Radish
Tuna, canned
Sardines, canned
Blue mussel, canned
Balsamic vinegar
Wine vinegar
Tea leaves
Cocoa beans
Cocoa, manufactured
Chocolate products
Concentration
<10 - 60 ppm
<10 ppm
<10 ppm
<10 ppm
<10 - 36 ppm
<10 ppm
22-24 ppm
<10 ppm
10 ppm
21 ppm
12-18 ppm
0.1 ppm (max.)
0.2 ppm (max.)
0.3 ppm (max.)
15-307 ug/L
36-50 ug/L
0.59-4.49 ppm
0.5 ng/g
230 ng/g
70 ng/g
Location
Chicago, IL
Chicago, IL
Chicago, IL
Chicago, IL
Chicago, IL
Chicago, IL
Chicago, IL
Chicago, IL
Chicago, IL
Chicago, IL
Chicago, IL

Zhejiang Province, China
Nigeria
Nigeria
Nigeria
Description
Edible portion, urban garden
Edible portion, urban garden
Edible portion, urban garden
Edible portion, urban garden
Edible portion, urban garden
Edible portion, urban garden
Edible portion, urban garden
Edible portion, urban garden
Edible portion, urban garden
Edible portion, urban garden
Edible portion, urban garden

Commercial tea producing areas

Reference
Finster et al. (2004)
Finster et al. (2004)
Finster et al. (2004)
Finster et al. (2004)
Finster et al. (2004)
Finster et al. (2004)
Finster et al. (2004)
Finster et al. (2004)
Finster et al. (2004)
Finster et al. (2004)
Finster et al. (2004)
Lourenco et al. (2004)
Lourenco et al. (2004)
Lourenco et al. (2004)
Ndung'u et al. (2004)
Ndung'u et al. (2004)
Jin et al. (2005)
Rankin et al. (2005)
Rankin et al. (2005)
Rankin et al. (2005)
-------
fruits, mixtures and sweets; dairy foods contributed at least 10% of Pb intake by infants and
children to age 16, and beverages contributed between 8 and 18% of Pb intakes by adults (Egan
et al., 2002). Ryan et al. (2001) assessed the contribution from various foods to dietary Pb
intake, and found that numerous different items, including both canned and fresh fruits and
vegetables, cereal and hamburger, were predictive of blood-Pb concentrations. Using the
Dietary Exposure Potential Model, Moschandreas et al. (2002) reported that drinking water,
coffee and tea (likely reflecting Pb from the drinking water source) were major contributors to
estimated daily dietary Pb intake.
One review article concluded that, since the elimination of Pb solder in U.S. canned food,
the primary source of Pb in U.S. food is atmospheric deposition (Flegal et al., 1990). Overall,
anthropogenic aerosols account for an estimated 40% of Pb in food, while the bulk of the
remainder is derived from harvesting, transporting, processing, packaging, or preparing the food
(Flegal et al., 1990; Juberg et al., 1997; Dudka and Miller, 1999). Lead contamination in poultry
and livestock is also primarily atmospheric in origin. Lead deposits on forage or feed or onto
soil that is directly ingested (Flegal et al., 1990). Lead concentrations in food have been reported
to increase by a factor of 2 to 12 between harvest and consumption (Flegal et al., 1990). A food
production facility in Turkey was shown to contaminate pasta with Pb (Demirozii and Saldamli,
2002), as indicated by Pb concentrations in the semolina being 14.2 to 36.5 ng/g compared with
the finished pasta product where concentrations ranged from 107.1 to 147.6 ng/g (Demirozii and
Saldamli, 2002). An increase (from an average value <0.5 ng/g to average values between 11.9
and 69.8 ng/g) between raw and finished cocoa products has also been observed (Rankin et al.,
2005). In this case, contamination seems to occur during shipping and/or processing.
Other significant sources of dietary Pb are calcium-supplemented food where calcium is
derived from limestone and tin coatings that contain Pb.
Lead concentrations in vegetables may be increased by soil amendments such as mine
wastes, slag, or fly ash. Historically, mine tailings were often disposed in streambeds, and this
poses an exposure risk when stream sediments are used to boost productivity in gardens (Cobb
et al., 2000). Slag is sometimes used for constructing agricultural and forestry roads or for
landfill. This can be an additional source of Pb contamination for nearby crops (Bunzl et al.,
2001). Fly ash is applied to land infrequently for alkaline adjustment, as cover for landfills, or to
amend agricultural soils. Elevated Pb levels in fly ash can subsequently contaminate crops
5-48
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(Brake et al., 2004). Although soil contamination may be important on a local scale, overall
atmospheric deposition is a more significant source of food Pb than uptake from soil. For
example, more than 52% of the total Pb present in citrus fruits was removed by washing,
indicating that surface deposits make up the bulk of Pb contamination in unprocessed fruit
(Caselles, 1998).
3.5 LEAD-BASED PAINT
Lead-based paint was the dominant form of house paint for many decades, and a
significant percentage of homes still contain Pb-based paint on some surfaces. As discussed in
previous sections, Pb-based paint can be an important source of Pb in house dust and soil, thus
contributing to human exposure through those routes. Lead-based paint also poses a potential
exposure risk due to inhalation during renovation or demolition projects, or due to ingestion from
hand-to-mouth activities and pica, which are common in children. Lead-based paint exposure is
one of the most common causes of clinical Pb toxicity.
In a 1970 study, it was observed that for children with blood-Pb levels >50 |ig/dL, more
than 80% were reported to ingest paint chips or broken plaster (Sachs et al., 1970). A later study
by McElvaine et al. (1992) showed that children with blood-Pb levels above 55 |ig/dL were ten
times more likely to have paint chips observable on abdominal radiographs than children with
blood-Pb below this value. Shannon and Greaf (1992) noted that the majority of preschool-aged
children with blood-Pb over 25 |ig/dL were reported to put paint chips in their mouth.
As Pb-based paint degrades, it becomes incorporated into house dust, which was
discussed in depth earlier in this chapter. Lead-based paint can pose an inhalation risk during
renovation and demolition activities. As described in Section 3.1 of this document, renovation
projects often involve abrasive blasting techniques to remove old layers of paint. This forms Pb
particles that are easily inhaled (Chute and Mostaghim, 1991; Mickelson and Johnston, 1995;
Jacobs, 1998; Mielke et al., 2001). At industrial sites, exposure is limited primarily to workers.
However, during residential renovation or abatement projects, residents may be unduly exposed
to very high levels of airborne Pb. Blood-Pb levels were shown to increase in children who lived
in houses with a significant (>1.5 mg/cm2 on at least one surface) amount of Pb paint that had
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undergone some sanding, scraping, or other indoor surface refinishing in the preceding six
months (Rabinowitz et al., 1985b).
Additionally, exterior Pb-based paint can degrade and contaminate nearby soil. Paint and
surface soil samples collected in and around Oakland, CA households of children having
elevated blood-Pb levels had nearly identical isotopic ratios as the children's blood (Yaffe et al.,
1983). This suggests that weathered, Pb-based paint was a major exposure route for these
children who played outside near the most highly contaminated areas.
3.6 OTHER ROUTES OF EXPOSURE
3.6.1 Calcium Supplements
Potentially toxic levels of Pb were measured in calcium supplements in studies
undertaken in the 1960s through the early 1990s (Scelfo and Flegal, 2000). An analysis of
136 different brands of supplements showed that two-thirds of the supplements did not meet the
1999 California criteria for acceptable Pb levels: 1.5 jig Pb/daily dose of calcium (Scelfo and
Flegal, 2000). The lowest concentrations were observed in calcium products that were
nonchelated synthesized and/or refined. These corresponded to antacids and infant formulas.
Antacids and infant formulas had Pb concentrations ranging from below the detection limit to
2.9 jig Pb/g calcium (Scelfo and Flegal, 2000). Natural calcium supplements derived from
bonemeal, dolomite, or oyster shell were much more likely to be in exceedance of the 1999
standard. Pb levels reported elsewhere showed comparable Pb levels in supplements and cow
milk (Juberg et al., 1997). Whole milk, 2% milk, and calcium supplements had Pb
concentrations in the range of 1.7 to 6.7 jig Pb/g calcium, 0.8 to 9.0 jig Pb/g calcium, and 3.1 to
6.9 jig Pb/g calcium, respectively (Juberg et al., 1997). A clamshell powder commonly known
as hai gen fan that is added to tea has shown detectable levels of Pb contamination as well (CDC,
1999).

3.6.2 Glazes
Lead glazes have been commonly used throughout history. Kitchen glassware cannot
have a Pb solubility in excess of 2.5 to 7 |ig/mL according to a 1980 rule by the U.S. Food and
Drug Administration (Flegal et al., 1990). However, Pb glazes on imported pottery may persist.
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Foods with low pH (acidic ones) are particularly susceptible to solubilizing Pb and thereby
contaminating food during storage in Pb glazed glassware. Lead glazes may be especially
problematic when low temperature fluxes and glazes are used. This is more common in
traditional charcoal kilns than gas-fired kilns.

3.6.3 Miniblinds
Some imported vinyl miniblinds form Pb dust upon disintegration (Juberg et al., 1997).
This exposure route was responsible for several cases of Pb poisoning in Arizona and North
Carolina in the mid 1990s. Lead stabilizers are not used in vinyl miniblinds manufactured in the
United States (Juberg et al., 1997).

3.6.4 Hair Dye
The analysis of Mielke et al. (1997) shows that some hair dyes contain Pb acetate in the
range of 2300 to 6000 jig Pb/g of product. This Pb can be easily transferred via hand-to-mouth
and hand-to-surf ace activity, and an estimated 3 to 5% of Pb acetate can be transferred through
the skin. Hair dyes tested in this study contain 3 to 10 times more Pb than is allowable for paint
(Mielke et al., 1997).

3.6.5 Other Potential Sources of Lead Exposure
Additional consumer products that may pose a risk of Pb exposure include Pb crystal,
pool cue chalk (Miller et al., 1996), cosmetics, and folk remedies, which purposefully contain Pb
(such as alarcon, alkohl, azarcon, bali goli, coral, gliasard, greta, kohl, KooSo or KooSar pills,
liga, pay-loo-ah, rueda, and surma of East Indian, Pakistani, Chinese, and Latin American
origins) (CDC, 1999). Unintentional or malicious Pb contamination has also been found for the
following products: ground paprika, ayurvedic metal-mineral tonics, and Deshi Dewa (a fertility
drug) (CDC, 1999; Kakosky et al., 1996).
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3.7 MEASUREMENT METHODS
Methods for measurement of Pb in environmental media were discussed extensively in the
1986 Lead AQCD, and the reader is referred to that document for details regarding the main
methods employed and associated detection limits. Some of the most commonly employed
methods are concisely noted below.
The concentration of Pb in air can be measured through several different methods. Use of
filter media and inertial impactors are two of the most common methods, and in both cases
particles can be separated by size. An additional method involves mounting a particle separation
device in the stack along with gas flow control and metering equipment. The collected particles
are then analyzed for mass and Pb content (Clarke and Bartle, 1998).
Sampling of airborne particles to determine concentration and chemical species can be
performed via direct-reading instruments, which include optical counters, electrical counters,
resonant oscillation aerosol mass monitors, and beta radiation detectors (Koutrakis and Sioutas,
1996). Additionally, particles may be collected in cyclones and denuder systems.
Collected particles can be analyzed for Pb using x-ray fluorescence analysis (XRF),
proton-induced x-ray emission (PIXE), neutron activation analysis (NAA), atomic absorption
(AA), or inductively-coupled plasma mass spectrometry (ICP-MS) (Koutrakis and Sioutas,
1996).
Lead concentrations in soil, dust, food, and other environmental media are determined
using similar techniques. Generally, substances undergo acid digestion in an HC1 or HNOs
solution before analysis via XRF, PIXE, NAA, AA, or ICP-MS. Special care should be taken in
all cases to avoid external contamination of samples, especially when measuring very low Pb
concentrations.
For detailed discussions of methods for determining Pb speciation and isotopic ratios, see
Chapter 7 of this document. Also, see Chapter 4 for information related to measurement of Pb in
biological tissues.
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3.8 SUMMARY
Lead concentrations in various environmental media have been discussed throughout this
chapter. Concentrations of Pb in all environmental media are present in detectable quantities.
In general, Pb exposure in the United States has fallen with the elimination of leaded gasoline,
Pb-based paint and Pb solder in cans. However, the potential for high Pb exposures remains,
particularly in areas near major Pb sources or with exposures to Pb-based paint or high Pb levels
in drinking water.
As airborne concentrations of Pb have fallen in the United States, a corresponding
decrease in blood-Pb levels of the U.S. population has been observed. In a meta-analysis of
19 studies from six continents, a strong linear correlation was observed between blood-Pb levels
and gasoline Pb levels (Thomas et al., 1999). As gasoline Pb was reduced to zero in the study
countries, airborne Pb concentrations declined and converged to less than 0.2 |ig/m3, and blood-
Pb levels also declined, converging to a median of 3 |ig/dL.
A U.S. exposure study, the National Human Exposure Assessment Survey (NHEXAS),
has included assessment of the relationship between blood-Pb concentrations and exposure
measurements in various media as well as activity or household variables. In the NHEXAS in
Arizona, total daily Pb intake was found to range from 11 to 107 |ig/day, with a mean of
36 |ig/day (O'Rourke et al., 1999).
The relative contribution of Pb different media to human exposure varies, particularly for
different age groups. In the NHEXAS-Arizona study of a largely adult population, food was the
predominant source of Pb exposure, followed by consumption of beverages and tap water; most
measurements of Pb in indoor air, outdoor air, soil and dust were below the detection limit
(O'Rourke et al., 1999). In a study of children 6 to 24 months of age, Pb in floor dust had the
largest association with PbB concentrations, followed by Pb in drinking water and Pb in soil
(Lanphear et al., 2002).
The highest air, soil, and road dust concentrations are found near major Pb sources, such
as smelters, mines, and heavily trafficked roadways. While airborne Pb concentrations have
declined dramatically with the phase out of leaded gasoline, soil concentrations have remained
relatively constant, reflecting the generally long retention time of Pb in soil. Soil-Pb
concentrations decrease both with depth and distance from roadways and sources such as
smelters or mines. In another study of 831 homes in the United States, 7% of housing units were
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found to have soil-Pb levels exceeding 1200 ppm, the U.S.EPA/HUD standard for soil-Pb
concentration outside of play areas (Jacobs et al., 2002).
Drinking water is susceptible to Pb contamination primarily through leaching from pipes,
solder, and faucets. Water that has been stagnant in pipes, has been disinfected with
chloramines, has a low pH, or has a low alkalinity is particularly high risk for leaching Pb.
Lead-contaminated food can be an important exposure route. Deposition of airborne Pb
and house dust are the major sources of Pb in food. Data from the Food and Drug
Administration's Total Diet Studies show that estimates of daily Pb intake from food has
dropped substantially in recent years; for example, estimated dietary Pb intake dropped 96
percent in 2- to 5-year old children (from 30 |ig/day to 1.3 jig/day) between 1982-1984 and
1994-1996 study periods (Egan et al., 2002). Across all age groups, estimated Pb intake ranged
from 0.8 to 19.6 jig/day, with the lowest intake estimates in infants aged 6-11 months (Egan
et al., 2002).
Lead-based paint is still prevalent in older homes. This can be a major exposure route if
paint has deteriorated or undergone careless renovation. Lead-based paint also can be an
important source of Pb exposure through soil or house dust.
Other sources of Pb exposure vary in their prevalence and potential risk. These include
calcium supplements, Pb-based glazes, certain types of miniblinds, hair dye, and other consumer
products.
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4. TOXICOKINETICS, BIOLOGICAL MARKERS, AND
MODELS OF LEAD BURDEN IN HUMANS
4.1 INTRODUCTION
The preceding two chapters presented important background information on the physical
and chemical properties of lead (Pb) and its inorganic and organic compounds, sources and
emissions of Pb into the ambient air and other environmental media, and concentrations of Pb in
ambient air and other components of multimedia human exposure pathways (e.g., water, food,
soil, exterior and interior/dusts, etc.). This chapter deals predominately with the relationship
between human Pb exposure and Pb burden in the body.
With exposure, mainly by ingestion and inhalation, a portion of Pb is absorbed and
distributed to various body compartments from which it is eliminated at various rates.
Conceptually, the body's Pb burden may be considered to be divided between a fast (soft tissues)
and a slow (skeletal) compartment. Lead in blood exchanges with both of these compartments.
The contribution of bone Pb to the blood Pb changes with the duration and intensity of the
exposure, age, and various physiological variables (e.g., nutritional status, pregnancy, and
menopause).
In Pb toxicologic and epidemiologic studies, dose-response relationships for nearly all of
the major health effects of Pb are typically expressed in terms of an index of internal Pb dose.
Blood Pb concentration is extensively used in epidemiologic studies as an index of exposure and
body burden mainly because of the feasibility of incorporating its measurement into human
studies relative to other potential dose indicators, e.g., Pb in kidney, plasma, urine, or bone.
Blood Pb is determined by both the recent exposure history of the individual, as well as the long-
term exposure history that leads to accumulation in bone. The benefits and limitations of blood
Pb concentration as an indicator of Pb body burden were discussed in Section 13.3.2 of the 1986
Lead AQCD. Application of internal dose-response information to the assessment of risks from
Pb exposures requires means for estimating the resultant internal doses. Approaches to
estimating external Pb exposure impacts on internal tissue concentrations, including various
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types of regression analyses and complex biokinetic modeling, are thusly topics of much
importance here.
This chapter begins by providing an overview of the toxicokinetics of Pb, focusing on our
current understanding of the routes of Pb exposure, uptake, distribution, and elimination in
humans. Next, there is a detailed discussion of biological markers used to assess human Pb
burden and exposure. Subsequently, models for assessing Pb exposure-burden relationships in
humans are presented. The modeling discussion begins with recent developments in
epidemiological models of Pb exposure-blood Pb concentration relationships in humans.
The evolution of Pb biokinetics modeling and other major modeling advances during the past
25 years or so is then discussed.
4.2 TOXICOKINETICS OF LEAD
Information on Pb toxicokinetics was extensively summarized in the 1986 Lead AQCD
(U.S. Environmental Protection Agenyc, 1986) and has been the subject of several recent
reviews (e.g., ATSDR, 2005; Mushak, 1991, 1993). Since the completion of the 1986 AQCD,
knowledge of the Pb toxicokinetics has advanced in several areas. For example, new studies
have been published on the kinetics of Pb movement into and out of bone (based on analysis of
stable Pb isotope profiles) that have demonstrated the importance of bone Pb stores as an internal
source of Pb to the blood, fetus, and nursing infant. New animal and human experimental
models have been developed for studying dermal and gastrointestinal (GI) bioavailability of Pb;
the latter studies have provided a more quantitative understanding of the GI bioavailability of Pb
in soils. Also, several new models of the toxicokinetics of Pb in humans have been developed
that incorporate simulations of bone growth and resorption in the distributional kinetics of Pb in
humans.
The summary provided below discusses the major features of Pb absorption, distribution,
metabolism, and excretion. Information specific to route of exposure (e.g., inhalation, oral,
dermal) is discussed under separate subsections. Distinguishing features of inorganic and
organic Pb (e.g., alkyl Pb compounds) are also discussed.
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4.2.1 Absorption of Lead
Inhalation Exposure
Inorganic Lead
Inorganic Pb in ambient air consists of aerosols of Pb-bearing particulate matter (PM).
Amounts and patterns of deposition of inhaled particulate aerosols in the respiratory tract are
affected by the size of the inhaled particles, age-related factors that determine breathing patterns
(e.g., relative contributions of nose and mouth breathing), airway geometry, and air-stream
velocity within the respiratory tract (James et al., 1994). Publicly available models can be used
to predict the deposition and clearance of Pb-bearing particles in the respiratory tract of children
and adults. Two such models currently in wide use were developed by (1) the International
Commission on Radiological Protection (ICRP, 1994) and (2) the CUT Centers for Health
Research (CUT), USA, in collaboration with the National Institute of Public Health and the
Environment (RIVM), the Netherlands, and with the Ministry of Housing, Spatial Planning and
the Environment, the Netherlands. The CUT model or the Multi-Path Particle Dosimetry
(MPPD, available at http://www.ciit.org/mppd) model has been described in detail elsewhere
(Anjilvel and Asgharian, 1995; Asgharian et al., 2004; Brown et al., 2005; de Winter-Sorkina
and Cassee, 2002). Both the ICRP and the MPPD model can be used to predict deposition for
particles of-0.01 jim to 20 jim in diameter. The reader is referred to ICRP (1994) for detailed
information on factors affecting particle deposition and clearance in the human lung as well as
on breathing patterns as a function of age and activity. 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. Fine 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 (Bailey and Roy, 1994). Larger particles
(>2.5 jim) that are primarily deposited in the ciliated airways (nasopharyngeal and
tracheobronchial regions) can be transferred by mucociliary transport into the esophagus and
swallowed.
Quantitative studies of the deposition and clearance of inhaled inorganic Pb in humans
have been limited to studies of adults and to exposures to relatively small particles (Chamberlain
et al., 1978; Hursh and Mercer, 1970; Hursh et al., 1969; Morrow et al., 1980; Wells et al.,
1977). In these studies, exposures were to Pb-bearing particles having mass median
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aerodynamic diameters (MMAD) below 1 jam. Deposition of inhaled Pb particles of this size
can be assumed to have been primarily in the bronchiolar and alveolar regions of the respiratory
tract (James et al., 1994), where mucociliary transport is likely to have been a relatively minor
component of particle clearance (Hursh et al., 1969), compared to the fate of larger particles.
Approximately 25% of inhaled Pb chloride or Pb hydroxide (MMAD 0.26 and 0.24 jim,
respectively) was deposited in the respiratory tract in adult subjects who inhaled an inorganic Pb
aerosol through a standard respiratory mouthpiece for 5 min (Morrow et al., 1980), whereas
-95% of deposited inorganic Pb inhaled as submicron particles was absorbed (Hursh et al., 1969;
Wells et al., 1977). Clearance from the respiratory tract of inorganic Pb inhaled as submicron
particles of Pb oxide or Pb nitrate was multiphasic, with approximate half-times of 0.8 (22%),
2.5 (34%), 9 (33%), and 44 h (12%) (Chamberlain et al., 1978). Given the submicron particle
size of the exposure, these rates are thought to represent, primarily, absorption from the
bronchi olar and alveolar regions of the respiratory tract. As noted previously, amounts and rates
of absorption of inhaled Pb particles that are larger than 2.5 |im, and which may be more typical
of certain human exposure scenarios, will be mainly determined by rates of mucociliary transport
to the GI tract.
While no quantitative studies of the deposition and absorption of inhaled Pb in children
have been reported, the behavior of Pb-containing particles in the respiratory tract can be
inferred from experimental studies of inert particle deposition in children and particle dosimetry
models. The effect of an individual's age on extrathoracic particle deposition is not well
characterized. For example, Bennett et al. (1997) reported -50% greater deposition of 4.5-|im
(MMAD, mass median aerodynamic diameter) particles in children relative to adults. Xu and Yu
(1986) also predicted increasing deposition of 2-|im MMAD particles with decreasing age.
In contrast, Becquemin et al. (1991) reported nasal deposition to increase with age for particles
between 1 and 3 |im MMAD. Both experimental and modeling studies show little difference in
thoracic deposition fraction between adults and children at natural resting breathing patterns
(Asgharian et al., 2004; Bennett and Zeman, 1998; Hofman et al., 1989; Phalen and Oldham,
2001). Of importance with regard to surface doses is that particle deposition is greater in the
tracheobronchial region and lower in the pulmonary region of children compared to adults
(Hofman et al., 1989; Phalen and Oldham, 2001). Normalized to lung volume, total deposition is
predicted to be greatest in infants and decreases with age for particle between 0.01 and 10 jam
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(Asgharian et al., 2004). This increase in deposition per unit lung volume was predominately
attributable to particle losses in the tracheobronchial airways. Bennett and Zeman (1998) also
noted that because children have smaller lungs and higher minute ventilation relative to lung
size, they would be expected to receive greater particle doses per lung surface area compared
to adults.

Organic Lead
Alkyl Pb compounds can exist in ambient air as vapors. Inhaled tetraalkyl Pb vapor is
nearly completely absorbed following deposition in the respiratory tract. Following a single
exposure to vapors of radioactive (203Pb) tetraethyl Pb (~1 mg/m3 breathed through a mouthpiece
for 1 to 2 min) in four male subjects, 37% of inhaled 203Pb was initially deposited in the
respiratory tract, of which -20% was exhaled in the subsequent 48 h (Heard et al., 1979). One
hour after the exposure, -50% of the 203Pb burden was associated with liver, 5% with kidney,
and the remaining burden widely distributed throughout the body, suggesting near complete
absorption of the Pb that was not exhaled. 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 h. The distribution of 203Pb 1 h after the exposure was similar to
that observed following exposure to tetraethyl Pb.

Oral Exposure
Inorganic Lead
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 and iron
status, pregnancy) and physicochemical characteristics of the Pb-bearing material ingested
(e.g., particle size, mineralogy, solubility, Pb species). Lead absorption may also vary with the
amount of Pb ingested.
Effect of Age. Gastrointestinal absorption of water-soluble Pb appears to be higher in
children than in adults. Estimates derived from dietary balance studies conducted in infants and
children (ages 2 weeks to 8 years) indicate that -40 to 50% of ingested Pb is absorbed
(Alexander et al., 1974; Ziegler et al., 1978). In adults, estimates of absorption of ingested
water-soluble Pb compounds (e.g., Pb chloride, Pb nitrate, Pb acetate) ranged from 3 o 10% in
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fed subjects (Heard and Chamberlain 1982; James et al., 1985; Rabinowitz et al., 1980; Watson
et al., 1986); absorption fraction in fasted adults is high (-50% increase, see Effect of Fasting).
Data available on Pb absorption between childhood and adulthood ages are very limited. While
no absorption studies have been conducted on subjects in this age group, the kinetics of the
change in stable isotope signatures of blood Pb in mothers and their children as both come into
equilibrium with a novel environmental Pb isotope profile, suggests similar ingested Pb
absorption fractions in children (ages 6 to 11 years) and their mothers (Gulson et al., 1997).
The mechanisms for an apparent age difference in GI absorption of inorganic Pb have not been
completely elucidated and may include both physiological and dietary factors (Mushak, 1991).
Studies in experimental animals provide additional evidence for an age-dependency of GI
absorption of Pb; however, not all studies have completely segregated possible age-differences in
absorption from age differences in elimination of absorbed Pb. Absorption of Pb, administered
as Pb acetate (6.37 mg Pb/kg, oral gavage), was higher in juvenile rhesus monkeys (38% of
dose) compared to adult female monkeys (26% of the dose) (Pounds et al., 1978). Rat pups
absorb -40 to 50 times more Pb via the diet than do adult rats (Aungst et al., 1981; Forbes and
Reina, 1972; Kostial et al., 1978). This age difference in absorption may be due, in part, to the
shift from the neonatal to adult diet, and to postnatal physiological development of the intestine
(Weis and LaVelle, 1991).
Effect of Fasting. The presence of food in the GI tract decreases absorption of water-
soluble Pb (Blake and Mann, 1983; Blake et al., 1983; Heard and Chamberlain, 1982; James
et al., 1985; Maddaloni et al., 1998; Rabinowitz et al., 1980). In adults, absorption of a tracer
dose of Pb acetate in water was -63% when ingested by fasted subjects and 3% when ingested
with a meal (James et al., 1985). Heard and Chamberlain (1982) reported nearly identical
results. The arithmetic mean of reported estimates of absorption in fasted adults was 57% (based
on Blake et al., 1983; Heard and Chamberlain, 1982; James et al., 1985; Rabinowitz et al., 1980).
Reported fed:fasted ratios for absorption in adults range from 0.04 to 0.2 (Blake et al., 1983;
Heard and Chamberlain, 1983; James et al., 1985; Rabinowitz, et al., 1980). Mineral content is
one contributing factor to the lower Pb absorption when Pb is ingested with a meal; i.e., the
presence of calcium and phosphate in a meal depresses absorption of ingested Pb (Blake and
Mann, 1983; Blake et al., 1983; Heard and Chamberlain, 1982).
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Effect of Nutrition. Lead absorption in children is affected by nutritional iron status.
Children who are iron-deficient have higher blood Pb concentrations than similarly exposed iron-
replete children, suggesting that iron deficiency may result in higher Pb absorption or, possibly,
other changes in Pb biokinetics that contribute to altered blood Pb concentrations (Mahaffey and
Annest, 1986; Marcus and Schwartz, 1987). Evidence for the effect of iron deficiency on Pb
absorption has been derived from animal studies. In rats, iron deficiency increases the GI
absorption of Pb, possibly by enhancing binding of Pb to iron-binding proteins in the intestine
(Bannon et al., 2003; Barton et al., 1978a; Morrison and Quatermann, 1987).
Dietary calcium intake also appears to affect Pb absorption. An inverse relationship has
been observed between dietary calcium intake and blood Pb concentration in children,
suggesting that children who are calcium-deficient may absorb more Pb than calcium-replete
children (Mahaffey et al., 1986; Ziegler et al., 1978). An effect of calcium on Pb absorption is
also evident in adults. In experimental studies of adults, absorption of a single dose of Pb (100 to
300 jig Pb chloride) was lower when the Pb was ingested together with calcium carbonate (0.2 g
calcium carbonate) than when the Pb was ingested without additional calcium (Blake and Mann,
1983; Heard and Chamberlain, 1982). A similar effect of calcium occurs in rats (Barton et al.,
1978b). In other experimental animal models, Pb absorption from the GI tract has been shown to
be enhanced by dietary calcium depletion or administration of vitamin D (Mykkanen and
Wasserman, 1981, 1982).
Effect of Pregnancy. Absorption of Pb may increase during pregnancy. Although there
is no direct evidence for this in humans, an increase in Pb absorption may contribute, along with
other mechanisms (e.g., increased mobilization of bone Pb), to the increase in blood Pb
concentration observed during the later half of pregnancy (Gulson et al., 1997, 1998b, 2004;
Lagerkvist et al., 1996; Rothenberg et al., 1994; Schuhmacher et al., 1996).
Effect of Dose. Lead 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. However, available studies to date do not provide a firm basis for discerning if the GI
absorption of Pb is limited by dose. 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 (Pocock et al., 1983; Sherlock et al., 1982, 1984; Sherlock and Quinn, 1986).
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However, in immature swine that received oral doses of Pb in soil, Pb dose-blood Pb
relationships were curvilinear, whereas dose-tissue Pb relationships for bone, kidney, and liver
were linear (Casteel et al., 2006). The same pattern (nonlinearity for blood Pb concentration and
linearity for tissues) was observed in swine administered Pb acetate intravenously (Casteel et al.,
1997). These results suggest that the nonlinearity in the Pb dose-blood Pb concentration
relationship may derive from an effect of Pb dose on some aspect of the biokinetics of Pb other
than absorption (e.g., saturation of binding of Pb in red blood cells). In fasted rats, absorption
was estimated at 42 and 2% following single oral administration of 1 and 100 mg Pb/kg,
respectively, as Pb acetate, suggesting a limitation on absorption imposed by dose (Aungst et al.,
1981). Saturable mechanisms of absorption have been inferred from measurements of net flux
kinetics of Pb in the in situ-perfused mouse intestine, the in situ-ligated chicken intestine, and the
in vitro-isolated segments of rat intestine (Aungst and Fung, 1981; Barton, 1984; Flanagan et al.,
1979; Mykkanen and Wasserman, 1981). 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.
Effect of Particle Size. Particle size influences the degree of GI absorption (Ruby et al.,
1999). In rats, an inverse relationship was found between absorption and particle size of Pb in
diets containing metallic Pb particles that were <250 jim in diameter (Barltrop and Meek, 1979).
Tissue Pb concentration was a 2.3-fold higher when rats ingested an acute dose (37.5 mg Pb/kg)
of Pb particles <38 jim in diameter than when rats ingested particles having diameters in the
range of 150 to 250 |im (Barltrop and Meek, 1979). Dissolution kinetics experiments with Pb-
bearing mine waste soil suggest that surface area effects control dissolution rates for particles
sizes of <90 jim diameter; however, dissolution of 90- to 250-|im particle size fractions appeared
to be controlled more by surface morphology (Davis et al., 1994). Similarly, Healy et al. (1992)
found that the solubility of Pb sulfide in gastric acid in vitro was much greater for particles
30 jim in diameter than for particles 100 jim in diameter.
Absorption from Soil. Lead in soil can exist in a variety of mineralogical contexts,
which can affect Pb solubility in the GI tract and, potentially, Pb absorption from the GI tract.
In adult subjects who ingested soil (particle size <250 jim) collected from the Bunker Hill NPL
site, 26% (SD 8.1) of the resulting 250-|ig/70-kg body weight Pb dose was absorbed when the
soil was ingested in the fasted state, and 2.5% (SD 1.7) was absorbed when the same soil Pb dose
4-8
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was ingested with a meal (Maddaloni et al., 1998). The dominant Pb minerals in the sample
(relative Pb mass) contained Pb oxides (-40%), Pb sulfates (-25%), and Pb sulfides (-11%).
Absorption reported for fasted subjects (26%) was approximately half that reported for soluble
Pb ingested by fasting adults, i.e., -60% (Blake et al., 1983; Heard and Chamberlain, 1983;
James et al., 1985; Rabinowitz et al., 1980). Measurements of the absorption of ingested soil Pb
in infants or children have not been reported.
Relative bioavailability (RBA) of Pb in soils (i.e., ratio of estimated absorbed fraction of
ingested soil Pb to that of a water-soluble form of Pb, based on measurements of ingested Pb
recovered in blood and/or other tissues) has been more extensively studied in animal models.
The gastric function of juvenile swine is thought to be sufficiently similar to that of human
children to justify use of juvenile swine as a model for assessing RBA of Pb in soils (Casteel
et al., 1996; 2006; Weis and LaVelle, 1991). In immature swine that received oral doses of soil-
like materials from various mine waste sites (75- or 225-jig Pb/kg body weight), relative
bioavailability of soil-borne Pb ranged from 6 to 100% compared to that of a similar dose of
highly water-soluble Pb acetate (Figure 4-1; Casteel et al., 2006). Electron microprobe analyses
of Pb-bearing grains in the various test materials revealed that the grains ranged from as small as
1 to 2 |im up to a maximum of 250 jim (the sieve size used in preparation of the samples) and
that the Pb was present in a wide range of different mineral associations (phases), including
various oxides, sulfides, sulfates, and phosphates. These variations in size and mineral content
of the Pb-bearing grains are the suspected cause of variations in the rate and extent of GI
absorption of Pb from different samples of soil. Based on these very limited data, the relative
oral bioavailability of Pb mineral phases were categorized into "low" (<0.25), "medium" (0.25 to
0.75), and "high" (>0.75) relative bioavailability categories (Figure 4-2; Casteel et al., 2006).
Mineral phases observed in mineralogical wastes can be expected to change over time (i.e.,
weathering), which could change the relative bioavailability of the Pb in soils.
Studies conducted in rats also indicate that the bioavailability of Pb in soils can be lower
than that of water-soluble forms of Pb (e.g., Pb acetate) and that the ingestion of soil can lower
the bioavailability of water-soluble Pb (Freeman et al., 1992; 1994, 1996).
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1.2

1.0

0.8

1 0.6

0.4

0.2

0.0
0
5 10
Test Material
15
20
Test Material
1 California Gulch Fe/Mn PbO
2 Jasper Co Low Lead Yard
3 Jasper Co High Lead Mill
4 Aspen Residential
5 Aspen Berm
6 California Gulch Phase 1 Residential Soil
7 NIST Paint
8 Jasper Co High Lead Smelter
9 Palmerton Location 2
10 Murray Smelter Soil
11 Palmerton Location 4
12 Murray Smelter Slag
13 Bingham Creek Residential
14 Bingham Creek Channel
15 Calonria Gulch AV Slag
16 Midvale Slag
17 Bute Soil
18 California Gulch Oregon Gulch Tailings
19 Galena-enriched Soil
Figure 4-1. Relative bioavailability (RBA) is the bioavailability of the lead in the test
material compared to that of lead acetate (test material lead acetate). The test
material numbers on the horizontal axis refer to the numbered test materials
in the legend.

Source: Casteel et al. (2006).
Dermal Exposure

Inorganic Lead

Dermal absorption of inorganic Pb compounds is generally considered to be much less

than absorption by inhalation or oral routes of exposure; however, few studies have provided

quantitative estimates of dermal absorption of inorganic Pb in humans, and the quantitative

significance of the dermal absorption pathway as a contributor to Pb body burden in humans

remains an uncertainty. Lead was detected in the upper layers of the stratum corneum of

Pb-battery workers prior to their shifts and after cleaning of the skin surface (Sun et al., 2002),

suggesting adherence and/or possible dermal penetration of Pb. Following skin application of

203Pb-labeled Pb acetate in cosmetic preparations (0.12 mg Pb in 0.1 mL or 0.18 mg Pb in

0.1 g of a cream) to eight male volunteers for 12 h, absorption was <0.3%, based on whole-body,
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m
tt
u
Jt
'o
o
Q.
Q.
O
O
n
«
HI
1.2
0.8-

0.6-

0.4-

0.2-

-------
either in the skin or had been absorbed by 24 h; the amount that remained in or on the skin and
the fate of this Pb (e.g., exfoliation) was not determined.
Exfoliation has been implicated as an important pathway of elimination of other metals
from skin (e.g., inorganic mercury; Hursh et al., 1989). The concentrations of Pb in sweat
collected from the right arm increased 4-fold following the application of Pb to the left arm,
indicating that some Pb had been absorbed (amounts of sweat collected or total Pb recovered in
sweat were not reported). In similar experiments with three subjects, measurements of 203Pb in
blood, sweat and urine, made over a 24-h period following dermal exposures to 5 mg Pb as 203Pb
nitrate or acetate, accounted for <1% of the applied (or adsorbed) dose. This study also reported
that absorption of Pb could not be detected from measurements of Pb in sweat following dermal
exposure to Pb as Pb carbonate.
Studies conducted in animals suggest that dermal penetration of inorganic Pb may vary
with Pb species. Dermal absorption of Pb applied as Pb arsenate appeared to be less than of Pb
acetate, based on measurements of kidney Pb levels following application of either compound to
the shaved skin of rats (Laug and Kunze, 1948).

Organic Lead
Relative to inorganic Pb and organic Pb salts, tetraalkyl Pb compounds have been shown
to be rapidly and extensively absorbed through the skin of rabbits and rats (Kehoe and Thamann,
1931; Laug and Kunze, 1948). A 0.75-mL amount of tetraethyl Pb, which was allowed to spread
uniformly over an area of 25 cm2 on the abdominal skin of rabbits, resulted in 10.6 mg of Pb in
the carcass at 0.5 h and 4.4 mg at 6 h (Kehoe and Thamann, 1931). In a comparative study of
dermal absorption of inorganic and organic salts of Pb conducted in rats, -100 mg of Pb was
applied in an occluded patch to the shaved backs of rats. Based on urinary Pb measurements
made prior to and for 12 days following exposure, Pb compounds could be ranked according to
the relative amounts absorbed (i.e., percent of dose recovered in urine): Pb naphthenate (0.17%),
Pb nitrate (0.03%), Pb stearate (0.006%), Pb sulfate (0.006%), Pb oxide (0.005%), and metal Pb
powder (0.002%). This rank order (i.e., Pb naphthalene > Pb oxide) is consistent with a rank
ordering of penetration rates of inorganic and organic Pb salts through excised skin from humans
and guinea pigs: Pb nuolate (Pb linoleic and oleic acid complex) > Pb naphthenate > Pb acetate
> Pb oxide (nondetectable) (Bress and Bidanset, 1991).
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4.2.2 Distribution
Inorganic Lead
Lead in Blood. Blood Pb concentrations vary considerably with age, physiological state
(e.g., pregnancy, lactation, menopause), and numerous factors that affect exposure to Pb. Lead
in blood is found primarily (-99%) in the red blood cells (Bergdahl et al., 1997b, 1998, 1999;
Hernandez-Avila et al., 1998; Manton et al., 2001; Schutz et al., 1996; Smith et al., 2002). Most
of the Pb found in red blood cells is bound to proteins within the cell rather than the erythrocyte
membrane. ALAD is the primary binding ligand for Pb in erythrocytes (Bergdahl et al., 1997b,
1998; Sakai et al., 1982; Xie et al., 1998). Lead binding to ALAD is saturable; the binding
capacity (Kmax) has been estimated to be -850 |ig/dL for red blood cells and the apparent
dissociation constant (Kd) has been estimated to be -1.5 jig/L (Bergdahl et al., 1998). Two other
Pb-binding proteins have been identified in the red cell, a 45 kDa protein (Kmax, 700|ig/dL; Kd
5.5 |ig/L) and a smaller protein(s) having a molecular weight <10 kDa (Bergdahl et al., 1996,
1997b, 1998). Of the three principal Pb-binding proteins identified in red blood cells, 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 blood Pb levels below 40 |ig/dL
(Bergdahl etal., 1996, 1998; Sakai etal., 1982).
Approximately 40 to 75% of Pb in the plasma is bound to plasma proteins, of which
albumin appears to be the dominant ligand (Al-Modhefer et al., 1991; Ong and Lee, 1980).
Lead may also bind to y globulins (Ong and Lee, 1980). Lead 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). Free ionized Pb (i.e.,
Pb2+) in plasma represents an extremely small percentage of total plasma Pb. The concentration
of Pb2+ in fresh serum, as measured by an ion-selective Pb electrode, was reported to be 1/5,000
of the total serum Pb (Al-Modhefer et al., 1991).
Lead in Bone. In human adults, -94% of the total body burden of Pb is found in the
bones. In contrast, bone Pb accounts for 73% of the body burden in children (Barry 1975). Lead
concentrations in bone and bone Pb burden (mass) increase with age throughout the lifetime,
indicative of a relatively slow turnover of Pb in adult bone (Barry 1975, 1981; Gross et al., 1975;
Schroeder and Tipton, 1968). The age-related changes in bone Pb concentration do not exactly
correspond to the change in Pb burden as a result of skeletal growth during childhood and
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adolescence, which results in a dilution of the bone Pb burden in a larger skeletal mass.
The large pool of Pb in adult bone can serve to maintain blood Pb levels long after external
exposure has ended (Fleming et al., 1997; Inskip et al., 1996; Kehoe, 1987; O'Flaherty et al.,
1982; Smith et al., 1996). It can also serve as a source of Pb transfer to the fetus when maternal
bone is resorbed for the production of the fetal skeleton (Franklin et al., 1997; Gulson et al.,
1997, 1999b, 2003). See Section 4.3.2.3 for a more complete discussion of these topics.
Lead is not distributed uniformly in bone. Lead accumulates in those bone regions
undergoing the most active calcification at the time of exposure. During infancy and childhood,
bone calcification is most active in trabecular bone, whereas in adulthood, calcification occurs at
sites of remodeling in cortical and trabecular bone. This suggests that Pb accumulation will
occur predominantly in trabecular bone during childhood, but in both cortical and trabecular
bone in adulthood (Aufderheide and Wittmers, 1992). A portion of Pb in mature bone is
essentially inert, having an elimination half-time of several decades. A labile compartment
exists as well that allows for maintenance of an equilibrium between bone and soft tissue or
blood (Rabinowitz et al., 1976). Although a high bone formation rate in early childhood results
in the rapid uptake of circulating Pb into mineralizing bone, bone Pb is also recycled to other
tissue compartments or excreted in accordance with a high bone resorption rate (O'Flaherty
1995). Thus, most of the Pb acquired early in life is not permanently fixed in the bone
(O'Flaherty, 1995). In general, bone turnover rates decrease as a function of age, resulting in
slowly increasing bone Pb levels among adults (Barry, 1975; Gross et al., 1975; Schroeder and
Tipton, 1968). An X-ray fluorescence study of tibial Pb concentrations in individuals older than
10 years showed a gradual increase in bone Pb after age 20 (Kosnett et al., 1994). In 60- to
70-year-old men, the total bone Pb burden may be >200 mg, while children <16 years old have
been shown to have a total bone Pb burden of ~8 mg (Barry, 1975). However, in some bones
(i.e., mid femur and pelvic bone), the increase in Pb content plateaus at middle age and then
decreases at higher ages (Drasch et al., 1987). This decrease is most pronounced in females and
may be due to osteoporosis and the release of Pb from resorbed bone to blood (Gulson et al.,
2002). Bone Pb burdens in adults are slowly lost by diffusion (heteroionic exchange) as well as
by resorption (O'Flaherty, 1995). Bone Pb stores can contribute substantially to blood Pb, and
maternal bone Pb can be transferred to the fetus during pregnancy and to breast milk and nursing
infants during lactation (see Sections 4.3.2.4 and 4.3.2.5 for further discussion).
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Lead in Soft Tissues. Several studies have compared soft tissue concentrations of Pb in
autopsy samples (Barry, 1975, 1981; Gross et al., 1975; Schroeder and Tipton, 1968). These
studies were conducted in the 1960s and 1970s and, therefore, reflect burdens accrued during
periods when ambient and occupational Pb exposure levels where much higher than current
levels. Methods for quantifying Pb in tissues (including avoidance of contamination of the
tissues) have also advanced considerably since these studies were reported. Average blood Pb
concentrations reported in the adults subjects were -20 |ig/dL in the Barry (1975) and Gross
et al. (1975) studies, whereas more current estimates of the average for adults in the United
States are below 5 |ig/dL (Centers for Disease Control and Prevention, 2005). Levels in other
soft tissues also appear to have decreased substantially since these studies were reported.
For example, average Pb concentrations in kidney cortex of male adults were 0.78 jig/g wet
tissue and 0.79 |ig/g, as reported by Barry (1975) and Gross et al. (1975), respectively (samples
in the Barry study were from subjects who had no known occupational Pb exposures). In a more
recent analysis of kidney biopsy samples collected in Sweden, the mean level of Pb in kidney
cortex among subjects not occupationally exposed to Pb was 0.18 |ig/g (maximum, 0.56 |ig/g)
(Barregard et al., 1999).
In spite of the downward trends in soft tissue Pb levels, the autopsy studies provide a
basis for describing the relative soft tissue distribution of Pb in adults and children. Most of the
Pb in soft tissue is in liver. Relative amounts of Pb in soft tissues as reported by Schroeder and
Tipton (1968), expressed as percent of total soft tissue Pb, were liver, 33%; skeletal muscle,
18%; skin, 16%; dense connective tissue, 11%; fat, 6.4%; kidney, 4%; lung, 4%; aorta, 2%; and
brain, 2% (other tissues were <1%). The highest soft tissue concentrations in adults also occur in
liver and kidney cortex (Barry, 1975; Gerhardsson et al., 1986, 1995b; Gross et al., 1975;
Oldereid et al., 1993). The relative distributions of Pb in soft tissues in males and females,
expressed in terms of tissue:liver concentration ratios, were liver, 1.0 (~1 jig/g wet weight);
kidney cortex, 0.8; kidney medulla, 0.5; pancreas, 0.4; ovary, 0.4; spleen, 0.3; prostate, 0.2;
adrenal gland, 0.2; brain, 0.1; fat, 0.1; testis, 0.08; heart, 0.07; and skeletal muscle, 0.05 (Barry,
1975; Gross et al., 1975). In contrast to Pb in bone, which accumulates Pb with continued
exposure in adulthood, concentrations in soft tissues (e.g., liver and kidney) are relatively
constant in adults (Barry 1975; Treble and Thompson 1997), reflecting a faster turnover of Pb in
soft tissue relative to bone.
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Maternal-Fetal-Infant Transfer. The maternal:fetal blood Pb concentration ratio, based
on cord blood Pb measurements, is -0.9 (Carbone et al., 1998; Goyer, 1990; Graziano et al.,
1990). In one of the larger studies of fetal blood Pb concentration, maternal and cord blood Pb
concentrations were measured at delivery in 888 mother-infant pairs; the cord/maternal ratio was
relatively constant, 0.93, over a blood Pb concentration range of ~3 to 40 |ig/dL (Graziano et al.,
1990). A study of 159 mother-infant pairs also found a relatively constant cord:maternal ratio
(0.84) over a maternal blood Pb range of ~1 to 12 |ig/dL (Carbone et al., 1998). As noted in the
discussion of the distribution of Pb in bone, maternal bone Pb is transferred to the fetus during
pregnancy and can be transferred to breast milk and nursing infants during lactation (see
Sections 4.3.2.4, 4.3.2.5 for further discussion). Breast milk:maternal blood Pb concentration
ratios are, in general, <0.1, although values of 0.9 have been reported (Gulson et al., 1998b).

Organic Lead
Information on the distribution of Pb in humans following exposures to organic Pb is
extremely limited. One hour following 1- to 2-min 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 of these subjects showed an initial declining phase during the first
4 h (tetramethyl Pb) or 10 h (tetraethyl Pb) after the exposure, followed by a phase of gradual
increase in blood Pb concentration that lasted for up to 500 h after the exposure. Radioactive Pb
in blood was highly volatile immediately after the exposure and transitioned to a nonvolatile
state thereafter. These observations may reflect an early distribution of organic Pb from the
respiratory tract, followed by a redistribution of dealkylated Pb compounds.
In a man and woman who accidentally inhaled a solvent containing 31% tetraethyl Pb
(17.6% Pb by weight), Pb concentrations in the tissues, from highest to lowest, were liver,
kidney, brain, pancreas, muscle, and heart (Bolanowska et al., 1967). In another incident, a man
ingested a chemical containing 59% tetraethyl Pb (38% Pb w/w); the Pb concentration was
highest in the liver followed by kidney, pancreas, brain, and heart (Bolanowska et al., 1967).
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4.2.3 Metabolism
Inorganic Lead
Metabolism of inorganic Pb consists of formation of complexes with a variety of protein
and nonprotein ligands. Major extracellular ligands include albumen and nonprotein sulfhydryls.
The major intracellular ligand in red blood cells is ALAD. Lead in other soft tissues such as
kidney, liver, and brain exists predominantly bound to protein. High affinity cytosolic
Pb-binding proteins (PbBPs) generally only observed in male rats have been identified in rat
kidney and brain (DuVal and Fowler, 1989; Fowler, 1989). The PbBPs of rat are cleavage
products of a2|i globulin, a member of the protein superfamily known as retinol-binding proteins
that are generally observed only in male rats (Fowler and DuVal, 1991). Other high-affinity
Pb-binding proteins (Kd -14 nM) have been isolated in human kidney, two of which have been
identified as a 5 kD peptide, thymosin 4 and a 9 kD peptide, acyl-CoA binding protein (Smith
et al., 1998). Lead also binds to metallothionein, but does not appear to be a significant inducer
of the protein in comparison with the inducers cadmium and zinc (Eaton et al., 1980; Waalkes
and Klaassen, 1985). In vivo, only a small fraction of the Pb in the kidney is bound to
metallothionein, and it appears to have a binding affinity that is less than Cd2+, but higher than
Zn2+ (Ulmer and Vallee, 1969); thus, Pb will more readily displace zinc from metallothionein
than cadmium (Goering and Fowler, 1987; Nielson et al., 1985; Waalkes et al., 1984).

Organic Lead
Alkyl Pb compounds undergo oxidative dealkylation catalyzed by cytochrome P450 in
liver and, possibly, in other tissues. Few studies of the metabolism of alkyl Pb compounds in
humans have been reported. Occupational monitoring studies of workers who were exposed to
tetraethyl Pb have shown that tetraethyl Pb is excreted in the urine as diethyl Pb, ethyl Pb, and
inorganic Pb (Turlakiewicz and Chmielnicka, 1985; Vural and Duydu, 1995; Zhang et al., 1994).
Trialkyl Pb metabolites were found in the liver, kidney, and brain following exposure to the
tetraalkyl compounds in workers; these metabolites have also been detected in brain tissue of
nonoccupational subjects (Bolanowska et al., 1967; Nielsen et al., 1978). In volunteers exposed
by inhalation to 0.64 and 0.78 mg Pb/m3 of 203Pb-labeled tetraethyl and tetramethyl Pb,
respectively, Pb was cleared from the blood within 10 h, followed by a reappearance of
radioactivity back into the blood after -20 h (Heard et al., 1979). The high level of radioactivity
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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).

4.2.4 Excretion
Inorganic Lead
The kinetics of elimination of Pb from the body reflects the existence of fast and slow
pools of Pb in the body. Blood, which comprises -1% of body burden, exchanges with both
slow and fast pools and exhibits multiphasic elimination kinetics. The dominant phase, exhibited
shortly after a change in exposure occurs, has a half-life of-20 to 30 days. A slower phase
becomes evident with longer observation periods following a decrease in exposure. The half-life
of this slow phase has been estimated to be ~3 to 30 years and appears to correlate with finger
bone Pb levels and is thought to reflect the release of Pb from bone stores to blood.
Independent of the route of exposure, absorbed Pb is excreted primarily in urine and
feces; sweat, saliva, hair, nails, and breast milk are minor routes of excretion (Chamberlain et al.,
1978; Griffin et al., 1975; Hursh and Suomela, 1968; Hursh et al., 1969; Kehoe, 1987; Moore
et al., 1980; Rabinowitz et al., 1976; Stauber et al., 1994). Fecal excretion accounts for
-one-third of total excretion of absorbed Pb (fecal :urinary excretion ratio of-0.5), based on
intravenous injection studies conducted in humans (Chamberlain et al., 1978). A similar value
for fecal:urinary excretion ratio, -0.5, has been observed following inhalation of submicron Pb
particles (Chamberlain et al., 1978; Hursh et al., 1969). Estimates of blood-to-urine clearance
range from 0.03 to 0.3 L/day with a mean of 0.12 L/day (Araki et al., 1990; Berger et al., 1990;
Chamberlain et al., 1978; Gulson et al., 2000; Koster et al., 1989; Manton and Malloy, 1983;
Rabinowitz et al., 1973, 1976; Ryu et al., 1983; see Diamond, 1992 for an analysis of these data).
Much of the available information on the excretion of ingested Pb in adults derives from
studies conducted on five male adults who received daily doses of 207Pb nitrate for periods up to
210 days (Rabinowitz et al., 1976). The dietary intakes of the subjects were reduced to
accommodate the tracer doses of 207Pb without increasing daily intake, thus preserving a steady
state with respect to total Pb intake and excretion. Total Pb intakes (diet plus tracer) ranged from
-210 to 360 |ig/day. Urinary excretion accounted for -12% of the daily intake (range for five
4-18
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subjects, 7 to 17%) and fecal excretion, -90% of the daily intake (range, 87 to 94%). Based on
measurements of tracer and total Pb in saliva, gastric secretions, bile, and pancreatic secretions
(samples collected from three subjects by intubation), GI secretion of Pb was estimated to be
-2.4% of intake (range, 1.9 to 3.3%). In studies conducted at higher ingestion intakes, 1 to
3 mg/day for up to 208 weeks, urinary Pb excretion accounted for -5% of the ingested dose
(Kehoe, 1987).

Organic Lead
Lead absorbed after inhalation of tetraethyl and tetramethyl Pb is excreted in exhaled air,
urine, and feces (Heard et al., 1979). Following 1- to 2-min inhalation exposures to 203Pb
tetraethyl (1 mg/m3), in four male subjects, 37% of inhaled 203Pb was initially deposited in the
respiratory tract, of which -20% was exhaled in the subsequent 48 h (Heard et al., 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 h. Lead that was not
exhaled was excreted in urine and feces. Fecal:urinary excretion ratios were 1.8 following
exposure to tetraethyl Pb and 1.0 following exposure to tetramethyl Pb (Heard et al., 1979).
Occupational monitoring studies of workers who were exposed to tetraethyl Pb have shown that
tetraethyl Pb is excreted in the urine as diethyl Pb, ethyl Pb, and inorganic Pb (Turlakiewicz and
Chmielnicka, 1985; Vural and Duydu, 1995; Zhang et al., 1994).
4.3 BIOLOGICAL MARKERS OF LEAD BODY BURDENS
AND EXPOSURE
4.3.1 Lead in Blood
4.3.1.1 Summary of Key Findings from the 1986 Lead AQCD
The extensive use of blood Pb concentration as a dose metric mainly reflects the greater
feasibility of incorporating blood Pb measurements into clinical or epidemiologic studies,
compared to other potential dose indicators such as Pb in kidney, plasma, urine, or bone (Flegal
and Smith, 1995; Graziano, 1994; Skerfving, 1988). However, blood Pb measurements have
several limitations as measures of Pb body burden (Mushak, 1989, 1993), as were noted in
Section 13.3.2 of the 1986 Lead AQCD, which discussed attributes and limitations of blood Pb
4-19
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concentration as an indicator of internal exposure. Since the 1986 Lead AQCD was completed,
relevant developments include numerous studies of determinants of Pb levels in bone (see
Section 4.3.2), which further support the importance of bone Pb on blood Pb as an index of Pb
exposure. The enhanced understanding of Pb biokinetics has also been consolidated into
exposure-biokinetics models, which not only serve to illustrate exposure-blood-body burden
relationships, but also provide a means for making predictions about these relationships that can
be tested experimentally or in epidemiologic studies. The basic concepts laid out in the 1986
Lead AQCD, that the concentration of Pb in blood is largely determined by the relatively recent
exposure history of the individual and that it reflects the level of Pb in a relatively mobile and
small compartment, remain valid. In children, who experience a more rapid turnover of bone
mineral than adults, blood Pb concentrations closely parallel changes in total body burden.

4.3.1.2 Analytical Methods for Measuring Lead in Blood
Analytical methods for measuring Pb in blood include flame atomic absorption
spectrometry (AAS), graphite furnace atomic absorption spectrometry (GFAAS), anode stripping
voltammetry (ASV), inductively coupled plasma atomic emission spectroscopy (ICP-AES), and
inductively coupled plasma mass spectrometry (ICP-MS). GFAAS and ASV are generally
considered to be the methods of choice (Flegal and Smith, 1995). Background correction, such
as Zeeman background correction that minimizes the impact of the absorbance of molecular
species, must be applied. Limits of detection for Pb using AAS are on the order of 5 to 10 |ig/dL
for flame AAS measurements, -0.1 |ig/dL for flameless AAS measurements, and -0.005 |ig/dL
for GFAAS (Flegal and Smith, 1995; National Institute for Occupational Safety and Health,
1994). Standard methods that have been reported for blood Pb analysis are summarized in
Annex Table AX4-1. Sample preparation usually consists of wet ashing in heated strong acid
(National Institute for Occupational Safety and Health, 1977a,b,c,d,e); however, preparation
methods not requiring wet ashing have also been reported (Aguilera de Benzo et al., 1989;
Delves and Campbell, 1988; Manton and Cook, 1984; National Institute for Occupational Safety
and Health, 1977f; Que Hee et al., 1985; Zhang et al., 1997). The presence of phosphate,
ethylenediaminetetraacetic acid (EOTA), or oxalate can sequester Pb and cause low readings in
flame AAS (National Institute for Occupational Safety and Health, 1984). A comparison of
4-20
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IDMS, ASV, and GFAAS showed that all three of these methods can be used to quantify Pb
levels in blood (Que Hee et al., 1985).

4.3.1.3 Levels of Lead in Blood
Blood Pb concentrations in the U.S. general population have been monitored in the
National Health and Nutrition Examination Survey (NHANES) conducted by the Centers for
Disease Control and Prevention. Samples for Phase 2 of NHANES III were collected during
1991 to 1994. Geometric mean blood Pb concentration of U.S. adults, ages 20 to 49 years,
estimated from the NHANES III Phase 2, were 2.1 jig/dL (95% CI: 2.0, 2.2) (Pirkle et al.,
1998). Among adults, blood Pb concentrations were highest in the strata that included ages
70 years and older (3.4 |ig/dL; 95% CI: 3.3, 3.6). Geometric mean blood Pb concentration of
children, ages 1 to 5 years, were 2.7 (95% CI: 2.5, 3.0) for the 1991 to 1994 survey period;
however, the mean varied with socioeconomic status (SES) and other demographic
characteristics that have been linked to Pb exposure (e.g., age of housing) (Pirkle et al., 1998).
Central estimates from the NHANES III Phase 2 (1991 to 1994), when compared to those from
Phase 1 of the NHANES III (1988 to 1991) and the NHANES II (1976 to 1980), indicate a
downward temporal trend in blood Pb concentrations in the United States over this period. Data
from the most recent survey (NHANES IV, Centers for Disease Control and Prevention, 2005)
are shown in Tables 4-1 and 4-2. For survey years 2001-2002, the geometric mean blood Pb
concentration for ages >1 year (n = 8,945) was 1.45 |ig/dL (95% CI: 1.39, 1.52), with the
geometric mean in males (n = 4,339) being 1.78 |ig/dL (95% CI: 1.71, 1.86) and in females
(n = 4,606) being 1.19 |ig/dL (95% CI: 1.14, 1.25). Blood Pb concentrations in the U.S. general
population have decreased over the past three decades as regulations regarding Pb paint, leaded
fuels, and Pb-containing plumbing materials have decreased exposure. Changes in children over
time are shown in Figure 4-3.
Yassin et al. (2004) analyzed occupational category strata from NHANES III (1988 to
1994; Table 4-3). The geometric mean for all adults (n = 11,126) included in the analysis was
2.42 |ig/dL (GSD 6.93), with the highest means estimated for vehicle mechanics (n = 169;
GM 4.80 |ig/dL [GSD 3.88]) and construction workers (n = 122; GM 4.44 |ig/dL [GSD 7.84]).
See Annex Table AX4-2 for a summary of selected measurements of blood Pb levels in humans.
4-21
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Table 4-1. Blood Lead Concentrations in United States by Age, NHANES IV (1999-2002)
Age
Survey Period
n
Blood Lead
(Mg/dL)a
1-5 years
1999-2000 2001-2002
723 898
2.23 1.70
(1.96,2.53) (1.55,1.87)
6-11 years
1999-2000 2001-2002
909 1,044
1.51 1.25
(1.36, 1.66) (1.14, 1.36)
12-19 years
1999-2000 2001-2002
2,135 2,231
1.10 0.94
(1.04,1.17) (0.90,0.99)
^20 years
1999-2000 2001-2002
4,207 4,772
1.75 1.56
(1.68, 1.81) (1.49, 1.62)
aBlood Pb concentrations presented are geometric means (95% CI).
to
to
Table 4-2. Blood Lead Concentrations in United States by Gender, NHANES IV (1999-2002)
Gender
Survey Period
n
Blood Lead
(Mg/dL)a
Males
7999-2000
3,913
2.01
(1.93,2.09)
2001-2002
4,339
1.78
(1.71, 1.86)
Females
7999-2000
4,057
1.37
(1.32, 1.43)
2007-2002
4,606
1.19
(1.14, 1.25)
Blood Pb concentrations presented are geometric means (95% CI).
-------
Table 4-3. Blood Lead Concentrations by Occupation, NHANES III (1988-1994)
to
Blood Lead (ug/dL)
Occupation
Vehicle mechanics
Food service workers
Management, professional, technical, and sales workers
Personal service workers
Agricultural workers
Production workers: machine operators, material movers, etc.
Laborers other than in construction
Transportation workers
Mechanics other than vehicle mechanics
Construction trades people
Construction laborers
Health service workers
All workers
n
169
700
4,768
1,130
498
1,876
137
530
227
470
122
499
11,126
GM
4.80
2.00
2.13
2.48
2.76
2.88
3.47
3.49
3.50
3.66
4.44
1.76
2.42
GSD
3.88
2.69
4.05
4.52
4.02
4.24
3.36
5.19
4.91
4.64
7.84
2.24
6.93
Maximum
28.1
27.0
39.4
25.9
23.4
52.9
21.8
22.3
16.6
16.9
36.0
22.4
52.9
Data from Yassin et al. (2004).
-------
18
16-
? 14
J; 12J
re 10
o
&
DQ
8
6-I
4
24
0
1976-1980 1988-1991 1991-1994 1999-2000
Survey Period
2001-2002
Figure 4-3. Blood lead concentrations in U.S. children, 1-5 years of age. Shown are
geometric means and 95% confidence intervals as reported from the
NHANES II (1976-1980) and NHANES III Phase 1 (1988-1991; Pirkle et al.,
1994); NHANES III Phase 2 (1991-1994; Pirkle et al., 1998); and NHANES IV
(1999-2000, 2001-2002; Centers for Disease Control and Prevention, 2005).
4.3.1.4 Blood Lead as a Biomarker of Lead Body Burden
Considerable recent effort has been directed at evaluating possible associations between
Pb body burden and health outcomes, including neurodevelopmental outcomes in children
(Wasserman et al., 1994) and renal/cardiovascular outcomes in adults (Cheng et al., 2001; Gerr
et al., 2002; Glenn et al., 2003; Hu et al., 1996; Korrick et al., 1999; Rothenberg et al., 2002;
Tsaih et al., 2004). Conceptually, measurement of long-term Pb body burden may be a preferred
dose metric if the effects of Pb on a particular outcome are lasting and cumulative. However, if
the effects of Pb on the outcome represent the acute effects of current exposure, then long-term
body burden may not be the preferred exposure metric. In the absence of clear evidence as to
which averaging time (current versus long-term) is most relevant to a particular outcome, both
long-term and short-term dose metrics need be explored.
A simple conceptual representation of the Pb body burden is that it contains a fast
turnover pool, comprising mainly soft tissue, and a slow pool, comprising mainly skeletal tissues
(Rabinowitz et al., 1976). The rapid pool has an elimination half-life of-28 days and comprises
4-24
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<1% of the Pb body burden. The slow pool has an elimination half-life of several decades and
comprises >90% of the total Pb body burden. Blood, which comprises -1% of body burden,
exchanges with both the slow and fast pools and exhibits multiphasic elimination kinetics.
The dominant phase, exhibited shortly after a change in exposure occurs, has a half-life of -20 to
30 days. A slower phase becomes evident with longer observation periods following a decrease
in exposure. The half-life of this slow phase has been estimated to be ~3 to 30 years and appears
to correlate with finger bone Pb levels and is thought to reflect the release of Pb from bone stores
to blood. Children who have been removed from a relatively brief exposure to elevated
environmental Pb exhibit faster slow-phase kinetics than children removed from exposures that
lasted several years, with half-times of ~9 and -30 months, respectively (Manton et al., 2000).
This characterization is supported by measurements of Pb content of cadaver tissues (Barry,
1975; Schroeder and Tipton, 1968), Pb isotope and stable Pb kinetics in adults (Chamberlain
et al., 1978; Rabinowitz et al., 1976; Griffin et al., 1975), and measurements of blood and bone
Pb levels in retired Pb workers (Schiitz et al., 1987a; Chri staffers son et al., 1986).
As a consequence of a relatively large fraction of the body burden having a relatively slow
turnover compared to blood, a constant Pb uptake (or constant intake and fractional absorption)
gives rise to a quasi-steady state blood Pb concentration, while the body burden continues to
increase, largely as a consequence of retention of Pb in bone (Figure 4-4). As a result, the
contribution of blood Pb to body burden decreases over time. 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 to
100 days (i.e., 3 to 4 times the blood elimination half-life). In the hypothetical simulation shown
in Figure 4-4, body burden has approximately doubled (from 5 to 10 mg) as a result of a 5-year
period of increased Pb uptake; however, the blood Pb concentration prior to and 1 year following
cessation of the increased uptake has not changed (-2 |ig/dL). Therefore, a single blood Pb
concentration measurement, or a series of measurements taken over a short-time span, can be
expected to be a relatively poor index of Pb body burden unless exposure over the lifetime and,
thereby, body burden has been constant. On the other hand, an average of individual blood Pb
concentrations measured over a longer period of time (long-term average blood Pb
concentrations) can be expected to be a better index of body burden. In the hypothetical
simulation shown in Figure 4-4, both the long-term average blood Pb concentration and the Pb
body burden have approximately doubled.
4-25
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30 40
Age (year)
50
0)

II
< o>
•o .5
BT3
JC «
»3
f ^
^ o
O O
E 5
30
40
Age (year)
50
Figure 4-4. Simulation of relationship between blood lead concentration and body burden
in adults. A constant baseline intake gives rise to a quasi-steady state blood
lead concentration, while the body burden continues to increase, largely as a
consequence of retention of lead in bone (upper panel). An abrupt change
in lead uptake gives rise to a relatively rapid change in blood lead, to a new
quasi-steady state, and a relatively small change in body burden. The long-
term average blood lead concentration more closely tracks the pattern of
change in body burden (lower panel). Simulation based on lead biokinetics
model of Leggett (1993).
4-26
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The disparity in the kinetics of blood Pb and body burden has important implications for
the interpretation of blood Pb concentration measurements in epidemiology studies. By design,
cross-sectional studies sample blood Pb concentration at one time or over relatively narrow
windows of time. In these samples, the blood Pb concentration may or may not reflect well the
body burden; it is more likely to do so if the measured value is a reflection of the long-term
average blood Pb concentration. However, in cross-sectional samples, this cannot be
ascertained. Longitudinal sampling provides a means for estimating average blood Pb
concentrations over time, and such estimates are more likely to be more strongly influenced by
differences in body burden than by differences in short-term variability in exposure. The degree
to which repeated sampling will reflect the actual long-term time-weighted average blood Pb
concentration will depend on the sampling frequency in relation to variability in exposure. High
frequency variability in exposures can produce episodic (or periodic) oscillations in blood Pb
concentration and body burden that may not be captured with low sampling frequencies.
The same basic concepts described above regarding Pb biokinetics of adults also apply to
children. The empirical basis for the understanding of the biokinetics of Pb in children is much
weaker than that for adults. However, based on the understanding of bone mineral kinetics and
its importance as a mechanism for uptake and loss of Pb from bone (Leggett, 1993; O'Flaherty,
1991a,b,c, 1993, 1995), the slow pool, described above for adults, is thought to be much more
labile in children, reflecting a more rapid turnover of bone mineral in children. As a result, while
bone growth will contribute to accumulation of Pb in bone in children, changes in blood Pb
concentration in children are thought to more closely parallel changes in total body burden
(Figure 4-5). Empirical evidence in support of this comes from longitudinal studies in which
relatively high correlations (r = 0.85) were found between concurrent (r = 0.75) or lifetime
average blood Pb concentrations (r = 0.85) and tibia bone Pb concentrations (measured by XRF)
in a sample of children in which average blood Pb concentrations exceeded 20 |ig/dL;
the correlations was much weaker (r < 0.15) among children who had average blood Pb
concentration < 10 |ig/dL (Wasserman et al., 1994). Nevertheless, in children, as in adults, the
long-term time-weighted average blood Pb concentration is more likely to provide a better
reflection of Pb body burden than a single sample (the exception to this would be if exposure
and, thereby, body burden was relatively constant throughout the lifetime).
4-27
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m
4 6
Age (year)
10
00
O
a.
03
c
a
3
(O
23456
Age (year)
7 8 9 10
Figure 4-5. Simulation of relationship between blood lead concentration and body burden
in children. Blood lead concentration is thought to parallel body burden more
closely in children than in adults, due to more rapid turnover of bone and
bone-lead stores in children (upper panel). Nevertheless, the long-term
average blood lead concentration more closely tracks the pattern of change
in body burden (lower panel). Simulation based on Leggett (1993) lead
biokinetics model.
4-28
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Lead that distributes from blood to tissues does so through the plasma. Therefore,
theoretically, Pb concentrations in plasma should be a reflection of Pb in tissues and body
burden. However, the concentration of Pb in plasma is extremely difficult to measure accurately,
because levels in plasma are near the lower limits of most analytical techniques (e.g., -0.4 |ig/L
at blood Pb concentration of 100 |ig/L (Bergdahl and Skerfmg, 1997; Bergdahl et al., 1997b),
because hemolysis that occurs with typical analytical practices can contribute substantial
measurement error (Bergdahl et al. 1998; Cavalleri et al. 1978; Smith et al. 1998), and because
anticoagulant agents used in stabilization of blood samples for preparation of plasma can affect
the distribution of Pb between blood cells and plasma (Barton, 1989; Al-Modhefer et al., 1991;
Simons, 1993). Recent advances in ICP-MS offer sensitivity sufficient for measurements of Pb
in plasma (Schiitz et al. 1996). While the technique has been applied to assessing Pb exposures
in adults (Cake et al. 1996; Hernandez-Avila et al. 1998; Manton et al. 2001; Smith et al. 2002;
Tellez-Rojo et al. 2004), it has not received widespread use in epidemiologic studies.

4.3.1.5 Blood Lead as a Biomarker of Lead Exposure
Characterizing quantitative relationships between external Pb exposures and blood Pb
concentrations has become central to concentration-response analyses for human populations
exposed to Pb. The 1986 Lead AQCD summarized the empirical basis for this as it stood at the
time. A summary of empirically-derived regression slope factors relating Pb exposures and
blood Pb is provided in Abadin and Wheeler (1997). More recent meta-analyses, based on
structure equation modeling, provide further support for quantitative relationships between Pb
exposures and blood Pb concentrations in children (e.g., U.S. Environmental Protection Agency,
2001; Lanphear et al., 1998; Succop et al., 1998).
The elimination half-time of Pb from blood has been estimated to be -25 to 30 days in
adult males whose blood Pb concentrations were >20 |ig/dL (Chamberlain et al., 1978;
Rabinowitz et al., 1976; Griffin et al., 1975). In the latter studies, the elimination half-times
were estimated from measurements of the time to achieve a new quasi-steady state blood Pb
concentration following an increase in exposure (Griffin et al., 1975) or from measurement of the
rate of change in blood concentration of an administered isotope of Pb (Chamberlain et al., 1978;
Rabinowitz et al., 1976). However, the half-time for a change in blood Pb concentration (or
stable isotope ratio) after an abrupt change in exposure can be much longer. Gulson et al. (1995,
4-29
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1999a) estimated the half-time for the change in stable Pb isotope ratio (206Pb:204Pb) in blood,
after an abrupt change instable isotope exposure, to be -25 to 80 days in adult females (blood Pb
concentration range 3 to 20 |ig/dL). Manton et al (2000) estimated the half-time for the decline
in blood Pb concentration after an abrupt decrease in exposure to be -200 to 1000 days in
children (age range: 8 to 60 months; blood Pb concentration: 7 to 5 jig/dL). The longer half-
times measured under the latter conditions reflect the contribution of bone Pb stores to blood Pb
following a change in exposure. Studies of the bone (or chelatable Pb) and blood Pb kinetics in
retired Pb workers also demonstrate a slow phase to the elimination kinetics of blood Pb that
reflects the continuing redistribution of Pb from bone to blood (Alessio, 1988; Nilsson et al.,
1991; Schiitz et al, 1987a; Christoffersson et al., 1986).
Based on these observations, a single blood Pb concentration may reflect the near-term or
longer-term history of the individual to varying degrees, depending on the relative contributions
of internal (e.g., bone) and external sources of Pb to blood Pb, which in turn will depend on the
exposure history and possibly age-related characteristics of bone turnover.
Analyses of serial blood Pb concentrations measured in longitudinal epidemiologic
studies have found relatively strong correlations (e.g., r = 0.5 to 0.8) between individual blood
Pb concentrations measured after 6 to 12 months of age (Dietrich et al., 1993; McMichael et al.,
1988; Otto et al., 1985; Rabinowitz et al., 1984; Schnaas et al, 2000). These observations
suggest that, in general, exposure characteristics of an individual child (e.g., exposure levels
and/or exposure behaviors) tend to be relatively constant across age. However, a single blood Pb
measurement may not distinguish between a history of long-term lower-level Pb exposure from a
history that includes higher acute exposures (Mushak, 1998). This is illustrated in Figure 4-6.
Two hypothetical children are simulated. Child A has a relatively constant Pb intake from birth,
whereas Child B has the same long-term Pb intake as Child A but with a 1-year elevated intake
beginning at age 24 months (Figure 4-6, upper panel). The absorption fraction is assumed to be
the same for both children. Blood Pb samples 1 and 5, or 2 and 4, will yield similar blood Pb
concentrations (-3 or 10 |ig/dL, respectively), yet the exposure contexts for these samples are
very different. Two samples (e.g., 1 and 2, or 4 and 5), at a minimum, are needed to ascertain if
the blood Pb concentration is changing over time. The rate of change can provide information
about the magnitude of change in exposure, but not necessarily about the time history of the
change (Figure 4-6, lower panel). Here again, time-integrated measurements of Pb concentration
4-30
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0 12 24 36 48 60 72
Age (months)
40
30-
"8 20 -I
0)
T3
O
.2 10
CO
0
0 12 24 36 48 60
Age (months)
72 84
Figure 4-6. Simulation of temporal relationships between lead exposure and blood lead
concentration in children. Child A and Child B have a relatively constant
basal lead intake (ug/day/kg body weight) from birth; Child B experiences
1-year elevated intake beginning at age 24 months (upper panel). Blood lead
samples 1 and 5, or 2 and 4, will yield similar blood lead concentrations (~3 or
10 ug/dL, respectively), yet the exposure scenarios for these samples are very
different. As shown in the example of Child C and Child D, two samples can
provide information about the magnitude of change in exposure, but not
necessarily the temporal history of the change (lower panel).
4-31
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may provide a means for accounting for some of these factors and, thereby, provide a better
measure of long-term Pb exposure. The same concepts apply to estimation of long-term
exposure based on blood Pb measurements in adults (Gerhardsson et al., 1992, 1995a; Roels
etal., 1995).
An additional complication is that the relationship between Pb intake and blood Pb
concentration is curvilinear, i.e., the increment in blood Pb concentration per unit of Pb intake
decreases with increasing blood Pb concentration, both in children (Lacey et al., 1985; Ryu et al.,
1983; Sherlock and Quinn, 1986) and in adults (Kehoe, 1987; Laxen et al., 1987; Pocock et al.,
1983; Sherlock et al., 1982, 1984). The nonlinearity is evident even at blood Pb concentrations
below 25 |ig/dL (Figure 4-7). The nonlinearity in the Pb intake-blood Pb concentration
relationship is derived, at least in part, from a capacity limitation in the accumulation of Pb in
erythrocytes (Bergdahl et al., 1997a, 1998, 1999; Manton et al., 2001; Smith et al., 2002).
A capacity-limited process may also reside at the level of intestinal absorption; however, the
dose at which absorption becomes appreciably limited in humans is not known. Lead intake-
blood Pb relationships also vary (a) with age, as a result of age-dependency of GI absorption of
Pb, and (b) with diet and nutritional status (Mushak, 1991).
The blood Pb concentration is also influenced by Pb in bone. Evidence for the exchange
of bone Pb and soft tissue Pb stores comes from analyses of stable Pb isotope signatures of Pb in
bone and blood. As noted earlier, bone Pb likely contributes to the slow phase of elimination of
Pb from blood that has been observed in retired Pb workers (Christoffersson et al., 1986; Schiitz
et al., 1987a). Bone Pb stores may contribute 40 to 70% of the Pb in blood (Manton, 1985;
Gulson et al., 1995; Smith et al., 1996). This contribution increases during pregnancy, when
mobilization of bone Pb increases, apparently as the bone is resorbed to produce the fetal
skeleton (Gulson et al., 2003). The mobilization of bone Pb during pregnancy may contribute,
along with other mechanisms (e.g., increased absorption), to the increase in Pb concentration that
has been observed during the later stages of pregnancy (Gulson et al., 1997; Lagerkvist et al.,
1996; Schuhmacher et al., 1996). In addition to pregnancy, other states of increased bone
resorption appear to result in release of bone Pb to blood; these include lactation, osteoporosis,
and menopause (Gulson et al., 2003). These observations are consistent with epidemiologic
studies that have shown increases in blood Pb concentration after menopause and in association
with decreasing bone density in postmenopausal women (Hernandez-Avila et al., 2000;
4-32
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Leggett Model
T3
CO
O
T3
O
m
80
70
60
50
40
30
20
10
0
0 5 10 15
Lead Intake (pg/kg/day)
20
25

5
1 2 3
Lead Intake (pg/kg/day)
a)
(0
0)
O
£
03
80
70
60
50
40
30
20
10
0
O'Flaherty Model
10 20 30
Lead Intake (ng/kg/day)
40
25

0
1234
Lead Intake (|jg/kg/day)
Figure 4-7. Simulation of relationships between lead intake and blood lead concentration
in adults and children. The relationship between lead intake and blood lead
concentration is curvilinear in adults and children. Predictions are for a
2-year-old child and 30-year-old adult, for a constant lead intake (ug/kg/day).
Predictions are based on Leggett (1993, upper panel) and O'Flaherty (1993,
1995, lower panel). Right panels provide an enlarged view of lower intakes for
each model.
Nash et al., 2004; Symanski and Hertz-Picciotto, 1995). The relationship between blood and
bone Pb is discussed further in Section 4.3.2 on bone Pb as a biomarker of Pb exposure.
4-33
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4.3.1.6 Summary of Blood Lead as a Biomarker of Lead Body Burden and Exposure
The blood Pb concentration measured in an individual will be determined by the recent
exposure history of the individual as well as by the long-term exposure history that gives rise to
accumulated bone Pb stores. The contribution of the latter to blood Pb may change with the
duration and intensity of the exposure, age, and various physiological variables (e.g., nutritional
status, pregnancy, and menopause). Longitudinal measurements of blood Pb can be expected to
provide a more reliable measure of exposure history of an individual (and will more closely
parallel body burden) compared to a single measurement; however, the degree to which this will
apply will depend on the sampling frequency with respect to the temporal pattern of exposure.
In general, higher blood Pb concentrations can be interpreted as indicating higher
exposures (or Pb uptakes); however, they do not necessarily predict appreciably higher body
burdens. Similar blood Pb concentrations in two individuals (or populations) do not necessarily
translate to similar body burdens or similar exposure histories.

4.3.2 Lead in Bone
4.3.2.1 Summary of Key Findings from the 1986 Lead AQCD
In the 1986 Lead AQCD, the discussion on the distribution of Pb in bone was fairly
limited and mostly based on postmortem studies. The distribution between the two major
compartments of cortical and trabecular bone was specifically addressed based on the pioneering
isotopic work of Rabinowitz et al. (1977). Estimates of the amount of Pb in bone were also
provided. There was limited discussion of the half-life of Pb in bone as being on the order of
several decades.
One of the major conclusions of the 1986 Lead AQCD regarding bone Pb was that the
traditional view that the skeletal system was a total sink for body Pb was now giving way to the
notion that there were at least several bone compartments for Pb, with different mobility profiles.
The possibility of bone Pb serving as a source of long-term internal exposure was also
considered.
Since 1986, the main focus of Pb in bone studies has been on occupationally exposed
subjects, because of concern until more recent times about the ability to measure lower levels of
Pb in bone from environmentally exposed subjects. Furthermore, most of the focus has been on
4-34
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adult males, with very few studies on females and children. The newly available studies of Pb in
bone are discussed in the following sections.

4.3.2.2 Methodology of Bone Lead Analysis
Analytical Methods for Measuring Lead in Bone
Bone is comprised of two main types (cortical and trabecular) that have distinct rates of
turnover and Pb release, resulting in potential differences in implications with respect to toxicity
aspects (further discussed in Section 4.3.2.3). The most commonly measured bones are the tibia,
calcaneus, patella, and finger bone. For cortical bone, the midpoint of the tibia is measured.
For trabecular bone, both the patella and calcaneus are measured. Recent studies favor
measurement of the patella, because it has more bone mass and may afford better measurement
precision than the calcaneus. The advantages and disadvantages of patella and calcaneus sites
have not been thoroughly investigated. Bone Pb measurements in cadavers, environmentally
exposed subjects, and occupationally exposed subjects are presented in Annex Tables AX4-3,
AX4-4, and AX4-5, respectively.
Bone analysis methods for in vivo measurements have included AAS, ASV, ICP-AES,
ICP-MS, laser ablation inductively coupled plasma mass spectrometry (LA-ICP-MS), thermal
ionization mass spectrometry (TIMS), synchrotron radiation induced X-ray emission (SRIXE),
particle induced X-ray emission (PIXE), and X-ray fluorescence (XRF). Since the 1986 Lead
AQCD, there have been many new papers published on bone Pb using XRF. The upsurge in
popularity of the XRF method has paralleled a decline in the use of the other methods.
In the past, two main approaches for XRF measurements have been used to measure Pb
concentrations in bone, the K-shell and L-shell methods. The K-shell method is now the most
widely used, as there have been no further developments in L-shell devices since the early 1990s.
The K-shell methods using 57Cd and 109Cd have been described in detail by Somervaille et al.
(1989). Briefly, the K-shell XRF method uses 88.034 keV gamma rays from 109Cd to fluoresce
the K-shell X-rays of Pb.
Plaster-of-Paris "phantoms" with varying Pb concentrations measured by ICP-MS or AAS
are used to calibrate the K-shell systems. Differences and corrections in the use of the phantoms
have been discussed in, for example, Gordon et al. (1994), Kondrashov and Rothenberg (2001a),
Todd et al. (2000), Todd and Chettle (2003), and Chettle et al. (2003). Todd et al. (2002b) also
4-35
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commented that the calibration of 109Cd-based K-shell XRF equipment with standards that
consist of a Pb-doped Plaster-of-Paris matrix are not valid because Plaster of Paris is not true
bone matrix. The problem of contamination from various sources (external or the matrix of the
Plaster-of-Paris phantom) and its impact on variance was addressed by Todd (2000).
There have been several publications focused on the calculating estimates of Pb
concentrations and uncertainties for XRF measurements (Kondrashov and Rothenberg, 2001a,b;
Gordon et al. 1993; Todd, 2000; Todd and Chettle, 2003; Chettle et al., 2003; Todd et al., 2003)
culminating in an agreed position in 2003 for these uncertainties (Chettle et al., 2003). Todd
et al. (2000) provide a detailed discussion of the influences on the variability and measurement
uncertainty including repositioning, sample measurement duration, overlying tissue, operator
expertise, detector resolution, and changes to measurement process over time. Some of these
aspects were also discussed by Hu et al. (1995). In a K-shell XRF study of 210 children aged
11 to l2l/2 years from a smelter town in Yugoslavia, Todd et al. (2001) (ER) decided that the
methodological uncertainty in children was comparable to that in adults.
Apart from the recent study of the L-shell method by Todd et al. (2002a), there have been
several investigations of the reproducibility and accuracy of the K-shell method. The main
approaches have been repeated measurements on the same individuals within a limited time
frame, repeated measurements on the same individuals over an extended time frame, repeated
measurements of cadaver legs at the same location (and compared with AAS analyses), and
extended measuring times (Somervaille et al., 1986; Aro et al., 2000; Gordon et al., 1994;
Hoppin et al., 2000; Todd et al., 2000). In one of the earliest validation studies, Somervaille
et al. (1986) compared K-shell and AAS measurements of 30 dissected tibia whose Pb values
ranged from 6.5 to 83 jig/g of ashed bone. They found no evidence of a systematic difference
between the two measurement techniques of more than 1 jig/g.
Short-term variability of XRF results was investigated in two recent studies of cadaver
legs. In the first study, Aro et al (2000) compared Pb levels in 8 cadaver legs: XRF measures of
intact bone with skin and hair, bare bones, and then ashed bone by ICP-MS. The XRF
measurements involved 10 consecutive 30-min measurements on each bone. In the tibia, Pb
concentrations by XRF showed standard deviations for each bone ranging from 6 to 58% for
intact bone (mean 27%) and 13 to 36% for bare bone. Patella Pb concentrations had standard
deviations ranging from 9 to 88% (mean 36%) for intact bone and 6 to 64% (mean 19%) for bare
4-36
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bone. ANOVA results showed that after controlling for sampling variation contributed by age,
gender and leg sections, no significant difference was found for mean Pb concentrations
measured on intact bone and bare bone and by ICP-MS methods.
In the other cadaver study, Todd et al. (2000) measured 10 intact adult legs and the bare
tibiae dissected from 9 of those legs for investigation of the short-term variability in XRF results.
Each bare tibia was measured 6 times over 3 hours (without repositioning the tibia
measurements) at the same location as when measured intact. In addition, 4 volunteers
underwent monthly single measurements of the left tibia for 1 year. As a further check for
reproducibility, half-hour measurements on the bone standard, NIST 1400, were measured:
30 measurements were taken over a period of 4 days repositioning the sample between each
measurement, and 60 consecutive measurements were taken over a 30-h period without
repositioning the sample. They concluded that "the uncertainty in an individual measurement is
an underestimate of the standard deviation of replicate measurements, suggesting a
methodological deficiency probably shared by most current 109Cd-based K-shell XRF Pb
measurement systems."
As part of the same study of cadaver legs, 9 tibia were divided into cross-sectional
segments 2 cm apart, which were further separated into the tibia core and surface samples for the
AAS measurements (Todd et al. 2002b). The authors found no statistically significant difference
between mean XRF-measured concentrations and mean surface Pb concentrations measured by
AAS, but the XRF-measurements for tibia core Pb concentrations were significantly
overestimated by between 5 and 8 jig/g bone mineral. That is, XRF more closely reflects Pb
concentrations at tibia surface than in tibia core.
Another aspect of the cadaver studies of Todd et al. (2001) measured multiple locations
on the 10 intact legs and on the 9 bare tibiae. For example, each intact leg underwent single
XRF measurements at each of 10 locations along a middle track, extending to 10 cm above and
below the vertical midpoint of the tibia. Each of the 9 bare tibia underwent at least 6 XRF
measurements without repositioning the tibia between measurements at each centimeter location
on a 9 cm x 3 cm grid that covered the upper half of the tibia (27 locations). In bare tibia, mean
XRF results increased up and down from the vertical midpoint of the tibia, consistent with the
idea that the ends of tibia contain a larger component of trabecular bone than the middle section,
thereby increasing the 109Cd-based XRF result. This finding contrasted with those of Hoppin
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et al. (2000) and Todd et al. (2001) attributed the difference to the smaller range of vertical
displacement (±1 cm) in 4 bare bones measured by Hoppin et al. (2000). However, in single
XRF spectra taken at multiple locations on intact limbs, there was no detectable effect of vertical
location on XRF result; Todd et al. (2000) suggested that measuring away from the vertical
midpoint of the tibia should not be of practical concern when performing in vivo XRF
measurements. Aro et al. (2000) also concluded that bone Pb was reasonably homogeneously
distributed.
A comparison of Pb concentrations between right and left tibia in 12 human skeletons to
assess natural homogeneity of Pb content revealed natural symmetry with a calculated
correlation coefficient of 0.9 (Wittmers et al., 1988). K-shell XRF measurements on left and
right tibiae in 14 subjects showed no significant differences between legs (Gordon et al., 1994).
The reproducibility of XRF measurements over extended intervals has been investigated
in several studies. Gordon et al. (1994) measured 5 subjects five times on two occasions
10 months apart and found mean standard deviations of 3.4 and 5.1 |ig/g bone mineral for males
and females, respectively. Armstrong et al. (1992) measured tibia Pb concentrations on two
occasions separated by 5 years for a group of workers occupationally exposed to Pb. In 1983,
the average uncertainty in a single measurement was 9.3 |ig/g bone mineral with a mean Pb
concentration in 15 subjects of 54.4 |ig/g bone mineral. In 1988, the uncertainty was 4.9 |ig/g
and the mean value for 11 subjects was 44.2 jig/g. They suggested that the difference in
measurements separated by 5 years could be accounted for by counting statistics. Todd et al.
(2000) performed 27 replicate measurements on 10 intact cadaver legs on the same location over
a period of 4% months. The found the average difference between the (average) XRF results
from short term and longer term measurements was 1.2 |ig/g "showing there is a reassuringly
small amount of variability in the XRF results over a sustained period of time" (p. 3743). They
also performed monthly measurements on 4 adult volunteers (2 male, 2 female) over 1 year.
Tibia Pb varied from 6.4 to 12.9 jig/g bone mineral and standard deviations of the measurements
ranged from 4.9 to 9.9 |ig/g.
Attenuation of X-rays by skin and hair can affect the bone Pb measurements.
For example, normalization of the Pb X-rays to the elastic scatter was considered to render the
accuracy of measurements insensitive to variations in overlying tissue thickness (Chettle et al.,
1991; Hu et al., 1995). Todd et al. (2000, p 3737) state that: "The principal factor influencing a
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human subject's bone Pb measurement uncertainty is the thickness of tissue overlying the bone
that is measured" but concluded (page 3742) from the difference in average XRF results for
intact and bare bone that the bone Pb measurements were qualitatively independent of the
presence of overlying tissues.
However, in a study of young adults, McNeill et al. (1999) found that the measurement
uncertainties were greater than uncertainties for occupationally exposed males and attributed this
to inclusion of obese subjects and females in their study; uncertainties increased with bone body
mass index and were poorer for females than males.
In a study of 108 former female smelter employees and 99 referents, Popovic et al. (2005)
suggest that a high body mass index can give distorted values.
In the most recent study using nine leg phantoms of different soft tissue thickness, Ahmed
et al. (2006) found that by increasing the overlying tissue thickness from 3.2 to 14.6 mm, there
was an increase in average measurement uncertainty by a factor of 2.4 and an increase in
minimum detectable limit by a factor of 2.46.
Since 1986, several investigators have reported refinements to hardware and software to
improve the precision and accuracy of XRF measurements and there have been a number of
investigations into the precision, accuracy and variability in XRF measurements (e.g., Aro et al.,
2000; Todd et al., 2000, 2001, 2002). Todd et al. (2000) provided a detailed discussion of
factors that influence the variability and measurement uncertainty, including repositioning,
sample measurement duration, overlying tissue, operator expertise, detector resolution, and
changes to measurement process over time. Some of these aspects were also discussed by
Hu et al. (1995). From their cadaver and in vivo measurements, Todd et al. (2000) concluded
that the uncertainty in an individual measurement was an underestimate of the standard deviation
of replicate measurements, suggesting a methodological deficiency probably shared by most
current 109Cd-based K-shell XRF Pb measurement systems. In examining the reproducibility of
the bone Pb measurements over a 41/2 month period, Todd et al. found the average difference
between the XRF results from short term and longer term measurements was 1.2 |ig/g, indicating
only a small amount of variability in the XRF results over a sustained period of time.
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Statistical Methods for Analyzing Bone Lead Concentrations in Epidendologic Studies
In the literature, XRF bone Pb data have typically been reported in two ways: one that
involves a methodological approach to assessing the minimum detection limit and the other
termed an epidemiologic approach by Rosen and Pounds (1998). In the methodological
approach, a minimum detection limit is defined using various methods, including two or three
times the square root of the background counts; one, two, or three times the SD of the
background; and two times the observed median error. This approach relies upon the minimum
detection limit to define a quantitative estimate that is of sufficient precision to be included in the
statistical analysis. The following are examples of methodological minimum detection limits for
bone Pb analyses. Bellinger et al. (1994) observed minimum detection limits, equivalent to the
SD, of 5.4 |ig/g for tibia and 9.2 jig/g for patella. Using twice the median observed error,
Gerhardsson et al. (1993) observed minimum detection limits of 9.8 |ig/g for tibia and 19.1 |ig/g
for calcaneus. For finger bone Pb measurements, Chri staffers son et al. (1986) observed a
minimum detectable limit of 20 |ig/g, which was equivalent to three times the square root of the
background counts.
With the epidemiologic approach, to determine the minimum detection limit of an
instrument, all values are used (including negative values)—which results in extremely low
detection limits. Rosen and Pounds (1998) noted that this approach yields population bone Pb
averages that they considered artificially low and inconsistent with observations from many other
earlier studies. However, not including values that are negative or below the detection limit, or
assigning these values a fixed number for the statistical analysis is also of concern. To examine
and compare the two methods used to analyze data at low levels of bone Pb concentration,
Kim et al. (1995) performed serial measurements on phantoms containing spiked amounts of Pb.
The results demonstrated that the use of methodological minimum detection limits to recede
low-level observations reduced the efficiency of the analysis and the ability to distinguish
between the phantoms. Using the epidemiologic approach of retaining all point estimates of
measured bone Pb concentrations provided less bias and greater efficiency in comparing the
mean or median levels of bone Pb of different populations.
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4.3.2.3 Bone Lead as a Biomarker of Lead Body Burden
Uptake of Lead in Bone
The dominant compartment for Pb in the body is in bones. In human adults, more than
90% of the total body burden of Pb is found in the bones, whereas bone Pb accounts for -70% of
the body burden in children (Barry, 1975). Bone is comprised of two main types, cortical and
trabecular. The tibia consists of more than 95% cortical bone, the calcaneus and patella
comprise more than 95% trabecular bone, and finger bone is a mixed cortical and trabecular bone
although the second phalanx is dominantly cortical. The cortical and trabecular bones have
distinct rates of turnover and Pb release, as well as potentially different associated toxicity
implications (Hu et al., 1998). For example, adult tibia has a turnover rate of about 2% per year
whereas trabecular bone has a turnover rate of more than 8% per year (Rabinowitz, 1991). The
proportion of cortical to trabecular bone in the human body varies by age, but on average is
about 80 to 20 (International Commission on Radiological Protection, 1973). Although not so
important for certain types of measurements, the periosteum is of limited dimension and may
reflect a bone compartment of more rapid deposition and turnover of Pb than the other two types
(Skerfving et al., 1993), which would also likely have implications for toxicity, especially for
chelation therapy.
Much of the understanding of bone structure and metal deposition comes from studies of
radioactive elements (e.g., International Commission on Radiological Protection, 1996). Durbin
(1992, page 823) suggests that there is "an initial deposition of Pb on anatomical bone surfaces
with some skewing to the well nourished trabecular surfaces in red marrow, intense deposits at
bone growth sites, and later on, a nearly diffuse labeling throughout the bone volume.
For constant intake of Pb during growth, it is expected that Pb will be nearly uniformly
distributed in the mineralized bone. Single or irregular intakes during growth are expected to
result in residual buried lines and hotspots superimposed on a relatively uniform diffuse
concentration in bone mineral volume. For example, periosteal and subperiosteal Pb deposits in
the long bones, including those of the hands and feet, are likely to be greater than at many other
sites, since bone growth continues at the periosteal surface while the endosteal surface is
resorbed."
The importance of bone marrow was also stressed by Salmon et al. (1999), with a key
factor affecting Pb uptake into bone being the fraction of bone surface in trabecular and cortical
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bone adjacent to active bone marrow. The fraction of total marrow that is red and active
decreases from 100% at birth to about 32% in adulthood (Cristy, 1981). Early Pb uptake is
greater in trabecular bone due to its larger surface area and higher metabolic rate. Of the total
bone surface against red marrow, 76% is trabecular and 24% is cortical endosteal (Salmon et al.,
1999). Bone marrow has much lower Pb concentrations than bone matrix (Skerfving
etal., 1983).

Half-life of Lead in Bone
Estimates of the half-life of Pb in trabecular bone are partly dependent on the tissue
analyzed and the "purity" of the trabecular component (e.g., patella, calcaneus, and phalanx).
Earlier estimates of the half-life of Pb in trabecular bone ranged from 12 to 19 years (Bergdahl
et al., 1998; Gerhardsson et al., 1993). For cortical bone, estimates for the half-life of Pb
were on the order of 13 to 27 years (Bergdahl et al., 1998; Gerhardsson et al., 1993;
Rabinowitz, 1991).
With respect to half-lives in bone, recent K-shell XRF bone studies have indicated that
earlier concepts of a constant rate of removal of Pb from bone throughout adulthood assumed in
models of human metabolism (Leggett, 1993; O'Flaherty, 1993) may be incorrect. In a study of
active and retired smelter workers, Brito et al. (2001) suggested that people less than 40 years old
had a shorter half-life for the release of Pb from the tibia than those older than 40 years, 4.9 years
(95% CI: 3.6, 7.8) compared to 13.8 years (95% CI: 9.7, 23.8), respectively. Also, they
suggested that less intensely exposed subjects with a lifetime averaged blood Pb of <25 |ig/dL
had a shorter half-life in the tibia (6.2 years [95% CI: 4.7, 9.0]) than those with a lifetime
averaged blood Pb >25 |ig/dL (14.7 years [95% CI: 9.7, 29.9]).
Even by the end of the sixth decade, -35 to 40% of skeletal mass consists of
unremodelled first generation bone acquired during childhood and adolescence (International
Commission on Radiological Protection, 1973). This statement contrasts with that of O'Flaherty
(1993) who suggested that because of the relatively short half-life of Pb in the bones of children
that much of the Pb incorporated during active growth would not persist into adulthood. In a
comparison of Pb in tooth dentine and the tibia from young adults who were followed up after a
period of 13 years, Kim et al. (1996) suggested that "pockets" of Pb acquired in childhood may
persist into adults. Likewise, McNeill et al. (2000) compared tibia Pb levels and cumulative
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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, Idaho smelter. They concluded that Pb from exposure in early
childhood had persisted in the bone matrix until adulthood.

Changes in Bone Lead Concentrations with Age
Conventional and XRF analyses of bone have shown significant increases in bone Pb with
age (Hu et al., 1990, 1996; Kosnett et al., 1994; Morgan et al., 1990). Kosnett et al. (1994)
observed no significant change in bone Pb concentrations up to age 20 years, but found an
increasing trend with the same slope for men and women between the ages of 20 to 55 years and
an increase to a faster rate in men older than 55 years. Kosnett et al. reanalyzed earlier cadaver
cortical bone data of Drasch et al. (1987) and found that male bone Pb values increased
significantly after age 40 years, whereas female values slightly declined. A similar analysis of
the post-mortem data of Barry (1975) showed an upward inflection for all males after age
35 years. Kosnett et al. (1994) found no significant slope to the relationship between age and
bone Pb for the 10 to 20 year old subjects, in contrast to Barry (1975) and Drasch et al. (1987).
Kosnett et al. (1994) further noted that relatively high environmental Pb levels characterized
various populations in the past and would have resulted in higher levels of bone Pb deposition;
a portion of the increase of bone Pb with age is thus likely attributable to an exposure cohort
effect.
Annual increments of Pb to bone vary, although no attempt has been made to determine
whether the differences are significant. For example, the annual increment of 0.46 |ig/g bone
mineral/year found by Gordon et al. (1993) was slightly lower than that found by Somervaille
et al. (1989), but the difference was not significant. After age 20 years, Kosnett et al. (1994)
found the annual increment to be 0.38 jig/g bone mineral/year. Hu et al. (1990) reported a value
of 0.31 |ig/g bone mineral/year for subjects ranging in age from 20 to 58 years. Thus,
interpreting variations in bone Pb as a function of age is complex.

4.3.2.4 Distribution of Lead from Bone into Blood and Plasma
Contribution of Bone Lead to Blood Lead
Although the skeleton was recognized as a potentially significant contributor to blood Pb
in the 1986 Lead AQCD, there have been several investigations using both bone Pb XRF and
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stable Pb isotope methods that have helped quantify the contribution. The earlier estimation of
skeletal contribution to blood Pb was 70% by Manton (1985) and -65% ranging up to 100% by
Schiitz et al. (1987b). The more recent isotope studies confirmed these estimates. Using female
immigrants to Australia and their children, Gulson et al. (1995, 1997, 1999a) found a mean value
of 50% (range 16-73%) deriving from the skeleton. Smith et al. (1996) found a range of 40-70%
in five patients who underwent total hip or knee joint replacement. Gwiazda et al. (2005)
observed a range of 40-65% in two children and >90% in one child. Studies examining the bone
Pb contribution to blood Pb are presented in Annex Table AX4-6.
The contribution of skeletal Pb to blood Pb was further examined in females from varying
environments. In middle-aged to elderly subjects (46-74 years), an increase of 19 |ig/g of Pb in
tibia bone mineral was associated with an increase in blood Pb of 1.7 |ig/dL, which corresponds
to a 0.09 |ig/dL increase in blood Pb per 1 |ig/g bone mineral (Korrick et al., 2002). A study of
108 former workers at the Bunker Hill smelter in northern Idaho and 99 referents from the
Spokane, WA area examined the endogenous bone Pb release rate of postmenopausal and
premenopausal women (Popovic et al., 2005). The results indicated that the endogenous release
rate in postmenopausal women (0.13 |ig/dL per jig/g bone) was greater than the rate found in
premenopausal women (0.07 |ig/dL per |ig/g bone). In a Mexico City study, the endogenous
bone Pb release rate in postmenopausal women also was observed to be double that in
premenopausal women (Garrido-Latorre et al., 2003). A change of 10 jig/g bone mineral was
associated with an increase in blood Pb of 1.4 |ig/dL in postmenopausal subjects, compared to an
increase of 0.8 |ig/dL in premenopausal women. Lactation was also found to affect the
endogenous bone Pb release rate. After adjusting for patella Pb concentration, an increase in
blood Pb levels of 12.7% (95% CI: 6.2, 19.6) was observed for women who practiced partial
lactation and an increase of 18.6% (95% CI: 7.1, 31.4) for women who practiced exclusive
lactation compared to those who stopped lactation (Tellez-Rojo et al., 2002).
The mean cortical Pb to current blood Pb ratios for occupationally-exposed subjects are
shown in Figure 4-8. Box plots were calculated using data from the following studies: Bergdahl
et al., 1998; Brito et al., 2002; Christoffersson et al., 1984; Erfurth et al., 2001; Erkkila et al.,
1992; Fleming et al., 1998; Gerhardsson et al., 1993; Hanninen et al., 1998; Juarez-Perez et al.,
2004; Popovic et al., 2005; Roels et al., 1995; Schwartz et al., 2000a,b; Somervaille et al., 1988,
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6.0

o
_o
QQ
O
*•"
•o
TO
5
_J 20
03
O
"t
O
o
0.0
T
T
T
Active
Retired
Referent
Figure 4-8. Cortical lead to blood lead ratios for occupationally-exposed subjects
(both active and retired) and referents. Data compiled from several studies.
See text for more details.
1989; Todd et al., 2001. The mean cortical Pb to current blood Pb ratio is about 1.2 (range
0.4-2.6) for active employees (n = 17). For retired employees (n = 7), the mean is 3.2 (range
2.0-5.3), while for environmentally-exposed referent subjects from these industries (n = 7) the
mean ratio is about 1.3 (range 1-2.2). The differences in the cortical Pb to blood Pb ratio
between active and retired employees and retired employees and referents are significant
(p < 0.01) but not between active employees and referents. Several investigators have pointed
out the weak association between bone Pb and blood Pb in active employees in comparison with
the stronger association with retired employees (e.g., Erkkila et al., 1992; Fleming et al., 1997;
Gerhardsson et al., 1993). This is likely because circulatory Pb of active employees reflects
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mainly ongoing exposure, whereas that in retired employees is more dependent on Pb released
from the skeleton.
The mean tibia Pb to current blood Pb ratios for environmentally-exposed subjects are
shown in Figure 4-9. The box plot for pregnancy-related subjects was calculated using data from
the following studies: Brown et al., 2000; Chuang et al., 2001; Ettinger et al., 2004; Gomaa
et al., 2002; Gonzalez-Cossio et al., 1997; Hernandez-Avila et al., 1996, 1998, 2002, 2003;
Hu et al., 1996; Moline et al., 2000; Rothenberg et al., 2000; Sanin et al., 2001; Tellez-Rojo
et al., 2002, 2004. The box plot for middle-aged and elderly subjects included the following
studies: Berkowitz et al., 2004; Cheng et al., 1998; Garrido-Lattore et al., 2003; Hu et al., 1996,
2001; Korrick et al., 2002; Kosnett et al., 1994; Oliveira et al., 2002; Schafer et al., 2005; Tsaih
et al., 2004; Webber et al., 1995. The box plot for the younger subjects (age range 1-30 years)
included Farias et al., 1998; Kim et al., 1996; Rosen et al., 1989; Stokes et al., 1998. The mean
tibia Pb to blood Pb ratio for pregnancy-related subjects (n = 21) is 1.5 (range 1.0-4.2) and is
statistically significantly different (p < 0.001) from the mean ratio of 3.4 (range 1.6-5.4) for
middle-aged to elderly subjects (n = 27). Similar relationships are observed for the patella Pb to
blood Pb ratios for pregnancy-related subjects and middle-aged to elderly subjects.
In several other studies of environmentally-exposed subjects, there is a stronger
relationship between patella Pb and blood Pb than tibia Pb and blood Pb (e.g., Hernandez-Avila
et al., 1996; Hu et al., 1996, 1998). Hu et al. (1998) suggest that these relationships indicate that
trabecular bone is the predominant bone type providing Pb back into circulation under steady-
state and pathologic conditions. The stronger relationships 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.

Partitioning of Bone Lead into Plasma
Although most of the Pb in whole blood is associated with erythrocytes (-99%), it has
been suggested that the small fraction of Pb in plasma (<0.3%) may be the more biologically
labile and lexicologically active fraction of the circulating Pb. Several authors have proposed
that Pb released from the skeleton was preferentially partitioned into serum compared with red
cells (Cake et al., 1996; Hernandez-Avila et al., 1998; Tsaih et al., 1999) with one explanation
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6.0
.2
o:
o
_o
DO
eg
In
4.0
2.0
0.0
O
T
T
T
Pregnancy Middle-aged Young
to Elderly
Figure 4-9. Tibia lead to blood lead ratios for environmentally-exposed pregnancy-related
subjects, middle-aged to elderly subjects, and younger subjects.
Data compiled from several studies. See text for more details.
being that the Pb from endogenous sources was in a different form to that from exogenous
sources. In the latter study, Tsaih et al. (1999) suggested that urine was a satisfactory proxy for
serum. However, this concept has been withdrawn by its main proponents. In matched blood
and urine samples from 13 migrant subjects to Australia who were monitored prior to and during
pregnancy, there was no statistically significant difference in the 206Pb/204Pb and 207Pb/206Pb
ratios over pregnancy and the urine results for the postpartum period were in the opposite
relationships to those predicted for a preferential partitioning hypothesis (Gulson et al., 2000).

4.3.2.5 Mobilization of Lead From Bone
Although earlier investigators such as Brown and Tompsett (1945), Ahlgren et al. (1976)
and Christoffersson et al. (1984) suggested that the skeleton was a potential endogenous source
of Pb poisoning, the opposing concept of the skeleton as a "safe" repository for Pb persisted until
the mid-1980s and early 1990s. Potential mobilization of Pb from the skeleton could occur at
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times of physiological stress associated with enhanced bone remodeling such as during
pregnancy and lactation (Hertz-Picciotto et al., 2000; Manton, 1985; Silbergeld, 1991),
menopause or in the elderly (Silbergeld, 1991; Silbergeld et al., 1988), extended bed rest
(Markowitz and Weinberger, 1990), hyperparathyroidism (Kessler et al., 1999), and
weightlessness. The Pb deposited in the bone of adults can serve to maintain blood Pb levels
long after exposure has ended (Fleming et al., 1997; Gulson et al., 1995; Inskip et al., 1996;
Kehoe, 1987; Manton, 1985; Nilsson et al., 1991; O'Flaherty et al., 1982; Schutz et al., 1987b;
Smith etal., 1996).
In the 1986 Lead AQCD, there was a comprehensive summary of chelation therapies and
the recognition that there was limited release of Pb from bones. The potential role of bone Pb as
an endogenous source of Pb in blood (resulting in elevated levels for former Pb employees) was
mentioned, although data to support this hypothesis were limited.

Mobilization of Lead from Bone during Pregnancy and Lactation
Bone Pb studies of pregnant and lactating subjects are summarized in Annex
Table AX4-7. Most of the bone XRF studies on pregnancy and lactation have focused on
subjects from Mexico City and Latin subjects from Los Angeles, California. Relationships
and/or health outcomes from these investigations include: patella bone as a significant
contributor to blood Pb (Brown et al., 2000; Hernandez-Avila et al., 1996); a positive association
between plasma Pb and bone Pb in the highest bone Pb group of pregnant women (Tellez-Rojo
et al., 2004); a positive association of tibia and calcaneus Pb with prenatal Pb concentration, and
calcaneus Pb with postnatal Pb (Rothenberg et al., 2000); a positive association of tibia Pb and
seasonal variations in blood Pb (Rothenberg et al., 2001); maternal tibia and patella Pb as
significant predictors of fetal exposure determined using cord blood (Chuang et al., 2001);
a positive association of calcaneus Pb and increased systolic and diastolic blood pressure in the
third trimester (Rothenberg et al., 2002); an inverse relationship between maternal tibia and
patella Pb, and birth weight (Gonzalez-Cossio et al., 1997; Sanin et al., 2001); an inverse
association between tibia Pb and birth length, and patella Pb and head circumference
(Hernandez-Avila et al., 2002); an inverse association of maternal patella bone and Mental
Development Index (Gomaa et al., 2002); increased bone resorption during lactation (Tellez-
4-48
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Rojo et al., 2002); increased Pb in breast milk with an increase in patella and tibia Pb (Ettinger
et al., 2004).
Lead isotope studies on immigrant women to Australia (Gulson et al., 1997, 1998a)
confirmed the earlier work of Manton (1985) of increased blood Pb during pregnancy. Gulson
et al. reported that, during pregnancy, blood Pb concentrations in the first immigrant cohort
(n = 15) increased by an average of about 20% compared to non-pregnant migrant controls
(n = 7). The percentage change in blood Pb concentration was significantly greater during the
postpregnancy period than during the second and third trimesters (p < 0.001). Skeletal
contribution to blood Pb, based on the isotopic composition for the immigrant subjects, increased
in an approximately linear manner during pregnancy. The mean increases for each individual
during pregnancy varied from 26% to 99%. Skeletal Pb contribution to blood Pb was
significantly greater during the postpregnancy period than during the second and third trimesters.
The contribution of skeletal Pb to blood Pb during the postpregnancy period remained essentially
constant at the increased level of Pb mobilization. In a follow-up study using a different
immigrant cohort of 12 women with calcium supplementation at the recommended level of
-1,000 mg/day (National Institutes of Health, 1994), Gulson et al. (2004) found increased
mobilization of Pb occurred in the third trimester rather than in the second trimester as observed
with first cohort. In addition, the extra flux released from bone during late pregnancy and
postpartum varied from 50 to 380 jig (geometric mean 145 jig) compared with 330 jig in the
previous cohort.
Also examing blood Pb during pregnancy, Manton et al. (2003) observed that blood Pb
concentrations decreased in early pregnancy and rose during late pregnancy. These investigators
attributed their results to changes in bone resorption with decoupling of trabecular and cortical
bone sites.

Transplacental Transfer of Lead and Transfer through Breast Milk
Transplacental transfer of Pb in humans has been suggested by a number of studies based
on cord blood to maternal blood Pb ratios ranging from about 0.6 to 1.0 at the time of delivery.
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 (Macaco, fascicularis). Approximately 7 to
4-49
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39% of the maternal Pb burden transferred to the fetus appears to derive from the maternal
skeleton (Franklin et al., 1997; O'Flaherty et al., 1998). Further evidence for maternal-to-fetal
transfer of Pb in humans has been gained from stable Pb isotope measurements. For example,
a 0.99 correlation in Pb isotopic ratios for maternal and cord blood (Manton, 1985; Gulson et al.,
1998b) and the similarity of isotopic ratios in maternal blood and in blood and urine of newly-
born infants provide strong evidence for placental transfer of Pb to the fetus (Gulson et al.,
1999b).
Breast milk can also be a pathway of maternal excretion of Pb. However, given the very
low Pb concentrations and analytical difficulties arising from high fat contents in breast milk,
their analyses require careful attention. Selected studies appear to show a linear relationship
between breast milk and maternal whole blood with the percentage of Pb in breast milk
compared with whole blood of <3% in subjects for blood Pb concentrations ranging from 2 to
34 |ig/dL. Blood Pb concentrations in breastfed newborn infants decreased in spite of the
maternal blood Pb concentrations having risen or remained elevated postpartum compared to
lower levels during prepregnancy or in the first trimester (Gulson et al., 1999b). Similar trends
were noted by Manton et al. (2000). However, in a Mexico City study, an association between
patella Pb and blood Pb concentrations was higher for women with partial lactation than for
those who stopped lactation, and it was increased among women who breastfed exclusively
(Tellez-Rojo et al., 2002). In another Mexico City study, Ettinger et al. (2004) concluded that an
interquartile increase in patella Pb was associated with a 14% increase in breast milk Pb, whereas
for tibial Pb the increase was -5%.
In conclusion, there is evidence that maternal-to-fetal transfer of Pb occurs, likely
resulting from the mobilization of Pb from the maternal skeleton during pregnancy. Breast-fed
infants appear to be at greater risk only if the mother is exposed to high Pb concentrations either
from exogenous sources or endogenous sources such as the skeleton.

Mobilization of Lead in Bone during Menopause and in the Elderly
Increases in blood Pb for postmenopausal women have been attributed to release of Pb
from the skeleton associated with increased bone remodeling during menopause. Many of the
studies have been based on blood Pb concentration. Bone Pb studies of menopausal and middle-
aged to elderly subjects are summarized in Annex Table AX4-8.
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Overall, the various studies of bone and blood Pb levels, as well as hormone replacement
therapy, have provided conflicting outcomes. Hormone replacement therapy alone or combined
with calcium supplementation prevents bone resorption and increases the bone mineral density in
trabecular and cortical bones of women with or without metabolic bone disease. The effect of
hormone replacement therapy may result in a decrease of Pb mobilization from bone along with
a reduction in blood Pb levels. Several studies have found that tibia bone Pb levels were higher
in women who used hormone replacement therapy (Popovic et al., 2005; Webber et al., 1995).
In contrast, other investigators have found no association between bone Pb and use of estrogens
(Berkowitz et al., 2004; Korrick et al., 2002). In addition, some studies observed a decrease in
blood Pb concentrations associated with hormone replacement therapy (Garrido-Latorre et al.,
2003), whereas others observed no association (Webber et al., 1995).
The endogenous release rate of Pb from bone in postmenopausal women was double the
rate in premenopausal former smelter employees (Popovic et al., 2005) and environmentally-
exposed women from Mexico (Garrido-Latorre et al., 2003). In middle-aged to elderly males
from the Normative Aging Study, patella Pb accounted for the dominant portion of variance in
blood Pb(Huetal., 1996).

Effect of Nutritional Status on Mobilization of Lead from Bone
Most studies that investigated the effect of nutritional status on the mobilization of Pb
from the skeleton have examined the effects of calcium supplementation. Several studies have
suggested that dietary calcium may have a protective role against Pb by decreasing absorption of
Pb in the gastrointestinal tract and by decreasing the mobilization of Pb from bone stores to
blood, especially during periods of high metabolic activity of the bone such as pregnancy,
lactation, and menopause. An inverse association between patella Pb and low calcium intake in
postpartum women has been found (Hernandez-Avila et al., 1996). In contrast, Rothenberg et al.
(2000) observed that dietary calcium intake had no effect on calcaneus Pb in women monitored
during the third trimester and 1 to 2 months postpartum. Likewise, no effect from calcium
supplementation on bone Pb was found amongst lactating women from Mexico City (Tellez-
Rojo et al., 2002), although in a follow-up study, Hernandez-Avila et al. (2003) reported a
16.4% decrease in blood Pb concentration among women with the highest patella bone Pb levels
who were taking supplements. Gulson et al. (2004) observed that calcium supplementation was
4-51
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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 calcium intake was 850 mg per day showed lower amounts of bone resorption during late
pregnancy and postpartum than those whose intake was 560 mg calcium per day (Manton et al.,
2003). Tellez-Rojo et al. (2004) observed that plasma Pb levels were inversely related to dietary
calcium intake. Results for whole blood Pb were similar but less pronounced.
Some researchers have noted concerns regarding potential Pb toxicity resulting from
calcium supplementation. However, Gulson et al. (2001) observed that Pb in calcium or vitamin
supplements did not appear to increase blood Pb concentrations. No information was available
on the effects of other nutritional supplements (e.g., iron or zinc) on Pb body burden.

4.3.2.6 Summary of Bone Lead as a Biomarker of Lead Body Burden and Exposure
Bone accounts for more than 90% of the total body burden of Pb in adults and 70% in
children. In addition, the longer half-life of Pb in bone, which largely depends on the bone type
but is generally estimated in terms of years compared to days for blood Pb, allows a more
cumulative measure of Pb dose. The more widespread use of in vivo XRF Pb measurements in
bone and indirect measurements of bone processes with stable Pb isotopes since the 1986 Lead
AQCD have enhanced the use of bone Pb as a biomarker of Pb body burden.
In addition to considering bone Pb as an indicator of cumulative Pb exposure, Pb in the
skeleton can also be regarded as a source of Pb. Key studies have examined the contribution of
bone Pb to blood Pb; the preferential partitioning of bone Pb into plasma; mobilization of Pb
from bones during pregnancy, lactation, and menopause; and the role of nutritional
supplementation in bone mobilization.

4.3.3 Lead in Teeth
4.3.3.1 Summary of Key Findings from the 1986 Lead AQCD
The importance of dentine as a potential indicator of Pb exposure was noted in the 1986
Lead AQCD. There was more emphasis and optimism on using dentine to assess Pb exposure in
this document as the bone XRF method was in its infancy. The issue of deciduous tooth type
was addressed but there was little information on permanent teeth. The portion of the tooth
analyzed (i.e., whole tooth or circumpulpal dentine) was also addressed. In the 1990
4-52
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Supplement, the use of tooth Pb as an exposure metric was described in a number of the
longitudinal and cross-sectional studies.

4.3.3.2 Analytical Methods for Measuring Lead in Teeth
Analytical methods for tooth analysis vary from the most widely used AAS, to energy-
dispersive XRF, laser ablation inductively coupled plasma mass spectrometry (LA-ICP-MS), and
high precision Pb isotopes.
As a standard analytical method has yet to be established for tooth Pb analysis, some of
the discrepancies in findings between studies could arise from several factors, including
differences in tooth type, part of the tooth analyzed, and tooth location. Any real differences
among populations are unlikely to be the result of physiological factors such as blood supply to
teeth or mineralization rates. As enamel and dentine in different teeth calcify at overlapping but
different times (Orban, 1953), they could retain varying amounts of Pb.
In a systematic evaluation of the magnitude of random errors associated with dentine Pb
measurements, Fergusson et al. (1989) measured Pb concentrations in two samples of dentine
from 996 New Zealand children. They estimated that 15 to 20% of the variance was
unexplained. Tests of differences of means and variances showed no significant differences
between the two samples.
Lead measurements in deciduous teeth in individuals from urban and remote
environments and from polluted environments are presented in Annex Tables AX4-9 and
AX4-10, respectively. Based on the limited number of studies, it would appear that the range in
whole deciduous tooth Pb for environmentally exposed subjects is about 1-10 jig/g, but the most
likely levels are <5 jig/g and probably even <2 jig/g. Studies of whole deciduous teeth from
industrial environments, including those in urban settings, are also commonly much less than
10 |ig/g.
The utility of circumpulpal dentine (Shapiro et al., 1973) as the metric of Pb exposure in
deciduous teeth has not been enthusiastically received. This is likely due to the separation
difficulties, as well as the limited amount of circumpulpal dentine that may be present when the
teeth are resorbed, prior to exfoliation.
In another approach to gain more information about exposure during pregnancy and early
childhood, the teeth may be sectioned into dominantly enamel or dominantly dentine. These
4-53
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samples can then be analyzed for Pb isotopic ratios and Pb concentrations (Gulson and Wilson,
1994). Even for children living in Pb mining and smelting communities, Pb levels in the enamel
are generally low (<5 |ig/g) and are consistent with other studies of whole teeth. However,
higher levels are observed in the dentine samples (e.g., 32 |ig/g), which likely reflect the early
childhood exposure. Permanent teeth tend to have up to three times the level of Pb compared
with deciduous teeth, but the number of studies is very limited.

4.3.3.3 Tooth Lead as a Biomarker of Lead Body Burden
Compared with the amount of Pb in the skeleton, tooth Pb is a minor contributor to the
body burden of Pb. Most of the tooth Pb information is based on analyses of deciduous teeth.
There is still controversy over the amounts of Pb in different whole teeth but it appears that the
highest concentrations are in central incisors, with decreasing amounts in lateral incisors,
canines, first molars, and second molars. Teeth from the upper jaw tend to have higher Pb
concentrations than those from the lower jaw.
As teeth accumulate Pb, tooth Pb levels are generally considered an estimate of
cumulative Pb exposure. Rabinowitz et al. (1993) found that tooth Pb was a better measure of
exposure than current blood Pb levels; however, it was not a good measure of the child's
cumulative exposure from birth to exfoliation due to the mobilization of Pb from dentine.
Teeth are composed of several tissues formed over the years. 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). This may allow investigators to elucidate
the history of Pb exposure in a child.
Gulson and Wilson (1994) advocated the use of sections of enamel and dentine to obtain
additional information compared with analysis of the whole tooth (e.g., Fosse et al., 1995;
Tvinnereim et al., 1997). For example, deciduous tooth Pb in the enamel provides information
about in utero exposure whereas that in dentine from the same tooth provides information about
postnatal exposure until the tooth exfoliates at about 6 to 7 years of age.

4.3.3.4 Relationship Between Tooth Lead and Blood Lead
As with bone Pb-blood Pb relationships, there is interest in understanding more about
potential relationships between tooth Pb and blood Pb. The tooth Pb-blood Pb relationship is
4-54
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more complex than the bone Pb-blood Pb relationship because of differences in tooth type,
location, and analytical method.
Rabinowitz (1995) used studies which reported values for dentine, whole shed teeth, or
crowns, but discarded those measuring circumpulpal dentine because of the higher values in this
medium. The mean tooth Pb levels varied from 2.8 to 12.7 jig/g and blood Pb levels from 6.5 to
17 |ig/dL. In a plot of blood versus tooth Pb, Rabinowitz found a good fit (R2 = 0.97;
p < 0.0001) with the relationship:

Tooth Lead (|ig/g) = P x [Blood Lead (jig/dL)], where P = 0.49 (SE 0.04) (4-1)

In an earlier Boston study, Rabinowitz et al. (1989) found that the association between
tooth and blood Pb increased with age, first achieving statistical significance at 18 months; by
57 months, the correlation coefficient was 0.56. A correlation of 0.47 was found between
current blood Pb and incisor Pb concentrations amongst 302 German children (Ewers et al.,
1982).

4.3.3.5 Mobilization of Lead from Teeth
Although mobilization of Pb from bone appears well established, this is not the case for
Pb in teeth. Conventional wisdom has Pb fixed once it enters the tooth. Although that may be
the case for the bulk of enamel, it is not true for the surface of the enamel and dentine.
In evaluating deciduous teeth data, Rabinowitz et al. (1993) suggested that their data were
compatible with a model that allows Pb to be slowly removed from dentine. Such a process may
be associated with resorption of the root and dentine that precedes exfoliation, which allows
reequilibration of dentine Pb with blood Pb.
In children exposed to Pb sources from mining, paint, or petrol in communities such as
the Broken Hill Pb mining community, Gulson and Wilson (1994) and Gulson (1996) showed
that the source of Pb from the incisal (enamel) sections was different from the source of Pb in the
cervical (dentine) sections of deciduous teeth, reflecting the change in Pb from in utero exposure
to early childhood. Based on changes in the isotopic composition of enamel and dentine in
deciduous teeth sections from the Broken Hill mining community children, Gulson (1996)
estimated that Pb is added to dentine at a rate of-2-3% per year.
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Stable Pb isotopes and Pb concentrations were measured in the enamel and dentine of
permanent (n = 37) and deciduous teeth (n = 14) from 47 European immigrants to Australia to
determine whether Pb exchange occurs in teeth and how it relates to Pb exchange in bone
(Gulson et al., 1997). The authors concluded that enamel exhibited no exchange of its European-
origin Pb with Pb from the Australian environment, whereas dentine Pb exchanged with
Australian Pb to the extent of ~1 ± 0.3% per year.

4.3.3.6 Summary of Tooth Lead as a Biomarker of Lead Body Burden and Exposure
Tooth Pb is a minor contributor to the total body burden of Pb. Moderate-to-high
correlations have been observed between tooth Pb levels and blood Pb levels. Differences in
tooth type, part of the tooth analyzed, and tooth location may contribute to some of the
discrepancies in findings between studies of tooth Pb. As teeth are composed of several tissues
formed over the years, if a child's Pb exposure during the years of tooth formation varied widely,
different amounts of Pb would be deposited at different rates. Deciduous tooth Pb in the enamel
provides information about in utero exposure, whereas that in dentine provides information about
postnatal exposure until the tooth exfoliates.

4.3.4 Lead in Urine
4.3.4.1 Summary of Key Findings from the 1986 Lead AQCD
The 1986 Lead AQCD provided an extensive discussion of the physiological basis for
"chelatable" urinary Pb. Also discussed was Pb excretion provoked by EDTA, including the
pools of Pb in the body that might be mobilized in the EDTA provocation test, and the
relationship between the outcome and blood Pb concentration. The 1986 Lead AQCD noted
observations that formed the basis for application of the EDTA provocation test for detecting
elevated Pb body burden.

4.3.4.2 Analytical Methods for Measuring Lead in Urine
Standard methods that have been reported for urine Pb analysis are summarized in Annex
Table AX4-1 and are, in general, the same as those analyses noted for determination of Pb in
blood. Reported detection limits are -50 |ig/L for AAS, 5-10 |ig/L for ICP AES, and 4 |ig/L for
ASV for urine Pb analyses. Sample preparation usually consists of wet ashing; however,
4-56
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chelation and solvent extraction has also been reported (National Institute for Occupational
Safety and Health, 1994, 1977a).

4.3.4.3 Levels of Lead in Urine
A summary of selected measurements of urine Pb levels in humans can be found in Annex
Table AX4-11. Urine Pb concentrations in the U.S. general population have been monitored in
NHANES. Data from the most recent survey (NHANES IV, Centers for Disease Control, 2005)
for subjects >6 years of age are shown in Table 4-4. The geometric mean for the entire sample
(n = 2,689) was 0.64 |ig/g creatinine (95% CI: 0.60, 0.68). The geometric means for males
(n = 1,334) and females (n = 1,335) were 0.64 |ig/g creatinine (95% CI: 0.61, 0.67) and
0.64 |ig/g creatinine (95% CI: 0.59, 0.69), respectively. These values correspond to an excretion
rate of-1-1.3 jig Pb/day for an adult, assuming a daily creatinine excretion rate of-1.5 g/day in
adult females, a body weight of 70 kg for males and 58 kg for females, and a lean body mass
fraction of 0.88 for males and 0.85 for females (Forbes and Bruining, 1976; International
Commission on Radiological Protection, 1975).

Table 4-4. Urine Lead Concentrations in U.S. by Age, NHANES IV (1999-2002)
Age
Survey Period
N
Urine Lead3
6-11 years
1999-2000
340
1.17
(0.98, 1.41)
2001-2002
368
0.92
(0.84, 1.00)
12-19 years
1999-2000
719
0.50
(0.46, 0.54)
2001-2002
762
0.40
(0.38,0.43)
^20 years
1999-2000
1406
0.72
(0.68, 0.76)
2001-2002
1559
0.66
(0.62, 0.70)
"Urine Pb concentrations presented are geometric means (95% CI) of ug-Pb/g-creatinine.
Geometric mean urinary Pb excretion rates of 7-10 |ig/g creatinine (maximum 43) have
been reported in groups of children living in areas impacted by Pb smelting operations
(Brockhaus et al., 1988). Daily urinary Pb excretion can exceed 200 |ig/day in association with
occupational exposures (Biagini et al., 1977; Cramer et al., 1974; Lilis et al., 1968; Lin et al.,
2001;Wedeenetal., 1975).
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4.3.4.4 Urine Lead as a Biomarker of Lead Body Burden
Urine is a major route of excretion of absorbed Pb (Chamberlain et al., 1978; Griffin
et al., 1975; Kehoe, 1987; Rabinowitz et al., 1976). 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 blood plasma. Evidence for this is the observation that
urinary Pb excretion is strongly correlated with the rate of glomerular filtration of Pb
(i.e., glomerular filtration rate x plasma Pb concentration; Araki et al., 1986). Estimates of
urinary clearance of Pb from serum (or plasma) range from 13-22 L/day, with a mean of
18 L/day (Araki et al., 1986; Chamberlain et al., 1978; Manton and Cook, 1984; Manton and
Malloy, 1983). Estimates of blood-to-urine clearance, on the other hand, range from
0.03-0.3 L/day with a mean of 0.12 L/day (Araki et al., 1990; Berger et al., 1990; Chamberlain
et al., 1978; Gulson et al., 2000; Koster et al., 1989; Manton and Malloy, 1983; Rabinowitz et al.,
1973, 1976; Ryu et al., 1983; see Diamond, 1992 for an analysis of these data), consistent with a
plasma to blood concentration ratio of-0.005-0.01 L/day (U.S. Environmental Protection
Agency, 2003b). Based on the above, 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).
Plasma Pb makes a small contribution (<1%) to the blood Pb concentration and a
negligible contribution to total Pb body burden. Furthermore, the kinetics of elimination of Pb
from plasma is fast, relative to Pb in bone, where most of the Pb burden resides. Therefore, the
basic concepts described for blood as a biomarker for Pb body burden also apply to urine.
A single urine Pb measurement, or a series of measurements taken over short-time span, is likely
a relatively poor index of Pb body burden (Figure 4-10). On the other hand, long-term average
measurements of urinary excretion can be expected to be a better index of body burden. In the
hypothetical simulation shown in Figure 4-10, both the long-term average urinary Pb excretion
rate and the body burden have approximately doubled.
The above considerations do not exclude the potential utility of urine Pb as a dose metric
in epidemiological studies. Some effect outcomes may be more strongly associated with plasma
concentrations of Pb (e.g., ferrochelatase inhibition) than Pb body burden. Given the technical
difficulties in accurately measuring the concentrations of Pb in plasma, especially at low blood
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30 40
Age (year)
50
OJ
en
2 >:
Q) CO
T3

-------
Pb levels (e.g., <10 |ig/dL), measurements of urinary Pb may serve as a more feasible surrogate
for measurements of plasma Pb concentration.

4.3.4.5 Relationship Between Lead in Blood and Urine
Assuming first-order kinetics, a plasma-to-urine clearance (UClp) of 13-22 L/day
corresponds to half-time for transfer of Pb from plasma to urine of 0.1-0.16 day for a 70 kg adult
who has a plasma volume (VP) of ~3 L:
(4'2)
This translates to a very rapid steady-state, much faster than observed for blood Pb after a change
in exposure level. The kinetics of change in urinary Pb excretion in response to a change in
exposure, therefore, will be determined by variables that affect the plasma Pb level, including
partitioning of Pb into erythrocytes and exchanges with Pb in soft tissues and mobile pools
within bone (e.g., bone surface). Here again, the basic concepts that apply to blood Pb as a
biomarker of exposure also apply to urine Pb. Urinary Pb excretion reflects, mainly, the
exposure history of the previous few months; thus, a single urinary Pb measurement cannot
distinguish between a long-term low level of exposure or a higher acute exposure. The
relationship between urinary Pb concentration and Pb uptake is thought to be linear, unlike that
for blood Pb concentration, although there are no direct empirical tests of this assumption in
humans. This assumption predicts a linear relationship between Pb intake (at constant absorption
fraction) and urinary Pb excretion rate. Figure 4-11 presents a simulated relationship between Pb
intake and urinary Pb excretion in adults and children using both the Leggett (1993) model and
O'Flaherty (1993, 1995) model. The major difference between the Leggett model and the
O'Flaherty model is in the assignment of the time dependence of bone Pb residence. The
Leggett model assumes a slow accumulation of a nonexchangable Pb pool, whereas the
O'Flaherty model assumes a gradual distancing of Pb from bone surfaces by diffusion
throughout the bone volume (O'Flaherty, 1998).
4-60
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140

120
•5 100

I 80
0)
u
X
UJ
60

8> 4°-l
1 20
_
z> 0
0
Adult .•*
Child
5 10 15
Lead Intake (pg/kg/day)
20
•^ 140
i^S
co
5 120
en
-r 1oo^
-53 so
u
X
UJ
T3
re
CD
c
60

40 ^

0
0
Adult
Child
10 20 30
Lead Intake (fjg/kg/day)
40
Figure 4-11. Simulation of relationship between lead intake and urinary lead excretion
in adults and children. Predictions are for a 2-year-old child and 30-year-
old adult, for a constant lead intake (jig/kg/day). The relationship is linear,
for intake and plasma lead concentration (not shown). Predictions
are based on Leggett (1993, upper panel) and O'Flaherty (1993,1995,
lower panel).
4-61
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It is important to emphasize that the above concepts apply to urinary Pb excretion rate, not
to urinary Pb concentration. The concentration of Pb in urine (Upb) is a function of the urinary
Pb excretion (UEPb) and the urine flow rate (UFR, L/day):

UEph=UPh-UFR (4-3)

Urine flow rate can vary by a factor or more than 10, depending on the state of hydration
and other factors that affect glomerular filtration rate and renal tubular reabsorption of the
glomerular filtrate. All of these factors can be affected by Pb exposure at levels that produce
nephrotoxicity (i.e., decreased glomerular filtration rate, impaired renal tubular transport
function; see Section 6.4 for discussion of effects of Pb on the renal system). Therefore, urine Pb
concentration measurements provide little reliable information about exposure (or Pb body
burden), unless they can be adjusted to account for unmeasured variability in urine flow rate
(Arakietal., 1990).
A determination of urinary Pb excretion rate requires measurement of two variables, urine
Pb concentration, and urine flow rate; the latter requires collection of a timed urine sample,
which is often problematic in epidemiologic studies. Collection of un-timed ("spot") urine
samples, a common alternative to timed samples, requires adjustment of the Pb measurement in
urine to account for variation in urine flow (Diamond, 1988). Several approaches to this
adjustment have been explored, including adjusting the measured urine Pb concentration by the
urine creatinine concentration, urine osmolality, or specific gravity (Araki et al., 1990).
The measurement of Pb excreted in urine following an injection (intravenous or
intramuscular) of the chelating agent calcium disodium EDTA (EDTA provocation) has been
used to detect elevated body burden of Pb in adults (Biagini et al., 1977; Lilis et al., 1968;
Wedeen, 1992; Wedeen et al., 1975) and children (Chisolm et al., 1976; Markowitz and Rosen,
1981). EDTA-provoked urinary Pb excretion has been shown to correlate with tibia bone Pb
measurements (Wedeen, 1992). Given difficulties associated with parenteral administration of
EDTA, XRF measurements of bone Pb, offer a more feasible alternative to the EDTA
provocation test for assessment of bone Pb stores in epidemiologic studies. More recently,
DMSA (DMSA-provocation) has been used as an orally-effective alternative to EDTA and has
been applied to epidemiologic studies as dose metric for Pb body burden (e.g., Lee et al., 2001;
Schwartz et al., 2001, 2000a,b).
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4.3.4.6 Summary of Urine Lead as a Biomarker of Lead Body Burden and Exposure
Similar to blood Pb concentration measurements, urinary Pb excretion measured in an
individual at a single point in time will reflect the recent exposure history of the individual and
physiological variables that determine the plasma Pb concentration time profile. As a result,
measurement of urinary Pb may serve as a more feasible surrogate for plasma Pb concentration,
and may be useful for exploring dose-response relationships for effect outcomes that may be
more strongly associated with plasma Pb concentration than Pb body burden. Longitudinal
measurements of urinary Pb excretion can be expected to provide a more reliable measure of
exposure history of an individual and will more closely parallel body burden than will single
measurements; however, the degree to which this will apply will depend on the sampling
frequency with respect to the exposure temporal pattern.
Although, in general, higher urinary Pb excretion can be interpreted as indicating higher
exposures (or Pb uptakes), it does not necessarily predict appreciably higher body burdens.
Similar urinary Pb excretion rates in two individuals (or populations) do not necessarily translate
to similar body burdens or similar exposure histories.
Measurement of the urinary Pb excretion rate requires either a timed urine sample, or an
approach to adjusting measured urinary Pb concentrations for variability in urine flow rate,
which by itself may be affected by Pb exposure (i.e., Pb-induced nephrotoxicity). Both
approaches, timed urine samples or adjustment of concentration, introduce complications into the
assessment and uncertainties into the interpretation of urinary Pb measurements as biomarkers of
Pb body burden or exposure. The EDTA-provocation test provides a more reliable indicator of
elevated Pb body burden than do measurements of basal Pb excretion; however, it is not feasible
to apply this test for epidemiologic investigations. The DMSA-provocation test may provide a
more feasible alternative.

4.3.5 Lead in Hair
4.3.5.1 Summary of Key Findings from the 1986 Lead AQCD
The 1986 Lead AQCD did not discuss applications of hair Pb measurements for assessing
Pb body burden or exposure.
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4.3.5.2 Analytical Methods for Measuring Lead in Hair
Methods used for hair Pb analysis are summarized in Annex Table AX4-1. Wilhelm et al.
(1989) reported a detection limit of 0.16 |ig/g for GFAAS; use of GFAAS for hair Pb
measurements has been reported elsewhere (Annesi-Maesano et al., 2003). Gerhardsson et al.
(1995a) reported a detection limit of 0.5 jig/g for XRF of the hair shaft; but Campbell and
Toribara (2001) found XRF to be unreliable for hair root Pb determinations. Use of other
methods has been reported, including ICP (Tuthill, 1996), ET/AAS (Drasch et al., 1997), and
AAS (Sharma and Reutergardh, 2000; Esteban et al., 1999).

4.3.5.3 Levels of Lead in Hair
A summary of selected measurements of hair Pb levels in humans can be found in Annex
Table AX4-12. Reported hair Pb levels vary considerably. Esteban et al. (1999) reported a
geometric mean level of 5.4 ng/g (range 1-39) for a sample of 189 children (aged 1.9 to
10.6 years) residing in Russian towns impacted by smelter and battery plant operations.
By contrast, Tuthill (1996) reported much higher levels in a sample of Boston, MA children
(aged 6.5 to 7.5 years, n = 277). Approximately 41% had levels that ranged from 1 to 1.9 |ig/g.
DiPietro et al. (1989) reported a geometric mean hair Pb level of 2.42 jig/g (10-90th percentile
range <1.0-10.8) in a general population sample of U.S. adults (aged 20 to 73 years, n = 270).
In a post-mortem sample of the general population from Germany (aged 16 to 93 years, n = 150),
the median hair Pb level was 0.76 |ig/g (range 0.026-20.6) (Drasch et al., 1997). Also,
Gerhardsson et al. (1995a) reported median values for postmortem samples of 8.0 jig/g
(range 1.5-29,000) in active workers (n = 6), 2.6 |ig/g (range 0.6-9.3) in retired workers (n = 23),
and 2.1 |ig/g (range 0.3-96) in a reference group (n = 10).

4.3.5.4 Hair Lead as a Biomarker of Lead Body Burden
Lead is incorporated into human hair and hair roots (Bos et al., 1985; Rabinowitz et al.,
1976) and has been explored as a possibly noninvasive approach for estimating Pb body burden
(Gerhardsson et al., 1995a; Wilhelm et al., 1989, 2002). Hair Pb measurements are subject to
error from contamination of the surface with environmental Pb and contaminants in artificial hair
treatments (i.e., dyeing, bleaching, permanents) and are a relatively poor predictor of blood Pb
concentrations, particularly at low levels (<10 to 12 |ig/dL) (Campbell and Toribara, 2001;
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Drasch et al., 1997; Esteban et al., 1999). Studies evaluating quantitative relationships between
hair Pb and Pb body burden have not been reported. Nevertheless, hair Pb levels have been used
as a dose metric in some epidemiologic studies (e.g., Annesi-Maesano et al., 2003; Esteban et al.,
1999; Gerhardsson et al., 1995a; Powell et al., 1995; Sharma and Reutergardh, 2000; Tuthill,
1996).

4.3.5.5 Hair Lead as a Biomarker of Lead Exposure
Rabinowitz et al. (1976) measured hair Pb levels in two adult males who received a stable
Pb isotope supplement to their dietary intake for 124-185 days. Approximately 1% of the daily
Pb intake was recovered in hair. Temporal relationships between exposure levels and kinetics
and hair Pb levels, and kinetics of deposition and retention of Pb in hair have not been evaluated.
Higher hair Pb levels were observed in Pb workers than in reference subjects with lower blood
Pb levels (Mortada et al., 2001).

4.3.5.6 Summary of Hair Lead as a Biomarker of Lead Body Burden and Exposure
Although hair Pb measurements have been used in some epidemiologic studies, an
empirical basis for interpreting hair Pb measurements in terms of body burden or exposure has
not been firmly established. Hair Pb measurements are subject to error from contamination of
the surface with environmental Pb and contaminants in artificial hair treatments (i.e., dyeing,
bleaching, permanents) and, as such, are relatively poor predictor of blood Pb concentration,
particularly at low levels (<10 to 12 |ig/dL).
4.4 MODELING LEAD EXPOSURE AND TISSUE DISTRIBUTION
OF LEAD
4.4.1 Introduction
Models are essential for quantifying human health risks that derive from exposures to Pb.
Models come in various forms. Multivariate regression models, commonly used in
epidemiology, provide estimates of the contribution of variance in the internal dose metric to
various determinants or control variables (e.g., surface dust Pb concentration, air Pb
concentration). Structural equation modeling links several regression models together to
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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). A strength of regression models is that
they are empirically verified within the domain of observation and have quantitative estimates of
uncertainty imbedded 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. 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 which are expected to
vary spatially and temporally. Thus, extrapolation of regression models to other spatial or
temporal contexts, which is often necessary for regulatory applications of the models, can be
problematic.
An alternative to regression models are 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 more complex than regression models;
this added complexity introduces challenges in terms of their mathematical solution and
empirical verification. However, by incorporating parameters that can be expected to vary
spatially or temporally, or across individuals or populations, mechanistic models can be
extrapolated to a wide range of exposure scenarios, including those that may be outside of the
domain of paired predictor-outcome data used to develop the model. Exposure-intake models, a
type of mechanistic models, are highly simplified mathematical representations of relationships
between levels of Pb in environmental media and human Pb intakes (e.g., jig 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); and they 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-biokinetics models (i.e., integrated
exposure-intake-biokinetics models) provide an approach for predicting blood Pb concentrations
(or Pb concentrations in other tissues) that corresponds to a specified exposure (medium, level,
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and duration). Detailed information on exposure and internal dose can be obtained from
controlled experiments, but almost never from epidemiological observations or from public
health monitoring programs. Exposure intake-biokinetics 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 to applying
epidemiologically-based information on blood Pb-response relationships to the quantification
and characterization of human health risk. They are also critical for assessing the potential
impacts of public health programs directed at mitigation of Pb exposure or of remediation of
contaminated sites.
Models (both regression models and mechanistic models) also have several other
important features that are useful for risk assessment and for improving our basic understanding
of Pb exposures and biokinetics. They organize complex information on Pb exposure and
biokinetics into a form that provides predictions that can be quantitatively compared to
observations. By analyzing the relationships between model assumptions and predictions
(i.e., sensitivity analysis) and by comparing predictions to observations (i.e., model evaluation),
such models can contribute to the identification of important gaps in our understanding of Pb
exposure, biokinetics, and risk. Thus, these models provide a consistent method for making,
evaluating and improving predictions that support risk assessment and risk management
decisions.
Modeling of human Pb exposures and biokinetics has advanced considerably during the
past several decades. Among the most important new advances are development, evaluation, and
extensive application of the Integrated Exposure Uptake Biokinetic (IEUBK) Model for Lead in
Children (U.S. Environmental Protection Agency, 1994a) and the development of models that
simulate Pb biokinetics in humans from birth through adulthood (Leggett, 1993; O'Flaherty
1993, 1995). While these developments represent important conceptual advances, several
challenges remain for further advancements in modeling and applications to risk assessment.
The greatest challenge derives from the complexity of the models. Human exposure-biokinetics
models include large numbers of parameters, which are required to describe the many processes
that contribute to Pb intake, absorption, distribution, and excretion. The large number of
parameters complicates the assessment of confidence in parameter values, many of which cannot
be directly measured. Statistical procedures can be used to evaluate the degree to which model
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outputs conform to "real-world" observations and values of influential parameters can be
statistically estimated to achieve good agreement with observations. Still, large uncertainty can
be expected to remain about many, or even most, parameters in complex exposure-biokinetic
models such as those described below. Such uncertainties need to be identified and their impacts
on model predictions quantified (i.e., through use of sensitivity analysis, probabilistic methods).
Given the difficulty in quantitatively assessing uncertainty in values of all of the
individual parameters in an exposure-biokinetics model, assurance that the model accurately
represents the real-world in all aspects is virtually impossible. As consequence of this, Oreskes
(1998) noted, " ...the goals of scientists working in a regulatory context should be not validation
but evaluation, and where necessary, modification and even rejection. Evaluation implies an
assessment in which both positive and negative results are possible, and where the grounds on
which a model is declared, good enough are clearly articulated. " In this context, evaluation of
confidence in a given exposure-intake or intake-biokinetics model rests largely on assessment of
the degree to which model predictions, based on model inputs appropriate for a situation,
conform to observations and/or expectations; and, most importantly, the degree to which this
conformity does or does not satisfy requirements of model application to a specific context.
Because of limitations in observations of predicted outcomes, it may be possible to evaluate
confidence in some uses of a model, but not others. Similarly, it is possible for confidence in a
model to be judged acceptable for a given use, but not for others. The concept of validation of
highly complex mechanistic models, outside of the context of a specific use of the model, has
little meaning. In this chapter, discussions of specific models include references to sources of
information on quantitative assessments of uncertainty of specific model applications. These
assessments have been limited to assessments of model performance for prediction of impacts of
surface dust Pb exposures on blood Pb concentrations in children.
In the ensuing discussion of specific models, reported efforts to evaluate the models are
noted. In most cases, however, the relevance of these evaluations to the assessment of
confidence in a specific use of that model (e.g., predicting average blood Pb concentrations in
children who live in areas that have certain cross-sectionally measured environmental Pb levels)
cannot be ascertained from the reported literature. Nevertheless, as a framework for qualitatively
comparing the various evaluative procedures that have been applied, the following general
classification of model evaluations has been adopted:
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• Sensitivity analyses have been conducted and most influential parameters identified and
uncertainty characterized.
• Model predictions have been compared qualitatively to observations.
• Predictions have been compared quantitatively to observations (i.e., a statistical model
has been applied for estimation of "goodness of fit" and uncertainty).
• Confidence in model predictions for specific uses has been quantitatively evaluated.
• Accuracy of model implementation code has been verified.
Descriptions of the individual models are intended to provide only brief snapshots of key
features of each model, with particular attention to conceptual features that are unique to each
model. Key references are cited in which more complete specifications of model parameters can
be found.
4.4.2 Slope Factor Models
Empirically-based relationships between blood Pb concentrations and Pb intakes and/or
Pb levels in environmental media have provided the basis for what has become known as slope
factor models. Slope factor models are highly simplified representations of empirically-based
regression models in which the slope parameter represents the change in blood Pb concentration
projected to occur in association with a change in Pb intake or uptake. The slope parameter is
factored by exposure parameters (e.g., exposure concentrations, environmental media intake
rates) that relate exposure to blood Pb concentration (Maddaloni et al., 2005; U.S. Environmental
Protection Agency, 2003c; Abadin and Wheeler, 1997; Stern, 1996; Bowers et al., 1994; Stern,
1994; Carlisle and Wade, 1992). In slope factor models, Pb biokinetics are represented as a
linear function between the blood Pb concentration and either Pb uptake (uptake slope factor,
USF) or Pb intake (intake slope factor, ISF). The models take the general mathematical forms:

PbB = E-ISF (4-4)

PbB = E-AF-USF (4-5)

where PbB is the blood Pb concentration, E is an expression for exposure (e.g., soil intake x soil
Pb concentration) and AF is the absorption fraction for Pb in the specific exposure medium of
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interest. Intake slope factors are based on ingested rather than absorbed Pb and, therefore,
integrate both absorption and biokinetics into a single slope factor, whereas models that utilize
an uptake slope factor include a separate absorption parameter. In contrast to mechanistic
models, slope factor models predict quasi-steady state blood Pb concentrations that correspond to
time-averaged daily Pb intakes (or uptakes) that occur over sufficiently long periods to produce a
quasi-steady state (i.e., >75 days, ~3 times the ti/2 for elimination of Pb in blood).
The U.S. EPA Adult Lead Methodology (ALM) is a example of a slope factor model that
has had extensive regulatory use in the EPA Superfund program (Maddaloni et al., 2005); it is
the model recommended by EPA for assessing health risks to adults associated with non-
residential exposures to Pb in contaminated soils (U.S. Environmental Protection Agency,
1996a). The model was developed to predict maternal and fetal blood Pb concentrations that
might occur in relation to maternal exposures to soils contaminated. The model is implemented
with the following algorithms (see Table 4-5 for explanation of parameters):

PhR PhR PbS-BKSF-IRs-AFs-EFs
Pt>h'adult, central = PbBadult,0 + I4'6)
• R (4-7)
fetal,95thpercentile ~ adultrcentral
PKR PbBfetal.95.goa
rt)r>adult,central,goal ~
fetal,maternal
adult,central.goal "" adult.O ) '
BKSF-IRS-AFS-EFS
,* QN
The ALM partitions the contributions of Pb exposure on blood Pb concentration of adults
(PbBaduit,centrai) into two sources: exposure to non-residential site soil, including outdoor soil and
indoor soil-derived dust and off-site (e.g., residential) exposures to all other media; the latter
contribute to a baseline blood Pb concentration (PbBaduit,o). If the risk-based goal is to ensure
that there is no more than a 5% probability that a fetus will have a blood Pb concentration of
10 ng/dL, then a risk-based goal for the blood Pb concentration in the adult (PbBaduit,centrai!goai) is
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Table 4-5. Recommended Parameter Values for the Adult Lead Methodology (ALM)
and Corresponding Risk-based Soil Lead Concentrations (RBCs)a
Parameter
PbBfetal)95) goal

po-pj 1.645
*J"-L' i.adult

-^Metal/maternal
PbBadult,central

PbBaduit,0

Pbs

BKSF

Value
10

2.1-2.33

0.9
Calculated
value
1.5-1.7b

User
Specified
value
0.4

Unit
ug/dL

—

—
ug/dL

ug/dL

ppm

ug/dL per
ug/day
Comment
Goal for the 95th percentile blood Pb concentration (ug/dL) among
potential fetuses of women exposed to the RBCPbs.
Inter-individual geometric standard deviation of PbBs among women of
child-bearing age that have exposures to similar on-site Pb
concentrations.
Proportionality between fetal PbB at birth and maternal PbB.
Typical PbB concentration (ug/dL) in women of child-bearing age at
the site in the absence of exposures to the site that is being assessed.
Baseline PbB, the typical PbB in women of child-bearing age at the site
in the absence of exposures to the site.
Concentration of Pb in soil (ug/g)

BKSF = biokinetic slope factor; ratio of (quasi-steady state) increase in
typical adult PbB concentration to average daily Pb uptake (ug/dL PbB
IRS

AFS

EFS

AT
RBCS
increase per ug/day Pb uptake)
0.05 g/day Combined intake rate of soil, including both outdoor soil and indoor
soil-derived dust.
0.12 — Absolute gastrointestinal absorption fraction for ingested Pb in soil and
Pb in dust derived from soil.
219 days/year Exposure frequency for contact with assessed soils and/or dust derived
in part from these soils (days of exposure during the averaging period).
365 days/year Averaging time, the total period during which soil contact may occur.
Calculated ppm Risk-based soil Pb concentration (RBC) that would be estimated to
value result in a specified central tendency PbB concentrations in adults (i.e.,
women of child-bearing age) at the site (PbBaduit,centrai,goai) and
corresponding 95th percentile fetal PbB concentration (PbBfetai 0 95,g0ai)-
"Based on U.S. EPA (1996a).
bGSD; adult and PbBadulto are for women of ages 17^5 years as reported in the combined NHANES III Phases 1 and 2
(U.S. EPA, 2002).
given by Equation (4-8), and the corresponding risk-based site soil Pb concentration (RBCs) is

given by Equation (4-9). Figure 4-12 shows the predicted relationship between soil Pb

concentration and 95th percentile fetal blood Pb concentration. The predictions correspond to

values GSDi=2.1 andPbBo=1.5 ug/dL, the central estimates of these parameters for the

U.S. adult population (age range: 17-45 years, U.S. EPA 2002). The soil Pb concentration that

corresponds to 95th percentile=10 |_ig/dL is -1250 ppm.
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TJ
"gj
CQ
.O
•4-1
0>
LL
Q)
a
Q)
a.
10
O)
18
16
14
12
10
8
6
4
2
0
0
500 1000 1500
Soil Lead Concentration (pg/g)
2000
Figure 4-12. Adult lead model (ALM) predictions of the relationship between soil lead
concentration and 95th percentile fetal blood lead concentration. The line
represents predictions for assumed values of GSDi=2.1 and PbBo=1.5 ug/dL.
All other parameter values are from Table 4-5. The soil lead concentration
that corresponds to 95th percentile=10 ug/dL is -1250 ppm.
4.4.3 Empirical Models of Lead Exposure-Blood Lead Relationships
The 1986 Lead AQCD described epidemiological studies that explored models of
relationships between Pb exposures and blood Pb concentrations in children. A more recent
summary of this literature can be found in Abadin et al. (1997). Key studies reported since the
completion of the 1986 Lead AQCD are summarized here. Although varying widely in exposure
scenarios, blood Pb concentration ranges, and modeling approaches, most studies have found
significant associations between surface dust Pb levels (interior and exterior) and blood Pb
concentrations. These outcomes support the general concept that contact with Pb in surface dust
(e.g., surface dust-to-hand-to-mouth) is a major contributor to Pb intake in children.
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One of the largest analyses of relationships between environmental Pb levels and blood Pb
concentrations in children was conducted at the Bunker Hill Superfund site, a former Pb mining
and smelting site. Although not an epidemiological study, per se, as a part of public health
monitoring and remedial investigation studies connected with the site, extensive surveying was
conducted of Pb levels in residential soil and interior dust, and blood Pb concentrations in
children residing at these residences. These data provide a basis for exploring quantitative
relationships between soil and dust exposure variables and blood Pb concentrations, as well as
for evaluating mechanistic models that simulate these relationships (see discussion of calibration
and evaluation of the IEUBK Model for Lead in Children, Section 4.4.5.2).
TerraGraphics (2000) conducted an analysis of data on environmental Pb levels and child
blood Pb concentrations in children, as part of a 5-year review of the clean-up at the Bunker Hill
Superfund site, a former Pb mining and smelting site. The analysis included -4,000 observations
of blood Pb concentrations in children between the ages of 9 months and 9 years of age,
collected over an 11-year period (1988-1999). The number of children for which blood Pb
concentrations were available each year ranged from 230 in 1988 to 445 in 1993; -54 to 88% of
the child population was sampled each year. Blood Pb concentrations (annual geometric mean)
ranged from 4.4 to 9.9 |ig/dL. Environmental Pb levels (e.g., dust, soil, paint Pb levels) data
were collected at -1300 residences. Interior dust Pb concentrations (annual geometric mean)
ranged from -400 to 4200 ppm. Yard soil Pb concentration (annual geometric mean) ranged
from -100 to 2600 ppm. Several multivariate regression models relating environmental Pb
levels and blood Pb concentration were explored; the model having the highest R2 (0.23) is
shown in Table 4-6. The model predicts significant associations between blood Pb
concentration, and the (natural log-transformed) community soil Pb concentration (P = 1.76),
neighborhood soil Pb concentration (P = 0.73; geometric mean soil Pb concentration for areas
within 200 ft of the residence), or interior dust Pb concentration (P = 0.84).
The model predicted a 1.2 |ig/dL decrease in blood Pb concentration in association with a
decrease in community soil Pb concentration from 2000 to 1000 ppm. The same decrease in
neighborhood soil Pb concentration, or interior dust Pb concentration, was predicted to result in a
0.5 or 0.6 |ig/dL decrease in blood Pb concentration, respectively. Regression models (R2 = 0.86
to 0.94), based on repeated blood Pb measurements made on the same children from this data set,
predicted much stronger associations between current blood Pb concentration and the blood Pb
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Table 4-6. General Linear Model Relating Blood Lead Concentration in Children and
Environmental Lead Levels—Bunker Hill Superfund Site
Parameter
Intercept
Age (years)
/^(interior dust Pb) (ppm)
/w(yard soil Pb) (ppm)
/w(GM soil Pb within 200 ft of residence)
(ppm)
/w(GM community soil Pb) (ppm)
Coefficient
-0.22877
-0.44803
0.83723
0.21461
0.73100
1.76000
P-value
0.7947
0.0001
0.0001
0.0080
0.0001
0.0001
Standardized
Coefficient
0.00000
-0.25541
0.15677
0.06466
0.12938
0.19709
R2 = 0.231; p < 0.0001; based on data from Bunker Hill Superfund Site collected over the period 1988-1999
GM, geometric mean; In, natural log
Source: TerraGraphics (2000).
concentration measured in the previous year (P = 0.62) for the same child than to the
corresponding community soil Pb concentration (P = 0.095) or interior dust Pb concentration
(P = 0.1). Structural equation modeling was applied to the larger data set, utilizing the model
structures shown in Figure 4-13. Model 1 included a direct pathway connecting community soil
Pb to blood Pb. Both models yielded similar R2 values (0.89) and predicted a relatively large
influence of interior dust Pb on blood Pb (Tables 4-7 and 4-8).
A subsequent analysis was conducted of paired environmental Pb and blood Pb levels in
children (n = 126, ages 9 months to 9 years), collected during 1996 to 1998 at various locations
in the Coeur d'Alene Basin (outside of the Bunker Hill Site, TerraGraphics, 2001). The annual
geometric mean of blood Pb concentrations for the study area was ~4 |ig/dL, with the blood Pb
concentration range of individuals included in the regression analysis being ~1 to 23 |ig/dL.
Yard soil Pb concentrations ranged from <100 to 7350 ppm. A model that included all
significant (p < 0.05) variables is shown in Table 4-9. The model predicted a 0.7 |ig/dL decrease
in blood Pb concentration per 1000 ppm decrease in exterior soil Pb, and a 0.16 |ig/dL decrease
in blood Pb per 1 mg/cm2/day decrease in entryway dust Pb loading rate. Entryway dust Pb
loading rate was estimated from measurements of the amount of Pb (and dust) recovered from
doormats placed at each residence for a known duration. Regression models (general linear
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Figure 4-13. Structural equation model for relationships between dust and soil lead and
blood lead concentration in children, based on data collected at the Bunker
Hill Superfund Site (1988-1999). Neighborhood soil lead is represented in
the model as the mean soil lead within 200 feet of the residence, whereas
community soil lead is the mean for the city. The pathway between
community soil lead and blood lead was included in Model 1 and excluded
from Model 2. Units: blood lead, ug/dL; dust and soil lead (ug/g); age,
years. See Tables 4-7 and 4-8 for estimated regression coefficients.

Source: TerraGraphics et al. (2000).
model) relating entryway mat Pb loading rate, or mat Pb concentration, and environmental

variables were also developed. The strongest predictor of both outcome variables (natural log-

transformed) was soil Pb concentration (natural log-transformed, P: -0.4; R2 = 0.36-0.46).

Lanphear et al. (1998) conducted a pooled analysis of data on environmental Pb levels and

blood Pb concentrations in children (n = 1861) collected as part of 12 epidemiologic studies

(conducted over a 15-year period, from 1982 to 1997). Seven of the studies were of

communities near Pb mining and/or smelting sites (Bingham Creek, UT; Butte, MT; Leadville,

CO; Magna, UT; Midvale, UT; Palmerton, PA; Sandy, UT); and 5 studies were of urban
communities (2 in Boston, MA; 2 in Cincinnati, OH; Rochester, NY). The mean age of children

included in the analysis was 16 months; the inter-study range was 6 to 24 months. The

geometric mean blood Pb concentration for the subjects in the pooled analysis was 5.1 |ig/dL;
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Table 4-7. Structural Equation Model (1) Relating Blood Lead Concentration in
Children and Environmental Lead Levels—Bunker Hill Superfund Site
Parameter
Coefficient
Standardized
P-value Coefficient
Contribution
(%)'
Model for In(bloodPb) fag/dL): R2 = 0.892
Error
Intercept
Age (years)
/«(interior dust Pb) (ppm)
/n(yard soil Pb) (ppm)
/«(AM soil Pb within 200 ft of
residence) (ppm)
/n(AM community soil Pb) (ppm)
Model for
Error
Intercept
/w(yard soil Pb) (ppm)
/w(AM soil Pb within 200 ft of
residence) (ppm)
/w(AM community soil Pb) (ppm)
1.000
-0.519
-0.065
0.159
0.051
0.067
0.095
In (Interior dust Pb)
1.000
3.237
0.129
0.133
0.235
—
0.05
0.05
0.05
0.05
0.05
0.05
(ppm): R2 = 0.986
—
0.05
0.05
0.05
0.05
0.329
-0.171
-0.210
0.597
0.171
0.267
0.389

0.117
0.487
0.114
0.141
0.256
—
—
—
42
12
19
27

—
—
22
28
50
Based on data from the Bunker Hill Superfund Site collected over the period 1988 to 1999.
Largest standardized residual, 0.183; Chi-Square, 21.309; P, 0.0001; Comparative fit index, 0.9993;
Normed fit index, 0.9993; Non-normed fit index, 0.9863.
GM, geometric mean; In, natural log
aBased on sum of standardized coefficients for dust and soil Pb parameters.
Source: TerraGraphics (2000).
95% were within the range 1.2 to 26 |ig/dL and 19% were >10 ng/dL. The geometric mean
interior dust Pb loading was 13.5 |ig/ft2 (95% range: 1 to 4500 |ig/ft2) and geometric mean
exterior soil or surface dust Pb level was 508 ppm (95% range: 8 to 10,200 ppm). A regression
model was developed relating natural log-transformed blood Pb concentration to log-transformed
environmental Pb variables, and categorical demographic or behavioral variables (Table 4-10).
The R2 for the final model was 0.53 (uncorrected for measurement error). Measurement error
was included in the model as variance estimates for each environmental Pb variable as follows
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Table 4-8. Structural Equation Model (2) Relating Blood Lead Concentration in
Children and Environmental Lead Levels—Bunker Hill Superfund Site
Parameter

Error
Intercept
Age (years)
/^(interior dust Pb) (ppm)
/w(yard soil Pb) (ppm)
/w(AM soil Pb within 200 ft of
residence) (ppm)
Coefficient
Model for In(bloodPb)
1.000
-0.206
-0.064
0.165
0.051
0.115
Standardized
P-value Coefficient
(jug/dL): R2 = 0.892
—
0.05
0.05
0.05
0.05
0.05

0.329
-0.116
-0.208
0.619
0.171
0.456
Contribution
(%)'

—
—
—
50
14
37
Model for In (interior dust Pb) (ppm): R2 = 0.986
Error
Intercept
/w(yard soil Pb) (ppm)
/w(AM soil Pb within 200 ft of
residence) (ppm)
/w(AM community soil Pb) (ppm)
1.000
3.237
0.129
0.133
0.235
—
0.05
0.05
0.05
0.05
0.117
0.487
0.114
0.141
0.256
—
—
22
28
50
Based on data from the Bunker Hill Superfund Site collected over the period 1988 to 1999.
Largest standardized residual, 0.183; Chi-Square, 21.309; P, 0.0001; Comparative fit index, 0.9993;
Normed fit index, 0.9993; Non-normed fit index, 0.9863.
AM, arithmetic mean; In, natural log.
aBased on sum of standardized coefficients for dust and soil Pb parameters.

Source: TerraGraphics (2000).
(log-transformed values): dustPb loading, 1.00; exterior Pb concentration, 1.00; water Pb

concentration, 0.75; maximum XRF, 0.75. Of the model variables listed above, significant

variables (p < 0.05, after correction for measurement error) were as follows: interior dust Pb

loading (P = 0.183, p < 0.0001), exterior soil/dust Pb (P = 0.02116, p = 0.00025), age

(P = 0.02126, p = 0.0044), mouthing behavior (P = -0.0323, p = 0.0004), and race (P = 0.123,

p = 0.0079). Significant interactions in the model included: age and dust Pb loading, mouthing

behavior and exterior soil/dust level, and SES and water Pb level. Predicted relationships

between interior dust Pb loading or exterior Pb concentrations and blood Pb concentration are

shown in Tables 4-11 and 4-12. The model predicted a geometric mean blood Pb concentration
4-77
-------
Table 4-9. General Linear Model Relating Blood Lead Concentration in Children and
Environmental Lead Levels—Coeur d'Alene Basin
Parameter
Intercept
Age (years)
Soil Pb (ppm)
Entryway (mat) Pb loading rate
(mg/cm2/day)
Median exterior paint Pb (mg/cm2)
Minimum interior paint condition
(categorical: 1-3)
Coefficient
2.8644
-0.3351
0.0007
0.1638
0.5176
1.9230
P-value
0.0032
0.0007
0.0012
0.0006
0.0005
0.0008
Standardized
Coefficient
0.00000
-0.2056
0.2249
0.3212
0.2742
0.2313
N: 126 (ages 9 mo to 9 years), R = 0.597, p < 0.0001; based on data from the Coeur d'Alene Basin collected
over the period 1996 to 1999.
Source: TerraGraphics (2001).
of 4.0 |ig/dL (4% probability of exceeding 10 |ig/dL) assuming the study median environmental
Pb levels to be as follows: dust Pb, 5.0 |ig/ft2; soil Pb, 72 ppm; maximum interior paint Pb,
1.6 mg/cm2; water Pb, 1 ppb.
Succop et al. (1998) conducted a meta-analysis of relationships between environmental
Pb levels and blood Pb levels in children (n = 1855, age <72 months) based on data from
11 epidemiologic studies (conducted over a 13-year period, 1981 to 1994). All but 2 of the
studies (Cincinnati prospective study, 1981 to 1985; Cincinnati soil Pb study, 1989 to 1991) were
of communities near Pb mining and/or smelting sites (Bingham Creek, UT; Butte, MT;
Leadville, CO; Magna, UT; Midvale, UT; Palmerton, PA; Sandy, UT; Telluride, CO; Trail,
B.C.). The inter-study age range was 15 to 39 months, and the inter-study range of the geometric
mean blood Pb concentration was 2.6 to 12.9 |ig/dL; 7.5% of children were > 10 |ig/dL. The
9
inter-study geometric mean ranges were: interior dust Pb loading, 31 to 976 |ig/m ; interior dust
Pb concentration, 110 to 1548 ppm; handwipe Pb, 2 to 9 jig; exterior entry dust Pb concentration,
72 to 1830 ppm. Structural equation modeling was applied to the data from each study.
The same generic model was initially applied to each dataset, followed by backward
elimination of pathways and co-variables until a model for each study evolved in which all
predictors and co-variables were significant (p < 0.05). The generic model is shown in
4-78
-------
Table 4-10. Multivariate Regression Model Relating Blood Lead Concentration in
Children and Environmental Lead Levels—Multi-study Pooled Analysis
Parameter
Intercept
Dust Pb loading (ug/ft2)
Water Pb (ppb)
Soil or exterior dust Pb (ppm)
Soil or exterior exposure dust Pb * type of
sample
Soil or exterior exposure dust Pb * type of
sample * location
Type of exterior exposure sample
Soil or exterior exposure dust location
Paint Pb content (mg/cm3)
CLN(MAX XRF) * paint condition
Paint condition
Age
Age 2
Age 3
Study

Race

Level

Boston
Butte
Bingham Creek
Cincinnati Program
Cincinnati Soil
Leadville
Magna
Rochester
Longitudinal
Rochester LID Study
Sandy
Midvale
Palmerton
Other
White
Estimate
1.496
0.183
0.01398
0.02116
0.005787
0.4802
-0.1336
0.5858
-0.02199
0.03811
-0.0808
0.02126
-0.001399
0.00007854
-0.3932
-0.01167
0.2027
0.2392
0.5383
0.05717
0.1761
-0.04209
0.07257
-0.3712
0.1777
0
0.123
0
P-value

0.0001
0.2067
0.0025
0.9247
0.0409
0.2805
0.0455
0.3402
0.3888
0.1685
O.0001
0.0044
0.0022
o.ooor

0.00793

4-79
-------
Table 4-10. (cont'd). Multivariate Regression Model Relating Blood Lead Concentration
in Children and Environmental Lead Levels - Multi-study Pooled Analysis
Parameter
Socioeconomic status (SES)

Mouthing behavior

Dust Pb loading * Age
Dust Pb loading * Age 2
Dust Pb loading * Age 3
Exterior Pb exposure * mouthing behavior

Water Pb levels (ppb) * SES

Age * race

Age * SES

Standard deviation of the prediction error
Level
1
2
3
4
5
Often
Rarely
Sometimes
Unknown

Often
Rarely
Sometimes
Unknown
1
2
3
4
5
Other
White
1
2
3
4
5

Estimate
0.3175
0.2138
0.1799
0.1691
0
-0.03233
-0.2454
-0.1397
0
0.002649
-0.0003381
-0.00001281
0.2212
0.07892
0.1663
0
0.5305
-0.0136
0.1033
-0.09098
0
0.01192
0
-0.01023
0.003849
0.00008468
-0.01679
0
0.5425
P-value
o.iosr

0.00043

0.1860
0.0573
0.6185
0.04193

0.09983

0.01293

0.00613

Interactions are indicated by asterisks. Blood Pb concentration (ug/dL) and all environmental Pb variables were
natural log-transformed. R2 for blood Pb concentration was 0.53 (uncorrected for measurement error).

"Overall factor significance.

Source: Lanphearetal. (1998).
4-80
-------
Table 4-11. Children's Predicted Blood Lead Levels for Floor Dust Lead Loading
(ug/ft2) and Exterior Lead Exposures (ppm)a
Geometric mean blood Pb levels (ug/dL) with 90% Confidence Intervals3 in parentheses
Dust Pb
loading
(Jig/ft2)
1
5
10
15
20
25
40
55
70
100
Exterior Pb exposure (ppm)
10
2.3
(0.9, 5.7)
3.2
(1.3,8.0)
3.7
(1.5,9.2)
4.0
(1.6, 10.0)
4.2
(1.7, 10.6)
4.4
(1.8,11.2)
4.9
(1.9,12.3)
5.2
(2.1, 13.2)
5.5
(2.2, 13.8)
5.9
(2.3, 14.9)
72"
2.8
(1.1,7.0)
4.0
(1.6,9.8)
4.6
(1.8,11.3)
5.0
(2.0, 12.3)
5.3
(2.1,13.0)
5.5
(2.2, 13.6)
6.1
(2.4, 15.0)
6.5
(2.6, 16.1)
6.8
(2.7, 16.9)
7.3
(2.9, 18.2)

(1
(1.
(1.
(2.
(2.
(2.
(2.
(2.
(2.
(3.
100
2.9
.2,7.3)
4.1
7,10.1)
4.7
9,11.7)
5.1
1, 12.7)
5.4
2,13.5)
5.7
3,14.1)
6.3
5, 15.6)
6.7
7, 16.6)
7.0
8, 17.5)
7.6
1,18.9)
500
3.5
(1.4,8.7)
4.9
(2.0, 12.0)
5.6
(2.3, 13.9)
6.1
(2.5,15.1)
6.5
(2.6, 16.0)
6.8
(2.8, 16.8)
7.5
(3.0, 18.5)
8.0
(3.2, 19.7)
8.4
(3.4,20.7)
9.0
(3.7,22.3)

(1
(2.
(2.
(2.
(2.
(3.
(3.
(3.
(3.
(3.
1000
3.8
.5,9.4)
5.3
1,13.0)
6.1
5, 15.0)
6.6
7, 16.3)
7.0
8, 17.3)
7.3
0,18.1)
8.1
3, 19.9)
8.6
5,21.3)
9.1
7,22.3)
9.7
9,24.1)
1500
4.0
(1.6,9.8)
5.5
(2.2, 13.6)
6.3
(2.6, 15.7)
6.9
(2.8, 17.0)
7.3
(3.0, 18.0)
7.7
(3.1,18.9)
8.4
(3.4,20.8)
9.0
(3.7,22.2)
9.5
(3.8,23.4)
10.2
(4.1,25.2)

(1.
(2.
(2.
(2.
(3.
(3.
(3.
(3.
(4.
(4.
2000
4.1
,6,10.1)
5.7
,3, 14.0)
6.5
,7, 16.2)
7.1
,9, 17.6)
7.6
1,18.6)
7.9
,2, 19.5)
8.7
,5,21.5)
9.3
,8,22.9)
9.8
,0,24.1)
10.5
,3,26.0)
4000
4.4
(1.8,11.0)
6.1
(2.5, 15.2)
7.1
(2.9, 17.5)
7.7
(3.1,19.0)
8.1
(3.3,20.1)
8.5
(3.5,21.1)
9.4
(3.8,23.2)
10.0
(4.1,24.8)
10.5
(4.3,26.0)
11.3
(4.6,28.0)
Confidence interval is estimated to cover 90% of the observed blood Pb levels with 5% above and 5% below the interval.
bValues for an exterior Pb paint exposure of 72 ppm were estimated median levels based on U.S. Housing and Urban
Development national survey, 1989-1990
Source: Lanphear et al. (1998).
Figure 4-14, along with the percent of studies in which a given pathway was found to be
significant. The most common exposure pathway influencing blood Pb concentration
(i.e., significant in models of most studies) was exterior soil, operating through its effect on
interior dust Pb and hand Pb. Paint Pb was also a significant influential variable on the soil and
interior dust-to-blood pathway in -40% of the studies. Significant co-variables varied across
studies and included: child age, mouthing frequency, time spent outdoors, SES, house age and
condition, home renovation, parental occupation, bare soil in yard, and presence of pets.
The relative strength of the influence of various environmental sources of Pb in the structural
4-81
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Table 4-12. Likelihood of a Child's Blood Lead >10 ug/dL for Floor Dust Lead Loadings
and Exterior Exposure Levels (ppm)a
Probability of blood Pb greater than 10
Dust Pb
loading
(Jig/ft2)
1
5
10
15
20
25
40
55
70
100
ug/dL

Exterior Pb exposure (ppm)
10
0.33%
(0.05,2.24)
1.8%
(0.4, 7.9)
3.3%
(0.8, 12.6)
4.5%
(1.2,16.2)
5.7%
(1.5,19.2)
6.7%
(1.8,21.8)
9.4%
(2.7, 27.8)
12%
(3, 32)
13%
(4, 36)
17%
(5,41)
72"
1.0%
(0.3,3.8)
4.4%
(1.7,11.0)
7.4%
(3.1,16.5)
9.8%
(4.3,20.7)
12%
(5, 24)
14%
(6,27)
18%
(9, 33)
21%
(10,38)
24%
(12,42)
28%
(14,48)
100
1.2%
(0.3,4.2)
5.0%
(2.0,11.8)
8.3%
(3.8, 17.5)
11%
(5,22)
13%
(6,25)
15%
(7,28)
20%
(10,35)
23%
(12,40)
26%
(14,44)
31%
(16,49)
500
2.7%
(0.9, 7.4)
9.3%
(4.7, 17.6)
14%
(8,24)
18%
(11,29)
21%
(13,33)
24%
(15,36)
30%
(19,43)
34%
(22,48)
37%
(24, 52)
43%
(28, 58)
1000
3.7%
(1.3,9.7)
12%
(6,21)
18%
(10,29)
22%
(14, 34)
26%
(16,38)
28%
(18,41)
35%
(23,48)
39%
(27, 53)
43%
(29, 57)
48%
(34, 63)
1500
4.4%
(1.6,
11.5)
14%
(7,24)
20%
(12, 32)
25%
(15,37)
28%
(18,41)
31%
(20, 45)
38%
(25, 52)
42%
(29, 57)
46%
(32, 60)
51%
(37, 66)
2000
4.9%
(1.8,
12.8)
15%
(8,26)
22%
(13,35)
27%
(16,40)
30%
(19,44)
33%
(22, 47)
40%
(27, 54)
45%
(31,59)
48%
(34, 63)
54%
(39, 68)
4000
6.5%
(2.3,
16.9)
18%
(9, 32)
26%
(15,41)
31%
(19,47)
35%
(22,51)
38%
(25, 54)
45%
(31,61)
50%
(35,65)
54%
(38, 69)
59%
(43, 73)
aAll other variables held at their national median.
bEstimated median levels based on U.S. Housing and Urban Development national survey, 1989 to 1990.
Source: Lanphear et al. (1998).
equation model, on blood Pb concentration, was evaluated by applying a simple linear regression
model to the geometric mean values for environmental variables (natural log-transformed) and
blood Pb concentrations (natural log-transformed), from the individual studies (Table 4-13).
The strongest relationships were obtained for interior dust Pb loading ([P = 0.474, R2: 0.96] and
handwipe Pb [P = 1.184, R2 = 0.90]). The models predicted a 8.6 jig/dL decline in blood Pb
concentration (from -15 |ig/dL) for a 1000 jig/cm2 reduction in interior dust Pb loading (from
1100 jig/cm2), and a 14.4 |ig/dL decline in blood Pb concentration (from 10 jig) for a 10 jig
reduction in handwipe Pb.
4-82
-------
Figure 4-14. Structural equation model for relationships between dust and soil lead
and blood lead concentration in children. Numbers are the percentage of
11 studies included meta-analysis for which the pathway was significant
(p = 0.05). Units: blood lead, ug/dL; dust and soil lead, ug/g; handwipe
lead (jig); pant lead, mg/cm2
Source: Succopetal. (1998).
Lanphear and Roghmann (1997) collected data on blood Pb concentrations for 205
children residing in Rochester, NY (1991-1992) paired with their residential environmental Pb
levels. The mean age of the children was 20 months (range: 12 to 30 months). Mean blood Pb
concentration was 7.7 |ig/dL (SD: 5.1), with 23% of children having a blood Pb > 10 |ig/dL.
Geometric mean interior dust Pb loading was 106 |ig/ft2 (+SD: 10, 1167) and soil Pb level was
981 ppm (+SD: 225, 4267). Data on the following variables were used for structural equations
modeling: serum ferritin (ng/dL), blood Pb concentration (|ig/dL), hand Pb (jig), interior dust Pb
loading (jig/ft2), paint Pb loading (XRF, mg/cm2), water Pb (ppb), soil Pb (ppm), race, parent
marital status, household income, maternal cleaning behaviors, and child exposure behaviors
(e.g., time spent outside, mouthing, dirt ingestion). A structural equation model, shown in Figure
4-15, yielded an R2 of 0.41 for blood Pb concentration. The exposure pathway most influential
on blood Pb was interior dust Pb loading, directly or through its influence on hand Pb. Both soil
and paint Pb influenced interior dust Pb; with the influence of paint Pb greater than that of soil
4-83
-------
Table 4-13. Meta-analysis of the Relationship Between Log-transformed Blood Lead and
Various Environmental Lead Sources3
Independent variable
In(handwipe Pb)
In(interior dust Pb loading)
In(interior dust Pb loading)0
In(interior dust Pb
concentration)
In(exterior entry dust Pb
concentration)
In(perimeter soil Pb
concentration)
In(maximum interior paint
Pb loading)
In(maximum exterior paint
Pb loading)
Units
ug
ug/m2
ug/m2
ppm
ppm
ppm
mg/cm2
mg/cm2
Intercept
0.009
-0.479
-0.782
-1.502
-1.101
-0.015
1.562
1.502
Slope
Estimate
1.184
0.444
0.474
0.529
0.435
0.233
0.232
0.152
Squared
Correlation
0.90
0.55
0.96
0.58
0.72
0.65
0.07
0.07
No. of
Studies
6
10
8
10
10
6
8
9
Predicted
Decline in
Blood Leadb
14.4
9.1
8.6
6.5
4.5
2.2
2.1
1.3
aThese are simple relationships unadjusted for covariates.
Predicted decline in blood Pb for a reduction in hand Pb of 10-1 ug; dust Pb loading of 1100 to 100 ug/m2;
dust Pb or soil Pb concentration of 1100-100 ppm; or paint Pb loading of 3.0-0.5 mg/cm2 as calculated
from the fitted linear regression equation: In(blood Pb) = intercept + slope x In(environmental Pb).
Excluding the Trail and Cincinnati soil project studies, which appear to be outliers. The exposure in these two
studies appears to be primarily from exterior dust Pb.
Source: Succopetal. (1998).
Pb. Other influential variables were Black race (direct), family income (direct), and outside play
(indirect) through dirt ingestion behavior. Simple correlation analysis also revealed relatively
strong (significant, p < 0.5) associations between dust Pb loading (r = 0.41), soil Pb
concentration (0.31) and Black race (r = 0.44).
Bornschein et al. (1985a) applied structural equation modeling to paired environmental Pb
and blood Pb data collected on a subset of children (n = 45) from the Cincinnati Prospective
Study (1981-1985). The age range of children included in the study was 9 to 24 months. Group
statistics for the blood Pb concentrations of children included in the analysis were not reported in
Bornschein et al. (1985a); however, a subsequent analysis of data from the study reported a
geometric mean of 12.9 |ig/dL (n = 149; Succop et al. 1998). Similarly, dust and soil Pb levels
4-84
-------
Interior
Paint Lead
0.414
Interior
Dust Lead
Black Race
Income
Soil
Lead
0.229
0.178
0.229
Handwipe
Dust Lead
0.221
0.319
-0.143
Plays
Outside
0.154
Blood
Lead
0.186
0.152
Eats Soil
Figure 4-15. Structural equation model for relationships between dust and soil lead
and blood lead concentration in children, based on data collected in the
Rochester (NY) Lead in Dust Study. Numbers are model coefficients. Units:
blood lead, ug/dL; dust lead ug/ft2; soil lead, ug/g; handwipe lead (ug); pant
lead, mg/cm2, plays outside, categorical: 0-1; eats soil, categorical: 0-1. R2
values: blood lead, 0.41; hand lead Pb, 0.14; dust lead, 0.25.
Source: Lanphear and Roghmann (1997).
were not reported in Bornschein et al. (1985a), but were reported in Succop et al. (1998) for a
larger study group (n = 149) as follows (geometric means): interior dust Pb loading, 976 |ig/m2;
interior dust Pb concentration, 1548 ppm; handwipe Pb, 7 jig; and exterior entry dust Pb
concentration, 1830 ppm. A structural equation model (Bornschein et al., 1985a), shown in
Figure 4-16, yielded R2 values that ranged from 0.44 to 0.59 across age groups from
9 (R2 = 0.59) to 24 months (R2 = 0.44). The exposure pathway most influential on blood Pb was
interior dust Pb concentration, directly or through its influence on hand Pb (exterior soil Pb
concentration and internal paint Pb was excluded from the model, as was race). Blood Pb
concentration was also influenced directly by SES. Interior dust Pb loading was influenced by
housing condition variables. Hand dust was directly influenced by maternal involvement with
4-85
-------
-1.07
Figure 4-16. Structural equation model for relationships between dust and soil lead and
blood lead concentration in children, based on data collected in the
Cincinnati (OH) Prospective Child Study. Numbers are model coefficients.
Units: blood lead, ug/dL; dust and soil lead, ug/g; handwipe lead (jug);
pant lead, mg/cm2; maternal involvement, categorical: 0-6; responsivity of
mother, categorical: 0-11; variety in daily stimulation, categorical: 0-5;
housing characteristics, categorical: 0-1. R2 values: blood lead, 0.41;
hand lead, 0.14; dust lead, 0.25.

Source: Bornschein et al. (1985a).
the child. Based on the above model, the relationship between blood Pb concentration, interior
dust Pb, and hand Pb, at 18 months of age, was as follows:
PbB = 1.94 - 0.02(SES) + 0.l5(PbD) + 0.l5(PbH)
(4-10)
PbH= 0.52 - Q.36(material involvement) + 0.50(PbD)
(4-11)
where PbB, PbD, PbH are the natural log-transformed blood Pb concentration (jig/dL), dust Pb

concentration (ppm), and hand Pb (jig), respectively. The above relationship predicts a decline

in blood Pb concentration ranging from 8 |ig/dL (maternal involvement score, 6) to 11 |ig/dL

(maternal involvement score, 0), for a reduction in interior dust Pb concentration from 1100 ppm
4-86
-------
to 100 ppm (assuming SES score of 17, based on geometric mean reported for the Cincinnati
child study in Succop et al. 1998).
The Urban Soil Lead Abatement Demonstration Project (USLADP) was a study
conducted to determine if urban soil Pb abatement would affect the Pb exposures and blood Pb
concentrations of urban children (U.S. Environmental Protection Agency, 1996b). The study
included measurement of blood Pb concentrations and environmental Pb prior to and following
removal of Pb-contaminated soils and surface dusts from selected urban neighborhoods in
Baltimore (Farrell, 1988), Boston (Aschengrau et al., 1994; Weitzman et al., 1993) and
Cincinnati (Clark et al., 1988, 1991, 1996). The numbers of children included in each study
were -182 in the Baltimore study, 92 in the Boston study, and 169 in the Cincinnati study.
Pre-abatement blood Pb concentrations (geometric mean) were -11 |ig/dL in the Baltimore
study, 12 |ig/dL in the Boston study, and 10 |ig/dL in the Cincinnati study. Pre-abatement soil
and interior floor dust Pb concentrations (geometric mean), respectively, were -420 and
1700 ppm in the Baltimore study; 2300 and 2200 ppm in the Boston study; and 400 and 300 ppm
in the Cincinnati study. Measurements of paired environmental Pb levels and blood Pb
concentrations provided the basis for the development of regression models relating blood Pb
concentration to Pb levels in interior dust and exterior soil. An extensive analysis of the data
collected in each study is reported in U.S. Environmental Protection Agency (1996b), from
which selected examples are provided here.
Structural equation modeling was applied to the data from the Boston and Cincinnati
studies; the generic model applied to these data is shown in Figure 4-17 and parameters for
selected models (based on cross-sectional data) are presented in Table 4-14. The model based on
the Cincinnati data showed a stronger association between interior dust Pb and blood Pb
concentration, compared to the model based on the data from the Boston study. Soil Pb level
influenced blood Pb concentration directly and secondarily, through its influence on interior dust.
Repeated measure models and longitudinal structural equation models were also developed
based on these data, and are described in detail in U.S. Environmental Protection Agency
(1996b).
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Window
Dust Lead
0.0656
Soil Lead
0.0756
0.046C
0.22C
Floor
Dust Lead
0.14B
4.10C
Blood
Lead
0.16B
0.28C
Figure 4-17. Structural equation model for relationships between dust and soil lead and
blood lead concentration in children, based on pre-abatement cross-sectional
data collected in the Urban Soil Lead Abatement Demonstration Project.
Numbers are model coefficients for the Boston study (B) or Cincinnati study
(C). Units: blood, ug/dL; dust, soil lead, ug/g; blood coefficient, ug/dL lead
in blood per 1000 ug/g lead in soil.
Source: U.S. Environmental Protection Agency (1996b).
4.4.4 Historic Overview of Mechanistic Models of Lead Biokinetics
4.4A.I Rabinowitz Model
Early Pb modeling applications presented Pb biokinetics in classical pharmacokinetics
terms. Compartments represented kinetically homogeneous pools of Pb which might be
associated with individual organs or groups of organs. Among the first of such models was one
proposed by Rabinowitz et al. (1976) based on a study of the kinetics of ingested stable Pb
isotope tracers and Pb mass balance data in five healthy adult males (Figure 4-18). The
Rabinowitz model has three compartments: (1) a central compartment representing blood and
other tissues and spaces in rapid equilibrium with blood (e.g., interstitial fluid); (2) a shallow
tissue compartment, representing soft tissues and rapidly exchanging pools within the skeleton;
and (3) a deep tissue compartment, representing, primarily, slowly exchanging pools of Pb
within bone. Excretion pathways include urinary (from the central compartment) and bile,
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Table 4-14. Structural Equation Models Relating Blood Lead Concentration in
Children and Pre-abatement Environmental Lead Levels—Lead in Urban Soil
Abatement Demonstration Project
Boston Cincinnati
Parameter Study Study
Model for blood Pb fag/dL)
Intercept 10.97a 7.55a
Floor dust Pb (ppm) 0.14 4.1 Od
SoilPb(ppm) 0.16 0.28
Model for floor dust Pb (jug/g)
Intercept 1008a 99.9b
Soil Pb (ppm) 0.075 0.2247C
Window dust Pb (ppm) 0.065 la 0.0458C

N = 126 (ages 9 mo to 9 years), R2 = 0.597, P < 0.0001; based on data from the Urban Soil Lead Abatement
Demonstration Project. Blood coefficients are expressed as ug/dL per ug/g; floor dust Pb coefficients are
expressed as ug/g per ug/g.
aP = <0.0001
bP = 0.0002-0.0019
CP = 0.002-0.0099
dP = 0.01-0.0499
Source: U.S. Environmental Protection Agency (1986).
sweat, hair, and nails (from the shallow tissue compartment). The model predicts pseudo-first
order half-times for Pb of-25, 28, and 7000 days in the central, shallow tissue, and deep
compartments, respectively (these values were calculated based on reported residence times, the
reciprocal of the sum of the individual elimination rate constants). The slow kinetics of the deep
tissue compartment led to the prediction that it would contain most of the Pb burden following
chronic exposures (e.g., for years), consistent with Pb measurements made in human autopsy
samples (Barry, 1975; Gross et al., 1975; Schroeder and Tipton, 1968). Note that this model did
not simulate the distribution of Pb within blood (e.g., erythrocytes and plasma), nor did it
simulate subcompartments within bone or physiological processes of bone turnover that might
affect kinetics in the deep tissue compartment.
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DIET + AIR

3
DEEP TISSUE
(BONE)
=200 rng
Tla=700Qdays

=7 pg/day
==7 pg/day
i
=48 pg/day
1
BLOOD
1.9mg
Tic=25 days
i
UR
=36 |j
NE
g/day

=15 pg/day
=2 |j g/day
SHALLOV
(SOFT T
=0.6
T1C=2J
1
BILE,
SWEAT,
=12 pg/
>
if TISSUE
"ISSUE)
mg
3 days
r
HAIR,
NAILS,.,
day
Figure 4-18. Lead biokinetics based on Rabinowitz et al. (1976). Half-times are based on
reported mean residence times for compartments 1, 2, and 3: 36, 40, and
104 days, respectively (half-time = 0.693*residence time).
4.4.4.2 Marcus Model(s)
Marcus (1985a) reanalyzed the data from stable isotope tracer studies of Rabinowitz et al.
(1976) and derived an expanded multicompartment kinetic model for Pb (Figure 4-19).
The model included separate compartments with different Pb turnover rates for cortical (slow,
tl/2 = 1.2 x 104 to 3.5 x 104 days) and trabecular (fast, tl/2 = 100 to 700 days) bone, an approach
subsequently adopted in several other models (O'Flaherty, 1995; U.S. Environmental Protection
Agency, 1994a,b; Leggett, 1993; O'Flaherty, 1993; Bert et al., 1989). A more complex
representation of the Pb disposition in bone included explicit simulation of Pb diffusion within
the bone volume of the osteon and exchange with blood at the canaliculus (Marcus, 1985b;
Figure 4-20). Lead diffusion in bone was based on Pb kinetics data from studies conducted in
dogs. A similar approach to simulating radial diffusion of Pb in bone, expanded to include eight
concentric diffusion shells, was implemented by O'Flaherty (1993, 1995). Marcus (1985c) also
introduced nonlinear kinetics of exchange of Pb between plasma and erythrocytes. The blood
kinetics included four blood subcompartments: diffusible Pb in plasma, protein-bound Pb in
plasma, a "shallow" erythrocyte pool, and a "deep" erythrocyte pool (see Figure 4-21).
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Figure 4-19. Lead biokinectics based on Marcus (1985a). Bone is represented as a slow
turnover (cortical) compartment and a faster (trabecular) compartment.
The Marcus (1985c) model predicted the curvilinear relationship between plasma and blood Pb
concentrations that has been observed in humans (DeSilva, 1981).

4.4.4.3 Bert Model
Bert et al. (1989) adopted the bone model from Marcus (1985a), in which the bone
compartment is subdivided into slow cortical bone and faster trabecular bone compartments
(Figure 4-22). The central compartment (denoted as blood) is assumed to be 1.5 times the
volume of whole blood (Rabinowitz et al., 1976), with the whole blood volume varying in direct
proportion with body weight. The model includes a discrete pathway for excretion of
unabsorbed Pb from the GI tract into feces. Secretion of Pb in bile, gastric secretions, and saliva
are represented as transfers from the soft tissue compartment to the GI tract. Compartment
transfer coefficients were based on average values estimated for four individuals from the
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Figure 4-20. Lead biokinetics based on Marcus (1985b). Bone is represented as an
extended cylindrical canalicular territory. The canalicular territory has
a radius b, and surrounds the canaliculus of radius a. Lead diffuses
across radius library, between the fluid in the canaliculus (which is in
communication with blood in the Haversian canal, not shown) and the
bone volume of the canalicular territory.
Rabinowitz et al. (1976) study. Initial average values for Pb in cortical bone for a given age at
the start of a simulation were derived from Barry (1975).

4.4.4.4 Contemporary Models
Additional information on Pb biokinetics, bone mineral metabolism, and Pb exposures has
led to further refinements and expansions of these earlier modeling efforts. In particular, three
pharmacokinetic models are currently being used or are being considered for broad application in
Pb risk assessment: (1) the Integrated Exposure Uptake BioKinetic (IEUBK) model for Pb in
children developed by EPA (U.S. Environmental Protection Agency, 1994a,b; White et al.,
1998); (2) the Leggett model, which simulates Pb kinetics from birth through adulthood
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Figure 4-21. Lead biokinetics based on Marcus (1985c). Blood is represented with a
plasma (central exchange) compartment and a red blood cell compartment,
the latter having shallow and deep pools.
(Leggett, 1993); and (3) the O'Flaherty model, which simulates Pb kinetics from birth through
adulthood (O'Flaherty, 1993, 1995). Of the three approaches, the O'Flaherty model has the
fewest Pb-specific parameters and relies more extensively on physiologically based parameters
to describe volumes, flows, composition, and metabolic activity of blood and bone that
determine the disposition of Pb in the human body. Both the IEUBK model and the Leggett
model are more classical multicompartmental models; that is, the values for the age-specific
transfer rate constants for Pb are based on kinetics data obtained from studies conducted in
animals and humans and may not have precise physiological correlates. Thus, the structure and
parameterization of the O'Flaherty model is distinct from both the IEUBK model and Leggett
model. All three models represent the rate of uptake of Pb (i.e., amount of Pb absorbed per unit
of time) as relatively simple functions (f) of Pb intake:
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Air Lead
Diet, Other
Gastrointestinal
tract
Figure 4-22. Lead biokinetics based on Bert et al. (1989).
Uptake = Intake • AF
(4-12)
Uptake = Intake • f(Intai
(Intake)
(4-13)
Values assigned to absorption factor (AF) or other variables in f(intake) are, in general,
age-specific and environmental medium-specific in some models. However, the models do not
modify the representation of uptake as functions of the many other physiologic variables that
may affect Pb absorption (e.g., nutritional status). While one can view this approach as a
limitation of the models, it also represents a limitation of the data available to support more
complex representations of Pb absorption.
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The IEUBK model simulates multimedia exposures, uptake, and kinetics of Pb in children
ages 0 to 7 years; the model is not intended for use in predicting Pb pharmacokinetics in adults.
The O'Flaherty and Leggett models are lifetime models, and include parameters that simulate
uptake and kinetics of Pb during infancy, childhood, adolescence, and adulthood. Lead exposure
(e.g., residence-specific environmental Pb concentrations, childhood activity patterns) is not
simulated by current versions of the O'Flaherty and Leggett models; however, this is not
necessarily a limitation since existing exposure models can be used to derive exposure inputs
(in terms of Pb intakes) for these models. By contrast, the IEUBK model includes parameters for
simulating exposures and uptake to estimate daily uptake of Pb (jig/day) among populations of
children potentially exposed via soil and dust ingestion, air inhalation, Pb-based paint chip
ingestion, tap water ingestion, and/or diet.
The above three models have been individually evaluated, to varying degrees, against
empirical physiological data on animals and humans and data on blood Pb concentrations in
individuals and/or populations (U.S. Environmental Protection Agency, 1994a,b; Leggett, 1993;
O'Flaherty, 1993). However, applications in risk assessment typically require that the models
accurately predict blood Pb distributions in real populations (U.S. Environmental Protection
Agency, 1994a), in particular those values or percentages falling in the "upper tails"
(e.g., >95th percentiles) of the distributions, when input to the models consists of data that
describe site-specific exposure conditions, e.g., environmental Pb concentrations or
physicochemical properties of soil and dust (Beck et al., 2001; Griffin et al., 1999a,b).
In evaluating models for use in risk assessment, exposure data collected at hazardous waste sites
have been used to drive model simulations (Bowers and Mattuck, 2001; Hogan et al., 1998).
The exposure module in the IEUBK model makes this type of evaluation feasible.

4.4.5 Integrated Exposure Uptake Biokinetic (IEUBK) Model for
Lead in Children
4.4.5.1 Model Structure
The IEUBK model for Pb in children (see Figure 4-23) is a multicompartmental
pharmacokinetics model linked to an exposure and probabilistic model of blood Pb concentration
distributions in children (U.S. Environmental Protection Agency, 1994a,b; White et al., 1998).
The model simulates exposure and biokinetics of Pb from birth to age 7 years (84 months) and
4-95
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Respiratory
tract
Uptake Component
I
Respiratory
tract

Gastrointestinal tract

Plasma extra -cellular fluid

/Ot
1
1
1
1
1
o
CL

O
o
o
Plasma extra-cellular fluid
O
CD
Elimination pools of
the body

I ntake from
environmental
Media (pgtead/day)
Body compartment
Body compartment or
elimination pool required in
more than one component
Figure 4-23. Structure of the integrated exposure uptake biokinetics model for lead
in children (U.S. Environmental Protection Agency, 1994a,b; White
et al., 1998).
was developed for predicting average quasi-steady state blood Pb concentrations corresponding

to daily average exposures, averaged over periods > 1 year.
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The model has four major components or submodels:
• Exposure model, in which average daily intakes of Pb (jig/day, averaged over a 1 year time
increment) are calculated for each inputted exposure concentration (or rates) of Pb in air,
diet, dust, soil, and water;
• Uptake model, which converts environmental media-specific Pb intake rates calculated from
the exposure model into a media-specific time-averaged rates of uptake (jig/day) of Pb to the
central compartment (blood plasma);
• Biokinetic model, which simulates the transfer of absorbed Pb between blood and other body
tissues, elimination of Pb from the body (via urine, feces, skin, hair, and nails), and predicts
an average blood Pb concentration for the exposure time period of interest; and
• Blood Pb probability model, which simply applies a log-normal distribution (with specific
geometric mean and geometric standard deviation parameters) to predict probabilities for the
occurrence of a specified blood Pb concentration in a population of similarly exposed
children.
Exposure Model. The exposure model simulates intake of Pb (jig/day) for exposures to
Pb in air (|ig/m3), drinking water (|ig/L), soil-derived dust (|ig/g), and diet (jig/day).
The temporal resolution of the exposure model is 1 year; exposure inputs are intended to
represent annual averages for an age-year time step (e.g., ages 1, 2, 3...years). Exposure inputs
that represent the average daily value for an age-year will yield corresponding daily average
intakes for the same age-year. The spatial resolution of the exposure model was intended to be a
child's residence (e.g., the home and yard). The model accepts inputs for media intake rates
(e.g., air volume breathing rates, drinking water consumption rate, soil and dust ingestion rate).
The air exposure pathway partitions exposure to outdoor air and indoor air; with age-dependent
values for time spent outdoors and indoors (hours/day). Exposure to Pb in soil derived dust is
also partitioned into outdoor and indoor contributions. The intakes from all ingested exposure
media (diet, drinking water, soil-derived dust) are summed to calculate a total intake to the
gastrointestinal tract, for estimating capacity-limited absorption (see description below of the
Uptake Model).
Uptake Model. The uptake model simulates Pb absorption in the gastrointestinal tract as
the sum of a capacity-limited (represented by a Michaelis-Menten type relationship) and
unlimited processes (represented by a first-order, linear relationship). These two terms are
intended to represent two different mechanisms of Pb absorption, an approach that is in accord
with limited available data in humans and animals that suggest a capacity limitation for Pb
4-97
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absorption (Mushak, 1991). One of the parameters for the capacity-limited absorption process
(that represents that maximum rate of absorption) is age-dependent. The above representation
gives rise to a decrease in the fractional absorption of ingested Pb as a function of total Pb intake
as well as age. Absorption fractions are also medium-specific (Figure 4-24).
At 30 months of age, at low intakes (<200 jig/day), below the rates at which capacity-
limitation has a significant impact on absorption, the fraction of ingested Pb in food or drinking
water that is absorbed is 0.5 and decreases to 0.1 at high intake (5000 jig/day). For Pb ingested
in soil or dust, fractional absorption is 0.3 at low intake (<200 jig/day) and decreases to 0.1 at
high intake (5000 |ig/day).
The uptake model has a default absorption rate of 32% (percent of inhaled Pb that is
absorbed). This default absorption rate was intended to be an intermediate value between the
absorption rate of 25% to 45% for young children living in non-point source areas and a rate of
42% for children living near point sources (U.S. Environmental Protection Agency, 1994b).
In derivation of these absorption rates, it was assumed that Pb deposited in the alveolar region is
completely absorbed, whereas, Pb deposited in the nasopharyngeal and tracheobronchial regions
is transported to the gastrointestinal tract where -40% absorption occurs. Furthermore, this
default absorption rate was based on a respiratory particle deposition fraction for the particle size
distribution to which children were assumed to be exposed. If sufficient information about a
child's exposure is available, it is possible to calculate particle deposition fractions using
publicly available particle dosimetry models (see Section 4.2.1). Using these new particle
deposition fraction data, a recalculated absorption rate can be used in place of the default value.
Biokinetics Model. The biokinetics model includes a central compartment, plasma and
extracellular fluid combined (plasma-ECF), six peripheral body compartments, and three
elimination pathways. The temporal resolution of the biokinetics model is 1 month and, as
discussed below, parameter values for bone-plasma-ECF exchanges were assigned with the
objective of simulating the quasi-steady state condition of months, rather than short-term kinetics
of days. The body compartments include kidney, liver, trabecular bone, cortical bone, and other
soft tissue. The model simulates growth of the body and tissues, compartment volumes, and Pb
masses and concentrations in each compartment. Blood Pb concentration at birth (neonatal)
is assumed to be 0.85 of the maternal blood Pb (see Section 4.2.2. for discussion of observed
4-98
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0.55
0.50
0,45
a
u.
I
0,40
0.35
0.30
10 20 30 40 50 60 70
Food or Water Lead Intake (ug/day)
-•-Age 6m
-B-Age 18m
—i— Age 30m
-*-Age42m
-*-Age54m
-*-~Age66m
-*-Age78m
80
90
100
0.35
0,30
-•-Age 6m
-~B-Age 18m
—t— Age 30m
-*-Age42m
0.25
S
u.
0.20
0,15
0.10
1000 2000 3000 4000 5000 6000 7000
Soil or Dust Lead Concentration (ppm)
8000
9000
10000
Figure 4-24. Age-dependency of absorption fraction for ingested lead in the IEUBK model
for lead in children. Absorption fraction for food and water (top panel); soil
and dust (bottom panel).
4-99
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neonatal-maternal blood Pb concentration relationships). Neonatal Pb masses and concentrations
are assigned to other compartments based on a weighted distribution of the neonatal blood Pb
concentration. Exchanges between the central compartment and tissue compartments are
simulated as first-order processes, which are parameterized with unidirectional, first-order rate
coefficients. Rate coefficients are allometrically scaled as a power function of body weight
(BW033). Saturable uptake of Pb into erythrocytes is simulated, with a maximum erythrocyte Pb
concentration of 120 |ig/L. Excretory routes simulated include urine, from the central
compartment; bile-feces, from the liver; and a lumped excretory pathway representing losses to
skin, hair and nails, from the "other soft tissue" compartment.
Bone is simulated as a trabecular bone compartment (20% of bone volume) and a cortical
bone compartment (80% of bone volume). Rate constants for transfer from plasma to the two
bone compartments are assigned values that result in a 4:1 cortical Pb:trabecular Pb mass ratio
within one biokinetic time step (one month). This is achieved by assigning the two bone
compartments identical rate coefficients for transfer of Pb from bone to plasma-ECF (half-time
8.5 days, at age 2 years), and faster (cortical, half-time 0.0083 days) and slower transfer
(trabecular, half-time 0.035 days) from the plasma-ECF (cortical:trabecular rate ratio is -4:1).
Note, this approach is different from previous and subsequent modeling approaches, in which
cortical bone-to-plasma (or blood) transfer is assumed to occur slowly, relative to trabecular
bone-to-plasma transfer (Marcus, 1985a; Bert et al., 1989; Leggett, 1993; O'Flaherty, 1993,
1995). For predictions of quasi-steady state conditions and the intended use of the IEUBK
Model, the two general approaches can be expected to yield similar distributions of Pb between
the cortical and trabecular bone compartments.
Blood Lead Probability Model. Inputs to the IEUBK model are exposure point estimates
that are intended to represent time-averaged central tendency exposures. The output of the
model is a central tendency estimate of blood Pb concentration for children who might
experience the inputted average exposures. However, within a group of similarly exposed
children, blood Pb concentrations would be expected to vary among children as a result of inter-
individual variability in media intakes (e.g., daily average intakes of soil-derived dust, drinking
water, or food), absorption, and biokinetics. The model simulates the combined impact of these
sources of variability as a lognormal distribution of blood Pb concentration for which the
geometric mean (GM) is given by the central tendency blood Pb concentration outputted from
4-100
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the biokinetics model, and the geometric standard deviation (GSD) is an input parameter.
The resulting lognormal distribution also provides the basis for predicting the probability of
occurrence of given blood Pb concentration within a population of similarly exposed children:

PIO = probability of exceeding a blood Pb concentration of 10 |ig/dL (4-14)

The model can be iterated for varying exposure concentrations (e.g., a series of increasing
soil Pb concentration) to predict the media concentration that would be associated with a
probability of 0.05 for the occurrence of a blood Pb concentration exceeding 10 |ig/dL
(Pio=0.05).

4.4.5.2 Model Calibration and Evaluation
Evaluations of the IEUBK model have been carried out by comparison of model
predictions of blood Pb concentrations in children with observations from epidemiologic studies
of hazardous waste sites (Hogan et al., 1998; Bowers and Mattuck, 2001; TerraGraphics, Inc.,
2000, 2001). Data characterizing residential Pb exposures and blood Pb concentrations in
children living at four Superfund National Priorities List (NPL) sites were collected in a study
designed by the Agency for Toxic Substances and Disease Registry (ATSDR) and EPA.
The residential exposure data were used as inputs to the IEUBK model and predicted blood Pb
concentration distributions were compared to the observed distributions in children living at the
same residences. The IEUBK model predictions of geometric mean blood Pb concentrations for
children whose exposures were predominantly from their residence (i.e., no more than
10 hours/week away from home) were within 0.7 |ig/dL of the observed geometric mean at each
site (Table 4-15). The prediction of the percentage of children expected to have blood Pb
concentrations exceeding 10 |ig/dL were within 4% of the observed percentage at each site
(Table 4-16). A similar type of empirical comparison was conducted by Bowers and Mattuck
(2001) based on data from 4 mining and/or smelting sites. The results from both studies
(Hogan et al., 1998 and Bowers and Mattuck, 2001) are shown in Figure 4-25. TerraGraphics,
Inc. (2000) reported predicted and observed blood Pb concentrations in 2-year old children who
resided at the Bunker Hill, ID site during the period 1988-1998. Comparisons of observed and
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Table 4-15. Comparison of Observed and Predicted Geometric Mean Blood Lead for
Three Community Blood Lead Studies
Dataset

Galena, KA
Jasper Co, MP
Madison Co, ILa
Palmerton, PAb
N

111
333
34
Observed Blood Lead
(|ig/dL)
GM
5.2
5.9
6.8
95% CI
4.5-5.9
5.5-6.4
5.6-8.2
Model Predictions
(|ig/dL)
GM
4.6
5.9
7.5
95% CI
4.0-5.3
5.4-6.3
6.6-8.6
CI = confidence interval; GM = geometric means
aChildren away from home < 10 hours/week
bChildren away from home <20 hours/week
Table 4-16. Comparison of Observed and Predicted Probability of Exceeding a Blood
Lead Concentration of 10 ug/dL Lead for Three Community Blood Lead Studies
Dataset

Galena, KA
Jasper Co, MP
Madison Co, ILa
Palmerton, PAb
N

111
333
34
Observed Blood Lead
(|ig/dL)
Percent
20
19
29
95% CI
13-27
15-23
14-44
Model Predictions
(|ig/dL)
Percent
18
23
31
95% CI
11-25
19-28
16-47
CI, confidence interval
aChildren away from home < 10 hours/week
bChildren away from home <20 hours/week
predicted GM blood Pb concentration and PIO are shown in Figure 4-26. Empirical comparisons
have shown that agreement or disparity between IEUBK model predictions and observed blood
Pb concentrations at specific locations is influenced by numerous factors, including (a) the extent
to which the exposure and blood Pb measurements are adequately matched and (b) site-specific
factors (e.g., soil characteristics, behavior patterns, bioavailability) that may affect Pb intake or
4-102
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Observed GM Blood Lead (M9/dL)
1*1
12
10
8
6
4
2
n

• Hoganetal., 1998 T
• Bowers and Mattuck, 2001 ! " '
JL
f
*
f
*
*
f
+
+
__ 4>
^
*•
*
I Ik * I
,i_3- ,'
T ^i'T* ' ,''
^ ^V', i ,
,' ' 1 '
^-"
0
4 6 8 10 12

Predicted GM Blood Lead (|jg/dL)
14
Figure 4-25. Comparisons of IEUBK model predictions and observed blood lead
concentrations. Data from Hogan et al. (1998) include observations from
3 sites (left to right): Galena, KS and Jasper Co, MO, combined, N=lll;
Madison Co, IL, N=333; and Palmerton, PA, N=34. Data from Bowers and
Mattuck (2001) include observations from 4 sites (left to right): Sandy, UT
(reported as Sandy, UT-2), N=36; Midvale, UT, N=151; Palmerton, PA,
N=110; and Cincinnati, OH, N=95. Bars are 95% confidence limits; dashed
lines represent simple linear regression models based on means (not
individual child data) from each study (Hogan et al., r2>0.99; Bowers and
Mattuck, r2=0.84; these r2 are the fit to the group means and exceed the r2
that would be expected for the complete datasets).
uptake in children. Error in measurement of exposure, by itself, can be expected to attenuate the
predicted slope of the relationship between exposure and blood Pb concentration (Carrol and
Galindo, 1998; Marcus and Elias, 1998). In the absence of a suitable dataset of paired
environmental Pb and blood Pb measurements at a given site, it is not possible to ascertain the
degree to which the model predictions will represent the exposure-blood Pb concentration
relationships at that site (Bowers and Mattuck, 2001).
4-103
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20
O)
3.
ID 15
CD
CD
O
O 10
CD
5
O
"O c
ffl =>

0)
-Q
O n
5 10 15

Predicted GM Blood Lead (MQ/dL)
20
60
50
o
"S 30
2
0
en on
O
10
10 20 30 40
Predicted P10 (%)
50
60
Figure 4-26. Comparison of IEUBK model predictions and observed blood lead
concentrations. Data are annual (1988-1998, N = 23 to 57 each year) site-
wide assessments of 2-year old children who resided at the Bunker Hill site in
northern Idaho. Soil and dust lead bioavailability was assumed to be 18%.
Relative contributions of soil and house dust to total soil/dust ingestion were
assumed to be 42% interior dust; 27% community soil; 10% neighborhood
soil (within 100-ft radius of residence); and 12% yard soil. Dashed lines
represent simple linear regression models applied to the geometric means
(r2 = 0.92*) or PlOs (r2 = 0.91*). *Note that this r2 is for the fit of the model
to the GM and does not consider the additional variability of the complete
data sets.
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4.4.5.3 Model Applications
The U.S. EPA has recommended that the IEUBK model be used to assess health risks to
children exposed to contaminated soils at hazardous waste sites (U.S. Environmental Protection
Agency, 1994c, 1998).
Biomarkers Simulated. The IEUBK model computes masses of Pb in bone and various
soft tissues, and excretion of Pb, which are used in the computation of blood Pb concentration.
However, the model was not developed for the purpose of predicting Pb masses in these tissues
or excreta. Blood Pb concentration is the only Pb biomarker output that is accessible to the user.
Exposure Inputs. The IEUBK model was developed to predict the probability of elevated
blood Pb concentrations in children exposed to user-specified annual average exposures to Pb in
air, food, drinking water, soil, and dust. As noted above, the exposure model has an age-year
time step (the smallest time interval for a single exposure event) and, therefore, is more suited to
applications in which long-term (i.e., > 1 year) average exposures and quasi-steady state blood Pb
concentrations are to be simulated. Intermittent exposures occur for brief periods of time
(e.g., a weekend at the beach), or in cases where significant seasonal variations are different from
the typical residential or occupational exposure. Intermittent exposures can be simulated as
time-weighted average exposures (U.S. Environmental Protection Agency, 2003a). Shorter-term
dynamics of blood Pb concentration, that may result from exposures that are highly variable on
time scales of days or weeks, will not be captured with this approach (Lorenzana et al., 2005;
Khoury and Diamond, 2003).
Modeling Variability and Uncertainty. As noted above, the IEUBK model uses a
lognormal probability model to simulate inter-individual variability in blood Pb concentrations
attributable to variability in media intakes, absorption, and biokinetics. The model provides a
generic default value of 1.6 for the geometric standard deviation (GSD;) of blood Pb
concentrations. This value, which can be altered by the user, was derived from an analysis of
exposure (soil Pb)-stratified variability in blood Pb concentrations in various cohorts of children
(U.S. Environmental Protection Agency, 1994a; White et al., 1998). Griffin et al. (1999b) also
explores various statistical methods for estimating an appropriate GSD; (regression, box
modeling, structural equation modeling).
A Monte Carlo approach has been used to simulate and propagate variability and
uncertainty in exposure and absorption through IEUBK model simulation of blood Pb
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concentrations (Goodrum et al., 1996). This extension of the model provides an alternative to
the generic blood Pb probability approach for incorporating explicit estimates of variability (and
uncertainty in variability) in exposure and absorption into predictions of an expected probability
distribution of blood Pb concentrations. A quantitative uncertainty analysis of IEUBK model-
based estimates of the PIO for a smelter site in Utah revealed that parameters specifying soil
ingestion rate were a dominant contributor to uncertainty in the PIO; however, the contribution of
soil ingestion uncertainty, relative to uncertainty in other model parameters (i.e., mean soil Pb
concentration, absorption fraction) varied across individual locations (Initial Study Zones) at the
site (Griffin et al., 1999a).

4.4.5.4 Implementation Code
The IEUBK model was initially released to the public in 1994 as a compiled DOS-based
C program (IEUBK v99d). This version was subjected to an independent code validation and
verification study which verified that the code accurately implement the model (Mickle, 1998;
Zaragoza and Hogan, 1998). A 32-bit C++ (IEUBKwin32) version of the model is available for
download from an EPA website (http://www.epa.gov/superfund/programs/Pb/ieubk.htm).
The IEUBKwin32 program outputs blood Pb concentrations and probabilities of exceeding a
given blood Pb concentration. Levels of Pb in other tissue compartments are computed but are
not accessible to the user.

4.4.6 Leggett Model
4.4.6.1 Model Structure
The Leggett model was developed from a biokinetic model originally developed for the
International Commission on Radiological Protection (ICRP) for calculating radiation doses
from environmentally important bone-seeking radionuclides, including radioisotopes of Pb
(Leggett, 1985, 1992a,b). The model 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; U.S. Environmental Protection Agency, 1998a). The model
includes a central exchange compartment, 15 peripheral body compartments, and 3 elimination
pools (Figure 4-27). The central exchange compartment is the diffusible pool of Pb in plasma.
The model simulates a bound pool in plasma (i.e., Pb bound to plasma proteins); that has an
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1
1
I
I
I
1
I
1
1
1
1
1
I
I
1

Bound
Plasma
Figure 4-27. Structure of the Leggett Lead Biokinetic Model (Leggett, 1993). The central
exchange compartment is diffusible plasma. Bone is represented as having
surface (which rapidly exchanges with plasma) and volume compartments;
the latter stimulates slow exchange with the surface and slow return of lead
to the plasma from bone resorption.
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equilibrium ratio (bound:free) of ~5. Transport of Pb from plasma to tissues is assumed to
follow first-order kinetics. The temporal resolution of the model is 1 day. Transfer rate
constants vary with age and blood Pb concentration. The latter adjustment accounts for the
limited uptake of plasma Pb into red blood cells and the resulting shift in distribution of Pb from
plasma-ECF to other tissues. Above a nonlinear threshold concentration in red blood cells
(assumed to be 60 |ig/dL), the rate constant for transfer to red blood cells declines and constants
to all other tissues increase proportionally (Leggett, 1993). This replicates the nonlinear
relationship between plasma and red blood observed in humans (Smith et al., 2002; Manton
et al., 2001; Bergdahl et al., 1997a, 1998, 1999). The model simulates blood volume as an age-
dependent function, which allows simulation of plasma and blood Pb concentrations. However,
volumes of other tissues are not simulated; therefore, only Pb masses in these tissues, and not
concentrations are simulated.
First-order transfer coefficients (day-1) between compartments were developed for six age
groups, and intermediate age-specific values are obtained by linear interpolation (Leggett, 1993).
The total transfer rate from diffusible plasma to all destinations (TPALL) combined is assumed
to be 2000 day-1, based on isotope tracer studies in humans receiving Pb via injection or
inhalation. Values for transfer coefficients from plasma to tissues and tissue compartments are
based on measured deposition fractions (DF) or instantaneous fractional outflows of Pb between
tissues compartments (Leggett, 1993), where the transfer coefficient to a specific tissue or
compartment (TP;) is given by:
TPALL (4-15)
This approach establishes mass balance with respect to the transfer rates from plasma:

(4-16)

The model simulates both rapid exchange of Pb with plasma via bone surface and slow
loss by bone resorption. Cortical bone volume (80% of bone volume) and trabecular bone
volume (20% of bone volume) are simulated as bone surface compartments, which rapidly
exchanges with Pb in plasma, and bone volume, within which are exchangeable and
nonexchangeable pools. Lead enters the exchangeable pool of bone volume via the bone surface
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and can return to the bone surface, or move to the nonexchangeable pool, from where it can
return to the plasma only when bone is resorbed. Transfers from plasma to bone surface, return
from bone surface to plasma, and bone surface to exchangeable bone volume are assumed to be
relatively fast processes (adult ti/2 = 3.85, 1.4, and 1.4 days, respectively). Return of Pb from the
exchangeable bone volume is slower (adult ti/2 = 30 days); however, the dominant transfer
process determining long-term accrual of bone Pb burden are slow rate coefficients for transfer
of Pb from the nonexchangeable pools of trabecular and cortical bone to plasma (adult ti/2 = 3.8
and 23 years, respectively). Bone transfer coefficients vary with age (faster in children) to
reflect the age-dependence of bone turnover. The slow, nonexchangeable, bone volume
compartment is much more labile in infants and children than in adults (e.g., cortical ti/2 =
68 days at birth and 1,354 days at age 15 years; trabecular ti/2 = 68 days at birth and 725 days at
age 15 years). Other physiological states (such as pregnancy and menopause) that affect bone
turnover and, therefore, bone Pb kinetics are not simulated, although such states could
conceivably be accommodated with adjustments to tissue (e.g., bone) transfer coefficients.
The liver is simulated as two compartments; one compartment has a relatively short
removal half-life for transfers to plasma and to the small intestine by biliary secretion (adult
ti/2 = 10 days); a second compartment simulates a more gradual transfer to plasma of-10% of Pb
uptake in liver (adult ti/2 = 365 days). The kidney is simulated as two compartments, one that
exchanges slowly with blood plasma and accounts for Pb accumulation in kidney tissue (adult
ti/2 = 365 days) and a second compartment that receives Pb from blood plasma and rapidly
transfers Pb to urine (adult ti/2 = 5 days), with essentially no accumulation (urinary pathway).
Other soft tissues are simulated as three compartments representing rapid, intermediate, and slow
turnover rates, without specific physiologic correlates (adult ti/2 = 0.3, 100, and 1824 days,
respectively). Other excretory pathways (hair, nails, and skin) are represented as a lumped
pathway from the intermediate turnover rate of the soft tissue compartment.
The Leggett model simulates Pb intakes from inhalation, ingestion, or intravenous
injection. The latter was included to accommodate model evaluations based on intravenous
injection studies in humans and animal models. The respiratory tract is simulated as four
compartments into which inhaled Pb is deposited and absorbed with half-times of 1, 3, 10,
and 48 hours. Five percent of the inhaled Pb is assumed to be transferred to the GI tract. These
parameter values reflect the data on which the model was based, which were derived from
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studies in which human subjects inhaled submicron Pb-bearing particles (Morrow et al., 1980;
Chamberlain et al., 1978; Wells et al., 1977; Hursh and Mercer, 1970; Hursh et al., 1969).

These assumptions would not necessarily apply for exposures to larger airborne particles

(see Sections 2.3.1 for a discussion of atmospheric transport of Pb particles). Absorption of

ingested Pb is simulated as an age-dependent fraction of the ingestion rate, declining from

0.45 at birth to 0.3 at age 1 year (to age 15 years), and to 0.15 after age 25 years (Figure 4-28).
60
EUBK Food and Drinking Water
Figure 4-28. Age-dependency of absorption fraction for ingested lead in the Leggett and
O'Flaherty models. The IEUBK model projects absorption only through age
seven (84 mo). At intakes below those which approach the limit on "active"
absorption of lead, absorption is constant with age, with default valves of
50% for diet and drinking water, 30% for soil and dust. Fractional
absorption via the active pathway decreases with age and lead intake
(see Figure 4-23).
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4.4.6.2 Model Calibration and Evaluation
Leggett (1993) and Pounds and Leggett (1998) describe various qualitative empirical
comparisons of model predictions against observations made on adults (e.g., Skerfving et al.,
1985; Campbell et al., 1984; Manton and Cook, 1984; Barry, 1981; DeSilva, 1981; Chamberlain
et al., 1978; Rabinowitz et al., 1976; Barry, 1975; Griffin et al., 1975; Gross et al., 1975; Hursh
and Mercer, 1970; Booker et al., 1969; Hursh et al., 1969; Schroeder and Tipton, 1968).
Age-specific changes in parameter values that specify the biokinetics of Pb in children were
assigned values that resulted in agreement between predicted age-specific Pb distribution
(fraction of body burden) in blood, bone, brain, kidney, liver, and other tissues, and reported
postmortem values (Schroeder and Tipton, 1968; Barry, 1975, Gross et al. 1975; Barry, 1981).
Comparisons of model predictions to observed relationships between plasma and red blood cell
Pb levels are reported in U.S. Environmental Protection Agency (2003b).

4.4.6.3 Model Applications
Biomarkers Simulated. The Leggett model simulates the concentrations of Pb in blood
and plasma, masses of Pb in bone and various soft tissues, and excretion of Pb in urine that
correspond to lifetime exposures (in terms of daily Pb intakes).
Exposure Inputs. The model does not contain a detailed exposure module (although it can
be linked to an exposure model); Pb exposure estimates are incorporated into the simulations as
age-specific point estimates of daily intake (jig/day) from ingestion, inhalation, or injection.
The model operates with a Pb intake time step of 1 day, which allows simulation of rapidly
changing (i.e., daily) intermittent exposures (Lorenzana et al., 2005; Khoury and Diamond,
2003). Assumptions of blood Pb concentrations at birth can also be introduced into the
simulations, from which levels in other tissue in the first time step after birth are calculated.
Dose reconstruction is possible with this model, since intakes, and corresponding tissue
Pb burdens accrued at any period in the lifetime, prior to an exposure event of interest, can be
simulated. Pounds and Leggett (1998) illustrate this in a study of a childhood Pb poisoning case,
in which the exposure is followed by chelation. Chelation was simulated as a short-duration
increase in the plasma Pb deposition fraction to urine, with corresponding proportional decreases
in deposition fractions to other tissues.
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4.4.6.4 Implementation Code
The Leggett model was initially developed as a Fortran code, which can be run, without
compiling, from various platforms, including DOS and Windows (see Pounds and Leggett, 1998
for a description). A version compiled in Advanced Continuous Simulation Language (ACSL)
has also been reported (Lorenzana et al., 2005). Confirmation of the Leggett model code was
carried out by a panel of experts (ICRP, 1989, 1993).

4.4.7 O'Flaherty Model
4.4.7.1 Model Structure
The O'Flaherty model simulates Pb exposure, uptake, and disposition in humans, from
birth through adulthood (O'Flaherty, 1993, 1995, 2000). Figure 4-29 shows a conceptualized
representation of the model. Important novel features of the O'Flaherty model are the simulation
of growth, bone formation, and resorption. A growth curve is simulated with a logistic
expression relating body weight to age in males or females. The full expression relating weight
to age has five parameters (constants), so that it can readily be adapted to fit a range of
standardized growth curves for males and females. Tissue growth and volumes are linked to
body weight; this provides explicit modeling of Pb concentrations in all tissues simulated. Other
physiologic functions (e.g., bone formation) are linked to body weight, age, or to both.
The model can be implemented with a temporal resolution of 1 day; however, as originally
configured, the rate parameters are expressed in time units of years.
Rates of bone formation and resorption are simulated as age-dependent functions
(Figure 4-30). Uptake and release of Pb from trabecular bone and metabolically active cortical
bone are functions of bone formation and resorption rates, respectively; this establishes the age-
dependence to the Pb kinetics in and out of bone. Lead exchange between blood plasma and
bone is simulated as parallel processes occurring in cortical (80% of bone volume) and trabecular
bone (20% of bone volume). The model simulates an age-related transition from immature bone,
for which bone turnover (formation and resorption) rates are relatively high, to mature bone,
for which turnover is relatively slow. Changes in bone mineral turnover associated with aging
and senescence (e.g., postmenopausal osteoporosis) can be simulated by introducing an age-
dependent increase rate of bone resorption (O'Flaherty, 2000). Metabolically active regions of
bone, in which Pb uptake and loss is dominated by bone formation and loss, a region of slow
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Respiratory Tract
o
o
Elimination pools of
the body

Intake from environmental
Media (|jg/lead/year)

Body compartment
Blood Plasma
Well-Perfused Tissues
Poorly Perfused Tissues
Cortical Bone
Trabecular Bone
Liver
Kidney
Diet, Dust,
Paint, Soil,
Water
Gastrointestinal Tract
Figure 4-29. Structure of the O'Flaherty Lead Exposure Biokinetics Model (O'Flaherty,
1993,1995, 2000). The central exchanges compartment is diffusible plasma.
Lead distribution is represented by flows from blood plasma to liver,
kidney, richly-perfused tissues, poorly-perfused tissues, and cortical and
trabecular bone. The model simulates tissue growth with age, including
growth and resorption of bone mineral.
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10 15 20
Age (year)
25
30
0.6
0.5
— 0.4
0)
E
0)
c
o
0.3
0.2
0.1
0.0
Juvenile cortical
bone
Juvenile trabecular
bone
10 15 20
Age (year)
25
30
Figure 4-30. Bone growth as simulated by the O'Flaherty Lead Exposure Biokinetics
Model (O'Flaherty, 1993,1995, 2000). The model simulates an age-related
transition from juvenile bone, in which bone turn-over (formation and
resorption) rates are relatively high, to mature bone, in which turn-over is
relatively slow. Cortical bone comprises -80% of total bone volume.
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kinetics in mature cortical bone is also simulated, in which Pb uptake and release to blood occur
by heteroionic exchange with other minerals (e.g., calcium). Heteroionic exchange is simulated
as a radial diffusion in bone volume of the osteon. All three processes are linked to body weight
or the rate of change of weight with age. This approach allows for explicit simulation of the
effects of bone formation (e.g., growth) and loss, changes in bone volume, and bone maturation
on Pb uptake and release from bone. Exchanges of Pb between blood plasma and soft tissues
(e.g., kidney and liver) are represented as flow-limited processes. The model simulates saturable
binding of Pb in erythrocytes (maximum capacity is 2.7 mg Pb/L cell volume); this replicates the
curvilinear relationship between plasma and erythrocyte Pb concentrations observed in humans
(Smith et al., 2002; Manton et al., 2001; Bergdahl et al., 1997a, 1998, 1999). Excretory routes
include kidney to urine and liver to bile. Total excretion (clearance from plasma attributable to
bile and urine) is simulated as a function of age-dependent glomerular filtration rate. Biliary and
urinary excretory rates are proportioned as 70 and 30% of the total plasma clearance,
respectively.
The O'Flaherty model simulates Pb intake from inhalation and ingestion. Inhalation rates
are age-dependent. Absorption of inhaled Pb is simulated as a fraction (0.5) of the amount
inhaled and is independent of age. Gastrointestinal absorption of Pb in diet and drinking water is
simulated as an age-dependent fraction, declining from 0.58 of the ingestion rate at birth to
0.08 after age 8 years (Figure 4-28). These values can be factored to account for relative
bioavailability when applied to absorption of Pb ingested in dust or soil.

4.4.7.2 Model Calibration and Evaluation
The O'Flaherty model was initially calibrated to predict blood, bone, and tissue Pb
concentrations in rats (O'Flaherty, 1991a,b,c) and was later modified to reflect anatomical and
physiological characteristics in children (O'Flaherty, 1995), adults (O'Flaherty, 1993), and
Cynomolgus monkeys (M. fasicularis) (O'Flaherty et al., 1998). Model parameters were
modified to correspond with available information on species- and age-specific anatomy and
physiological processes. Empirical comparisons (largely qualitative) of model predictions
against observations made in adults (e.g., Van De Vyver et al., 1988; Kehoe, 1987; Marcus,
1985c; Manton and Malloy, 1983; Sherlock et al., 1982; DeSilva, 1981; Moore et al., 1977;
Cools et al., 1976; Rabinowitz et al., 1976; Azar et al., 1975) are provided in O'Flaherty (1993);
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and comparisons against observations made in children (e.g., Sherlock and Quinn, 1986;
Bornschein et al., 1985b; Chisolm et al., 1985; Lacey et al., 1985) and adults are described in
O'Flaherty (1995, 1998, 2000).

4.4.7.3 Model Applications
Biomarkers Simulated. The O'Flaherty model simulates Pb concentrations in blood and
plasma, bone, and various soft tissues, and excretion of Pb in urine that correspond to lifetime
exposures (in terms of daily Pb intakes). Lead in feces is a mixture of unknown proportions of
unabsorbed Pb in food, drinking water, ingested dust, a small amount of inhaled Pb entering the
GI tract by the mucociliary clearance from the respiratory tract, and a small amount of absorbed
Pb eliminated with the red blood cells passing along the bile duct to the GI tract. In this respect,
Pb in feces represents a poorly defined measure of Pb exposure.
Lead in perspiration represents Pb in extracellular plasma, but the concentration is low
and difficult to measure in a small volume (1 drop = 0.05 mL) and is potentially contaminated
with Pb in dust on the skin surface.
The model predicts blood Pb concentrations for a broad age range (infants to adults),
which allows for simulated dose reconstruction, since intakes and corresponding tissue Pb
burdens accrued at any period in the lifetime, prior to an exposure event of interest can be
simulated. Physiological states (such as pregnancy and menopause) that affect bone turnover
and, therefore, bone Pb kinetics are not simulated, although such states could be accommodated
with adjustments to the physiological bone formation and resorption rates.
Exposure Inputs. The O'Flaherty model simulates Pb intake by inhalation and ingestion.
The model simulates ingestion exposures from infant formula, soil, dust, and drinking water.
Rates of soil and dust ingestion are age-dependent, increasing to -130 mg/day at age 2 years and
declining to <1 mg/day after age 10 years. However, the ACSL implementation code allows
constructions of simulations with an exposure time step as small as 1 day, which would allow
simulation of rapidly changing intermittent exposures (e.g., an acute exposure event).
Modeling Variability and Uncertainty. The O'Flaherty model, as described in O'Flaherty
(1993, 1995), utilizes point estimates for parameter values and yields point estimates as output;
however, a subsequent elaboration of the model has been reported that utilized a Monte Carlo
approach to simulate variability in exposure, absorption, and erythrocyte Pb binding capacity
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(Beck et al., 2001). This approach could be used to predict the probability that children exposed
to Pb in environmental media will have blood Pb concentrations exceeding a health-based level
of concern (e.g., 10 |ig/dL).

4.4.7.4 Implementation Code
The O'Flaherty model was developed in ACSL and published in O'Flaherty (2000).
A compiled C program has also been developed (personal communication, E. O'Flaherty).
The extent to which code verification and validation studies have been conducted for the
O'Flaherty model is unclear at this time. However, analogs of certain components of the
O'Flaherty model (e.g., parameters related to bone growth) have been incorporated into the EPA
All Ages Lead Model (AALM) (see Section 4.4.7) as a potential option for evaluation.

4.4.8 EPA All Ages Lead Model
4.4.8.1 Model Structure
The EPA All Ages Lead Model (AALM), currently under development, simulates lifetime
Pb exposures and biokinetics in humans (Figure 4-31). The model is expected to simulate
exposure and biokinetics of Pb from birth to age 90 years and should also incorporate, at some
near-future time, a pregnancy module that simulates transplacental transfer of Pb from the
mother to the fetus.
Exposure Module. The exposure component of the AALM incorporates and extends the
exposure component of the IEUBK model. The AALM exposure model defines an individual in
terms of age, sex, date of birth, and activity pattern profile. The age specification establishes up
to nine age ranges (e.g., infant, child, adolescent, adult, etc.) for which various exposure (and
biokinetic) parameter values can be applied. This provides a means for varying parameter values
with age. The sex specification links the modeled individual to the appropriate growth algorithm
(O'Flaherty 1993, 1995), and the date specification links the individual to historical exposure
levels (e.g., air, diet) for the selected age range. The activity pattern specification sets the
relative amount of time the individual spends in various exposure settings (e.g., residential,
school, recreational, occupational) for which exposure concentrations can be specified.
The diet exposure module allows input values (current or historical) for Pb levels (|ig/g)
in market basket fruits, vegetables, meat and fish; recreational- or subsistence-harvested fish and
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Resp
Tr

ratory
set
Skeleton
Gastrointestinal Tract
Plasma Extracellular Fluid
Kidneys

Other
Kidney
Tissue

Urinary
Path

Bladder
Contents

Urine
Figure 4-31. Structure of the All Ages Lead Model. The AALM adds a comprehensive
exposure component and an uptake component to a revised and recoded
version of the Leggett model to produce a model with fully selectable
exposure, uptake, and biokinetic parameters.
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meat; and corresponding food intakes for each food type (jig food/day). Lead intake from
drinking water is calculated from concentrations (|ig/L) in tap water (first draw and/or flushed),
fountain water, and/or bottled water; and corresponding source water intake rates (L/day).
The dust exposure module accepts input values for dust concentrations (|ig/g) in various
settings (e.g., residential, school, recreational, occupational) or dust loadings (|ig/m2) and
corresponding dust ingestion rates (jig dust/day) or contact rates (m2/day), the Pb ingestion rate
for a given loading being calculated as the product of loading and contact rate. Pica ingestion for
soil and/or paint chips can be simulated with input values for Pb levels in soil (ng/g) or paint
(|ig/cm2) and corresponding pica ingestion rates (g soil/day, cm2 paint/day). Dermal exposure to
Pb in dust can also be simulated with input values for dust Pb level (ng/g), dust loading on the
skin (mg/cm2), and skin exposure rate (cm2/day).
Calculated Pb intakes for each exposure pathway are summed to calculate total intakes
(jig/day) to the respiratory tract, gastrointestinal tract, and dermal pathway, respectively.
The exposure model time step is 1 day (the smallest time interval for a single exposure event).

Biokinetics Module. The biokinetics module of the AALM is based on Leggett (1993)
with the following modifications and enhancements:

1. A simulation of dermal absorption is implemented that calculates transfer of
Pb from the skin to the central plasma compartment, as a function of rate of
dermal contact with Pb (jig/day) and a dermal absorption fraction.

2. Male and female growth algorithms for body weight, soft tissues, and
cortical and trabecular bone are implemented, based on O'Flaherty (1993,
1995). This allows simulation of tissue growth and volumes, as well as Pb
concentrations in all tissues simulated.

3. A simulation of maternal-fetal transfer is implemented that simulates Pb
levels in fetal tissues, and establishes blood and tissue Pb levels for a
postnatal simulation. This provides a means for multigeneration simulation
of exposure and Pb biokinetics.

4.4.9 Model Comparisons
Table 4-17 summarizes the major features of various models of human exposure that
predict tissue Pb burdens. The slope factor models give similar predictions of quasi-steady state
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Table 4-17. Summary of Models of Human Exposure that Predict Tissue Distribution of Lead
to
o
Model
U.S. Environmental
Protection Agency
IEUBK Model
White etal. (1998)

U.S. Environmental
Protection Agency
AALM (2005)

Leggett (1985)

O'Flaherty (1993, 1995)

Age
Range Exposure Pathways
0-7 yr Air
Diet
Soil/dust
Water
Other
0-90 yr Air
Diet
Soil/dust
Water
Other

0-Adult Intakes (inhaled,
ingested, injected)

0-Adult Air
Diet
Soil/dust
Water
Other

Exposure Biomarkers
Time Step Biokinetics Simulation Predicted
1 year Multicompartmental Blood Pb

1 day Multicompartmental Blood
Bone
Brain
Fetus
Kidney
Liver
Urine
1 day Multicompartmental Blood
Bone
Brain
Kidney
Liver
Urine
1 year Multicompartmental Blood
(code supports Bone
1 day) Brain
Kidney
Liver
Urine
Variability and
Uncertainty Simulation
Variability: blood PbGSD,
Variability/uncertainty :
MCA (Griffin et al., 1999b)

Variability and uncertainty
determined by independent
assessment of multiple runs
of the model.

Beck etal. (2001)

-------
Table 4-17 (cont'd). Summary of Models of Human Exposure that Predict Tissue Distribution of Lead
to
Model
U.S. Environmental
Protection Agency ALM
Maddaloni et al. (2005)
California Environmental
Protection Agency,
Carlisle and Wade
(1992)
Bowers et al. (1994)
Stern (1994, 1996)
Age
Range
Adult
Child
Adult

Adult
Child
Adult
Exposure Pathways
Soil (supports
other pathways)
Air
Diet
Soil/dust
Water
Air
Soil/dust
Water
Dust/soil
Exposure Time Step
>3 months
(quasi-steady state)
>3 months
(quasi-steady state)

<3 months
(quasi-steady state)
>3 months
(quasi-steady state)
Biokinetics
Simulation
Uptake slope factor
Intake slope factor

Uptake slope factor
Intake slope factor
Biomarkers
Predicted
Blood
Blood

Blood
Blood
Variability and
Uncertainty Simulation
Variability: blood PbGSD,
Variability: blood PbGSD,

Variability: blood PbGSD
Variability: blood PbGSD,:
MCA
-------
blood Pb concentration when similar inputs and parameter values were applied to each model
(Maddaloni et al., 2005). Of the models presented in Table 4-17, Bowers et al. (1994) and
U.S. EPA (2003c) implement uptake slope factors. The slope factors used in both models
(-0.4 |ig/dL per jig Pb/day) are similar to biokinetic slope factors predicted from the O'Flaherty
model (0.65 |ig/dL per jig Pb uptake/day) and Leggett model (0.43 |ig/dL per jig Pb uptake/day)
for simulations of adult exposures (Maddaloni et al., 2005). A review of reported intake slope
factors relating medium-specific exposures and blood Pb concentrations derived from
epidemiologic studies can be found in the 1986 AQCD and in Abadin and Wheeler (1997).
Lead uptake-blood Pb concentration relationships in children, predicted by the IEUBK,
Leggett, and O'Flaherty models are shown in Figure 4-32. In the range of uptakes shown (0.1 to
100 jig Pb absorbed/day), nonlinearity of the relationship is apparent in both the Leggett and the
O'Flaherty models simulations. This reflects assumptions in each model regarding the limited
capacity of red blood cells to take up Pb, which has also been observed in humans (Bergdahl
et al., 1997a, 1998, 1999; Manton et al., 2001; Smith et al., 2002; see Section 4.3.1 for further
discussion of curvilinear relationship between Pb intake and blood Pb concentration).
Regression slopes (|ig/dL blood per |ig/day uptake) for the predictions < 10 |ig/dL are: Leggett
model, 0.88; IEUBK model, 0.36; O'Flaherty model, 0.29. The models predict an average blood
Pb concentration of 10 |ig/dL for the age range 2 to 3 years, in association with average Pb
uptakes (jig/day) for the same period of approximately: Leggett model, 12; IEUBK model, 29;
O'Flaherty model, 36.
A similar comparison of uptake-blood Pb concentration relationships predicted in adults is
shown in Figure 4-33. Regression slopes for adults predicted by the Leggett and O'Flaherty
models (at blood Pb concentrations < 10 |ig/dL) are more similar for adults (Leggett model, 0.54;
O'Flaherty model, 0.72) than for children (see Figure 4-32 versus Figure 4-33). The models
predict an average blood Pb concentration of 10 |ig/dL for the age range 31 to 32 years, in
association with average Pb uptakes, for the same period, of-18 and 13 |ig/day, Leggett and
O'Flaherty models, respectively. The nonlinearity in both children and adults is due largely to
assumptions made in the models about the limited capacity of red blood cells to take up Pb at
concentrations above 15 to 20 |ig/dL. The IEUBK model (for children) does not include this
nonlinearity feature.
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60
50-
40
35-
Leggett - Child (2-3 yr)
O'Flaherty - Child (2-3 yr)
. IEUBK - Child (2-3 yr)
20
40 60

Lead Uptake (pg/day)
80
100
Leggett - Child (2-3 yr)
O'Flaherty - Child (2-3 yr)
. IEUBK - Child (2-3 yr)
10
Lead Uptake (jjg/day)
Figure 4-32. Model comparison of predicted lead uptake—blood lead concentration
relationship in children. In the range of uptakes shown, the nonlinearity
of the relationship is apparent in the Leggett and O'Flaherty Models
simulations, reflecting the simulation of the limited capacity of red blood cells
to take up lead. Regression slopes (ug/dL blood per ug/day uptake) for the
predictions <;10ug/dL are: Leggett Model, 0.88; IEUBK Model, 0.36;
O'Flaherty Model, 0.29.
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60
50
Leggett - Adult (30-31 yr)
O'Flaherty - Adult (30-31 yr)
20
40 60
Lead Uptake (pg/day)
Figure 4-33. Model comparison of predicted lead uptake—blood lead concentration
relationships in adults. The nonlinearity of the relationship is apparent in
both the Leggett and O'Flaherty Models. Regression slopes (ug/dL blood per
jig/day uptake) for the predictions ^lOug/dL are: Leggett Model, 0.54;
O'Flaherty Model, 0.72.
Comparisons of predicted bone and soft tissue Pb burdens are shown in Figure 4-34.
Leggett and O'Flaherty models predict bone Pb burdens. Both the Leggett and O'Flaherty
models predict a bone Pb burden in adults of-90 and 98% of total body burden, respectively.
Regression slopes (mg Pb in bone per jig uptake/day) are 1.2 for the Leggett model and 2.1 for
the O'Flaherty model.
Figures 4-35 and 4-36 compare model predictions for blood Pb concentration for
hypothetical childhood or adult Pb exposures. The hypothetical child (Figure 4-35) has a
blood Pb concentration of 2 |ig/dL at age 2 years and then experiences a 1-year exposure to
100 jig Pb/day. All three models (Leggett, IEUBK, and O'Flaherty) predict a similar temporal
pattern of increase in blood Pb concentration at the start of exposure, then attainment of a quasi-
steady state, followed by a decrease in blood Pb concentration, with fast and slower phases of the
4-124
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240
200^
2> 160
TJ
TO
0)

0)
c
o
CO
120-
80-
O>
E
(0

-------
Base
Leggett Model
IEUBK Model
O'Flaherty Model
Age (year)

Figure 4-35. Comparison of model predictions for childhood lead exposure. The
simulations are of a hypothetical child who has a blood lead concentration
of 2 ug/dL at age 2 years, and then experiences a 1-year exposure to 100 ug
lead/day. Default bioavailability assumptions were applied in all three
models.
30 31 32
Age (year)
33
34
35
Figure 4-36. Comparison of model predictions for adult lead exposure. The simulations
are a hypothetical adult who has a blood lead concentration of 2 ug/dL at age
30 years and then experiences a 1-year exposure to 100 ug lead/day. Default
bioavailability assumptions were applied in the Leggett and O'Flaherty
models.
4-126
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decline in blood Pb concentration after the exposure ceases. However, differences in the
predicted kinetics of the blood Pb changes and the predicted quasi-steady state blood Pb
concentrations are evident. For this hypothetical scenario, the Leggett model predicts the highest
blood Pb concentrations (23 |ig/dL) compared to the O'Flaherty (12 |ig/dL) and IEUBK
(10 |ig/dL) models. These differences are not solely the result of different values for the
absorption fraction in 2 to 3 year old children: Leggett model, 30%; O'Flaherty model, 45%
(descending from 49% at age 2 years to 39% at age 3 years); IEUBK model, 25% (at a soil Pb
intake of 100 jig/day). A similar pattern is evident in the simulation of the same exposure
(100 jig/day for 1 year) in an adult (age 30 years; Figure 4-36). The Leggett model predicts
a quasi-steady state blood Pb concentration of-8.2 |ig/dL and the O'Flaherty model predicts
5.4 |ig/dL. However, most of this difference can be attributed to the different absorption fraction
values used for adults in the two models; 15% in the Leggett model and 8% in the O'Flaherty.
A comparison of predictions of quasi-steady state blood Pb concentrations from various models
was reported in Maddaloni et al. (2005). Results of comparisons between the U.S. ALM,
Leggett model and O'Flaherty model are presented in Table 4-18. When similar exposure inputs
are used in the three models, similar blood Pb concentrations are predicted. Much of the
difference between the Leggett model and ALM predictions can be ascribed to differences in
assumed bioavailability of Pb in soil: 12% in the U.S. EPA ALM, and 15% in the Leggett
model.

4.4.10 Conclusions and Future Directions
Modeling of relationships between Pb exposures and Pb levels in tissues has advanced
considerably during the past 25 years or so. Three mechanistic models have been developed and
evaluated to varying degrees for predicting associations between exposure and body burden
(IEUBK model, Leggett model, O'Flaherty model). A fourth model, the All Ages Lead Model,
is still under development and may resolve some of the issues regarding discrepancies between
other models, while at the same time adding new features directly applicable to risk assessment.
The IEUBK model has had the most extensive application in the regulatory context,
as EPA guidance recommends that, where possible, risk estimates for residential exposures to Pb
at hazardous waste sites be based on IEUBK model predictions of blood Pb concentrations
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to
oo
Table 4-18. Inputs and Results of Simulations Comparing the U.S. EPA Adult Lead Methodology (ALM)
With Multicompartmental Models
Parameters
Soil Pb concentration
IRs
AF
Baseline blood Pb
Exposure frequency
Exposure duration
Predicted quasi-steady state PbB
SoilPb: 1000 ug/g
SoilPb: 10,000 ug/g
ALM
1000 ug/g
0.05 g/day
0.12
2ug/dL
5 days/week
(260 days/year)1
NA
(jug/dL)
3.7
19
Rabinowitz model
1000 ug/ga'b
0.05 g/day a
0.12a
2 ug/dLe
5 days/week
(260 days/year)
365 days

4.1
23
Bert model
1000 ug/g ^
0.05 g/day a
0.12(0.08)d
2 ug/dLf
5 days/week
(260 days/year)
365 days

4.0 (2.6)k
21 (14)k
Leggett model
1000 ug/ga'c
0.05 g/day a
0.12(0.15)d
2 ug/dL8
5 days/week
(260 days/year)
17-45 years

4.1 (4.9)k
23 (29)k
O'Flaherty model
1000 ug/ga'c
0.05 g/day a
0.08
2 ug/dLh
5 days/week
(260 days/year)
17-45 years'

4.6
24
aNot a parameter in the model.
bSimulatedas an increment in daily uptake of 6 jig/day (i.e., 1000 x 0.05 x 0.12) above baseline.
°Simulated as an increment in daily intake of 50 ng/day (i.e., 1000 x 0.05) above baseline.
dDefault values are shown in parenthesis.
eA daily uptake of 4.1 ng/day yielded a quasi-steady state PbB of 2 |ig/dL.
fA daily intake of 38.7 jig/day yielded a quasi-steady state PbB of 2 |ig/dL.
gA daily intake that varied from 12 to 25 jig/day yielded pre-adult PbBs within the ranges reported from Phase I of NHANES III (Brody et al. 1994) and an adult PbB
of 2 |ig/dL.
hSetting all Pb concentrations and intakes to null and food Pb ingestion by adults born in 1980 to 25 ng/day yielded pre-adult PbBs within the ranges reported from
Phase I of NHANES III (Brody et al. 1994) and an adult PbB of 2 |ig/dL.
'The default exposure frequency for the ALM is 219 days/year; however, the assumption of 260 days/year in the simulations would not change the outcome of the
model comparisons.
JAdults born in 1980.
Predictions based on the default value for the AF are shown in parenthesis.
-------
in children. Although, these models are constructed very differently (e.g., the O'Flaherty
biokinetics model has only 17 Pb parameters, compared to 65 in the Leggett biokinetics model,
and 47 in the IEUBK biokinetics model), the three models yield remarkably similar predictions
of blood Pb concentration for similar hypothetical exposure scenarios. The three models predict
similar kinetics of change in blood Pb concentrations in association with a change in Pb exposure
(e.g., Figures 4-35 and 4-36). Both the Leggett and O'Flaherty models predict similar rates of Pb
accumulation in bone, for the same rates of uptake of Pb into the body. Predictions of quasi-
steady state blood Pb concentrations for the scenarios are simulated in Figures 4-35 and 4-36 and
differ across models by a factor of ~2. This magnitude of difference is substantial in the context
of certain regulatory uses of the models (e.g., for establishing cleanup goals at hazardous waste
sites); however, it is not surprising, given the various approaches taken to reduce the complex
biokinetics of Pb to tractable, and relatively simple, mathematical expressions.
Several major challenges remain to be confronted in further developing our ability to
simulate Pb exposure-tissue level relationships in real individuals or populations. The three
earlier mechanistic models described above do not simulate the kinetics of Pb in pregnancy or in
senescence (e.g., menopause). Only one of these three earlier models (Leggett) simulates Pb
levels in brain, a potential target organ for Pb toxicity. None of the models have been rigorously
evaluated for accuracy of predictions of bone Pb levels in humans, for which there is a rapidly
expanding set of observations of importance to dose-response assessment. Of great importance
for regulatory uses of the models, for example, is the need for more rigorous quantitative
assessment of confidence (i.e., uncertainty) in model predictions. To date, such assessments
have not been applied uniformly in a manner that allows cross-model comparisons of confidence
for specific regulatory uses.
The IEUBK Model has undergone the most extensive and thoroughly reported evaluation
of a regulatory use of the model, i.e., (a) quantitative evaluation of predicted distributions of
blood Pb concentrations in children who live in areas for which cross-sectional measurements of
environmental Pb levels were available and (b) independent verification of the IEUBK model
implementation code (Hogan et al., 1998; Zaragoza and Hogan, 1998). However, a similar level
of evaluation of the Leggett and O'Flaherty models has not been reported, although specific
predictions of the models have been evaluated against observations (e.g., experimentally-
observed kinetics of change in blood Pb following a change in intakes).
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To a large extent, the important information gap regarding evaluation of model confidence
derives from a lack of observational data and/or public access to observational data on which
predictions could be evaluated. An additional challenge for applications of the models in a
regulatory context relates to uncertainties in exposure data from which exposure model inputs
are derived. Model development and uncertainty assessment could be substantively advanced by
assembling verified (for accuracy) sets of data on Pb biokinetics against which models could be
uniformly evaluated. Examples of the types of data that would be valuable include data on the
kinetics of change in blood or tissue Pb concentrations, or stable Pb isotope ratios, in response to
a change in exposure. Also, access to large data bases that include reported Pb exposure
measurements for various media that are paired with blood or tissue Pb measurements for
individuals affected by pertinent exposure scenarios would also be extremely valuable for cross-
model evaluations.
4.5 SUMMARY
At the time of the 1986 Lead AQCD, it was recognized that Pb distributed to and
accumulated in several bone compartments which exhibited differing mobility profiles. It was
also recognized that a larger fraction of total body burden of Pb is found in the bones of adults
relative to children. The possibility of bone Pb serving as a source of long-term internal
exposure was considered. Several models of Pb pharmacokinetics of Pb in humans were
developed to simulate the multiphasic elimination kinetics of Pb from blood, bone, and soft
tissues (Marcus, 1985a,b,c; Rabinowitz et al., 1976). New studies have been published on the
kinetics of Pb movement into and out of bone which demonstrate the importance of bone Pb
stores as a source of Pb to the blood in retired Pb workers and during pregnancy. Additional
information regarding Pb absorption, distribution, and elimination in humans is available.
The main pathway of intake of Pb in the general human population is ingestion (diet,
ingestion of surface dust). Inhalation can be an important pathway in some occupational
settings. Dermal absorption of inorganic Pb compounds is generally considered to be much less
than absorption by inhalation or oral routes of exposure; however, few studies have provided
quantitative estimates of dermal absorption of inorganic Pb in humans, and the quantitative
significance of the dermal absorption pathway as a contributor to Pb body burden in humans
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remains an uncertainty. Absorption of ingested Pb is affected by numerous factors, including the
individual's age, diet, and nutritional status, as well as chemical and physical properties of Pb.
Lead absorption appears to be increased by both iron and calcium deficiency. Fasting also
increases the absorption of ingested Pb (Blake et al., 1983; Heard and Chamberlain, 1983; James
et al., 1985; Rabinowitz et al., 1980; Maddaloni et al., 1998). Gastrointestinal (GI) absorption of
Pb in humans may be a capacity limited process such that the fraction of ingested Pb that is
absorbed may decrease with increasing rate of Pb intake. The available studies to date, however,
do not provide a firm basis for discerning if the GI absorption of Pb in humans is limited by
dose. The size of ingested Pb particles also affects GI absorption, with absorption decreasing
with increasing particle size (Ruby et al., 1999; Healy et al., 1992). Exposures to airborne
inorganic Pb are usually in the form of particulate aerosols. Deposition and clearance of Pb
particles from the respiratory tract are affected by numerous factors, including the individual's
age and activity, particle size, and physical-chemical properties of the inhaled Pb.
In general, the burden of Pb in the body may be viewed as divided between a dominant
slow compartment (bone) and a smaller fast compartment (soft tissues). In human adults, more
than 90% of the total body burden of Pb is found in the bones, whereas bone Pb accounts for
-70% of the body burden in children (Barry, 1975). The highest soft tissue concentrations in
adults also occur in liver and kidney cortex (Barry, 1975; Gerhardsson et al., 1986, 1995b; Gross
et al., 1975; Oldereid et al., 1993). Lead in blood (i.e., plasma) exchanges with both of these
compartments. The contribution of bone Pb to blood changes with the duration and intensity of
the exposure, age, and various physiological variables (e.g., nutritional status, pregnancy, and
menopause).
Lead accumulates in bone regions having the most active calcification at the time of
exposure. Lead accumulation is thought to occur predominantly in trabecular bone during
childhood and in both cortical and trabecular bone in adulthood (Aufderheide and Wittmers,
1992). Lead concentrations in bone increase with age throughout the lifetime, indicative of a
relatively slow turnover of Pb in adult bone (Barry 1975, 1981; Gross et al., 1975; Schroeder and
Tipton, 1968). Some bones (i.e., mid femur and pelvic bone) increase in Pb content plateaus at
middle age and then decreases at higher ages (Drasch et al., 1987). This decrease is most
pronounced in females and may be due to osteoporosis and release of Pb from resorbed bone to
blood (Gulson et al., 2002). Lead in adult bone can serve to maintain blood Pb levels long after
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external exposure has ceased (Fleming et al., 1997; Inskip et al., 1996; Kehoe, 1987; O'Flaherty
et al., 1982; Smith et al., 1996). During pregnancy, bone Pb can also serve as a Pb source as
maternal bone is resorbed for the production of the fetal skeleton (Franklin et al., 1997; Gulson
etal., 1997, 1999b, 2003).
The highest soft tissue concentrations in adults also occur in liver and kidney cortex
(Barry, 1975; Gerhardsson et al., 1986, 1995b; Gross et al., 1975; Oldereid et al., 1993).
In contrast to Pb in bone, which accumulates with continued exposure in adulthood, Pb
concentrations in soft tissues (e.g., liver and kidney) are relatively constant in adults (Barry
1975; Treble and Thompson 1997), reflecting a faster turnover of Pb in soft tissue relative to
bone. Pb in soft tissues exists predominantly bound to protein. High affinity cytosolic Pb
binding proteins (PbBPs) have been identified in rat kidney and brain (DuVal and Fowler 1989;
Fowler 1989). Other high-affinity Pb binding proteins (Kd -14 nM) have been isolated in
human kidney, two of which have been identified as a 5 kD peptide, thymosin 4 and a 9 kD
peptide, acyl-CoA binding protein (Smith et al., 1998a).
Lead in blood is found primarily (-99%) in the red blood cells (Bergdahl et al., 1997b,
1998, 1999; Hernandez-Avila et al., 1998; Manton et al., 2001; Schutz et al., 1996; Smith et al.,
2002). 5-aminolevulinic acid dehydratase (ALAD) is the primary binding ligand for Pb in
erythrocytes (Bergdahl et al., 1997b, 1998; Sakai et al., 1982; Xie et al., 1998). Lead binding to
ALAD is saturable; the binding capacity has been estimated to be -850 |ig/dL red blood cells
(or -40 |ig/dL whole blood) and the apparent dissociation constant has been estimated to be
-1.5 |ig/L (Bergdahl et al., 1998). It has been suggested that the small fraction of Pb in plasma
(<0.3%) may be the more biologically labile and lexicologically active fraction of the circulating
Pb. Several authors have proposed that Pb released from the skeleton was preferentially
partitioned into serum compared with red cells (Cake et al., 1996; Hernandez-Avila et al., 1998;
Tsaih et al., 1999). Approximately 40 to 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). Lead
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).
Blood Pb concentration is extensively used in epidemiologic studies as an index of
exposure and body burden due mainly the feasibility of incorporating its measurement into
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human studies relative to other potential dose indicators, e.g., Pb in kidney, plasma, urine, or
bone. A single blood Pb measurement may not distinguish between a history of long-term lower
level Pb exposure from a history that includes higher acute exposures (Mushak, 1998).
An additional complication is that the relationship between Pb intake and blood Pb concentration
is curvilinear; that is, the increment in blood Pb concentration per unit of Pb intake decreases
with increasing blood Pb concentration, both in children (Lacey et al., 1985; Ryu et al., 1983;
Sherlock and Quinn, 1986) and in adults (Kehoe, 1987; Laxen et al., 1987; Pocock et al., 1983;
Sherlock et al., 1982, 1984). In general, higher blood Pb concentrations can be interpreted as
indicating higher exposures (or Pb uptakes); however, they do not necessarily predict higher
body burdens. Similar blood Pb concentrations in two individuals (or populations) do not
necessarily translate to similar body burdens or similar exposure histories. The disparity in the
kinetics of blood Pb and body burden may have important implications for the interpretation of
blood Pb concentration measurements in some epidemiology studies, depending on the health
outcome being evaluated.
In addition to blood Pb, hair and urine Pb have also been used as biomarkers of Pb
exposure. An empirical basis for interpreting hair Pb measurements in terms of body burden or
exposure has not been firmly established. Hair Pb measurements are subject to error from
contamination of the surface with environmental Pb and contaminants in artificial hair treatments
(i.e., dyeing, bleaching, permanents) and, as such, are relatively poor predictor of blood Pb
concentration, particularly at low levels (<10 to 12 |ig/dL). Urine Pb concentration
measurements provide little reliable information, unless they can be adjusted to account for
unmeasured variability in urine flow rate (Araki et al., 1990). Similar to blood Pb concentration
measurements, urinary Pb excretion measured in an individual at a single point in time will
reflect the recent exposure history. As a result, measurement of urinary Pb may serve as a
feasible surrogate for plasma Pb concentration, and may be useful for exploring dose-response
relationships for effect outcomes that may be more strongly associated with plasma Pb
concentration than Pb body burden.
Several new studies have investigated the relationship between Pb exposure and blood Pb
in children. These studies support the concept that contact with Pb in surface dust (interior and
exterior) are a major contributor to Pb intake in children. In the meta-analysis by Succop et al.
(1998), the most common exposure pathway influencing blood Pb concentration was exterior
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soil, operating through its effect on interior dust Pb and hand Pb. Using a structural equation
model, Lanphear and Roghmann (1997) also found the exposure pathway most influential on
blood Pb was interior dust Pb loading, directly or through its influence on hand Pb. Both soil
and paint Pb influenced interior dust Pb; with the influence of paint Pb greater than that of soil
Pb. Interior dust Pb loading also showed the strongest influence on blood Pb in a pooled
multivariate regression analysis (Lanphear et al., 1998).
New information on Pb biokinetics, bone mineral metabolism, and Pb exposures has led
to refinements and expansions of earlier mechanistic models of Pb biokinetics. In particular,
three pharmacokinetic models are currently being used or considered for broad application in Pb
risk assessment: (1) the Integrated Exposure Uptake BioKinetic (IEUBK) model for Pb in
children developed by EPA (U.S. Environmental Protection Agency, 1994a,b; White et al.,
1998); (2) the Leggett model, which simulates Pb kinetics from birth through adulthood
(Leggett, 1993); and (3) the O'Flaherty model, which simulates Pb kinetics from birth through
adulthood (O'Flaherty, 1993, 1995). The above three models have been individually evaluated,
to varying degrees, against empirical physiological data on animals and humans and data on
blood Pb concentrations in individuals and/or populations (U.S. Environmental Protection
Agency, 1994a,b; Leggett, 1993; O'Flaherty, 1993). In evaluating models for use in risk
assessment, exposure data collected at hazardous waste sites, where exposures to contaminated
soils are the dominant contributors to exposure, have been used to drive model simulations of
corresponding blood Pb distributions (Bowers and Mattuck, 2001; Hogan et al., 1998;
TerraGraphics, Inc, 2000, 2001). The exposure module in the IEUBK model makes this type of
evaluation feasible. These empirical comparisons have shown that agreement or disparity
between IEUBK model predictions and observed blood Pb concentrations at specific locations is
influenced by numerous factors, including (a) the extent to which the exposure and blood Pb
measurements are adequately matched and (b) site-specific factors (e.g., behavior patterns, soil
characteristics, bioavailability) that may affect Pb intake or uptake in children. Exposure-
biokinetics models serve not only to illustrate exposure-blood-body burden relationships,
but also provide a means for making predictions about them that can be tested experimentally
or in epidemiologic studies.
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5. TOXICOLOGICAL EFFECTS OF LEAD IN
LABORATORY ANIMALS AND IN VITRO
TEST SYSTEMS

5.1 INTRODUCTION
As noted in Chapter 1, U.S. EPA air quality criteria documents evaluate scientific
knowledge of relationships between pollutant concentrations and their effects on the
environment and public health. Chapters 2 and 3 of this document discussed the chemistry and
physical properties of lead (Pb); sources, emissions, transport, and deposition of Pb; and
environmental concentrations and pathways to human exposure. Chapter 4 discussed biokinetics
of external Pb exposure impacts on internal distribution of lead to body tissues and models of
human exposure that predict tissue distribution of Pb. This chapter (Chapter 5) assesses
information regarding the toxicological effects of Pb in laboratory animals and in vitro test
systems. Emphasis is not only placed here on qualitative characterization of various Pb-induced
effects, but also on attempts to define dose-effect relationships for key health effects thought
likely to occur at ambient exposure levels encountered by the general population of the United
States. Chapter 6 follows with a discussion of epidemiologic studies of ambient Pb-exposure
effects. The environmental effects of Pb are then discussed in Chapter 7. Lastly, Chapter 8
provides an overall integrative synthesis of information on Pb exposures, health effects and their
potential public health significance, and environmental (especially ecologic) effects of Pb.
The framework used here for assessing the toxicologic effects of Pb is subdivided mainly
according to organ systems. As noted in the last previous Lead Air Quality Criteria Document
(Lead AQCD) published in 1986, this facilitates presentation of the information, but it must also
be stressed that all systems are interdependent, functioning in delicate concert to preserve the
physiological integrity of the whole organism.
The information discussed in this chapter is derived from a very wide body of literature on
studies in laboratory animals and in vitro test systems of animal cell lines and organ systems that
may mimic responses in intact animals. This chapter is not intended to be a compendium of all
that is known about Pb effects; rather, it is an update assessment of the reported biological
effects from the 1986 Lead AQCD (U.S. Environmental Protection Agency, 1986a), the
Addendum to that document (Lead Effects on Cardiovascular Function, Early Development, and
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Stature) (U.S. Environmental Protection Agency, 1986b), and the Supplement to the 1986
Addendum (U.S. Environmental Protection Agency, 1990). The historical Pb literature is briefly
summarized at the opening of each section or subsection and is intended as a very concise
overview of previous work. The reader should refer to the previous documents listed above for
more detailed discussion of the literature prior to the late 1980s. Each section then continues
with brief, evaluative discussion of key studies published since 1986. Longer discussions of
newly available studies are included where warranted. Sections also include comparisons of
findings from the 1986 AQCD to those derived from the new data and statements of key
bottomline conclusions regarding Pb effects on given types of health endpoints. More detailed
descriptive summaries of newly available studies and results are provided in Annex AX5.
5.2 EFFECTS OF LEAD ON HEME SYNTHESIS
5.2.1 Effects of Lead on Erythrocyte Biology and Function
Lead poisoning is one of the most common acquired environmental diseases, because of
physical properties of the metal and its widespread environmental distribution. It is a complex
disorder affecting several organs in the body, including developing erythrocytes (red blood cells
[RBCs]). Anemia is frequently observed with Pb poisoning and is thought to result from the
shortening of erythrocyte life span and the effects of Pb on hemoglobin synthesis. However, the
exact mechanisms by which Pb affects the red blood cell (RBC) life span and heme synthesis are
not clear. It is postulated that the mechanisms may be due to the effects of Pb on iron uptake and
several other interactions of Pb and iron-mediated cellular processes. It has been demonstrated
for well over three decades now that conditions of iron deficiency and Pb poisoning can
independently occur and coexist. Both conditions are capable of independently producing
microcytosis and anemia. In erythrocytes, they affect several cellular processes, and their
combined effect on hemotology is an interesting area of study. The increase in Pb absorption in
iron deficient rats was first demonstrated in 1972 with several more studies following later,
including one by Wright et al. (1998). A common iron-Pb transporter, DMTI, has been
postulated that could lead to potential competition for binding sites that could affect intestinal
absorption; and DTMI has a higher afficity for Pb. It has also been observed that anemia
accompanying the combination of both conditions is more severe and more hypochromic than
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uncomplicated iron deficiency anemia. When heme synthesis is inhibited at the final step in
heme synthesis, the net effect is that zinc (instead of iron) is incorporated into protoporphyrin,
resulting in elevated levels of zinc protoporphyin (ZPP). Both iron deficiency and Pb poisoning
are capable of inhibiting heme synthesis at this final step, and combined conditions lead to
dramatic ZPP elevations. Lead poisoning also causes increased urinary excretion of porphyrins
and 5-aminolevulinic acid (ALA), the first precursor for heme synthesis. Additional evidence
for striking similarities between Pb poisoning and acute intermittent porphyria (the disease
associated with lesions in the heme biosynthetic enzyme, porphobilinogen deaminase) strongly
suggests that one major site of Pb intoxication is the heme biosynthetic pathway.
The 1986 Lead AQCD assessed literature available at that time from both animal and
human studies indicating potential effects of Pb intoxication on enzymes and precursors involved
in heme synthesis, erythrocyte morphology and function, as well as the influence of these
perturbations on the nervous system and vitamin D metabolism and associated physiological
process. In summary, those studies reported an association between increased Pb exposure and
increased ALAS activity (which is increased in kidney with acute exposure and in spleen with
chronic exposure, while it decreased in liver tissue in both the exposure scenarios). The activity
of ALAD appeared to be inversely correlated to blood Pb values and was found to be inhibited in
several tissues. It was also inferred from several animal studies that the effect of Pb on heme
formation involved both ferrochelatase inhibition and impaired mitochondrial transport of iron.
Human studies indicated that occupational exposure to Pb results in decreased erythrocyte cell
survival and alterations in erythrocyte membrane integrity and energetics. The vast scientific
literature on the effects of Pb on various aspects of heme metabolism in diverse organ systems,
both in humans and animals, has expanded over the past two decades. This chapter is primarily
concerned with discussions of data from animal and in vitro studies, while the human studies are
dealt with in Chapter 6.

5.2.2 Effects of Lead on Erythrocyte Functions
The cellular membrane is one of the main targets for toxic effects of heavy metals such as
Pb. Anemia, one of the clinical symptoms of severe Pb intoxication, can develop because of
impairment of hemoglobin synthesis and damage of erythrocyte membranes by Pb ions.
Although the erythrocyte membrane is not as specialized as other cell membranes are, it carries
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out important functions common to other cell membranes, such as active and passive transport
and the production of ionic and electric gradients. Changes in erythrocyte membrane lipid and
protein profiles can alter the membrane fluidity, potentially affecting enzymatic activity and the
functionality of receptors and ion channels present on the plasma membrane and can also
influence the ionic and molecular composition of intracellular spaces.

Lead Uptake, Binding, and Transport
Studies by Simons (1986a) indicated that the uptake of Pb into human RBCs is a passive
process, i.e., it does not require the use of energy in the form of ATP. In addition, Pb may be
able to cross the membrane passively in either direction. This process involves anion transport
mechanisms, as characteristic anion exchange inhibitors have been found to inhibit passive
uptake of Pb by RBCs (Simons, 1986a,b). It has also been shown that the transport of Pb across
the membrane depends on the presence of another anion, the bicarbonate ion, and Pb is
transported as Pb-carbonate (Simons, 1986a). When Pb enters the cell, it binds mainly to
hemoglobin, and the ratio of bound to free Pb in cytoplasm has been estimated to be 6000:1.
Simons (1986a,b) carried out studies using citrate buffers, which may cause hemolysis of RBCs.
To avoid the influence of a citrate buffer, Sugawara et al. (1990) measured the uptake of Pb into
human RBCs by adding Pb directly into plasma. These investigators also found that the
transport of Pb across the erythrocyte membrane is energy-independent (passive) and carrier
mediated. Little release of Pb from the cells was seen, suggesting absence of any hemolysis of
the cells in this protocol. Furthermore, the progressive accumulation of Pb was not observed.
More than 98% of the Pb was found accumulated in the cytoplasm in protein-bound form,
whereas only 2% was found in the membrane fraction. Sugawara et al. (1990) also reported
finding 45 Pb-binding sites on human hemoglobin. On the other hand, studies reported by
Bergdahl et al. (1997), using liquid chromatography coupled with inductive plasma mass
spectrometry analysis, suggested aminolevulinic acid dehydratase (ALAD), the enzyme involved
in the heme synthesis pathway, to be the principle Pb-binding protein, not hemoglobin, as earlier
thought.
Additional studies carried out by Simons (1993a) evaluated the transport of Pb into RBCs
for cell Pb contents in the range of 1 to 10 jiM and showed that 203Pb uptake was mediated by an
anion exchanger and that the efflux was mediated through a vanadate-sensitive pathway
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identified with the calcium pump (Simons, 1988). He further concluded that the high ratio of
RBC to plasma Pb observed in vivo was due to a labile Pb-binding component within the
cytoplasm. Simons (1993a) also observed that exit of Pb ions from the RBC was much lower
than expected based on his earlier work with erythrocyte ghosts. Utilizing a group of drugs that
modify anion exchange and thiol groups in the cytoplasm, Lai et al. (1996) showed that anion
exchange mechanisms and thiol groups were critical factors in how Pb stimulates calcium-
dependent processes in erythrocytes. Once the role of anion exchanger proteins had been
implicated in Pb transport in erythrocytes, Bannon et al. (2000) investigated whether similar
anion exchange processes are involved in the uptake and transport of Pb in other cells, such as
Madin-Darby canine kidney epithelial cells. Based on a comparative in vitro study using human
erythrocytes and canine kidney epithelial cells, these authors reported transport of Pb in kidney
epithelial cells, suggesting similar anion exchange involvement.

Erythrocyte Survival, Mobility, and Membrane Integrity
It is well recognized that Pb intoxication interferes with RBC survival by shortening the
life span and altering the mobility of erythrocytes; however, the molecular mechanisms behind
these effects of Pb on erythrocyte functions are not well understood. The shape and
deformability of the human erythrocyte, or RBC, is maintained by several factors, including low
concentration of free intracellular Ca2+ (<0.1 jiM) and a replenished ATP level. An elevated
interfacial Ca2+ concentration inside the RBC activates the passive ion efflux via a K+ selective
(voltage independent) channel and a concomitant water transport (Gordos effect). Low
concentrations of Pb ions can mimic Ca2+ and activate the same channel in the RBC.
Intraperitonially injected Pb significantly decreases rat erythrocyte membrane mobility
(Terayama et al., 1986), an effect evident to some extent even below blood Pb concentrations of
100 jig/100 mL. This decrease in rat erythrocyte mobility was found simultaneous or prior to
changes in hematological parameters such as hemoglobin (Hb) levels and hematocrits (Hct).
The same group (Terayama and Muratsugu, 1988) also reported a significant decrease in
erythrocyte membrane sialic acid content at the same blood Pb levels with exposure to Pb
(20 mM Pb acetate once a week for 5 weeks). Additional studies by the same group reported
that other hematological parameters, such as mean corpuscular volume (MCV), mean
corpuscular hemoglobin (MCH), and mean corpuscular hemoglobin concentration (MCHC),
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were also significantly decreased upon Pb exposure, along with decreased mobility, sialic acid
content, and deformability of rat RBCs. However, the blood Pb levels reported in these studies
range from 100 to 800 |ig/dL and, at best, point out newer mechanistic details of erythrocyte
membrane alterations affecting their survival and mobility. Still, it should be noted that these
changes were seen to a minor extent even at blood Pb levels <100 |ig/dL. It was speculated that
Pb-induced decreases in sialic acid content and deformability of RBCs shorten RBC survival
time and may lead to anemia in Pb poisoning. Jehan and Motlag (1995) reported that Pb
exposure caused significant change in RBC membrane cholesterol and phospholipid contents
along with sialic acid. Coexposure to Zn was found to reduce these alterations.
Lead-induced morphological changes in human RBCs were studied by Eriksson and
Bering (1993), using electron paramagnetic resonance imaging. These authors reported that Pb
ions (a) induced time-dependent changes in MCV and cell shrinkage and (b) inhibited the Gardos
effect. Trialkyl-Pb compounds have also been reported to induce hemolytic activity in
erythrocytes, with intensity increasing with hydrophobicity of the compounds (Kleszcyriska
et al., 1997). Serrani et al. (1997) reported that Pb ions confer protection against RBC lysis in
hypotonic low ionic strength media, presumably due to interaction of Pb with certain constituents
in the cell membrane. This resistance to erythrocyte lysis was found to significantly increase
with Pb (20 to 25 jiM) compared to other metals such as Al, Cd, and Zn (Corchs et al., 2001).
The Pb-induced reduction in MCV (RBCs derived from umbilical cord) was found to be reversed
when the cells were treated with quinidine, an inhibitor of a potassium channel activator, without
any effect on resistance to cell lysis, suggesting changes in cell membrane structure. This effect
may also be involved in membrane deformability (Mojzis and Nistiar, 2001).
Heavy metals, including Cd, Zn, and Pb, have been found to alter RBC membrane
microviscosity and fluidity (Amoruso et al., 1987). These authors labeled RBC membranes with
fluorescent lipid probe all trans 1, 6-diphenyl-l,3,5-hexatriene (DPH) and demonstrated
increased polarization with increased membrane lipid viscosity upon exposure to heavy metals.
They also postulated that such alterations in cell membrane lipid, and possibly also protein
fluidity, may contribute to abnormal cellular function. Similar changes in RBC fluidity were
observed in the RBC collected from workers exposed to Pb (Cook et al., 1987). The RBC ghost
membranes isolated from Pb-exposed workers exhibited a significant increase in the
phosphotidylcholine to phosphotidylethanolamine ratio (an established correlate of membrane
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fluidity) along with an increase in RBC cholesterol levels, as also reported by Jehan and Motlag
(1995) discussed above. These authors predict that such alterations in phospholipid composition
of the membrane are responsible for biochemical instability of RBC in Pb-exposed workers.
Zimmermann et al. (1993) investigated the potential of such membrane lipid alterations to cause
resistance to oxidation. These investigators induced hyperlipidemia by treating Pb-exposed
Wistar rats with triton and observed an increase in erythrocyte choline phospholipid levels,
together with a significant decrease in membrane lipid resistance to oxidation. They postulated
that such a decrease in resistance might cause RBC fragility and ultimate destruction, leading to
anemic conditions. It has been also reported that exposure to Pb may also increase the levels of
fatty acids, e.g., arachidonic acid, in the RBC membrane in humans exposed to Pb (Osterode and
Ulberth, 2000). Based on the negative correlation between serum calcium and increased
arachidonic acid content, these authors postulated that Pb ions might have substituted for calcium
in the activation of phospholipase enzymes, leading to increased synthesis of arachidonic acid.
The fact that these biochemical and molecular changes were reported at somewhat higher blood
Pb levels (70 |ig/dL) probably does not undermine these observations made from the RBCs of
humans exposed to Pb over a period of time and enhances our understanding of the several
molecular facets that may play a role in the altered erythrocyte mobility. Suwalsky et al. (2003)
investigated the interaction of Pb with the RBC membrane, utilizing intact as well as isolated
unsealed RBC membrane models (representing phospholipids present in the inner and outer
layers of the membrane). Electron microscopy, fluorescence spectroscopy, and X-ray diffraction
analyses of these models by Suwalsky et al. (2003) indicated that Pb particles adhere to both
external and internal surfaces of the membrane. Pb ions also have been found to disturb the
lamellar organization by causing considerable molecular disorder within lipid layers.
Recently, it has been shown that osmotic shock, oxidative stress, and/or energy depletion
activate Ca2+-sensitive erythrocyte scramblase, leading to the exposure of phosphotidylserine at
the cell surface. This exposure of phosphotidylserine had been implicated in the phagocytosis of
RBC by macrophages that can be measured by annexin binding, as determined by fluorescence
activated cell sorting analysis. Kempe et al. (2005) carried out experiments to investigate
whether anemic conditions reported in Pb intoxication are the result of the decreased life span of
RBCs due to the above-mentioned mechanisms. These investigators reported that when human
RBCs were exposed to Pb nitrate (above 0.3 |iM), it caused a significant increase in Pb annexin
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binding, indicative of phosphotidylserine exposure. Using inhibitors for Ca2+-sensitive
potassium channels and whole cell patch clamp experiments, they concluded that Pb exposure
increased activation of potassium channels, leading to shrinkage of cells and also activation of
scramblase, resulting in the exposure of phosphotidylserine on the cell membrane surface. These
authors further postulated that this exposure of phosphotidylserine on the membrane might have
led to them being engulfed by macrophages and the ultimately decreased life span of RBCs in Pb
intoxication.

Membrane Proteins
Earlier, Fukumoto et al. (1983) reported the differential profile for RBC-membrane
polypeptides determined by SDS-PAGE analysis. These investigators found decreased levels of
polypeptides in band 3 and increases in the levels of four other bands (i.e., bands 2, 4, 6, and 7)
in the RBCs of human workers exposed to Pb. From these observations, they postulated that
such Pb-induced alteration in RBC membrane proteins may lead to membrane permeability
changes. Apostoli et al. (1988) also observed similar changes in RBC membrane polypeptides in
Pb-exposed workers and suggested that band 3 may represent an anion channel protein; they also
found that these changes occurred at blood Pb levels >50 jig/100 mL.
Lead exposure has been known to increase the amount of membrane-bound protein
kinase C in rat brain, endothelial, and glial cells. Belloni-Olivi et al. (1996) reported an
increased phosphorylation of RBC membrane proteins on Pb exposure. When human RBCs
were incubated with Pb acetate (>100 nM) for 60 min, increased phosphorylation of membrane
cytoskeletal proteins (120, 80, 52 and 45 kDa) was found, accompanied by increased protein
kinase C activity. Membrane proteins were not phosphorylated when treated with protein kinase
C inhibitors. Calcium and diacylglycerol were found not to be involved in this process. The
authors suggested that this activation of protein kinase was a direct interaction of the enzyme
protein with Pb. Slobozhanina et al. (2005) reported that incubation of human RBCs with Pb
acetate (1 to 10 jiM for 3 h) caused differential binding of fluorescent probes to the membrane,
suggesting alterations in the physicochemical state of the membrane proteins and lipids. Based
on these observations, the authors postulated that such alterations in membrane molecular
composition may influence the activity of membrane enzymes and the functioning of receptors
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and channels present on the membrane. These and related studies are summarized in Annex
Table AX5-2.1.

5.2.3 Effects of Lead on Erythrocyte Heme Metabolism
Enzyme studies of the heme pathway have shown that Pb is an inhibitor of several
enzymes involved in heme synthesis, including 5-aminolevulinic acid dehydratase (ALAD),
coproporphyrinogen oxidase, and ferro chelatase (see Figure 5-1 for a schematic representation
of heme biosynthesis). ALAD is a cytoplasmic enzyme that catalyzes the second, rate-limiting
step of the heme biosynthesis pathway; that is, ALAD catalyzes formation of porphobilinogen
through the conjugation of two molecules of 5-aminolevulinic acid. ALAD is a Zn-dependent
enzyme, and thiol groups are essential for its activity (Bernard and Lauwerys, 1987). Decreased
erythrocyte ALAD is the most sensitive indicator of human Pb exposure, to the extent that
measurement of ALAD activity reflects well Pb levels in the blood. Similarly, erythrocyte
ALAD activity measurements have been used to assess Pb toxicity in other species.

Erythrocyte ALAD
Terayama et al. (1986) reported decreased ALAD activity in rat RBCs at blood Pb
levels of 10 |ig/dL. Scheuhammer (1987) studied the usefulness of the ALAD ratio
(activated/nonactivated enzyme activity) to study Pb effects in avian RBCs. The ALAD activity
ratio is a sensitive, dose-responsive measure of Pb exposure regardless of the mode of Pb
administration. For example, dietary Pb concentrations as low as 5 ppm (dry weight) can be
estimated through the use of the ALAD enzyme activity ratio method. A highly significant
positive correlation was observed between dietary Pb concentration over the 5 to 100 ppm range
and the ALAD activity ratio. The author concluded that RBC ALAD ratio may be a useful
method for estimating average dietary concentrations of Pb over an environmentally relevant
range, in situations where diet is the major source of exposure to Pb or where accurate
estimations of dietary Pb are not possible. Redig et al. (1991) reported heme synthetic pathway
alterations upon chronic exposure (3 or 11 weeks) to Pb in red-tailed hawks. This treatment
resulted in a severe decrease in RBC ALAD activity, which did not return to normal levels until
5 weeks after termination of Pb treatment. Lead exposure also decreased ALAD activity in the
bone marrow and in the liver but did not alter aminolevulinic acid synthase activity.
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Protoporpnynnogen Ix oxidase
Uroporphyrinogen
cosynthase
Uroporphyrinogen I
synthetase
Cytoplasm
Porphobilinogen
Uroporphyrinogen I
Uroporphyrinogen I
Coproporphyrinogen I
Coporphyrinogen \
Figure 5-1. Schematic presentation of the enzymatic steps involved in heme synthesis
pathway. Potential lead (Pb) interacting sites are indicated by curved arrows
(ft increased, .[(.decreased)
Source: Modified from U.S. EPA (1986a).
Dorward and Yagminas (1994), using comparative enzyme kinetic analysis of ALAD in
Pb-exposed female cynomolgus monkeys and human erythrocyte ALAD, found similar
inhibition profiles and concluded that ALAD could be a useful model for measuring the
biological response in monkeys. Santos et al. (1999) reported that rat RBC heme biosynthesis
was affected by either Pb treatment alone or Pb in combination with ethanol, due to the
inhibition of ALAD activity.
Analysis of blood ALAD activity had been used as a powerful clinical biomarker in
evaluating Pb toxicity in occupational exposure. Fontanellas et al. (2002) further suggested that
this enzyme assay be used in identifying even subclinical Pb poisoning in chronic renal failure
(see Section 5.7 for details).
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Other Heme Metabolism Enzymes
Taketani et al. (1985) studied the heme synthesizing activity of ferric ion using purified
ferrochelatase from rat liver mitochondria and reported that Pb reduced NAD(P)H-dependent
heme synthesis by 50% at ICf5 M, but that it had no effect when ferrous ion was used as the
substrate. Based on these results, the authors concluded that heme synthesis from ferric ion was
more susceptible to Pb than the ferrous ion. These studies also revealed that the NAD(P)H
oxidizing system reduces ferric ion to ferrous ion, which in turn was used for heme synthesis by
ferrochelatase.
The effect of various metals, including Pb, on RBC porphobilinogen synthase (PBG-S)
was studied using human RBC hemolysate. Farant and Wigfield (1987) reported that the effect
on the enzyme depends on the affinity of the metal for thiol groups at its active sites. Additional
studies carried out by the same group utilizing rabbit erythrocyte PBG-S indicated that Pb acts as
a potent effector of this enzyme both in vitro and in vivo (Farant and Wigfield, 1990). Human
RBC porphobilinogen synthetase activity was found to be inhibited by Pb, whereas Zn ions
activated this enzyme (Simons, 1995). Another enzyme involved in the heme synthetic pathway,
porphobilinogen deaminase, was inhibited in human RBC by Pb nitrate (100 mM) in in vitro
studies, but had no effect in vivo (Tomokuni and Ichiba, 1990). Rossi et al. (1992) reported no
inhibition of coproporphyrinogen oxidase activity in human lymphocytes on exposure to Pb.
Heme synthesis can also be affected in Pb intoxication by interference with Fe transport into
reticulocytes. Using a rabbit reticulocyte model, Qian and Morgan (1990) reported that
inhibitory effects of Pb on transferrin endocytosis and iron transport across the membrane may
also contribute to altered heme metabolism in RBCs. These and other related studies are
summarized in Annex Tables AX5-2.2 and AX5-2.3.

5.2.4 Effects of Lead on Other Hematological Parameters
The RBC pyrimidine 5-nucleotidase (P5N) catalysis of the hydrolytic dephosphorylation
of pyrimidine 5-monophosphates is sensitive to inhibition by Pb. Tomokuni et al. (1989)
evaluated the activity of RBC and bone marrow 5-nucleotidase (P5N) and RBC ALAD in mice
exposed to drinking water Pb (200 to 500 ppm) for 14 or 30 days. These authors reported that Pb
exposure decreased both P5N and ALAD activities in erythrocytes. Additional studies from this
group, using a similar exposure regimen, indicated no change in levels of urinary coporphyrins.
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Lead exposure (4 mg/kg and 6 mg/Kg body wt/30 days) in splenectomized rats was found
to cause depletion of RBC Hb content, to increase numbers of reticulocytes in peripheral blood,
and to increase urinary delta aminolevulinic acid excretion (Gautam and Chowdhury, 1987).
These authors further reported that the increased number of reticulocytes found in the blood may
be due to induced acceleration of the erythropoeitic cell series. Redig et al. (1991) reported
biphasic effects of Pb on hematological parameters from their chronic exposure studies in
red-tailed hawks over 3 or 11 weeks. These authors observed a rapid and relatively brief
increase in RBC free protoporphyrin and a slower, but more prolonged, increase in its Zn
complex with 3-week exposure to Pb (0.82 mg/kg body wt). On the other hand, exposure to a
higher dose of Pb (1.64 mg/kg body wt) for a longer duration (11 weeks) resulted in a decrease
in the Hct and Hb. Panemangalore and Bebe (1996) reported that Zn deficiency increased the
Pb-induced accumulation of porphyrin in RBCs to a lesser extent compared to its accumulation
in the liver in weaning rats.
The effects of Pb on RBC number and other Hct parameters appear to be dose dependent.
lavicoli et al. (2003) investigated these effects by feeding mice with eight different doses of Pb
below (0.6 to <2.0 |ig/dL) and above (>2.0 to!3 |ig/dL) normal background levels. These
authors reported that mice receiving below normal background levels of dietary Pb displayed
enhanced RBC counts and increased Hb and Hct values, whereas a marked decrease in RBC
number occurred when blood Pb levels approached 10 |ig/dL. Sivaprasad et al. (2003) also
reported significant reductions in RBC Hb content and Hct on Pb exposure (0.02% Pb acetate in
drinking water for 5 weeks). Toplan et al. (2004) observed significant decreases in RBC Hb
content and Hct and increases in blood viscosity in Wistar rats after 5-week exposure to Pb.
Studies cited above are summarized in Annex Table AX5-2.4.

5.2.5 Effects of Lead on Erythrocyte Enzymes
The toxic effects of Pb on RBCs result from its complexation with the sulfhydryl,
carboxyl, and imidazole groups of proteins, particularly enzymes, by competitive binding of Pb2+
with Zn2+ or Mg2+ in metalloenzymes. This binding of Pb to enzyme proteins can inhibit
enzymes involved in the glycolytic and pentose phosphate pathway, both of which are sources of
energy compounds and intermediates of purine conversion, thus causing a disruption of energy
metabolism. Along with these changes, Pb-induced changes in the membrane integrity, as
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discussed earlier (Section 5.2.1), may also affect the enzymes' associated ion channels and other
transport mechanisms.

Energy Metabolism
Erythrocytes generate high-energy ATP by anerobic glycolysis and cycle oxidized and
reduced nicotinamide adenine dinucleotide phosphate (NADP) by the aerobic pentose phosphate
pathway. Anemic conditions associated with Pb poisoning, along with the inhibitory effects of
Pb on heme synthesis, may result in increased RBC destruction due to the inhibitory effects of
Pb on the activities of the enzyme, pyramidine 5-nucleotidase (P5N). Deficiency of this enzyme
is characterized by intracellular accumulation of pyramidine-containing nucleotides, leading to
hemolysis. Inhibition of this enzyme, along with perturbations in heme metabolism, creates
imbalances in the energy currency of the erythrocyte. Perturbations in energy metabolism can be
followed by changes in the concentration of purine nucleotides. In erythrocytes, these
compounds cannot be synthesized de novo; they can only be reconstructed from preexisting free
purine bases on nucleosides through salvage type reactions. The cell energy content can be
measured by adenylate (ATP + ADP + AMP) and guanylate (GTP + GDP + GMP) nucleotides,
and by their sum total. The concentrations of nucleoside monophosphates increase in cases of
cell energy deficit, but they quickly degrade to nucleosides and bases.
Cook et al. (1987) compared P5N and deoxypyramidine-5-nucleotidase levels in the RBC
of Pb-exposed workers and matched controls and reported significantly lower levels of P5N in
Pb-exposed workers. Konantakieti et al. (1986) reported similar observations in neonatal rat
RBCs. These authors further indicated that the low levels of nucleotides were due to inhibition
of P5N activity by Pb, as the depression in enzyme activity was correlated with blood Pb levels.
This was further validated by in vitro inhibition of P5N in a dose-dependent manner. Tomokuni
and Ichiba (1988) found similar results with human RBCs both in vitro and in vivo. They
reported activation of Pb-exposed human RBCs. Antonowicz et al. (1990) observed significantly
higher levels of glycolytic enzymes and increased production of lactic acid and 2,3-diphospho
glycerol, when human RBCs were incubated with Pb. Based on their observations, these authors
suggested that Pb exposure may result in anaerobic glycolysis activation in human RBCs.
In contrast, Grabowska and Guminska (1996) reported that Pb exposure diminished the ATP
levels in human RBCs by inhibiting aerobic glycolysis.
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Erythrocyte energy metabolism in workers exposed to heavy metals, but without clinical
manifestations of toxicity, was found to intensify and become more pronounced when they were
occupationally exposed to Pb. Nikolova and Kavaldzhieva (1991) measured the exposed
workers and reported higher ratios of ATP/ADP in Pb-exposed workers. Because the RBC
energy pool is perturbed due to Pb exposure, Morita et al. (1997) evaluated the effect of Pb on
NAD synthetase and reported an apparent dose-dependent decrease in NAD synthetase activity
in the erythrocytes of Pb exposed workers.
Baranowska-Bosiacka and Hlynczak (2003) evaluated Pb effects on distribution profiles
of adenine, guanine nucleotide pools, and their degradation products in human umbilical cord
RBCs. In vitro exposures (Pb acetate; 100 to 200 |ig/dL) equivalent to Pb exposure for 20 h
were found to significantly lower the levels of nucleotide pools, including NAD and NADP,
accompanied by a significant increase in purine degradation products (adenosine, guanosine,
inosine, and hypoxanthine). Associated morphological RBC alterations were also observed, with
marked significant increases in stomatocytes, spherocytes, and echinocytes. These investigators
also observed similar alterations in the nucleotide pools in Wistar rat RBCs with short-term
exposure to Pb (Baranowska-Bosiacka and Hlynczak, 2004). Based on these observations, the
authors postulated that decreases in NAD and NADP concentrations in RBCs may be a good
indicator of Pb-induced disturbance in the energy process and can serve as a useful marker for
chronic Pb exposure. If NAD synthetase activity had been measured in these studies, it might
have provided experimental support for the observation of inhibition of NAD synthetase reported
by Morita et al. (1997).

Other Enzymes
Lead-induced efflux of K+ from human RBCs had been recognized as being due to the
ability of Pb to selectively increase the membrane permeability for this cation. Studying the
efflux of 86Rb using inside-out RBC vesicles, Alvarez et al. (1986) demonstrated that Pb
promoted the selective efflux of K+ ions by altering the sensitivity of Ca2+ binding site on the
membrane either by direct binding or by altering Mg2+-mediated modulation. Fehlau et al.
(1989) indicated that this modulation of the Ca2+-activated K+ channel in human RBCs coincides
with the activation of RBC membrane-bound oxidoreductase. These authors suggested that,
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even though these two are independent events, the oxidoreductase enzyme activity may influence
K channel gating.
Earlier studies by Mas-Oliva (1989) on the potential effects of Pb on the RBC membrane
(using RBC ghosts) indicated that Pb has inhibitory effects on Ca2+-Mg2+-ATPase. Further
investigations on the role of calmodulin in the inhibition of Ca2+-Mg2+-ATPase indicated that the
inhibitory activity on the enzyme may be due either to the effect of Pb on sulfhydryl groups on
the enzyme or by direct binding to calmodulin.
Jehan and Motlag (1995) reported that when albino rats were administered Pb i.p (5 or
20 mg/kg body wt) for 14 consecutive days either alone or in combination with Cu (2 mg/kg
body wt) or zinc (5 mg/kg body wt), there were severe decreases in RBC membrane enzyme,
acetylcholine esterase (AChE), NADH dehydrogenase, and Na+-K+ ATPase levels along with
decreases in phospholipid content, hexose, and hexosamine. Of the combined metal treatment
exposure regimens, Zn was found to considerably reduce such changes. Grabowska and
Guminska (1996) assayed three ATPase activities (i.e., Na+-K+ ATPase, Mg2+-ATPase, and
Ca2+-ATPase) in human RBC in vitro and reported RBC Na+-K+ ATPase to be the only enzyme
inhibited by Pb, while Ca2+ or Mg 2+ATPases were not sensitive to Pb. On the other hand,
Sivaprasad et al. (2003) observed Pb-induced reductions in RBC activities for all three of those
types of ATPase activities.
Two reports by Calderon-Salinas et al. (1999a,b) indicated Pb effects on Ca transport in
human RBC. Initial studies by this group indicated that Pb and Ca are capable of inhibiting the
passive transport of other metals in a noncompetitive way. Inhibition studies using N-ethyl-
maleimide indicated that Pb and Ca share the same permeability pathway in human RBCs and
that this transport system is electrogenic (Calderon-Salinas et al., 1999a). Additional studies by
the same group reported that Pb is capable of inhibiting Ca efflux by inhibiting Ca-ATPase
(Calderon-Salinas et al., 1999b). These authors further suggested that under physiological
conditions, Pb, via Ca2+-ATPase, alters Ca influx, while chronic Pb intoxication inhibits Ca
efflux by altering RBC Ca homeostasis. Silkin et al. (2001) reported Pb-induced activation of K
channels in the RBCs of the teleost fish S. porcus. Exposure of teleost fish RBCs to 1 to 2 jiM
Pb led to a minor loss in cellular K+; but, at 20 to 50 jiM Pb, about 70% of cellular K+ was lost.
Based on their observations of Pb-induced K+ efflux from RBCs under competitive and
inhibitory regimens, these authors suggested that Pb activates RBC K+ channels.
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Eder et al. (1990) and Loipfiihrer et al. (1993) investigated activity levels of Ca2+-ATPase
and Ca accumulation, respectively, in Pb-depleted rat RBCs. No alteration in Ca2+-ATPase
activity or Ca accumulation was observed in the PO generation (Eder et al., 1990). On the other
hand, significant reduction in Ca-ATPase activity was observed in the Fl generation. It was
suggested that Pb-induced alterations in the metabolism of phosphoproteins and glycoproteins
result from Pb depletion and may be responsible for the reduced enzyme activity. Both of the
groups postulated that the decreased MCV observed in Pb depleted rat RBCs could be due to
reduced Ca2+-ATPase activity in the RBCs. These and other related studies are summarized in
Annex Tables AX5-2.5 and 5-2.6.

5.2.6 Erythrocyte Lipid Peroxidation and Antioxidant Defense
Although several mechanisms have been proposed to explain Pb toxicity, no mechanisms
have been defined explicitly. Recent literature on Pb toxicity suggests oxidative stress as one of
the important mechanisms for toxic effects of Pb in various organs. Because RBCs accumulate
major amounts of Pb compared to other tissues, oxidative stress may also result in the
accentuation of lipid peroxidation, with concomitant inhibition of antioxidant enzymes, such as
superoxide dismutase (SOD), catalase, GSH peroxidase, GSH reductase, and simultaneous
increases in oxidized GSH (GSSG) and reduced GSH/GSSG ratios. Below, Pb-induced lipid
peroxidation and the mitigating effects of experimental chelation therapy are discussed with
relevance to each tissue or organ within this chapter. The discussion focuses on the available
literature with reference to studies on erythrocytes.
Patra and Swarup (2000) reported significant changes in RBC lipid peroxide levels and
anti oxidant defense (SOD and catalase) levels in RBC hemolysates from male calves exposed to
Pb (7.5 mg/kg body wt for 28 days). These authors suggested the potential role for increased
peroxide levels in Pb-induced alterations in RBCs. Mousa et al. (2002) investigated the levels of
various antioxidant enzymes, thiols, lipid peroxide in erythrocytes, and total thiol status of
plasma in goats exposed to Pb (Pb acetate, 5.46 mg/kg body wt for 2 weeks). These authors
reported that all the parameters referred to above were significantly increased in RBCs by day 7
and receded to normal levels by day 14, while peroxides remained significantly increased even
by day 14. Based on these observations, it was suggested that Pb-induced lipid peroxide
generation in RBCs appears to be a continuous process and can lead to persistent oxidative stress
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in RBCs with chronic exposure. The effects of chelative agents on RBC lipid peroxidation are

summarized in Annex Table AX5-2.7.

5.2.7 Summary

• The 1986 Lead AQCD reported that the activity of ALAD appeared to be inversely
correlated to blood Pb values and was found inhibited in several tissues. Human studies
reviewed in 1986 Lead AQCD also indicated that occupational exposure to Pb results in
decreased RBC survival along with alterations in RBC membrane integrity and energetics.

• More recent studies reviewed in this AQCD indicate that the transport of Pb across the
RBC membrane is energy-independent, carrier-mediated, and that the uptake of Pb is
mediated by an anion exchanger through a vanadate-sensitive pathway.

• Lead intoxication interferes with RBC survival and alters RBC mobility. Hematological
parameters, such as mean corpuscular volume (MCV), mean corpuscular hemoglobin
(MCH), and mean corpuscular hemoglobin concentration (MCHC), are also significantly
decreased upon exposure to Pb. These changes are accompanied by decreased membrane
sialic acid content.

• Morphological analyses using electron paramagnetic resonance imaging and spin labeling
techniques indicate that changes occur in RBC morphology upon Pb exposure.

• Lead-induced RBC membrane lamellar organization and decreases in membrane lipid
resistance to oxidation in rats appear to be mediated by perturbations in RBC membrane
lipid profiles. Similarly, Pb-induced altered phosphorylation profiles of RBC membrane
proteins have been reported.

• Erythrocyte ALAD activity ratio (ratio of activated/non activated enzyme activity) has
been shown to be a sensitive, dose-responsive measure of Pb exposure, regardless of the
mode of administration of Pb. Competitive enzyme kinetic analyses in RBCs from both
human and Cynomolgus monkeys indicated similar inhibition profiles by Pb.

• Consistent observation of Pb-mediated inhibition of pyramidine 5'-nucleotidase (P5N)
suggests this enzyme as a potential biomarker for Pb exposure.

• Significant reductions in levels of nucleotide pools (e.g., NAD and NADP) accompanied
by significant increase in purine degradation products have been implicated in the
Pb-induced altered energetics of RBCs.

• Lead-induced increased permeability for K+ in RBCs appears to be due to the selective
efflux of K+ ions on the RBC membrane due to altered sensitivity of the Ca2+-binding site
on the membrane. Erythrocyte Na+-K+ ATPase appears to be more sensitive to Pb-induced
inhibition than Ca2+-Mg2+ ATPase.
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The newly available (since 1986) scientific evidence presented in this section clearly
demonstrates deleterious Pb effects on erythrocyte cell morphology and function, as well as Pb
uptake and alterations in certain enzymes involved in heme synthetic pathways. However, some
of the interesting and important conclusions are derived mainly from in vitro studies, often using
short time incubations. It would be useful to substantiate such findings further by more
systematic studies employing meaningful experimental designs for in vivo evaluation of
laboratory animal models.
5.3 NEUROLOGIC/NEUROBEHAVIORAL EFFECTS OF LEAD
5.3.1 Introduction
Since the initial description of Pb encephalopathy in the developing rat in the mid-1960s
(Pentschew and Garro, 1966), a continuing research focus has been on defining the extent of
CNS involvement at subencephalopathic, environmentally relevant levels of exposure. These
efforts have primarily addressed the developing animal, consistent with the primary public health
concerns for neurotoxicity from Pb during this period. While significant research advances have
been made in animal studies over the last four decades, relating these findings to neurotoxicity in
children has been challenging and difficult. The barriers to greater progress have primarily been
due to Pb's multiple toxic mechanisms of action in brain tissue, which encompass variable,
overlapping, and, at times, opposing dose-effect relationships. The goal of this section is to
bring greater clarity to the current state of knowledge.
Discussions of the biologic effects of Pb in the 1986 Lead AQCD focused on general
questions relating to (1) the internal exposure levels, as measured by blood Pb concentrations, at
which neurotoxic effects occur; (2) the persistence or reversibility of these effects; and (3) the
populations that are especially sensitive to the neurotoxic effects of Pb. The state of knowledge
at publication of the 1986 AQCD provided answers for these questions as follows.
At very high levels of exposure producing blood Pb levels of 100 to 120 |ig/dL in adults
and 80 to 100 |ig/dL in children, serious neurotoxic effects occur, including acute Pb
encephalopathy that can progress to convulsions, coma, and sudden death. Less severe
exposures creating blood Pb levels of 40 to 60 |ig/dL produce both central and peripheral nerve
dysfunction, including slowed nerve conduction velocity and overt signs and symptoms of
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neurotoxicity. Decrements in IQ are observed in children with blood Pb concentrations of 30 to
50 |ig/dL, with some studies showing effects at lower blood Pb levels. Neurobehavioral effects
are observed in rats and monkeys at levels <20 |ig/dL.
(1) Human studies provide little information on the persistence of effects. Animal studies
show that alterations in neurobehavioral function can persist long after lead exposure
has stopped and PbB levels have returned to normal. Persistent learning deficits occur
in both rats and monkeys, consistent with morphologic, electrophysiologic, and
biochemical endpoints that indicate lasting changes in synaptogenesis, dendritic
development, myelin and fiber track formation, ionic mechanisms of
neurotransmission, and energy metabolism.
(2) Animal studies show that the order of susceptibility of neurotoxic effects of lead is
young > adults and female > male.
At the time of publication of the 1986 AQCD, one line of evidence concerned the effects
of acute exposure to Pb2+ in vitro on voltage-sensitive Ca2+ channel function in the nerve cell
membrane, developed to a great extent by Cooper and co-workers (Kober and Cooper, 1976;
Cooper and Manalis, 1984; Suszkiw et al., 1984).
In the ensuing two decades, the Pb neurotoxicity literature has reflected an increased
focus on cognitive-related mechanisms and the refinement of approaches and methodologies.
Exposure-induced alterations at glutamatergic synapses have received considerable attention.
Synaptic plasticity models (e.g., long-term potentiation [LTP]) came into use in the 1990s for Pb
studies in laboratories around the world. Behavioral paradigms, refined to more consistently
discriminate Pb effects, aided in identifying optimal testing conditions and developmental
periods for exposure. These advances have led to a clearer understanding of the likely
mechanisms underlying Pb-induced cognitive impairments in exposed children.
The evidence for Pb neurotoxicity reviewed in this section is organized largely according
to scientific discipline: neurochemical alterations involving glutamatergic, cholinergic, and
dopaminergic function; mechanisms defined by neurophysiologic approaches; changes in
auditory and visual function; identification of altered components of behavioral function;
induced alterations in cellular morphology; and findings on cellular disposition of Pb. This type
of organization permits a more focused analysis of an extensive and broad literature. In each
section below, a brief description of work previously described in the 1986 AQCD introduces
each section. An integrative synthesis of the health effects of Pb exposure based on toxicologic
and epidemiologic findings is presented in Chapter 8.
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5.3.2 Neurochemical Alterations Resulting from Lead Exposure
Earlier work demonstrated that Pb interfered with chemically mediated synaptic
transmission, probably due to its resemblance to endogenous divalent cations. At that time, a
selective vulnerability of any particular neurotransmitter system to the effects of the metal was
not apparent. Some generalizations made in the 1986 AQCD regarding neurochemical effects at
blood Pb levels of-50 to -90 |ig/dL were as follows.
(1) Synthesis, turnover, and uptake of dopamine and norepinephrine are depressed in the
striatum and elevated in the midbrain, frontal cortex, and nucleus accumbens. These
changes were paralleled by concomitant increases in dopamine receptor binding in the
striatum and decreases in dopamine receptor binding in the nucleus accumbens, possibly
involving a specific subset (D2) of dopamine receptors.
(2) The findings for pathways utilizing y-aminobutyric acid (GABA) showed similar
parallels. Increases in GABA synthesis in the striatum are coupled with decreases in
GAB A receptor binding in that region, while the converse holds true for the cerebellum.
In these cases, cyclic GMP activity mirrors the apparent changes in receptor function.
The following areas of investigation discussed below have been accorded notable
attention in the Pb neurotoxicity field over the last 20 years, as reflected by the number of papers
published and number of investigators with these research foci.

Lead and Neurotransmitter Release Processes
By the mid-1980s, it was evident that acute exposure to Pb2+ in vitro reduced the
magnitude of depolarization-induced transmitter release, apparently by inhibiting Ca2+ influx
into the nerve ending through voltage-sensitive Ca2+ channels (Kober and Cooper, 1976; Cooper
and Manalis, 1984; Suszkiw et al., 1984). Since then, several investigators utilizing various
preparations (Shao and Suszkiw, 1991 [cortical synaptosomes]; Tomsig and Suszkiw, 1993
[bovine chromaffin cells]; Braga et al., 1999a,b [cultured hippocampal cells]; Westerink and
Vijverberg, 2002 [PC12 cells]) have demonstrated that in the absence of Ca2+, Pb2+ exhibits
Ca2+-mimetic properties in stimulating exocytosis and is substantially more potent in doing so.
That is, in the absence of Ca2+ and depolarization, nanomolar concentrations of Pb2+ alone
stimulate transmitter release. Many investigators have proposed that this action, in conjunction
with the ability of Pb2+ to suppress evoked release of neurotransmitters, produces a higher noise
level in synaptic transmission in Pb-exposed animals.
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Lead has been shown to diminish stimulated transmitter release in intact chronically
exposed animals with blood Pb values in the range of-20 to 40 |ig/dL via intracerebral
microdialysis (Kala and Jadhav, 1995; Lasley and Gilbert, 1996; Lasley et al., 1999). More
recently, Lasley and Gilbert (2002) used Ca2+-free perfusate containing a Ca2+ channel antagonist
for microdialysis to identify the Ca2+-independent component of neurotransmitter release. Under
these conditions, high K+-stimulated release of glutamate and GAB A was elevated in chronic Pb-
exposed animals, suggesting a Pb2+-induced enhancement of evoked release at higher exposure
levels. It was concluded that this pattern of results indicated the presence of two actions of Pb on
transmitter release in vivo: (1) a more potent suppression of stimulated release seen at lower
exposure levels (associated with blood Pb values of 27 to 62 |ig/dL) combined with (2) Ca2+-
mimetic actions that independently induce the exocytosis seen at higher exposure levels
(associated with blood Pb values of >62 jig/dL). Together, these two actions produce a biphasic
dose-effect relationship (see Figure 5-2). Thus, there is good correspondence between findings
of in vitro and in vivo studies with respect to the actions of Pb on transmitter release.

Lead and Glutamatergic NMDA Receptors
Because of the established importance of the TV-methyl-D-aspartate (NMDA) subtype of
glutamate receptor in synaptic plasticity and learning, these receptors have been a focus of
intense interest in Pb neurotoxicity for the last 15 years. Using whole cell and single channel
patch clamp methodologies, Alkondon et al. (1990) were the first to report that Pb2+ inhibited the
function of the NMDA receptor channel complex. Guilarte and Miceli (1992) reported similar
findings using nominal Pb2+ concentrations and receptor binding techniques and drew parallels
between Zn2+-, Ca2+-, and Pb2+-induced inhibition of the channel. However, Lasley and Gilbert
(1999), using free Pb2+ ion concentrations and radioligand binding, demonstrated that despite
similarities to the actions of Zn2+, Pb2+ did not inhibit the NMDA receptor channel complex by
binding to the Zn2+ allosteric site. Furthermore, these workers indicated that the Pb2+ ICso of
0.55 jiM for inhibition of the channel complex was likely about two orders of magnitude greater
than the extracellular fluid concentrations of Pb2+ associated with environmentally relevant
exposure. This does not imply that NMDA receptor function does not change after Pb exposure,
but it strongly suggests that the alterations are not based on a direct Pb2+ action.
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5.0
4.5
4.0
3.5
3,0
2.5
2.0
1.5
1.0
0.5
0.0
150 mM K+
Total Release
-O— Control
•- - • 0.2% Pb
l I
-120 -90 -60 -30 0 20 40 60 80
Time, min
110 140
tf)
m
•0
9
.*:
O
5.0
4.5
4.0
3.5
3.0
2.5
2.0
1.5
1.0
0.5
0.0
*
*
T
*
*
#
J_
#
Control
0.1%
0.2%
0.5%
1.0%
Figure 5-2. Time course and magnitude of response of extracellular GLU concentration as
a result of chronic lead exposure.
*** p < 0.001; ** p < 0.01 relative to the GLU concentration in control animals;
*p < 0.0001 relative to the GLU concentration in the 0.2% Pb group.
Source: Lasley and Gilbert (2002).
5-22
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Unfortunately, a consensus on the effects of chronic Pb exposure on NMDA receptor
expression and function has not been achieved. Extensive effort has been invested to assess
NMDA receptor subunit mRNA and protein expression in exposed animals with blood Pb values
in the range of 25 to 45 ng/dL (Guilarte and McGlothan, 1998; Nihei and Guilarte, 1999;
Guilarte et al., 2000; Nihei et al., 2000; Toscano et al., 2002; Guilarte and McGlothan, 2003),
but consistent findings have not emerged. A possible exception was the work of Nihei et al.
(2000, 25 to 32 ng/dL) who found decreases in hippocampal NR1 subunit mRNA and protein
expression associated with animals that exhibited deficits in LTP and spatial learning after
chronic exposure. Correlations of this type with functional measures are valuable in validating
biochemical observations. However, it should also be noted that such correlations do not
confirm a direct relationship between LTP or the behavior and the NMDA receptor subunit
changes.
While exposure-induced alterations of NMDA receptor binding have been observed in
multiple laboratories, there has been no uniform agreement as to the direction of change.
Upregulation of NMDA receptor density has been observed in rats continuously exposed
throughout development with blood Pb values in the range of 39 to 62 |ig/dL (Ma et al., 1997;
Lasley et al., 2001), but receptor downregulation has also been reported when exposure was
begun immediately postweaning and blood Pb levels achieved 16 to 28 |ig/dL (Cory-Slechta
et al., 1997a). The results of behavioral investigations are most parsimoniously explained by
increases in NMDA receptor density. Cohn and Cory-Slechta (1993, 1994a), using a repeated
learning component of a multiple reinforcement schedule, observed enhanced performance
sensitivity to exogenous NMDA administration and diminished sensitivity to MK-801, an
NMDA receptor antagonist, in exposed animals with blood Pb values of 25 to 74 |ig/dL.
The same findings resulted when a drug discrimination paradigm was utilized (Cory-Slechta,
1995; Cory-Slechta et al., 1996b): enhanced sensitivity to NMDA and reduced sensitivity to
MK-801 in Pb-exposed groups in the presence of blood Pb values in the range of 13 to 36 |ig/dL.
A decreased sensitivity to MK-801 can result from either increased numbers of NMDA receptors
or a diminished access of the antagonist to its binding site in the ion channel. Thus, all these
behavioral observations may be accounted for by Pb-induced increases in NMDA receptor
density resulting in increased sensitivity to agonists coupled with decreased sensitivity to
antagonists. That is, the functional measures suggest that an NMDA receptor upregulation
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occurs. This interpretation should not preclude the possibility that experimental outcomes can
change significantly in the presence of apparently small modifications in exposure parameters.

Pb2+-Ca2+ Interactions
At the time of publication of the 1986 AQCD/Addendum, one of the most reproducible
lines of evidence concerned the effects of acute exposure to Pb2+ in vitro on voltage-sensitive
Ca2+ channel function in the nerve cell membrane, developed to a great extent by Cooper and
co-workers (Kober and Cooper, 1976; Cooper and Manalis, 1984; Suszkiw et al., 1984). Using
neuromuscular endplate or synaptosomal preparations, these studies demonstrated that Pb2+
interfered with Ca2+ influx through voltage-sensitive channels. Subsequent work has replicated
and extended these findings (e.g., Tomsig and Suszkiw, 1993; Westerink and Vijverberg, 2002),
and has demonstrated that Pb2+ exhibits Ca2+-mimetic properties in stimulating transmitter
exocytosis. While acute exposure in vitro has been assumed to bear little resemblance to
environmentally relevant routes and magnitudes of exposure, recent findings nonetheless suggest
that inhibition of Ca2+ influx through voltage-sensitive Ca2+ channels and the Ca2+-mimetic
properties of Pb2+ are important neurotoxic mechanisms in intact animals across a range of
chronic exposure levels (Lasley and Gilbert, 2002).
Simons (1993b) has reviewed the ability of Pb2+ to disturb intracellular Ca2+ homeostasis,
and has emphasized the importance of utilizing free Pb2+ concentrations to define Pb2+-Ca2+
interactions clearly. Multiple laboratories have investigated the inhibition of depolarization-
induced Ca2+ currents produced by acute exposure of cultured cells using this approach, resulting
in free Pb2+ ICso values in the range of 0.3 to 1.3 jiM (e.g., Audesirk and Audesirk, 1991; Sun
and Suszkiw, 1995). Other workers examined the stimulation of spontaneous transmitter release
by acute exposure of permeabilized synaptosomes or cultured cells (Shao and Suszkiw, 1991;
Tomsig and Suszkiw, 1996) and reported a free Pb2+ EC50 of 4 nM. Westerink and Vijverberg
(2002) addressed this same question using fluorescent dyes and confocal laser scanning
microscopy of permeabilized PC 12 cells, an independent approach also based on determination
of free Pb2+ concentrations. They observed a threshold for acute Pb2+ to induce exocytosis of
between 10 and 20 nM. Suszkiw (2004) has reviewed this literature and has suggested that
Pb2+-induced augmentation of spontaneous release may involve stimulation of vesicle
mobilization consequent to Pb2+ activation of CaMKII-dependent phosphorylation of synapsin I
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and/or stimulation of asynchronous exocytosis via direct Pb2+ activation of the putative
exocytotic Ca2+-sensor protein synaptotagmin I. Other Ca2+-dependent proteins whose actions
are stimulated by Pb2+ include calmodulin and calmodulin-dependent phosphodiesterase
(reviewed by Goldstein, 1993), calcineurin (Kern and Audesirk, 2000), and Ca2+-ATPase
(Ferguson et al., 2000). These actions of Pb2+, shown in Figure 5-3, are proposed to be the
points of initiation of much of the metal's cellular toxicity.

Pb2+ and Protein Kinase C
As mentioned above, another intriguing focus area for Pb neurotoxicity research has been
the interactions of Pb2+ with protein kinases. Protein kinase (PKC), a family of serine/threonine
protein kinases, has been shown to be targets of Pb2+ neurotoxicity. Markovac and Goldstein
(1988a) were the first to report that Pb2+ directly stimulated PKC activity at picomolar
concentrations, thereby exhibiting greater potency for this action than Ca2+ by 4 to 5 orders of
magnitude. Long et al. (1994) made similar observations using free Pb2+ and Ca2+ ion
concentrations and nuclear magnetic resonance spectroscopy, finding an EC50 of 55 pM for
Pb2+ stimulation of PKC. These workers also presented evidence suggesting that the maximal
efficacy of Pb2+ was less than that of Ca2+, despite its greater potency. Tomsig and Suszkiw
(1995) further elucidated multiple interactions of Pb2+ with PKC, identifying both stimulatory
(affinity in the picomolar range) and inhibitory (affinities in the nanomolar and micromolar
range) binding sites on the kinase. They also showed that on the basis of these interactions, Pb2+
induced a peak efficacy for stimulation of PKC that was only -40% of the maximal efficacy
produced by Ca2+, leading them to refer to Pb2+ as a partial agonist of the kinase as reflected in
Figure 5-4.
The effects of chronic Pb exposure on PKC signaling have been more difficult to evaluate.
Most investigators have utilized broken cell preparations and measures of either kinase
translocation or enzyme activity; however, the broken cell preparation has not been shown to
simulate the intracellular milieu of a chronically exposed intact animal. In the preparation of a
tissue extract for determination of kinase activity, the unbound Pb2+ is removed or greatly
diluted, so that the resulting activity measure largely reflects changes in total PKC expression
resulting from the exposure. That is, this measure does not identify a synaptic pool of PKC or
necessarily represent the pool of kinase involved in signal transduction. Alternatively, the
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Presynaptic Terminal
Post Synaptic Dendritic Spine

Figure 5-3. Simplified diagram showing the actions of lead at a synapse. Lead decreases
release of GABA, dopamine, and glutamate and also decreases Ca2+ movement
through voltage-sensitive Ca2+ channels. Low levels of Pb increase PKC, while
higher concentrations inhibit the enzyme. GABA, y-aminobutyric acid; ER,
endoplasmic reticulum; DA, dopamine; Glu, glutamate; PKC, protein kinase
C; NMDA, 7V-methyl-D-aspartate.
translocation of kinase from a cytosolic to membrane cellular fraction is a somewhat nonspecific
measure, and observed changes should be independently confirmed. From the effects of acute
Pb2+ exposure in vitro, it seems clear that PKC is a lexicologically significant intracellular target
for Pb2+. However, various investigators have been unable to define how this acute effect
translates, if at all, to chronic exposure in the intact animal. Neither is it evident how one could
discriminate inhibition of PKC activity (due to decreased efficacy relative to that associated with
Ca2+, for example) from downregulation of the enzyme due to prolonged stimulation. Resolution
of these issues awaits the development of more specific methodologies.
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0.0
14 13 12 11 10 9 8 7 6 5 4
-log [free metal]
Figure 5-4. PKC activity as a function of Ca2+ and Pb2+ concentrations.
Source: Tomsig and Suszkiw (1995).
Lead Exposure and Cholinergic Neuronal Systems
The actions of chronic Pb exposure have also been studied with respect to changes in CNS
cholinergic systems as another substrate thought to underlie cognitive function. Bielarczyk et al.
(1996) reported (1) decreased functional cholinergic innervation in the hippocampus and
(2) depression of choline acetyltransferase activity in the hippocampus and cortex of young adult
rats exposed to Pb only during early development. This model produced a blood Pb level of
22 |ig/dL at the end of exposure. Similar changes were reported by Bourjeily and Suszkiw
(1997) in which blood Pb levels were -20 |ig/dL, leading to the conclusion that perinatal Pb
exposure results in a loss of septohippocampal cholinergic projection neurons that persists until
testing in young adulthood. Tian et al. (2000) exposed PC12 cells to Pb2+ for <48 h and found
that the downregulation of choline acetyltransferase activity reflected the effects of the metal at
the level of gene expression. Consistent with these findings, Jett et al. (2002) employed a similar
perinatal exposure protocol, producing a blood Pb of 47 |ig/dL at the end of exposure. They
observed increased nicotinic receptor binding in multiple brain regions. Zhou and Suszkiw
(2004) found that acute systemic nicotine reversed a deficit in spatial learning observed in the
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offspring of maternally Pb-exposed rats, presumably by compensating for deficient nicotinic
function. However, blood Pb levels in the Pb-exposed animals were not reported. These reports
reinforce the belief that Pb exposure during early development impacts cholinergic function and
suggest that these actions may comprise a component of the cognitive impairment resulting from
exposure to the metal.

5.3.3 Actions of Lead Exposure Defined by Neurophysiologic Approaches
An important advance in Pb neurotoxicity research over the past two decades is the
widespread application of synaptic plasticity models to study of the effects of exposure. These
plasticity models have served as an intermediate link between biochemical and behavioral
assessments in that they demonstrate the functional importance of underlying neurochemical
mechanisms. Moreover, these models are thought to involve the same physiological substrates
as do behavioral paradigms examining cognitive function. Key studies are summarized in
Table AX5-3.1.

Chronic Lead and Models of Synaptic Plasticity
About 1990, the LTP model of synaptic plasticity began to be used to study Pb
neurotoxicity to evaluate the synaptic processes involved in learning and cognitive function.
These investigations have characterized the actions of chronic exposure across several
experimental parameters (see Table 5-1). Furthermore, there was uniform agreement as to the
alterations that resulted in the hippocampal CA1 and dentate gyrus subregions.
Chronic developmental Pb exposure decreased the magnitude of LTP and increased the
threshold for LTP induction (Altmann et al., 1993; Gilbert et al., 1996; Gutowski et al., 1998;
Ruan et al., 1998). Simultaneous assessments of paired-pulse functions also uncovered
reductions in paired-pulse facilitation, indicating reduced glutamate release (Lasley and Gilbert,
1996; Ruan et al., 1998). It was also shown that the potentiation produced in Pb-exposed
animals decayed more rapidly than in controls (Gilbert and Mack, 1998). Blood and brain Pb
values reported in these studies are shown in Table 5-1; Lasley and Gilbert (1996) reported the
same values as shown for Gilbert et al. (1996).
Gilbert et al. (1999a) compared the effects on LTP when exposure occurred during
different developmental periods (see Table 5-1 for blood and brain Pb levels). These workers
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Table 5-1. Chronic Lead Exposure and LTP
Exposure Effect of Exposure on
Recording Site Period1 Blood Pb2 Brain Pb3 Preparation LTP

Hippocampal Dentate Gyrus

Gilbert et al. (1996) PO - P90-120

Ruanetal. (1998) PO-P90-115

Gilbertetal. (1999a) G16-P130-210
P30-P130-210

Gilbert etal. (1999b) G16-P120-180
Gilbert and Mack
(1998)

Hippocampal CA1

Altmann et al. (1993)
Gl 6-P210-540
GO-P70-210
Gutowski et al. (1998) GO - P90-130

Hippocampal CAS

Gutowski et al. (1997) GO - PI 3-140

Gutowski et al. (1998) GO - P90-130
37.2

30.1

40.2
38.7

26.84
40.2
61.8

ND
14.3

16.0

28.5

16.0
ND In vivo Elevated induction threshold

180 In vivo Diminished magnitude

378 In vivo Elevated induction threshold
350 and diminished magnitude

220 In vivo Elevated induction threshold
378 and diminished magnitude
670

ND In vivo Accelerated decay
160 slices Blocked, required exposure
during early development

135 Slices Diminished magnitude
180 Slices No effect across 4 ages

135 Slices No effect
'Exposure duration in terms of gestational (G) or postnatal (P) days of age; PO = day of birth.
2Values expressed as ug/dL.
3Values expressed as ng/g tissue.
4Different blood Pb values generated by differing levels of exposure.
ND = Not determined
found that animals whose exposure began shortly after weaning exhibited the same impairments

in LTP as animals continuously exposed from late gestation when testing in both groups

occurred well into adulthood. A smaller effect on potentiation was observed when exposure was

restricted to the period from late gestation to weaning. In this study, exposure duration varied

significantly, from 26 to >200 days among treatment groups, which suggests the relative

importance of duration of exposure in addition to period of exposure for the creation of

Pb-induced deficits.
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Gilbert et al. (1999b) also examined the effects of Pb on LTP as a function of chronic
exposure level, using a range of 0.1 to 1.0% Pb in the drinking water (corresponding to PbB
values of 27 to 118 |ig/dL; brain Pb measures are shown in Table 5-1). A reduced capacity for
LTP was found at all exposure levels except in the 1.0% group, indicating a biphasic dose-effect
relationship (Figure 5-5). The 1.0% Pb-exposure level was clearly less effective than the lower
exposure groups in reducing LTP magnitude and did not differ significantly from control values.
Blood Pb values were elevated as a function of increasing exposure and could not account for the
lack of effect in the 1.0% exposure group.
Zhao et al. (1999) utilized low frequency electrical stimulation in the paradigm of long-
term depression (LTD) and found that chronic Pb exposure, producing a blood Pb level of
30 |ig/dL, depressed the magnitude of this form of synaptic plasticity in both hippocampal CA1
and dentate gyrus subregions. The authors also concluded that in combination with the reduced
magnitude of LTP as reported by other workers, the decrease in LTD magnitude results in a
reduced range of synaptic plasticity in chronically exposed subjects.
1
0 0.1 0.2 0.5 1.0
Pb Concentration (%)

Figure 5-5. Difference score measure of population spike amplitude.
Source: Gilbert et al. (1999b).
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While the effects of Pb on synaptic plasticity are quite similar in the CA1 and dentate
gyrus, they are not uniformly present throughout this region of the hippocampus. Gutowski et al.
(1997, 1998) were unable to find any effect of chronic Pb exposure on LTP in hippocampal CAS
(i.e., mossy fiber LTP), even when the investigation was extended across multiple ages (see
Table 5-1 for blood and brain Pb values). The bases for this regional distinction await future
investigation.

Lead Exposure, Glutamatergic Transmission, and Synaptic Plasticity
Investigation of the synaptic processes underlying LTP has provided insight into the bases
for Pb exposure-induced impairment of potentiation and cognitive ability (reviewed by Lasley
and Gilbert, 2000). Biochemical and neurophysiologic approaches (Lasley and Gilbert, 1996;
Gilbert et al., 1996; Ruan et al., 1998) have found stimulated glutamate release to be diminished
in the hippocampus at PbB values where deficits in LTP have been observed. Multiple actions
of Pb may be involved at this exposure level, because animals exposed postweaning exhibited
similar decrements in evoked glutamate release to those exposed continuously from conception
(Lasley et al., 1999; adult blood Pb values of 39 to 45 jig/dL), similar to the observations for
measures of LTP (Gilbert et al., 1999a). A biphasic dose-effect relationship was also found in
which stimulated glutamate release in the hippocampus was decreased at intermediate exposures
(blood Pb of 27 to 40 |ig/dL), but not at higher levels (blood Pb of 62 to 117 |ig/dL) (Lasley and
Gilbert, 2002). On the basis of these observations, it appears that decreases in stimulated
glutamate release may contribute to the biphasic dose-effect relationship in LTP.
In comparison to the high concordance across laboratories with regard to the effects of
chronic Pb exposure on LTP and the notable similarities to its actions on glutamate release, there
is no general agreement as to the exposure-induced changes in the NMDA receptor. Alterations
in receptor function occur readily in response to externally applied treatments and might be
expected to vary in a dynamic fashion as a function of exposure parameters, e.g., Lasley et al.
(2001) reported receptor upregulation at blood Pb levels in the range of 39 to 62 |ig/dL.
However, most studies have involved measures of NMDA receptor expression binding in adult
animals exposed to constant levels of Pb for at least 3, and more commonly for 6 to 15 months,
so that receptor-mediated effects should have stabilized. Consequently, the following alternative
conclusions could be proposed regarding the actions of Pb exposure on the NMDA receptor that
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are related to its effects on LTP. First, changes in NMDA receptor function may depend on
specific Pb exposure conditions. For example, a postweaning exposure protocol may not
necessarily produce similar effects to an exposure protocol initiated during earlier development.
Alternatively, effects on LTP may be produced at signal transduction or other cellular loci that
exert regulatory influences on the NMDA receptor. This latter conclusion implies that changes
in the NMDA receptor do not mediate the primary action of Pb on LTP. Furthermore, this
suggests that identification of some site of direct Pb effect that has regulatory influence on the
receptor would produce more consistently observable findings.

Lead and Electrophysiologic Changes in Dopaminergic/Cholinergic Systems
Electrophysiologic approaches have been employed to delineate other interesting findings
in Pb-exposed animals not directly related to synaptic plasticity. Using standard extracellular
recording methods, Tavakoli-Nezhad et al. (2001) identified an exposure-dependent decrease in
the number of spontaneously active dopamine cells in the substantia nigra and ventral tegmental
area in the presence of blood Pb levels of 29 to 54 |ig/dL, but they found no evidence that this
decrease was related to a physical loss of cells. In subsequent work, Tavakoli-Nezhad and Pitts
(2005) determined that the decrease in the number of active dopamine cells in the presence of
blood Pb values of 31 |ig/dL was not based on depolarization inactivation. However, they
discerned a reduced impulse flow in dopamine neurons and a diminished sensitivity of DI
receptors in the nucleus accumbens. The functional importance of these observations remains to
be determined.
The actions of Pb2+ on cholinergic nicotinic receptors have been investigated in acutely
dissociated or cultured hippocampal cells using the patch clamp technique in whole cell mode
(Ishihara et al., 1995). These workers found that Pb2+ potently inhibits activation of fast-
desensitizing nicotinic currents in a noncompetitive and voltage-dependent manner. The
nicotinic receptors affected (methyllycaconitine-sensitive) were more sensitive to Pb2+ than other
nicotinic subtypes and are known to be highly permeable to Ca2+. This latter observation likely
explains the potency for their inhibition by Pb2+.
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5.3.4 Lead Exposure and Sensory Organ Function
Research assessed in the 1986 AQCD demonstrated that the visual system is sensitive to
perturbation by neonatal Pb exposure. Suckling rats exposed through dams' milk creating blood
Pb values of 65 |ig/dL at postnatal day (PND) 21 had significant alterations in their visual-
evoked responses and decreased visual acuity, indicating depressed conduction velocities in
visual pathways. It was hypothesized that neonatal lead exposure increases the ratio of
excitatory to inhibitory systems in the developing cerebrospinal axis and decreases the number of
cholinergic receptors, leading to lasting decreases in visual acuity and spatial resolution.
Sensory organ function has continued to be a productive focus area for Pb neurotoxicity
research, generating important scientific findings. Visual and auditory systems have received the
most attention, have generated results closely resembling clinical observations, and have been
successful in defining some of the mechanisms underlying the exposure-induced alterations.
These studies are summarized in Table AX5-3.3.

Sensory Organ Assessments in Nonhuman Primates
Lilienthal and Winneke (1996) tested monkeys continuously exposed to Pb from gestation
through 8 to 9 years of age, producing blood Pb values of 33 to 56 |ig/dL before termination of
exposure. They found increased latencies for waves I, II, and IV in brainstem auditory evoked
potentials. These effects persisted for at least 18 months after exposure was terminated and
blood Pb values had declined nearly to control levels. Rice (1997) determined pure tone
detection thresholds in monkeys exposed continually from birth to 13 years of age, resulting in
blood Pb levels of 30 |ig/dL from 3 to 9 years of age and 50 to 170 |ig/dL around the time of
testing. Half of the subjects exhibited thresholds outside of the control range at some
frequencies. These findings are consistent with reported alterations in auditory function in
humans developmentally exposed to Pb (reviewed by Otto and Fox, 1993). Moreover, these
authors concluded that evidence from human and animal studies indicate that Pb exposure
impairs auditory function. In both developing and mature humans and experimental animals, the
cochlear nerve and more central structures appear to be preferentially sensitive. At low to
moderate levels of Pb exposure, consistent findings include elevations in hearing thresholds and
increased latencies in brainstem auditory evoked potentials. Thus, there is good correspondence
between human and animal studies in the effects of chronic Pb on auditory function.
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Visual pathology was assessed by Reuhl et al. (1989) in monkeys by exposing low- and
high-dose groups from birth to 6 years of age. Blood Pb values were 10 and 50 |ig/dL,
respectively, except for a 3- to 4-month period, when values rose to 20 and 220 |ig/dL. This
investigation uncovered a decrease in neuronal volume density in cortical areas VI and V2 in the
high-exposure compared to the low-exposure group, and also a decrease in dendritic arborization
in pyramidal neurons in these brain areas. These authors concluded that chronic developmental
Pb exposure produces changes in cytoarchitecture in visual projection areas. Lilienthal et al.
(1988) continuously exposed monkeys to 350- or 600-ppm Pb acetate beginning prenatally,
producing blood Pb values of-40 and 50 |ig/dL, respectively. Visual evoked potentials and
electroretinograms (ERG) were recorded at ~7 years of age. Exposure-related decreases in
amplitudes and increases in latencies were observed. In Pb-exposed monkeys, the effects on
amplitude were greater in dark conditions, and the effects on latencies were greater in bright
conditions. Electroretinograms, studied during dark adaptation, showed greater increases in
amplitude of the b-wave in exposed animals. Thus, visual function in primates is also impaired
as a result of exposure.

Retinal Function in Rodents
The actions of Pb on retinal cells have been a focus of research for more than two
decades. It has long been recognized that Pb2+ exhibits a selective effect on rod cells (Fox and
Sillman, 1979) and, more recently, that the associated loss of rod and bipolar cells was due to
exposure-induced apoptotic changes (e.g., Fox et al., 1997, in the presence of blood Pb values of
19 to 59 |ig/dL at the termination of exposure). These observations have been linked with
exposure-related alterations in rod-mediated visual function. In vitro studies using free Pb2+ ion
concentrations have done much to elucidate the mechanistic bases of these observations.
These latter efforts have established the concentration-dependent inhibition of cyclic
GMP (cGMP) hydrolysis by free Pb2+, in addition to increases in retinal cGMP and rod Ca2+
levels (e.g., Srivastava et al., 1995). Kinetic studies have shown that picomolar Pb2+
concentrations competitively inhibit rod cGMP phosphodiesterase relative to the millimolar
concentrations that are required for Mg2+ cofactor activity, thus binding with 104- to 106-fold
higher affinity than Mg2+ and preventing cGMP hydrolysis (Srivastava et al., 1995). When
retinas are incubated in Ca2+ and/or Pb2+ in vitro, the rods selectively die by apoptosis associated
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with mitochondrial depolarization, release of mitochondrial cytochrome c, and increased caspase
activity (He et al., 2000). He et al. (2003) have proposed that apoptosis is triggered by Ca2+ and
Pb2+ overload resulting from translocation of cytosolic Bax to the mitochondria, which likely
sensitized the overloaded mitochondria to release cytochrome c. This effect occurred at a blood
Pb level of 26 |ig/dL at the end of exposure. Subsequent work found the elevations in free Ca2+
and Pb2+ to be localized to photoreceptors and determined that the effects of the two ions were
additive and blocked by a mitochondrial permeability transition pore inhibitor (He et al., 2000).
This suggested that the two ions bind to the internal metal binding site of this pore and, thereby,
initiate the apoptosis cascade.
These mechanisms are consistent with ERG changes observed in animals chronically
exposed during early development: decreases in maximal ERG amplitude, decreases in absolute
ERG sensitivity, and increases in mean ERG latency that were selective for rod photoreceptors in
the presence of blood Pb values of 59 |ig/dL at the termination of exposure (Fox and Farber,
1988). Also in agreement with these mechanisms were observed elevations in retinal cGMP
levels and reductions in light-activated cGMP phosphodiesterase activity. Moreover, the
degenerating rod and bipolar cells exhibited the classical morphological features of apoptotic cell
death (Fox et al., 1997). Other measures of visual function in chronically exposed animals also
have been found to be consistent with the mechanistic data. Long-term dose-dependent
elevations in response thresholds were observed only at scotopic (i.e., rod-mediated) levels of
illumination, and dark adaptation was delayed (Fox et al., 1994; in the presence of blood Pb
values of 19 to 59 |ig/dL at the termination of exposure). In addition, exposure-induced
decreases in rhodopsin content that were proportional to the loss of rod cells have been reported
(Fox et al., 1997) as well as dose-dependent decreases in retinal Na, K- ATPase activity (Fox
et al., 1991a; blood Pb levels as above in Fox et al., 1994).
The studies investigating rod photoreceptors are perhaps the best examples of the ability
to correlate data obtained in vitro with findings derived from in vivo exposure and with changes
in visual physiology. In multiple instances, the same cellular mechanisms were affected with
each approach and are consistent with ERG and rod-mediated functional measures. These
relationships are summarized in Table 5-2.
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Table 5-2. Mechanisms of Lead-Induced Impairment of Retinal Function
In Vitro Evidence
In Vivo Evidence
Physiologic Changes
Competitive inhibition of
cGMP PDE

Increased retinal cGMP
Increased rod [Ca2+]

Apoptosis from increased
photoreceptor Ca2+/Pb2+ via
binding to mitochondrial
permeability transition pore
Decreased retinal Na
K+-ATPase activity
Decreased stimulated cGMP PDE
activity

Increased retinal cGMP
Decreased maximal ERG amplitude
Decreased absolute ERG sensitivity
Increased mean ERG latency
Morphological features of apoptotic rod, Increased response thresholds at
bipolar cell death scotopic backgrounds
Decreased rhodopsin proportional to Delayed dark adaptation
cell loss
Translocated cytosolic Bax to the
mitochondria, cytochrome c released

Decreased retinal Na+,K+-ATPase
activity
Abbreviations: PDE, phosphodiesterase; ERG, electroretinogram.
5.3.5 Neurobehavioral Toxicity Resulting from Lead Exposure

As discussed elsewhere in this chapter, the young are vulnerable to the effects of Pb

exposure due to their greater absorption and retention of Pb. The developing state of the nervous

system makes the perinatal period a particularly critical time for the initiation of neurobehavioral

perturbations by exposure to Pb. However, work reviewed in the 1986 Lead AQCD showed that

behavioral effects in animals are found with both perinatal exposures and with exposures after

weaning or during adulthood.

Very early research on the effects of Pb on learning ability failed to adequately report

exposure regimen or the resulting body burden. Studies reviewed in the 1986 AQCD were more

useful; they reported this information and further attempted to control for the confounding

factors of litter size and undernutrition. Thirty-four rat studies were evaluated, from which it

was possible to ascertain that learning was altered at blood Pb levels of 15 to 30 |ig/dL. Test

methods that revealed learning deficits in rats included radial arm maze testing and fixed-interval

(FI) operant conditioning. Rats in these studies tended to respond more rapidly (i.e., higher
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response rates, shorter interresponse times, or shorter response latencies) or to respond even
when inappropriate (i.e., when no reward is provided for responses or when reward is
specifically withheld for responding). Impaired acquisition of discrimination and performance in
other tests has been demonstrated with similar blood Pb levels in rats.
Thirteen nonhuman primate studies previously reviewed showed that exposures from birth
impaired learning ability, even after current blood Pb levels had dropped to control values.
Studies using operant conditioning tasks demonstrated that learning ability was impaired when
monkeys' blood Pb levels reached 15 |ig/dL and steady state levels were 11 |ig/dL. Other
significant findings in these studies, consistent with the rat findings mentioned above, were the
tendency for Pb-induced excessive or inappropriate responses in the monkeys and higher
response rates and shorter interresponse times on FI operant schedules. The neural mechanisms
responsible for this "hyperreactive" behavior were thought to originate in the hippocampus, as
similar behaviors have been shown in animals with lesions of that brain region. These increased
response tendencies were shown to change to decreased responding with sufficiently high
exposure levels. An explanation of this curvilinear dose-response was that there are differences
in the time required for response rates to reach their maximum as a function of different exposure
levels. At sufficiently toxic Pb concentrations, responding declines due to the inability to
perform the necessary motor responses.
A survey of the major animal studies published since the 1986 AQCD that characterized
Pb-induced neurobehavioral deficits that may correlate with behavioral deficits observed in
humans are presented below, organized by endpoint, test method, and species. Summaries of
key animal neurobehavioral studies are presented in Annex Table AX5-3.4.

Effects ofPb on Learning Ability
In the past 20 years, major advances have occurred in the understanding of the effects of
Pb on learning ability, which is impacted throughout the life cycle. Assessments of Pb-induced
deficits in learning ability in both rats and primates are discussed below.

Schedule-Controlled Behavior
Schedule-controlled behavior studies, such as fixed interval (FI) and fixed ratio (FR)
operant conditioning, have been used with both rats and monkeys to assess cognitive ability
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(integrated with sensory and motor abilities). The effects of prolonged Pb exposure on FR
performance was evaluated in Long Evans (LE) rats exposed throughout the study to 50 or
500 ppm from weaning, producing blood Pb levels of 30.3 and 58 to 94 |ig/dL, respectively
(Cory-Slechta, 1986). At PND 55, assessment of FR performance was started using increasing
ratio values. No effects were seen in the 50-ppm group. In the 500-ppm group, response rates
initially decreased, then reached control levels, primarily because of longer interresponse times
(TRTs). In comparing these data to earlier studies with similar Pb exposures, the author
concluded that FI response rates are more sensitive to perturbation by Pb than FR response rates.
To evaluate the effects of exposure duration on FI performance, PND 21 LE rats were
exposed to 50 ppm Pb for 8 to 11 months, then tested using an FI 1-min schedule of food
reinforcement (Cory-Slechta, 1990a). These rats demonstrated decreased FI response rates
(i.e., longer IRTs and lower running rates) compared to controls. The author suggested that this
suppression of FI response rates, which contrasted with earlier studies showing increased
response rates with shorter exposures but similar blood Pb levels (-20 jig/dL), was due to the
greater body and brain burdens of Pb. Following changes in schedule parameters, Pb-treated rats
demonstrated a delay in acquisition. In the same study, adult rats (6 to 8.5 months at exposure)
were trained on FI schedules and then exposed to 50 or 500 ppm Pb for 3 to 5 months. These
animals demonstrated no consistent changes in FI performance, suggesting that once a behavior
has been acquired, it may be resistant to disruption by subsequent Pb exposure.
To examine old age as a possible vulnerable period for Pb exposure, Cory-Slechta and
Pokora (1991) dosed Fischer 344 (F344) rats at PND 21, 8-months of age (adult), and 16-months
of age (old) with 2 or 10 mL/kg/day Pb acetate for 9.5 months. Training began 2.5 months after
the start of exposure. Steady state blood Pb levels of 13 to 18 |ig/dL were obtained. Young and
old rats demonstrated increased variable-interval (VI) and FI response rates, while adult rats
showed decreased response rates on both schedules. Effects on FI responding were seen with the
2 mg dose and on VI with only the 10 mg dose. Additionally, these data suggest that F344 rats
are less sensitive to Pb effects than the LE rats used in most of the previous schedule-controlled
behavior studies.
To characterize neurotransmitter system involvement in Pb-induced changes in FI
performance, rats were exposed from weaning to 0, 50, or 150 ppm Pb acetate, resulting in blood
Pb levels of ~<5, 15 to 25, and 30 to 50 |ig/dL, respectively (Cory-Slechta et al., 1996b).
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Behavior was shaped at PND 40 to 45 days, followed by imposition a FI 2-min schedule of
reinforcement. Dopaminergic (DA) agonists quinpirole (02), SKF38393 (Di), and SKF82958
(Di); |i-opioid agonist morphine; muscarinic cholinergic agonist arecoline; glutamate agonist
NMD A; and NMDA antagonist MK801 were administered after 50 FI sessions. FI performance
was altered by all drugs tested except NMDA. Pb exposures attenuated the decrements in rates
produced by the two DI agonists and, at 150 ppm Pb exposure, altered the rate change associated
with the low dose (0.03 mg/kg) of quinpirole. The DA agonists' effects were not concentration-
dependent. These data suggest that Pb-induced changes in behavior were mediated by DI
receptors. Additional evidence suggesting Pb's attenuation of DA activity was obtained using
the D2 agonist quinpirole and the D2 antagonist eticlopride (Areola and Jadhav, 2001). Post-
weaning, rats that had been exposed to 50-ppm lead acetate, producing a blood Pb level of
15.1 |ig/dL, were tested on an FI 1-min schedule. Quinpirole at 0.05 mg/kg reversed the effects
of Pb, while eticlopride (0.01 and 0.05 mg/kg) had no effect on response rates in Pb-treated
animals.
To test the hypothesis that elevated nucleus accumbens (NAC) DA is a mechanism of
Pb-induced changes in FI performance, NAC DA activity was evaluated using the DA antagonist
N-ethoxycarbonyl-2-ethoxy-l,2-dihydroquinone (EEDQ) (Cory-Slechta et al., 1998). LE rats
were exposed to 0, 50, or 500 ppm Pb acetate continuously from weaning, creating blood Pb
levels of 2.1/0.5, 7.2/9.6, and 49.1/49.4 |ig/dL (~3 months of exposure/end of experiment).
After shaping (lever press), rats performed an FI 1-min schedule of reinforcement for at least
50 sessions. DA increased FI rates in the 0 and 50 ppm groups and decreased rates in the
500-ppm group. Intra-NAC administration of EEDQ suppressed FI response rates. At the
highest EEDQ dose, Pb at 500 ppm delayed recovery of response rates to control level rates,
suggesting that NAC DA activity may be one mechanism mediating FI response rates. Using a
similar exposure paradigm, Cory Slechta et al. (2002a) examined the involvement of the
dorsomedial striatum in Pb-induced increases in FI response rates. Both DA and EEDQ,
microinjected into the dorsomedial striatum, increased or decreased FI response rates, which
depended on baseline FI overall rates. DA mimicked the effects of Pb in this region. At this
point, it is unclear whether this area of the striatum modulates Pb-induced changes in FI
performance. Changes in FI performance were also used to characterize interactions between
chronic Pb exposure and intermittent stress (Virgolini et al., 2005), discussed below.
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Rice (1988a) orally dosed cynomolgus monkeys from birth with 2 mg/kg/day of Pb
acetate continuously throughout the study. At PND 100, blood Pb levels peaked at 115 |ig/dL
and declined to 33 |ig/dL by PND 270. At PND 60, the monkeys were tested on an FR schedule,
learning to respond by pushing a button to receive a reinforcement. The Pb-treated monkeys at
2.5 to 5.0 months of age demonstrated increased mean FR pause times compared to controls.
In later, but not earlier sessions, FI pause was decreased in Pb-treated monkeys. Following FR
training, the monkeys were tested on a discrimination reversal, then a chain FR Imin-FI 2 min
operant schedule, which required the monkeys to complete the FR, followed by the FI to receive
one reinforcer (chain FI-FR). The monkeys were then tested as juveniles (3 years of age) on a
multiple FI-FR schedule of reinforcement. Pb-treated juvenile monkeys demonstrated increased
FI run rate, pause time, and index of curvature. At both ages, the treated monkeys showed
increased variability of performance (both within and between sessions, and between subjects)
compared to controls.
To evaluate the effect of Pb exposure during different developmental periods, Rice
(1992a) exposed cynomolgus monkeys to 1.5 mg/kg/day Pb acetate either continuously from
birth, from birth to PND 400, or from PND 300 onward. These exposures resulted in a steady
state blood Pb level of 20 to 35 |ig/dL during dosing. Tested at 3 years of age on a multi FI-FR,
the Pb-treated monkeys showed no effects on FI rate. Tested at 7 to 8 years of age, all three
groups of treated monkeys demonstrated increased run rates and decreased interresponse times
on the FI. To explain the negative results in the juveniles and the positive results in the adults,
the author postulated a possible interaction of Pb with the behavioral history. The monkeys had
been tested first with a multi FI-FR, then a differential reinforcement of low rate (DRL)
schedule, a series of nonspatial discrimination reversal tasks, a delayed spatial alternation (DSA)
task, and then a second multi FI-FR at 7 to 8 years of age. Additionally, the author stated that FI
performance can be affected even without exposure to Pb during infancy and that exposure only
during infancy is sufficient to affect responses.
A concurrent schedule of reinforcement was used to test squirrel monkeys exposed
gestationally to Pb (Newland et al., 1994). Maternal blood Pb levels ranged from 21 to
79 |ig/dL. At 5 to 6 years of age, the monkeys were tested using a concurrent reinforcement with
VI schedules. The monkeys were allowed to respond on either of two levers, one of which had a
greater density of reinforcement than the other. The ratio of reinforcement density was changed
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within the test session trial. Control monkeys learned to follow the reinforcement density by
responding to a greater degree to the lever associated with the richer density. Monkeys whose
blood Pb level in utero had been >40 |ig/dL changed their responses more slowly or in the wrong
direction to the changing reinforcement, suggesting to the authors that this faulty response to
changes in reinforcement may be one mechanism of learning impairment.
Schedule-controlled behavior in squirrel monkeys was assessed at ages 3 to 7 years
following in utero-only exposure to Pb acetate (Newland et al., 1996). Doses were adjusted
individually to provide a maternal blood Pb level of 21 to 70 |ig/dL. The monkeys were trained
to pull a 1-kg weighted bar on FR and FI schedules of reinforcement. Pb-treated monkeys
demonstrated an increase in the number of responses that failed to adequately displace the bar.
This increase in incomplete responses occurred on FR schedules, but not on the FI schedule.
The enhanced sensitivity of the FR schedule may be attributed to it requiring a fixed number of
responses for reinforcement, unlike the FI schedule, which reinforces only one response for
reinforcement. Because the monkeys had to use greater physical force to complete the response
than the monkeys in the studies discussed above, this study identified a deficit in the physical
execution of the response. The lack of increased response rate could also be related to the
physical effort required. These data suggested to the authors that gestational exposure to Pb can
produce motor impairments long after exposure has ended and that these motor impairments
accompany deficits in acquisition behavior.
In both rats and monkeys, an increased rate of FI responding has been seen with Pb
exposures producing blood Pb levels as low as 11 |ig/dL. Figure 5-6 shows a graph summarizing
studies examining Pb-induced changes in FI response rates (Cory-Slechta, 1994). This figure
summarizes the dose effect function for Pb-induced changes in FI performance for several
species. Low-level Pb exposures increase FI response rates and high-level Pb exposures
decrease FI response rates. These data extend earlier findings of a curvilinear dose-response
relationship for this endpoint.

Differential Reinforcement of Low Rates
On the basis of the results of the FI testing done on monkeys described above, Rice
(1992b) used a DRL to assess whether the monkeys could learn to inhibit inappropriate
responding. Monkeys were exposed to 2 mg/kg/day Pb acetate, creating a steady state blood Pb
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300i
P
E
R
C
E
N
T

O
F

C
O
N
T
R
O
L
250'
1.0
10.0 100.0
Pb DOSE (mg/kg)
1000
Figure 5-6. Dose-effect function for lead-induced changes in fixed-interval performance.
The lead effect (response rate, interresponse time, or percentage of
reinforcement) was plotted as a percentage of the control group value for
sessions in which peak effects were observed. Different symbols represent
different experimental species: circles, rats; triangles, monkeys; squares,
sheep; diamonds, pigeons. Numbers next to curves or selected points
represent data from the following studies: (1) Cory-Slechta et al., 1983;
(2) Cory-Slechta and Thompson, 1979; (3) Van Gelder et al., 1973; (4)
Barthalmus et al., 1977;(5) Zenick et al., 1979; (6) Rice et al., 1979; (7) Angell
and Weiss, 1982; (8) Cory-Slechta and Pokora, 1991; (9) Cory-Slechta et al.,
1985; (10), Cory-Slechta and Weiss, 1989; (11) Nation et al, 1989; (12) Rice
1992a; (13) Rice, 1988b.

Source: Copyright 1994 from Principles ofNeurotoxicology by Louis W. Chang, Ed. Reproduced by permission
of Routledge/Taylor & Francis Group, LLC.
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level of 33 |ig/dL after withdrawal of infant formula. Compared to controls, the Pb-treated
monkeys demonstrated greater non-reinforced responding, less reinforced responding, and had a
shorter average time between responses. These results suggested to the author that Pb exposure
may cause a failure to inhibit inappropriate responding, a poorer ability to use internal cues for
timing, or an alteration in timing capabilities.

Radial Arm Maze and Passive Avoidance
The radial arm maze evaluates spatial acquisition and retention by measuring animals'
retrieval of food from the arms of the maze. A study was done to determine if Pb-induced
behavioral deficits in this learning endpoint are related to injury of hippocampal neurons (Munoz
et al., 1988). Female Wistar rats were fed 750-ppm Pb acetate in their diet and then bred after
50 days. Pups were either continued on the same Pb diet for permanent exposure or fed control
chow for maternal-only exposure. Blood Pb values at PND 16 were 17.3 |ig/dL in Pb-exposed
rats and ranged from 32 to 39 |ig/dL in continuously dosed animals. Brain Pb levels were
7.3 |ig/g at PND 16 in Pb-exposed animals. Hippocampal lesions, consisting of complete
bilateral depletion of granular and pyramidal cells in dorsal hippocampus, were induced in other
rats by stereotaxic injection of ibotenic acid. The lesioned animals showed no effects on
acquisition of learning in the radial arm maze, while the Pb-exposed animals did. Tested
4 weeks later, both lesioned and Pb-treated animals showed impaired retention, suggesting to the
authors that Pb may damage the dorsal region of the hippocampus and may be associated with
the retention component of learning. A subsequent study by the same group (Munoz et al.,
1989), using a similar Pb exposure protocol and ibotenic acid lesions to the amygdala, was done
to determine if that brain region was involved in Pb-induced learning deficits. Both treatments
impaired both acquisition of food-retrieval behavior in the maze and passive avoidance behavior,
but neither treatment affected locomotor activity. The permanently exposed rats showed greater
deficits, indicating possible reversibility of Pb-induced effects in prenatal-only exposures or the
cumulative effects of the chronic exposure creating a greater body burden of Pb. Further, these
data point to the amygdala as another Pb target.
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Discrimination
Early studies demonstrated Pb-induced impairments in tests of discrimination of relevant
environmental stimuli. To study the effects of chronic Pb exposure on discrimination, Morgan
et al. (2000) exposed LE rats continuously from the beginning of gestation. Lead acetate at
0, 75, or 300 ppm in drinking water produced adult blood Pb levels of <5, 20, and 36 |ig/dL.
At ages 7 to 9 weeks, an automated, three-choice visual discrimination task revealed a dose-
dependent slowing of learning and an increased incidence of "impaired" animals. The authors
concluded that chronic developmental Pb exposure results in associative deficits and an
increased tendency to respond rapidly. In another study, Morgan et al. (2001) evaluated Pb-
induced alterations in visual discrimination in LE rats exposed only during early development.
One group received Pb acetate in drinking water throughout gestation and lactation (GL300),
other groups received 300 or 600 ppm Pb during lactation only (L300 and L600). Blood Pb
levels were <5 (controls), 36-43 atPND 8, 27-34 atPND24, 131-158 atPND 53, and
16-18 |ig/dL at PND 53 (treated animals). Pb-treated animals showed no differences in learning
rate, motivation, or response latency for correct or incorrect responses.

Discrimination Reversal
Discrimination reversal studies examine the ability to alter behavior in response to a
change in reinforcement contingencies. Earlier studies showed that chronic low-level exposures
in monkeys, creating a steady state blood Pb level of 11 to 15 |ig/dL, produced deficits in
nonspatial discrimination reversal tests, with and without irrelevant cues. Reversal of previous
learning appears to be consistently affected by Pb exposure, often resulting in perseverative
behavior. Additionally, in early cynomolgus monkey studies, the Pb-treated monkeys were
found to be more distracted by irrelevant cues than control monkeys.
Using this same cohort of cynomolgus monkeys, Gilbert and Rice (1987) examined
spatial discrimination reversal at 9 to 10 years of age. The monkeys had been exposed to 50 or
100 jig/kg/day Pb acetate, resulting in blood Pb peaks of 15.4 and 25.4 |ig/dL, respectively.
Steady state blood Pb levels were 10.9 and 13.1 |ig/dL, respectively. Compared to controls, the
treated monkeys were impaired in the presence, but not the absence, of irrelevant cues. In the
lower-dose group monkeys (blood Pb 10.9 |ig/dL), impairment ended when the irrelevant stimuli
became familiar.
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To evaluate the effects of timing of exposure on this learning task, monkeys were exposed
to one of three protocols: (1) 1.5 mg/kg/day Pb acetate continuously from birth; (2) during
infancy only (birth to PND 400); or (3) beginning after infancy (Pb from PND 300 and
thereafter) (Rice and Gilbert, 1990a). These exposures resulted in blood Pb levels of 32 to
36 |ig/dL during dosing and given infant formula and levels of 19 to 26 |ig/dL during Pb
exposure in the post-infancy group. At age 5 to 6 years, the monkeys were tested on a nonspatial
discrimination reversal task. Monkeys exposed continuously and those exposed beginning after
infancy demonstrated a dose-dependent impairment of learning. The monkeys exposed only
during infancy showed no impaired learning of the task. At 7 to 8 years of age (adults), the
monkeys were tested on spatial discrimination tasks (Rice, 1990). Using no irrelevant cues or
irrelevant form and color cues, results showed that the continuously exposed monkeys were
impaired in the absence of irrelevant cues. All three treatment groups were impaired when
irrelevant cues were present. These results are in contrast to the results on the nonspatial tasks
described by Rice and Gilbert (1990a) and suggested to the author that the developmental period
of exposure may differentially affect spatial and nonspatial tasks.
The effects of Pb on olfactory reversal discrimination have been examined in two rodent
studies. Hilson and Strupp (1997) exposed LE rats chronically from conception to 0, 75, or
300 ppm Pb acetate in water. Blood Pb levels were, respectively, <5, 26, and 51 |ig/dL on
PND 1; <5, 22, and 37.5 |ig/dL on PND 17; and <5, 27.5, and 51 |ig/dL in adulthood. At 20, 14,
and 22 weeks of age for the three replicates, testing consisted of acquisition of the original
discrimination (i.e., learning to respond to an odor), then reversal learning in which the other
odor became correct. Pb treatment did not affect learning the original discrimination; however, it
did impair learning the reversals in the high dose group by prolonging the postperseverative
phase, which the authors note is similar to the effects seen with lesions of the amygdala. Both
groups of Pb-treated rats also showed impairment during an extradimensional shift in which the
rats had to learn the correct spatial location of the odor for reward. Garavan et al. (2000) further
investigated these responses in LE rats exposed to: (1)0 ppm Pb; (2) 300 ppm Pb acetate during
both gestation and lactation (GL300); (3) 300 ppm during lactation (L300); or (4) 600 ppm
during lactation to (L600). At testing on PND 53, blood Pb levels were <5, 16, 12, and
18 |ig/dL, respectively. Compared to controls in a two-choice olfactory serial reversal task, all
Pb-treated groups needed more trials to reach the point at which perseverative responding to the
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previously correct cue ended. The authors hypothesized that this deficit was not due to
perseverative responding, but rather to a Pb-related spatial response bias and a concurrent, but
independent, associative impairment.

Learning Set Formation
Learning set formation tasks evaluate an animal's ability to "learn to learn"
discriminations by presenting a series of visual discrimination problems and quantifying the rate
at which each successive problem is learned. To ascertain Pb's effects on this learning endpoint,
Lilienthal et al. (1986) exposed rhesus monkeys to either 0 (control), 350 (low,) or 600 ppm
(high) Pb in utero, resulting in blood Pb levels of 50 and 110 |ig/dL, respectively. At age 12 to
15 months, only the high-dose group exhibited deficits in simple discrimination learning, which
the authors attributed to possible Pb-induced reduced attention, higher frustration levels, and
lowered adaptability. Both groups showed impairments in forming a learning set, which the
authors hypothesize was due to Pb-induced cognitive deficits.

Concurrent Discrimination
Concurrent discrimination tests an animal's ability to learn a second set of problems after
a first set was learned concurrently to criterion. Rice (1992c) further evaluated the monkeys
used in the discrimination reversal studies described by Rice and Gilbert (1990a). At 8 to 9 years
old, the monkeys were tested on two sets of concurrent discrimination tasks. Pb-treated monkeys
in all three exposure groups learned more slowly, although there was less impairment in the
monkeys exposed only during infancy. Perseverative behavior was also demonstrated in the
treated monkeys. These data are consistent with the other discrimination studies.

Repeated Acquisition and Performance Schedule
Another test method effectively utilized to distinguish changes in chronically Pb-exposed
animals is the repeated acquisition and performance schedule (Cohn et al., 1993). The purpose
of this test is to determine the selectivity of Pb-induced changes in learning, as distinct from
nonspecific or performance effects, and to explore the nature of the underlying error patterns
contributing to any learning deficits. This schedule required completion of a sequence of three
responses for reinforcement, with the correct sequence for the learning component changing with
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each successive experimental session (i.e., repeated acquisition). In contrast, the correct
sequence remained constant across sessions in the performance component; thus, once learned,
this component did not require further learning to complete successfully.
This schedule was used in animals chronically exposed from weaning to Pb acetate at
0, 50, or 250 ppm in drinking water, producing blood Pb levels of 2.8, 25.1, and 73.5 |ig/dL,
respectively. Pb treatment caused significant decrements in accuracy on the learning component,
but not on the performance component, compared to controls (Cohn et al., 1993). A detailed
analysis of the subjects' behavior indicated that Pb exposure impaired learning by increasing
perseverative responding on a single lever, even though such repetitive responding was not
directly reinforced. In a subsequent study using the same exposures, dose-effect curves for the
NMDA receptor antagonist MK-801 were determined in controls and animals in which chronic
Pb exposure began at weaning (Cohn and Cory-Slechta, 1993). The decline in learning accuracy
and the increases in perseverative responding produced by MK-801 were attenuated by Pb
exposure, and dose-effect curves relating MK-801 dose to changes in rates of responding were
shifted to the right in Pb-exposed rats compared to control animals. These observations
demonstrate a sub sensitivity of Pb-exposed animals to both the accuracy-impairing and response
rate-altering properties of the antagonist. An additional investigation used the same Pb exposure
protocol and administration of doses of NMDA as a receptor agonist to rats performing this test
(Cohn and Cory-Slechta, 1994a). In control animals, NMDA was found to decrease accuracy of
response in both the repeated acquisition and performance components of this multiple schedule
and to suppress response rates as well. Pb exposure potentiated the accuracy-impairing effects of
NMDA by further increasing the frequencies of errors and likewise potentiated the drug's
rate-suppressing effects. Thus, as stated earlier in this section, the Pb-induced potentiation of the
agonist effects and reduced sensitivity to the antagonist effects in this test are consistent with a
functional upregulation of NMDA receptors in Pb-exposed brain. In other work, Cohn and
Cory-Slechta (1994b) were unable to distinguish any evidence of dopaminergic modulation of
responding in this behavioral paradigm. Thus, the repeated acquisition and performance
schedule proved valuable not only in providing a finer dissection of the animal's behavior, but
also in elucidating important mechanistic aspects of Pb neurotoxicity. It also provides an
unambiguous indication of an adverse effect (learning impairment) in the absence of sensory and
motor deficits.
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Avoidance Learning
Altmann et al. (1993) examined the effects of Pb on active avoidance learning (AAL) in
the offspring of Wistar rats fed 750-ppm Pb acetate for 50 days prior to mating. Deficits in AAL
were demonstrated in rats exposed to Pb either during pre-weaning or pre- and postweaning,
creating a blood Pb level of 15 |ig/dL and a brain Pb level of 0.09 to 0.16 jig/g wet weight. Rats
that received only postweaning exposure (blood Pb of 16 |ig/dL and brain level of 0.09 |ig/g) had
reduced deficits in AAL. Another study of AAL (Chen et al., 1997a) used SD rats exposed to
0.2% Pb acetate either during gestation and lactation (gestation day to until PND 21), during
postweaning only (PND 21 until testing at PND 56), or continuously. Blood Pb levels at
PND 56 were <2, 3.8, 25.3, and 29.9 |ig/dL, respectively. No Pb-associated effects in learning
were seen with just maternal or postweaning exposure. Compared to controls, continuously
exposed rats showed a tendency of lower avoidance and higher no response levels in the two-
way active avoidance tasks. Chen et al. (2001) tested step-down passive avoidance learning in
SD rats similarly exposed. All three Pb-treated groups tested at PND 55 to 56 demonstrated
impaired learning but unimpaired retention. Results of parallel autoradiographic analyses
suggested that the Pb-induced deficits in acquisition were associated with alterations in AMPA
receptor binding.

Open Field Performance
Salinas and Huff (2002) compared learning in Pb-exposed spatially trained and cue-
trained F344 rats using an open field arena. The chronically exposed rats were tested at
-29 weeks of age when blood Pb levels were -42 |ig/dL. The Pb-treated rats trained to find food
using extra-maze spatial cues demonstrated better performance than either controls or the
Pb-exposed rats trained using intra-maze discrete cues. Additionally, by the seventh day of
testing, both groups of Pb-treated rats spent less time on the periphery of the maze. The authors
(a) hypothesized that "a Pb-induced overflow of mesolimbic DA may have facilitated the
expression of rearing behaviors," which assisted in the spatial learning and (b) suggested that this
overflow may cause "impulsivity" that results in less time spent by the Pb-treated rats on the
periphery.
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Effects ofPb on Memory
Studies reported in the 1986 AQCD showed that monkeys with steady state blood Pb
levels of 33 |ig/dL and tested at 3 to 4 years of age had impaired delayed matching-to-sample in
both spatial- and nonspatial-based paradigms. However, Cory-Slechta (2003) pointed out that
much of this work was done with animals that had previous behavior testing done with them
and that may have altered results. New studies characterizing Pb-induced deficits in
matching-to-sample tests have not been found. Other studies evaluating Pb-induced deficits in
memory are described below.

Morris Water Maze
The Morris water maze tests both learning and memory by requiring rats to learn and
remember the position of a platform hidden in opaque water. The effect of chronic
developmental Pb exposure on water maze performance was tested in LE rats (Jett et al., 1997).
Dams were dosed with 250-ppm Pb acetate in feed from 10 days before breeding through
lactation. Offspring continued on the same Pb-dosed chow through testing. Blood Pb levels
were not reported. Hippocampal Pb levels were 1.73 (PND 21), 1.02 (PND 56), and 0.91 |ig/g
wet weight (PND 91). Reference (long-term) memory and working (short-term) memory were
tested on PND 21, 56, and 91, with different rats in each test. Pb-exposure had no effect on
working memory at any age tested, but did affect reference memory (significant in females and
nearly significant in males) in the PND 21 rats. This group (Jett et al., 1997) also demonstrated
an increase in escape latency in adult LE rats injected with Pb acetate directly into the dorsal
hippocampus. A study examined the effects of timing of Pb exposure (Kuhlmann et al., 1997) in
which rats were exposed to 750 ppm Pb acetate in diet either maternally (gestation and lactation),
permanently (gestation onward), or postweaning (at 750 or 100 ppm). Blood Pb levels at
PND 100 were 1.8, 21.3, 22.8, and 26.3 |ig/dL, respectively. Compared to controls, maternal
and permanent exposure groups were impaired in water maze performance, with maternal
exposure producing both the greatest escape latency and longest escape path length. There were
no effects on performance in the postweaning exposure groups. This study provides evidence
that early exposure can produce long-term deficits in learning and memory.
Yang et al. (2003) exposed Wistar rats gestationally to 0.03% (low), 0.09% (middle), or
0.27% (high) Pb acetate in food. Pups were fostered by control dams. Blood Pb levels were
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-30, -33, and -42 |ig/dL, respectively at PND 0 and -2 |ig/dL in all treatment groups at testing
(PND 49). By measuring swim path and time spent in target quadrants, it was shown that Pb
exposure at all three doses impaired memory retrieval in males. In female offspring, only the
low dose affected memory retrieval, suggesting a greater impact of low level Pb exposure on
females. Results on the fixed location/visible platform tasks showed that motor performance and
vision were not affected by Pb treatment and suggested to the authors that gestational exposure is
sufficient to cause memory deficits in young adults.
Results from recent studies have provided evidence that environmental enrichment during
development may protect against Pb-induced effects on learning and memory deficits. Schneider
et al. (2001) exposed male LE rats to 0 or 0.2% Pb acetate from PND 25 until later testing at
PND 100. At PND 25, some of the rats were raised in isolation with no access to any stimulus
objects, while the "enriched" rats were raised in groups of 8 with stimulus objects or toys. Blood
Pb levels were -5 |ig/dL in controls and -30 in Pb-treated animals. Pb-exposed rats raised in
isolation demonstrated spatial learning deficits in the Morris water maze, whereas the
Pb-exposed rats raised in the enriched environment performed better than the isolated Pb group.
Additionally, Pb-exposed rats had diminished hippocampal levels of the neurotrophic factors
brain-derived neurotrophic growth factor (BDNF), nerve growth factor-p, neurotrophin-3, and
basic fibroblast growth factor. This suggested to the authors a possible relationship between Pb
levels, neurotrophic factor levels, and diminished hippocampal development and function.
Earlier exposures to Pb showed some similar effects of environmental enrichment (Guilarte
et al., 2003); LE rats were exposed during gestation and lactation to 0- or 1500-ppm Pb acetate.
Pb exposure was stopped at PND 21 and rats were placed in isolation or in an enriched
environment. At PND 50 when blood Pb levels were 0.25 and 3.9 |ig/dL, respectively, testing in
a water maze showed enhanced performance of the Pb-treated rats raised in the enriched
environment. The environmental enrichment was accompanied by increased gene expression in
the hippocampus of NMD A receptor subunit 1 and BDNF. These results demonstrate that
Pb-induced learning impairments and molecular changes in hippocampus can be reversed by
environmental enrichment, even after the exposure has occurred.
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Delayed A Iternation
Delayed alternation trials assess both memory and attention by requiring the animal to
alter a previous response following a delay period. Typically, the longer the delay between
choices, the greater the inaccuracy of the choice. A number of studies have examined Pb effects
on delayed alternation tasks in rodents. To compare the effects of Pb exposure on memory
during different stages of life, rodent studies of this endpoint were completed (Cory-Slechta
et al., 1991). PND 21 (young), 8-month-old (adult), and 16-month-old (old) F344 rats were
exposed to 0, 2, or 10 mg/kg Pb acetate, resulting in blood Pb levels of-23 at the 2 mg dose and
of-42 (adult), -48 (old), and -58 |ig/dL (young) at the 10 mg dose. Training began after
4 months of exposure and testing continued until 8.5 months of exposure. Rats were trained with
many sessions using a cue light alternating between two positions and tested after 4 months of
exposure. Aging itself caused impaired accuracy. In both young and old rats, Pb exposure
increased accuracy, at the longest delay periods (12 sec) in young rats and at the short delay
periods in old rats. In adult rats, performance was not affected by Pb exposure. The authors
hypothesized that the improved performance was due to perseveration of alternation behavior
learned during the training and that both young and old animals may have enhanced vulnerability
to Pb. The effect of chronic postweaning Pb exposure (0, 75, or 300 ppm in drinking water
starting at PND 25) on DSA was evaluated in LE rats (Alber and Strupp, 1996). Blood Pb levels
were 19 and 39 |ig/dL. At 22 weeks of age, nose poke training was followed by cued alternation,
spatial, and DSA training. Learning the alternation rule was not affected by Pb treatment.
Across all delay periods, the treated rats performed more poorly than controls, suggesting that
memory was not affected by Pb exposure. An additional finding was that Pb induced side bias, a
strategy commonly adopted by rats in response to an insoluble problem.
Nonhuman primate DSA studies have also been completed. Levin and Bowman (1986)
evaluated DSA in 5- to 6-year-old rhesus monkeys exposed to two early pulses of 10 mg/kg Pb
acetate and a chronic exposure of 0.7 mg/kg/day for the first year. The pulses of Pb were done to
simulate brief periods of higher-level exposure that can occur in children with Pb pica. Peak
blood Pb levels were 250 to 300 |ig/dL, which decreased to 80 |ig/dL for the remainder of the
first year. Deficits occurred most commonly with short inter-trial delays, suggesting that
memory was not affected by Pb, but that deficits in attention or strategy may have been present.
Most Pb-induced deficits were accounted for by lose-shift errors, possibly due to perseveration
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on the alternation strategy despite loss of reinforcement. The authors hypothesized a "window of
sensitivity" to Pb, wherein exposures during the first year of life can create long-term cognitive
deficits. Another cohort of rhesus monkeys was dosed with 1.0 mg/kg/day Pb acetate for the
first year of life, producing a blood Pb level of 70 |ig/dL (Levin and Bowman, 1989). They were
tested on DSA at 4 years of age, by which time blood Pb levels had returned to control levels.
The monkeys performed better than controls on choice accuracy, which suggested to the authors
that a Pb-induced decrease in attentiveness may make the monkeys less susceptible to irrelevant
stimuli and thus able to perform better.
Rice and Karpinski (1988) evaluated the effects of low-level lifetime exposure on delayed
alternation tasks. Cynomolgus monkeys were dosed with 0, 50, or 100 jig/kg/day Pb acetate
from PND 1 through the completion of the study. Blood Pb levels peaked at PND 100 (~3, 15.4,
and 25.4 |ig/dL, respectively) and reached steady state by PND 300 (3, 10.9, and 13.1 |ig/dL).
When tested at 7 to 8 years of age, using delay values that increased over the course of the trial,
Pb-induced impairment of initial acquisition of tasks was observed. Longer delays between
alternations resulted in poorer performance by the Pb-treated monkeys and perseverative
behavior, sometimes lasting for hours. Further, the treated monkeys repeatedly pounded buttons
indiscriminately, suggesting to the authors a failure to inhibit inappropriate responding. Rice and
Gilbert (1990b) attempted to evaluate the effect of timing of exposure on this endpoint. The
cynomolgus monkeys described above (Rice and Gilbert, 1990a) that were exposed to Pb either
continuously from birth, during infancy only, or beginning after infancy were given a set of DSA
tasks at 7 to 8 years of age. Monkeys had to push two buttons alternately with delays increasing
from 0.1 to 15 sec. All Pb-treated groups had the same impairments of initial acquisition,
indiscriminate responding, greater impairment with longer delays, and perseverative responses.
All three exposure groups had similar degrees of impairment, indicating to the authors a possible
lack of sensitive period for Pb's affects on this endpoint.
Levin et al. (1987) examined potential involvement of cholinergic (ACh) and DA
neurotransmitter systems in Pb-induced impairments of DSA performance. Rhesus monkeys
were exposed during the first year of life to daily doses of 0.7 mg/kg/day plus two early high
pulses (10 mg/kg/day during the second and fourth weeks of life). Blood Pb levels were 63 at
weeks 1 to 4, 174 at weeks 5 to 10, 68 at weeks 11 to 52, 4 at 6 years of age, and 2 |ig/dL by the
time of testing (7 to 9 years of age). Thirty min prior to testing, one of the following drugs was
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administered: the DA receptor blockers, haloperidol and sulpiride; the muscarinic ACh receptor
blocker, scopolamine; or the DA agonist, amphetamine. In addition, amphetamine and L-dopa
were injected for 2 weeks for chronic testing. In both controls and Pb-treated monkeys,
scopolamine caused a dose-related decline in performance. Chronic L-dopa ameliorated the
Pb-induced DSA deficits, which returned following cessation of L-dopa administration. These
data again show long-term cognitive impairments resulting from early Pb exposures and
implicate DA mechanisms as a causal factor in these impairments.

Avoidance
A shuttle avoidance task evaluated retention in rats exposed during gestation and lactation
to 0.5, 2.0, or 4.0 mM lead acetate in drinking water (Rodrigues et al., 1996a). Blood Pb levels
ranged from 11 to 50 |ig/dL. The Pb-treated rats demonstrated no increases in avoidance
response between sessions, suggesting less retention in the treated animals compared to controls.
Murphy and Regan (1999) exposed Wistar rats from PND 1 to PND 30 to 400 mg of Pb
chloride/L drinking water, producing blood Pb levels of 10 to 15 |ig/dL by PND 8, -45 |ig/dL by
weaning, and 2 to 4 |ig/dL by PND 80. At PND 80, the rats were trained on a 1-trial, step-
through, light-dark passive avoidance test. At 48 h postexposure, the rats showed no Pb-induced
changes in recall; but at 5 days postexposure, the rats exhibited a decline in recall. The authors
hypothesized that the Pb exposure affected long-term memory storage.

Discrimination Retention
Munoz et al. (1986) used visual discrimination learning and spatial learning in a retesting
approach to evaluate changes in long-term memory storage in Wistar rats. The rats were
exposed to 750-ppm Pb acetate either through PND 16 (maternal exposure) or chronically
through testing. Blood Pb in males tested at PND 110 was <1 |ig/dL in controls and maternally
exposed rats and 34 |ig/dL in chronically exposed rats. Males were tested at PND 100 in visual
discrimination and then retested 42 days later. Both Pb-treated groups learned the original
discrimination comparably to controls, but both groups showed a deficit in retention of the
discrimination. Females were tested at PND 180 in spatial discrimination, then retested 4 weeks
later. The Pb-treated female rats took longer to reach criterion in the acquisition learning and
longer to eat the pellets in the retention phase. These data suggest that early gestational Pb
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exposure can cause long-lasing retention deficits and that subsequent direct dietary exposure
does not potentiate these effects.

Effects ofPb on Attention
Human studies have suggested that Pb exposure is associated with deficits in attention
(e.g., inattentiveness, impulsivity, distractibility, attention deficit disorder); however, only scant
evidence from early animal studies pointed to attention deficits as a causal factor in
neurocognitive dysfunction. Related to deficits in attention are increased response rates and
response perseveration, both of which are associated with Pb exposure as discussed above.
Several laboratories have examined the effect of Pb on attention specifically. Brockel and
Cory-Slechta (1998) exposed male LE rats to 0-, 50-, or 150-ppm Pb acetate in water from
weaning. After 3 months of exposure, blood Pb levels were <5, 10.8, and 28.5 |ig/dL,
respectively. After 40 days of exposure, the rats were trained on a FR/waiting-for-reward
behavioral baseline, learning to produce food delivery by pressing a lever 50 times. Additional
food could be earned by withholding lever presses (i.e., by waiting); free food was given at
increasing time intervals after completion of the FR. In the FR component of the study, the
high-dose animals had significantly higher response rates and more frequent resets of the waiting
period than the low dose group and controls. Wait time was significantly lower in both treated
groups compared to controls in the waiting behavior component. The high-dose animals also
had an increased number of reinforcers and a higher response to reinforcement ratio than low
dose and controls. These data suggest that blood Pb levels as low as 11 |ig/dL are associated
with inefficient response patterns and an inability to manage delays of reinforcement. The
authors hypothesized that this pattern of behavior in humans could have the consequence of
eventual dissipation of effort or lack of motivation.
Involvement of dopamine-like receptors in the Pb-induced decrements in waiting behavior
was tested using this FR/waiting-for-reward schedule (Brockel and Cory-Slechta, 1999a). The
same 0, 50, and 150 Pb dosing was done, resulting in blood Pb levels of <5, 9.7, and 26.2 |ig/dL
after both 3 and 7 months of exposure. Following performance stabilization, drugs were
administered IP 30 min prior to behavioral testing: the D2 agonist quinpirole, the DI agonist
SKF 82958, the D2 antagonist eticlopride, or the DI antagonist SCH 23390. The drugs
administered to control rats did not cause Pb-like effects. All the drugs decreased FR response
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rates, but only quinpirole reversed the Pb-induced effects on FR response rate, FR resets, wait
reinforcers, and wait time. This suggested to the authors a role for D2 receptors in Pb's effects,
a dissociation of Pb's effects on FR and waiting time, and a possibility that decreased waiting
behavior is a direct effect of Pb exposure.
A similar postweaning Pb exposure resulting in blood Pb levels of <5, 16.0, and
28.0 |ig/dL was used to evaluate sustained attention (Brockel and Cory-Slechta, 1999b). Food
rewards were obtained by the rats for discriminating correctly between a target and distracter
light. A 13-sec time-out was given for incorrect responses. Lead produced no effects on
sustained attention despite broad modifications of many parameters of the task, suggesting to the
authors that Pb affects other aspects of attention and that Pb-induced attention deficits are
modified by time-out contingencies and reinforcement. Morgan et al. (2001) evaluated changes
in attention in a study with LE rats exposed to Pb acetate in drinking water. One group (GL300)
received Pb throughout gestation and lactation (GL), other groups received 300 or 600 ppm Pb
during lactation only (L300 and L600). Blood Pb levels were <5 for controls, 36 to 43 at PND 8,
27 to 34 at PND 24, 131 to 158 at PND 53, and 16 to 18 |ig/dL at PND 53 (treated animals).
Lead-exposed animals (both GL and L) committed more errors of omission when a delay was
imposed prior to cue presentation and in trials that followed an incorrect response. Response
initiation was also impaired in Pb-treated animals in a sustained attention task in which the onset
and duration of the visual cue varied randomly across trials. These data suggested to the authors
that early Pb exposure caused long-lasting increased reactivity to errors and impairment of
sustained attention. Inconsistencies in these results compared to those of Brockel and Cory-
Slechta (1998,1999b) may be accounted for by the differences in exposure period and the higher
blood Pb level and the small percentage change in endpoints in the later study. In a review of the
attention literature, Cory-Slecta (2003) stated that impulsivity and waiting-for-reward behavior
may be more strongly affected by Pb than are sustained attention and hyperkinesis. Further, the
Pb-induced impulsivity as a behavioral dysfunction may lead to cognitive impairments.

Effects ofPb on Motor Function, Locomotor Activity, and Vocalization
Evaluations of the effects of Pb on development of motor function and reflexes discussed
in the 1986 AQCD showed that Pb affects the air righting reflex in rat pups with blood Pb levels
of 35 |ig/dL and rotarod performance at 175 |ig/dL. Developmental lags in gross activity were
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produced in rat pups with blood Pb as low as 14 |ig/dL. Studies ruled out the possible
contribution of Pb-exposed dams to their offspring's slowed development. In the early
evaluations of spontaneous activity, numerous issues were apparent, including a lack of
consensus for the definition of "activity" and activity being affected by confounding variables
such as age, sex, estrous cycle, time of day, novelty of environment, and food deprivation.
Discrepant findings summarized in the 1986 Lead AQCD included 11 studies showing increased
activity with Pb exposure, 6 studies showing decreased activity, and 28 studies showing age-
dependent, qualitative, mixed or no changes. Thus, no conclusions were reached regarding Pb
effects on these endpoints.
Lilienthal et al. (1986) evaluated activity in the rhesus monkeys exposed prenatally to Pb
as discussed above. Activity, measured in unfamiliar environments at age 12 to 15 months,
showed no Pb-related effects in these monkeys. In the study of schedule-controlled behavior
discussed above, Newland et al. (1996) identified a deficit in the physical execution of pulling a
weighted bar in Pb-treated monkeys, suggesting a motor impairment occurring long after the
exposure period. Open field behavior was assessed in rhesus monkeys exposed to a
"pulse-chronic" dosing paradigm (two pulses of 10 mg/kg the first month of life and chronic
level of 0.7 mg/kg/day for the rest of the first year of life) (Ferguson and Bowman, 1990). Blood
Pb levels peaked at 5 weeks of age (55 |ig/dL), averaged 36 |ig/dL for the remainder of the first
year, and were <5 |ig/dL for at least 1.5 years prior to testing at 4 years of age. The Pb-treated
monkeys demonstrated a longer latency to enter the open area, increased durations of activity
and environmental exploration, as well as a failure to habituate.
Laughlin et al. (1999) evaluated the effects of Pb on behavior of neonatal rhesus monkeys
using the Schneider Neonatal Assessment for Primates (SNAP). Monkeys were dosed orally
with Pb acetate daily to eventually achieve a target blood Pb level of 35 |ig/dL. Blood Pb levels
for controls was <5 |ig/dL, and for treated animals was 15 to 20 |ig/dL during week 3 and
22 to 28 |ig/dL during week 4. Testing was done during the first four weeks of life, at which
time few differences between control and Pb-exposed monkeys were seen. The authors reported
less stability in SNAP performance in the Pb-exposed monkeys compared to controls, which they
suggested may be caused by a disruption of continuity of development by Pb. The same animals
were evaluated for exploration behavior starting at the second postnatal week (Lasky and
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Laughlin, 2001). The Pb-exposed monkeys demonstrated more agitation, climbing, fear, and
exploration of the periphery than controls.
In the Rodrigues et al. (1996a) study cited above, rats were also subjected to open-field
testing, where a Pb-associated increase in locomotor activity was observed. To assess the effects
of developmental exposure on a number of neurobehavioral endpoints, Wistar rats were exposed
during gestation and lactation to 500-ppm Pb acetate (Moreira et al., 2001). Blood Pb levels
were 41 |ig/dL (dams), 21 |ig/dL (PND 23 pups), and <0.1 |ig/dL (PND 70). The PND 23 pups
exhibited Pb-induced increased ambulation in the open-field tests, decreased exploratory
behavior in the holeboard tests, and no differences from control in the elevated maze tests.
The PND 70 rats showed a Pb-induced increase in head dipping in the holeboard test. No
differences were noted in the rotarod tests. Ferguson et al. (1998) evaluated activity in SD rats
exposed to 350-ppm Pb acetate (through dams' drinking water) from birth until weaning. Blood
Pb was 46 |ig/dL in pups at weaning. No Pb-related effects were seen in the following
behavioral assessments: play (PND 38 and 45), burrowing (PND 49-54), dominance (PND 58
and 65), residential running wheel (PND 67-80), residential figure 8 maze (PND 70-84),
complex maze (PND 83-94), acoustic startle (PND 98), emergence (PND 121 or 128), and
prepulse inhibition (PND 177). The authors suggested that at these Pb exposure levels, the
development of Pb-induced functional changes may require substantial demands on the system
for detection.
To evaluate the effects of early Pb exposure on activity and vocalization, De Marco et al.
(2005) exposed female Wistar rats to 8, 16, or 24 mg/mL lead acetate in water, allowed them to
breed, and then crossfostered pups to them to allow exposure during pregnancy, during
pregnancy and lactation, or during lactation. In the treated rats, blood Pb levels ranged from
5.7 to 36.6 |ig/dL, with levels dropping to 0.5 |ig/dL in adults. In all three exposure groups, the
PND 7 pups showed a dose-dependent decrease in ultrasonic vocalization, whereas the PND 14
pups showed an increase compared to controls. These results contrast with the normal
developmental pattern of vocalization and suggested to the authors a Pb-induced alteration in the
maturation pattern of this behavior. Additionally, the PND 14 pups showed higher activity levels
than controls.
One animal study has shown behavioral deficits in offspring resulting from paternal
exposure to Pb (Nelson et al., 1997). Male Dutch Belted rabbits were dosed with Pb acetate for
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15 weeks to produce blood Pb levels of <5 (control), 20, 40 and 80 |ig/dL. They were then
mated with nonexposed females and the offspring were tested for exploratory behavior at
PND 15, 20, 25, and 30 in a figure-eight activity monitor. For those F2 offspring of male rabbits
that had paternal blood Pb >40 |ig/dL, exploratory behavior was affected at PND 25, the time of
peak activity in rabbits. This is the only animal study available showing these effects. Another
study demonstrated that behavioral effects of Pb can extend to a second generation (Trombini
et al., 2001). First, Pb-induced behavioral effects were assessed in Wistar rats exposed during
gestation and lactation. Pregnant rats were dosed with 750 ppm lead acetate in drinking water.
blood Pb levels in offspring were 25 ng/dL at PND 30 and 0.1 ng/dL at PND 90. No
Pb-associated changes in elevated maze behavior were noted in 30- and 90-day-old rats. In
open-field behavior studies, PND 30 Pb-treated rats showed decreased freezing, increased
ambulation, and increased grooming. PND 90 Pb-treated rats showed decreased freezing and
increased ambulation. Offspring of Pb-treated females were mated with nonexposed males and
evaluated at PND 30 and 90. At both ages, the F2 generation rats demonstrated increased
ambulation and decreased grooming, suggesting evidence of intergenerational effects of Pb.
In an attempt to discern the mechanism of motor deficits resulting from Pb exposure,
Morley et al. (2003) exposed Drosophila instar larvae to 100 jiM Pb acetate and examined the
neuromuscular junction. They observed nonuniform matching between the size of the motor
terminal and the muscle area, suggesting a possible mechanism of Pb's effects on motor
function.

Effects ofPb on Social Behavior
Early work evaluating Pb-induced alterations in social behavior or behavioral interactions
showed inconsistent results. In both rats and monkeys, Pb tended to reduce aggressive behavior.
Pb-treated mice demonstrated gender differences in sexual interaction and social investigation,
which could have been attributed to differences in brain Pb concentrations. Interactions between
mothers and offspring were shown to be affected by Pb exposure. These included increased
clinging by infants, less food-seeking activity by pups, suppressed play behavior, increased time
spent in the nest by dams, and increased retrieval of pups to the nest by dams.
Studies published since the 1986 Lead AQCD have examined the effects of Pb on social
behavior in greater detail. Donald et al. (1986) exposed male and female BK:W mice to
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0.13% Pb acetate in drinking fluid before breeding. At weaning, the pups were continued on the
0.13% Pb chronically. At 18 weeks, brain and femur Pb concentrations in males were 27.6 and
998 jiM Pb/g ash, respectively (controls were 7.5 and 78). At 34 weeks, brain and femur
concentrations in males were 445 and 5364 jiM Pb/g ash (controls 40 and 100). At 34 weeks,
brain and femur concentrations in females were 787 and 4026 jiM Pb/g ash (controls 88 and
334). Blood Pb levels were not reported. Young Pb-exposed female mice habituated more
slowly to the environment, while young Pb-exposed males habituated more rapidly. In general,
in adults, Pb caused an enhancement of social and sexual investigation. The same laboratory
(Donald et al., 1987) evaluated activity resulting from a higher Pb exposure (0.25% chronically
in drinking fluid from conception). Exploratory behavior and social investigation were increased
in both sexes of Pb-treated mice at age 3 to 4 weeks compared to controls. At age 7 to 8 weeks,
social investigation was increased, but exploratory behavior was decreased in Pb-treated mice.
At age 15 to 16 weeks, nonsocial activity was decreased in females, but increased in males. At
age 17 to 18 weeks, Pb-treated males demonstrated shorter latencies to aggression than controls.
Holloway and Thor (1987) tested social behavior in LE rats exposed to 500-ppm Pb
chloride during lactation. They estimated blood Pb to be 42 |ig/dL on PND 20 based on similar
exposures. AT PND 11, they found that Pb induced no sex differences, no effects on pup
activity, and no differences in pup retrieval by dams. At PND 26, Pb treatment influenced all
social behavior tested (i.e., investigation duration and frequency, crossover frequency, pinning)
but did not change activity levels compared to controls. At PND 36, Pb-treated pups
demonstrated increased crossover frequencies but no change in activity levels compared to
controls. The authors hypothesized that low-level Pb exposure increases social investigation and
"rough and tumble" play behaviors, due to increased behavioral reactivity to stimuli, but does not
increase aggression. Pb-induced changes in aggression were assessed in golden hamsters
exposed to 100 ppm lead acetate from gestational day 8 until PND 42 (Delville, 1999). Blood Pb
concentration at PND 42 was 10 to 15 |ig/dL. At PND 19 to 20, the Pb-exposed hamsters were
significantly smaller than controls and exhibited less play fighting. At PND 45, the treated and
control animals' weights were not significantly different, and the treated animals displayed more
aggression as measured by attacking and biting an intruder put in the cage. In the assessment of
developmental Pb exposure discussed above (Moreira et al., 2001), the early Pb exposure
resulted in a decrease in social interaction time in the PND 70 rats.
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Levin et al. (1988) exposed rhesus monkeys to a pulse-chronic Pb exposure protocol as
described above (Levin and Bowman, 1986) resulting in a blood Pb level of 56 |ig/dL during
week 5 and 33 to 43 |ig/dL for the remainder of the first 6 months of life. During the first
6 weeks after birth, results of the Early Infant Behavioral Scale showed Pb-induced lowered
muscle tonus and greater agitation, but no effects on sensorimotor measures. Beginning at
PND 14, monkeys were tested on a Piagetian object permanence task, which revealed no Pb-
related effects. Starting at 2 months of age, monkeys were tested on a visual exploration task,
which showed decreased visual attentiveness in Pb-treated monkeys. Laughlin et al. (1991)
evaluated the effects of Pb exposure and diet in rhesus monkeys exposed to 1 mg/kg/day Pb
acetate from PND 5 until PND 365. Monkeys were given either low-milk or high-milk diets
because of milk's ability to enhance tissue levels of Pb. Blood Pb levels reached a plateau of
-70 |ig/dL during the first year when initial testing occurred and then decreased to ~ 35 |ig/dL at
16 months postexposure. During the first year of life, Pb-induced disruption of social play and
increases in both self-stimulation and fearful behavior were observed. At 16 months of age,
these changes were still present. Differences in milk intake had little effect on behavior in this
study.

Pb Exposure and the Stimulus Properties of Neuropharmacologic Agents
The drug discrimination paradigm has been utilized to characterize postsynaptic receptor
status for multiple neurotransmitter systems. Rats chronically exposed to Pb beginning at
weaning and tested as adults were trained to discriminate either a systemically administered DI
or D2 receptor agonist (Cory-Slechta and Widzowski, 1991). Exposed rats learned the
discrimination task more rapidly than controls and exhibited greater levels of responding to
lower doses of the training drugs and less blockade by a D2 receptor antagonist, consistent with
generalized dopaminergic receptor supersensitivity. In groups of animals exposed only from
birth to weaning and trained to discriminate the same drugs, the D2-D3 subtype receptor
supersensitivity in exposed animals was again present, but no changes in responding to the DI
agonist were apparent (Cory-Slechta et al., 1992). Further work with this test employing the
postweaning exposure protocol failed to demonstrate any Di-D2 receptor interactions in the
supersensitivity displayed by Pb-exposed animals (Cory-Slechta et al., 1996a).
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To test cholinergic sensitivity in animals chronically exposed after weaning, rats were
trained to discriminate a muscarinic agonist (Cory-Slechta and Pokora, 1995) and were tested in
the added presence of a muscarinic antagonist. The results suggest an increased sensitivity to at
least one subtype of muscarinic receptor in Pb-exposed rats.
Glutamatergic functioning also has been assessed by use of the drug discrimination
paradigm. Rats chronically exposed beginning at weaning and tested as adults exhibited
diminished responsiveness to an NMDA subtype receptor antagonist (Cory-Slechta, 1995), but
enhanced responding to lower doses of NMDA (Cory-Slechta et al., 1996a). When exposure
was limited to the period between birth and weaning, the diminished sensitivity to the NMDA
receptor antagonist was less evident but still present (Cory-Slechta, 1997).
Thus, the drug discrimination method appears to have provided useful insights into the
status of some neurotransmitter systems in chronically exposed animals. The reports cited above
indicate an upregulation of dopaminergic, cholinergic, and glutamatergic receptors that are
generally consistent with findings of diminished presynaptic function described earlier in this
section. Nonetheless, this paradigm has some limitations. As all drugs in the cited studies were
administered systemically, the results provide no evidence on brain regional sites of action.
In addition, the chronic intermittent administration of the training drug has the potential to
induce compensatory neuronal changes by itself and may thusly mask or otherwise alter the
manifestation of the effects of Pb exposure.

5.3.6 Lead-Induced Changes in Cellular Development and Disposition
of the Metal
Alterations in cellular differentiation and morphology can be important structural
components of the manifestations of Pb neurotoxicity in neurons and glia. While these issues
have not been thoroughly addressed by research investigations, important observations have
nonetheless been made, as discussed below in the following subsections.

Lead Exposure and Neural/Glial Progenitor Cells
Recent Pb neurotoxicity studies have evaluated the effects of Pb exposure on neural and
glial progenitor cells. Chronic Pb exposure of rats, beginning at PND 25 and producing blood Pb
levels of 20 |ig/dL at the termination of exposure, was found to significantly decrease
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proliferation of new cells in the dentate gyrus compared to control animals (Schneider et al.,
2005). Other workers determined that continuous exposure from birth to adulthood producing
blood Pb levels of 35 to 40 |ig/dL reduced the total number of labeled cells in the hippocampal
dentate gyrus at 28 days after the last administration of a DNA synthesis marker (Gilbert et al.,
2005). Rats whose exposure was terminated at weaning showed no changes in cellular labeling
or survival, indicating that chronic exposure reduces the capacity for hippocampal neurogenesis.
Studies have also been conducted to investigate the effects of Pb exposure on glial
progenitor cells. Deng et al. (2001) examined cultured oligodendrocytes and their progenitor
cells acutely exposed to Pb2+ in vitro; they observed an exposure-induced delay in the
differentiation of the progenitors and that the progenitor cultures were more sensitive to Pb2+
than the mature oligodendrocytes. These findings suggested interference with the timely
developmental maturation of the progenitor cells. A subsequent study found that a low
concentration of Pb2+ in vitro inhibited proliferation and differentiation of these progenitors
without affecting cell viability (Deng and Poretz, 2002). Proliferative capability was decreased
and cell-intrinsic lineage progression was inhibited at a late progenitor stage. Thus, acute Pb2+
exposure suppresses both the proliferation and differentiation of progenitor cells.

Lead Exposure andNeurite Outgrowth
Neurite initiation is highly sensitive to neurotoxic compounds and has been the focus of
studies examining morphological alterations caused by in vitro exposure to Pb2+. Kern and
Audesirk (1995) found that 100 nM Pb2+ inhibited neurite initiation in cultured rat hippocampal
neurons and, on the basis of results with kinase inhibitors, concluded that this occurred by
inappropriate stimulation of protein phosphorylation by Ca2+-calmodulin-dependent or cyclic
AMP-dependent protein kinases, possibly through stimulation of calmodulin. Intracellular free
Ca2+ concentrations were not altered by up to 48 h exposure to nominal 100 nM Pb2+, suggesting
that the stimulation of the above kinases or calmodulin were not via increased Ca2+, but instead
were attributable to intracellular Pb2+ concentrations. Evidence of Pb2+-induced inhibition of
neurite outgrowth is in general agreement with results seen after chronic exposure to Pb
employing in vivo models. Cline et al. (1996) employed an exposure protocol of 0.1 nM to
100 jiM nominal Pb2+ for 6 weeks localized to the retinotectal system of frog tadpoles; they
observed that the area and number of retinal ganglion cell axon arborizations within the optic
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tectum was reduced at nanomolar Pb2+ concentrations. As discussed in Section 5.3.4, Reuhl
et al. (1989) exposed primates to 2 mg Pb/kg/day from infancy to 6 years of age and found that
neuronal volume density was reduced in primary visual area VI and in visual projection area V2,
compared to a group exposed to 25 jig Pb/kg/day. Moreover, a relative decrease in the number
of arborizations among pyramidal neurons in both areas VI and V2 was observed in the higher-
dose group. Thus, there was good correspondence between reports that acute Pb2+ exposure in
vitro and extended exposure in animal models in vivo results in diminished neuronal growth and
differentiation at Pb levels of apparent environmental relevance. Studies employing intact
animals have not investigated specific cellular mechanisms underlying these effects.

Lead Exposure and Neural Stem Cells
Given considerable contemporary interest in the use of neural stem cells to treat various
neurological diseases, the efforts of Huang and Schneider (2004) to examine the actions of
exposure to Pb2+ in vitro on these cells is noteworthy. Pb exposure produced no effect on
neurosphere viability, but caused a significant dose-dependent inhibition of proliferation.
In addition, the number of neurons differentiated from Pb2+-exposed neurospheres was
significantly decreased versus control, as were the number of oligodendrocytes obtained.
However, Pb exposure increased the number of astrocytes obtained. These observations suggest
an important Pb2+-induced influence on stem cell proliferation and differentiation.

Lead and the Blood-Brain Barrier
Early work demonstrated that the capillary epithelium in the brain is a target for Pb and
that Pb intoxication can disrupt the blood-brain barrier (BBB). Pb-exposed capillary endothelial
cells isolated from rat cerebral cortex showed deposits of Pb preferentially sequestered in
mitochondria, suggesting Pb-induced disruption of transepithelial transport of Ca2+ and other
ions. Furthermore, the developing CNS is especially sensitive to Pb-induced vascular damage.
Cerebral endothelial cells are known to accumulate Pb much more than other cell types and the
choroid plexus in both humans and animals accumulates much higher Pb concentrations than
other brain regions. However, these studies employed high exposure levels and, thus, are of
limited utility in evaluating the effects of environmentally relevant exposures.
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Bradbury and Deane (1986) examined the rate of uptake of 203Pb into brain and other soft
tissues of the rat at constant radiotracer levels in plasma. Uptake of 203Pb in the brain was linear
up to 4 h, and the authors hypothesized that the capillary endothelium was the rate-limiting
component of Pb transport into the brain because of its relatively small area. The rate of uptake
in the brain contrasted with that into cisternal cerebrospinal fluid, which plateaued at -5% of that
of plasma at 1 to 2 h, and into the choroid plexus, in which Pb accumulated to -350% of plasma
levels at 4 h suggesting a 10-fold greater uptake. Bradbury et al. (1991) found that albumin
rarely enters the brain from blood, suggesting that Pb transport into brain is likely as free Pb2+ or
Pb in the form of inorganic complexes (e.g., PbOH+, PbHCC>3+, or PbCl+) or Pb bound to a low
molecular weight organic ligand (e.g., cysteine). Further work (Bradbury and Deane, 1993)
using short vascular perfusion of a cerebral hemisphere determined that 203Pb enters the brain
very quickly in the absence of an organic ligand but that transport is abolished in the presence of
albumin, L-cysteine, or EDTA. It was proposed that PbOH+ or some other simple organic Pb2+
complex passively enters the endothelium, and that the entry is mitigated by active back
transport of Pb2+ into blood by Ca-ATPase pumps.
To evaluate mechanisms by which Pb increases the permeability of the BBB, Dyatlov
et al. (1998a) dosed BALB/cByJ suckling male mice with 2.5 jig/g body weight of Pb acetate,
LPS (100 ng/g body weight), recombinant IL-6 (5 ng/body weight), Pb + IL-6, or sodium acetate
+ LPS. Following five injections over 10 days, they measured the transendothelial electrical
resistance across the BBB. Pb at this level alone had no effect, but did potentiate the increases
due to LPS. Pb plus IL-6 also caused a delay in the increase in arteriole resistance. Thus, Pb
potentiates the actions of both IL-6 and LPS. Glutamate topically applied to the cerebrum
caused a reversible decrease in resistance, whereas Pb caused this decrease to be irreversible.
The authors hypothesize that this disruption of the BBB allows glutamate to enter the brain,
further disrupting the BBB and irreversibly potentiating brain injury.
Lead also compromises the function of the barrier between the cerebrospinal fluid and
systemic circulation, allowing transfer between tight junctions of the choroid plexus. Zheng
et al. (1996) demonstrated that chronic Pb exposure (50 or 250 mg/mL in drinking water for
30, 60, or 90 days) reduced transthyretin levels in cerebrospinal fluid of male weanling rats in the
presence of blood Pb levels of 18.2 and 48.9 |ig/dL, respectively. Transthyretin, which is
expressed in early fetal development, is produced in the choroid plexus and is the major thyroid
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hormone binding protein, allowing transfer of thyroxine from choroid plexus into cerebrospinal
fluid. The authors proposed that this Pb-induced reduction in transthyretin may be a factor in
Pb-induced alterations in brain development.
Thus, the BBB allows significant entry of Pb into both the adult and developing brain.
This transport is dependent on the chemical form of Pb, interactions of Pb with proteins and
other components of the blood, and other biochemical and physiologic factors that are not fully
defined. Information regarding Pb-binding proteins in brain are presented in Section 5.11.2.

Tissue Uptake of Lead
Pb2+ appears to be taken up into cultured cells by multiple ion channel-based mechanisms,
including influx through channels activated by depletion of intracellular Ca2+ stores, non-L-type
Ca2+ channels, and NMDA receptor-associated channels (Kerper and Hinkle, 1997; Mazzolini
et al., 2001). Astroglia are well-known to act as Pb sinks and in culture and can accumulate up
to 24 times more of the metal than neuronal cells (Lindahl et al., 1999). There is also evidence
that glutathione may regulate Pb uptake into astroglia.
Histologic studies (e.g., Struzyriska et al., 1997) have demonstrated the transport of Pb
into the brains of chronically exposed adult animals. Weanling rats were dosed with 2 g/L lead
acetate in water for 3 months, creating blood Pb levels of 39 |ig/dL. Using horseradish
peroxidase as a tracer of vascular permeability, leaky microvessels were demonstrated by both
light and electron microscopy. Focal leakage of tracer was observed in the short segment wall of
microvessels, the surrounding neuropil, and regions of parenchyma near microvessels. Staining
was also evident in the cytoplasm of pericytes and on the basement membrane of endothelial
cells.

Accumulation ofPb in Blood and Brain
Early animal studies often neglected to include blood Pb and concomitant tissue levels
achieved by the exposure protocols. The 1986 Lead AQCD was able to draw some limited
conclusions about the relationship of exposure levels to blood and brain Pb concentrations.
In general, at exposure concentrations of >0.2% Pb in drinking water and for exposure durations
extending beyond the birth to weaning period, the ratio of blood to brain Pb concentration is
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<1, suggesting that even as blood Pb peaks and then falls due to excretion or removal, the high-
affinity binding of the metal to brain proteins promotes further accumulation in the brain.
The decline in blood Pb after exposure is terminated has been shown to depend on the
level and duration of exposure (O'Flaherty et al., 1982; Hryhorczuk et al., 1985). Widzowski
and Cory-Slechta (1994) exposed rat pups to various concentrations of Pb in the drinking water
from birth to weaning, and they found that blood Pb declined rapidly with a half-life of <20 days.
Wedeen (1992) reported a half-life of ~1 month in children, but Manton et al. (2000) found
half-lives of 10 to 38 months in young children, depending somewhat on duration of Pb exposure
from home remodeling. Because of the dependence of decreases in blood Pb concentrations on
exposure level and duration, estimates of half-life vary significantly from one exposure scenario
to another.
Using the same exposure protocol, Widzowski and Cory-Slechta (1994) measured brain
regional half-lives for Pb and found that these values averaged -20 days and did not vary
between brain regions. Because of binding to brain tissue proteins, Pb concentrations in brain
decline or respond to chelation more slowly than do blood Pb levels (Stangle et al., 2004).
Other studies describing chelation of Pb are included in Table AX5-3.5.

5.3.7 Susceptibility and Vulnerability Factors Modifying Lead Exposure
and Thresholds for CNS Effects
The effects of chronic Pb exposure may be modified by a number of physiologic variables
and exposure parameters that can significantly enhance or diminish the toxic response. These
susceptibility factors impact the toxic manifestations observed in the organism. A few of these
factors are considered below.

Aging
The neurotoxic effect of chronic exposure to low level Pb with advancing age is becoming
an important public health social issue (Yun et al., 2000a). Physiologic conditions associated
with bone resorption, e.g., pregnancy, lactation, and aging, can potentiate the CNS effects of Pb
and enhance exposure of adults. Such demineralization conditions can also add to the in utero
exposure of the fetus, and postmenopausal resorption can increase PbB levels in women by 25%
(Silbergeld, 1990).
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Bone Pb levels serve as a marker of the burden of Pb in middle-aged and elderly men.
During the demineralization that occurs during aging, blood Pb levels increase and may lead to
declines in cognitive function (Weisskopf et al., 2004a,b; Payton et al., 1998). Supporting data
have been obtained from animal studies. Cory-Slechta (1990b) administered various amounts of
Pb for a constant duration beginning in young, adult, or old rats. An increased vulnerability to
Pb was observed in older animals due to increased exposure from an elevated resorption from
bone and an apparently greater sensitivity to the biochemical effects of the metal. Yun et al.
(2000b) have presented evidence suggesting that this increased vulnerability may also be due to
cerebral energy depletion in exposed animals. However, such findings have not been uniformly
observed. Rice and Hayward (1999) exposed monkeys to Pb chronically and tested temporal
visual function and contrast sensitivity at two different ages. Pb-exposed subjects exhibited
differences in temporal function at the first, but not the second assessment, and there was no sign
of accelerated decline in contrast sensitivity at either testing period. Thus, the deleterious
Pb-aging interactions would appear to be dependent on the CNS function under study.
Recent studies in rats have demonstrated the importance of Pb exposure during early
development in promoting the emergence of Alzheimer's-like changes in old age. Basha et al.
(2005) exposed rat offspring to Pb through the dam during the lactational period and monitored
gene expression of beta-amyloid precursor protein (APP). APP mRNA was induced in neonates,
and exhibited a delayed overexpression 20 months after exposure was terminated. At this point,
APP protein and its beta-amyloidogenic product were increased, and a rise in the activity of the
transcription factor Spl, one of the regulators of the APP gene, was also present. The changes
induced by early exposure to Pb could not be reproduced by exposure to the metal during
senescence. It was concluded that environmental influences occurring during brain development
predetermined the expression and regulation of APP later in life. Subsequent work observed the
same responses in monkeys who had been exposed to Pb as infants (Zawia and Basha, 2005),
arguing for both an environmental trigger and a developmental origin of Alzheimer's disease.

Gender
Relatively few animal studies have attempted a clear demonstration of gender differences
for toxic responses to chronic Pb exposure. For example, Donald et al. (1986) reported enhanced
social investigatory behavior in exposed mice that differed in time course between males and
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females. A subsequent report (Donald et al., 1987) showed Pb-induced nonsocial activity
decreased in females, but increased in males and also caused shorter latencies to aggression in
males. Also, Yang et al. (2003) found a gender difference in rats in memory retrieval, wherein
higher doses of Pb affected memory in males than in females.
A significant functional impact of such gender differences has recently emerged.
Cory-Slechta et al. (2004) evaluated the effects of interactions of chronic Pb exposure and
maternal restraint stress on offspring. Corticosterone levels, neurotransmitter changes, and FI
operant behavior were measured in animals exposed only during gestational and lactational
periods, creating blood Pb levels of 30 to 40 |ig/dL. These workers found that synergistic effects
of Pb and stress were observed more frequently in female offspring, and that Pb with (in
females) or without (in males) stress permanently elevated corticosterone levels. It was proposed
that these findings uncovered a new mechanism by which exposure could enhance susceptibility
to diseases. Virgolini et al. (2005) used continuous Pb exposure in the drinking water to males
beginning at weaning, creating a maximal blood Pb of 27 |ig/dL, in combination with variable
intermittent stress challenges and found that Pb alone decreased basal plasma corticosterone
levels and glucocorticoid receptor binding. Novelty stress in combination with exposure was
found to modify FI behavior. These findings support the results of Cory-Slechta et al. (2004) in
suggesting the potential for hypothalamic-pituitary-adrenal axis-mediated effects of Pb on CNS
function.
Virgolini et al. (2006) extended this line of investigation by exposing rats to Pb prior to
breeding and continuing throughout gestation and lactation. The exposure created maximal
blood Pb levels of 33 to 43 |ig/dL at termination of exposure at weaning. Maternal restraint
stress was applied on gestational days 16 and 17 in some of the animals. Female offspring were
then tested for responsiveness to various stressors (i.e., restraint, novelty, cold) as measured by
FI operant performance. The combination of exposure and maternal stress produced more
changes in responsiveness than either factor alone: operant behavior was altered following both
restraint and cold, and the corticosterone response was modified by cold. It thusly appears that
maternal Pb exposure can permanently alter stress responsivity and does so with a profile of
effects that differ from those produced by either exposure alone or maternal stress alone. These
studies are also among the first to make clear delineations of the effects of exposure across
gender.
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Developmental Period of Exposure
While early development is well-known to represent a period of particular sensitivity to
the neurotoxic effects of Pb, several studies have made comparisons to the CNS effects produced
by exposure encompassing later periods of development. There has been some consistency
across these behavioral, neurophysiologic, and neurochemical observations.
As discussed in Section 5.3.5, Rice and Gilbert (1990a) employed a nonspatial
discrimination reversal task in monkeys using form and colors as cues. The group continuously
exposed to Pb from birth exhibited the most impairment on this task, whereas the group exposed
continuously beginning at 300 days of age displayed impairment, but not as great as that of the
preceding group. Another group exposed to Pb from birth but discontinued at 400 days of age
did not exhibit an effect when tested as adults, suggesting that beginning Pb exposure after
infancy results in altered performance, while exposure also during infancy exacerbates the effect.
Rice (1990) later tested these same monkeys on a spatial discrimination reversal task and again
found that the group continuously exposed from birth exhibited the most impairment.
Rice (1992a) utilized the same exposure protocols, with blood Pb levels of 20 to
35 |ig/dL, and tested monkeys on a multi FI-FR schedule of reinforcement at two different ages.
Few effects were seen during the first testing period at 3 years of age, but at 7 to 8 years of age
increased response run rates and decreased interresponse times in FI responding were evident in
all three exposed groups. These results indicated that exposure during infancy was not required
for a Pb effect on this task, and that exposure only during infancy was sufficient to produce
alterations.
Gilbert et al. (1999b) used a model of synaptic plasticity, i.e., LTP in the hippocampal
dentate gyrus, to examine the effects of Pb exposure encompassing various developmental
periods in intact animals with maximal blood Pb levels of 35 to 40 |ig/d. Similar effects were
seen in groups continuously exposed from birth or continuously exposed from PND 30: the
magnitude of population spike LTP was diminished and the threshold for induction of the
phenomenon was elevated. A group exposed only during the lactational period displayed a
diminished magnitude of excitatory postsynaptic potential (EPSP) LTP also, but the threshold for
EPSP slope LTP induction was higher only in the group continuously exposed from birth. Thus,
exposure restricted to the lactational period was less disruptive to LTP in adult animals than
exposure beginning around birth or weaning.
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Lasley et al. (1999) examined stimulated release of glutamate and GABA in the
hippocampus to evaluate the effects of developmental Pb exposure in intact animals with
maximal blood Pb levels of 39 to 45 |ig/dL. Similar decreases in Ca2+-dependent depolarization-
induced glutamate and GABA release were observed in groups continuously exposed from
conception or continuously exposed from PND 35. The pattern of changes induced in the group
continuously exposed from the beginning of gestation was similar to that observed previously in
a group continuously exposed from birth (Lasley and Gilbert, 1996), indicating that gestational
exposure did not further enhance the impact of Pb beginning at birth when exposure in both
groups extended into adulthood. The changes found in the group continuously exposed
beginning in the early postweaning period demonstrated that exposure during early development
is not required to produce changes in glutamate and GABA release. Reductions in stimulated
release were also observed in a group exposed only during the gestational and lactational
periods, indicating that Pb limited to early development is also sufficient to produce deficits in
evoked transmitter release.
Altmann et al. (1993) also examined synaptic plasticity and learning in rats chronically
exposed during early development, producing maximal blood Pb levels of 14 to 16 |ig/dL, but
the results were in contrast to the others cited in this section. These workers found impaired LTP
and AAL in groups continually exposed from the beginning of gestation, even if Pb was
terminated in the early developmental period. Another group whose exposure began just prior to
weaning did not display any Pb-related differences from controls.
These studies indicate that continuous exposure begun pre- or perinatally consistently
results in a Pb effect as great as or greater than that produced by exposure over any other
developmental period. Continuous exposure begun postweaning also is consistently potent in
producing alterations, but whether the magnitude of the effect is similar to that of the preceding
group, and whether exposure limited to the gestational and/or lactational period elicits an effect,
appear to be task- or process-dependent.

Nutrition
It has long been known that diets sufficiently high in minerals such as zinc, iron, and
calcium offer some protection from Pb exposure by competing with Pb2+ for absorption from the
gastrointestinal tract. A full discussion of gastrointestinal absorption in humans is found in
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Chapter 4, and information on absorption in animals is contained in Section 5.10.2. In fact, this
dietary influence may comprise one component of the socioeconomic status (SES) factor
repeatedly identified as important in studies of Pb neurotoxicity in children, as maternal dietary
habits influence risk of exposure in infants. More recent studies have shown that vitamin D
administration can reduce blood and bone Pb values (Cheng et al., 1998; Cortina-Ramirez et al.,
2006), but whether this occurs as a result of increased Ca2+ uptake has not been determined.
Diets designed to reduce caloric intake and increase weight loss have also been associated with
increases in blood Pb values (Han et al., 1999).

Thresholds for CNS Effects
There is no evidence in the animal Pb neurotoxicity literature reflecting well-defined
thresholds for any of the toxic mechanisms of the metal. Most studies performed with in vivo
models report blood Pb values in the range of 15 to 35-40 |ig/dL or higher. Moreover, in view of
the complex and undefined speciation equilibria and distribution of Pb2+ in physiological
milieus, there is no way to directly relate a blood Pb value to the levels of free Pb2+ ion or to any
other complexed active form of the metal, either in extracellular or intracellular fluids. Generally
accepted estimates of the free Pb2+ ion concentrations produced in brain extracellular fluid by
environmentally relevant exposures fall in the low nanomolar range.
Nonetheless, changes induced by chronic Pb exposure have been reported on
neurochemical (e.g., Cory-Slechta et al., 1997b) and neurophysiologic (e.g., Altmann et al.,
1993) measures at blood Pb values of-15 |ig/dL. Recent studies have reported behavioral
changes at blood Pb concentrations of-10 |ig/dL (Brockel and Cory-Slechta, 1998; 1999a,b),
and the results on measures of attention have closely paralleled those observed in children at the
same blood Pb levels. These observations serve to validate the accuracy and usefulness of the
animal exposure models.

5.3.8 Summary
• Pb2+-Ca2+ interactions resulting from exposure, e.g., the Ca 2+-mimetic properties of Pb2+,
are important components of the cellular toxicity of the metal and are closely related to the
dose-dependent effects of exposure on neurotransmitter release. Some of these actions of
Pb2+ are shown in Figure 5-3.
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• Exposure-induced decreases in glutamatergic, cholinergic, and dopaminergic transmission
are important because of the purported role of these neuronal systems in brain development
and cognitive function.

• The majority of the data suggests an upregulation of NMD A receptors resulting from chronic
exposure, but a consensus on the effects of Pb on NMD A receptor subunit expression and
function remains to be attained. It is increasingly apparent that this glutamate receptor
subtype may not be a direct primary target of chronic exposure in the intact animal.

• In vitro interactions of Pb2+ and PKC have been carefully described and are broadly relevant
to cellular signaling pathways, but functional effects of these interactions in intact animals
have not been achieved.

• Using hippocampal models of synaptic plasticity, it has been demonstrated that Pb exposure
decreases the magnitude of LTP and increases the threshold for induction with a biphasic
dose-effect relationship, another indication of the nonlinear effects of Pb.

• Decreases in stimulated glutamate release are a significant factor contributing to Pb-induced
changes in LTP.

• Lead induces decreased activity of dopaminergic cells in substantia nigra and ventral
tegmental area and inhibits activation of nicotinic currents in cultured hippocampal cells.

• Evidence from nonhuman primate studies demonstrates convincingly that Pb exposure
producing blood Pb values as low as 33 |ig/dL impairs auditory function by increasing
latencies in brainstem auditory evoked potentials and elevating hearing thresholds.

• It has been shown that Pb exhibits selective effects on rod and bipolar cells in rats at a blood
Pb of 19 |ig/dL, causing decreases in maximal ERG amplitude, decreases in ERG sensitivity,
and increases in mean ERG latency.

• Mechanisms for Pb-induced deficits in visual function include concentration-dependent
inhibition of cGMP phosphodiesterase, increased rod Ca2+ levels, decreased retinal
Na-K-ATPase activity, and apoptotic death of rod cells.

• Developmental exposure to Pb, creating steady state blood Pb levels of-10 |ig/dL, results in
behavioral impairments that persist into adulthood in monkeys. There is no evidence of a
threshold and Pb-induced deficits are, for the most part, irreversible.

• In monkeys, permanent neurobehavioral deficits are observed both with in utero-only
exposure and with early postnatal-only exposure when peak blood Pb levels did not exceed
15 |ig/dL and steady state levels were ~1 l|ig/dL. Exposure started at -300 of age creates
deficits similar to those created with exposure from birth onward. In rats, permanent deficits
are observed with prenatal, preweaning, and postweaning exposure.

• The developmental period of exposure is critical to the type of behavioral deficit produced.
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• Lead affects the performance on a number of neurobehavioral tasks by inducing (1) reduced
ability to inhibit inappropriate responding, (2) distractibility, (3) reduced ability to adapt to
changes in behavioral contingencies and, (4) perseveration.

• Response perseveration, insensitivity to changes in reinforcement density or contingencies,
and deficits in attention appear to be important mechanisms of Pb-induced learning deficits.
Impulsivity and waiting-for-reward behavior may be more strongly affected by Pb than are
sustained attention and hyperkinesis. Pb-induced impulsivity may be factor in cognitive
impairments.

• Pb impairs learning at blood Pb levels as low as 11 |ig/dL as tested in FI tasks. Performance
on spatial and nonspatial discrimination reversal tasks is affected following developmental
exposures in nonhuman primates. Distracting irrelevant stimuli potentiate these impairments.
Pb also impairs performance on olfactory reversal tasks in rats. Discrimination reversal
appears to be a more sensitive indicator of Pb-induced learning impairment than simple
discrimination. Repeated-acquisition tests show that these deficits are unlikely to be caused
by sensory or motor impairment.

• The effects of lead on memory are not as clear-cut: in some cases, the animals showed
impairment of memory at blood Pb levels of as low as 10 |ig/dL while in other studies
animals demonstrated improved memory following Pb exposure. Short-term memory does
not appear to be affected by low level Pb exposure. Some behavioral deficits in tests of
working memory (e.g., DSA) appear to result from impaired attention rather than memory.

• Lead has been demonstrated to affect reactivity to environmental stimuli and social behavior
in both rodents and nonhuman primates at blood Pb levels of 15 to 40 |ig/dL.

• Other test paradigms such as drug discrimination and repeated acquisition/performance tasks
have provided useful assessments of the integrity of CNS neurotransmitter systems in Pb
exposed animals. Evidence from both methods has been in general agreement in indicating
upregulated neurotransmitter receptor systems.

• Pb has been shown to decrease cell proliferation in vivo at 20 |ig/dL and to decrease
proliferation, differentiation, and neurite outgrowth in vitro.

• In both the adult and developing brain, the blood brain barrier (BBB) allows entry of Pb into
the brain. Entry is dependent upon the chemical form of the Pb and interactions with
proteins and other components of blood.

• The rate at which Pb accumulates in the brain depends upon the level and duration of
exposure. The half-life of Pb in brain tissue is -20 days in rats and appears to be
homogenous across regions.

• Susceptibility and vulnerability factors that modify responses to lead include (1) age,
(2) gender, (3) socioeconomic status, (4) period of exposure, and (5) nutrition.
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A well-defined threshold for any toxic effects in animals has not been identified.
Neurochemical and neurophysiologic effects have been seen at blood Pb values of-15 |ig/dL
and neurocognitive effects have been observed at -10 |ig/dL.
5.4 REPRODUCTIVE AND DEVELOPMENTAL EFFECTS OF LEAD
5.4.1 Summary of Key Findings on the Developmental and Reproductive
Effects of Lead in Animals from the 1986 Lead AQCD
The 1986 Lead AQCD presented unequivocal evidence for effects of Pb on reproduction
and development in laboratory animals, derived principally from studies of rodents. Fetotoxic
effects (spontaneous abortion and fetal death) were reported following chronic exposures to
relatively high doses (600 to 800 ppm) of inorganic Pb in the diet, and more subtle effects (such
as changes in ALAD activity or hematocrit in offspring) at lower doses (5 to 10 ppm in drinking
water and 10 |ig/m3 in air). The 1986 Lead AQCD reported that the lowest observed adverse
effect level (LOAEL) for reproductive and developmental effects was 64 ng/kg per day (multiple
exposures by gavage).
The 1986 Lead AQCD also reported evidence for a variety of sublethal effects on
reproduction and development in experimental laboratory animals following Pb exposure.
Sublethal effects included changes in levels or function of reproductive hormones as well as
effects on the gonads (both male and female) and conception. The animal data also suggested
more subtle effects on hormone metabolism and reproductive cell structure. Stowe and Goyer
(1971) classified the reproductive effects of Pb as gametotoxic, whether intrauterine or
extrauterine.
The data reported in the 1986 Lead AQCD, and more recent studies conducted in
experimental animal models, provide convincing evidence that Pb induces temporary and long-
lasting effects on male and female reproductive and developmental function. The newer
literature supports the earlier conclusions presented in the 1986 Lead AQCD that Pb disrupts
endocrine function at multiple points along the hypothalamic-pituitary-gonad axis (Sokol et al.,
1985; Stowe and Goyer, 1971; Vermande-Van Eck and Meigs, 1960; Junaid et al., 1997;
McGivern at al., 1991; Ronis et al., 1996, 1998b,c; Sokol, 1987; Sokol et al., 1985, 1994, 1998;
Sokol andBerman, 1991; Kempinas atal., 1988, 1990, 1994; Tchernitchin etal., 1998b; Sant'
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Ana et al., 2001; Srivastava et al., 2004). A schematic representation of the hypothalamic-
pituitary-gonadal (HPG) axis is shown in Figure 5-7.
Testosterone or
Estradiol
Figure 5-7. Data from male and female experimental animals suggests that lead
has multiple targets in the hypothalmic-pituitary-gonadal axis.
The majority of the experimental animal studies on developmental and reproductive
effects of Pb examined effects due to inorganic forms of lead; very little is known about the
reproductive and developmental effects due to organic forms. In general, the few available
studies suggest that effects of organic forms of Pb are similar to those produced by inorganic
forms. Administration of triethyl-Pb-chloride during early gestation reduces pregnancy rates in
mice (Odenbro and Kihlstrom, 1977). Growth retardation following organolead exposure has
been reported (Kennedy et al., 1975; McClain and Becker, 1972). More recent studies have
demonstrated that exposure of mice to triethyl-Pb-chloride during late gestation reduces perinatal
growth rate (Odenbro et al., 1988).
This section summarizes the evidence for effects of Pb exposure in developing organisms
exposed during the period from conception to maturity that has been reported since 1986.
Effects on neurological, immunological, or renal endpoints in developing organisms are
discussed in Sections 5.3, 5.9 and 5.7, respectively.
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5.4.2 Effects on Male Reproductive Function
The 1986 Lead AQCD reported convincing evidence based on experimental animal
studies that Pb acts as an endocrine disrupter in males. Those studies demonstrated an
association between reduced male fertility and repeat-dose Pb exposure. Lead exposure had
been reported to alter sperm development and function; however, the mechanism underlying
these effects was not completely understood. These effects were attributed to either alterations
in testicular enzymes important for hormone production or to changes in the hormone receptors.
More recent research supports the conclusion that mechanisms for endocrine disruption in males
involve lead acting at multiple sites along the hypothalamic-pituitary-gonadal axis (see
Figure 5-7).
Reported effects of Pb on male reproduction differ substantially across studies, with some
studies finding severe adverse effects and other studies finding no or minimal effects. The
variable findings have been attributed to the complex mechanisms involved in hormone
regulation and the multiple sites of action for Pb. Sokol et al. (2002) suggested that differences
in results among studies may be, in part, attributed to an adaptive mechanism in the
hypothalamic-pituitary-gonadal axis that may render the expression of some toxic effects
dependent on dose and exposure duration. The mechanisms underlying this possible adaptation
have not been completely elucidated. Lead exposure produces, a dose (blood Pb)-related
suppression of serum testosterone levels and spermatogenesis, with an increase in GnRH mRNA
in hypothalamus (at PbB <50). The latter effect is attenuated at higher exposures (>50 |ig/dL)
and with increasing exposure duration (Sokol et al., 2002). Sokol and Berman (1991) found that
timing of exposure was critical to Pb-induced male reproductive toxicity in rats. Studies
conducted in nonhuman primates supported the importance of timing, found that the adverse
effects of Pb on male reproduction are dependent upon age (i.e., developmental stage at time of
exposure) and duration of exposure (Foster et al., 1993; Singh et al., 1993a). It is currently
unclear whether these effects reflect a physiological adaptation to the stress of lead exposure, or
reflect the combined outcome of distinct dose-duration-response relationships for multiple
effects on the FtPG axis.
The adverse effects of Pb on male reproduction may be expressed as perturbations in
sexual development and maturation, changes in fertility, endocrine disruption, and alterations in
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structure of reproductive cells or tissue. Each of these effects is discussed in greater detail in the
sections that follow.

5.4.2.1 Effects on Male Sexual Development and Maturation
The 1986 Lead AQCD reported adverse effects of Pb on male sexual development and
maturation. Experimental studies conducted in animals demonstrated that high-dose (e.g.,
dietary exposure to 0.08 to 1.0% (800-1000 ppm) Pb acetate in mice and to 100 ppm in dogs)
preadolescent Pb exposure can produce long-lasting detrimental effects on male sexual
development. Numerous more recent studies conducted in experimental animals support the
earlier findings that Pb exposure during early development can delay the onset of male puberty
and alter reproductive function later in life (McGivern et al., 1991; Al-Hakkak et al., 1988;
Chowdhuri et al., 2001; Dearth et al., 2002, 2004; Gandley et al., 1999; McGivern et al., 1991;
Ronis et al., 1998a,c; Sokol et al., 1994; Yu et al., 1996). Studies that provide the strongest
evidence for the dose-response range for typical effects in rodents are discussed below
(Table 5-3).
McGivern et al. (1991) found that male rats born to dams that received Pb acetate in
drinking water beginning on gestation day 14 and through parturition (blood Pb 73 |ig/dL)
exhibited reduced sperm counts, altered male reproductive behavior, and enlarged prostates later
in life. Prepubertal exposure of male Sprague-Dawley rats (age 24 to 74 days) to Pb acetate in
drinking water (blood Pb 30 to 60 |ig/dL) resulted in significant reduction in testis weight and in
the weight of secondary sex organs; however, these effects were not observed in rats exposed
postpubertally (day 60 to 74; Ronis et al., 1996). A dose-dependent delay in sexual maturation
was found in male rats, following prenatal Pb exposure that continued until adulthood (age
85 days) (Ronis et al., 1998a,b,c). In these studies, blood Pb levels in the pups between the ages
of 21 and 85 days were >100 |ig/dL. Additional details concerning these studies are provided in
Table 5-3.
One possible explanation for the observed persistent effects of Pb exposure on the male
reproductive system is a disruption in pulsatile release of sex hormones during early
development (Ronis et al., 1998c). Lead effects on sex hormones are discussed in
Section 5.4.2.3.
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Table 5-3. Selected Studies Showing the Effects of Lead on Reproductive Function in Males
Citation
Species/
Strain
Dose/Route/Form/Duration/Group Size Endpoint/Magnitude of Effect/p-value Blood Lead Concentration (PbB)
Foster et al.
(1993)
Foster et al.
(1996a)
fj\

250 ng/dL reduced circulating
testosterone levels in male rats 40-50%
(p < 0.05); reduction in male secondary sex
organ weight (p < 0.005); delayed vaginal
opening (p < 0.0001); disrupted estrous cycle in
females (50% of rats); increased incidence of
stillbirth (2% control vs. 19% Pb) (p < 0.005).
Lifetime group 3-26 ng/dL at 4-5
years
Infancy group 5-36 ng/dL at 100-300
days, 3-3 ng/dL at 4-5 years
Post-infancy group 20-35 ng/dL

PbB10±3or56±49|ig/dL
PbB -35 ng/dL
Control PbB <5 ng/dL at birth
Maternal PbB 73 ng/dL at birth
Pup PbB 64 ng/dL at birth
Pubertal PbB 30-60 ng/dL
Post-pubertal PbB 30-60 ng/dL
Mean PbBs in male rats 30-60 ng/dL,
respectively
-------
Table 5-3 (cont'd). Selected Studies Showing the Effects of Lead on Reproductive Function in Males
Species/
Citation Strain Dose/Route/Form/Duration/Group Size Endpoint/Magnitude of Effect/p-value Blood Lead Concentration (PbB)
Ronis et al. Rat/Sprague- 0.6% Pb acetate in drinking water ad libitum for
(1998a) Dawley various durations as follows: GDStoPNDl;
GD 5 to weaning; PND 1 to weaning; 3 control
litters, 2 gestation exposure litters, 2 lactation
exposure litters, 2 gestation and lactation
exposure litters, 2 postnatal exposure litters,
2 chronic exposure litters; 4 male and 4 female
pups per litter

Suppression of adult mean serum testosterone
levels was only observed in male pups exposed
to Pb continuously from GD 5 throughout life
(p<0.05).

Group

Naive

Control

Gest

Lact
Gest+Lact
Postnatal
Chronic
Male PbB

5. 5 ±2.0 ug/dL

1.9 ±0.2 ug/dL

9.1 ±0.7 ug/dL

3.3 ±0.4 ug/dL
16.1 ±2.3 ug/dL
226.0 ± 29 ug/dL
316.0 ±53 ug/dL
Ronis et al. Rat/Sprague- Lead acetate in drinking water (0.05% to 0.45%
(1998b) Dawley w/v); dams exposed until weaning; exposure of
pups which continued until PND 21, 35, 55, or
85; 5 control litters (0%), 10 low-dose litters
(0.05%), 8 mid-dose litters (0.15%), 9 high-dose
litters (0.45%); 4 male and 4 female pups per
litter
Ronis et al. Rat/Sprague- Lead acetate 0.05, 0.15, or 0.45% in drinking
(1998c) Dawley water beginning GD 5 continuing until PND 21,
35, 55, or 85; 5 control litters (0%), 10 low-dose
litters (0.05%), 8 mid-dose litters (0.15%), 9
high-dose litters (0.45%); 4 male and 4 female
pups per litter
Dose-response reduction in birth weight
(p < 0.05), more pronounced in male pups;
decreased growth rates in both sexes (p < 0.05)
were accompanied by a statistically significant
decrease in plasma concentrations of IGF 1
through puberty PND 35 and 55 (p < 0.05);
increase in pituitary growth hormone during
puberty (p < 0.05).

Dose-responsive decrease in birth weight
(p < 0.05); dose-responsive decrease in crown-
to-rump length (p < 0.05); dose-dependent delay
in sexual maturity (p < 0.05); decrease in
prostate weight (p < 0.05); decrease in plasma
concentration of testosterone during puberty
(p < 0.05); decrease in plasma LH (p < 0.05);
elevated pituitary LH content (p < 0.05);
decrease in plasma testosterone/LH ratio at high
dose (p < 0.05).
Mean PbB in offspring at 0.05% (w/v)
49 ± 6 ug/dL

Mean PbB in offspring at 0.15% (w/v)
126 ±16 ug/dL

Mean PbB in offspring at 0.45% (w/v)
263 ± 28 ug/dL
Dams: 0, 48, 88, or 181 ug/dL
Pups PND 1: <1, 40, 83, or 120 ug/dL
Pups PND 21: <1,46, 196, or
236 ug/dL
Pups PND 35: <1,20, 70, or
278 ug/dL
Pups PND 55: <1,68, 137, or
379 ug/dL
Pups PND 85: <1,59, 129, or
214 ug/dL
-------
Table 5-3 (cont'd). Selected Studies Showing the Effects of Lead on Reproductive Function in Males
oo
o
Citation
Singh et al.
(1993a)
Sokol and
Berman
(1991)

Species/
Strain Dose/Route/Form/Duration/Group Size
Monkey/ 0-1500 |ig Pb acetate/kg-d in gelatin capsules for
Cynomolgus various durations: 3 control monkeys, 4 monkeys
in infancy group (exposure first 400 days), 5 in
post-infancy group (exposure 300 days to 9 years
of age), 4 in lifetime group (exposure from birth
until 9 years)
Rat/Wistar 0,0.1, or 0.3% Pb acetate in drinking water for 30
days beginning at 42, 52, or 70 days old; 8-1 1
control rats for each age, 8-1 1 rats for each age in
0.1% group, 8-11 rats for each age in 0.3% group

Endpoint/Magnitude of Effect/p-value Blood Lead Concentration (PbB)
Degeneration of seminiferous epithelium in all Chronic PbB
exposed groups (frequency not specified); <40-50 |ig/dL
ultrastructural alterations in seminal vesicles,
most prominent in infancy and post-infancy
groups (frequency not specified).
Dose-related suppression of spermatogenesis Group
(decreased sperm count and sperm production
rate) in the exposed rats of the two highest age 0%
groups (p < 0.05); dose-related suppression of
serum testosterone in 52-day old rats (p = 0.04)
and in 70-day old rats (p < 0.003).
0.1%

0.3%

Age
All
42 d
52 d
70 d
42 d
52 d
70 d

PbB
<7 ng/dL
25 |ig/dL
35 |ig/dL
37 ng/dL
36 ng/dL
60 ng/dL
42 ng/dL
FSH, follicle stimulating hormone; GD, gestational day; GnRH, gonadotropin releasing hormone; IGF], insulin-like growth factor 1; LH, luteinizing hormone; PbB, blood Pb

concentration; PND, post-natal day
-------
5.4.2.2 Effects on Male Fertility: Effects on Sperm Production and Function
The 1986 Lead AQCD presented evidence that Pb exposure affects male fertility in
various animal species, including rabbits (Cole and Bachhuber, 1915), guinea pigs (Weller,
1915), rats (Ivanova-Chemishanska et al., 1980), and mice (Schroeder and Mitchener, 1971).
Several more recent studies, conducted in various animal species, have demonstrated
Pb-induced alterations of sperm parameters (e.g., count, motility, number of abnormal) (Sokol
et al., 1985; and eight other studies). These effects, however, have not been reproduced in all
studies. For example, Foster et al. (1996a) reported that 15- to 20-year-old cynomolgus monkeys
receiving Pb acetate for their lifetime (mean blood Pb 56 |ig/dL) showed no significant
alterations in sperm parameters (i.e., sperm count, viability, motility, and morphology) or
circulating levels of testosterone (see Section 5.4.2.3 for discussion of Pb-induced changes in
testosterone levels). Adaptive (Sokol et al., 2002) or multiple effects on the FtPG axis having
different dose-duration-response relationships may explain the apparent inconsistency in
reported effects on circulating testosterone levels, sperm count, and sperm production following
Pb exposure. As a result, changes in testosterone levels and certain sperm parameters may not
always serve as reliable endpoints for assessing the effects of Pb on male fertility and
reproductive function for all exposure durations.
Although gross changes in sperm parameters were not observed in monkeys in which
chronic blood Pb was -56 |ig/dL, Foster et al. (1996a) reported that monkey sperm exhibited a
statistically significant, dose-related reduction in chromatin structure (as determined by
susceptibility to weak acid denaturation). These changes may have adverse impacts on fertility,
and they are thought to be related to dominant lethal effects of Pb (similar to the effects reported
by al-Hakkak et al. [1988] in mice). Additional details concerning Foster et al. (1996a) are
provided in Table 5-3.
The data from Foster et al. (1996a), demonstrating a change in monkey sperm chromatin
suggestive of a subtle lead-induced reduction in male fertility (in the absence of gross changes in
sperm parameters), are consistent with observations of reduced in vitro fertilization capacity of
sperm collected from other mammalian species. Sokol et al. (1994) reported that exposure of
adult male rats to Pb acetate in drinking water for 14 to 60 days (blood Pb 33 to 46 |ig/dL)
resulted in reduced in vitro fertilization of eggs harvested from unexposed females. No
differences were observed in sperm ultrastructure or in the DNA histogram of sperm obtained

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from Pb-exposed rats compared to controls. Consistent with this finding are reports of reduced
fertilization capacity of rabbit sperm exposed to high concentrations (25 jiM) of Pb chloride
in vitro (Foote, 1999) and reduced in vitro fertilization capacity of sperm from mice exposed to
Pb in drinking water at 1 g/L for 4 months (blood Pb concentration not reported) (Johansson
etal., 1987).
Two modes of action have been proposed for Pb-induced alterations in sperm capacity for
fertilization. The affinity of Pb for sulfhydryl groups may explain some of the Pb-induced
alterations in sperm structure and function. Mammalian sperm possess high concentrations of
sulfhydryl groups, including chromatin stabilizing protamines, which are critical for maintenance
of normal function (Johansson and Pellicciari, 1988; Quintanilla-Vega et al., 2000). Reyes et al.
(1976) demonstrated that binding of Pb to membrane thiols inhibits sperm maturation.
In addition, recent experimental data also suggest that Pb-induced generation of reactive oxygen
species (ROS) may contribute to the injury of tissues responsible for sperm formation (see
Section 5.4.2.4).

5.4.2.3 Effects on Male Sex Endocrine System
The 1986 Lead AQCD reported that, although the mode of action for the adverse effects
of Pb on the male reproductive system was not understood, effects on hormone production or
hormone receptors were likely contributors. More recent studies provide convincing evidence
that Pb acts as an endocrine disrupter in males at various points along the hypothalamic-
pituitary-gonadal axis (Figure 5-7). In rats, Pb exposures that decreased serum testosterone
levels increased mRNA levels of GnRH and LH in the hypothalamus and pituitary, respectively,
and increased levels of LH in pituitary; these changes can occur in the absence of a change in
serum gonadotropin levels (Klein et al., 1994; Ronis et al., 1998c; Sokol et al., 2002).
In monkeys, chronic Pb exposures (blood Pb 20 to 35 |ig/dL) suppressed GnRH-induced
secretion of LH and decreased serum testosterone:LH and inhibin:FSH ratios (Foster et al.,
1993). The mechanisms underlying the effects on the hypothalamic-pituitary-gonadal axis have
not been elucidated but may involve a suppression of GnRH secretion (Bratton et al., 1994;
Sokol, 1987; Sokol etal., 1998).
Although there is evidence for a common mode of action, consistent effects on circulating
testosterone levels are not always observed in Pb-exposed animals. Rodamilans et al. (1988) and

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Kempinas et al. (1994) attributed these inconsistencies to the normal biological variation
(circannual and seasonal) of testosterone secretion in rats and monkeys. Observations of
Pb-induced reductions in testosterone levels in some studies, but not others, may be due to
enhanced sensitivity to inhibition of the testosterone secretory system during certain periods of
development. In addition, compensatory mechanisms in the hypothalamic-pituitary-gonad axis
may attenuate some effects of Pb during prolonged Pb exposure (Sokol et al., 2002). Taken
together, the sensitivity of testosterone secretion during certain periods and potential for
modulation of the effects during long-term exposures studies, may explain some of the apparent
inconsistencies in the reported effects of Pb exposure on circulating testosterone levels.

5.4.2.4 Effects on Morphology and Histology of Male Sex Organs
The 1986 Lead AQCD reported evidence for histological changes in the testes or prostate
in rats, in association with relatively high doses of Pb (Chowdhury et al., 1984; Hilderbrand
et al., 1973; Golubovich et al., 1968). More recent studies conducted in animal models provide
persuasive support for testicular damage, i.e., ultrastructural changes in testes and cytotoxicity in
Sertoli cells (Foster et al., 1998; Singh et al., 1993a; Batra et al., 2001; Chowdhury et al., 1986,
1987; Corpas et al., 1995; Pinon-Lataillade et al., 1993; Saxena et al., 1990). Studies conducted
in nonhuman primates warrant particular attention. These studies found ultrastructural changes
in the testes (Sertoli and other spermatogenic cells) of monkeys at blood Pb 35 to 40 |ig/dL
(Foster et al., 1998; Singh et al., 1993a).
Foster et al. (1998) reported that chronic Pb exposure (blood Pb -35 |ig/dL), beginning in
infancy, resulted in persistent ultrastructural changes in the testes of cynomolgus monkeys.
Electron microscopy showed disruption of the general structure of the seminiferous epithelium
involving Sertoli cells, basal lamina, and spermatids in the groups exposed for lifetime and
during infancy only (no duration difference in severity). Chronic exposures to Pb beginning
after infancy, that achieved similar blood Pb levels, did not produce these effects.
Similarly, Singh et al. (1993a) demonstrated ultrastructural changes in testicular basement
membrane and Sertoli cell morphology (seminiferous tubules) in cynomolgus monkeys exposed
chronically to Pb (blood Pb <40 to 50 |ig/dL); the effects were most prominent when dosing
began in infancy or post-infancy. These results suggest that, in monkeys, Pb exposure during
5-83
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certain periods of development produces persistent testicular alterations. Additional details
concerning Foster et al. (1998) and Singh et al. (1993a) are provided in Table 5-3.
A possible mode of action for Pb-induced testicular injury is oxidative stress. Foster et al.
(1998) suggested that Pb-induced oxygen free radical generation was a plausible mechanism of
testicular injury in primates. This oxygen radical hypothesis is supported by studies conducted
in rodents (Chowdhury et al., 1984; Acharya et al., 2003; Adhikari et al., 2001; Batra et al.,
2001; Bizarro et al., 2003; Chowdhury et al., 1984; Gorbel et al., 2002; Mishra and Acharya,
2004). Also supporting the oxidative stress hypothesis are observations of increases in the
percentage of apoptotic cells in the testes of rodents in response to Pb exposure (Pace et al.,
2005; Gorbel et al., 2002; Adhikari et al., 2001).
Studies in experimental animals (assessed in the 1986 Lead AQCD and others published
subsequent to the 1986 Lead AQCD) provide convincing evidence that Pb acts as an endocrine
disrupter in males. The majority of present studies support the conclusion that endocrine
disruption in males involves Pb acting at multiple sites along the hypothalamic-pituitary-gonadal
axis. The adverse effects of Pb on male reproduction include perturbations in sexual
development and maturation, changes in fertility, changes in male sex hormone levels, and
alterations in gonad tissues and cell structure.

5.4.3 Effects on Female Reproductive Function
Lead has been shown to disrupt the hypothalamic-pituitary-gonadal axis and to produce
ovarian atrophy and reproductive dysfunction in females (Figure 5-7). The 1986 Lead AQCD
reported that Pb exposure was associated with inhibition of menstruation, ovulation, and
follicular growth in monkeys (Vermande-Van Eck and Meigs, 1960), and, in rodents, Pb
exposure delayed vaginal opening, decreased frequency of implantation, and reduced rates of
pregnancy (Kimmel et al., 1980; Odenbro and Kihlstrom, 1977, respectively).
Data from more recent experimental animal studies support these findings. Lead effects
on female reproduction may be classified as alterations in female sexual maturation, effects on
fertility and menstrual cycle, endocrine disruption, and changes in morphology or histology or
female reproductive organs as well as the placenta. Recent literature concerning each of these
effects is summarized below.
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5.4.3.1 Effects on Female Sexual Development and Maturation
The 1986 Lead AQCD reported that Pb exposure in rodents produced delays in sexual
maturation. Grant et al. (1980) reported delayed vaginal opening in female rats exposed in utero
and during lactation and maturation (blood Pb -20 to 40 |ig/dL). More recent studies in
experimental animals (primarily rodent studies) provide convincing evidence that Pb exposure
before puberty (particularly prenatal and early postnatal exposure) delays the maturation of the
female reproductive system (Dearth et al., 2002, 2004; Ronis et al., 1996, 1998b,c).
The study of Dearth et al. (2002) is of particular interest because it employed a
cross-fostering design (to allow comparison of pups exposed during gestation only, lactation
only, or both) and because maternal and offspring blood Pb were monitored throughout gestation
and lactation. Fisher 344 dams were exposed to Pb by gavage beginning 30 days before mating
until weaning of the pups at 21 days of age (gavage exposure removes possible confounding of
exposure by consumption of Pb in drinking water by pups in those studies where drinking water
is the route of exposure for dams). Mean maternal blood Pb level was -40 |ig/dL. Pups exposed
during gestation and lactation had the highest blood Pb (38.5 |ig/dL) on day 10; at this time, the
blood Pb levels in pups exposed during gestation only or lactation only were 13.7 and
27.6 |ig/dL, respectively. By postnatal day (PND) 30, all three groups had blood Pb <3 |ig/dL.
Dearth et al. (2002) reported a statistically significant delay in the onset of puberty (vaginal
opening and days at first diestrus) in rats exposed during lactation, gestation, or during lactation
and gestation (with no differences among the groups). In addition, statistically significant
reductions in the circulating levels of insulin-like growth factor 1 (IGFi), LH, and estradiol (E2)
were reported on PND 30 in all three treatment groups (with no differences among treatment
groups). Additional details concerning Dearth et al. (2002) are provided in Table 5-4.
A subsequent study in both Sprague-Dawley and F344 rats (Dearth et al., 2004) showed
that the F344 strain is more sensitive to maternal Pb exposure than Sprague-Dawley rats to
Pb-induced delayed puberty, which could, in part, explain the inconsistencies with effect levels
observed in Sprague Dawley rats (e.g., Ronis et al., 1998a,b,c; McGivern et al., 1991). Ronis
et al. (1998c) suggested that the delayed onset of puberty may arise from a Pb-induced disruption
of pulsatile release of sex hormones (see Section 5.4.3.3).
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Table 5-4. Selected Studies Showing the Effects of Lead on Reproductive Function in Females
Species/ Dose/Route/Form/Duration/
Citation Strain Group Size
Endpoint/Magnitude of Effect
(% or incidence)/p-value
Blood Lead Concentration
(PbB)
Dearth
etal.
(2002)
oo
Foster
(1992)
Foster et al.
(1992)
Foster et al.
(1996b)
Rat/Fisher
344
Monkey/
Cynomolgus
Monkey/
Cynomolgus
Monkey/
Cynomolgus
12 mg/mL Pb acetate gavage from
30 days prior breeding until pups were
weaned 21 day afterbirth; 10-32 litters
per group, control group, gestation and
lactation exposure, gestation only
exposure, lactation only exposure
Daily dosing for up to 10 years with
gelatin capsules containing Pb acetate
(1.5 mg/kg); 8 control group monkeys,
8 lifetime exposure (birth-10 years),
8 childhood exposure (birth-400 days),
and 8 adolescent exposure
(PND 300-10 years of age)

Daily dosing for up to 10 years with
gelatin capsules containing Pb acetate
(1.5 mg/kg); 8 control group monkeys,
8 childhood (birth-400 days),
7 adolescent (PND 300-10 years),
7 lifetime (birth-10 years)

Chronic exposure to Pb acetate 50 to
2000 ug/kg-day p.o. beginning at birth
for 15-20 years; 20 control monkeys,
4 monkeys in 50 ug/kg-d group,
3 monkeys in 100 ug/kg-d, 2 monkeys
in 500 ug/kg-d group, and 3 monkeys in
2000 ug/kg-d group
Delay in onset of puberty (p < 0.05); reduced
serum levels of IGFj (p < 0.001), LH
(p< 0.001), and E2 (p < 0.001).
Statistically significant reductions in circulating
levels of LH, (p < 0.042), FSH (p < 0.041), and
E2 (p < 0.0001) during menstrual cycle;
progesterone concentrations were unchanged and
menstrual cycle was not significantly affected.
No effect on endometrial response to gonadal
steroids as determined by ultrasound.
Reduced corpora luteal production of
progesterone (p = 0.04), without alterations in
E2, 20-alpha-hydroxyprogesterone, or menstrual
cyclicity.
Maternal PbB -40 ug/dL
Pups PbB as follows:
Gest+lact -38 ug/dL PND 10
Gest+lact -15 ug/dL PND 21
Gest+lact -3 ug/dL PND 30
Gest -14 ug/dL PND 10
Gest~3 ug/dL PND 21
Gest~l ug/dL PND 30
Lact-28 ug/dL PND 10
Lact-15 ug/dL PND 21
Lact-3 ug/dL PND 30

PbB <40 ug/dL
PbB <40 ug/dL
PbB 10-15 ug/dL in low group
(50 or 100 ug/kg-day)
PbB 25-30 ug/dL in moderate
group (500 or 2000 ug/kg-day)
-------
Table 5-4 (cont'd). Selected Studies Showing the Effects of Lead on Reproductive Function in Females
Species/ Dose/Route/Form/Duration/
Citation Strain Group Size
Endpoint/Magnitude of Effect
(% or incidence)/p-value
Blood Lead Concentration
(PbB)
oo
Franks Monkey/ Lead acetate in drinking water (2-8
etal. Rhesus mg/kg-d) for 3 3 months; 7 control
(1989) and 10 Pb monkeys
Laughlin Monkey/ Lead acetate in drinking water at 3.6,
et al. Rhesus 5.9, or 8.1 mg/kg-day for 1-2 years

7 control and 10 experimental monkeys
per group

Logdberg Monkey/ Lead acetate (varying concentrations
et al. Squirrel <0.1% in diet) maternal dosing from
(1988) 5-8.5 weeks pregnant to PND 1

11 control monkeys, 3 low Pb exposure
group (PbB 24 ug/dL), 7 medium Pb
group (PbB 40 ug/dL, 5 high Pb group
(PbB 56 ug/dL)
Reduced circulating concentration of
progesterone (p < 0.05); treatment with Pb did
not prevent ovulation, but produced longer and
more variable menstrual cycles and shorter
menstrual flow.

Reductions in cycle frequency (p < 0.01); fewer
days of flow (p < 0.01); longer and more variable
cycle intervals (p < 0.025).
Dose-dependent reduction in placental weight
(p < 0.0007); various pathological lesions were
seen in the placentas (n = 4), including
hemorrhages, hyalinization of the parenchyma
with destruction of the villi and massive
vacuolization of chorion epithelium.
PbB 68.9 ± 6.54 ug/dL
PbB 44-89 ug/dL
51.2 ug/dL (low dose)
80.7 ug/dL (mid dose)
88.4 ug/dL (high dose)

Mean maternal PbB 37 ug/dL
(22-82 ug/dL)
24 (22-26) ug/dL (low dose)
40 (35-46) ug/dL (mid dose)
56 (43-82) ug/dL (high dose)
E2, estradiol; FSH, follicle stimulating hormone; GD, gestational day; IGF!, insulin-like growth factor 1; LH, luteinizing hormone; PbB, blood Pb
concentration; PND, post-natal day
-------
5.4.3.2 Effects on Female Fertility
The 1986 Lead AQCD reported convincing evidence from experimental animal studies
for Pb-induced alterations in female fertility, including interference with implantation and
pregnancy (Odenbro and Kihlstrom, 1977; Wide and Nilsson, 1977). More recent studies have
confirmed these effects. In general, Pb exposure does not produce total sterility, although Pb
exposure clearly disturbs female fertility (Taupeau et al., 2001). Studies in nonhuman primates
and rodents have shown that exposure of gravid females to Pb produces implantation dysfunction
and reduces litter size and newborn survival (Logdberg et al., 1987; Flora and Tandon, 1987;
Johansson and Wide, 1986; Pinon-Lataillade et al., 1995; Piasek and Kostial, 1991; Ronis et al.,
1996). See Section 5.4.4.1 for details.

5.4.3.3 Effects on the Female Sex Endocrine System and Menstrual Cycle
The 1986 Lead AQCD described numerous studies that found effects of Pb on the female
endocrine system and menstrual cycle in various species, including nonhuman primates, and that
supported the conclusion that Pb was an endocrine disrupter in females (Grant et al., 1980;
Maker et al., 1975; Vermande-Van Eck and Meigs, 1960). Observations of delayed vaginal
opening (see Section 5.4.3.1) were attributed to the endocrine disruption effects of Pb on the
hypothalamic-pituitary-gonadal axis (Stowe and Goyer, 1971; Vermande Van Eck and Meigs,
1960).
More recent studies have provided convincing support for endocrine-mediated
alterations of the female reproductive system in rats (Srivastava et al., 2004; Dearth et al.,
2002; Ronis et al., 1998a,b,c; Junaid et al., 1997; Ronis et al., 1996), guinea pigs (Sierra and
Tiffany-Castiglioni, 1992), and nonhuman primates (Foster et al., 1992, 1996b; Foster, 1992;
Franks et al., 1989; Laughlin et al., 1987). The nonhuman primate studies are particularly
relevant to extrapolations to humans and provide dose-response information for effects of Pb on
female sex hormones and menstrual cycle.
Laughlin et al. (1987) found that Pb exposure (blood Pb 44 to 89 |ig/dL) alters menstrual
cycles (specifically, causing reductions in cycle frequency, fewer days of menstrual flow, and
longer and more variable cycle intervals) in female rhesus monkeys. Consistent with these
observations, Franks et al. (1989) found that chronic exposure to Pb in the drinking water (blood
Pb 70 |ig/dL) reduced circulating concentrations of progesterone (suggesting impaired luteal
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function), produced longer and more variable menstrual cycles and temporally shorter menstrual
flow in female rhesus monkeys. Additional details concerning these studies are provided in
Table 5-4.
At lower blood Pb levels (<40 |ig/dL), female cynomolgus monkeys exhibited statistically
significant reductions in circulating levels of LH, FSH, and £2 during the menstrual cycle;
however, serum progesterone concentrations were unchanged and menstrual cycle was not
significantly affected (Foster, 1992). Similar exposures and blood Pb levels were shown to have
no effect on endometrial response to gonadal steroids in cynomolgus monkeys as determined by
ultrasound analysis (Foster et al., 1992). At lower blood Pb concentrations (25 to 30 |ig/dL),
reduced corpora luteal production of progesterone occurred in the absence of alterations in E2,
20-alpha-hydroxyprogesterone, or menstrual cyclicity (Foster et al., 1996b). In contrast to Foster
et al. (1992), this study (Foster et al., 1996b) found no statistically significant effect of Pb on
serum progesterone levels in cynomolgus monkeys that had lower blood Pb levels (10 to
15 jig/dL). Additional details concerning these studies are provided in Table 5-4.
Several modes of action for Pb-induced, endocrine disruption-mediated alterations in
female reproduction have been proposed, including changes in hormone synthesis or metabolism
at the enzyme level (Wiebe and Barr, 1988; Wiebe et al., 1988) and changes in hormone receptor
levels (Wiebe et al., 1988; Wide and D'Argy, 1986). In addition, Pb may alter sex hormone
release and imprinting during early development (Ronis et al., 1998c; Tchernitchin et al.,
1998a,b). The latter effects would be consistent with observations of persistent changes in
estrogen receptor levels in the uterus (Wiebe and Barr, 1988) and LH function in the ovary
(Srivastava et al., 2004) in Pb-exposed animals.

5.4.3.4 Effects on Morphology and Histology of Female Sex Organs and the Placenta
Lead-induced changes in morphology or histology in female sex organs and the placenta
may explain reduced fertility and impaired female reproductive success (see Sections 5.4.3.2 and
5.4.4.1.). Logdberg et al. (1988) reported a dose-dependent reduction in placental weight and an
increase in pathological lesions of the placenta in squirrel monkeys that received oral doses of Pb
acetate (0.001 to 0.1% in diet) during the last three-fourths or two-thirds of pregnancy (mean
maternal blood Pb 37 |ig/dL; range: 22 to 82 |ig/dL). These effects occurred without overt
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toxicity in the mothers. Additional details concerning Logdberg et al. (1988) are provided in
Table 5-4.
Similar effects on placental weight and histology were observed in mice (Fuentes et al.,
1996; Nayak et al., 1989a). These effects on the placenta may explain the reduced birth weight
that has been associated with prenatal Pb exposure (see Section 5.4.5). Exposure to Pb in early
pregnancy also produces structural changes in the epithelium of the uterus of mice (Nilsson
et al., 1991; Wide and Nilsson, 1979). These changes in uterine tissue may impair successful
implantation of the blastocysts (see Section 5.4.4.1).
The 1986 Lead AQCD reported that Pb exposure (blood Pb 20 to 40 |ig/dL) in rodents
produced delays in sexual maturation. More recent studies in experimental animals (primarily
rodent studies) provide convincing evidence that Pb exposure before puberty (prenatal and early
postnatal blood Pb -40 |ig/dL) delays maturation of the female reproductive system (Dearth
et al., 2002, 2004; lavicoli et al., 2004; McGivern et al., 1991; Ronis et al., 1998a,b,c,). Ronis
et al. (1998c) suggested that lead-induced disruption of pulsatile release of sex hormones may
result in delayed onset of puberty.

5.4.4 Effects on Embryogenesis
Lead exposure can increase fetal mortality, produce a variety of sublethal effects, and
disrupt the growth and development of the offspring. Many of the lead-induced sublethal
developmental effects occur at maternal blood Pb levels that do not result in clinical toxicity in
the mothers.

5.4.4.1 Embryo/Fetal Mortality
The 1986 Lead AQCD concluded that that acute exposure to high doses of Pb interfered
with implantation and pregnancy (Wide, 1985; Odenbro and Kihlstrom, 1977; Wide and Nilsson,
1977; Vermande-Van Eck and Meigs, 1960). This conclusion is supported by results of more
recent studies (Logdberg et al., 1987; Giavini et al., 1980; Jacquet, 1976, 1977; Jacquet et al.,
1975, 1976; Johansson and Wide 1986; Johansson et al., 1987; Johansson, 1989; Maisin et al.,
1978;Pinon-Latailladeetal., 1995; Wide and Nilsson, 1977, 1979).
Logdberg et al. (1987) reported an increase in pre- and perinatal mortality in squirrel
monkeys that received Pb acetate orally during the last two-thirds of pregnancy (45% versus
5-90
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7 to 8% among controls). The mean maternal blood Pb level was 54 |ig/dL (39 to 82 |ig/dL).
These fetotoxic effects occurred without overt toxicity in the mothers. Additional details
concerning Logdberg et al. (1987) are provided in Table 5-5. These effects are consistent with
data from rodent studies, wherein gestational exposure to Pb (blood Pb 32 to >70 |ig/dL) resulted
in smaller litters and fewer implantation sites (e.g., Pinon-Lataillade et al., 1995; Singh et al.,
1993b; Piasek and Kostial, 1991). Numerous studies have been performed to elucidate possible
mechanisms by which Pb causes prenatal death (Maisin at al., 1978; Jacquet, 1977, 1976;
Jacquet et al., 1976, 1975). The available data suggest that Pb may alter blastocyst development
and impair implantation. Hanna et al. (1997) demonstrated that in vitro exposure of 2- and 4-cell
mouse embryos to 200 jiM Pb acetate resulted in reduced cell proliferation and blastocyst
formation. Additional evidence for an effect on blastocysts is provided by data from in vitro
fertilization studies (Chowdhuri et al., 2001; Johansson, 1989; Johansson et al., 1987).
Johansson and co-workers (1989, 1987) reported that Pb delayed the timing of escape of
spermatozoa from the zona pellucida and induced a premature acrosome reaction. These effects
could disrupt attachment and implantation of the blastocyst if they were to occur in vivo.
Observations from more recent experimental animal studies support these findings. The
effects of Pb on female reproduction may be classified as alterations in female sexual maturation,
effects on fertility and menstrual cycle, alterations in levels of female sex hormones, and changes
in morphology or histology of female reproductive organs as well as the placenta.

5.4.4.2 Effects on embryo/fetal morphology
The 1986 Lead AQCD summarized numerous reports that found associations between
prenatal exposure to high doses of Pb and increased incidences of teratogenic effects
(particularly tail stunting) in rodents (Perm and Carpenter, 1967, Wide, 1985). More recent
studies provide additional support for the teratogenic effects of lead in experimental animals
(Dey et al., 2001; Flora and Tandon, 1987; Ronis et al., 1996). However, the few studies
(including those described in the 1986 Lead AQCD and more recent reports) that have
demonstrated teratogenic effects of Pb exposure are confounded by maternal toxicity.
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Table 5-5. Selected Studies Showing the Effects of Lead on Mammalian Embryogenesis and Development
Citation
Species/
Strain
Dose/Route/Form/Duration/Group size Endpoint/Magnitude of Effect/p-value
Blood Lead Concentration
(PbB)
Cory-Slechta
et al. (2004)
Rat/Long-
Evans
Dearth et al. Rat/Fisher
(2002) 344
vo
to
Flora and
Tandon
(1987)
Fox et al.
(1991a)
Rat/Albino
(NOS)
Rat/Long-
Evans
hooded
Lead acetate in drinking water (150 ppm);
2 months before breeding until the end of
lactation; 14 rats no maternal stress with Pb
exposure, 15 rats no maternal stress with Pb
exposure, 18 rats maternal stress without Pb
exposure, 23 rats maternal stress and Pb
exposure

12 mg/mL Pb acetate gavage during gestation
and lactation exposure
4 groups: control group, gestation and
lactation exposure, gestation only exposure,
lactation only exposure
10-32 litters per group (NOS)
Lead nitrate dissolved in water 2-20 mg/kg-d
i.v. on day 9, 10, 11 of gestation; 6 rats in
each group (0, 5, 10 , 20, 40 mg/kg lead)
Lactation exposure via dams exposed to
0.02 or 0.2% Pb in drinking water from PND
1 through weaning (PND 21); 8 female pups
per litter (number of litter unspecified) control
pups, 8 pups for litter (number of litter
unspecified) low-level exposure pups, 8 pups
per litter (number of litter unspecified)
moderate level exposure pups
Pb alone (in male) (p < 0.05) and Pb plus stress
(in females) (p < 0.05) permanently elevated
corticosterene levels in offspring
Delayed onset of puberty (p < 0.05); suppressed
serum levels of IGFb LH, and E2 (p < 0.001);
Pb altered translation and/or secretion of IGF!
(p< 0.001).
Dose-dependant increase in external malformations
at all doses (p < 0.001), particularly tail defects;
dose-dependant decrease in number of live births
at 20 and 400 mg/kg (p < 0.001); dose-dependent
increase in number of resorptions per dam at
< 10 mg/kg (p< 0.01).

Long-term, dose-dependent decreases retinal Na/K
ATPase activity in the female offspring (only
female pups were used) (-11%; -26%) (p < 0.05).
PbB 30^10 ng/dL
Maternal PbB -40 ng/dL
Pups PbB as follows:
Gest+lact~38 ng/dL PND 10
Gest+lact -15 ng/dL PND 21
Gest+lact ~3 ng/dL PND 30
Gest~14|ig/dLPND10
Gest~3|ig/dLPND21
Gest~l |ig/dLPND30
Lact~28|ig/dLPND10
Lact~15|ig/dLPND21
Lact ~3 ng/dL PND 30

PbB 4.13 ± 0.61 ng/dL 0 mg/kg
PbB 10.21 ± 0.61 ng/dL 5 mg/kg
PbB 13.13 ± 0.27 ng/dL 10 mg/kg
PbB 29.41 ± 0.41 ng/dL 20 mg/kg
PbB 45.03 ± 0.31 ng/dL 40 mg/kg
PbB 18.8 ng/dL (0.02%) or
59.4 ng/dL (0.2%) at weaning
-------
Table 5-5 (cont'd). Selected Studies Showing the Effects of Lead on Mammalian Embryogenesis and Development
Citation
Species/
Strain
Dose/Route/Form/Duration/Group Size Endpoint/Magnitude of Effect/p-value
Blood Lead Concentration
(PbB)
Fox et al.
(1997)
Rat/Long-
Evans
hooded
vo
lavicoli et al.
(2003)
Mouse/Swiss
0.02 or 0.2% Pb acetate in drinking water
from PND 0-PND 21; 8 female pups per litter
control pups; 8 pups per litter low-level
exposure; 8 pups per litter moderate level
exposure (number of litters per dose
unspecified)
Lead acetate in food (0.02, 0.06, 0.11, 0.2, 2,
4, 20, 40 ppm); exposure began 1 day after
mating until litter was 90 days old; one litter
of mice exposed to each dietary concentration
Developmental and adult Pb exposure for 6 weeks
produced age and dose-dependent retinal
degeneration such that rods and bipolar cells were
selectively lost; at the ultrastructural level, all
dying cells exhibit the classical morphological
features of apoptotic cell death; decrease in the
number of rods was correlated with the loss of
rhodopsin content per eye confirming that rods
were directly affected by Pb (p < 0.05); single-
flash rod ERGs and cone ERGs obtained from
lead-exposed rats demonstrated that there were
age- and dose-dependent decreases in the rod a-
wave and b-wave sensitivity and maximum
amplitudes without any effect on cones; in adult
rats exposed to Pb for three weeks, qualitatively
similar ERG changes occurred in the absence of
cell loss or decrease in rhodopsin content
(p < 0.05); developmental and adult Pb exposure
for three and six weeks produced age- and dose-
dependent decreases in retinal cGMP
phosphodiesterase (PDE) activity resulting in
increased cGMP levels (p < 0.05); retinas of
developing and adult rats exposed to Pb exhibit
qualitatively similar rod mediated ERG alterations
as well as rod and bipolar apoptotic cell death
(p<0.05).
Similar biochemical mechanism such as the
inhibition of rod and bipolar cell cGMP PDE,
varying only in degree and duration, underlies both
the lead-induced ERG rod-mediated deficits and
the rod and bipolar apoptotic cell death (p < 0.05).
Low-level Pb exposure (PbB 2-13 ug/dL) reduced
red cell synthesis (p < 0.05); high-level exposure
(PbB 0.6-2 ug/dL) enhanced red cell synthesis
(p<0.05).
PbB weanlings 19 ± 3 (low exposure)
or 59 ± 8 ug/dL (moderate exposure),
adult 7 ± 2 ug/dL (at PND 90)
PbB 0.6 to <2.0 ug/dL or
>2.0-13 ug/dL
-------
Table 5-5 (cont'd). Selected Studies Showing the Effects of Lead on Mammalian Embryogenesis and Development
Citation Species/Strain
Dose/Route/Form/Duration/Group
size
Endpoint/Magnitude of effect/p-value
Blood Lead Concentration
(PbB)
Logdberg Monkey/Squirrel Lead acetate (5-20 mg/kg daily to maintain
et al. PbB) maternal dosing from 5-8.5 weeks
(1987) pregnant to PND1 20 control; 11 lead-
exposed monkeys
Ronis Rat/Sprague- 0.6% Pb acetate in drinking water for
et al. Dawley various durations: PND 24-74 (pubertal
(1996) exposure); PND 60-74 (post pubertal
exposure); 11 males and females in pubertal
exposure group (10 each in control pubertal
group)
6 males and females post-pubertal exposure
and control groups
Ronis Rat/Sprague- 0.6% Pb acetate in drinking water ad libitum
et al. Dawley for various durations; GD 5 to PND 1; GD 5
(1998a) to weaning; PND 1 to weaning; 3 control
litters, 2 gestation exposure litters, 2
lactation exposure litters, 2 gestation and
lactation exposure litters, 2 postnatal litters,
2 chronic litters (4 male and 4 female pups
per litter)
Increase in pre- and perinatal mortality among
squirrel monkeys receiving Pb acetate p.o. during
the last two-thirds of pregnancy (45% vs. 7-8%
among controls). Statistically significant
reductions in mean birth weight (p < 0.05) were
observed in Pb exposed monkeys as compared to
controls. Effects occurred without clinical
manifestation of toxic effects in the mothers.

Reduction in serum testosterone levels in male, not
female; in female suppression of circulating E2
(p < 0.05) and LH (p < 0.05); reduction in male
secondary sex organ weight (p < 0.0005); delayed
vaginal opening and disrupted diestrous in females
(p < 0.005); increased incidence of stillbirth (2%
control vs. 19% Pb) (p < 0.005).
Dose-dependent delay in sexual maturation
(delayed vaginal opening) (p < 0.0002) following
prenatal Pb exposure that continued until
adulthood (85 days old); reduced birth weight
(p < 0.05), more pronounced among male pups.
PbB 54 ng/dL (39-
82
In utero PbB 250-
300 ng/dL
pre-pubertal PbB
30-60 ng/dL
post pubertal PbB
30-60 ng/dL
PbBs in the dams
and offspring in this
experiment were
>200 ng/dL
Naive

Control

Gest

Lac

Gest+Lac

Postnatal

Chronic
-------
Table 5-5 (cont'd). Selected Studies Showing the Effects of Lead on Mammalian Embryogenesis and Development
Species/
Citation Strain
Dose/Route/Form/Duration/Group Size Endpoint/Magnitude of Effect/p-value
Blood Lead Concentration
(PbB)
Ronis
etal.
(1998b)
Rat/Sprague-
Dawley
vo
Ronis
etal.
(1998c)
Rat/Sprague-
Dawley
Ronis
etal.
(2001)
Rat/Sprague-
Dawley
Lead acetate in drinking water (0.05% to
0.45% w/v); dams exposed until weaning;
exposure of pups which continued until PND
21,35, 55, or 85
5 control litters (0%), 10 low-dose litters
(0.05%), 8 mid-dose litters (0.15%), 9 high-
dose litters (0.45%)
(4 male and 4 female pups per litter)

Lead acetate 0.05, 0.15, or 0.45% in drinking
water beginning GD 5 continuing until PND
21, 35, 55, or 85; 5 control litters (0%),
10 low-dose litters (0.05%), 8 mid-dose litters
(0.15%), 9 high-dose litters (0.45%)
(4 male and 4 female pups per litter)
Lead acetate in drinking water to 825 or
2475 ppm ad libitum from GD 4 to GD 55
postpartum; 1 male and female pup/litter
(5 litters per group) control group, 1 male and
female pup/litter (5 litters per group) 825 ppm
Pb acetate group, 1 male and female pup/litter
(5 litters per group) 2475 ppm Pb acetate
group
Prenatal Pb exposure that continues until adulthood
(85 days old) delays sexual maturation in female pups
in a dose-related manner (p < 0.05); birth weight
reduced (p < 0.05), more pronounced among male
pups; decreased growth rates (p < 0.05) in both sexes
accompanied by decrease in plasma concentrations of
IGF] through puberty (p < 0.05) and a significant
increase in pituitary growth hormone during puberty
(p<0.05).

Dose-responsive decrease in birth weight (p < 0.05),
and crown-to-rump length (p < 0.05); dose-responsive
delay in sexual maturity in male (p < 0.05) and female
(p < 0.05); neonatal decrease in sex steroids
(p < 0.05); pubertal decrease in testosterone (male)
(p < 0.05), and E2 (female) (p < 0.05); decrease estrous
cyclicity at high dose (p < 0.05).
Dose-dependent decrease of the load of failure in male
(p < 0.05); no difference in plasma levels of vitamin D
metabolites; reduced somatic growth (p < 0.05),
longitudinal bone growth (p < 0.05, and bone strength
during the pubertal period (p < 0.05); sex steroid
replacement did not restore skeletal parameters in Pb
exposed rats; L-Dopa increased plasma IGF]
concentrations, rates of bone growth, and bone
strength measures in controls while having no effect in
Pb exposed groups; DO gap x-ray density and
proximal new endostreal bone formation were
decreased in the distration gaps of the lead-treated
animals (p < 0.01); distraction initiated at 0.2 mm
30 to 60 days of age.
PbBs in the pups between the ages of
21 and 85 days were >100 ug/dL and
reached up to 388 ug/dL
Dams: 0, 48, 88, or 181 ug/dL
Pups PND 1: <1, -40, -70, or
>120 ug/dL
Pups PND 21: <1,>50,>160, or
-237 ug/dL
Pups PND 35: <1,~22, >70, or
>278 ug/dL
Pups PND 55: <1,>68, >137, or
-380 ug/dL
Pups PND 85: <1,>43, >122, or
>214 ug/dL

PbB at 825 ppm was 67-192 ug/dL
PbB at 2475 ppm was 120-388 ug/dL
-------
Table 5-5 (cont'd). Selected Studies Showing the Effects of Lead on Mammalian Embryogenesis and Development
Species/
Citation Strain
Dose/Route/Form/Duration/Group Size Endpoint/Magnitude of Effect/p-value
Blood Lead Concentration
(PbB)
Rat/Sprague- Lead in drinking water at 34 ppm from
Dawley weaning of mothers through gestation and
weaning of offspring until birth; 6 pups
control group, 6 pups experimental group
Reduced body weight (p = 0.04); parotid function was
decreased by nearly 30% (p = 0.30); higher mean
caries scores than the control pups (p = 0.005); pre-
and perinatal Pb exposure had significantly increased
susceptibility to dental caries (p = 0.015).
PbB48±13|ig/dL
GMP, cyclic guanosine—3',5'-monophosphate; DO, distraction osteogenesis; E2, estradiol; ERG, electroretinographic; GD, gestational day; IGF!, insulin-like growth factor 1;
LH, luteinizing hormone; NOS, not otherwise specified; PbB, blood Pb concentration; PDE, phosphodiesterase; PND, post-natal day
vo
-------
5.4.5 Effects on Growth and Endocrine Regulation of Growth
Studies conducted in rodents provide convincing evidence for an association between
gestational Pb exposure and reduced birth weight and postnatal growth at doses that produce no
clinical toxicity in the mothers (Dearth et al., 2002; Hamilton et al., 1994; Logdberg et al., 1987;
Piasek and Kostial, 1991; Pinon-Lataillade et al., 1995; Ronis et al., 1998a,b,c; Singh et al.,
1993b; Watson et al., 1997). In squirrel monkeys, Logdberg et al. (1987) reported a statistically
significant reduction in mean birth weight following oral exposure to Pb acetate during the latter
trimesters of pregnancy (mean maternal blood Pb 54 |ig/dL [39 to 82 |ig/dL]). Additional details
concerning Logdberg et al. (1987) are provided in Table AX5-4.3.
In addition, the literature provides convincing support for Pb-induced impairment of
postnatal growth. Although some early studies (Minnema and Hammond, 1994; Hammond
et al., 1993, 1990) ascribed the reduction in postnatal growth to reduced food consumption
(suggesting an effect of Pb on the satiety endpoint), more recent studies report impaired growth
unrelated to changes in food consumption. Ronis et al. (1996, 1998a,b,c) reported Pb-induced
reductions in birth weight and postnatal growth that occurred in the absence of a significant
alteration in food consumption. Han et al. (2000) found a reduction in the birth length of pups
(pup blood Pb -16 |ig/dL on PND 1) whose mothers had been exposed to Pb up to 1 month
before mating (maternal blood Pb on GD 9, 16, and 21 <40 |ig/dL). Berry et al. (2002) reported
depressed growth in rats exposed to lead for six weeks beginning at weaning, even though food
consumption was higher in the lead-exposed rats.
Ronis et al. (2001) showed that in rats, pre- and postnatal (through PND 55) exposure to
Pb reduced somatic longitudinal bone growth and bone strength during the pubertal period
(blood Pb >67 |ig/dL). These effects could not be reversed by stimulation of the growth
hormone axis by supplemental sex hormone. These results suggest that Pb exposure may impair
growth through a mechanism that involves a suppressed pituitary response to hypothalamic
stimulation. The mechanism may be related to a reduction in plasma concentrations of IGFi
following Pb exposure (Dearth et al., 2002; Ronis et al., 1998b). Dearth et al. (2002) exposed
F344 rats to Pb by gavage beginning 30 days before mating and continuing until weaning of the
pups at 21 days of age. By PND 30, all three groups had blood Pb <3 |ig/dL and all lead-
exposed groups exhibited decreased serum levels of IGFi, LH, and £2. Since liver IGFi mRNA
was not affected, it appeared that Pb altered the translation and/or secretion of IGFi, which in
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turn decreased LH-releasing hormone at the hypothalamic level. Additional details concerning
Dearth et al. (2002) are provided in Table AX5-4.3. An effect on IGFi also been demonstrated
byRonisetal. (1998b).

5.4.6 Effects on Other Endocrine Systems during Development
Recent experimental animal studies provide evidence for an interaction between Pb
exposure and stress hormones, including glucocorticoids and catecholamines (Cory-Slechta
et al., 2004; Yu et al., 1996; Vyskocil et al., 1991; Saxena et al., 1990). Lead has been reported
to increase stress hormone levels (Vyskocil et al., 1991).
Cory-Slechta et al. (2004) reported a persistent effect of maternal Pb exposure (blood
Pb 30-40 |ig/dL) on corticosteroid levels in adult offspring. Both male and female offspring
born to dams exposed to lead exhibited elevated corticosteroid levels as adults. In female
offspring, the Pb effect was potentiated when maternal Pb exposure occurred in combination
with environmental stress (administered as restraint). These data suggest that brief exposures to
Pb during development may result in persistent changes in the hypothalamic-pituitary-adrenal
axis (e.g., fetal glucocorticoid programming). Additional details concerning Cory-Slechta et al.
(2004) are provided in Table AX5-4.3.
The interplay between Pb and stress hormones is consistent with the findings of Yu et al.
(1996) wherein neonatal exposure to Pb (blood Pb 70 |ig/dL) decreased cold-water swimming
endurance (a standard test for stress endurance). The enhancement of Pb-induced toxicity by
stress was also reported by Saxena et al. (1990) in adult male rats. Saxena et al. (1990) reported
enhanced testicular injury when rats were exposed to immobilization stress in combination with
Pb exposure (blood Pb >200 (ig/dL).

5.4.7 Effects on Other Organ Systems during Development
5.4.7.1 Developmental Effects on Blood and Liver
Recent data provide evidence for Pb-induced alterations in developing hematopoietic and
hepatic systems. However, the data concerning Pb exposure effects on the developing
hematopoietic system are limited. The 1986 Lead AQCD proposed that alterations in blood
ALAD activity and erythrocyte protoporphyrin were possible biomarkers for subtle, prenatal
effects of Pb on heme synthesis (Hayashi 1983a,b; Jacquet et al., 1977; Prigge and Greve, 1977;
5-98
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Hubermont et al., 1976). A more recent study (lavicoli et al., 2003) of Pb effects on RBC
production, Hb concentration, and Hct was not able to clearly establish a dose-response
relationship for these endpoints. Although limited by small group size (one litter per dose),
dietary exposure during conception, lactation, and through weaning to 90 days of age increased
red blood cell synthesis, blood hemoglobin concentration, and hematocrit in offspring that had
blood Pb levels in the range of 0.6 to 2 |ig/dL; and decreased red cell blood cell synthesis, blood
hemoglobin concentration, and hematocrit in offspring in that had blood Pb concentrations in the
range of 2 to 13 |ig/dL. More data are needed to clarify the effect of low-dose Pb exposure on
blood endpoints.
Two rodent studies provide limited suggestive evidence that Pb exposure during
development produces changes in hepatic enzymes and other biomarkers of hepatic function.
Pillai and Gupta (2005) reported that long-term exposure of rats (pre-mating, gestation, and
lactation) to Pb acetate (subcutaneous injections of 0.05 mg/kg-day; blood Pb not reported)
resulted in reduced activities of maternal hepatic steroid (£2) metabolizing enzymes (17-
p-hydroxy steroid oxidoreductase and UDP glucuronyl transferase) and decreased hepatic
CYP450 content. Corpas et al. (2002a) reported that exposure to Pb in drinking water exposure
during gestation and lactation (pup blood Pb -22 |ig/dL at PND 12 and PND 21) resulted in
alterations in the hepatic systems of neonates (PND 12) and pups (PND 21). The effects
manifested as alterations in several biochemical indicators of hepatic toxicity: reductions in Hb,
iron, alkaline and acid phosphatase levels, and hepatic glycogen, and elevated blood glucose.
These data suggest that Pb may alter hepatic function during development; however, more data
are needed to determine whether these effects are persistent.

5.4.7.2 Developmental Effects on Skin
Recent data provide limited evidence of altered soft tissue development resulting from Pb
exposure. The literature includes one report of Pb-induced abnormalities in skin development.
Dey et al. (2001) reported that the pups of mice exposed orally to Pb citrate (5 jig/kg-day)
throughout gestation exhibited a variety of skin anomalies, including perforations, cell deformity,
and disordered collagen bundles. Blood Pb levels of mothers and pups were not reported.
Although detailed biochemical studies are required to elucidate the mechanism for structural
5-99
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abnormalities, it appears that covalent binding of Pb ions to the sulfate group of
glycosaminoglycans may be involved.

5.4.7.3 Developmental Effects on the Retina
Several studies have found that Pb exposure during early postnatal development impairs
retinal development in female Long-Evans (LE) hooded rats (Fox et al., 1997, 1991a,b; Fox and
Rubenstein, 1989; Fox and Chu, 1988). Of these, two studies are particularly important.
Fox et al. (199 la) demonstrated that lactational exposure of LE hooded rats (blood Pb 18.8
or 59.4 |ig/dL) resulted in long-term, dose-dependent decreases retinal Na/K ATPase activity in
the female offspring (only female pups were used). Fox et al. (1997) subsequently demonstrated
that lactational exposure to female LE hooded rats (blood Pb 19 ± 3 or 59 ± 8 |ig/dL) or drinking
water exposure to adult females (blood Pb 56 ± 9 |ig/dL) resulted in differential age- and
dose-dependent alterations in retinal structure and function following low (blood Pb <20 |ig/dL)
and moderate (blood Pb <60 |ig/dL) Pb exposures during lactation or long-term (-60 days)
exposure during adulthood. The mode of action for the effects of Pb on retinal development may
be related to impaired Na/K ATPase activity (Fox et al., 1991a). The observation of reduced
enzyme activity in the retina, but not in the kidney, suggests specificity for the retinal alpha-3
isozyme of Na/K ATPase, rather than the renal alpha-1 isozyme of Na/K ATPase. The authors
suggested that this specificity may play a role in the target organ-specific toxicity of Pb (Fox
etal., 1991a).

5.4.8 Summary
The 1986 Lead AQCD presented unequivocal evidence (derived principally from studies
of rodents) for effects of Pb on reproduction and development in laboratory animals. This
included evidence for lethal effects in developing organisms exposed to Pb during gestation and
in the neonatal period, as well as a variety of sublethal effects on reproduction and development.
Sublethal effects included changes in levels or function of reproductive hormones, effects on
maturation of reproductive systems, persistent toxic effects on the gonads (both male and
female), and adverse effects on the conceptus. More subtle effects on hormone metabolism and
reproductive cell structure of developing organisms were also documented.
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• More recent studies support earlier conclusions, presented in the 1986 Lead AQCD, that Pb
can produce temporary and persistent effects on male and female reproductive function and
development and that Pb disrupts endocrine function at multiple points along the
hypothalamic-pituitary-gonadal axis.

• Studies conducted in male experimental animals unequivocally demonstrate that Pb exposure
during early development (blood Pb >30 |ig/dL) can delay the onset of puberty and alter
reproductive function later in life.

• Persistent effects of Pb exposure on the male reproductive system may derive from
disruption in pulsatile release of sex hormones during early development (Ronis et al.,
1998c).

• Experimental animal studies provide convincing evidence that Pb acts as an endocrine
disrupter in males at various points along the hypothalamic-pituitary-gonadal axis. Although
there is evidence for a common mode of action, consistent effects on circulating testosterone
levels are not always observed in Pb-exposed animals. The inconsistency in the reports of
circulating testosterone levels complicates the derivation of a dose-response relationship for
this endpoint.

• More recent studies in animals provide additional support for testicular damage (i.e.,
ultrastructural changes in testes and cytotoxicity in Sertoli cells) following exposure to Pb
and demonstrated ultrastructural changes in testes of monkeys at blood Pb levels of 35 to
40 |ig/dL. Lead-induced oxygen free radical generation is the plausible mechanism of
testicular injury in primates and rodents.

• Recent studies of various mammalian species provide convincing support for Pb-induced
endocrine-mediated alterations of the female reproductive system. The nonhuman primate
studies provide dose-response information concerning the effects of Pb on female sex
hormones and menstrual cycle.

• Exposures of monkeys to Pb resulting in chronic blood Pb levels <20 |ig/dL produce few
effects on circulating hormone levels and do not alter the menstrual cycle. Higher exposures
of monkeys to Pb (blood Pb >40 |ig/dL) alter circulating hormone levels and the menstrual
cycle, with more marked changes in these endpoints occurring at higher blood Pb levels.

• Several modes of action for Pb-induced alterations in female reproduction have been
proposed, including changes in hormone synthesis or metabolism and changes in hormone
receptor levels. In addition, Pb may alter sex hormone release and imprinting during early
development.

• More recent studies have confirmed that Pb exposure disturbs female fertility; however, Pb
exposure does not generally produce total sterility.

• Studies in nonhuman primates and rodents have also demonstrated reductions in litter size,
implantation dysfunction, and decreased postnatal survival following Pb exposure of gravid
female experimental animals (blood Pb >30 jig/dL).
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• Lead-induced changes in morphology or histology in female sex organs and placenta may
explain reduced fertility and impaired female reproductive success.

• Exposure to Pb in early pregnancy also produces structural changes in the epithelium of the
uterus of mice. These changes in uterine tissue may impair successful implantation of the
blastocysts.

• Histological and morphological effects on the uterus and placenta may explain the reduced
birth weight that has been associated with prenatal Pb exposure (possibly due to placental
insufficiency).

• Pre- and postnatal exposure to Pb has been demonstrated to result in fetal mortality and
produce a variety of sublethal effects in the offspring. Many of these Pb-induced sublethal
developmental effects occur at maternal blood Pb levels that do not result in clinical (overt)
toxicity in the mothers. The few studies that have reported teratogenic effects resulting from
Pb exposure are confounded by maternal toxicity.

• Studies conducted in rodents and primates provide convincing evidence for an association
between Pb exposure and reduced birth weight and postnatal growth at doses that produce no
clinical toxicity in the mothers (maternal blood Pb >40 |ig/dL).

• Recent experimental animal studies provide evidence for an interaction between Pb exposure
during development (blood Pb 30 to 40 |ig/dL) and stress hormones, including both
glucocorticoids and catecholamines.

• Lead exposure during early postnatal development (blood Pb -20 |ig/dL) impairs retinal
development in female Long-Evans hooded rats.

• In addition, recent studies provide limited evidence for Pb-induced alterations in developing
skin, and hematopoietic and hepatic systems.
5.5 CARDIOVASCULAR EFFECTS OF LEAD

5.5.1 Introduction

Numerous large and small epidemiological studies have attempted to examine the link
between Pb exposure and development of hypertension (HTN) in the general population and
occupationally exposed individuals. In addition, a number of studies have reported on other
Pb-associated cardiovascular effects in Pb-exposed humans (U.S. Environmental Protection
Agency, 1990). While several studies have demonstrated a positive correlation between blood
pressure and blood Pb concentration, others have failed to show such association when
controlling for confounding factors such as tobacco smoking, exercise, body weight, alcohol
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consumption, and socioeconomic status. Thus, the studies that have employed blood Pb level as
an index of exposure have shown a relatively weak association with blood pressure. In contrast,
the majority of the more recent studies employing bone Pb level have found a strong association
between long-term Pb exposure and arterial pressure (Chapter 6). Since the residence time of Pb
in the blood is relatively short but very long in the bone, the latter observations have provided
rather compelling evidence for a positive relationship between Pb exposure and a subsequent rise
in arterial pressure. This section reviews the published studies pertaining to the cardiovascular
effects of Pb exposure in experimental animals, isolated vascular tissues, and cultured vascular
cells.

5.5.2 Lead Exposure and Arterial Pressure in Experimental Animals
Numerous studies have shown that exposure to low levels of Pb for extended periods
results in a delayed onset of arterial HTN that persists long after the cessation of Pb exposure in
genetically normal animals (see Tables AX5-5.1 to AX5-5.5). In addition, Pb exposure during
gestation has been reported to significantly increase arterial pressure in the third trimester of
pregnancy in SD rats given a low calcium diet (Bogden et al., 1995). Taken together, these
observations provide irrefutable evidence that extended exposure to low levels of Pb can result in
the subsequent onset of HTN in experimental animals.
Many studies have been conducted to explore the mechanisms by which chronic Pb
exposure may cause HTN. Most of these studies have examined various blood-pressure
regulatory and vasoactive systems in animal models of Pb-induced HTN. In addition, several
studies have investigated the direct effect of Pb on vascular tone or the ability of Pb to modify
the response to vasoconstrictor/vasodilator agents in isolated vascular tissues. Finally, a number
of studies have explored the effect of Pb on cultured endothelial and vascular smooth muscle
cells. An overview of the findings of these studies is provided below.

5.5.2.1 Effect of Lead on Production of Reactive Oxygen Species and Nitric
Oxide Metabolism
Reactive oxygen species (ROS), such as superoxide ((V), hydroxyl radical (OH) and
hydrogen peroxide (H2O2) are normally produced in the course of metabolism and are safely
contained by the natural antioxidant defense system. Excess production and/or diminished
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containment of ROS can lead to oxidative stress in which uncontained ROS can attack and
denature functional/structural molecules and, thereby, promote tissue damage, cytotoxicity, and
dysfunction. In fact, oxidative stress has been implicated in the pathogenesis of HTN,
atherosclerosis, neurodegenerative disorders, aging, and neoplasm among other afflictions.
During the past decade, several studies have demonstrated that Pb exposure causes oxidative
stress, particularly in the kidney and cardiovascular tissues, as well as in cultured endothelial and
vascular smooth muscle cells (VSMC). The in vivo studies have further shown that Pb-induced
oxidative stress is, at least in part, responsible for the associated HTN in experimental animals.
Relevant published studies pertaining to this issue are summarized below and listed in Annex
Table AX5-5.1.
Khalil-Manesh et al. (1994) were among the first to suggest that oxidative stress may be
involved in the pathogenesis of Pb-induced HTN. This assumption was based on the observation
that chelation therapy with dimethyl succinic acid (DMSA) rapidly ameliorated HTN and raised
plasma cGMP level in rats with Pb-induced HTN. They further demonstrated that DMSA
possesses strong antioxidant properties in vitro. Accordingly, they theorized (a) that Pb exposure
may increase the generation of ROS, which, in turn, elevate arterial pressure by reacting with and
inactivating endothelium-derived-relaxing factor (EDRF), and (b) that by scavenging ROS,
DMSA rapidly lowers blood pressure prior to significantly affecting body Pb burden.
In a subsequent study, Gonick et al. (1997) showed a marked increase in renal tissue
content of lipid peroxidation product malondialdehyde (MDA) coupled with significant
upregulations of endothelial (eNOS) and inducible (iNOS) nitric oxide synthases. Thus, the
study provided evidence for the occurrence of oxidative stress and compensatory upregulation of
NOS isotypes in the kidney of animals with Pb-induced HTN.
In another study, Ding et al. (1998) showed that infusion of NOS substrate, L-Arginine,
lowers blood pressure to a much greater extent in rats with Pb-induced HTN than that seen in
either control animals or DMSA-treated Pb-exposed animals. The data, therefore, provided
indirect evidence for the role of depressed NO availability in the pathogenesis of Pb-induced
HTN. The study further suggested that oxidative stress may be responsible for diminished NO
availability in this model. It should be noted that administrating cell-impermeable native SOD
did not lead to a further reduction of blood pressure beyond that seen with L-Arginine alone.
As with the previous study (Khalil-Manesh 1994), oral DMSA therapy for 2 weeks significantly
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lowered blood pressure in the Pb-exposed animals. This was accompanied by a significant
reduction of blood Pb concentration. In an attempt to explore whether the observed amelioration
of Pb-induced HTN was due to the reduction of Pb burden or alleviation of oxidative stress by
DMSA, Vaziri et al. (1997) carried out a study in which rats with Pb-induced HTN were treated
with a lazaroid compound, a potent, non-chelating antioxidant. The study revealed marked
elevation of blood pressure and oxidative stress (increased lipid peroxidation) and reduced NO
availability (depressed urinary NC>2 + NOs excretion) in the untreated rats with Pb-induced HTN.
Antioxidant therapy with the lazaroid compound resulted in a significant alleviation of oxidative
stress, improved NO availability, and a marked attenuation of HTN without affecting blood Pb
concentration. Thus, the latter study provided convincing evidence for the role of oxidative
stress as a major mediator of Pb-induced HTN. The study further demonstrated that Pb-induced
HTN is associated with diminished NO availability and that the latter was mediated by oxidative
stress. The reduction in NO availability observed in rats with Pb-induced HTN (Pb acetate,
100 ppm in drinking water for 12 weeks) was recently confirmed by Dursun et al. (2005) in rats
treated with daily IP injection of Pb acetate (8 mg/Kg) for 2 weeks. The authors showed that the
increase in arterial pressure was accompanied by a significant reduction of urinary NO2 + NOs
excretion and a significant fall in renal blood flow (indicating increased renal vascular
resistance), mimicking the effect of the NOS inhibitor LNAME.
To further explore the cause for the observed reduction of NO availability, Vaziri et al.
(1999a) subsequently studied the expression of eNOS and iNOS in the kidney and cardiovascular
tissues of rats with Pb-induced HTN. The study showed that the reduction in NO availability is
paradoxically associated with a significant upregulation of NOS isotypes. Moreover, in vitro
incubation experiments revealed no significant change in NOS activity in the presence of lead.
Interestingly, antioxidant therapy with pharmacological doses of vitamin E and ascorbic acid
reversed the upregulation of NOS isotypes and paradoxically raised NO availability in the
subgroup of rats with Pb-induced HTN (Vaziri et al., 1999a). These observations were
subsequently confirmed by Vaziri and Ding (2001) who showed marked reduction of NO
availability despite significant upregulations of eNOS, nNOS, and iNOS in the aorta, heart,
kidney, and brain of rats with Pb-induced HTN and their normalization with the administration
of superoxide-scavenger tempol (15 mg/Kg IP/day) for 2 weeks. It is noteworthy that tempol
administration had no effect on the measured parameters in the control animals. Taken together,
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these observations indicated that ROS-mediated NO inactivation and, hence, depressed NO
availability, results in a compensatory upregulation of NOS isotypes in animals with Pb-induced
HTN. This phenomenon is consistent with other studies from this group, which have
demonstrated the presence of a negative-feedback regulation of eNOS by NO (Vaziri and Wang,
1999; Vazirietal., 2005).
The occurrence of compensatory upregulation of NOS by oxi dative stress in Pb-exposed
intact animals described above was subsequently replicated by Vaziri and Ding (2001) in
cultured human endothelial cells incubated in media containing different concentrations of Pb
acetate (versus control media containing sodium acetate). Once again, co-incubation with
tempol prevented this phenomenon. This study confirmed the ability of Pb to affect endothelium
independently of its effects on humoral or hemodynamic factors, which are operative in vivo.
Taken together, these observations suggest that Pb-induced reduction of biologically active NO
is not due to the reduction of NO-production capacity. Instead, it is linked to oxidative stress.
In an attempt to explore this supposition, in a separate study, Vaziri et al. (1999b), tested the
hypothesis that avid inactivation and sequestration of NO by ROS may be, in part, responsible
for the reduction of NO availability in animals with Pb-induced HTN. To this end, they tested
for the presence of immunodetectable nitrotyrosine in kidney, brain, and cardiovascular tissues
harvested from untreated and antioxidant-treated (vitamin E + vitamin C) rats with Pb-induced
HTN and normal control rats. Nitrotyrosine was used as a marker of NO oxidation by ROS
(NO + O2 —»• ONOO , ONOO + tyrosine —»• nitrotyrosine). The study showed an
overabundance of nitrotyrosine in all plasma and tested tissues in the untreated rats with
Pb-induced HTN. Antioxidant therapy reduced nitrotyrosine abundance, attenuated HTN, and
simultaneously raised NO availability in the subgroup of rats with Pb-induced HTN but had no
effect on the normal control group. These observations provided compelling evidence that
Pb-induced HTN causes oxidative stress, which, in turn, promotes functional NO deficiency via
ROS-mediated NO inactivation. The latter, in turn, participates in the development and
maintenance of HTN and cardiovascular abnormalities. In addition, the formation of the highly
cytotoxic reactive nitrogen species, peroxynitrite (ONOO ), from the NO-ROS interaction and
the associated nitrosative stress could potentially contribute to the long-term cardiovascular,
renal, and neurological consequences of Pb exposure.
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In subsequent studies, Vaziri et al. (2003) explored the expression of NAD(P)H oxidase
(which is a well-recognized source of ROS in, not only, the immune cells but also in renal,
cardiovascular, and neuronal tissues) in animals with Pb-induced HTN. In addition, expression
of the main antioxidant enzymes, namely Mn and CuZn-superoxide dismutases (SOD), catalase
and glutathione peroxidase were investigated. The study revealed significant upregulation the
gp91phox subunit of NAD(P)H oxidase in the brain as well as a trend for higher levels in the renal
cortex and left ventricle of rats with Pb-induced HTN. This was accompanied by a significant
compensatory upregulation of CuZn SOD in the kidney and brain, and of Mn SOD in the heart,
of rats with Pb-induced HTN. In contrast, despite the presence of oxidative stress, catalase and
glutathione peroxidase activity levels were unchanged. In a more recent study, Farmand et al.
(2005) showed a significant increase in CuZn SOD activity with no change in either catalase or
glutathione peroxidase activity in the aorta of rats with Pb-induced HTN compared with control
animals. Since the latter enzymes are responsible for the reduction of H2O2 and lipoperoxides,
the lack of an appropriate rise in their tissue levels may contribute to the severity of oxidative
stress in Pb-exposed animals.
The contribution of oxidative stress in the pathogenesis of HTN in this model was
confirmed by experiments that demonstrated normalization of arterial pressure with the infusion
of superoxide-scavenger, tempol, in rats with Pb-induced HTN (but no change was observed in
the blood pressure in the control rats) (Vaziri et al., 2003). As noted above, the relative
reduction of tissue catalase and glutathione peroxidase, which are responsible for the reduction
of H2O2 to water and molecular oxygen (2H9O9 —CAT )2H9O + O9), can result in accumulation of
GPX
H2O2. H2O2 serves as a cellular growth signal, as well as a substrate for hydroxyl radical ('OH)
generation. The former action can potentially contribute to cardiovascular remodeling, whereas
the latter can promote oxidative injury. In a recent study, Ni et al. (2004) demonstrated a
transient rise in O2 production followed by a sustained rise in H2O2 production by human
coronary endothelial and vascular smooth muscle cells cultured in media containing Pb acetate
versus the control media containing Na-acetate. This was accompanied by, and primarily due to,
upregulation of NAD(P)H oxidase and SOD together with reduced or unchanged catalase and
glutathione peroxidase levels. Accordingly, the results of this in vitro study confirmed the
findings of the in vivo studies and validated the anticipated accumulation of H2O2.
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As noted above, H2O2 is the substrate for the Fenton and Haber-Weiss reactions, which
culminate in formation of the highly cytotoxic-OH (H2O2 + e~ —>• OH + OH ). Thus, the
accumulation of H2O2 in animals with Pb-induced HTN can facilitate-OH production and,
thereby, promote oxidative stress and tissue injury. This supposition was confirmed in a series
of studies by Ding et al. (2001), who showed increased hydroxyl radical production in rats with
Pb-induced HTN. Oxidative stress, HTN, and excess hydroxyl radical production were all
reversed with IV infusion of the reputed hydroxyl radical scavenger, DMTU, in the Pb-exposed
animals. Increased hydroxyl radical production observed in intact animals with Pb-induced HTN
was confirmed in Pb-treated cultured endothelial cells (Ding et al., 2000). The role of oxidative
stress in the pathogenesis of HTN and endothelial dysfunction (depressed NO availability) has
been substantiated by a number of other investigators. For instance, Attri et al. (2003) showed
that exposure to Pb for up to 3 months resulted in a significant rise in arterial pressure, which
was substantially ameliorated by coadministration of the antioxidant vitamin ascorbic acid (20
mg/rat) in Wistar-Kyoto rats. The rise in arterial pressure in Pb-treated rats was accompanied by
diminished NO availability (low plasma NO2 + NOs) and biochemical evidence of oxidative
stress, i.e., elevations of plasma MDA, a DNA oxidation product (8-hydroxyguanosine), and
diminished ferric-reducing antioxidant power, as well as electrophoretic evidence of DNA
damage. Amelioration of HTN by antioxidant therapy was accompanied by improved NO
availability (plasma NO2 + NO3), marked attenuation of oxidative stress, and partial reduction of
DNA damage in this model. In another study, Malvezzi et al. (2001) showed partial amelioration
of HTN in Pb-exposed rats with the administration of either DMSA or L-arginine and showed a
much greater response with the combination thereof. These observations support the role of
interaction of ROS and NO in the pathogenesis of Pb-induced HTN in the rat.
As cited above, Pb-induced HTN is associated with and is, at least in part, due to
ROS-mediated inactivation and hence, reduced availability of biologically active NO. Many of
the biological actions of NO are mediated by cGMP, which is produced from the substrate GTP
by the cytosolic enzyme soluble guanylate cyclase (sGC). sGC is expressed in VSMC and
several other cell types. The enzyme is activated by NO to produce cGMP, which, in turn,
promotes vasorelaxation by lowering cytosolic Ca2+ concentrations. In an earlier study,
Khalil-Manesh et al. (1993a) demonstrated a significant reduction of plasma and urinary
cGMP in rats with Pb-induced HTN. These observations prompted a number of studies to
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evaluate the effect of Pb on sGC expression and cGMP production in vascular tissues obtained
from rats with Pb-induced HTN or in normal vascular tissues incubated in Pb-containing media.
For instance, Marques et al. (2001) found significant reductions of acetylcholine- and
Na-nitroprusside-induced vasorelaxation, despite upregulation of eNOS, in the aorta of rats with
Pb-induced HTN. This was associated with marked downregulation of sGC abundance and
diminished cGMP production in the aorta. In an attempt to explore the possible role of oxidative
stress in Pb-induced downregulation of sGC, they included a group of rats that were co-treated
with Pb and the antioxidant vitamin ascorbic acid. Antioxidant therapy ameliorated HTN,
restored vasorelaxation response to acetylcholine and Na-nitroprusside, and normalized sGC
expression and cGMP production. The authors, therefore, identified diminished sGC as another
mechanism by which Pb exposure can promote endothelial dysfunction and HTN. They further
showed that Pb-induced downregulation of sGC is mediated by oxidative stress, as evidenced by
its prevention with antioxidant therapy. Downregulation of sGC protein abundance in the aorta
of Wistar rats with Pb-induced HTN was recently confirmed by Farmand et al. (2005) in the
Pb-exposed Sprague-Dawley rats. In another study, Courtois et al. (2003) showed that 24-h
incubation of normal rat aorta in the Pb-containing media resulted in a concentration-dependent
downregulation of sGC (beta subunit), with the maximum effect observed at 1 ppm
concentration. This was associated with increased C>2 production and upregulation of
cyclooxygenase-2 (COX-2) expression. Co-incubation with ascorbic acid reduced COX-2
expression and C>2 production and attenuated, but did not fully prevent, the Pb-induced
downregulation of sGC. Similarly, addition of COX-2 inhibitor Rofecoxib or of protein kinase
A inhibitor (H-89) partially mitigated the Pb-induced downregulation of sGC in vitro. However,
the COX-2 inhibitor failed to reduce Q^ production in Pb-exposed vascular tissues. Based on
these observations, the authors concluded that Pb exposure downregulates vascular tissue sGC
abundance via induction of oxidative stress and upregulation of COX-2.
Oxidative stress and altered NO metabolism can potentially trigger a cascade of events
that work in concert to promote HTN and cardiovascular disease in Pb-exposed organisms.
Some of these potential links are illustrated in Figure 5-8.

5.5.2.2 Protein Kinase C, Inflammation, NFidJ Activation, and Apoptosis
Protein kinase C (PKC) isoforms belong to a family of serine-threonine kinases, which
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Chronic Lead Exposure
Oxidative Stress
NO inactivation
I NO production
(eNOS uncoupling,
BH4 depletion)
.[ NO bioavailability
Inflammation
Oxidation of
arachidonic acid
Vasoconsfriction,
Renal Na retention
+
t Sympathetic activity
Isoprostanes
Vasoconstriction,
Renal Na retention
RAS
Activation
Endothelial dysfunction
Cardiovascular
remodeling,
Platelet
activation
Hypertension
Arteriosclerosis,
Thrombosis
Figure 5-8. This illustration depicts some of the potential mechanisms by which
oxidative stress may participate in the pathogenesis of lead-induced HTN
and cardiovascular complications. In the presence of oxidative stress,
uncontained reactive oxygen species (ROS) inactivate nitric oxide (NO),
deplete NO synthase cofactor (tetrahydrobiopterin), uncouple eNOS, promote
generation of isoprostanes by oxidizing arachidonic acid, and activate the
redox-sensitive transcription factor NFidJ. Together, these events can cause
vasoconstriction, salt retention, sympathetic system activation, renin-
angiotensin system stimulation, platelet adhesion, and, thereby, endothelial
dysfunction, hypertension (HTN), inflammation, arteriosclerosis, and
thrombosis.
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serve numerous diverse cellular functions. For instance, PKC is involved in regulating vascular
contractility, blood flow, permeability, and cell growth. In this regard, the activation of PKC has
been shown to cause vascular contraction and Pb exposure has been found to raise PKC activity.
For example, Hwang et al. (2002) found increased PKC activity in the erythrocytes of a
group of Pb-exposed Korean workers, and Markovac and Goldstein (1988b) showed a significant
increase in PKC activity in rat brain micro vessels following exposure to micromolar Pb
concentrations. Also, Watts et al. (1995) demonstrated that Pb acetate (10~10 to 10~3 M) caused
contraction in an isolated rabbit mesenteric artery preparation. This Pb-induced vasoconstriction
was unaffected by denudation of endothelium, while it was significantly potentiated by PKC
agonists and attenuated by a PKC inhibitor. Calcium channel blockade with verapamil
attenuated, but did not abolish, Pb-induced vasoconstriction. These findings were considered to
indicate that activation of PKC is, in part, responsible for Pb-induced vasoconstriction,
independently of endothelium or extracellular influx of calcium. Taken together, these
observations suggest that the activation of PKC in the vascular smooth muscle cells may, in part,
contribute to the pathogenesis of Pb-induced HTN by enhancing vascular contractility. It should
be noted, however, that Pb-induced contraction has been shown to be unaffected by a PKC
inhibitor in rat aorta rings (Valencia, 2001). Thus, the contribution of PKC activation to the
Pb-induced alteration of vascular contractility appears to be both vessel- and species-specific.
It is of note that at high concentrations, Pb can reduce PKC activity in certain cell types,
including mouse macrophages and rat brain cortex (reviewed by Watts et al. [1995]).
As noted earlier, Pb exposure results in oxidative stress in cultured VSMC and endothelial
cells, as well as in intact animals. Oxidative stress can promote the activation of the nuclear
transcription factor kappa B (NFicB) and, thereby, trigger inflammation and apoptosis. In this
context, Ramesh et al. (2001) showed that exposure to low Pb levels (50 ppm in drinking water)
for 90 days activates NFicB and capsases in the rat brain. It is of note that several studies have
revealed the presence of renal tubulointerstitial infiltration of activated T cells, macrophages, and
angiotensin II (Ang-II) producing cells in various forms of genetic and acquired HTN in
experimental animals. Moreover, the associated tubulointerstitial inflammation has been shown
to contribute to the pathogenesis of HTN in these disorders (Rodriguez-Iturbe, 2004). These
abnormalities are accompanied by activation of the redox-sensitive NFicB, which can account for
the associated inflammation (reviewed by Rodriguez-Iturbe et al. [2004]). The NFicB activation,
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the accompanying inflammation, and HTN are ameliorated by antioxidant therapy in these
models, pointing to the role of oxidative stress in this process. In a recent study, Rodriguez-
Iturbe, et al. (2005) observed marked activation of NFicB coupled with tubulointerstitial
accumulation of activated T-cells, macrophages, and Ang-II-producing cells, as well as increased
apoptotic cells in the kidneys of Pb-exposed rats (100 ppm Pb acetate in water for 3 months).
This was associated with increased nitrotyrosine staining (a marker of NO/ROS interaction) in
the kidney tissue. Since tubulointerstitial inflammation plays a crucial role in the pathogenesis
of HTN in various other models of HTN, its presence in the Pb-exposed animals may contribute
to the associated HTN. Inflammation in Pb-induced HTN is not limited to the kidney. In fact,
lymphocyte infiltration is reported in the periaortic tissues in rats with Pb-induced HTN
(Carmignani et al., 2000). The inflammatory response to Pb exposure in the renal and vascular
tissues outlined above parallels observations reported for the immune system in Section 5.9 of
this chapter.

5.5.2.3 Effect of Lead Exposure on the Adrenergic System
The adrenergic system plays an important role in regulating arterial pressure, renal and
systemic hemodynamics, and cardiac function in health and disease. For this reason, a number
of clinical and animal studies have focused on the sympathetic system as a possible mediator of
Pb-induced HTN and cardiovascular abnormalities. For instance, in a study of a group of
Pb-exposed workers, Chang et al. (1996) found elevated plasma norepinephrine (NE), but
normal plasma dopamine and epinephrine, levels. The constellation of these biochemical
abnormalities points to increased sympathetic nervous system activity in Pb-exposed humans.
The impact of Pb exposure on the sympathetic nervous system activity has been substantiated in
experimental animals. For example, Chang et al. (1997) showed that administration of Pb (Pb
acetate 0.5% in drinking H2O) for 2 months resulted in significant rises in arterial pressure and
plasma NE (but not epinephrine) in Wistar rats. This was coupled with significant reductions of
the aorta P adrenergic receptor density and isoproterenol (P agonist)-stimulated cAMP
production. In a subsequent study, Tsao et al. (2000) reported a significant rise in plasma NE
coupled with marked reductions of P receptor density as well as diminished basal and
isoproterenol-stimulated cAMP productions in the aorta and heart of Wistar rats with Pb-induced
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HTN. In contrast to the heart and aorta, P receptor density as well as basal and P agonist-
stimulated cAMP production were increased in the kidneys of Pb-exposed animals.
In another study, Carmignani et al. (2000) found significant elevations of blood pressure,
plasma catecholamines, and cardiac contractility (dP/dt), together with reduced carotid blood
flow in rats with Pb-induced HTN. The effect of Pb on the sympathetic nervous system activity
was examined by Lai et al. (2002) who tested the rapid response to intrathecal (IT) injection of
PbCb in vivo and its addition to the thoracic cord slices in vitro in the rats. They found
significant rises in arterial pressure and heart rate with IT injection of Pb-chloride. These effects
of Pb were abrogated by the administration of ganglionic blockade using hexomethonium.
The in vitro studies revealed a significant rise in excitatory and significant fall in inhibitory
post-synaptic potentials with the addition of Pb to the bathing medium and their reversal with
saline washout.
In a recent study, Chang et al. (2005) showed a gradual decline in blood, kidney, heart,
and aorta Pb contents toward the control values within 7 months following cessation of exposure
in rats with Pb-induced HTN. This was coupled with a parallel declines in arterial pressure,
plasma NE and renal tissue P receptor density as well as parallel rises in the aorta and heart
P receptors densities during the 7-month period following cessation of Pb exposure. However,
while HTN and P receptor abnormalities were significantly improved, they were not completely
reversed. It should be noted that bone Pb contents were not measured in this study and were
most likely elevated despite normalization of blood and soft tissue levels. These findings
provided evidence for the stimulatory effect of Pb on the sympathetic nervous system and for its
contribution to the cardiovascular effects of Pb exposure.

5.5.2.4 Effects of Lead on the Renin-Angiotensin-Aldosterone (RAAS) and
Kininergic Systems
The available data on the effects of Pb exposure on the RAAS are contradictory. This
appears to be primarily due to variability in the dosage and duration of Pb exposure, as well as
the age at which exposure is initiated or the animals studied. In addition, when present,
nephropathy can potentially affect the RAAS profile of Pb-exposed animals or humans. The
majority of animal studies of the effects of Pb on RAAS were conducted and published in the
late 1970s and 1980s. In a meta-analysis of the studies published in that period, Vander (1988)
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found increased plasma renin activity and renal tissue renin content in young rats after several
weeks of Pb exposure sufficient to achieve blood Pb levels in the range of 30 to 40 |ig/dL.
Similar results were found in rats exposed to Pb in utero and for 1 month after birth. In contrast,
plasma renin activity and renal renin contents were generally unchanged or even reduced in older
rats whose Pb exposure had commenced in utero.
In a more recent study, Carmignani et al. (1999) showed a significant increase in plasma
angiotensin converting enzyme (ACE) activity in the rats exposed to Pb (60 ppm Pb acetate in
water) for 10 months beginning at an early age (weaning). This was accompanied by a
significant increase in plasma kininase II, kininase I, and kallikrein activities. In a subsequent
study, Sharifi et al. (2004) examined plasma and tissue ACE activity in young adult rats
(weighing 200 g) exposed to Pb (100 ppm Pb acetate) for 2 to 8 weeks. They found significant
rises in plasma, aorta, heart, and kidney ACE activities, peaking at 2 to 4 weeks. This was
followed by a decline in plasma and tissue ACE activity to subnormal values by 8 weeks, at
which point arterial pressure was markedly elevated. The authors concluded that the elevated
ACE activity is involved in the induction of HTN but may not be necessary for maintaining
HTN in Pb-exposed animals. Finally, in a recent study, Rodriguez-Iturbe et al. (2005)
demonstrated a marked increase in the number of Ang-II positive cells in the kidneys of rats
treated with Pb acetate (100 ppm in water) for 3 months. This observation points to heightened
intra-renal Ang-II generation in rats with Pb-induced HTN.
Taken together, the data point to activation of the RAAS at some point in the course of
Pb-induced HTN. Further studies are needed to fully elucidate the effects of Pb exposure on
various other RAAS components.

5.5.3 Effects of Lead Exposure on Vasomodulators
In a study of a group of Pb workers with elevated blood Pb concentration, Cardenas et al.
(1993) found a significant increase in urinary excretion of the metabolite of vasoconstrictive
prostaglandin, thromboxan (TXB2), and significant reduction of the vasodilatory prostaglandin,
6-keto-PGFl, when compared with the control workers. Subsequently, Hotter et al. (1995)
confirmed the elevation of urinary TXB2 in another group of Pb-exposed workers. Based on
these observations, the authors suggested that Pb can alter the balance between vasoconstrictive
and vasodilatory prostaglandins in a way that may contribute to HTN and cardiovascular disease.
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In an attempt to examine such possible effects of Pb exposure in experimental animals, Gonick
et al. (1998) measured urinary excretion of the above metabolites in the rat model of Pb-induced
HTN. The study showed no significant difference in urinary excretion of the given prostaglandin
metabolites between the Pb-exposed and control rats. However, in a recent in vitro study,
Dorman and Freeman (2002) demonstrated that Pb promotes the release of arachidonic acid by
vascular smooth cells via activation of phospholipase A2. They further showed that, at low
concentrations, Pb augments Ang-II-induced VSMC proliferation, whereas at a high
concentration it reduces viability and cell count in unstimulated cells and reduces DNA
synthases in Ang-II and fetal calf serum (FCS)-stimulated VSMC. Thus, Pb can increase the
release of arachidonic acid (the substrate for prostaglandins) via activation of phospholipase A2.
Given the limited and contradictory nature of the published data, further in-depth studies
are needed to clarify the effects of Pb on regulation of arachidonic acid metabolism and the
synthesis of various classes of prostaglandins.

Endothelin
Endothelins (ET) represent a family of potent vasoconstrictive peptides that are produced
by endothelium and a number of other cell types. Excess production or increased sensitivity to
ET can raise arterial pressure. In an attempt to explore the possible contribution of ET to the
pathogenesis of Pb-induced HTN, Khalil-Manesh et al. (1993a) studied the effects of exposure to
low and high levels of Pb (100 ppm versus 5000 ppm) in the drinking water for 1 to 12 months in
rats. Rats exposed to low (but not high) levels of Pb exhibited HTN and a significant increase in
plasma ET-3 concentration. These findings were confirmed by these investigators in a
subsequent study of rats with Pb-induced HTN (Khalil-Manesh et al., 1994). Similarly, Gonick
et al. (1997) demonstrated a significant elevation of plasma concentration and urinary excretion
of ET-3 in rats with Pb-induced HTN. Courtois et al. (2003) showed that incubation in the
Pb-containing media resulted in the downregulation of soluble guanylate cyclase and cGMP
production in the isolated artery segment of normal rats. They further found that co-incubation
with an ET-A receptor antagonist can partially reverse this effect of Pb. These findings suggest
that the adverse effect of Pb exposure on cGMP production in the vascular tissue is, in part,
mediated by its ability to raise ET activity. It, thus, appears that exposure to low-levels of Pb can
raise activity or production of ET, which can, in turn, play a part in the pathogenesis of Pb-
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induced HTN in the rat. Further studies are required to carefully explore the effects of Pb on
various components of the ET system.

Atrial Natriuretic Factor
Atrial natriuretic factor (ANF) is produced and secreted by cardiac myocytes. Plasma
concentration of ANF rises with volume expansion and declines with volume contraction. ANF
serves as a vasodilator and a natriuretic agent and, as such, plays a role in regulating blood
volume, vascular resistance, and, hence, arterial pressure. Giridhar and Isom (1990) measured
ANF in rats treated with IP injection of Pb acetate (0.0 to 1.0 mg/kg/twice weekly for 30 days).
The Pb-exposed animals exhibited fluid retention, which was coupled with a paradoxical dose-
dependent decline in plasma ANF concentration. Based on these findings, they suggested that
Pb may interfere with the hormonal regulation of cardiovascular system, which may, in turn,
relate to the cardiovascular toxicity of this metal.

5.5.4 Effects of Lead on Vascular Reactivity
Addition of Pb acetate to the bathing medium has been shown to elicit a cumulative
concentration-dependent vasoconstriction in isolated rabbit mesenteric artery (Watts et al.,
1995). This effect was reported to be partly mediated by activation of PKC. In a more recent
study, Valencia et al. (2001) found a concentration-dependent vasoconstrictive response to Pb
acetate (0.1 to 3.1 mM) in Wistar rat thoracic aorta rings. The contractile response was observed
in both intact and endothelium-denuded rings. Likewise, Pb-induced vasoconstriction was
preserved in calcium-free medium and was unaffected by either a-1 blockade (prazosin), PKC
inhibition (Calphostin) or L-type calcium channel blockade (verapamil). However, Pb-induced
vasoconstriction was inhibited by lanthanum, which is a general calcium-channel blocker. These
observations suggest that Pb can promote an endothelium-independent vasoconstriction by a
direct effect on the vascular smooth muscle cells. The data further suggest that the effect of Pb is
Ca-independent and may depend on the entry of Pb to the cell via a lanthanum-blockable
channel. In contrast to the latter studies, addition of Pb acetate did not cause vasoconstriction in
the rat aorta rings used in a study reported by Shelkovnikov and Gonick (2001). Moreover, Pb
acetate at either high (10~4m) or low (10~8m) concentrations did not modify the response to NE,
phorbol ester, or isoproterenol. However, at 10~4M, Pb acetate augmented the contractile
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response to submaximal concentrations of calcium. Thus, the rapid action of Pb on vascular
reactivity in vitro seems to vary depending on the type of the vessel used, the Pb concentration
employed, and the animal species being studied.
A number of studies have endeavored to discern possible differences in vascular reactivity
to various agonists between animals with Pb-induced HTN and control animals. For instance,
Purdy et al. (1997) found no significant difference in vasoconstrictive response to NE and
phenylephrine or vasodilatory response to acetylcholine or nitroprusside in the aorta rings
obtained from Sprague-Dawley rats with Pb-induced HTN. In contrast, Marques et al. (2001)
showed a significant reduction of vasodilatory response to both acetylcholine and nitroprusside
in Wistar rats with Pb-induced HTN. It should be noted that the Wistar rats employed in the
latter study had been treated with 5 ppm Pb acetate in the drinking water for 1 month, whereas
those reported by Purdy et al. (1997) had been given a higher dosage (100 ppm) for a longer
period (3 months). Therefore, the magnitude and duration of exposure may account for the
differences observed between the two reports. Also, the effect of Pb on vascular reactivity may
vary from one tissue to the next, as clearly exemplified by studies (Oishi et al., 1996) that
showed significant endothelium-dependent vasorelaxation of mesenteric artery response to
acetylcholine in the presence of the NOS inhibitor L-NAME in tissues from rats exposed to Pb
acetate for 3 months. These observations suggest that chronic Pb exposure may impair
endothelium-dependent hyperpolarization in the rat mesenteric artery. However, no such effect
was noted in the aorta obtained from the same animals.

5.5.5 Lead-Calcium Interactions in Vascular Tissue
Changes in cytosolic Ca2+ concentrations are intimately involved in regulating vascular
tone and vascular smooth muscle contraction. Consequently, several studies have focused on the
interaction of Pb with cellular Ca2+ and Ca2+-dependent signaling pathways as a means to gain
insight into the pathogenesis of Pb-induced HTN (Piccini et al., 1977; Favalli et al., 1977; Webb
et al., 1981; Goldstein, 1993; Watts et al., 1995). Lead can potentially compete with Ca2+ in
transport systems (i.e., channels and pumps) involved in physiological movements of ions,
particularly Ca2+, into and out of the cell (Simons, 1993a,b). Moreover, Pb can alter the
intracellular distribution of Ca2+ between cytoplasm, endoplasmic reticulum, and mitochondria,
which normally regulates cytosolic Ca2+ concentration, (Simons 1993a,b). In addition, Pb can
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serve as a substitute for calcium in Ca2+-dependent signaling pathways by interacting with
calmodulin, PKC, and calcium-dependent potassium channels (Haberman, 1983; Richardt et al.,
1986; Chai and Webb, 1988; Simons, 1993a,b; Watts et al., 1995). Thus, interactions of Pb with
cellular Ca2+ via these complex mechanisms in the vascular cells may contribute to alterations of
vascular resistance and HTN. For example, Piccini et al. (1977) and Favalli et al. (1977) showed
that Pb exposure increases calcium content in the tail artery in rats. The authors attributed this
phenomenon to a possible Pb-induced inhibition of Ca2+ extrusion from the vascular cells. Using
rabbit mesenteric artery preparations, Watts et al. (1995), showed that blockade of either PKC or
voltage-gated Ca channels by verapamil substantially attenuated Pb-induced vasoconstriction in
both intact and endothelium-denuded preparations. Based on these observations, the authors
suggested that Pb promotes a vasoconstrictive response in rabbit mesenteric artery via a
Ca2+-dependent activation of PKC. In contrast, Valencia et al. (2001) using rat aorta rings
reported a vasoconstrictive response to Pb acetate in rat aorta rings bathed in either Ca2+-free or
Ca2+-containing media and in the presence or absence of the L-type calcium-channel blocker
verapamil or of the PKC inhibitor calphostin. Moreover, depletion of intracellular Ca2+ stores by
preincubation of rings in EOT A, while diminishing the intensity, did not abrogate Pb-induced
vasoconstriction in this system. In contrast, Pb-induced vasoconstriction was prevented by
lanthanum (a general blocker of calcium channels) in both Ca2+-containing and Ca2+-free media.
Based on these observations, the authors concluded that Pb can elicit a PKC-independent
contractile response in the rat aorta by entering VSMC via a non-voltage-gated Ca2+ channel and
mimicking the action of Ca2+. It, thus, appears that Pb exerts its effect by mechanisms that are
species- and vessel-specific.

5.5.6 Cardiotoxicity and Atherogenesis
Acute Pb exposure has been reported to affect cardiac function, and chronic exposure has
been linked to atherosclerosis and increased cardiovascular mortality by some, but not by all
investigators, in humans (See Chapter 6). In an attempt to assess the cardiotoxicity of Pb,
Prentice and Kopp (1985) carried out the in vitro perfusion of isolated rat heart preparations with
a perfusate containing 0.3 and 30 jiM Pb acetate for up to 60 min. At 30 jiM concentration, Pb
prolonged the AV node and His bundle conduction times, reduced coronary blood flow and heart
rate, and altered cardiac energy metabolism. Milder, and statistically insignificant, changes were
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also observed at 0.3 jiM Pb concentration in this model. These observations illustrate the direct
cardiotoxicity of Pb independently of its systemic and neuroendocrine actions in acute
intoxication. In an attempt to determine whether chronic exposure to Pb or cadmium can cause
atherosclerosis, Revis et al. (1981), studied male white pigeons that were exposed to Pb (0.8 ppm
in drinking water) for extended periods. Long-term low-level Pb exposure in this model resulted
in a significant rise in arterial pressure and a near doubling of the number of atheromatous
plaques in the aorta. These observations demonstrate the proatherogenic effects of chronic
exposure to low levels of Pb in pigeons.

5.5.7 Effects of Lead on Endothelial Cells
Endothelium is an important constituent of the blood vessel wall, which regulates
macromolecular permeability, vascular smooth muscle tone, tissue perfusion, and blood fluidity.
Endothelial damage or dysfunction results in atherosclerosis, thrombosis, and tissue injury.
Chronic Pb exposure has been shown to promote atherosclerosis in experimental animals (Revis
et al., 1981). Given the central role of endothelial injury/dysfunction in the pathogenesis of
atherosclerosis, numerous studies have explored the effect of Pb on cultured endothelial cells.
These studies have searched for evidence of Pb-mediated endothelial cell injury and the effects
of Pb on endothelial cell proliferation, tube formation (angiogenesis), monolayer wound repair,
and production of heparansulfate proteoglycans, plasminogen activator (IPA), and plasminogen
activator inhibitor-1 (PAI-1).
Using cultured bovine aorta endothelial cells, Kaji et al. (1995a) showed that incubation
with Pb nitrate at concentrations equal to or below 50 jiM for 24 h, results in mild
de-endothelialization of endothelial monolayers in vitro. They further showed that adding Pb at
10 jiM concentration markedly increased cadmium-induced endothelial injury.
Proliferation of endothelial cells is a critical step for the repair of injured endothelium.
Failure of the repair process can result in thrombosis, VSM cell migration and proliferation, and
atherosclerosis. In this regard, Pb (Pb nitrate 0.5 to 5 jiM) has been shown to significantly
reduce DNA synthesis and cell proliferation in growing cultured bovine aorta endothelial cells
(Kaji, 1995a). Similarly, the proliferative response to PFGF and oFGF is significantly attenuated
by Pb in this system (Kaji, 1995b). The reported inhibition of endothelial cell proliferation by Pb
can potentially diminish the repair process in response to endothelial injury. This supposition
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has been confirmed by Fujiwara et al. (1998) who showed that at 5 to 10 jiM concentrations, Pb
markedly inhibited the repair of the wounded endothelial monolayer in vitro. Moreover, Pb
severely mitigated the zinc-stimulated endothelial cell proliferation and repopulation of the
denuded sections in this system.
Endothelial cell proliferation is the primary step in angiogenesis, a phenomenon that is
essential for numerous physiological functions such as growth, development, wound repair, and
menstrual cycle as well as certain pathological events including diabetic retinopathy and tumor
growth. In view of the demonstrated inhibition of endothelial cell growth by Pb, it has been
postulated that Pb may impair angiogenesis. This assumption has been confirmed by a number
of studies testing the effect of Pb by angiogenesis assay (tube formation) in endothelial cells
cultured on matrigel (a laminin-rich basement membrane product) matrix in vitro. For instance,
Ueda et al. (1997) and Kishimoto et al. (1995) have shown that Pb acetate (1 to 100 |iM) results
in a concentration- and time-dependent inhibition of tube formation by human umbilical vein
endothelial cells cultured on a matrigel matrix.
Endothelial cell migration and proliferation are critical for angiogenesis and repair of the
damaged endothelium. PFGF is a powerful mitogen for endothelial cells as well as several other
cell types. Endothelial cells synthesize PFGF, which is released following injury or spontaneous
death of endothelial cells and acts in an autocrine fashion to facilitate the repair process by
promoting endothelial cell migration and proliferation. Binding of PFGF to its receptor on the
endothelial cell is facilitated by heparan sulfate proteoglycans (HSPGs) that are normally
produced and released by the endothelial cells for attachment to the cell surface as well as
incorporation in the extracellular matrix. As noted above, Pb significantly attenuates PFGF and
oFGF-mediated DNA synthesis and proliferation in cultured endothelial cells (Kaji et al.,
1995b). In this regard, Pb has been shown to reduce PFGF binding to the cell surface HSPGs
without changing the biosynthesis or intracellular abundance of PFGF in cultured bovine
endothelial cells (Fujiwara and Kaji, 1999). Moreover, Pb has been shown to significantly
reduce the synthesis of glycosamino-glycans (GAG, measured by sulfate incorporation into
heparan sulfate) in the growing endothelial cells.
The above observations suggest that Pb-induced reduction of pFGF-mediated proliferative
response in cultured endothelial cells is largely due to impaired production of HSPGs. This
supposition is further supported by observations that DNA synthesis can be restored by adding
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heparin in Pb-treated growing endothelial cells (Fujiwara et al., 1995). The reduction in the
production of GAGs by Pb in the growing endothelial cells (Fujiwara et al., 1995) is also seen in
confluent (quiescent) cells. For instance, Kaji et al. (1991) demonstrated a marked reduction of
GAG production following incubation with 10 jiM Pb nitrate in confluent endothelial cells in
vitro. The Pb-induced reduction of heparan sulfate production was more severe than that of the
other GAGs. Moreover, the reduction in the cell surface-associated GAGs was more severe than
that of the newly synthesized GAG found in the incubation media. GAGs combine with a series
of specific core proteins to form anionic macromolecular complexes known as proteoglycans,
which are widely distributed in the extracellular matrix of the mammalian tissues. Endothelial
cells produce two types of HSPGs, i.e., the high-molecular weight and low-molecular weight
classes. Perlecan is a high-molecular weight heparan-sulfate proteoglycan that is a component of
the basement membrane. Syndecan, glypican, ryudocan, and fibroglygan are among the low-
molecular weight subclass and are primarily associated with the cell surface. Proteoglycans play
an important role in regulating vascular function and structure. For instance, by providing a
negative electrostatic charge, these molecules constitute a major barrier against extravasations of
negatively charged plasma proteins. In addition, by interacting with antihthrombin-III and tPA,
these molecules serve as important endogenous anticoagulants. Moreover, perlecans facilitate
PFGF binding to its receptor on endothelial cells and, thus, contributes to the endothelial growth
and repair processes. In contrast, these molecules tend to inhibit migration and growth of
vascular smooth muscle cells and, thereby, help to prevent athero- and arteriosclerosis. Another
important function of HSPGs is their role in stabilizing and anchoring lipoprotein lipase and
VLDL receptors on the endothelial surface. Consequently, they play an important indirect part
in the clearance of VLDL and chylomicrons from the circulation, a process that has major
implications for energy metabolism and cardiovascular protection.
In a study of cultured bovine endothelial cells, Kaji et al. (1997) found that Pb chloride, at
10 jiM concentration, markedly lowers incorporation of precursors (glycosamine and sulfate)
into HSPG in confluent bovine aorta endothelial cells. The effect of Pb was more severe on
low-molecular than high-molecular weight HSPGs. However, Pb did not change the length of
heparan sulfate chains. It is of note that Pb slightly increased the abundance of the HSPG core
proteins. This observation excluded a reduction in core protein synthesis as a cause of
diminished HSPGs in the Pb-treated confluent endothelial cells. In a subsequent study, Fujiwara
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and Kaji (1999) investigated the effect of Pb nitrate on production of high- and low-molecular
weight subclasses of HSPGs in growing bovine aorta endothelial cells. In contrast to the
quiescent cells, Pb-treated growing cells exhibited a marked reduction in the high-molecular
weight with no change in production of low molecular weight (-50KD) HSPGs. They further
showed a significant reduction of the core protein of perlecan, which is a high-molecular weight
(400 KD) HSPG. Thus, Pb appears to affect productions of subclasses of HSPGs differently
depending on the cells' growth cycle. Accordingly, in the growing endothelial cells (a condition
that simulates the response to injury), Pb downregulates perlecan, which is involved in
pFGF-mediated migration and proliferation of endothelial cells and inhibition of migration and
proliferation of VSMC. This phenomenon may adversely affect endothelial repair and promote
athero- and arteriosclerosis. On the other hand, Pb-induced reduction of the cell surface-
associated low-molecular weight HSPGs (which are predominantly involved with lipolytic,
anticoagulant, and other functions of confluent endothelial cells (simulating intact endothelium)
can contribute to hyperlipidemia and thromboembolism, among other disorders.
One of the major properties of normal endothelium is its ability to prevent coagulation.
Several factors contribute to the thromboresistance of the endothelial lining. These include the
surface coating of HSPG (which confers heparin-like properties), nitric oxide (which inhibits
platelet adhesion and activation), and tPA (which promotes thrombolysis), thrombomodulin, and
prostacycline. As noted earlier, Pb exposure reduces HSPG-production (Kaji et al., 1995b, 1997)
and diminishes nitric oxide availability via ROS-mediated NO inactivation (Vaziri et al., 1999b).
In addition, Kaji et al. (1992) showed that incubation of confluent human umbilical vein
endothelial cells with Pb nitrate, at 0.01 to 1.0 jiM concentrations, significantly reduced basal
and thrombin-stimulated tPA release. It thus, appears that Pb exposure may confer a
thrombophilic diathesis.

5.5.8 Effects of Lead on Vascular Smooth Muscle Cells
Lead has been demonstrated to stimulate proliferation of bovine aorta VSMCs in a
concentration-dependent manner (Fujiwara et al., 1995). Moreover, the combination of Pb and
PFGF results in an additive effect on VSMC proliferation. As with bovine aorta VSMCs,
cultured rat aorta VSMCs exhibit hyperplasia in response to a low concentration of (100 |ig/L) of
Pb-citrate (Carsia et al., 1995). The reported hyperplasia is accompanied by phenotypical
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transformation of cells from the spindle or ribbon shape to cobblestone shape, simulating the
neointimal cell morphology. This was accompanied by a significant reduction in Ang-II receptor
but no change in a, P, or ANP receptor densities. It is of note that, in contrast to the low
concentration, a high concentration (500 |iM/L) of Pb resulted in growth arrest in this system.
Thus, the effect of low concentration of Pb on VSMC proliferation is opposite of its action on the
endothelial cells.
Under normal conditions, intact endothelial lining shields the cells residing in the
subendothelial tissue, i.e., fibroblasts and VSMCs, from coming into contact with the circulating
blood. However, this barrier is lost when the endothelium is injured, an event which can lead to
platelet adhesion and fibrin thrombosis formation. Propagation of fibrin thrombus is limited by
activation of the fibrinolytic system, which, in turn, depends on the balance between tPA and
plasminogen activator inhibitor-1 (PAI-1). In addition to endothelial cells, VSMCs and
fibroblasts express tPA and PAI-1. Using cultured human aorta VSMCs and fetal lung
fibroblasts, Yamamoto et al. (1997) investigated the effect of Pb chloride on the release of tPA
and PAI-1 in vitro. The authors found that Pb causes a significant inhibition of tPA release and a
significant increase in PAI-1 release in cultured fibroblasts in a dose-dependent manner. The
Pb-treated VSMC exhibited a significant dose-dependent decline in tPA release and to a lesser
extent of PAI-1 release. Taken together, exposure to Pb appears to evoke a negative effect on
fibrinolytic process by the cellular constituents of the subendothelial tissue.

5.5.9 Summary
• In vivo and in vitro studies published during the past 15 years have considerably expanded
our knowledge of the effects of Pb exposure on the cardiovascular system. However,
many questions remain unanswered and await further investigation.
• A number of in vivo and in vitro studies conducted during the review period have provided
compelling evidence for the role of oxidative stress in the pathogenesis of Pb-induced
HTN. Moreover, the effect of oxidative stress on blood pressure has been shown to be, in
part, mediated by avid inactivation of NO and downregulation of sGC. In addition, a
limited number of in vitro studies have provided indirect evidence that, via activations of
PKC and NFicB, Pb may raise vascular tone and promote inflammation.
• Based on several studies that evaluated the role of adrenergic system on Pb-toxicity,
chronic low-level Pb exposure appears to increase central sympathetic activity, reduce
cardiac and vascular and raise kidney P adrenergic receptor density. These events can, in
turn, increase peripheral vascular resistance and renal renin release/production and,
thereby, arterial pressure. Since sympathetic outflow is inhibited by NO, inactivation of
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NO by oxidative stress may be, in part, responsible for the increased sympathetic activity
in Pb-exposed animals.

• The renin-angiotensin-aldosterone system (RAAS) plays an important role in regulating
blood pressure and cardiovascular function and structure. The available new data suggest
that Pb exposure can raise plasma ACE and kininase activities at different points in the
course of Pb-induced HTN in experimental animals. This can, in turn, contribute to the
genesis and/or maintenance of HTN. Since renin release (which is responsible for
production of ACE substrate, i.e., Ang-1) is, in part, driven by P adrenergic activation,
upregulation of renal P adrenergic activity may, in part, account for increased RAAS
activity in the Pb-exposed animals.
• The balance in production of vasodilator and vasoconstrictor prostaglandins plays an
important role in regulation of blood pressure and cardiovascular function. Studies of the
Pb exposed humans have revealed an imbalance in production of prostaglandins favoring a
rise in arterial pressure. However, the animal and in vitro studies published during the
review period have been limited and inconsistent. Further studies are needed to address
this issue.

• Based on the available studies, Pb exposure appears to increase endothelin production in
experimental animals. This phenomenon can, in part, contribute to the rise in blood
pressure in the Pb-exposed animals. For instance, Pb has been shown to cause
vasoconstriction and to attenuate acetylcholine- and NO-mediated vasodilatation in some,
but not all vascular tissues and in some, but not all, studies. These effects have been
variably attributed to Pb-mediated activation of PKC and Ca2+-mimetic action of Pb,
among other possibilities.

• Finally, a number of studies have explored the effects of endothelial and vascular smooth
muscle cells to explore the possible atherogenic effect of Pb exposure. In this context, Pb
has been found to inhibit proliferation of the growing (non-confluent) endothelial cells
(mimicking in vivo response to injury), impair tube formation (angiogenesis), and the
repair of wounded endothelial monolayer in vitro. Likewise, Pb exposure was shown to
reduce production of HSPGs and tPA by confluent endothelial monolayers, events that
may favor thrombosis and hyperlipidemia. Lead exposure has been also shown to promote
vascular smooth muscle cell and fibroblast proliferation and phenotypic transformation in
ways that seem to favor arteriosclerosis and vascular remodeling.
• Among many questions awaiting clarification, a few are of particular interest. For
instance, it is not clear as to why low, but not high, levels of Pb exposure cause HTN in
experimental animals. Similarly, it is uncertain as to why HTN occurs long after the onset
of Pb exposure in the intact animals, whereas the effects on cultured cells and isolated
tissues are manifested within short periods of time.
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5.6 GENOTOXIC AND CARCINOGENIC EFFECTS OF LEAD
5.6.1 Introduction
The 1986 Lead AQCD (U.S. Environmental Protection Agency, 1986a) and its 1990
Supplement (U.S. Environmental Protection Agency, 1990) concluded that, at relatively high
concentrations, Pb may be carcinogenic to laboratory animals, particularly the rat. Cell culture
studies were considered to be supportive of these observations, but also indicated that Pb was not
particularly potent. Human data were considered to be of concern, but not definitive, and given
the animal data, the prudent choice was to consider Pb to be a possible human carcinogen.
This section reviews reports of Pb-induced carcinogenesis and DNA damage published
since 1986. More than 200 publications were read and considered and those that reported any
effect related to carcinogenesis or genotoxicity that was attributable to Pb are presented below.
This report follows the same format as the previous one (1986) and the explanations for
the relative importance of the various types of studies (e.g., epidemiology, animal and cell
culture) can be found in the original report and are not repeated here. Carcinogenesis studies are
presented first, followed by genotoxicity studies. Each of these sections is further subdivided
into human studies (considering adults and then children), animal studies, and then cell culture
studies (considering human, mammalian, and then nonmammalian). When appropriate, these
sections are followed by a section describing acellular (cell-free) model studies.
There are some differences with this new report. For one, each section is more distinctly
broken out. The epidemiology has been reviewed in more detail in Chapter 6 (Section 6.7) in
this document and, so, only a brief summary is presented here. Because of more recent concerns
about effects on childhood development, this issue was specifically considered in a separate
section. Following advances in hypotheses and technology, much more specific sections about
the possible epigenetic effects of Pb have also been added.

5.6.2 Carcinogenesis Studies
5.6.2.1 Human Studies
The human carcinogenesis studies are only briefly reviewed in this section; for a more
detailed review, see Chapter 6 (Section 6.7) in this document.
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Adults
The assessment of the carcinogen!city of Pb through human epidemiological studies
remains ambiguous. Several reports state that occupational exposure to Pb increases the risk of
lung, kidney, brain, stomach, and liver cancer (Fu and Boffetta, 1995; Kauppinen et al., 1992;
Gerhardsson et al., 1995; Ades and Kazantzis, 1988; Wicklund et al., 1988; Steenland et al.,
1992; Englyst et al., 2001; Gerhardsson et al., 1986; Anttila et al., 1995, 1996; Cocco et al.,
1998; Shukla et al., 1998). However, a full interpretation of the data in these studies is
complicated by the fact that the study participants also incurred coexposure to other known
carcinogens, such as arsenic, cadmium, and hexavalent chromium. Thus, it is difficult to
determine if the excess cancers observed were due to exposure to Pb, one of these other
carcinogens, or some combination of the various chemicals. In addition, other reports indicate
that occupational or environmental exposure to Pb did not alter cancer risk (Cocco et al., 1996;
Fanning, 1988; Jemal et al., 2002). Consequently, a definitive assessment of the carcinogenicity
of Pb from human studies cannot be made at this time.

Children
There have been no recent studies of Pb-induced cancers in children. This lack of data is
not unexpected and is largely because Pb has not been considered a likely cause of childhood
cancers. There have, however, been studies of cancers in children resulting from paternal
exposure. Here again, the same confounding problems encountered are as seen in the adult
population studies, and it is difficult to draw any definitive conclusions. For example, two
studies reported elevated childhood tumors (Wilm's tumor and acute nonlymphocytic leukemia)
in children whose fathers worked in Pb-related industries, such as welding, painting, and auto
repair (Buckley et al, 1989; Olshan et al., 1990). However, workers in these occupations also
experienced coexposure to arsenic, cadmium, and hexavalent chromium, and so the cancers
observed cannot be solely linked to Pb exposure. In addition, a report from the printing industry
in Norway found no link between paternal exposure and childhood cancers and, perhaps, even
found a possible reduction in the incidence of childhood cancers with paternal Pb exposure
(Kristensen and Andersen, 1992).
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The possible interaction of paternal occupation and childhood cancer is an important area
of concern. However, a definitive assessment of paternal exposure to Pb cannot be made at this
time and more research is needed.

5.6.2.2 Laboratory Animal Studies
Lead is a well-established animal carcinogen, as noted in the 1986 Lead AQCD.
Consequently, limited tumorigenesis studies have been conducted in animal models and the
focus has been more on the mechanism of neoplasia (e.g., the roles of calcium and
metallothionein) and possible immunomodulatory effects of Pb in the promotion of cancer.
These studies are summarized in Table AX5-6.1.
All of the studies exposed animals to Pb acetate except one, which focused on Pb
chromate. One study investigated the carcinogenicity of a series of chromate compounds, i.e.,
Pb chromate and several Pb chromate-based compounds were included as part of the group of
chromate compounds. The Pb chromate was administered by implantation into the lung after
being embedded within a cholesterol pellet. The authors indicated that in this design, Pb
chromate was not carcinogenic, but that 4 of the Pb chromate compounds did induce a very rare
tumor in the mice. Thus, there is some ambiguity about the carcinogenicity of Pb chromate in
the study, as the statistics calculated an expected tumor level based on any tumor and were not
based on the occurrence of this very rare (for rats) tumor. It is likely that had the expected value
been adjusted for the rare tumor, a conclusion would have been reached that either Pb chromate
was tumorigenic or that the study lacked the power to make any determination. The previous
EPA report had concluded that Pb chromate is tumorigenic. Thus, it is difficult to draw a firm
conclusion from this study.
The remaining five studies focused on Pb acetate (Schrauzer, 1987; Blakley, 1987; Teraki
and Uchiumi, 1990; Bogden et al., 1991; Waalkes et al., 2004). In most studies, this compound
was administered in drinking water at concentrations from 0.5 to 4000 ppm, but one study
considered effects from a subcutaneous (SC) injection both in mice and in rats. Consistent with
the findings in the 1986 Lead AQCD, Pb not only induced renal tumors, but also induced other
tumors, although the possible effect on mammary tumors is difficult to interpret, as important
study details were omitted, as discussed below. In a surprising development, during one lifetime
exposure study, Pb suppressed liver tumors (Waalkes et al., 2004).
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The key study in this group of studies was a lifetime exposure study that investigated
mice exposed to drinking water concentrations of 1,000 to 4,000 ppm Pb and also considered the
role of metallothionein. In wild-type mice, Pb acetate induced a low frequency of renal tumors,
but hyperplasia was common and exhibited overexpression of cyclin Dl. Lead inclusion bodies
were also common. Lead also suppressed liver tumors in this study.
By contrast, in metallothionein-deficient mice, Pb acetate induced a high frequency of
kidney tumors and severe inflammation. Both the tumors and the regions of inflammation
exhibited cyclin Dl overexpression. Lead also suppressed liver tumors in these animals.
In contrast to the wild-type mice, Pb inclusion bodies were not seen in these animals.
Another study focused on the ability of Pb to induce tumors in rats after SC injection of
Pb acetate (Teraki and Uchiumi, 1990). Tumors formed at the site of injection, and Pb
accumulated in the tumors, indicating that Pb is tumorigenic. However, full interpretation of the
data is complicated by the absence of data on control animals and the fact that only a single dose
was considered.
Three studies investigated compounds that might reduce or prevent Pb-induced cancers,
specifically selenium and calcium compounds (Schrauzer, 1987; Bogden et al., 1991). The first
study used a rather complex approach to study the possibly protective effects of selenium
(Shrauzer, 1987). In this study, mice were infected with the murine mammary tumor virus,
because they are known to develop mammary adenocarcinomas when maintained on a
low-selenium diet. The data indicated that Pb can induce tumors in these mice even when they
are maintained on a high-selenium diet. However, the data are difficult to interpret and the
impact of the study is uncertain, as the methods are incomplete, the data on control animals are
not provided, and the experimental results are stated but not presented in tables or figures.
The second study investigated the effect of calcium (Bogden et al., 1991). The main
focus of this study appeared to be blood pressure, but tumorigenesis was also considered.
It might be anticipated that calcium might reduce Pb tumorigenesis by competing for its binding
sites or blocking its uptake. However, in this study, calcium did not affect Pb levels in tissue and
actually exacerbated Pb-induced carcinogenesis. The full impact of this study is also difficult to
assess, as the calcium-treated animals incurred profound nephrocalcinosis.
The remaining study considered Pb-induced immunosuppression as a possible factor
contributing to the tumorigenesis induced by other agents, including viruses or chemicals
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(Blakley, 1987). The results indicated that Pb may suppress humoral immunity but not cellular
immunity. However, this is the only study of its kind and the results need to be repeated in other
settings. In addition, it is difficult to determine if these data are specific to the agents used (e.g.,
murine lymphocytic leukemia virus) or if they represent a class of agents (e.g., viruses in
general).
Overall, the above studies confirm that Pb is an animal carcinogen and extends our
understanding of mechanisms involved to include a role for metallothionein. Specifically, the
recent data show that metallothionein may participate in Pb inclusion bodies and, thus, serves to
prevent or reduce Pb-induced tumorigenesis. Much more work is needed to determine the
potential exacerbating or ameliorating roles of calcium and selenium and to determine what role
Pb-induced immunomodulation may play in the promotion of tumors.

5.6.2.3 Cell Culture Studies
Carcinogenesis is measured in cell culture systems through studies of neoplastic
transformation, where morphologically transformed cells are injected into athymic mice to see if
the cells can form a tumor in the host animal. Morphological transformation refers to cells that
incur a change in morphology, such as formation of a focus (or foci) of cell growth. In addition,
for faster study results and as a screening tool, the ability of cells to grow in agar without a
surface to attach to (anchorage independence) is often used as a short-term substitute measure for
transformation.

Human Cell Cultures
Since the 1986 Lead AQCD, only four studies have used human cell culture systems to
study the carcinogenesis of Pb compounds. One found that Pb acetate induced anchorage
independence in primary human foreskin fibroblasts (HFF) (Hwua and Yang, 1998). The full
impact of these data is uncertain, as previous studies of known metal carcinogens in primary
HFF found that these carcinogens induced anchorage independence, but those anchorage-
independent cells ultimately senesced. These studies are summarized in Table AX5-6.2. Further
study is needed to confirm that Pb can induce anchorage independence and to see if these cells
can progress to full neoplastic transformation.
In an effort to explore the importance of oxidative metabolism in inducing anchorage
independence, Hwua and Yang (1998) also co-treated some cells with 3-aminotriazole, a known

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catalase inhibitor. This co-treatment had no effect on Pb acetate-induced anchorage
independence, suggesting that catalase was not involved in this effect. It would be premature to
conclude that oxidative metabolism is not involved in anchorage independence, as these are the
only data available and are limited to catalase only. More data are needed to elucidate whether
oxidative metabolism is involved in this lead effect.
The remaining three studies focused on Pb chromate (Biedermann and Landolph, 1987,
1990; Sidhu et al., 1991). Two used similar HFF cells and found that Pb chromate-induced
anchorage independence (Biedermann and Landolph, 1987, 1990). However, these
anchorage-independent cells ultimately underwent senescence, suggesting that anchorage
independence may not be a suitable short-term marker for neoplastic transformation in primary
HFF. It should be noted that these studies were focused on the chromate component of this
compound and the potential contribution of Pb was not investigated or discussed. By contrast,
Sidhu et al. (1991) found that Pb chromate did not induce anchorage independence in a human
osteosarcoma cell line, while it did induce full neoplastic transformation of these cells and the
transformed cells did grow in agar. It should be noted that this study was also focused on the
chromate component of this compound and that the potential contribution of Pb was not
investigated or discussed.
The 1986 Lead AQCD did not include any studies of transformation in human cells.
Given that other chromate compounds have been shown to induce anchorage independence, it
seems quite possible that the data from Pb chromate exposures may represent effects from
chromate and not from Pb. Thus, the data currently seem to indicate that Pb can induce
anchorage independence in human cells, but its ability to induce neoplastic transformation of
human cells is uncertain. Further study of different Pb compounds and the full assessment of
their neoplastic potential (i.e., including studies of the ability of treated cells to form tumors in
experimental animal models) are needed before definitive conclusions can be drawn.

Animal Cell Cultures
The 1986 Lead AQCD presented several studies demonstrating that Pb compounds could
induce anchorage independence and morphological and neoplastic transformation in rodent cell
culture systems. Since that report, six studies have further considered the ability of Pb
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compounds to induce these effects. Three focused on Pb chromate and three on Pb compounds
without the confounding factor of chromate; these studies are summarized in Table AX5-6.3.
Four studies considered Pb acetate, Pb chloride, or Pb nitrate in Syrian hamster embryo
and C3H10T1/2 mouse embryo cells (Zelikoff et al., 1988; Patierno et al., 1988; Patierno and
Landolph, 1989; Elias et al., 1991). Three found that Pb compounds did not induce
transformation (Patierno et al., 1988; Patierno and Landolph, 1989; Elias et al., 1991); but the
third study (Zelikoff et al., 1988) indicated that Pb was weakly positive, though no statistics were
performed to validate this conclusion. Zelikoff et al. (1988) indicated that the observations were
repeated several times, but only showed data from one experimental run. It is unclear why the
studies were not averaged together, as multiple repeats would likely have provided the power to
detect whether the observed weak increase was significant.
Five studies considered Pb chromate, which induced neoplastic and morphological
transformation of Syrian hamster and mouse C3H10T1/2 embryo cells, as well as enhancing
viral transformation (Patierno et al., 1988; Patierno and Landolph, 1989; Schectman et al., 1986;
Elias et al., 1989, 1991). The focus on Pb chromate was based largely on concern about
chromate; but these studies found that Pb chromate was more potent than other chromate
compounds, suggesting that Pb may enhance or contribute to the carcinogenicity. Indeed, one
study found that combining Pb nitrate with soluble chromate was as potent as Pb chromate and
greater than soluble chromate alone (Elias et al., 1991).
Thus, all together, these studies suggest that Pb ions alone cannot transform rodent cells;
however, they may be co-carcinogenic or promote the carcinogenicity of other compounds.
These data are in contrast to findings described in the 1986 Lead AQCD that included a positive
study. One possible factor may be exposure duration; the study in question indicated that the
Pb-transformed cells were exposed for 9 days. The studies discussed here all exposed cells for
7 days or less. Further careful study of a time course of exposure is necessary to determine
whether Pb actually induces transformation in cultured rodent cells.

Nonmammalian Cell Cultures
No carcinogenesis studies were located that used nonmammalian cell culture models.
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5.6.2.4 Organ-Specific Studies
No organ-specific or organ culture studies concerning Pb carcinogenesis were located.

5.6.2.5 Carcinogenesis Summary
It remains difficult to conclude whether Pb is a human carcinogen. The assessment of the
carcinogenicity of Pb through human epidemiological studies remains unclear. By contrast, the
studies confirm that Pb is an animal carcinogen and further extend our understanding of the
mechanism to include a role for metallothionein. The cell culture data suggest that Pb can
induce anchorage independence, but whether it can induce full neoplastic transformation of
human cells is uncertain. Both IARC and NTP have recently upgraded Pb to a 2A classification
(probable human carcinogen); in keeping with the most recent EPA cancer guidelines, while
acknowledging the inadequacy of the human data, Pb would likely be characterized as a probable
human carcinogen.
To conclude, animal tumorigenicity studies clearly implicate Pb (primarily tested as Pb
acetate) as being carcinogenic, although i.v. administration has been the main route of exposure
employed in such studies. Based on neoplastic transformation in animal cell culture studies, Pb
has also been implicated as a carcinogen with chromate.

5.6.3 Genotoxicity Studies
The human genotoxicity studies are only briefly reviewed in this section. For a more
detailed review, see Chapter 6 (Section 6.7) in this document.

5.6.3.1 Human Studies
Adults
A number of studies investigating the potential genotoxicity of Pb have been conducted in
human populations. Endpoints considered include chromosome aberrations, sister chromatid
exchanges (SCE), micronuclei formation, DNA strand breaks, and hypoxanthine guanine
phosphoribosyl transferase (HPRT) mutations. In general, these studies were much more
specific than the carcinogenesis studies, as correlations with blood Pb levels could be made,
other confounders could be ruled out, and the endpoints were more short-term.
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The chromosome damage studies are ambiguous and contained some methodological
flaws. Four studies were positive (Huang et al., 1988; De at al., 1995; Bilban, 1998; Pinto et al.,
2000), while two were negative (Anwar and Kamal, 1988; Rajah and Ahuja, 1996). Moreover,
the four positive studies included two that could not rule out potential contributions from other
genotoxic metals and one that found a correlation only at very high blood Pb levels (>52 |ig/dL).
By contrast, the studies of micronucleus formation (Bilban, 1998; Vaglenov et al., 1998;
Pinto et al., 2000; Palus et al., 2003; Minozzo et al., 2004), SCE (Huang et al., 1988; Bilban,
1998; Pinto et al., 2000; Duydu et al., 2001; Palus et al., 2003), and DNA strand breaks
(Restrepo et al., 2000; Fracasso et al., 2002; Hengstler et al., 2003; Danadevi et al., 2003; Palus
et al., 2003) all consistently found clear correlations between Pb and genotoxicity. It should be
noted that there were two negative studies for SCE (Rajah and Ahuja, 1995, 1996), but both were
by the same group and considered the same very small population of workers (only 5 Pb-exposed
workers) and, thus, may not have had enough power to detect potential differences.
It is notable that one study found an interesting correlation of HPRT mutation rates and
blood Pb levels from environmental Pb exposure in Belgian women (Van Larebeke et al., 2004).
This study is the first and only one to consider Pb-induced mutations. Further research is needed
to assess the validity of these findings.
Thus, it appears from these studies that Pb is genotoxic to humans, although it may not
induce substantial amounts of chromosome damage. This conclusion is consistent with the
laboratory studies discussed below. For more in-depth consideration of the epidemiology
studies, see Chapter 6, Section 6.7.

Children
Two recent studies of Pb-induced genotoxicity in children have been published. One
study of children living in a high Pb-contamination area of Czechoslovakia found no increase in
chromosome damage in white blood cells compared with children living in an area with lower Pb
contamination (Smejkalova, 1990). Comparisons were not done with children living in an area
with little or no Pb contamination. Analyses of blood Pb levels indicated a statistical difference
in blood levels between the two groups but not necessarily a substantial, or biologically
significant, difference between them. (Typically, the control group levels were in the high 20s
compared to the low 30s |ig/dL in the exposed group). Thus, the possibility that each group was
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exposed to a Pb level that could induce a baseline level of damage cannot be ruled out, and it
cannot be conclusively stated that Pb was not clastogenic in this study.
The other study found an increase in Pb-induced strand breaks in white blood cells from
children living in an area of Mexico with high Pb contamination compared to children living in
an area with lower Pb contamination (Yanez et al., 2003). Blood Pb levels confirmed a
difference in exposure to Pb, but urinary arsenic levels also showed that these children were
exposed to higher levels of arsenic, too; and, thus, it cannot be determined which chemical was
responsible for the damage.
The possible genotoxicity of Pb for children is an important concern. However, there are
simply too few data to draw definitive conclusions, and more research is needed. See Chapter 6
(Section 6.7) for more in-depth discussion of the epidemiology of Pb in human populations.

5.6.3.2 Laboratory Animal Studies
Fourteen studies evaluated the genotoxicity of Pb compounds in animal models. The
majority of these studies focused on mice, and the Pb was administered by intraperitoneal (IP) or
intravenous (IV) injection. Several endpoints were considered, including SCE, chromosome
aberrations, micronucleus formation, and DNA strand breaks. Overall, the results are
ambiguous, due in part to study design and the various endpoints considered. These studies are
summarized in Table AX5-6.4.
Lead compounds appear to be able to damage chromosomes, if only weakly. Two studies
with well-performed analyses were positive (Fahmy, 1999; Aboul-Ela, 2002). The other positive
studies observed that Pb could induce karyotypic arrangements, indicating a possible clastogenic
response; however, these studies did not analyze very many cells (Chakraborty et al., 1987;
Nayak et al., 1989a,b; Dhir et al., 1990, 1992a,b; Nehez et al., 2000). Some found chromosome
damage, but it did not increase with dose (Chakraborty et al., 1987; Nayak et al., 1989a,b; Dhir
et al., 1990). Altogether, the data suggest some role for Pb in inducing chromosome damage, but
it may be a weak effect.
Similarly, the data for micronuclei and DNA damage are ambiguous. One study found
that Pb induced micronucleus formation in a dose-associated manner, but only considered two
doses (Roy et al., 1992). The other study found that Pb induced micronucleus formation but not
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in a dose-dependent manner (Jagetia and Aruna, 1998). This difference may reflect the
somewhat shorter exposure time in the second study.
One DNA damage study found that Pb nitrate could induce DNA strand breaks in the
white blood cells of mice (Devi et al., 2000); however, the damage was not dose-dependent.
Another found DNA damage in a number of organs, but only one dose was considered and the
authors described the effect as weak (Valverde et al., 2002). In both studies, the highest doses
caused less damage than the moderate to low doses. These data again suggest that Pb is only
weakly causing damage.
By contrast, the results for SCE have been consistently positive. The three studies that
were positive found that SCEs were induced in a dose-dependent manner (Fahmy, 1999; Nayak
et al., 1989a; Dhir et al., 1993).
The route of administration complicates the interpretation of all of these genetic studies.
All of the studies, except for three chromosome damage studies, used injection-based exposures.
It is unknown if exposures that reflect more realistic scenarios (e.g., Pb exposure via drinking
water) would cause any of these effects. Only one study of DNA strand breaks used a
physiologically relevant exposure (inhalation).
Four studies exposed animals by gavage, which is still a somewhat artificial exposure.
One was a DNA damage study that found weak activity (Devi et al., 2000). The other three
considered chromosome damage (Aboul-Ela, 2002; Dhir et al., 1992b; Nehez et al., 2000). Two
found a dose-response for a 24 h-exposure to Pb nitrate-induced chromosome aberrations in mice
(Aboul-Ela, 2002; Dhir et al., 1992b). The other found that a 4-week exposure to Pb acetate
induced aneuploidy, but not chromosome aberrations, in rats (Nehez et al., 2000). It is difficult
to reconcile these two studies, as they use different exposure times, chemicals, and species.
More work is needed using relevant doses and exposure conditions to Pb compounds in multiple
species to determine if Pb induces chromosome aberrations.
Some studies also tried to offset the effects of Pb with a variety of compounds. Potential
modulators included fruit extract from Phyllanthus emblica, ascorbic acid, calcium, and iron
(Aboul-Ela, 2002; Dhir et al., 1990, 1992a, 1993; Roy et al., 1992). Other studies sought to
determine if coexposure to other toxicants would potentiate the effects of Pb (Dhir et al., 1992b;
Nehez et al., 2000) and considered both zirconium and cypermethrin. The data indicated that the
fruit extract could block the toxic effects of Pb, an effect that may, in part, be attributable to
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ascorbic acid, but that other components must also be involved, because ascorbic acid alone
produced variable results. Iron also had an effect, but only if given just before, or with, the Pb
compound; post treatments with iron had no effect. Calcium had a strong effect.
The effects seen with zirconium and cypermethrin are less clear. Both were reported to
exacerbate the effects of Pb, but the effects for both are complicated by experimental design
problems. For example, zirconium only exacerbated Pb's effects when given simultaneously but
not when given 2 h before, or after, Pb. This seems rather unusual, as the total exposure to each
was 24 h and, thus, simultaneous exposure occurred in every circumstance. Hence, the data
seem to suggest that a 22-h coexposure had no effect, but that a 24-h exposure did. There may
have been an interaction of the two chemicals in the gut during coexposure, creating a more toxic
species.
Interpretation of the cypermethrin study is complicated by its design and the results. Only
20 metaphases were analyzed per animal, instead of the recommended 100. Also, the statistical
analyses were done relative to untreated controls and not to animals treated with Pb or
cypermethrin alone. Careful inspection of the tables reveals that actual exposure to Pb plus
cypermethrin induced less damage than that induced by Pb alone. Thus, the effects of them
together appear to be less than additive. More work is needed to explore the meaning of these
data and the importance of Pb mixtures.
The previous report revealed a similar amount of ambiguity; some animal studies were
positive for chromosome damage and others were negative. Other endpoints were not described
after Pb exposure in experimental animals. These data suggest that Pb can induce SCE, but that
it can induce chromosome damage, DNA damage, or micronuclei either weakly or not at all.

5.6.3.3 Cell Culture Studies
Few cell culture studies were reported in the 1986 Lead AQCD. Since 1986, a great deal
of theoretical and technological progress has allowed for a large number of cell culture studies to
be performed, as discussed below.
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Human Cell Culture
Mutagenicity
Two studies considered Pb acetate-induced mutagenesis in human cells. Both considered
mutations at the HPRT locus, with one using keratinocytes and the other skin fibroblasts (Ye,
1993; Hwua and Yang, 1998). These studies are summarized in Table AX5-6.5.
One study reported no Pb-induced mutagenesis (Hwua and Yang, 1998) but sought to
explore the importance of oxidative metabolism in Pb-induced mutagenesis by co-treatment with
3-aminotriazole, a known catalase inhibitor. This co-treatment did not increase Pb acetate-
induced mutagenesis, suggesting that either catalase was not involved in this effect or that Pb is
truly not mutagenic. It would be premature to conclude that oxidative metabolism is not
involved in anchorage independence, as these are the only data and are limited to catalase. More
data are needed to elucidate whether oxidative metabolism is involved in this effect of Pb, as
well as further studies of Pb-induced mutagenesis.
The other study reported that Pb acetate-induced mutagenesis (Ye, 1993). However,
interpretation of this study is hampered by its methodology. The study did not actually measure
HPRT mutations or colony formation, but rather it attempted a quicker methodology that
measured tritium incorporation. Although a shorter assay is highly desirable, the study did not
verify the observed effects with standard methods, and, thus, it is uncertain if the tritium
incorporation actually reflected Pb-induced mutations.
One study considered Pb chromate and found that it was not mutagenic (Biedermann and
Landolph, 1990).
There are insufficient data at this point to conclude whether Pb is mutagenic in human
cells or not, but the few data available are largely negative.

Clastogenicity
Ten studies investigated the ability of Pb compounds to induce chromosome damage in
cultured human cells. All but one were essentially from the same research group, and all but two
considered Pb chromate. All were done using normal, or nearly normal, human cells. These
studies are summarized in Table AX5-6.6.
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Only two of those studies focused on the clastogenicity of Pb itself (Wise et al., 2004b,
2005); the remainder used Pb compounds but focused on either chromate or radioactive particles
as the clastogenic species. These studies found that Pb glutamate was not clastogenic.
All of the Pb chromate studies found that Pb chromate induced chromosome damage in a
concentration-dependent manner. However, the effects were either attributed or demonstrated to
be caused by chromate ions. Lead ions were produced by Pb chromate, but they were not
clastogenic.
There was one study of radioactive Pb (Martins et al., 1993). The focus was on the
clastogenic activity of alpha particles, and the identity of the specific Pb salt was not provided.
The alpha particles were able to induce chromosome damage.
Overall, the data appear to indicate that Pb does not induce chromosome damage in
human cells, although more investigation of different compounds is needed.

DNA Damage
Studies of DNA damage in cultured human cells have considered DNA strand breaks,
Pb-DNA adducts, and DNA-protein crosslinks for a variety of Pb compounds. The only clear
positive damage induced by Pb was Pb-DNA adducts following Pb chromate exposure, although
the authors referred to them as Pb associated with DNA (Singh et al., 1999). It is uncertain if
these represent actual adducts or some weaker association. Two studies found no DNA strand
breaks induced by Pb (Hartwig et al., 1990; Snyder and Lachmann, 1989), and one study
involving several laboratories found no DNA-protein crosslinks after Pb exposure (Costa et al.,
1996). The other study found DNA double-strand breaks, but these were attributed to chromate
and not Pb (Xie et al., 2005). These studies are summarized in Table AX5-6.7.
One other study was positive (Wozniak and Blasiak, 2003), but the results were unusual
and their impact uncertain. Specifically, this study found that Pb acetate induced DNA
single-strand breaks but that the amount of damage decreased with concentration, and ultimately
the highest concentration had less damage than the control. DNA double-strand breaks were
observed, but were lowest at the highest concentration. DNA-protein crosslinks were seen only
at the highest concentration, and the authors attempted to explain the decrease in strand breaks
with this effect. This explanation may partially correct, but it does not entirely explain the
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decreased amount of damage at the middle concentration. These data need to be repeated by an
independent group before they can be fully assessed.
Together, these data suggest that Pb likely does not induce DNA damage; however, the
data are still too limited to allow any definitive conclusions.

Human Cell Genotoxicity Summary
The cumulative data suggest that Pb is not mutagenic and does not induce chromosome
aberrations or DNA damage in cultured human cells. It is interesting to note that Pb-induced
SCEs have not been considered in human cells.

5.6.3.4 Animal Cell Cultures
Mutagenicity
The potential mutagenicity of Pb compounds in rodent cells was considered in six studies.
In particular, three mutagenesis systems were considered: mutagenesis at the HPRT locus, the
gpt locus, and mutations in sodium-potassium ATPase. The results are highly variable and may
be specific to the Pb compound considered in each case. In particular, Pb chromate and Pb
acetate appear to be nonmutagenic. Lead acetate was positive but only at highly cytotoxic
concentrations. By contrast, Pb chloride and Pb sulfate appeared to be mutagenic at relatively
nontoxic concentrations. These studies are summarized in Table AX5-6.8. However,
insufficient data exist at this point to conclude whether or not Pb is mutagenic in animal cells.

Clastogenicity
Seven studies investigated the ability of Pb compounds to induce chromosome aberrations
in cultured mammalian cells (Table AX5-6.9). Four of these studies considered Pb chromate,
and further investigation revealed that chromate was responsible for the clastogenic effect (Wise
et al., 1992, 1993; Blankenship et al., 1997). Three of these studies considered other Pb
compounds (Wise et al., 1994; Lin et al., 1994; Cai and Arenaz, 1998). All but one were
negative and that one only found a small response at a single high dose (Wise et al., 1994).
Lower doses had no effect. Considered together, the studies indicate that Pb does not induce
chromosomal aberrations in cultured mammalian cells.
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Only two studies considered Pb-induced micronuclei in cultured mammalian cells. One
was negative (Lin et al., 1994) and the other positive (Bonacker et al., 2005).
Four studies considered Pb-induced SCE in cultured mammalian cells. The results were
predominately negative (three studies [Hartwig et al, 1990; Lin et al., 1994; Zelikoff et al.,
1988]). Interpreting these studies, however, is complicated by the fact that too few metaphase
cells (less than 30 per concentration) were analyzed in each study. The one positive study
considered 100 metaphases per concentration, making those data more reliable (Cai and Arenaz,
1998).

DNA Damage
Several measures of DNA damage in cultured human cells have been investigated,
including DNA single-strand breaks and DNA-protein crosslinks. Most Pb compounds did not
induce DNA single-strand breaks. The exception was Pb chromate, which did induce DNA
strand breaks, but this effect was likely a result of the chromate ion. These studies are
summarized in Table AX5-6.10.
Both Pb chromate and Pb nitrate induced DNA-protein crosslinks in cultured mammalian
cells. These data suggest that Pb is genotoxic in this manner; however, it is thought that the Pb
chromate-induced DNA-protein crosslinks result from the chromate and that the method used for
Pb nitrate is not sufficiently rigorous. Thus, while the data are certainly suggestive, they are
insufficient to make any definitive conclusion.

Nonmammalian Cell Cultures
Only one study was located considering Pb in a nonmammalian model (Table AX5-6.11).
This study found that Pb chromate was not mutagenic in a bacterial assay. The compound was
studied because of its chromate content and, given that it is the lone study, no definitive
conclusions can be reached.

5.6.3.5 Cell-Free Studies
No cell-free studies concerning Pb carcinogenesis or genotoxicity were located.
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5.6.3.6 Organ-Specific Studies
One study (Valverde et al., 2002) considered organ-specific effects (see Table AX5-6.4).
That study found a different pattern of DNA strand breaks in mice after inhalational exposure to
Pb acetate. DNA in the brain and lung were damaged the most, kidney and liver next, then nasal
epithelia and leukocytes, with no damage in testicle DNA. These data are intriguing, as they
suggest organ-specific responses after a physiologically relevant exposure (inhalation). More
research is needed, however, to fully assess the impact of these findings. Moreover, while the
damage was statistically significant, the authors described the effects as weak.

5.6.3.7 Genotoxicity Section Summary
There is some ambiguity in the genotoxicity results, as some endpoints were positive
while most were negative. Consistent with the animal study data, Pb can induce SCE in rodent
cells, but it is unknown if it can do so in human cells because this has not been tested. Lead also
seems to induce DNA-protein crosslinks in rodent cells.

5.6.4 Genotoxicity as it Pertains to Potential Developmental Effects
The human genotoxicity studies are only briefly reviewed in this section. For a more
detailed review, see Chapter 6 (Section 6.7). Only limited animal data and no cell culture studies
focused on this issue as a concern. The available data are described below.

Adults
One study was located that considered the effects of Pb on sperm quality and quantity.
This study considered Pb, cadmium, and selenium levels in 56 nonsmoking volunteers (Xu et al.,
2003). No effects on sperm quality were correlated with Pb exposure up to 10 |ig/L. Two other
studies were located on the effects of Pb on sperm morphology in animals (Fahmy, 1999; Aboul-
Ela, 2002). Both were positive, indicating that Pb may have an effect on sperm. They also
found that Pb induced DNA damage in the sperm (See Table AX5-6.4). These studies are
summarized in Table AX5-6.12.
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Children
No studies were analyzed that considered the genotoxic effects of Pb in children as a
developmental hazard. There are two studies that considered the genotoxic effects of Pb in
children. They were discussed in Section 5.6.3.1.
Three studies were located on the fetal effects of Pb nitrate on the fetus (Kristensen et al.,
1993; Nayak et al., 1989a,b). Lead induced an increase in resorptions and there were hints of
possible fetal chromosome damage, but the methods were poorly described and much more work
is needed before conclusions can be drawn. These studies are summarized in Table AX5-6.13.

5.6.5 Epigenetic Effects and Mixture Interactions
Lead has been proposed to be a co-mutagen or possibly a promoter. Thus, a number of
epigenetic mechanisms have been proposed. Epigenetic effects occur when a compound such as
Pb induces changes in cellular processes that do not result from changes in DNA sequence.
In other words, Pb has been proposed to alter cells in ways that may change the cell without
breaking or mutating DNA. There are three possible mechanisms: (1) alterations of gene
expression that can stimulate cells to grow (mitogenesis) and/or can interfere with DNA repair;
(this possibility has been investigated in several studies); (2) interaction with other metals; and
(3) alteration of oxidative metabolism. Neither of the latter two have been extensively studied.

5.6.5.1 Gene Expression
It has been argued that Pb may induce or co-induce carcinogenesis by altering cellular
metabolism or by altering the metabolism of another chemical. Both whole animal and cell
culture studies have been conducted to address this question and are described below.

Animal
Animal studies indicate that Pb can induce the expression of some phase I metabolizing
enzymes, such as cytochrome P4501 Al, and phase II metabolizing enzymes, such as glutathione
and glutathione-S-transferase. These studies are summarized in Table AX5-6.14. Thus, it is
plausible that through this mechanism, Pb may act as a co-carcinogen by affecting the
metabolism of other chemicals or possibly as a direct carcinogen by enhancing endogenously
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induced damage. However, no studies have directly shown that such Pb effects are linked to
cancer or alter the potency of another chemical; and, thus, it remains only a plausible hypothesis.

Human Cell Culture Studies
A few human cell culture studies have been done, and these generally confirm the animal
studies. These studies are summarized in Table AX5-6.15. Lead has been shown to affect the
induction of some phase I metabolizing enzymes (such as cytochrome P4501A1) and phase II
metabolizing enzymes (such as glutathione and glutathione-S-transferase and NAPDH oxidase).
These experiments also indicate that Pb can affect the metabolism of other carcinogenic
compounds, although they do not show that the genotoxic or carcinogenic effects change as a
result of these effects; and, thus, more work remains to make this more than just a plausible
explanation.

Animal Cell Culture Studies
No animal cell culture studies concerning the effects of Pb on the expression of metabolic
genes were located.

5.6.5.2 DNA Repair
It has been argued that Pb may induce or co-induce carcinogenesis by altering the repair
of DNA lesions induced by another agent. The greatest focus has been on damage induced by
ultraviolet (UV) light. Only cell culture and cell-free studies have been conducted to address this
question and are described below.

Human
Only one study considered Pb-induced effects on DNA repair in cultured human cells (see
Table AX5-6.16). This study found that coexposure to Pb caused persistence of strand breaks
induced by UV light. This persistence suggests that Pb interfered with the repair of these lesions,
but direct evidence of that interference was not provided. These are the only data in human cells;
and, thus, it cannot be determined if Pb inhibits DNA repair in human cells.
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Mammalian Cell Culture Models
Two studies considered Pb-induced effects on DNA repair in cultured mammalian cells.
These studies are summarized in Table AX5-6.17. Both found that Pb acetate increased
UV-induced DNA damage including SCE, mutagenesis, and cytotoxicity. Lead did not affect
strand breaks induced by UV. These data suggest that Pb may indeed inhibit repair, although
direct interactions with repair proteins were not demonstrated.

Cell Free Systems
One study considered the effects of Pb on DNA repair proteins (McNeill et al., 2004).
That study found that Pb can inhibit APE nuclease in cell-free systems.

5.6.5.3 Mitogenesis
It has been argued that Pb may induce or co-induce carcinogenesis by inducing cells to
grow when they should not. Both animal and cell culture studies have been conducted to address
this question and are described below.

5.6.5.3.1 Animal
Several studies have considered Pb-induced mitogenesis in animal models. These studies
are summarized in Table AX5-6.18. These studies found that Pb can stimulate cell growth, but
primarily in the liver. One study did consider TNF-a expression in brain cells, but it was not
demonstrated whether these effects were mitogenic. The interpretation of many of the studies is
complicated by the exposure method (IV injection), which does not reflect human exposure.
In general, the data indicate that Pb is mitogenic to the liver.

Human Cell Culture Studies
A number of studies have considered the potential growth-stimulatory effects of Pb in
cultured human cells (Table AX5-6.19). These studies all found that Pb did not stimulate cell
growth. Thus, mitogenesis is not a likely epigenetic effect for Pb in human cells.
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Mammalian Cell Culture Studies
A number of studies have considered the potential growth-stimulatory effects of Pb in
cultured mammalian cells other than the kidney. These studies all found that Pb did not
stimulate cell growth. Thus, mitogenesis is not a likely epigenetic effect of Pb in human cells.
One study found an increased mitotic index; however, it did not consider possible cell cycle
arrest (Lin et al., 1994). Indeed, another study found that Pb increased the mitotic index, because
it induced M-phase arrest (Wise et al., 2005).

Other
Lead-induced oxidative damage has been investigated as a potential cause of genotoxic or
carcinogenic effects. Generally, the results suggest that Pb only produces low levels of reactive
oxygen species, but that it may inhibit some enzymes involved in oxidative metabolism
(Table AX5-6.20). Thus, Pb may affect oxidative metabolism, but more work is needed to draw
meaningful conclusions.

5.6.5.4 Epigenetic Mechanisms Summary
The collective data support the hypothesis that Pb can induce an epigenetic effect. Lead
can alter the expression of metabolic genes in cultured cells and may alter DNA repair, although
much more study is needed. Lead may also affect oxidative metabolism or interact with other
metals, but again more study is needed. By contrast, it is unclear if Pb is mitogenic. It is
mitogenic to the liver in animals, but it is not mitogenic in cultured cells. More study is needed
to determine if this difference reflects differences between in vivo and cell culture models or if
this property is specific to only certain organs, e.g., the liver.

5.6.6 Summary
• Overall, the above studies confirm that Pb is an animal carcinogen and extends our
understanding of mechanisms involved to include a role for metallothionein. Specifically,
the recent data show that metallothionein may participate in Pb inclusion bodies and, thus,
serves to prevent or reduce Pb-induced tumorigenesis.
• Much more work is needed to determine the potential exacerbating or ameliorating roles of
calcium and selenium and to determine what role Pb-induced immunomodulation may
play in the promotion of tumors.
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All together, these studies suggest that Pb ions alone cannot transform rodent cells;
however, they may be co-carcinogenic or promote the carcinogen!city of other
compounds. These data are in contrast to findings described in the 1986 Lead AQCD that
included a positive study. One possible factor may be exposure duration; the study in
question indicated that the Pb-transformed cells were exposed for 9 days. The studies
discussed here all exposed cells for 7 days or less. Further careful study of a time course
of exposure is necessary to determine whether Pb actually induces transformation in
cultured rodent cells.

The previous report found a similar amount of ambiguity; some animal studies were
positive for chromosome damage and others were negative. Other endpoints were not
described after Pb exposure in experimental animals.

These data suggest that Pb can induce SCE, but that it can only induce chromosome
damage, DNA damage, or micronuclei either weakly or not at all.

Overall, the data appear to indicate that Pb does not induce chromosome damage in human
cells, although more investigation of different compounds is needed.

Together, these data suggest that Pb likely does not induce DNA damage; however, the
data are still too limited to allow any definitive conclusions.

There is some ambiguity in the genotoxicity results, as some endpoints were positive while
most were negative. Consistent with the animal study data, Pb can induce SCE in rodent
cells, but it is unknown if it can do so in human cells because this has not been tested.
Lead also seems to induce DNA-protein crosslinks in rodent cells.

The collective data support the hypothesis that Pb can induce an epigenetic effect.

Lead can alter the expression of metabolic genes in cultured cells and may alter DNA
repair, although much more study is needed.

Lead may also affect oxidative metabolism or interact with other metals, but again more
study is needed.

It is unclear if Pb is mitogenic. It is mitogenic to the liver in animals, but it is not
mitogenic in cultured cells. More study is needed to determine if this difference reflects
differences between in vivo and cell culture models or if this property is specific to only
certain organs, e.g., the liver.

The overall conclusions have not changed much from the 1986 Lead AQCD. Lead
remains an ambiguous carcinogen in humans and a clear carcinogen in animals.

Cell culture studies support these conclusions, as effects in rodent cells were not seen in
human cells.

Lead does appear to be genotoxic in human epidemiology studies.
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By contrast, the laboratory studies are more ambiguous in both animal and cell culture
studies. In these systems, the genotoxicity in culture is limited to SCE and, perhaps, to
DNA-protein crosslinks. For other endpoints, it is only weakly active, if at all.
Lead has not been evaluated sufficiently as a potential genotoxic hazard, but this probably
stems from the fact it appears to be weakly genotoxic.
The available data suggest that Pb can damage sperm and affect fetuses. More work is
urgently needed on this topic.
Cell culture studies do support a possible epigenetic mechanism or co-mutagenic effects.
5.7 LEAD AND THE KIDNEY
5.7.1 Review of Earlier Work
This section summarizes key findings from the 1986 Lead AQCD with regard to Pb
effects on the kidney in animals. Human studies published since 1986 are then reviewed in
Section 6.4.
Both in vivo and in vitro studies on several different animal species revealed that renal
accumulation of Pb is an efficient process that occurs in both proximal and distal portions of the
nephron and at both luminal and basolateral membranes (Victery et al., 1979a; Vander et al.,
1977). The transmembrane movement of Pb appears to be mediated by an uptake process that is
subject to inhibition by several metabolic inhibitors and the acid-base status of the organism.
Alkalosis increases Pb entry into tubule cells via both the luminal and basolateral membranes
(Victery et al., 1979b).
Goyer et al. (1970a) were principally responsible for defining the role of renal proximal
tubular nuclear inclusion bodies in the response to Pb intoxication. In addition to the early
reports of nuclear inclusion bodies appearing in the proximal tubule following Pb exposure
(Goyer et al., 1970b), biochemical studies on the protein components of isolated rat kidney
intranuclear inclusion bodies have shown that the main component has an approximate molecular
weight of 27 kDa (Moore et al., 1973) or 32 kDa (Shelton and Egle, 1982) and is rich in
glutamate and aspartate. Goyer et al. (1970c) suggested that the intranuclear inclusion body
sequesters Pb, to some degree, away from sensitive renal organelles and metabolic pathways.
Goyer and Wilson (1975) and Goyer et al. (1978) also showed that single or repeated
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administration of CaNa2EDTA leads to the disruption of the nuclear inclusion bodies and their
removal from the nuclei. Rats treated for 24 weeks with both Pb and CaNa2EDTA had no
inclusion bodies, but showed early interstitial nephropathy. As an extension of this study,
Cramer et al. (1974) examined renal biopsies from 5 Pb workers with 0.5 to 20 years of
exposure. The two workers with normal GFRs, and shortest exposure duration, showed
intranuclear inclusion bodies, whereas the remaining three workers had no intranuclear
inclusions but showed peritubular fibrosis.
Formation of intranuclear inclusion bodies was a common pathognomic feature for all
species examined. In addition, proximal tubular cytomegaly and swollen mitochondria with
increased numbers of cytosomes were also observed (Fowler et al., 1980; Spit et al., 1981). The
morphological changes were principally localized in the straight (S3) segments of the proximal
tubule. Goyer (1968) and Goyer et al. (1968) had demonstrated earlier that, after Pb exposure,
mitochondria were not only swollen but also had decreased respiratory control ratios (RCRs) and
inhibited state-3 respiration.
Aminoaciduria has been reported in several studies (Studnitz and Haeger-Aronson, 1962;
Goyer et al., 1970b; Wapnir et al., 1979). Other studies have reported increased urinary
excretion of electrolytes (e.g., sodium, potassium, calcium, water) following Pb administration
(Mouw et al., 1978). Victery et al. (1981, 1982a,b, 1983) found that zinc excretion increased
following Pb injection.
Wapnir et al. (1979) observed that Pb acetate administration caused a reduction in renal
alkaline phosphatase activity and an increase in Mg-ATPase activity, but no significant changes
in NaK-ATPase activity. On the other hand, Suketa et al. (1979) found marked a decrease in
renal NaK-ATPase activity following a single oral administration of Pb acetate at a dose of
200 mg/kg, but no change in Mg-ATPase.
Renal ALAD was found to be inhibited by Pb in both acute and chronic experiments
(Silbergeld et al., 1982). Renal ALAD was similar to control levels when GSH was present but
was significantly reduced in the absence of GSH (Gibson and Goldberg, 1970). Accumulation of
both ALA and porphobilinogen was also observed in kidney tissue of Pb-treated rabbits,
compared to controls. Other studies have not shown a reduction in renal ALAD following Pb
exposure (e.g., Fowler et al., 1980). Higher levels of Pb may be required to cause the reduction
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in ALAD reported by Silbergeld et al. (1982), and it may possibly involve Pb-binding proteins in
the kidney.

5.7.2 Markers of Renal Toxicity
The establishment and validation of new screening tests for nephrotoxic effects have been
principally due to the efforts of the Belgian group (Price et al., 1996; Price, 2000; Lauwerys
et al., 1992). They proposed that the following battery of tests be used to screen both
environmentally exposed and occupationally exposed individuals: (1) measures of glomerular
integrity, i.e., urinary high-molecular weight proteins (albumin, IgG, transferring (2) measures
of tubular absorption and secretion, i.e., low-molecular weight proteins (retinol binding protein,
a-1-microglobulin; (3) measures of tubular integrity, i.e., enzymes, lysosomal N-acetyl
P-D-glucosaminidase (NAG), brush border alanine aminopeptidase, brush border intestinal
alkaline phosphatase, nonspecific alkaline phosphatase, a-glutathione-S-transferase (GST), and
brush border antigens (BB50, BBA, HF5); (4) measures of glomerular and distal tubular
function, i.e., prostanoids (thromboxane B2, prostaglandin F2 alpha, 6-keto prostaglandin
Fl alpha); (5) measures of glomerular structural proteins (fibronectin and laminin fragments); and
(6) measures of distal tubular function, i.e., Tamm-Horsfall protein and u-GST. Other useful
markers include urinary p2-microglobulin, as a marker of proximal tubular integrity; PGE2 and
PGF2, distal nephron markers; kallikrein, a marker of the distal tubule; lysozyme, ribonuclease,
and y-glutamyl transferrase, enzymes reflecting proximal tubule integrity; and sialic acid, an
extracellular matrix marker (Pels et al., 1994; Pergande et al., 1994; Taylor et al., 1997). One or
several of these urinary markers have been used in screening tests for human Pb workers and in
animal studies of renal nephrotoxicity, although none has proved to be specific for Pb.
Questions have been raised about the usefulness of urinary NAG as a nephrotoxic marker
due to the absence of light or electron microscopic changes in low-dose Pb-treated animals that
showed substantial increases in NAG (vide infra) (Khalil-Manesh et al., 1993b). Furthermore,
Chia et al. (1994) found that urinary NAG in workers exposed to Pb correlated best with recent
blood lead changes, suggesting that the increased urinary NAG activity reflected an acute
response to a sharp increase in the renal Pb burden rather than to exocytosis. Questions have
also been raised about the value of measuring the vasoconstricting prostariod cytokine
thromboxane B2 (TXB2) and the vasodilating prostanoid 6-keto prostaglandin Fl alpha (PGF1
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alpha). Conflicting results have been reported in human Pb-exposed workers. Cardenas et al.
(1993) reported an elevation in TXB2 and a diminution in PGF1 alpha in 41 Pb-exposed workers
in contrast to 41 controls. Hotter et al. (1995), on the other hand, reported that both substances
were increased in 69 Pb-exposed workers in contrast to 62 controls. Blood Pb levels in the two
worker groups were comparable, i.e., 48 |ig/dL in the first group and 43 |ig/dL in the second. In
animal experiments (Gonick et al., 1998), the excretion of both prostanoids was equal in low-Pb
(100 ppm)-fed rats as contrasted to normal controls after 3 months, despite an elevation in blood
pressure in the Pb-fed rats. Blood Pb in the Pb-fed rats averaged 12.4 |ig/dL compared to
1 |ig/dL in the controls. Thus, measurements of these prostanoids remain of questionable value.
Attempts to validate nephrotoxic markers were conducted by Pergande et al. (1994),
utilizing Pb-exposed workers as contrasted to normal controls. They found that about 30% of the
Pb workers showed an increased excretion of ai-microglobulin, NAG, ribonuclease, and/or
Tamm-Horsfall protein, with positive correlations between these tubular indicators and blood Pb
concentration.

5.7.3 Biochemical Mechanisms of Lead Toxicity
Nolan and Shaikh (1992) summarized what was known about biochemical mechanisms
underlying Pb-induced toxicity at that time. A more detailed description based on recent animal
studies follows in the next section.
The initial accumulation of absorbed Pb occurs primarily in the kidneys. This takes place
mainly through glomerular filtration and subsequent reabsorption, and, to a small extent, through
direct absorption from the blood. Lead may be taken up by the renal tubular epithelial cells from
the basolateral side by active transport of the free ion. Smaller amounts can also cotransport
with low molecular weight organic anions. The uptake of Pb through the renal brush border does
not appear to occur via any specific carriers. Instead, the process may involve binding of Pb to
nonspecific surface sites on the brush border membrane, followed by internalization via
endocytosis. Acute kidney damage due to Pb manifests primarily in the proximal tubules. The
ultrastructural changes observed in acute experimental Pb nephropathy include both specific and
nonspecific effects on the proximal tubular epithelium, e.g., dilation of the endoplasmic
epithelium, blebbing of the nuclear membrane, enlargement of the autophagosomes, changes in
mitochondrial structure, formation of inclusion bodies. Chronic exposure to Pb affects
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glomerular filtration, renal clearance, and tubular reabsorption and can lead to renal failure from
interstitial nephritis.
Kidneys of chronically exposed individuals often show fewer or no nuclear inclusion
bodies compared to kidneys of acutely exposed individuals. The specific ultrastructural changes
associated with Pb nephropathy are the formation of cytoplasmic and nuclear Pb inclusion bodies
(discussed at greater length in Section 5.11). These inclusion bodies are not limited to the
proximal tubular epithelium, and they have also been observed in peritoneum, astrocytes,
neuroblastoma cells, lung cells, and osteoclasts upon Pb exposure. The inclusion bodies are
roughly spherical and typically consist of an electron-dense core, with a fibrillary network at the
periphery. Research has revealed that the formation of the nuclear inclusion bodies is preceded
by the synthesis of cytoplasmic inclusion bodies with a very similar structure. A protein unique
to these structures is rich in acidic amino acids and has an isoelectric point of 6.3 and a
molecular weight of 32 kDa. Two additional proteins with apparent molecular weights of
11.5 kDa and 63 kDa have been identified in kidney extracts. Both of these proteins have a high
affinity, but little capacity, for binding Pb. A Pb-binding protein of 12 kDa molecular weight
was identified in the supernatant of brain homogenate from Pb-treated rats. A Pb-binding
protein of 10 kDa has also been isolated from the erythrocytes of Pb-exposed workers.
Mitochondrial function, in addition to structure, is very sensitive to Pb. Changes include
the uncoupling of oxidative phosphorylation, decreased substrate oxidation, and modification of
ion transport processes. Other effects of Pb on cellular energetics include chelation of ATP and
inhibition of microsomal NaK-ATPase. These changes may account for the proximal tubular
dysfunction seen with acute Pb poisoning in children.
A new area of investigation of the mechanism of Pb toxicity was initially proposed by
Quinlan et al. (1988) and Hermes-Lima et al. (1991). Both investigators proposed that free
radicals, or ROS, stimulated by Pb, may accelerate iron-dependent lipid peroxidation, causing
tissue injury. Hermes-Lima et al. (1991) stated further that ALA, which is formed in large
amounts in Pb toxicity, may undergo enolization and autoxidation, yielding ROS. Autoxidation
of ALA, in the presence or absence of iron complexes, yields superoxide, peroxide, and hydroxyl
radicals. Gurer and Ercal (2000), based on several animal studies to be discussed below, have
proposed that antioxidant supplementation following Pb exposure may provide a partial remedy
by restoring the cell's antioxidant capacity.
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5.7.4 Animal Studies
Two excellent review articles have been written about the effects of heavy metals on, and
their handling by, the kidney (Barbier et al., 2005) as well as the mechanisms of kidney cell
injury from metals (Fowler, 1992). The interested reader is directed to these reviews, although
individual effects and mechanisms are discussed below.

5.7.4.1 Lead Toxicokinetics
DeVries et al. (1998) published a model for Pb toxicokinetics to be used in planning
treatment. The model is a four-compartment model with first-order kinetics. The four
compartments of this model are blood, bone, liver, and kidney. Soft tissues are represented by
the kidney and liver compartments. In addition, intake and excretion are included in the model.
Excretion of Pb is mainly via the kidneys (70 to 80%), via bile and feces (15%), via nails, hair,
and sweat (8%). The blood makes up the central compartment from which Pb is distributed after
uptake in the body. The blood compartment contains about 4% of the total Pb body burden and,
within this compartment, the Pb is mainly taken up by erythrocytes. The half-life of Pb in blood
is about 30 days. From the blood, Pb is distributed relatively quickly to the soft tissues and bone.
The distribution constant from blood to bone is much higher than the one from bone to blood,
resulting in the accumulation of Pb in bone. The half-life in the soft tissues is about 30 to
40 days. Most of the body burden of Pb can be found in the bone compartment (-94%), where
the half-life of Pb is several decades. Because of the vast amount of Pb in bone, a rebound in
blood Pb usually occurs after chelation therapy. This model can be compared with a
toxicokinetic model developed by Marcus (1985a,b,c) and further explored by Hogan et al.
(1998), as discussed in Chapter 4 of this document.
Dieter et al. (1993) examined the effect of the nature of the Pb salt on the oral intake of Pb
in male F344 rats. For 30 days, they administered doses of 0, 10, 30, and 100 ppm Pb in the
form of soluble Pb oxide, Pb acetate, Pb sulfide, and Pb ore. At 100 ppm of Pb acetate or soluble
Pb oxide, the rats developed -80 |ig/dL of blood and -200 jig/g of bone Pb levels, whereas rats
fed Pb sulfide or Pb ore developed -10 |ig/dL of blood Pb and 10 jig/g of bone Pb. In rats fed Pb
acetate or soluble Pb oxide, blood Pb progressively increased with increasing dose, whereas
measurable levels of Pb in the other two groups were observed only at the highest dose
(100 ppm).
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5.7.4.2 Pathology, Ultrastructural, and Functional Studies
Two important series of studies contrast the pathological and functional changes in the
kidney after prolonged exposure to Pb, with and without chelation therapy (i.e., DMSA or
CaNa2EDTA). In the first series of 3 long-term studies, Khalil-Manesh et al. (1992a,b, 1993b)
described the effects of Pb acetate on renal function and morphology in male Sprague-Dawley
rats fed a low-calcium diet. Lead acetate was used in concentrations of 0.5% (high dose) and
0.01% (low dose) in drinking water for periods from 1 to 12 months, and then Pb-exposed
animals were compared to pair-fed controls (12 rats in each group). In all studies, GFR was
measured as 125I-iothalamate clearance by a single injection technique. Urinary markers
included NAG, GST, and brush border antigens (BB50, HF5, and CG9) and were expressed as
units/g creatinine. Blood and urine Pb were measured prior to sacrifice in each group of animals.
Wet and dry weights of kidneys were determined, then the kidneys were processed for light,
electron, and immunofluorescent microscopy.
In the first study (Khalil-Manesh et al., 1992a), animals treated with continuous high-dose
Pb for 12 months reached a maximum blood Pb of 125.4 ±10.1 |ig/dL after 6 months, at which
time the dose of Pb was reduced from 0.5% to 0.1%. Blood Pb at the end of 12 months averaged
55 |ig/dL. Urine Pb remained above 100 jig/g creatinine at all times, but it was highest at
3 months, averaging 340 jig/g creatinine. In the Pb-treated animals, GFR was increased above
controls at 3 months (1.00 ± 0.14 versus 0.83 ± 0.26 mL/min/100 g body wt, p = 0.05), then
declined after 6 months to 0.78 ±0.16 versus 0.96 ± 0.08 mL/min/100 g body wt in controls
(Figures 5-9 and 5-10).
As indicated by the ratio of kidney dry/wet weight, increased kidney tissue mass was
observed during the first 3 months of Pb exposure, but decreased tissue mass was observed by
12 months. With regard to urinary markers, NAG was elevated above control levels at 3, 6, and
9 months of Pb exposure; GST was elevated at 3, 6, and 12 months of Pb exposure; and no
significant differences were observed in the brush border antigens. Proximal tubular nuclear
inclusion bodies were present at all time periods in Pb-treated animals. Enlargement of proximal
tubular cells and nuclei were seen beginning at 3 months. At 6 months, focal tubular atrophy and
interstitial fibrosis appeared, increasing in extent up to 12 months. Mitochondrial alterations,
consisting of rounding and elongation, appeared by 1 month and were persistent. Glomeruli
were normal through 9 months, but, at 12 months, they showed focal and segmental sclerosis.
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P<0.05
P<0.05
O)
O
0.0
1 3 6 9 12
Duration of exposure, months

Figure 5-9. Changes in GFR of experimental high-dose lead and control animals with
duration of exposure to lead. Open and closed bars represent GFR in
experimental and control rats, respectively.

Source: Khalil-Manesh et al. (1992a), with permission.
1.50
I 1.00-
o>
8
0.50 -
0.00
r = 0.703
0 15 30 45 60 75 90 105 120 135 150

Btood lead,,
Figure 5-10. Correlation between GFR and blood lead during the first 6 months of
high-dose lead exposure.

Source: Khalil-Manesh et al. (1992a), with permission.
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There were no electron-dense deposits and immunofluorescent studies were negative. Renal
arteries and arterioles were normal at all time point examined.
The second study (Khalil-Manesh et al., 1992b) consisted of the discontinuation of both
the high- and low-dose Pb exposure after 6 months, then treatment with three courses of DMSA
or discontinuation of high-dose Pb alone after 1, 6, and 9 months of Pb feeding. Controls were
pair-fed, exposed to Pb for 6 months, then removed from exposure for 6 months without
receiving DMSA. Low-dose Pb-treated rats showed no significant pathologically with or
without DMSA treatment but exhibited a significant increase in GFR after DMSA treatment
(1.09 ± 0.19 versus 0.88 ± 0.22 mL/min/100 g body weight; P < 0.03) (Figure 5-11). Urinary
markers remained unchanged, and there were no structural alterations by light or electron
microscopy. High-dose Pb-treated animals showed no functional or pathologic changes when Pb
exposure was discontinued after 1 month. However, when the duration of exposure was 6 or
9 months, GFR was decreased and serum creatinine and urea nitrogen were increased compared
to controls. Tubulointerstitial disease was severe. Administration of DMSA resulted in an
improvement in GFR (Figure 5-11) and a decrease in albuminuria, together with a reduction in
size and number of nuclear inclusion bodies in proximal tubules.
However, tubulointerstitial scarring was only minimally reduced. In conclusion, except
for a brief initial exposure, discontinuation of high-dose Pb exposure failed to reverse
Pb-induced renal damage. Treatment with the chelator, DMSA, improved renal function but had
less effect on pathologic alterations. Because GFR improved after DMSA treatment in both
low- and high-dose Pb-treated animals, irrespective of the degree of pathologic alterations, it
may be concluded that the DMSA effect is most likely mediated by hemodynamic changes.
The third study (Khalil-Manesh et al., 1993b) examined the course of events over
12 months in continuous low-level Pb-exposed animals. Maximum blood Pb levels in
experimental animals were reached at 3 months, averaging 29.4 ±4.1 |ig/dL. GFR was found
to be significantly increased above pair-fed controls at 1 and 3 months, but it was normal at
other time periods (1 month experimental, 1.18 ± 0.12 versus control, 0.76 ± 0.15mL/min/100 g;
p < 0.001; 3 month experimental, 1.12 ± 0.16 , versus control, 0.86 ± 0.10 mL/min/100 g;
p< 0.001) (Figure 5-12).
Levels of urinary NAG in Pb-exposed rats exceeded control levels at all time periods,
except at 12 months, when the normal increase with aging obscured differences between
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1.50
High Lead
I
1 1.00 -
0.50
Low Lead
ED6 C12
DMSA ED6

GROUPS
C12 DMSA
Figure 5-11. GFR in high-lead and low-lead experimental discontinuous (ED6)

and DMSA-treated rats (DMSA) as compared to controls (C12).

All rats were studied at 12 months.

*p < 0.01 when compared to ED6 and C12.
**p < 0.05 when compared to ED6.

Source: Khalil-Manesh et al. (1992b), with permission.
2.00
1.50
01
o
o
p<0 001
p<0 001
c

J. 1.00

tr
u.
o
0.50
t

Duration of Exposure (Months)

Figure 5-12. Changes in GFR in experimental and control rats, at various time periods.

Source: Khalil-Manesh et al. (1993b).
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experimental animals and controls (Figure 5-13). In contrast, urinary GST, a more specific
marker of metal-associated proximal tubular injury, was normal at all time periods. Proximal
tubular nuclear inclusion bodies were sparse and were observed only at 1 and 3 months.
600
500
~ 400
a
a
0
o> 300
O
<
2
200
100
p<0 001
p<0 001
p<0 05
p<0 001
369
Duralion of Exposure (Months)
Figure 5-13. Urinary NAG concentration in experimental and control rats at various
time periods.
Source: Khalil-Manesh et al. (1993b).
No other pathologic alterations were found in the kidneys until 12 months of exposure,
when mild tubular atrophy and interstitial fibrosis were seen. The absence of changes in urinary
GST accorded with the relative absence of morphologic changes, whereas the observed increases
in urinary NAG suggest that this enzyme may be an overly sensitive indicator of tubular injury,
more probably reflecting upregulation of the enzyme even in the absence of tubular injury.
It should be noted that both low-dose Pb-treated animals and high-dose Pb-treated animals
showed a "hyperfiltration" phenomenon during the first 3 months of Pb exposure. This
observation could be invoked as a partial explanation for the late changes of glomerulosclerosis
in the high-dose animals, but it cannot explain the lack of glomerular changes in the low-dose
animals. Thus, these studies join those of Roels et al. (1994) and Hu (1991) in humans that
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indicate that Pb nephropathy should be added to diabetic nephropathy as diseases that lead to
early hyperfiltration.
The second series of studies were performed by Sanchez-Fructuoso et al. (2002a,b).
Sanchez-Fructuoso et al. (2002a,b) evaluated the effect of CaNa2EDTA on tissue mobilization of
Pb in six-month-old Wistar rats initially treated with 500 ppm Pb acetate for 90 days, followed
by treatment with three courses of CaNa2EDTA 50 mg/kg/day for 5 days, separated by 9 days, or
placebo. Lead levels were measured in blood, urine, kidney, liver, brain, and femur. There was
no change in bone Pb after CaNa2EDTA compared to placebo, but Pb levels were significantly
reduced in all other tissues (Figure 5-14).
s
100

0
Eivd Pl> upotum
12
60 90 111 12S 139 d*y*
60 90 111 125 139 days
10

6
I
60
111 129 139 days
60
|
20
T End Pb *xpo*ur*
60 90 111 12S 139 (lays
Figure 5-14. Kidney, liver, brain, and bone lead levels in 56 Pb-exposed rats.
After 90 days of poisoning, animals were administered serum saline
(solid line) or calcium disodium EDTA (broken line).
Source: Sanchez-Fructuoso et al. (2002a), with permission.
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The authors emphasized that there was no redistribution to brain. Cory-Slechta et al.
(1987) had originally reported that, with one day of CaNa2EDTA chelation in Pb-exposed rats,
Pb is preferentially mobilized from bone and then redistributed to other organs, including brain,
but with further CaNa2EDTA treatment, brain levels return to baseline. The Sanchez-Fructuoso
et al. (2002a,b) findings stand in contrast, explained by the authors as being due to a 3-fold
higher level of CaNa2EDTA used by Cory-Slechta et al. (1987).
Sanchez-Fructuoso et al. (2002b) also evaluated pathologic changes, as well as the
response of ALAD activity before and after CaNa2EDTA treatment in the same rats. In the
90-day Pb-treated animals, the main findings were hypertrophy and vacuolization of medium and
small arteries (Figure 5-15); mucoid edema and muscular hypertrophy in arterioles; loss of cell
brush borders, cell loss, and intranuclear inclusion bodies in the proximal tubule; and fibrosis and
the presence of infiltrates in the interstitial component. Treatment with CaNa2EDTA slowed the
progression of most alterations (Figure 5-16) and resulted in a diminution in nuclear inclusion
bodies. ALAD activity was reduced from 3.18 ± 0.52 U/mL in controls, to 0.82 ± 0.16 U/mL in
the Pb-exposed rats. In the rats treated with CaNa2EDTA, ALAD returned to near control levels
(2.98 ± 0.41 U/mL) at 137 days. It is surprising that such remarkable vascular changes were
noted in this study, while none were noted in Khalil-Manesh et al. (1992a), even with high-dose
Pb for longer periods of time. The kidney content of Pb (mean 74.6 |ig/g) was also lower than
the mean kidney content at 12 months (294 |ig/g) in the Khalil-Manesh et al. (1992a) study.
The only explanation for these striking differences that can be offered is that different strains of
rats were employed, i.e., Wistar in the Sanchez-Fructuoso (2002b) study and Sprague-Dawley in
the Khalil-Manesh et al. (1992a) study. The presence or absence of hypertension cannot be
invoked as an explanation, because in another Khalil-Manesh et al. (1993a) study the low-dose
Pb animals became hypertensive whereas the high-dose animals did not. These and other related
studies are summarized in Table AX5-7.1.

5.7.4.3 Biochemical Mechanisms of Lead Toxicity
Role of Free Radicals (Reactive Oxygen Species)
Since the early 1990s, it has been appreciated that free radicals, now known as reactive
oxygen species (ROS), are involved in the manifestations of Pb poisoning, presumably via their
adverse effects on tissue integrity and/or their vasoconstrictive effects on vascular endothelium.
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100
90
III 125 139 125 139 Days
No (real men t EOT A
Figure 5-15. Percentage of moderate and severe hypertrophy and vacuolization lesions in
small and medium sized arteries in the kidney of lead-exposed rats.
Source: Sanchez-Fructuoso et al. (2002b), with permission.
100

60
60 00 111 125 B9 111 125 139 Days
No treatment EDTA

Figure 5-16. Percentage of moderate and severe muscular hypertrophy lesions in
arterioles of the kidney in lead-exposed rats.
Source: Sanchez-Fructuoso et al. (2002b), with permission.
Wolin (2000) produced an extensive review of individual ROS, and their interactions with
NO, the major endogenous vasodilator, which acts via a second messenger, cGMP. The
production of ROS often begins with a one-electron reduction of molecular oxygen to superoxide
anion (Oi) by various oxidases. NAD(P)H oxidases are the principal enzymes involved.
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Superoxide anion is a negatively charged free radical that can be broken down to hydrogen
peroxide (H2O2) by superoxide dismutase (SOD) or can interact with NO to form the highly
reactive peroxynitrite ion (ONOO ), which, because of its extremely short half-life, is measured
as its reaction product, tissue nitrotyrosine. Catalase and glutathione peroxidase (GSHPx)
metabolize H2O2 to Compound I and oxidized glutathione (GSSG), respectively, while
myeloperoxidase metabolizes H2O2 to hypochlorous acid (HOC1). The reaction of H2O2 with
ferrous ion results in the formation of hydroxyl ion (*OH). ROS can be scavenged by
endogenous thiols (e.g., GSH) or exogenous thiol, e.g., N-acetylcysteine (NAC). ROS can be
measured as the concentration of the lipid peroxidation product, malondialdehyde-thiobarbituric
acid (MDA-TBA) or by the more recently introduced F-2 isoprostanes.
Kumar and Das (1993) explored the involvement of ROS in the pathobiology of human
essential hypertension. They found that plasma levels of lipid peroxides were higher in subjects
with uncontrolled essential hypertension compared to normal controls. Angiotensin II, a potent
vasoconstrictor, was found to stimulate free radical generation in normal leukocytes, which was
thought to inactivate NO, and possibly prostacyclin, which can lead to increased peripheral
vascular resistance and hypertension.
Hermes-Lima et al. (1991) also explored the involvement of ROS in Pb poisoning. They
described the process of autoxidation of ALA in the presence or absence of iron complexes,
which yields free radicals. Free radicals are also produced by Pb-stimulated iron-dependent lipid
peroxidation, as determined by quantification of thiobarbituric acid-reactive species (TEARS).
Pereira et al. (1992) demonstrated that chronically ALA-treated rats (40 mg/kg body weight
every 2 days for 15 days) under swimming training reached fatigue significantly earlier than the
control group, as well as demonstrating decreased mitochondrial enzymatic activities. In vivo
prooxidant properties of ALA were also suggested by the observed increase of CuZnSOD in
brain, muscle, and liver of untrained rats submitted to chronic treatment with ALA.
Ercal et al. (1996) contrasted the effects of treatment with DMSA or NAC in Pb-exposed
C57BL/6 mice. Five weeks of Pb exposure was found to deplete GSH levels, increase GSSG,
and promote MDA production in both liver and brain samples. Glutathione levels increased and
GSSG and MDA levels decreased in groups of Pb-exposed mice that received 1 mmol/kg DMSA
or 5.5 mM/kg NAC for 7 days prior to sacrifice. Treatment with DMSA caused reduction in
blood, liver, and brain Pb levels consistent with its function as a chelating agent, while treatment
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with NAC did not reduce these Pb levels. However, NAC treatment reduced indices of oxidative
stress in both brain and liver samples. Blood Pb concentrations in controls were 0.5 ± 0.5 |ig/dL;
in Pb-treated mice, 36.5 ± 2.4 |ig/dL; in Pb + DMSA-treated mice, 13.7 ± 1.3 |ig/dL; and in Pb
+ NAC-treated mice, 36.0 ±3.5 |ig/dL. Thus, both DMSA and NAC acted as antioxidants,
presumably via their thiol groups, but only DMSA reduced Pb concentration.
Daggett et al. (1998) and Fowler et al. (2004) have explored the effects of Pb and Pb
mixed with cadmium and arsenic on oxidative stress in the rat kidney. Daggett et al. (1998)
found that a single injection of Pb failed to deplete GSH or alter MDA levels in the kidney
within 24 hours. All subunits of glutathione-S-transferase (GST), however, were increased,
apparently not as the result of oxidative stress. Fowler et al. (2004) reported preliminary studies
of oxidative stress produced at 30, 90, and 180 days by mixtures of Pb (Pb acetate at 25 ppm),
cadmium (cadmium chloride at 10 ppm), and arsenic (sodium arsenite at 5 ppm). These dosages
were at the lowest observed adverse effect levels (LOAEL). Kidney carbonyls (a marker of
protein oxidation) were increased at all time points in the combination group, but decreased at
90 days in individually administered metal groups. Kidney non-protein thiols (representing
glutathione) increased in all groups at 180 days, suggesting that the induction of glutathione or
metallothionine attenuated increases in oxidative stress.
Vaziri and co-workers (Gonick et al., 1997; Ding et al., 1998, 2000, 2001; Vaziri and
Ding, 2001; Vaziri et al., 1997, 1999a,b, 2000, 2001, 2003; Zhou et al., 2002; Ni et al., 2004)
have published a number of articles relating to the production of ROS and alterations in
enzymatic activities in Pb-induced hypertension. These were discussed in detail in Section 5.5
but are described briefly here. In the majority of studies, Pb-induced hypertension was
produced by the administration of Pb acetate, 100 ppm in drinking water, for 3 months to male
Sprague-Dawley rats. Early studies (Gonick et al., 1997) revealed that hypertension could occur
in the absence of changes in NO or cGMP but with an attendant rise in plasma and kidney
MDA-TBA, indicating an increase in ROS. In a second study, Ding et al. (1998) showed that
infusion of arginine, the precursor of NO, or DMSA, a thiol Pb chelator and antioxidant, reduced
blood pressure to or towards normal, while simultaneously increasing depressed urinary NO and
decreasing an elevated MDA-TBA. Ding et al. (2000, 2001) further showed that the ROS
species, *OH, measured as salicylate-trapped 2,3 dihydroxybutyric acid, was increased in plasma
and cultured rat aortic endothelial cells after exposure to Pb, and that dimethylthiourea, a reputed
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scavenger of *OH, returned blood pressure, MDA-TB A, 'OH, and nitrotyrosine to or towards
normal. Ni et al., in 2004, demonstrated in both human coronary endothelial (EC) and vascular
smooth muscle cells (VSMC) that Pb acetate also increased superoxide (demonstrated by flow
cytometry using hydroethidine) and H2O2 (demonstrated with dihydrorhodamine) production.
After long-term (60-h) exposure, detectable superoxide levels fell to near normal while H2O2
production remained high.
Vaziri et al. (1997) showed that lazaroids, a class of non-thiol antioxidant, also restored
blood pressure, NO, and MDA-TB A to normal. Vaziri et al. (1999a) studied rats treated for
12 weeks with either Pb acetate alone or Pb acetate + vitamin E-fortified food (5000 units/kg rat
chow). They measured urinary excretions of stable NO metabolites (NOX) and plasma and tissue
abundance of nitrotyrosine, the footprint of NO oxidation by ROS. The Pb-treated group showed
a marked rise in blood pressure; a significant increase in plasma and kidney, heart, liver, and
brain nitrotyrosine abundance; and a substantial fall in urinary NOX excretion. Concomitant
administration of high-dose vitamin E ameliorated hypertension and normalized both urinary
NOX excretion and tissue nitrotyrosine without altering tissue Pb content. Vaziri et al. (1999b)
also measured eNOS and iNOS in the aorta and kidney of Pb-treated and Pb + vitamin E-treated
rats. Lead treatment increased both isotypes in aorta and kidney, signifying increased NO
production, while Pb + vitamin E lowered aortic, but not kidney, expression of eNOS and iNOS.
Vaziri and Ding (2001) tested the effect of Pb, 1 ppm, on cultured human EC cells. Lead was
tested alone or with either the SOD-mimetic agent, tempol, or a potent antioxidant lazaroid
compound (both at 10"8 or 10"7mol/L) on eNOS expression and NO production. Lead-treated
cells showed a significant upregulation of endothelial eNOS, increase in protein abundance, and
increase in the production of NO metabolites. Treatment with either tempol or lazaroids
abrogated the Pb-induced upregulation of eNOS protein and NOX production. Vaziri et al.
(2001) also studied increases in NOS isoforms in vivo in Pb-induced hypertension and reversal
by tempol. Both eNOS and iNOS were increased in kidney, aorta, and heart, while NOS was
increased in cerebral cortex and brain stem, of Pb-treated rats; blood pressure and NOS isoforms
were returned to normal by tempol. Vaziri et al. (2003) determined whether the oxidative stress
in animals with Pb-induced hypertension is associated with dysregulation of the main antioxidant
enzymes (i.e., SOD, catalase, and GSHPx), or increases in the superoxide-producing enzyme
NAD(P)H oxidase. At the conclusion of the experiment, immunodetectable CuZnSOD,
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MnSOD, catalase, GSHPx, and the gp91phox subunit of NAD(P)H oxidase were measured by
Western analysis in the kidney, brain, and left ventricle of control and Pb-exposed rats. Lead
exposure resulted in a significant increase in kidney and brain CuZnSOD with a significant
increase in brain, and insignificant increase in kidney and heart, gp91phox. In contrast, MnSOD,
catalase, and GSHPx in the kidney, brain, and left ventricle were unchanged. Incubation with Pb
acetate did not alter SOD activity in vitro. Thus, animals with Pb-induced hypertension
exhibited oxidative stress, which was associated with mild upregulation of the superoxide-
generating enzyme NAD(P)H oxidase, with no evidence of quantitative SOD, catalase, or
GSHPx deficiencies.
Vaziri et al. (2000) demonstrated that induction of oxidative stress in normal animals
(by feeding the GSH synthase inhibitor, buthionine sulfoximine, 30 mmol/L in drinking water
for 2 weeks) led to an increase in blood pressure, a reduction of urinary NOX, a 3-fold decrease in
liver GSH, and an increase in nitrotyrosine in kidney, aorta, heart, liver and plasma. The
administration of vitamin E + ascorbic acid ameliorated hypertension and also mitigated
nitrotyrosine accumulation despite persistent GSH depletion. This experiment demonstrated the
importance of GSH in protecting against the adverse effects of ROS accumulation in normal
animals. The majority of the studies reported by Vaziri and co-workers indicated that low Pb
exposure induced hypertension to be primarily mediated by ROS-induced depletion of NO.
NO production, on the other hand, is stimulated, as shown by the increase in eNOS and iNOS.
Enzymatic control of ROS levels by low Pb is achieved by upregulation of NAD(P)H oxidase
with no decrease in SOD, catalase, or GSHPx, i.e., the enzymes that breakdown ROS.
Scavengers of ROS ameliorate the elevated blood pressure, while the depletion of the
endogenous methyl scavenger, GSH, increases blood pressure in normal animals. No studies to
date have addressed the question of why high-dose Pb administration does not lead to
hypertension.
Farmand et al. (2005) pursued enzymatic studies by activity measurements and measures
of protein abundance in the rat kidney and aorta when rats are fed Pb acetate 100 ppm for
12 weeks. They demonstrated that the activities of CuZnSOD and catalase were increased by Pb
administration in renal cortex and medulla, whereas GSHPx was unchanged. In the thoracic
aorta, Pb exposure resulted in significant upregulation of CuZnSOD activity, while catalase and
GSHPx activities were unchanged, CuZnSOD, MnSOD, and catalase protein abundance were
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likewise unchanged. However, guanylate cyclase protein abundance in the thoracic aorta was
decreased. The authors suggested that the Pb-induced compensatory upregulation of CuZnSOD
and catalase and the decrease in aortic guanylate cyclase may be related to Pb-induced
hypertension.
Gurer et al. (1999a) evaluated whether captopril, an ACE inhibitor, acted as an
antioxidant in Pb-exposed F344 rats. Lead acetate was given in drinking water for 6 weeks.
Group I were the controls; group II received 1100 ppm Pb for 5 weeks and plain water during the
week 6; group III received 1100 ppm Pb for 5 weeks and, during the week 6, received water
containing captopril (10 mg/day). Blood Pb concentrations in the control group measured
0.8 ng/dL; in the Pb treated group, 24.6 ± 20 ng/dL; and in the Pb + captopril group, 23.8 ±
1.6 |ig/dL. MDA concentrations in liver, brain, and kidney were increased by Pb administration
and reduced to or towards normal by the Pb + captopril treatment. GSH concentrations were
decreased by Pb administration and restored by Pb + captopril treatment, whereas GSSG
concentrations were increased by Pb administration and reduced by Pb + captopril treatment.
Thus, this study showed that captopril was capable of augmenting the reducing capacity of the
cells by increasing GSH/GSSG ratios without affecting blood Pb concentrations.
McGowan and Donaldson (1987) examined total nonprotein sulfhydryl and GSH
concentrations in liver and kidney as well as GSH-related free amino acid concentrations in liver,
kidney, and plasma in 3-week-old Pb-treated (2000 ppm dietary lead) chicks. Cysteine,
converted from methionine, is the rate-limiting amino acid in GSH formation. The availability
of glutamate, cysteine, and glycine becomes important in the restoration of depleted GSH.
GSH, nonprotein sulfhydryl groups, glycine, and methionine were increased versus controls in
the liver, but only nonprotein sulfhydryl, glycine, cysteine, and cystathionine increased in the
kidney. Plasma levels of cysteine, taurine, and cystathione were reduced. Thus, Pb, for short
periods of time, increases GSH turnover. These and other related studies are summarized in
Table AX5-7.2.

Effect of Lead on Selective Renal Enzyme Levels
Effects of Lead on Renal NAG
Dehpour et al. (1999) studied NAG release by the rat kidney perfused with Pb acetate at
10, 20, and 50 jig/dL for 120 min, or Pb + arginine (the substrate for NO), or Pb + L-NAME
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(an inhibitor of NOS). Lead acetate caused a time and concentration-dependent increase in
enzymuria. Addition of arginine decreased, while addition of L-NAME increased, Pb-induced
NAG release. Histologic studies showed damage to some of the proximal tubule epithelial cells
in rats treated with 50 |ig/dL Pb acetate, damage that which was increased further by the addition
of L-NAME.

Effect of Lead on Renal GST
Two studies have evaluated the effects of Pb administration on GST isoforms in
developing rat kidney. In the first study (Moser et al., 1995), rats were treated either acutely
(14- and 50-day old rats given three daily injections of Pb acetate, 114mg/kg) or chronically
(Pb levels of 0, 50, 250, and 500 ppm in drinking water for 1, 2, 3, 4, and 7 weeks postnatal).
Chronic treatment rats were also given a 0.66% low calcium diet or standard rat chow.
Essentially all kidney cytosolic GSTs (Ybl, Yb2, Yp, Ycl, Yl, Yb3, Yal, Ya2, Yk) increased in
the acute experiment (1.1- to 6.0-fold). In the chronic experiment, all but one isoform (Yb3)
increased, and these results were markedly exacerbated by placing the rats on a low-calcium diet
(Ybl and Yp increased >25-fold). In the second study (Oberley et al., 1995), pregnant rats were
given 250 ppm Pb from conception until weaning, then pups received 500 ppm from weaning
until termination at either 3 or 7 weeks of age. By 7 weeks, proximal tubular cells showed
intranuclear inclusions, tubular injury, and interstitial fibrosis. Creatinine clearances were
reduced (0.55 + 0.05 versus 1.05 + 0.07 mL/min/lOOg; P < 0.001). Treatment with Pb also
caused large increases in the immunoreactive protein of Yc, Yk, Ybl, and Yp GST subunits in
proximal tubules but did not increase in the antioxidant enzymes CuZnSOD, catalase, and
GSHPx.
Another experiment that examined the effect of an acute dose of Pb as Pb nitrate
(100 |imol/kg IV) on GST levels in rat liver and kidney was reported by Planas-Bohne and
Elizade (1992). Seventy hours after injection, there was a marked increase in GST activity in
both organs, accompanied by induction of the isoenzyme GST 7-7 in the liver.
The relationship between GST induction by acute exposure to Pb acetate and oxidative
stress was explored by Daggett et al. (1998). Rats in the 72-h and 7-day experimental groups
received three consecutive daily injections of 114 mg/kg body weight of Pb acetate. The level of
kidney GST was increased at 3, 6, 12, and 24 h after injection, but MDA levels remained
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unchanged. Immunohistochemical markers of oxidative stress and NO production (MnSOD,
eNOS, iNOS, and 4-hydroxy-2-nonenal) also did not change. The authors concluded that the
GST changes were not the result of oxidative stress.
Witzman et al. (1998) and Kanitz et al. (1999) utilized two-dimensional (2-D) gel
electrophoresis to explore protein markers of Pb exposure. Witzman et al. (1998) gave three
consecutive IP injections of Pb acetate (114 mg/kg) to Sprague-Dawley rats, sacrificed them on
the fourth day, and subjected the cytosolic fraction of kidney homogenate to 2-D gel
electrophoresis. Lead exposure caused detectable inductions in both GSTP1 and GSTM1 and
caused quantifiable charge modifications in GSTP1. Kanitz et al. (1999) examined kidney
protein expression in male rabbits injected with Pb acetate (260, 360, or 100 |ig/kg) designed to
produce blood Pb levels of 20, 40, or 80 |ig/dL. Injections were given during weeks 6 to 10,
followed by maintenance doses during study weeks 11 to 20. Kidney homogenates were
subjected to 2-D electrophoresis. Significant quantitative changes occurred in 12 proteins in a
dose-related manner. Four proteins cross-reacted with anti-rat GSTpl (rc-GST). Thus, both
studies confirmed GST induction by Pb.
Daggett et al. (1997) examined the effects of triethyl Pb administration on the expression
on GST isoenzymes and quinone reductase in rat kidney and liver. Fischer 344 rats were given
one IP injection of triethyl Pb chloride (10 mg/kg body weight) and subsequent changes in
enzyme expression were measured. There was a significant increase in GST activity in kidney;
and all GST subunits were significantly elevated, the largest increase being a 3.2-fold increase in
GST Ybl. In the liver, injection of triethyl Pb chloride resulted in decreased GST activity.
The largest decrease in subunits was a 40% reduction in GST Yal. The activity of quinone
reductase was elevated 1.5-fold in kidney and 2.7-fold in liver within 14 days after the injection
of triethyl Pb chloride.

Effects of Lead on Renal Heme Enzymes
Vij et al. (1998) explored Pb-induced alterations in male rats in the heme synthesizing
enzymes, ALAD and uroporphyrinogen I synthetase, and also the effect of ascorbic acid
supplementation in reversing these alterations. Lead-treated rats were injected IP with 20 mg/kg
of Pb acetate for 3 consecutive days and sacrificed 4 days later. A separate group of animals was
administered 100 mg/kg ascorbic acid PO for 3 days following Pb administration. Blood Pb
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concentration was 4.7 ±1.5 |ig/dL in control rats, 16.6 ± 4.7 |ig/dL in Pb-treated rats, and
7.8 ± 2.0 |ig/dL in the Pb + ascorbic acid treated rats. Lead content of liver and kidney followed
the same pattern. Blood ALAD activity was diminished in the Pb-treated rats but was restored in
the Pb + ascorbic acid-treated rats. Uroporphyrinogen I synthetase activity followed the same
pattern in blood but was not restored by ascorbic acid in liver. Total and nonprotein sulfhydryl
concentrations in blood were depressed by Pb administration and were not restored by ascorbic
acid. However, levels in liver and kidney were restored by ascorbic acid.
ALAD levels following administration of Pb were also investigated by Rodrigues et al.
(1996a) and Peixoto et al. (2004). The study by Rodrigues et al. (1996a) examined rats from
Pb-exposed mothers that were maintained after weaning on either 0.5 or 4.0 mM Pb acetate in
drinking water for 21 days or 6 months. At sacrifice, ALAD activity was measured in kidney,
forebrain, and cerebellum. Both 6-month-old Pb-exposed groups showed an increase in the
kidney-to-body weight ratio, suggesting Pb-induced cell proliferation in the kidney. Blood Pb
increased from 6.5 to 7.6 |ig/dL in the 21-day-old exposed rats compared to 6-month-old
controls. In the 0.5 mM Pb-treated group, blood Pb was 9.8 |ig/dL in the 21-day-old and
41.6 |ig/dL in 6-month-old rats, while in the 4.0 mM group, blood Pb was 44.4 |ig/dL in the
21-day-old and 116.9 |ig/dL in the 6-month-old group. ALAD activity was reduced at 6 months
in the forebrain of the 4.0 mM Pb-treated group, and in the kidneys at 6 months in both the
0.5 mM and 4.0 mM Pb-treated groups. The study by Peixoto et al. (2004) examined the in vitro
sensitivity (ICso) to Pb of ALAD activity of brain, kidneys, and liver from suckling rats aged
between 1 and 5, 8 and 13, or 17 and 21 days. The metal concentrations ranged from 0 to 50 jiM
for Pb acetate. Rats in the first age group showed the greatest sensitivity in all three organs.
Liver was the least sensitive to ALAD inhibition by Pb, while brain was the most sensitive.

Effects of Lead on NaK-A TPase
Fox et al. (1991b) explored the effect of in vivo Pb exposure on adult rat retinal and
kidney NaK-ATPase. Pups, exposed to Pb through the milk of dams consuming 0, 0.02, or
0.2% Pb solutions, had mean blood Pb concentrations of 1.2, 18.8, and 59.4 |ig/dL at weaning,
respectively, and 5 to 7 |ig/dL as 90 to 100-day-old adults. Prior Pb exposure produced
significant dose-dependent decreases in isolated retinal NaK-ATPase activity (-11%; -26%),
whereas activity in the kidney was unchanged. In contrast, NaK-ATPase from both isolated
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control tissues was inhibited by Pb in vitro. The half-maximal inhibitory dose of Pb for retinal
and renal NaK-ATPase was 5.2 x ICf7 and 1.3 x ICf5 M, respectively. Retinal and renal
NaK-ATPase were 20-fold and 1.1-fold more sensitive to inhibition by Pb than calcium.
The increased sensitivity of retinal, compared to renal, NaK-ATPase to inhibition following
in vivo or in vitro Pb exposure may be related to their different a subunit composition.
Kramer et al. (1986) had also explored the half-maximal inhibitory dose for Pb chloride
on renal cortical homogenate NaK-ATPase, and found it to be 7 x ICf5 M. There was a
competitive inhibition with regard to the substrate, ATP. Of several metals tested, Pb was
second only to Hg in potency as a NaK-ATPase inhibitor.
Weiler et al. (1990) studied the effect of Pb on the kinetics of purified (from hog cerebral
cortex) NaK-ATPase and potassium-stimulated p-nitrophenylphosphatase (K-pNPPase), which is
referred to as the E2 configuration of the NaK-ATPase system. ICso for Pb was found to be
8.0 x 10~5 M for NaK-ATPase and 5.0 x 10~6 M for K-pNPPase. Inhibition of NaK-ATPase by
Pb was found to be noncompetitive with respect to K, but competitive with respect to Na and
MgATP. Inhibition of K-pNPPase by Pb was competitive with respect to K.

Effects of Lead on Cardiovascular Hormones
Effects of Lead on Endothelin
Khalil-Manesh et al. (1993a) examined the role of endothelial factors in Pb-induced
hypertension. They found that low Pb administration (0.01%), but not high Pb administration,
(0.5%) resulted in increased blood pressure in rats treated for 12 months. In the low-Pb-treated
rats, measurement of plasma endothelins-1 and -3 revealed that endothelin-3 concentration
increased significantly after both 3 months (Pb, 92.1 ± 9.7 versus control, 46.7 ± 12.0 pmol/mL;
p < 0.001) and 12 months (Pb, 105.0 ± 9.3 versus control, 94.1 ± 5.0 pmol/mL; p < 0.01), while
endothelin-1 was unaffected. Plasma and urinary cyclic GMP concentrations, as a reflection of
endothelium-derived relaxing factor (EDRF), decreased significantly at 3 months (plasma Pb,
1.8 ± 0.9 versus control, 4.2 ± 1.6 pmol/mL; p < 0.001) and 12 months (plasma Pb 2.2 ± 0.7
versus control, 4.2 ± 0.9 pmol/mL; p < 0.001). High levels of Pb exposure did not result in
hypertension, perhaps related to the fact that plasma concentrations of endothelin-1,
endothelin-3, and cyclic GMP were unaltered at 3 months, while their concentrations were
significantly decreased at 12 months (plasma cyclic GMP at 12 months, 2.2 ± 0.7, Pb, versus
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4.2 ± 0.9 pmol/mL, control; p < 0.001). Thus, the path to development of hypertension in low-
Pb rats was thought to be through an increase in the concentration of the vasoconstrictor,
endothelin-3, and a decrease in the vasodilator hormone, endothelium-derived relaxing factor or
NO.
Novak and Banks (1995) studied the effects of Pb on the actions of endothelin. They
measured renal clearances and mean arterial pressure in rats in which endothelin-1 was infused at
110 ng/kg/min for 30 min. Lead was infused as Pb acetate throughout the experiment at 0.48,
4.8, and 24 nmoles/min. At the two higher doses, Pb significantly attenuated the endothelin-
induced increase in mean arterial pressure; Pb infused as 0.48 nmoles/min had no effect.
An endothelin-induced decrease in GFR in control rats was completely blocked at the higher
doses of Pb. In additional experiments, calcium chloride was infused at 500 nmoles/min for
105 min, and then calcium + Pb (4.8 nmoles/min) were infused for another 105 min. In these
experiments, there was no Pb-induced inhibition of the mean arterial pressure response to
endothelin. However, the GFR response to the peptide remained blocked. These data illustrate
that Pb inhibits the cardiorenal actions of endothelin and that a calcium-related process is
involved in the systemic, but not the renal, component of this inhibition.

Effects of Lead on the Catecholamine System
Carmignani et al. (2000) studied the effects of 10 months of low Pb exposure (60 ppm of
Pb acetate), on catecholamine and monoaminoxidase (MAO) levels. Plasma catecholamines
were measured by HPLC and MAO in aorta, liver, heart, kidney, and brain by a histochemical
technique. Plasma norepinephrine (NE) increased by 104% and adrenaline by 81%, with no
changes noted in L-DOPA and dopamine levels. MAO activity was increased in all organs.
These workers ascribed the low Pb-induced hypertension in part to raised catecholamines levels.
Tsao et al. (2000) and Chang et al. (2005) measured changes in the P-adrenergic system in
Wistar rats during and following Pb exposure. In Tsao et al. (2000), rats were chronically fed
with 0.01, 0.05, 0.1, 0.5, 1.0, and 2.0% Pb acetate for 2 months. Plasma catecholamine levels
were measured by HPLC; cAMP levels in heart, kidney, and aorta by radioimmunoassay; and
P-adrenergic receptors in heart, kidney, and aorta membranes by a radio ligand binding assay.
Blood Pb increased from 0.05 ± 0.05 |ig/dL in controls to 85.8 ± 4.1 |ig/dL in the
2.0% Pb-treated group. Plasma NE, but not E, levels increased with increasing Pb dosage.
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P-adrenoreceptor density of heart and kidney decreased progressively with increasing Pb dosage,
whereas kidney P-adrenoreceptor density increased up to the 0.5% Pb group and then remained
constant. Unstimulated cAMP was constant in all tissues, but cAMP stimulated by isoprotorenol
was lowered progressively in aorta and heart and increased in kidney. Chang et al. (2005)
continued these measurements in rats fed 2% Pb acetate for 2 months then withdrawn from Pb
for periods of 1, 2, 3, 4, 5, 6, and 7 months. Blood Pb levels, systolic and diastolic blood
pressure levels, and plasma NE were reduced after cessation of Pb exposure. This occurred in
conjunction with an increase in P-adrenoreceptor density in heart and aorta and a decrease in
P-adrenoreceptor density in kidney. (See Table AX5-5.5 for details on these studies).

Effects ofChelators (Single or Combined) on Lead Mobilization
These studies are summarized in Tables AX5-7.3 and AX5-7.4. For the sake of brevity,
they will not be discussed further here.

Effects of Other Metals on Lead Distribution
Lead and Calcium
Fullmer (1992) published a review of intestinal interactions of Pb and calcium. High
affinity Pb binding to intracellular calcium receptors and transport proteins, as well as the
involvement of Pb in calcium-activated and calcium-regulating processes, are thought to provide
a partial molecular basis for the cellular and systemic effects of Pb.
Maidonado-Vega et al. (1996) examined the intestinal absorption of Pb and bone
mobilization during lactation. All experiments were started with 3-week-old female Wistar rats.
Rats were impregnated at 16 weeks and were fed a 100 ppm solution of Pb acetate for 158 or
144 days (mid-lactation or before lactation). Rats were also exposed for only 14 days, from
144 to 158 days (i.e., only during lactation). Nonpregnant rats from the same litter were exposed
to Pb for periods equivalent to each of these groups. In the nonpregnant rats, blood Pb increased
to 27.3 |ig/dL from 5.2 |ig/dL in controls. Similarly, kidney Pb increased to 13.2 nmol/g from
0.5 nmol/g, and bone Pb increased to 88.9 nmol/g from 0.9 nmol/g. ALAD activity decreased to
410 nmol/h/mL from 1004 nmol/h/mL. Compared to nonpregnant rats, there was a moderate
increase in blood Pb in the lactating animals whether the Pb was given to mid-lactation or up to
the period before lactation. Similarly, when Pb was administered only during lactation, there
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was a much higher increase in blood Pb in the pregnant rats than in the nonpregnant rats. Bone
Pb concentration increased when Pb was given only during lactation, whereas bone Pb decreased
(compared to Pb-treated nonpregnant rats) when the Pb was given either before lactation or
before and during lactation. The authors considered that resorption of Pb from bone was the
main additional source of Pb during lactation. The data indicate that Pb stored in bone as a result
of prior maternal exposure should be considered as a major source of self intoxication and of Pb
in milk available to suckling pups.

Lead and Cadmium
Skoczyriska et al. (1994) compared the effects of the combined exposure to Pb and
cadmium to each metal singly on tissue composition of trace metals. Experiments were
performed on 5- to 6-week-old male Buffalo rats given Pb acetate (70 mg Pb/kg body weight
twice a week) and cadmium chlorate (20 mg Cd/kg body weight once a week) intragastrically for
7 weeks either singly or in combination. Blood Pb in the control group was 5.1 |ig/dL, compared
to 29.6 |ig/dL in the Pb-treated group. In contrast, the Pb + cadmium group showed a blood Pb
of 37.4 |ig/dL. After combined exposure to Pb and cadmium, the level of these metals in the
liver and kidney was lower than after the single administration of Pb or cadmium. Exposure of
the rats to cadmium resulted in an increase of kidney zinc and copper and liver zinc levels;
combined exposure to Pb + cadmium did not produce more extensive changes in tissue zinc and
copper concentrations.

Lead and Selenium
Othman and El Missiry (1998) examined the effect of selenium against Pb toxicity in
male rats. Male albino rats were given a single dose of Pb acetate (100 jimol/kg body weight)
and sacrificed 3 or 24 h later. Another group of animals was pretreated with sodium selenite
(10 jimol/kg body weight) 2 h before receiving Pb acetate and sacrificed 24 h later. Selenium is
well known as an antioxidant and cofactor for GSHPx. In this experiment, GSH content,
GSHPx, SOD activities, and the products of lipid peroxidation (i.e., TEARS) were determined.
It was found that lipid peroxidation was prevented and the reduction in GSH caused by Pb in
liver and kidney was diminished by selenium. Lead-induced diminution in SOD activity and
GSHPx activity was also returned to normal by selenium.
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Tandon et al. (1992) studied the effect of selenium supplementation during chelation of
Pb with CaNa2EDTA. Rats were given Pb acetate 10 mg/kg/day by gastric gavage for 6 weeks.
This was followed by a 5-day treatment course of CaNa2EDTA, 0.3 mmol/kg IP or of
CaNa2EDTA + sodium selenite, 0.5 mg/kg PO. Selenium had marginal effects on Pb removal
by CaNa2EDTA in blood, liver, and kidney and similar effects on ALAD activity.

Lead and Zinc
Flora et al. (1989) examined the role of thiamine, zinc, or their combination in the
prevention or therapy of Pb intoxication. Albino rats received the following treatments daily
through gastric gavage for 6 days each week over a six-week period: 10 mg/kg of Pb as Pb
acetate; or the same dose of Pb acetate + thiamine (25 mg/kg) zinc sulfate (25 mg/kg) or
Pb + thiamine and zinc. Rats exposed to Pb only were additionally divided into four groups
treated by gastric gavage daily for 6 days as follows: group I, water only; group II, thiamine
only; group III, zinc only; and group IV, combined zinc + thiamine. The activities of blood
ALAD, blood ZPP, blood Pb, and urine ALA were determined. Blood Pb concentrations
increased from 6.2 to 120.9 |ig/dL, contrasting normal controls with Pb-treated animals. There
was a slight reduction in blood Pb in animals treated with either thiamine or zinc and a greater
reduction in animals treated with thiamine + zinc. In the post-Pb-exposure treatment group,
thiamine + zinc was also the most effective treatment. Liver and kidney Pb levels followed the
same course, but brain Pb was not reduced by treatment. Blood ALAD activity was decreased
from a normal level of 7.63 (imol ALA/min/L to 0.69 in Pb-treated animals and restored to 7.52
in Pb + thiamine + zinc-treated rats. ZPP was increased from 1.78 |ig/g hemoglobin to 4.22 in
Pb-treated animals and reduced to 2.50 in Pb + thiamine + zinc-treated animals. Urine ALA
increased from 0.07 to 0.24 mg/dL in Pb-treated animals but decreased to 0.17 in the
Pb + thiamine + zinc-treated rats. Prevention was more effective than post-Pb-exposure
treatment. This was thought to be due mainly to the decrease in the absorption of Pb in the GI
tract in the presence of thiamine and/or zinc.
Flora et al. (1994) explored the dose-dependent effects of zinc supplementation during
chelation of Pb in rats. The chelator employed was CaNa2EDTA, whose toxic effects are known
to be mainly due to the depletion of endogenous zinc and, possibly, copper and manganese.
Male Wistar rats were started on exposure to Pb acetate, 10 mg/kg, administered through gastric
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gavage once daily for 56 days. Twenty-four hours later, the Pb-exposed animals were treated
daily for 5 days as indicated: group I, saline ; group II, CaNa2EDTA 0.3 mmol/kg, IP, once daily
for 5 days; group III, CaNa2EDTA + zinc sulfate, 10 mg/kg, PO once daily for 5 days; and group
IV, CaNa2EDTA + zinc sulfate, 50 mg/kg, PO once daily for 5 days. Blood ALAD decreased
from 6.30 to 1.44 nmol/min/mL erythrocyte in Pb-exposed animals, with no change after
CaNa2EDTA treatment and partial restoration after the CaNa2EDTA + zinc, 10 mg/kg treatment.
There was no improvement following zinc, 50 mg/kg. Blood Pb concentration increased from
4.6 |ig/dL to 43.0 |ig/dL in Pb exposed animals, decreased to 22.5 |ig/dL in CaNa2EDTA-treated
animals and decreased further to 16.5 |ig/dL in CaNa2EDTA plus zinc-treated animals. Zinc at
50 mg/kg led to an increase in blood Pb to 56.1 |ig/dL. Changes in the liver followed the same
pattern, whereas zinc increased the Pb levels further in the kidney, and zinc had no influence on
Pb content in the femur. Blood zinc decreased from 6.1 to 5.7 |ig/mL in Pb-exposed rats and
further to 5.0 |ig/mL in CaNa2EDTA-treated animals. There was an increase to levels of
6.6 |ig/mL on the 10 mg/kg supplement of zinc and a further increase to 8.1 |ig/mL on the
50 mg/kg zinc supplement.

Lead and Iron
Hashmi et al. (1989) examined the influence of dietary iron deficiency, Pb exposure, or
the combination of the two on the accumulation of Pb in vital organs of rats. Animals fed an iron
deficient diet for 2 weeks were also subjected to orbital plexus puncturing twice a week to allow
a Hb levels to decrease to 7 to 8 g/dL. Animals were thereafter treated for the next 6 weeks with
iron deficient diets or iron-deficient diets + 0.1% Pb acetate in drinking water. At the end of
3 and 6 weeks, animals from each group were sacrificed. Feeding of an iron-deficient diet
during Pb exposure enhanced the accumulation of Pb in soft tissues and flat bones. For example,
liver Pb content was 0.75 |ig/g in control animals, 8.43 in Pb treated animals, and 12.93 in iron-
deficient and Pb-treated animals. The sequence of events was similar in kidney, spleen, and
femur except that the Pb content in femur was reduced in the iron deficient and Pb-treated group.
Singh et al. (1991) conducted a study to ascertain the role of iron deficiency during
pregnancy in inducing fetal nephrotoxicity in mothers exposed to Pb. Rats were fed either a
normal iron diet or an iron free synthetic diet for 15 days, followed by a diet containing half of
the daily required iron (47 mg/100 g ferrous ammonium sulfate) for a further 15 days. Female
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animals were mated with healthy adult males. Lead doses of 250, 500, 1000, and 2000 ppm
were given in drinking water during pregnancy and lactation. Fetuses were removed by
Caesarean section on the 21st day. Maternal blood Pb levels in rats on an iron deficient diet
were higher than those in rats on a normal iron diet at all Pb dosing levels. Similarly, placental
Pb levels were higher in animals on an iron-deficient diet as compared to a normal diet. Lead
content in the fetuses were higher on the iron-deficient diet. Lead administration resulted in
dose-dependent hydropic degeneration of renal proximal tubular cells in the fetuses. At a dose of
2000 ppm Pb with iron deficiency, more Pb accumulated in maternal blood, placenta, and fetuses
and maximum pathological changes were seen in the fetal kidney as compared to other doses.

Lead and Aluminum
Shakoor et al. (2000) reported beneficial effects of aluminum on the progression of
Pb-induced nephropathy in rats. Male albino rats were treated with water only or Pb acetate
(125 mg/kg) and/or aluminum chloride (50 mg/kg or 100 mg/kg) for a period of 90 days.
Aluminum was found to prevent the Pb-induced increase in relative kidney weight in a dose-
dependent manner. Aluminum also prevented Pb-induced increases in plasma creatinine levels
of Pb-treated animals. The net deposition of Pb in kidneys was lower in animals that were given
both Pb acetate and aluminum chloride simultaneously. By day 90, plasma creatinine was
1.26 mg/dL in control animals, 1.88 mg/dL in Pb-treated animals, and 1.34 and 1.44 mg/dL in
Pb- and aluminum-treated animals. Similarly, kidney Pb increased from 5.4 |ig/g in control
animals to 220.0 jig/g in Pb-treated animals and decreased to 138.5 and 98.9 jig/g in Pb- and
aluminum-treated animals. These and other related studies are summarized in Table AX5-7.5.

Lead, Cadmium, and Arsenic
In their review of mechanisms of nephrotoxicity from metal combinations, Madden and
Fowler (2000) discuss the effects of Pb, cadmium, and arsenic combinations, given that such
combinations may be found in the industrial setting or at toxic dump sites. Cadmium has been
shown to interact with Pb, minimizing Pb kidney effects by lowering the renal Pb burden and
preventing the appearance of Pb inclusion bodies (Mahaffey et al., 1981). Thus, cadmium may
therefore affect the binding of Pb to Pb-binding protein (Mistry et al., 1985). Lead, cadmium,
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and arsenic combinations also increase the degree of porphyrinuria beyond that produced by Pb
alone (Fowler and Mahaffey, 1978).

5.7.4.4 Effect of Age on Lead Toxicity
Han et al. (1997) examined the hypothesis that the high rate of bone remodeling during
childhood and the consequent high calcium and Pb turnover would result in a substantial
reduction in bone Pb stores, so that much of the Pb incorporated in bone during childhood does
not persist into adulthood. They treated female Sprague-Dawley rats with 250 ppm of Pb in
drinking water for 5 weeks beginning at 5, 10, or 15 weeks of age. Organ harvesting occurred
4 weeks after the end of Pb exposure for all groups, as well as 8 and 20 weeks after cessation of
Pb ingestion in the rats exposed beginning at 5 weeks of age. Organs examined were brain,
kidney, liver, femur, and spinal column bone. Blood and organ Pb concentrations were
significantly higher in the rats exposed beginning at 5 weeks of age than in those exposed
beginning at 10 or 15 weeks of age. The results of this experiment rejected the hypothesis and
suggested instead that a younger age at Pb exposure is associated with greater Pb retention and
toxicity, even in the absence of continued Pb exposure.
Garcia and Corredor (2004) examined biochemical changes in the kidneys after perinatal
intoxication with Pb and/or cadmium. Lead acetate (300 ppm) and/or cadmium acetate (10 ppm)
were administered in drinking water to pregnant Wistar rats from day 1 of pregnancy to
parturition (day 0) or until weaning (day 21). The following kidney enzyme activities were
determined: alkaline and acid phosphatases, Mg-ATPase, and NaK-ATPase. Blood Pb was
measured in control pups as well as in pups exposed to Pb at parturition and at weaning. Control
pups showed 1.43 |ig/dL of blood Pb compared to 31.5 |ig/dL at day 0 and 22.8 |ig/dL at day 21
in pups exposed to Pb. In those rats receiving both cadmium and Pb, the blood Pb concentration
was 23.2 |ig/dL at day 0 and 13.2 |ig/dL at day 21. Lead caused a significant inhibition of
kidney alkaline phosphatase and kidney acid phosphatase. At parturition, Pb intoxication
produced a strong inhibition of NaK-ATPase (-80%) as well as of Mg-ATPase activities
(-24%); whereas, when Pb was given in combination with cadmium, these inhibitory effects
were attenuated. At weaning, Pb continued to produce a significant inhibition of Mg-ATPase but
had no effect on NaK-ATPase. Thus, simultaneous perinatal administration of both Pb and
cadmium seemed to protect against the toxicity produced by Pb separately.
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Cory-Slechta (1990b,c) published two articles on the effects of old age on the disposition
of Pb. In the study (1990c) male F344 rats, at the ages of 8 months (adult) and 16 months (old)
were exposed to concentrations of 0, 250, or 500 ppm Pb acetate in drinking water for 7 months.
At these Pb doses, prior studies had indicated that blood Pb levels ranged from 60 to 90 |ig/dL.
Blood Pb, ZPP, and urinary ALA levels were determined after both 3 and 7 months of exposure.
Organ weights, tissue Pb concentrations, and urinary excretion of Pb, calcium, copper, and zinc
were examined after 7 months of exposure. Tissue Pb distribution was markedly altered in old
rats: in bone and kidney, Pb levels were reduced while liver Pb was substantially increased.
Blood Pb levels in adult and old rats were comparable at both measurement intervals, as was
urinary Pb excretion at 7 months. Lead-induced elevation of ZPP exhibited differential changes
between 3 and 7 months; values in adults declined, while levels in old rats increased or remained
unchanged. In the adult group, Pb exposure increased calcium excretion primarily at the
500 ppm exposure level. In contrast, Pb exposure decreased urinary calcium excretion in old
animals at the higher exposure level. No effects of either age or Pb exposure were detected in
the comparison of adult versus old urinary excretion of zinc or copper.
In the second study, Cory-Slechta (1990b), young (21 days old), adult (8 months old), and
(16 months old) rats exposed to 0, 2, or 10 mg of Pb acetate/kg per day for 9.5 months were
evaluated. Differences in the tissue distribution of Pb with age included lower bone levels, but
increased concentrations in brain, liver, and kidney. Differences in blood Pb levels over the
course of exposure were not remarkable. Thus, these effects did not appear to reflect an
enhanced Pb absorption from the GI tract with age. Instead, the bone changes may reflect
enhanced bone resorption with a concurrent decline in bone apposition with age, combined with
altered patterns of urinary Pb excretion over time, i.e., elevated urinary Pb at 3 and 6 months, but
comparable Pb excretion at 9.5 months, as compared to young and adult rats.

5.7.5 Summary
Highlights of the previous 1986 Lead AQCD and of studies done between 1986 and the
present are outlined in this section.
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1986 Document

• In animal studies, nuclear inclusion bodies were found in proximal tubules, identified as
27 kDa or 32 kDa proteins in combination with Pb. Subsequently, a 63 kDa Pb-binding
cytosolic protein was described in kidney.

• Swollen mitochondria, with diminished mitochondrial function, were found in the
proximal tubules.

• Renal ALAD was the same in Pb-treated animals as in controls when GSH was present,
but was reduced when GSH was absent.
Newer studies

• Experimental studies have shown that early effects of Pb on tubular cells are generally
reversible, but with continued exposure, a chronic irreversible nephropathy is likely to
ensue.

• Hyperfiltration, when compared to age- and sex-matched normal controls, was found in
adults who had suffered from childhood Pb poisoning, in young occupationally exposed
Pb workers in Korea, and in both low-Pb-treated rats and high-Pb-treated rats up to
3 months of exposure. This is paralleled in animal experiments by an increase in
kidney weight.

• Various new urinary markers for Pb toxicity have been described. These include NAG,
p2-microglobulin, al-microglobulin, retinol binding protein, GST, lysozyme, y-glutamyl
transferase, alanine aminopeptidase, prostanoids, and brush border antigens. Information
on these markers is voluminous, but, on review, only GST and al-microglobulin seemed
to be appropriate urinary markers. NAG, which has been most extensively investigated,
appears in detailed-animal studies to be overly sensitive, increasing in low-Pb-treated
animals, despite an absence of pathological changes on ultrastructural study. Both
p2-Microglobulin, and possibly retinol binding protein, which are low-molecular weight
proteins reabsorbed by the proximal tubule, appeared to be elevated only with high blood
Pb levels (>80 jig/dL).

• Animal studies have implicated free radicals in the pathogenesis of Pb-induced
hypertension and renal disease. A sequence of free radicals can be demonstrated in
Pb-induced disease, as evidenced by an increase in superoxide radicals, hydroxyl
radicals, hydrogen peroxide, and peroxynitrite, together with a diminution in GSH in
liver, brain, and aorta. Nitric oxide is most commonly decreased (by free radicals) as is
urinary cyclic GMP. Aortic guanylate cyclase is decreased. The enzyme responsible for
an increase in the production of free radicals, NAD(P)H oxidase, is increased by Pb,
whereas eNOS and iNOS, the enzymes involved in the production of nitric oxide, are also
increased, attesting to the importance of free radical destruction of nitric oxide.
Antioxidants reverse these changes and diminish blood pressure.
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Norepinephrine and epinephrine are increased by Pb administration, whereas
P-adrenoreceptor density of heart and kidney are decreased. In a second study,
norepinephrine, but not epinephrine, was increased by Pb.
Various antioxidants have been used in conjunction with chelators, to both remove Pb
from tissue and to diminish free radicals. Taurine, lipoic acid, arginine, ascorbic acid,
vitamin E, thiamine, tempol, and lazaroids have been used in conjunction with DMSA,
all improving free radical diminution.
Metal combinations have also been employed to reduce tissue Pb and/or affect free
radicals. Cadmium increases Pb in blood when both are given, but diminishes Pb in liver
and kidney. Selenium, an antioxidant, improves both parameters, as does thiamine or
L-lysine plus zinc. Iron deficiency increases intestinal absorption of Pb and the Pb
content of soft tissues and bone. Aluminum decreases kidney Pb content and serum
creatinine in Pb-intoxicated animals.
Age also has an effect on Pb retention. There is higher Pb retention at a very young age
and lower bone and kidney Pb at old age, attributed in part to increased bone resorption
and decreased bone accretion.
5.8 EFFECTS ON BONE AND TEETH
5.8.1 Biology of Bone and Bone Cells
By weight, bone is composed of 28% collagen fibers (predominantly type I collagen) and
5% noncollagenous proteins (osteocalcin, osteonectin, and other proteoglycans), with crystals of
hydroxyapatite [Caio(PO4)e(OH)2] making up the remaining 67%. In addition to providing
mechanical support for the body and protection of vital organs, the skeletal system also functions
in a metabolic capacity. Historically, bones have been classified as either long or flat based on
their appearance, with long bones including limb bones, e.g., the femur and humerus, and flat
bones including the bones of the skull, sternum, pelvis, and scapula. Long and flat bones
originate by distinct methods of formation, endochondral and intramembranous, respectively,
with long bones eventually using both processes. In endochondral bone formation, a
mineralized, cartilaginous matrix precedes the transition into true bone, while in
intramembranous formation, the bone forming cells create bone directly without the cartilaginous
template.
Bone cells responsible for producing the bone matrix of collagen and ground substance
are called osteoblasts. Several signaling factors including growth factors and hormones
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influence pre-osteoblastic cells to differentiate into mature osteoblasts and subsequently
synthesize and mineralize the extracellular matrix to form mature bone. It is during the process
of bone mineralization that the Pb ion (Pb2+) can become incorporated by substituting for the
calcium ion (Ca2+). The bone cells responsible for bone resorption are the osteoclasts.
Osteoclasts, which are large and multicellular (4 to 20 cells), dissolve bone matrix and
hydroxyapatite by synthesizing and releasing lysosomal enzymes and acidifying the extracellular
surroundings. It is during the process of dissolving bone, or demineralization that Pb stored in
bone can be released locally and into the general system.
Bone cell function may be compromised both directly and indirectly by exposure to Pb.
Regulation of bone cells occurs by numerous local and systemic factors, including growth
hormone (GH), epidermal growth factor (EOF), transforming growth factor-beta I(TGF-PI), and
parathyroid hormone-related protein (PTHrP). As discussed further below in this section, the
presence of Pb can potentially interfere with each of these factors. The bones of the skeleton
serve as the primary reservoir for calcium and phosphate in the body and help to maintain
homeostasis of these ions in the serum through bone turnover or remodeling. Vitamin D
[1,25-(OH)2D3] maintains the normal range of calcium in the serum by increasing the efficiency
of calcium absorption in the intestines and facilitating differentiation of stem cells into
osteoclasts, which break down bone and mobilize calcium (and Pb) stores. Parathyroid hormone
(PTH), in turn, regulates the production of vitamin D in the kidney. Lead has been shown to
interfere with the action of both of these hormones. Other substances influenced by Pb and
discussed in this section are alkaline phosphatase, an enzyme necessary for mineralization of
bones and teeth, and osteocalcin, a noncollagenous protein whose spatial and temporal pattern of
expression suggests a role in bone mineralization. Both substances are also markers for
osteoblast activity and, by default, bone formation. Alkaline phosphatase is a potential carrier of
ionic calcium and is capable of hydrolyzing inhibitors of mineral deposition such as
pyrophosphates.

5.8.2 Summary of Information Presented in the 1986 Lead AQCD
Lead has been shown to become localized and accumulate in bones and teeth, with
accumulation beginning as early as fetal development. Lead administered to rats as a single dose
results in blood Pb concentrations that are initially elevated, but rapidly fall as Pb is transferred
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to bone or excreted. The dose of Pb administered does not apparently affect distribution to the
various body compartments; however, the rate-limiting step in the clearance of Pb from rats and
mice involves absorption into/clearance from the skeletal system. The loss of Pb from various
organs and tissues follows first-order kinetics, except from bone. More absorbed Pb is retained
by young animals compared with adult animals, leading to higher tissue levels. Moreover, once
Pb is incorporated into the young animal's body, the long-term rate of retention is greater than
that of adults. In Pb-exposed animals, Pb is distributed subcellularly, preferentially to the
nucleus and mitochondrial fractions.
During lactation in mice, a redistribution of tissue Pb occurs (mobilization), resulting in
the transfer of Pb and calcium from mother to pups via the milk and subsequent overall loss of
Pb in the mothers. Lead transfer to suckling rats via mother's milk has been reported to be
-3% of the maternal body burden or more, if Pb exposure continues during lactation. Eight days
after a single injection of Pb, the content of Pb in rabbit's milk was 8-fold higher than the
maternal blood level, suggesting Pb transfer can occur against a concentration gradient.
Transplacental transfer of Pb from mother to fetus also occurs in various animals.
In rats, a significant reduction of calcium in the diet leads to enhanced uptake of Pb into
the bones and other tissues. In general, an enhanced uptake of Pb into tissues is also seen in rats
fed diets deficient in iron, zinc, copper, or phosphorus, and in the presence of low or excess
vitamin D.

5.8.3 Bone Growth in Lead-Exposed Animals
Lead is readily taken up and stored in the bone of experimental animals, where it can
potentially manifest toxic effects that result in stunted skeletal growth. In experiments reported
since the 1986 Lead AQCD, Hac and Krechniak (1996) determined uptake and retention of Pb in
bone from rats exposed to plain water or water containing Pb acetate (41.7 to 166.6 mg/L) for
12 to 16 weeks. After 4 weeks, the skeletal Pb in animals receiving the lowest dose was almost
5 times higher than control animals (5.9 versus 1.2 jig Pb/g bone, respectively). Lead levels in
bones from animals receiving 83.3 mg/L and 166.6 mg/L were dose-dependently higher at
11.7 and 17.0 jig Pb/g bone, respectively, after 4 weeks of exposure. All bone Pb levels were
maintained essentially in a steady state until the completion of exposure, when all animals were
placed on control water. Approximately 64% of Pb remained in the bones of rats in the
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83.3 mg/L exposed group at 64 days postexposure. No blood levels of Pb were reported.
Similarly, airborne Pb can be inhaled and subsequently incorporated into bone. Grobler and co-
workers (1991) exposed 6-week-old rats to either "clean air" (0.05 jig Pb/m3) or air containing
77 jig Pb/m3 and found significant differences in the amount of Pb incorporated into the alveolar
bones of the animals. After 70 days, a mean of only 0.2 jig Pb/g of bone dry mass was found in
bone from control animals, while 16.9 jig Pb/g was present in bone from the 77 jig Pb/m3
exposure group. Exposure to air containing 249 jig Pb/m3 for 28 days or 1,546 jig Pb/m3 for
50 days, resulted in mean values of 15.9 and 158 jig Pb/g dry weight of Pb incorporation into the
bone, respectively, highlighting the fact that dose and length of exposure are determinates of
amount of Pb contained in the bones of these animals. Blood Pb levels were 2.6 |ig/dL in control
animals and ranged from 11.5 |ig/dL to 61.2 |ig/dL in the experimental groups. The uptake of Pb
by bone has the potential for immediate toxic effects on the cellular processes occurring during
bone growth, development, and maintenance, with the additional potential for delayed toxicity
from release of stored Pb during periods of normal or accelerated bone remodeling.
Numerous studies have examined growth suppression associated with developmental Pb
exposure. Hamilton and O'Flaherty (1994) examined the effects of Pb on growth in female rats,
and subsequently, on growth and skeletal development in their offspring. Administration of
drinking water containing either 250 or 1,000 ppm Pb to weaning female rats for 49 days
produced no alteration in growth rate in these future dams. Blood Pb levels prior to mating were
2.7 ± 0.6 |ig/dL (control), 39.9 ±3.5 |ig/dL (250 ppm group), and 73.5 ± 9.3 |ig/dL (1000 ppm
group). The rats were then bred, with Pb exposure continuing through parturition and lactation.
Lead did not affect gestation time nor Day 1 suckling body weight, however, pup body weight
and tail length were subsequently decreased in both exposure groups. A 10% increase in tibial
growth plate width and disruption of chondrocyte organization were observed in offspring from
the high exposure group.
In male rats exposed to 100 ppm Pb in drinking water and a low calcium diet for up to one
year, bone density was significantly decreased after 12 months, while rats exposed to 5,000 ppm
Pb had significantly decreased bone density after 3 months (Gruber et al., 1997). Pb content of
femurs was significantly elevated over the content of control rats at all time points (1, 3, 6, 9, and
12 months). Blood Pb levels ranged from 1 to 4 |ig/dL in control animals, 17 to 29 |ig/dL in low
dose animals, and 45 to 126 |ig/dL in high dose animals. Trabecular bone from the low dose
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animals was significantly decreased from 3 months forward. Young female rats exposed to
17 mg of Pb acetate per kg of feed for 50 days showed no differences in the length of the femurs,
but the mean length of the 5th lumbar vertebra was significantly decreased (Gonzalez-Riola
et al., 1997; Escribano et al., 1997). The mean length of the femur growth plate cartilage was
also significantly decreased in Pb-exposed animals. Blood Pb levels were not reported.
In a dose-response study, Ronis et al. (1998a,b) exposed pregnant rats to Pb acetate in
drinking water (0.05% up to 0.45% w/v) beginning at gestation Day 5 and continuing through
weaning of offspring at Day 21. Early bone growth was significantly depressed in a
dose-dependent fashion in pups of all Pb-exposed groups, with growth suppression in male
offspring considerably greater than in females. Significant decreases in plasma insulin-like
growth factor and plasma sex steroids and increased pituitary growth hormone were also
observed. Blood Pb levels in offspring ranged from 49 ± 6 |ig/dL (0.05% group) to
263 ± 28 |ig/dL (0.45% group). This is somewhat in contrast to the findings of Camoratto and
co-workers (1993), who reported low exposure to 0.02% Pb nitrate (125 ppm Pb) did not
significantly affect growth, though males weighed significantly less than females. Note however
that the blood Pb levels in the rat pups were less (43.3 ± 2.7 |ig/dL at 5d and 18.9 ± 0.7 |ig/dL
at 49d) than in the Camorrato study. Between age 57 and 85 days, Ronis et al. (1998b) noted
that growth rates were similar in control and Pb-exposed pups, suggesting exposure at critical
growth periods such as puberty and gender may account for differences in growth reported by
various investigators. In a series of follow-up experiments, Ronis et al. (2001) reported a dose-
dependent decrease in load to failure in tibia from Pb-exposed (0.15% and 0.45% Pb acetate in
drinking water) male pups only. Hormone treatments (estradiol in females or L-dopa,
testosterone or dihydrotestosterone in males) failed to attenuate Pb deficits during the pubertal
period. Distraction osteogenesis experiments performed after stabilization of endocrine
parameters (at 100 days of age) found decreased new endosteal bone formation and gap x-ray
density in the distraction gaps of Pb-exposed animals (Ronis et al., 2001). Again blood Pb levels
were high, ranging from 67 to 388 |ig/dL in the offspring.
Hamilton and O'Flaherty (1995) found Pb disrupted mineralization during growth when
they implanted demineralized bone matrix subcutaneously into male rats. In the matrix that
contained 200 jig Pb/g of plaque tissue, alkaline phosphatase activity and cartilage
mineralization were absent, though calcium deposition was enhanced. Separate experiments
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found enhanced calcification and decreased alkaline phosphatase activity in rats implanted with a
control (no Pb) matrix and given 1,000 ppm Pb in drinking water for 26 days (blood Pb 96.4 to
129.8 |ig/dL).
In summary, results from animal studies suggest Pb exposure is capable of adversely
affecting bone growth and density, potentially manifesting its action through interference with
growth and hormonal factors as well as toxic effects directly on bone.

5.8.4 Regulation of Bone Cell Function in Animals - Systemic Effects
of Lead
Lead may exhibit multiple complex systemic effects that ultimately could influence bone
cell function. As discussed in the animal studies below, Pb can modulate alterations in calcium
binding proteins and in calcium and phosphorus concentration in the blood stream, in addition to
potentially altering bone cell differentiation and function by altering plasma levels of growth
hormone and calciotropic hormones such as vitamin D3 [1,25-(OH)2D3] and parathyroid
hormone.

5.8.4.1 Hypercalcemia/Hyperphosphatemia
Intravenous injection of Pb has been shown to produce both an acute hypercalcemia and
hyperphosphatemia in rats (Kato et al., 1977). Injection of a relatively high dose of 30 mg/kg Pb
resulted in maximum values of calcium (17 mg%) after one hour and maximum values of
phosphorus (13.5 mg%) after 30 minutes. After 12 hours, the levels of both calcium and
phosphorus had returned to baseline levels. Histochemical examination demonstrated deposition
of Pb into bone and dentin in the rats, suggesting a direct action of Pb on bone and/or teeth,
ultimately displacing calcium and phosphorus and thereby producing hypercalcemia and
hyperphosphatemia. Blood Pb levels were not reported.

5.8.4.2 Vitamin D [1,25-(OH)2D3]
As discussed above, vitamin D [1,25-(OH)2D3] modulates the normal range of calcium in
serum. In rats fed a low calcium or low phosphorus diet, ingestion of 0.82% Pb in the diet
reduced plasma levels of 1,25-(OH)2D3; however, this effect is lost when a high calcium or
normal phosphorus diet is given (Smith et al., 1981), suggesting a high calcium/phosphorus diet
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reduces the susceptibility of vitamin D system to the effect of Pb. No mobilization of calcium
from bone or elevation of inorganic phosphorus was seen. Ronis et al. (2001) also reported no
effects of Pb on plasma concentrations of vitamin D metabolites, 25-OH D3 or 1,25-(OH)2D3, in
pubertal male rats exposed to either 0.15% or 0.45% Pb acetate in drinking water and maintained
on an adequate diet. High blood Pb levels (over 350 |ig/dL) were reported in some animals in
both of these studies. Fullmer (1995) found vitamin D function was severely compromised in
young growing chicks given a diet low in calcium (0.1% calcium) for two weeks and then
exposed to 0.2% or 0.8% Pb in their diet for an additional one or two weeks. In chicks
maintained on an adequate diet (1.2% calcium), exposure to 0.2% or 0.8% Pb in the diet resulted
in increased plasma levels of 1,25-(OH)2D3 as well as significantly increased intestinal
Calbindin-D protein [a calcium binding protein induced by 1,25-(OH)2D3] and its associated
mRNA, when compared with unexposed control chicks. Levels of intestinal Calbindin-D mRNA
and protein and plasma levels of 1,25-(OH)2D3 were elevated during the first week of Pb
exposure to chicks fed a diet deficient in calcium, but were significantly decreased by the second
week of Pb exposure. The study suggested Pb was mediating its effect through 1,25-(OH)2D3,
rather than via a direct action on the Calbindin-D protein. Follow up studies by Fullmer et al.
(1996) confirmed dose dependent increases in serum 1,25-(OH)2D3 levels (and Calbindin-D
protein and mRNA) with increasing dietary Pb exposure (0.1% to 0.8%) in similar experiments
performed on Leghorn cockerel chicks fed an adequate calcium diet. No blood Pb levels were
reported in either study.

5.8.4.3 Parathyroid Hormone
At least one animal study has associated experimental Pb exposure with secondary
hyperparathyroidism. Szabo et al. (1991) exposed Wistar Kyoto rats to either 1% Pb acetate in
water for a short term (10 weeks) or varying concentrations (0.001 to 1% Pb acetate) for a longer
term (24 weeks) to assess the influence of Pb on the interaction of the parathyroids with
1,25-(OH)2D3. Blood Pb levels in the short term experiment were reported simply as less than
0.2 |ig/dL in control animals and greater than 50 |ig/dL in the Pb-exposed animals. No levels
were reported for the longer term experiment. Short term administration of 1% Pb resulted in
significant increases in bone Pb; however, total serum calcium and ionized serum calcium were
significantly decreased, as compared to controls. Circulating levels of 1,25-(OH)2D3 were also
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decreased, though the rats were maintained on a normal calcium diet (0.95%). Parathyroid
glands from rats exposed short term to Pb were significantly increased in size over those in
control animals (178 jig per gland versus 96 jig per gland) and specific binding of 1,25-(OH)2D3
to parathyroid and intestinal tissue was increased. Likewise, long term administration of 1% Pb
resulted in significant increases in bone Pb and normalized parathyroid gland weights, and a
significant decrease in the level of 1,25-(OH)2D3. In the long term study, a dose-dependent
increase in parathyroid weight occurred with increasing exposure to Pb in drinking water.
The authors concluded the secondary hyperparathyroidism was associated with, and/or a result
of, the hypocalcemia and decreased 1,25-(OH)2D3 levels secondary to Pb exposure.

5.8.4.4 Growth Hormone
As discussed in Section 5.8.3, exposure to Pb has been associated with altered bone
metabolism and decreased growth and skeletal development (Hamilton and O'Flaherty, 1994,
1995; Gruber et al., 1997; Gonzalez-Riola et al., 1997; Escribano et al., 1997; Ronis et al.,
1998a,b, 2001; Camoratto et al., 1993), suggesting perturbation of one or more endocrine factors
such as growth hormone. To examine the effect of exposure to low-level Pb on pituitary growth
hormone release, Camoratto et al. (1993) exposed pregnant female rats to 0.02% Pb nitrate
(125 ppm Pb) beginning on gestational day 5 and continuing in pups through postnatal day 48.
Basal release of growth hormone from control and Pb-exposed pups at age 49 days was not
significantly different. Growth hormone releasing factor-stimulated release of growth hormone
from pituitaries of Pb-exposed pups was smaller than the stimulated release of growth hormone
from pituitaries of control animals (75% increase over baseline versus 171% increase,
respectively), but the difference did not achieve significance (p = 0.08). Growth hormone
content of the pituitary glands was also not influenced by Pb exposure. Ronis et al. (1998b)
reported similar findings in rat pups exposed to 0.05%, 0.15%, or 0.45% Pb acetate in drinking
water from gestation day 5 through postnatal day 85, with the exception being significantly
elevated pituitary growth hormone levels at postnatal day 55. Blood Pb levels for both of these
studies were reported above in Section 5.8.3. Taken together, these rat studies suggest that
differences in growth seen with Pb exposure may not necessarily be the result of alterations in
secretion of growth hormone.
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5.8.5 Bone Cell Cultures Utilized to Test the Effects of Lead
5.8.5.1 Bone Organ Culture
In an early bone organ culture study utilizing incorporated radioactive Pb into fetal radii
and ulnae, Rosen and Wexler (1977) reported release of Pb as the concentration of calcium in the
media was reduced or with addition of parathyroid hormone, but that calcitonin inhibited the
release of Pb as expected, verifying the capacity of this model system. The bone organ system
was subsequently used to evaluate the efficacy of Pb chelating agents, such as D-Penicillamine
and CaNa2EDTA (Rosen and Markokwitz, 1980; Rosen et al., 1982).

5.8.5.2 Primary Cultures of Osteoclasts and Osteoblasts
The ability to isolate primary cultures of osteoclasts and osteoblasts from mouse calvaria
provided an additional experimental model system to study the effects of Pb on specific bone
cells. Using isolated osteoclasts and osteoblasts, Rosen (1983) reported that uptake of
radioactive Pb by osteoclasts was rapid, almost linear, while osteoblasts showed very little
increase in uptake of Pb at increasing media concentrations. Physiological concentrations of
parathyroid hormone markedly increased uptake of Pb and calcium by osteoclast cells and, once
loaded with Pb, osteoclasts were capable of releasing Pb slowly into the media. Further kinetic
analysis of cultured osteoclastic bone cells indicated that cellular Pb is primarily associated with
the mitochondrial fraction (-78%) and that this Pb is readily exchangeable with the outside
media (Pounds and Rosen, 1986; Rosen and Pounds, 1988). Experiments conducted to
characterize the steady-state kinetic distribution and metabolism of calcium and Pb supported the
concept that the two elements are metabolized similarly in the osteoclast cells (Rosen and
Pounds, 1989).

5.8.5.3 Rat Osteosarcoma Cell Line (ROS 17/2.8)
In recent years, the rat osteosarcoma cell line ROS 17/2.8 has been used extensively to
investigate the influence of Pb on various cellular processes and kinetics within these
osteoblast-like cells. The ROS 17/2.8 cell model is useful in that the cells are capable of
producing osteocalcin (a bone protein important for proper bone mineralization), have high
alkaline phosphatase activity (an enzyme normally associated with mineralization of cartilage),
possess vitamin D receptors, and respond to parathyroid hormone. In comparisons of cellular Pb
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toxicity and metabolism between primary cell culture from mouse calvaria and the rat
osteosarcoma cell line, Long and co-workers (1990) reported remarkable similarities in the
profile of radiolabeled Pb kinetics and intracellular Pb distribution. Using this cell line, Schanne
and co-workers (1989) simultaneously measured intracellular Pb and calcium concentrations and
found 5 and 25 micromolar Pb produced sustained 50% and 120% (respectively) increases in
intracellular calcium over a 5 hour period, and that measurable entry of Pb into the cells could be
demonstrated at the higher concentration. These findings advanced the hypothesis that
perturbation of intracellular calcium concentration may be the mechanism of Pb bone toxicity.
Schirrmacher and co-workers (1998) reported that calcium homeostasis is upset within
20 minutes of its addition to calvarial bone cell culture. Their results suggested that the calcium-
ATPases of intracellular stores were potentially poisoned by Pb entering the cells. Wiemann
et al. (1999) demonstrated that Pb was also capable of interfering with the calcium release
activated calcium influx (CRAC) in calvarial bone cell cultures. Pb was found to partially inhibit
the influx of calcium into the bone cells, plus influx of Pb into the cells was greatly enhanced
(2.7 fold) after CRAC had been induced. These effects of Pb were found to be independent of
any inhibitory effect on calcium-ATPase.
Miyahara et al. (1995) did a series of experiments in 45Ca-labeled bone organ culture to
determine whether the Pb-induced hypercalcemia was the result of the active process of
biological bone resorption or simply physiochemical mineral dissolution. Lead introduced into
the culture at concentrations of 50 jiM and above stimulated the release of calcium and
hydroxyproline into the medium; however, no release was elicited from bones inactivated by
freezing and thawing. Pb-stimulated 45Ca release was inhibited by bafilomycin AI, eel
calcitonin, and scopadulcic acid B, suggesting the release was secondary to osteoclastic bone
resorption. Further evidence to support this conclusion came from experiments examining the
influence of two inhibitors of cyclooxygenase on Pb-induced bone resorption. Lead was found
to stimulate prostaglandin E2 release and in cultures, there was a high correlation between
prostaglandin E2 released into the media and 45Ca release. In the presence of cyclooxygenase
inhibitors (blocking prostaglandin synthesis), Pb-stimulated 45Ca release was inhibited,
suggesting the mechanism of bone resorption in this instance was via a prostaglandin
E2-mediated mechanism.
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Lead has been demonstrated to directly impair production of osteocalcin by ROS
17/2.8 cells by 70% after 24 hours of exposure to 25 micromolar Pb (Long et al., 1990).
The resulting decrease in cell proliferation is in agreement with similar studies by Sauk et al.,
1992). Interestingly, exposure of dental pulp cells, which also produce osteocalcin, to a similar
concentration of Pb reduced osteocalcin production by 55% after 12 hours of exposure
(Thaweboon et al., 2002). Vitamin D has been shown to increase osteocalcin production in
ROS 17/2.8 cells; however, Pb inhibited the vitamin D-stimulated osteocalcin production
in a dose-dependent manner from 0 up to 25 micromolar concentrations, plus was shown to be
capable of attenuating basal (non-vitamin D-stimulated) osteocalcin production (Long et al.,
1990). Lead (5 to 20 micromolar) inhibition of vitamin D stimulation of osteocalcin in ROS
cells was also reported by Guity and co-workers (2002). Later studies suggested that Pb acts by
inhibiting vitamin D activation of calcium channels and interferes with regulation of calcium
metabolism (Schanne et al., 1992), though apparently this effect is not mediated via PKC (Guity
et al., 2002). Angle and co-workers (1990) reported that 24 hours of incubation with vitamin D
(10 nM) was capable of evoking a 4 to 5 fold increase in osteocalcin production and a
100% increase in cellular alkaline phosphatase activity in ROS cells. Osteocalcin production and
cellular DNA contents were increased 100% and 20% respectively by addition of insulin-like
growth factor (92.5 ng/mL). Consistent with a toxic effect of Pb on osteoblast function, the
addition of 1 to 10 jiM Pb to the system inhibited both basal and stimulated osteocalcin
secretion, alkaline phosphatase activity and DNA contents (Angle et al., 1990). Dose- and time-
dependent reduction in alkaline phosphatase activity with Pb exposure (2 to 200 micromolar) has
also been reported in osteosarcoma cells, along with parallel reductions in steady state levels of
alkaline phosphatase mRNA levels (Klein and Wiren, 1993). No effect on cell number or DNA
and protein synthesis was seen at these levels of Pb exposure.
Though the exact mechanism of Pb toxicity on osteocalcin was unclear, Pb was known to
inhibit some of the functional properties of osteocalcin including inhibition of osteocalcin
adsorption to hydroxyapatite. An investigation by Dowd and co-workers (1994) utilized the
ability of osteocalcin added to a solution of 43CaCl2 to broaden 43Ca resonance, as a method to
examine binding of calcium to osteocalcin and the influence of Pb on calcium binding. It was
determined that the dissociation constant of calcium for osteocalcin was 7 micromolar, while the
dissociation constant for Pb was determined by competitive displacement to be 2 nM, indicating
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more than three orders of magnitude tighter binding of Pb than calcium to osteocalcin and the
likelihood that even submicromolar levels of free Pb would significantly inactivate osteocalcin.
Circular dichroism indicated that upon binding, Pb induces a similar structural change in
osteocalcin to that found with calcium binding, but the binding with Pb occurs at 2 orders of
magnitude lower than with calcium (Dowd et al., 2001). Similarly, hydroxyapatite binding
assays indicated Pb causes an increased absorption to hydroxyapatite that is similar to calcium,
but again at 2 to 3 orders of magnitude lower concentration, potentially leading to low bone
formation rates and/or density (Dowd et al., 2001).
Besides perturbation of calcium metabolism, Pb has been shown to reduce intracellular
free magnesium concentrations by 21% in osteosarcoma cells incubated in 10 micromolar Pb for
2 hours (Dowd et al., 1990). Under these same conditions, the unidirectional rate of ATP
synthesis (i.e., P; to ATP) was reduced by a factor greater than 6 over control cultures.
Impairment of both of these processes by Pb could ultimately influence bone growth and
development.
Lead has also been show to perturb Epidermal Growth Factor's (EGF) control of
intracellular calcium metabolism and collagen production in ROS cells (Long and Rosen, 1992).
EGF is known to activate protein kinase C (PKC), resulting in increased calcium influx and
through this mechanism, decreased collagen synthesis. Incubation of ROS cells with
5 micromolar Pb and 50 ng/mL EGF for 20 hours resulted in a 50% increase in total cell calcium
versus the calcium increase seen in cells treated with EGF alone, suggesting more than one site
of action is involved in calcium messenger perturbation. A similar finding was reported by Long
and co-workers (1992) who found that treatment of Pb (25 micromolar) intoxicated
osteosarcoma cells with parathyroid hormone (PTH, 400 mg/mL) resulted in a greater increase in
cell calcium than with either treatment alone. Supplementary inhibition of collagen synthesis has
also been reported with the addition of 25 micromolar Pb plus 50 ng/mL EGF, suggesting more
than one site of action for the effect of Pb on collagen synthesis (Long and Rosen, 1992).
Additional study has since suggested that Pb activates PKC in ROS cells and that PKC mediates
the rise in intracellular calcium (Schanne et al., 1997). The observation that calphostin C, an
inhibitor of PKC, prevented the Pb-induced elevation of intracellular calcium supported this
hypothesis, as did the fact that free Pb at concentrations of 10"11 to 10"7 M directly activated PKC
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in the absence of activating concentrations of calcium. This would suggest Pb is capable of
activating PKC at concentrations -3,000 times lower than calcium.
Finally, Pb has been shown to be capable of inhibiting secretion of osteonectin, a bone
related protein found in areas of active morphogenesis (Sauk et al., 1992). Treatment of
ROS 17/2.8 cells with Pb (4.5 x 10"6 M to 4.5 x 10"7 M) demonstrated that intracellular
osteonectin levels were actually enhanced; however, the secretion of osteonectin into the media
was delayed or inhibited. Protein production of collagen and the endoplasmic reticulum protein,
Hsp47, were relatively unaffected by Pb at these concentrations. The intracellular retention of
osteonectin coincided with a decrease in levels of osteonectin mRNA, suggesting the processes
associated with translation and secretion of osteonectin are sensitive to Pb.

5.8.5.4 Human Osteosarcoma Cells (HOS TE 85)
Evidence exists that Pb is directly osteotoxic to bone cells in culture. Studies examining
the sensitivity of human osteosarcoma cells (HOS TE 85) to Pb found proliferation of the cells
was inhibited at Pb concentrations of 4 |imol/l, while cytotoxicity occurred at the 20 |imol/l Pb
concentration (Angle et al., 1993). In parallel experiments, rat osteosarcoma cells (ROS 17/2.8)
were found to be somewhat less sensitive to the effects of Pb with inhibition of proliferation
occurring at 6 |imol/l Pb concentration and cytotoxicity at Pb concentrations over 20 |imol/l.

5.8.5.5 Chick Chondrocytes
The effects of Pb on cartilage biology have been examined in isolated avian chondrocytes
obtained from 3 to 5 week old chicks (Hicks et al., 1996). Exposure to media containing 0.1 to
200 jiM Pb acetate or chloride were found to decrease thymidine incorporation, suppress alkaline
phosphatase, and suppress both type II and type X collagen expression at the mRNA and protein
levels. Cytotoxicity of the cultures from Pb exposure was dismissed as proteoglycan synthesis
was found to be augmented, suggested Pb selectively inhibits specific aspects of the chondrocyte
growth plate. Using the avian chondrocyte model, Zuscik et al. (2002) similarly reported Pb
exposure (1 to 30 jiM) causing a dose-dependent inhibition of thymidine incorporation into the
growth plate, with a 60% reduction in proliferation at the highest concentration. Addition of
TGF-P1 and PTHrP, regulators of growth plate, both separately stimulated thymidine
incorporation, an effect that was dose-dependently blunted in the presence of Pb. At the highest
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Pb concentration (30 |iM), inhibition was significantly less in the chondrocytes treated with
Pb + TGF-P1 (24%) and Pb + PTHrP (19%) than for Pb alone (60%), suggesting the interaction
of Pb with these growth factors may be independent of its primary action on the chondrocyte
cells. Support for a direct action of Pb on these growth regulators is supported by the finding
that normal TGF-pl and PTHrP suppression of type X collagen expression is significantly
reversed in a dose-dependent fashion in the presence of Pb. This effect evidently was not
mediated by BMP-6 (Bone Morphogenic Protein), an inducer of terminal differentiation known
to partially reverse the inhibitory effect of PTHrP, because in the presence of Pb, PTHrP
significantly suppressed BMP expression, while combined exposure to Pb and TGF-pl increased
BMP expression ~3-fold. Further experiments performed on chick sternal chondrocyte cultures,
utilized PTHrP responsive (AP-1) and non-responsive (NF-^B) reporter constructs to examine
potential effects of Pb on signaling. While having no effect on the basal activity of the AP-1
reporter, Pb dose-dependently enhanced PTHrP induction of the responsive AP-1 reporter. Lead
dose-dependently inhibited the basal activity of the non-PTHrP responsive, NF-KB reporter.
Taken together, these studies demonstrate that Pb has an inhibitory effect on the process of
endochondral bone formation and that the effects of Pb are likely from its modulation of growth
factors and second messengers involved in cell signaling responses.

5.8.6 Bone Lead as a Potential Source of Toxicity in Altered
Metabolic Conditions
Lead is avidly taken up by bone and incorporated into bone matrix, where a substantial
amount can remain over the lifetime of an organism. The uptake and incorporation of Pb into
bone during acute exogenous exposures may be of short term benefit by limiting the exposure of
other, more sensitive tissues; however, this does not eliminate Pb from the system. Subsequent
release of Pb from this endogenous storage can produce a lifetime of steady, low level Pb
exposure during periods of normal bone remodeling, while elevated Pb release during times of
increased bone metabolism and turnover (i.e., pregnancy, lactation, menopause, and
osteoporosis) can elevate blood levels of Pb significantly, potentially to toxic concentrations.
This is especially relevant when there is concurrent exogenous exposure to Pb, as current blood
Pb levels are a composite of current and past Pb exposure. Of greater concern is the mobilization
of Pb during pregnancy and subsequent transfer to the developing brain of the fetus across the
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poorly developed blood:brain barrier. Maternal Pb also appears in breast milk, providing further
exposure of the infant to Pb during lactation. Currently, the majority of animal studies
examining mobilization of Pb from bone stores have focused principally on elevation of Pb
levels or transfer of Pb, rather than reporting toxic effects associated with these exposures. Note
that in most instances the mobilization and elimination of Pb is much faster in laboratory animals
than in humans. For example, as discussed in Section 5.8.3, Hac and Kruchniak (1996) reported
-64% of Pb given over a 12 week period remained in the bones of rats 64 days post exposure.
Therefore, the caveats of experiments performed in small animals, especially when examining
mobilization of Pb stores, must be taken into consideration.

5.8.6.1 Pregnancy and Lactation
Pregnancy, and to a much greater extent, lactation, place significant calcium demands on
the mother as she provides all the necessary calcium requirements of the developing fetus/infant.
During these times of metabolic stress, increased demineralization of maternal bone occurs to
supplement demand, unfortunately accompanied by the concurrent mobilization and release of
Pb stored in the maternal skeleton from past exposure. Studies in several animal models have
shown that maternal bone Pb can be mobilized during pregnancy and lactation, ultimately being
transferred to the fetus during gestation and breast feeding. Keller and Doherty (1980)
administered radiolabeled Pb drinking water (200|ig/mL) to female mice for 105 days prior to
mating or 105 days prior to mating and during periods of gestation and lactation (total 160 days
of exposure). The results suggested very little Pb was transferred from mother to fetus during
gestation, however, Pb transferred in milk and retained by the pups accounted for 3% of the
maternal body burden of those mice exposed to Pb prior to mating only. No blood Pb levels
were reported for any of the animals. The amount of Pb retained in these pups exceeded that
retained in the mothers, suggesting lactation effectively transfers Pb burden from mother to
suckling offspring. Transfer of Pb from mothers was significantly higher when Pb was supplied
continuously in drinking water, rather than terminated prior to mating. Considerably higher
lactational transfer of Pb from rat dams compared to placental transfer has also been reported
(Palminger Hallen et al., 1996). Continuous exposure of rat dams to Pb until day 15 of lactation
resulted in milk Pb levels 2.5 times higher than in whole blood, while termination of maternal Pb
exposure at parturition yielded equivalent blood and milk levels of Pb, principally from Pb
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mobilized from maternal bone. Blood Pb concentrations at day 15 of lactation were
1.4 ± 0.4 |ig/dL (control), 32.0 ± 5.5 |ig/dL (Pb-exposed until parturition), and
126.0± 17.1 |ig/dL (Pb-exposed until day 15 of lactation).
Using rats chronically exposed to Pb in drinking water, Maidonado-Vega et al. (1996)
studied intestinal absorption of Pb, its mobilization, and redistribution during lactation. In rats
exposed to Pb 144 days prior to lactation, the process of lactation itself elevated blood Pb and
decreased bone Pb, indicating mobilization of Pb from bone as there was no external source of
Pb during the lactation process. Rats exposed to Pb for 158 days (144 days prior to lactation and
14 days during lactation) also experienced elevated BLLs and loss of Pb from bone. Lead
exposure only during the 14 days of lactation was found to significantly increase intestinal
absorption and deposition (17 fold increase) of Pb into bone compared to non-pregnant rats,
suggesting enhanced absorption of Pb takes place during lactation. As in other previous studies,
the highest concentration of Pb in bone was found in non-pregnant non-lactating control animals,
with significantly decreased bone Pb in lactating rats secondary to bone mobilization and transfer
via milk to suckling offspring. Blood Pb levels at day 14 of lactation or equivalent ranged from
24.7 to 31.2 |ig/dL. Follow-up studies examining the influence of dietary calcium found when
calcium was altered from the normal 1% to 0.05%, bone calcium concentration decreased by
15% and bone Pb concentration decreased by 30% during the first 14 days of lactation
(Maldonado-Vega et al., 2002). In non-lactating rats on the 0.05% calcium diet, there were also
decreases in bone calcium, but neither incremental bone resorption nor Pb efflux from bone,
suggesting the efflux from bone during lactation was related to bone resorption. Of interest,
enhancement of calcium (2.5%) in the diet of lactating rats increased calcium concentration in
bone by 21%, but did not decrease bone resorption, resulting in a 28% decrease in bone Pb
concentration and concomitant rise in systemic toxicity. Blood Pb levels were similar to those
reported in the prior study above. In both studies, the authors concluded that Pb stored in bone
should be considered a major source of self-intoxication and of exposure to suckling offspring.
In one of few studies showing a toxic effect, Han et al. (2000) demonstrated adverse
effects in rat offspring born to females whose exposure to Pb ended well before pregnancy. Five
week-old-female rats had been given Pb acetate in drinking water (250 mg/mL) for five weeks,
followed by a one month period without Pb exposure before mating. To test the influence of
dietary calcium on Pb absorption and accumulation, some pregnant rats were fed diets deficient
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in calcium (0.1%) while others were maintained on a normal calcium (0.5%) diet. As expected,
all Pb-exposed dams and pups had elevated blood Pb levels; however, pups born to dams fed the
diet deficient in calcium during pregnancy had higher blood (up to 24 |ig/dL) and organ Pb
concentrations compared to pups from dams fed the normal diet. Significantly, pups born to
Pb-exposed dams had lower mean birth weights and birth lengths than pups born to
non-Pb-exposed control dams (p < 0.0001), even after confounders such as litter size, pup sex,
and dam weight gain were taken into account. The authors concluded that while increases in
dietary calcium during pregnancy are capable of reducing Pb accumulation in the fetus, they
cannot prevent the decreases in birth weight and length associated with pre-maternal Pb exposure
and subsequent mobilization. This has relevance in human pregnancy, as many women
experience exposure to Pb during their lifetimes (especially during childhood) and mobilization
of the Pb from bone stores during pregnancy could present toxic complications.
Within the last decade, an invaluable method to explore the kinetics of Pb transfer from
bone to blood has been developed and evaluated (Inskip et al., 1996; O'Flaherty et al., 1998).
The method utilizes recent administration of sequential doses of Pb mixes enriched in stable
isotopes ( 204Pb, 206Pb, and 207Pb) to female cynomolgus monkeys (Macaca fascicularis) that
have been chronically (1,300 to 1,500 jig Pb/kg body weight per day for ten years or greater)
administered a common Pb isotope mix. The stable isotope mixes serve as a marker of recent,
exogenous Pb exposure, while the chronically administered common Pb serves as a marker of
endogenous (principally bone) Pb. From thermal ionization mass spectrometry analysis of the
Pb isotopic ratios of blood and bone biopsies collected at each isotope change, and using
end-member unmixing equations, it was determined that administration of the first isotope label
allows measurement of the contribution of historic bone stores to blood Pb. Exposure to
subsequent isotopic labels allowed measurements of the contribution from historic bone Pb
stores and the recently administered enriched isotopes that incorporated into bone (Inskip et al.,
1996). In general the contribution from the historic bone Pb (common Pb) to blood Pb level was
constant (-20%), accentuated with spikes in total blood Pb due to the current administration of
the stable isotopes. Blood Pb ranged from 31.2 to 62.3 jig/100 g in the animals. After cessation
of each sequential administration, the concentration of the signature dose rapidly decreased.
Initial attempts to apply a single-bone physiologically based model of Pb kinetics were
unsuccessful until adequate explanation of these rapid drops in stable isotopes in the blood were
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incorporated (O'Flaherty et al., 1998). Once revisions were added to account for rapid turnover
of the trabecular bone compartment and slower turnover rates of cortical bone compartment, an
acceptable model evolved. From this model it was reported that historic bone Pb from 11 years
of continuous exposure contributes -17% of the blood Pb at Pb concentrations over 50 |ig/dL,
reinforcing the concept that the length of Pb exposure and the rates of past and current Pb
exposures help determine the fractional contribution of bone Pb to total blood Pb levels
(O'Flaherty et al., 1998). The turnover rate for cortical (-88% of total bone by volume) bone in
the adult cynomolgus monkey was estimated by the model to be -4.5% per year, while the
turnover rate for trabecular bone was estimated to be 33% per year.
Using the method of sequential stable isotope administration, Franklin et al. (1997)
examined flux of Pb from maternal bone during pregnancy of 5 female cynomolgus monkeys
who had been previously exposed to common Pb (-1,100 to 1,300 jig Pb/kg body weight) for
about 14 years. In general, Pb levels in maternal blood (as high as 65 jig/100 g) attributable to
Pb from mobilized bone were reported to drop 29 to 56% below prepregnancy baseline levels
during the first trimester of pregnancy. This was ascribed to the known increase in maternal
fluid volume, specific organ enlargement (e.g., mammary glands, uterus, placenta), and increased
metabolic activity that occurs during pregnancy. During the second and third trimesters, when
there is a rapid growth in the fetal skeleton and compensatory demand for calcium from the
maternal blood, the Pb levels increased up to 44% over pre-pregnancy levels. With the
exception of one monkey, blood Pb concentrations in the fetus corresponded to those found in
the mothers, both in total Pb concentration and proportion of Pb attributable to each isotopic
signature dose (common = 22.1% versus 23.7%, 204Pb = 6.9% versus 7.4%, and 206Pb = 71.0%
versus 68.9%, respectively). From 7 to 25% of the Pb found in fetal bone originated from
maternal bone, with the balance derived from oral dosing of the mothers with isotope during
pregnancy. Of interest, in offspring from a low Pb exposure control monkey (blood Pb
<5 jig/100 g) -39% of Pb found in fetal bone was of maternal origin, suggesting enhanced
transfer and retention of Pb under low Pb conditions.
Clearly, the results of these studies show that Pb stored in bone is mobilized during
pregnancy and lactation, exposing both mother and fetus/nursing infant to blood/milk Pb levels
of potential toxicity. Of equal concern, a significant proportion of Pb transferred from the
mother is incorporated into the developing skeletal system of the offspring, where it can serve as
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a continuing source of toxic exposure. The above study by Franklin et al. (1997) illustrates the
utility of sequentially administered stable isotopes in pregnancy; however, its use may also be
applicable in studies of lactation, menopause, osteoporosis, and other disease states where
mobilization of bone and release of Pb stores occurs. Furthermore, given that isotopic ratios of
common Pbs vary by location and source of exposure, when humans migrate from one area and
source of exposure to another, it is possible to document changes in mobilized Pb, especially
during times of metabolic stress.

5.8.6.2 Age/Osteoporosis
The age of an animal at the time of exposure to Pb has been shown to influence the uptake
and retention of Pb by bone. In experiments to determine the influence of age on this process,
Han et al. (1997) exposed rats for five weeks to 250 mg/L Pb acetate in drinking water beginning
at 5 weeks of age (young child), 10 weeks of age (mid-adolescence), or 15 weeks of age (young
adult), followed by a 4 week period of without Pb exposure. An additional group of rats were
exposed to Pb beginning at 5 weeks, but examined following an 8 or 20 week period after
cessation of Pb. Significantly lower blood and bone Pb concentrations were associated with
greater age at the start of Pb exposure and increased interval since the end of exposure.
No blood Pb levels were greater than -30 |ig/dL. However, young rats beginning exposure to Pb
at 5 weeks and examined 20 weeks after cessation of exposure, still had bone Pb concentrations
higher than those found in older rats only 4 weeks after cessation of exposure. This showed that
exposure to Pb at a young age leads to significant skeletal Pb accumulation and retention, despite
the high rate of bone remodeling that occurs during growth and development at that time.
At the opposite end of the spectrum, Cory-Slechta et al. (1989) studied differences in
tissue distribution of Pb in adult and old rats. Adult (8 months old) and old (16 months old) rats
were exposed to 50 ppm Pb acetate in drinking water for 11 months, at which time the
experiment was completed. Bone (femur) Pb levels in older rats were found to be less than those
in younger rats; however, blood Pb levels were higher in the older rats. All levels of Pb in the
blood were reported to be 31 |ig/dL or less. Of interest, brain Pb concentrations in the older rats
exposed to Pb were significantly higher, and brain weights were significantly less than the brain
Pb concentration and weights of unexposed older control rats or adult rats exposed to Pb,
suggesting a potential detrimental effect. The authors suggested that a possibility for the
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observed differences in tissue concentrations of Pb was due to changes in the capacity of bone to
store Pb with advanced age.
In a subsequent study, Cory-Slechta (1990b) examined kinetic and biochemical responses
of young (21 day old), adult (8 months old), and old (16 months old) rats exposed to Pb at 0, 2,
or 10 mg Pb acetate/kg/day over a 9.5 month experimental period (blood Pb as high as
45 jig/dL). Results suggested that older rats may have increased vulnerability to Pb due to
increased exposure of tissues to Pb and greater sensitivity of the tissues to the effects of Pb.
As in the previous study (Cory-Slechta et al., 1989), lower bone Pb levels were present in older
rats with concomitant elevated levels of Pb in brain and other tissues, supporting the hypothesis
that exposure to Pb over a lifetime may contribute to deterioration of health in old age,
potentially during times of heightened bone remodeling such as occurs during osteoporosis.
In studies of bone Pb metabolism in a geriatric, female nonhuman primates exposed to Pb
-10 years previously (historic blood Pb concentration of 44 to 89 |ig/dL), McNeill et al. (1997)
reported no significant changes in bone Pb level over a 10 month observation period as measured
by 109Cd K X-ray fluorescence. The mean half-life of Pb in bone of these animals was found to
be 3.0 ± 1.0 years, consistent with data found in humans, while the endogenous exposure level
due to mobilized Pb was 0.09 ± 0.02 |ig/dL blood. Results examining Pb accumulation in the
bones of aging male mice suggest low levels of bone Pb contributing to the osteopenia observed
normally in C57BL/6J mice (Massie and Aiello, 1992). The mice were maintained on regular
diet (0.258 ppm Pb) and water (5.45 ppb Pb) from 76 to 958 days of age. While the Pb content
of femurs increased by 83%, no significant relationship was found between Pb and bone density,
bone collagen, or loss of calcium from bone. Blood Pb levels were not reported.

5.8.6.3 Weight Loss
The relationship between body mass and bone mass is highly correlated, and during times
of loss of body weight, such as dietary restriction, a concomitant loss of bone mass also occurs.
It is therefore possible that Pb stored in bone from prior exposures could be released into the
system as skeletal bone is mobilized and result in Pb toxicity. To examine the influence of
weight loss on release of stored Pb, Han et al. (1996) first exposed rats to Pb in drinking water
(250 mg/1 of Pb as acetate) for 5 weeks, followed by a 4 week washout period without Pb to
allow primarily accumulation in the skeleton. Rats were then randomly assigned to a weight
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maintenance group, a moderate weight loss group (70% of maintenance diet), or a substantial
weight loss group (40% of maintenance diet) for a four week period. At the end of this
experimental period the blood Pb (24 to 28 |ig/dL) and bone Pb levels did not differ between
groups, however, the amount and concentration of Pb in the liver increased significantly.
A follow up study in rats previous exposed to Pb for two weeks was undertaken to determine the
effect of weight loss and exercise on the distribution of Pb (Han et al., 1999). They found weight
loss secondary to dietary restriction to be the critical factor elevating organ Pb levels and,
contrary to their first study, elevated blood levels of Pb (as high as 42 |ig/dL). No significant
difference in organ or blood Pb concentrations was reported between the exercise versus no
exercise groups. These studies suggest Pb toxicity could occur in those previously exposed to Pb
during times of dietary restriction.

5.8.7 Teeth - Introduction
There was little information in the prior 1986 Lead AQCD relating Pb exposure to
adverse outcomes in the teeth of animals. At that time, the incorporation of Pb into teeth was
recognized, as was the fact that tooth Pb increased with age, proportional to the rate of exposure
and roughly proportional to the blood Pb concentration.
Teeth consist of a hard outer layer of enamel, supported by an underlying layer of dentin,
which itself is supported by a connective tissue known as the dental pulp. Enamel is the hardest
substance in the body and the most highly mineralized, consisting of-96% mineral (calcium
hydroxyapatite substituted with carbonate ions) and 4% other organic materials, while dentin is
only -70% mineral.
The formation of enamel (amelogenesis) occurs as a two stage process of organic matrix
production with -30% mineralization, followed by removal of water and proteins from the
matrix with concurrent further mineralization. As in bone, Pb ions are apparently capable of
substituting for calcium ions in the mineralizing tooth, becoming essentially trapped. However,
unlike bone, the tooth, with subtle exceptions, does not undergo a remodeling process. Dentin
formation (dentinogenesis) can be likened to endochondral bone formation, in that an
unmineralized matrix (predentin, rather than cartilage) is laid down first, followed by
mineralization to mature dentin. The cells responsible for amelogenesis and dentinogenesis,
called ameloblasts and odontoblasts respectively, are similar to osteoblasts in that they respond
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to various signaling factors, secrete matrix proteins, and create an environment favorable to
deposition of minerals. After enamel formation on a specific tooth is completed, ameloblasts are
lost and no additional enamel is laid down with the exception of certain teeth in rodents. These
teeth, typically incisors in rats, mice, and most other rodents, continuously erupt to offset the
attrition that occurs with daily use. Therefore, the process of amelogenesis is ongoing, albeit
confined to a localized area, throughout the life of the animals. For this reason, rodents have
been utilized extensively to examine the processes of amelogenesis and the influence of various
toxic agents, such as Pb, on tooth development.
Ameloblasts are especially sensitive to toxins and altered metabolic conditions and
respond to such insults with disruption of enamel formation. When disruption occurs, defects in
the enamel can occur, typically as a band of malformed or altered enamel. As described below,
exposure of animals to various concentrations of Pb during tooth development is not only
capable of creating distinctive marking of enamel ("Pb lines") but may also influence the
resistance of the enamel to dental decay. Within the dental pulp, a layer of odontoblasts
continues to reside against the inner layer of the primary dentin for the life of the tooth. During
this time, the odontoblasts are systematically slowly putting down thin layers of secondary
dentin, slowly decreasing the size of the pulp chamber with age. Lead present during this
process has been shown to be readily taken up by this dentin layer, providing a potential marker
of historic Pb exposure. Though the enamel is a non-living substance, it is not entirely inert.
The external surface of enamel is more or less in a continuous state of flux or turnover as it
chemically demineralizes from acids consumed or produced in the mouth by bacteria, followed
by remineralization of demineralized enamel when contact with saliva supersaturated with
calcium and phosphate ions occurs. Lead present during this process can easily be released from
enamel and/or incorporated initially or back into it depending on the circumstances.
In summary, Pb has the potential to disrupt the various processes associated with
formation of teeth, plus incorporate itself into all mineralized tooth tissues during formation.
Posteruptively, Pb can become incorporated into the secondary dentin, and can be taken up or
released from the outer surface layer of enamel during times of remineralization or
demineralization. As described below, exposure of animals to Pb has been associated with
adverse dental outcomes.
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5.8.8 Uptake of Lead by Teeth
As seen with bone, uptake of Pb into the teeth of animals has been demonstrated in a
number of studies and by multiple routes of administration. Twenty four hours after a single
intraperitoneal injection of radioactive Pb-203 (203Pb, 1 |ig/kg) to young (15 day suckling rats)
and old (120 day) female rats, 0.7% of the injected dose was present in the four incisor teeth of
the younger animals and 0.6% was present in the same teeth of the older animals (Momcilovic
and Kostial, 1974). These percentages jumped to 1.43% and 0.88%, respectively, 192 hours
after the injection, suggesting incorporation and retention of Pb by teeth is greater in younger
animals than in adults, as found in bone. Lead has also been shown to be incorporated into the
incisors of rats exposed to airborne Pb. Grobler and co-workers (1991) exposed 6-week-old rats
to either "Clean Air" (0.05 jig Pb/m3) or air containing 77 jig Pb/m3 and found significant
differences in the amount of Pb incorporated into the incisors of the animals. After 70 days, a
mean of only 0.8 jig Pb/g of incisor dry mass was found in incisors from control animals, while
11.0 jig Pb/g was present in incisors from the 77 jig Pb/m3 group. Exposure to air containing
249 jig Pb/m3 for 28 days or to 1,546 jig Pb/m3 for 50 days resulted in mean values of 13.8 and
153 jig Pb/g incisor dry weight of Pb incorporation, respectively, highlighting the fact that dose
and length of exposure are determinates of amount of Pb contained in the teeth of these animals.
Blood Pb levels were 2.6 |ig/dL (control), 11.5 |ig/dL (low exposure), 24.1 |ig/dL (middle
exposure), and 61.2 |ig/dL (high exposure). Lead has also been shown to be taken up into the
teeth of weanling rats whose mothers were exposed to Pb in drinking water. The offspring of
pregnant rats exposed during gestation and lactation until 21 days post partum to water
containing 0, 3, or 10 ppm Pb showed dose-dependent, significant increases in the Pb content of
incisors, first molars, and second molars (Grobler et al., 1985). No blood Pb levels were
reported. Taken together, these studies confirm the uptake of Pb into teeth as delivered by
various means and suggest that maternal exposure can result in uptake in offspring, during
gestation and/or lactation.

5.8.9 Effects of Lead on Enamel and Dentine Formation
Early microscopic studies by Eisenmann and Yaeger (1969) confirmed alterations in rat
incisor enamel formation 7 days after a single SC dose of Pb (0.15 or 1.5 mM/lOOg animal
weight); however, no effect was seen at the 0.075 mM/lOOg dose. Lead was found to have
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inhibited mineralization of both enamel and dentin, but only to a "mild to moderate" extent with
the mineralization of dentin more affected. It was speculated at the time that Pb could affect the
production of normal, mineralizable organic matrix; affect enzymes specific to enamel or dentin
formation; affect crystal structure and/or growth; or affect a combination of these factors. In
studies of dentinogenesis, incubation of fixed rat molar germs with Pb-pyrophosphate has shown
localization of Pb to the mineralization front of dentin (i.e., the area of recently formed dentin),
to the stratum intermedium, and to subodontoblastic cells, suggesting Pb may react with mineral
components located in the mineralization zone or have a high affinity for these incompletely
mineralized areas (Larsson and Helander, 1974). Localization of Pb was also seen at the area of
the dentino-enamel junction. Similar examination of first molar germs from 3-day-old rats
showed that Pb also localized to the periphery of dentinal globules (Larsson, 1974). A single,
high-dose injection of Pb acetate (30 mg/kg body weight) produces an immediate (within 6 h)
response in the growing dentin of the rat incisor, leading to the formation of a so-called "Pb line"
(Appleton, 1991). A transient rise in serum calcium and phosphorus accompanied the injection,
leading to speculation that Pb may have been replacing these minerals in the apatite structure.
However, backscattered electron imaging of the Pb line showed it to be composed of continuous
hypomineralized interglobular dentin with some incomplete fusion of calcospherites resulting in
uneven mineralization, but no localized concentration of Pb was detectable.
This is consistent with Featherstone and co-workers (1981) who reported that Pb
incorporation during apatite synthesis was widely dispersed, rather than concentrated in areas of
calcium deficiency. Once synthesis is complete, however, Pb is capable of entering calcium
deficient areas in enamel, substituting for calcium (Featherstone et al., 1979). This is essentially
the process that occurs during demineralization/remineralization of enamel. Appleton (1991,
1992) suggested that Pb has a direct effect on odontoblasts, creating a local disturbance of
calcium metabolism, a process similar to that described in bone (Pounds et al., 1991).
Interestingly, no ultrastructural changes were seen in ameloblasts from rat pups whose mothers
had been drinking water containing Pb.
During the normal process of amelogenesis, water and proteins contained within the
organic matrix are lost, leaving densely mineralized enamel. The removal of enamel proteins
during this phase is facilitated by enamel proteinases, which are believed to degrade the proteins
into smaller units capable of diffusing from the matrix. Using crude extracts from scrapings of
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rat incisor teeth, Gerlach and co-workers (2000a) demonstrated that Pb inhibited these
proteinases in vitro at micromolar concentrations. In rats given drinking water containing Pb at
either 0, 34, or 170 mg/L as Pb acetate for 70 days, increased amounts of proteins were found in
enamel matrix from animals exposed to Pb (Gerlach et al., 2002). Moreover, enamel
microhardness analysis of upper incisors revealed a significant decrease in microhardness in
regions of enamel maturation, but not in areas of fully mature enamel, suggesting Pb exposure
mediates a delay in enamel mineralization. In adult rats with incisors trimmed to remove
occlusal (biting) contact, a single IP dose of Pb acetate (40 mg/kg) significantly delayed the
continuous eruption of the incisor at all time points between 8 and 28 days after dosing,
compared with controls (Gerlach et al., 2000b). Blood Pb levels were -48 |ig/dL immediately
after injection and 16 |ig/dL 30d after injection. It is of interest that delayed eruption of teeth in
children living in areas of heavy metal contamination (Pb and zinc) has been reported previously
(Curzon and Bibby, 1970).

5.8.10 Effects of Lead on Dental Pulp Cells
Hampered by a general lack of cell cultures specifically for teeth, there remains a paucity
of information regarding both the cultures themselves and the effect of Pb upon such cultures. In
a single in vitro study using a human dental pulp cell culture obtained from teeth extracted for
orthodontic purposes, Thaweboon and co-workers (2002) examined the effects of three
concentrations (4.5 x 10"5 M, 4.5 x 10"6 M, 4.5 x 10"7 M) of Pb glutamate on cell proliferation,
protein production, and osteocalcin secretion. Under serum free conditions (DMEM only) all
concentrations of Pb significantly increased cell proliferation on day 1, day 3 and day 5 of
exposure, as measured indirectly by mitochondrial dehydrogenase enzyme assay. In the
presence of 2% fetal bovine serum only, the higher concentration of Pb significantly increased
protein production, suggesting an influence of serum constituents on cell growth or binding of
free Pb in the medium. Similar results were reported when rat osteosarcoma cells (ROS 17/2.8)
were exposed to identical concentrations of Pb over 2-, 4-, and 6-day time points (Sauk et al.,
1992). Concentrations of Pb less than 4.5 x 10"5 M concentration did not affect osteosarcoma
cell proliferation in the presence of serum but, in the absence of serum, 4.5 x 10"7 M Pb
increased cell proliferation at day 4, while at day 6, 4.5 x 10"6 M Pb inhibited proliferation.
Further testing of human dental pulp cells in serum-free conditions showed that Pb exposure
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caused dose-dependent decreases in intracellular protein and procollagen type I production over
the 5-day period experimental period (Thaweboon et al., 2002). Short-term exposure of the cells
to Pb significantly decreased osteocalcin production in a dose-dependent manner at 8- and 12-h
exposure time points. These results suggest that Pb is capable of exerting multiple toxic effects
on cells derived from human dental pulp.

5.8.11 Adverse Effects of Lead on Teeth—Dental Caries
In a recent review, Bowen (2001) highlighted 12 epidemiological studies that examined
the association between Pb exposure and dental caries (decay), reporting that 8 studies supported
the concept that Pb is a caries-promoting element. Unfortunately, the source and actual exposure
to Pb and measurement of prevalence of caries varied greatly, providing less than completely
satisfactory evidence in the opinion of the author. There is also a paucity of well-controlled
animal studies examining this issue.
In an early study examining the effect of drinking solutions containing various metallic
ions on dental caries in hamsters, Wisotzky and Hein (1958) reported post-eruptive ingestion of
drinking water containing 0.5 mEq of Pb significantly increased caries scores in molar teeth of
males after 84 days, but, perplexingly, not in females after 98 days of exposure. It should be
noted that in animal studies such as these it is routine to maintain the animals on cariogenic or
caries-promoting diets high in fermentable sugars. Clear evidence supporting Pb's role in
enhancing susceptibility to dental caries was reported by Watson and co-workers in 1997.
In their study, female rats were exposed to Pb in drinking water (34 ppm as Pb acetate) as young
adults, during pregnancy, and during lactation. Lead exposure of the subsequent offspring from
the dams was, therefore, from transfer of endogenous Pb from dam to pup during gestation and
lactation, with no further exposure after weaning. This pre- and perinatal exposure to Pb resulted
in a significant, almost 40%, increase in the prevalence of dental caries over control animals.
The study was significant for other reasons, as it mimicked the conditions found in many inner
cities, where young females are exposed to Pb in their environment and later transfer this Pb to
their own fetuses during the extensive bone remodeling that occurs during pregnancy and
lactation. The mean blood Pb level in the dams upon weaning was 48 |ig/dL, which is not unlike
upper levels reported in humans.
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The mechanisms by which Pb enhances susceptibility to caries remain uncertain, though
clearly altered mineralization and/or incorporation of Pb into enamel as described above could
enhance its solubility in acid. Lead also appears in the saliva of rats at about 5% of the whole
blood level and at about 61% of the plasma filtrate Pb level (Mobarak and P'an, 1984), providing
an avenue for post-eruptive interaction with the exposed enamel in the oral cavity. Notably,
decreased salivary flow has been reported in rats exposed to Pb, and decreased salivary function
is known to increase caries risk. Stimulated parotid function was decreased by nearly 30% in the
Pb-exposed offspring in the study by Watson and co-workers (1997), an effect that could have
been mediated by the salivary gland requirement of intact parasympathetic and sympathetic
nervous systems for normal development (Schneyer and Hall, 1970) and Pb's known adverse
effect on neurotransmitters (Bressler and Goldstein, 1991). Acute infusion of 4 jig of Pb per min
was reported to significantly reduce pilocarpine-stimulated salivary secretion in rats over a
50-min period (Craan et al., 1984), whereas 24-day administration of 0.05% Pb acetate
significantly reduced the concentration of protein and calcium in pilocarpine-stimulated rat
submandibular saliva (Abdollahi et al., 1997). Of potential interest, postnatal exposure of rats to
Pb (10 or 25 ppm in drinking water) and a caries-enhancing diet containing fluoride (sucrose
containing 15 ppm fluoride) was not associated with an increased risk of dental caries, which
suggests that Pb does not interfere with the protective effect of fluoride (Tabchoury et al., 1999).
Clearly though, the effect of Pb exposure on salivary gland function and the mechanism by
which Pb exposure enhances caries risk needs to be further explored.

5.8.12 Lead from Teeth as a Potential Source of Toxicity
Although no studies currently document the contribution of Pb incorporated into teeth as a
source of endogenous Pb exposure, the potential exists during the process of exfoliation of the
primary dentition. As described above (Section 5.8.9) Pb is avidly incorporated into the
developing dentin and enamel components of teeth. Like bone, the uptake and incorporation of
Pb into teeth during acute exogenous exposures may be of short-term benefit by limiting the
exposure of other, more sensitive tissues, but, unlike bone, teeth do not undergo a gross
remodeling process (the continuous, superficial demineralization/remineralization of the exposed
tooth surfaces, principally enamel, are assumed here to be insignificant). However, during the
exfoliative process, the erupting secondary tooth erodes away the root (composed of cementum
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and dentin) of the overlying primary tooth along with some surrounding alveolar bone. Any Pb

incorporated into these portions of bone and primary tooth would be released by the erosive

process, with the potential to produce highly elevated local concentrations of Pb in the proximity

of remodeling alveolar bone and developing secondary teeth. A more modest contribution to

circulating blood Pb would be predicted. Animal research in this area has been hampered, as

most common rodents (i.e., rats, mice) are monophyodonts (have only one set of teeth).

Although monkeys are an acceptable model, it is problematic as to how release of Pb stored in

teeth could be differentiated from that of remodeling skeletal bones formed at a similar time

point, plus the disproportionate size of the skeletal mass compared to the dentition may mask any
contribution of Pb mobilized by exfoliation.

5.8.13 Summary

• Pb substitutes for calcium and is readily taken up and stored in the bone of experimental
animals, potentially allowing bone cell function to be compromised both directly and
indirectly by exposure.

• Relatively short term exposure of mature animals to Pb does not result in significant
growth suppression, however, chronic exposure to Pb during times of inadequate nutrition
have been shown to adversely influence bone growth, including decreased bone density,
decreased trabecular bone, and growth plates.

• Systemically, Pb has been shown to disrupt mineralization of bone during growth, to alter
calcium binding proteins, and to increase calcium and phosphorus concentration in the
blood stream, in addition to potentially altering bone cell differentiation and function by
altering plasma levels of growth hormone and calciotropic hormones such as vitamin DS
[1,25-(OH)2D3].

• Bone cell culture studies have indicated that Pb is primarily taken up by osteoclasts and
likely perturbs intracellular calcium homeostasis secondary to osteoclastic bone resorption.

• Exposure of bone cell cultures to Pb has been shown to impair vitamin D-stimulated
production of osteocalcin, inhibit secretion of bone-related proteins such as osteonectin
and collagen, and suppress bone cell proliferation, potentially by interference with such
factors as Growth Hormone (GH), Epidermal Growth Factor (EGF), Transforming Growth
Factor-Beta I(TGF-PI), and Parathyroid Hormone-related Protein (PTHrP).
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• Periods of extensive bone remodeling, such as occur during weight loss, advanced age,
altered metabolic state, and pregnancy and lactation are all associated with mobilization of
Pb stores from bone of animals.

• Several animal studies have suggested Pb stored in bone can serve as a continuous,
endogenous source of exposure for an individual or can be transferred from mother to
offspring during pregnancy and/or lactation, with potentially toxic consequences.

• During pregnancy, transfer of Pb from mother to offspring has been documented, however,
available evidence suggests a more significant transfer from mother to offspring occurs
during lactation when the concentration of Pb in mother's milk can be several times higher
than corresponding blood levels.

• Despite the extensive remodeling of bone that occurs during growth and development of
young animals, a significant amount of Pb can be accumulated and retained during times
of exposure.

• Pb substitutes for calcium and is readily taken up and incorporated into the developing
teeth of experimental animals.

• Unlike bone, teeth do not undergo remodeling per se and, with few exceptions, most Pb
incorporated into tooth structure remains essentially in a state of permanent storage.

• Administration of high doses of Pb to animals has demonstrated the formation of a "lead
line," visible in both the enamel and dentin and localized to areas of recently formed tooth
structure. Within this Pb line, areas of inhibition of mineralization are evident in enamel
and dentin.

• Pb has been shown to decrease cell proliferation, intracellular protein, procollagen type I
production, and osteocalcin in human dental pulp cells in culture.

• Studies of Pb exposure in adult rats have reported inhibition of post-eruptive enamel
proteinases, delayed teeth eruption times, and decreased microhardness of surface enamel.

• During the process of enamel formation, Pb is apparently widely dispersed when first
incorporated into the developing apatite crystal; however, post-formation, Pb is capable of
entering and concentrating in areas of calcium deficiency within the enamel.

• Numerous epidemiologic studies and, separately, animal studies (both post-eruptive Pb
exposure and pre- and perinatal Pb exposure studies) have suggested that Pb is a
caries-promoting element; however, whether Pb incorporation into the enamel surface
compromises the integrity and resistance of the surface to dissolution and, ultimately
increases risk of dental decay, is unclear.

• No animal studies have examined the role exfoliation of the primary dentition in release of
Pb previously stored in tooth structure, though it is likely this process could serve as an
additional source of Pb exposure in childhood.
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5.9 EFFECTS OF LEAD ON THE IMMUNE SYSTEM
5.9.1 Introduction
The immune system, along with the nervous system, has emerged as one of the more
sensitive targets of Pb-induced toxicity. However, because Pb exposure at low to moderate
levels does not produce overt cytotoxicity of immune cells, immune-associated health effects
result from misregulation and shifts in functional capacity rather than profound lymphoid
deficiencies. As a result, the most sensitive biomarkers of Pb-induced immunotoxicity are those
associated with specific functional capacities as opposed to measures of cell enumeration and/or
lymphoid organ pathology. This distinguishes Pb from some other types of immunotoxicants.
The following sections provide a survey of the reported immune effects resulting from exposure
to Pb in humans and animal models. In general, the focus is on those studies that have been
reported since the 1986 Lead AQCD (U.S. Environmental Protection Agency, 1986) and have
altered our understanding of Pb-induced immunotoxicity.

5.9.2 Host Resistance
Host resistance to disease has been used as an effective measure of the impact of
environmental toxicants on immune function. Because different diseases require different
combinations of immune effector functions for host protection, analysis of environmental
modulation of host resistance across a spectrum of diseases can help identify clinically relevant
immunotoxicity.
The 1986 AQCD presented a range of studies in which exposure to Pb inhibited host
resistance to disease. Since the time of that report, few new infectious diseases have been added
to the list of those that Pb is known to influence. Instead, a much broader understanding of the
likely basis for the increased disease susceptibility to these pathogens has become evident.
Additionally, recognition of an increased risk for some atopic and autoimmune diseases arising
from Pb-induced immunotoxicity has occurred in recent years. This is discussed under
Section 5.9.8. Lead-induced alterations of host resistance against infectious and neoplastic
diseases are considered in the following sections.
To date, there has been either no effect or an increased susceptibility to disease resulting
from exposure to Pb for virtually every infectious agent examined. Given the capacity of Pb to
shift immune responses toward Th2, one might expect that enhanced resistance might occur for
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diseases where robust Th2 responses were required. For example, an increased resistance
against helminth parasitic disease might be hypothesized. However, this possible association has
not been widely examined to date.

5.9.2.1 Viral Diseases
In general, exposure to Pb increases the susceptibility to viral infections. Studies include
host resistance directed against the encephalomyocarditis virus (Gainer, 1977; Exon et al, 1979),
Langat virus (Thind and Khan, 1978), and Semliki Forrest virus (Gupta et al., 2002). In the last
example, oral dosing of Swiss mice with Pb acetate (250 mg/kg for 28 days) significantly
increased mortality to sublethal doses of the virus. Ewers et al. (1982) reported that occupational
exposure to Pb resulted in an increased incidence of influenza cases among workers. In chickens
administered Pb acetate orally (20 and 40 mg/lOOg body weight) for 56 days, antibody
production against Newcastle virus vaccine was reduced, while mortality against viral challenge
was increased (Youssef, 1996). It seems likely that the reduced Thl capacity (including
effective CTL generation) combined with increased TNF-a, ROI, and prostaglandin E2 (PGE2)
production by responding macrophages would contribute to increased tissue pathology but
reduce viral clearance for many infections.

5.9.2.2 Bacterial Diseases
Most of the Pb-associated host resistance research has been conducted on bacterial
diseases. Hemphill et al. (1971) first described the increased susceptibility of mice exposed to
Pb (250 jig given i.p. for 30 days) to Samonella typhimurium, while Selye et al. (1966) reported
increased susceptibility of rats to bacteria endotoxins. Cook et al. (1975) found increased
susceptibility of Pb-exposed rats (2 mg/lOOg body weight given i.v. once) to both Eschrichia coli
and Staphylococcus epidermidis.
The vast majority of studies have been conducted using the intracellular bacterium,
Listeria monocytogenes, in mice. Listeria infection and host resistance to the disease have been
well characterized. Essentially, this infection requires an effective antigen presentation
(probably involving toll-like receptor 2 involvement), a robust response by activated
macrophages leading to interlukin-12 (IL-12) and interferon-y (IFN-y) production and robust Thl
driven host protection (Torres et al., 2004; Lara-Tejero and Pamer, 2004). Ideally, activated
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macrophages would produce NO in an effective response against Listeria (Ito et al., 2005).
In the case of Pb-induced immunotoxicity, everything works against this type of response. First,
macrophages have severely suppressed NO production. Yet, overproduction of TNF-a, ROIs
and PGE2 leads to tissue inflammation and damage. The skewing of the response toward Th2
means that both IL-12 and IFN-y are lacking. Excessive production of IL-6 and other
pro-inflammatory cytokines results in what has been termed "sickness behavior" which involves
both the immune and central nervous systems (Dantzer et al., 1998; Dyatlov et al., 1998a,b;
Lawrence and Kim, 2000; Dyatlov and Lawrence, 2002). Lead-induced impairment in host
resistance to Listeria was reported by Lawrence (1981). CBA/J mice exposed orally to 80 ppm
or greater of Pb acetate for 4 weeks had 100% mortality (after 10 days) compared with no
mortality for mice exposed to 0 or 16 ppm Pb.
In an important study concerning individual variation in Pb-induced immunotoxicity and
host resistance, Kim and Lawrence (2000) demonstrated that neurological circuitry as it pertains
to brain-lateralized behavior could impact the effect of Pb on immune responses and host
resistance to Listeria. Not surprisingly, this suggests that host genotype and epigenetic factors
can be influenced by Pb exposure to the individual. Using female BALB/c mice, Kishikawa
et al., (1997) demonstrated that exogenously administered recombinant IL-12 (1 jig each for
three days i.p.) could enhance production of IFN-y as well as host resistance to Listeria in
Pb-exposed (2 mM in water for 3 weeks) mice. However, Pb-exposed mice continued to have
excess IL-6 production (part of the sickness behavior phenotype). The result with IL-12
validates the importance of the Th skewing and macrophage impairment induced by Pb on host
resistance to certain diseases.
Additional bacterial infections in which Pb exposure has been reported to reduce host
resistance include Serratia marcesens (Schlipkopter and Frieler, 1979) and Pasteurella
multocida (Bouley et al., 1977).

5.9.2.3 Parasitic Diseases
Few studies have been conducted to date regarding the effects of Pb on host resistance to
parasitic diseases. This is unfortunate as some parasitic disease challenges require effective Th2
responses for optimal resistance. Hence, it is not clear that Pb exposure would depress host
resistance in every case (e.g., for helminth parasites). Since the 1986 Lead AQCD, one study
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was conducted examining the effect of Pb on the killing ability ofLeishmania enriettii parasites
in vitro by mouse macrophages (Maue'l et al., 1989). The authors found that 30 to 100 mM of Pb
acetate interfered with the killing ability of macrophages without producing macrophage
cytotoxicity.

5.9.2.4 Tumors
The primary study concerning tumor immunity/tumor growth and Pb was already known
at the time of the 1986 AQCD. In this study, male C57B1/6 mice were exposed to Pb acetate in
the drinking water at concentrations of 0, 13, 130, or 1300 ppm. Moloney sarcoma virus
(MSV)-induced tumor formation and growth were compared following the exposure of mice to
Pb for 10 to 12 weeks. MSV-induced transplantable tumors were also used in this study.
Primary tumor growth was enhanced in animals that received 130 and 1300 ppm of Pb versus the
control. Still, all tumors regressed eventually. Most other studies involving Pb exposure and
tumors describe the fact that Pb can exacerbate the ability of other toxins to promote tumor
formation (Kobayashi and Okamoto, 1974; Hinton et al., 1979). Much of the tumor-promoting
activity of Pb would seem to involve depressed Thl and macrophage function, as well as the
promotion of excessive ROI release into tissues.

5.9.3 Humoral Immunity
The irony of Pb as an immunotoxicant is that the overall effects on humoral immunity are
reasonably modest compared to those reported for macrophages and T lymphocytes (McCabe
1994). McCabe et al. (1991) discussed the fact that Pb is not profoundly cytotoxic for most
immune cells yet can cause major functional shifts within the immune system as well as
decreased host resistance to disease. In many cases, antibody production can remain robust
in Pb-exposed animals and humans. However, the nature and spectrum of the antibodies
produced is the more significant cause for concern. Lead appears to alter the course of T
lymphocyte-driven B cell maturation such that class switching may be skewed in Pb-exposed
animals and humans. If Pb dosage and duration of exposure is sufficient, antibody production
may be depressed overall. However, with low-level Pb exposure, skewed isotype production is
the greater health risk.
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5.9.3.1 General Effects on B lymphocytes and Immunoglobulins
Despite the fact that T lymphocytes and macrophages appear to be the more sensitive
targets of Pb, the metal can alter B lymphocyte maturation and shift immunoglobulin production.
The 1986 AQCD describes the fact that some early studies reported no effect of Pb on antibody
production (Reigart and Graber, 1976; Ewers et al., 1982), while others reported a significant
decrease in the humoral immune response (Roller, 1973; Roller and Rovaic, 1974; Blakley et al.,
1980). In retrospect, this apparent discrepancy may have been caused by the various different Pb
concentrations administered, as well as by variations in the duration of exposure. Also, as
mentioned in the 1986 AQCD, the temporal relationship of Pb exposure to antigen challenge
may be important.
In studies measuring generation of plaque forming cells (PFCs) against sheep red blood
cells (SRBCs), Pb incubation with lymphocytes in vitro caused an increased response (Lawrence,
1981b). In a comprehensive study using several strains of mice, Mudzinski et al. (1986) reported
that Pb acetate administered in the drinking water (10 mM for 8 weeks) elevated the response in
the one strain (BALB/c mice) but failed to alter the humoral response to SRBCs (either PFCs or
antibody liters) in all other strains. McCabe and Lawrence (1990) reported that Pb caused an
elevation in B cell expression of Class II molecules, thereby influencing B cell differentiation.
Lead seemed to impact Class II molecule density at the cell surface via the levels of mRNA
translational and/or the posttranslational stages of cell surface protein synthesis (McCabe et al.,
1991).
Some human epidemiological and occupational studies have reported Pb-associated
differences in levels of circulating immunoglobulins. However, Tryphonas (2001) discussed the
pitfalls of relying on total serum immunoglobulin in assessing immunotoxic effects in humans.
Sun et al. (2003) reported that immunoglobulin M (IgM) and immunoglobulin G (IgG) were
lower but that IgE was higher among females within their high-Pb group. Basaran and Undeger
(2000) found that IgM, IgG, and some complement proteins were reduced among battery
workers with high Pb exposure. Results of Undeger et al. (1996) were similar as well.
In contrast, Sarasua et al. (2000) reported an elevation in immunoglobulin A (IgA), IgG, and
IgM associated with environmental Pb exposure. Pinkerton et al. (1998) found no major effects
but reported a significant Pb-associated decline in serum IgG and an elevation in B cell
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percentage. In a human in vitro study, Borella and Giardino (1991) showed that Pb exposure
caused an increased IgG production following stimulation of cells with pokeweed mitogen.
In more recent animal studies, Miller et al. (1998) and Chen et al. (1999) reported no
effect on antigen-specific IgG liters against keyhole limpet hemocyanin (KLH) protein in F344
strain rats that had been exposed in utero to Pb (0-500 ppm Pb acetate in drinking water).
It seems likely that Pb exposure may be capable of reducing serum immunoglobulin levels
given sufficient dose and duration of exposure. However, the more critical issue pertains to the
distribution of class and subclass of immunoglobulins produced after Pb exposure. Because Pb
can alter the development of T cells involved in specific antigen responses, this can impact the
spectrum of immunoglobulins produced in response to T-dependent antigens. As discussed in
the following section, production of IgE (a class of immunoglobulin that is poorly represented in
serum but of great clinical significance) is a central issue for Pb-induced immunotoxicity. One
additional health concern is the potential for Pb to enhance the likelihood of autoantibody
production (Lawrence and McCabe, 2002; Hudson et al., 2003). This latter concern is discussed
under Section 5.9.8.

5.9.3.2 IgE Alterations
One of the three predominant hallmarks of Pb-induced immunotoxicity is an increase in
IgE production. This can occur in the context of antigen-specific responses or as measured by
total serum IgE. For this endpoint, the human and animal findings are very similar. Virtually all
of the information concerning the capacity of Pb to elevate IgE production in humans and
animals has been obtained since the 1986 AQCD. As a result, this represents a relatively new
biomarker for Pb-induced immunomodulation, and one not included in most animal or human
studies conducted prior to 1990 (e.g., Wagerova et al., 1986).
Table 5-6 lists the studies reporting Pb-induced elevation of IgE. The disease
implications of Pb-induced increases in IgE production are potentially significant and may help
to address, in part, the allergy epidemic that has occurred in the last several decades (Isolauri
et al., 2004). A relationship has been established between relative Th2 cytokine levels, serum
IgE levels, and the risk of allergic airway inflammation (Maezawa et al., 2004; Cardinale et al.,
2005). In fact, attempts to manage allergic inflammation use IgE as one of the major targets
(Stokes and Casale, 2004). IgE levels are directly related to the production of Th2 cytokines
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Table 5-6. Recent Studies Reporting Lead-Induced Increase in IgE
Species
Human
Human
Human

Mouse
Human
Rat
Mouse
Human
Strain/Gender
Both genders
Both genders,
91% males
Females

Balb/c males
and females
Both genders,
56% male
F344 females
Balb/c females
Males
Age
Children
Adult
Children

Fetal
Juvenile
Embryo -
fetal
Adult
Adult
In vivo
Ex vivo
Yes
Yes
Yes

Yes
Yes
Yes
Yes
Yes
Lowest
Effective Dose
Not available
Not Available
Not Available

0.1 mM
Not Available
100 ppm
50 |j£ 3x per
week s.c.
Not Available
Exposure
Duration
Not Available
Not Available
Not Available

3 days
Not Available
5 weeks to dam (2
and 3 gestational)
3 weeks
Not Available
Reference
Karmaus
et al. (2005)
Heo et al.
(2004)
Sun et al.
(2003)
Snyder et al.
(2000)
Lutz et al.
(1999)
Miller et al.
(1998)
Heo et al.
(1996)
Horiguchi
etal. (1992)
such as interlukin-4 (IL-4), among others (Tepper et al., 1990; Burstein et al., 1991; Carballido
et al., 1995; Takeno et al., 2004; Wood et al., 2004). The relationship between Th2 cytokines
(e.g., IL-4), IgE levels, and allergic airway disease is supported through various pharmacological
interventions in both animals and humans that either induce Th2 cytokine and promote allergic
airway disease (Wu et al., 2004) or interfere with Th2 cytokine-driven IgE production and inhibit
allergic inflammation (Holgate et al., 2005; Ban and Hettich, 2005). The production of IgE is of
importance in terms of potential inflammation. Not only is the level of IgE a consideration, but
also the expression of the Fc receptor for the epsilon (e) chain of IgE on mast cells and basophils.
In humans, Karmaus et al. (2005) reported a positive association of blood Pb levels with
serum IgE concentration among second grade children living near a waste incinerator or other
lead-emitting industries. Sun et al. (2003) also found a positive association of blood Pb and
serum IgE levels among children in Taiwan. Lutz et al. (1999) reported a correlation of blood
Pb levels and serum IgE levels in children in Missouri from 9 months-6 years of age. This
association appears to hold not only for children but also for adults. Heo et al. (2004) recently
showed that battery workers with blood Pb >30 |ig/dL differed significantly in serum IgE levels
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from those with blood Pb <30 |ig/dL. Additionally, serum IgE concentration correlated with
blood Pb among the populations examined (r = 0.0872).
Animal data support this relationship between blood Pb concentration and IgE level and
further suggest that even very low-level Pb exposure early in development may produce elevated
IgE production in the juvenile offspring. Miller et al. (1998) found that gestational exposure of
rats to 100 ppm Pb acetate in the drinking water could produce elevated IgE in the adult
offspring. Snyder et al. (2000) showed that gestational and/or neonatal exposure of mice to Pb
acetate produced neonatal blood Pb levels that were not above background (5.0 |ig/dL) but still
could result in elevated IgE production in the juvenile mouse. In most cases, Pb exposures
associated with elevated IgE were also associated with increases in IL-4 production by T cells
(Chen et al., 1999; Snyder et al., 2000). This is consistent with the fact that high IL-4 production
can predispose B lymphocytes to undergo a specific class switch for the production of IgE.
For NK cells, activation can occur through various pathogenic components such as double
stranded RNA. However, recently Borg et al. (2004) showed that mature dendritic cells
produced a Thl-promoting cytokine, IL-12, and this in turn activates NK cells to produce the
further Thl-promoting cytokine, IFN-y. Interleukin-18 (IL-18) produced by macrophages is also
an activator of NK cells, facilitating Thl-promoting cytokine release while interleukin-2 and
interleukin-15 (IL-2, IL-15) are growth factors for NK cells. NK cells would appear to be
relatively resistant to the effects of Pb compared to some T lymphocytes and macrophages.
For a detailed consideration of the effects of Pb on NK cells, see Section 5.9.7.
Cytotoxic T lymphocytes are generated in response to antigen presentation delivered with
Thl cytokines. These cells are capable of mediating antigen specific destruction of neoplastic
and virally infected cells via binding and release of cytolytic proteins into the intracellular space.
Frequently, the most effective antigen targets of CTLs are the early viral proteins produced in the
first phase of host cell infection by viruses. IL-12, produced largely by dendritic cells, appears to
be important in the generation of antigen CTL cells and IFN-y produced by Thl lymphocytes.
NK cells are a potent regulator of CTL activity. Cell signaling via certain toll-like receptors on
antigen presenting cells seems to have a role in determining the nature of the Th activation (Thl
versus Th2) and can, therefore, influence the extent of CTL production.
Because T lymphocytes and their regulator and effector functions are so critical in CMI,
the maturation of thymocytes within the thymus microenvironment and the selection of
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repertoire among the maturing T lymphocytes are crucial issues for potential developmental
immunotoxicants. In fact, Pb seems to be capable of disrupting several aspects of T cell
maturation, activation, and repertoire usage (McCabe and Lawrence, 1991; Heo et al., 1998;
Miller et al., 1998; McCabe et al., 2001, Lee and Dietert, 2003).

5.9.4 General Effects on Thymocytes and T lymphocytes
In general, cells of the T cell lineage appear to be relatively sensitive to the toxic effects
of Pb compared to other lymphoid populations. At the time of the 1986 Lead AQCD, there was
some understanding of this sensitivity. However, there appear to be considerable differences in
sensitivity across various T cell subpopulations (McCabe and Lawrence, 1991; Heo et al., 1996;
1997; 1998). This was largely unknown when the 1986 AQCD was prepared, as the partitioning
of T helper cells into functionally distinct subpopulations (e.g., ThO, Thl, and Th2) was not
known until the latter part of the 1980s. The differential impact of Pb on T helper cell
populations and on immune balance was established during the 1990s. This has become one of
the hallmarks of Pb-induced immunotoxicity.
Original observations of both in vivo and in vitro T-dependent immune responses in the
presence of Pb suggest that T helper function, as well as the spectrum of cytokines produced, are
skewed toward the Th2. The cytokine skewing is discussed as well in Section 5.9.5.3. Smith
and Lawrence (1988) have shown that Pb can inhibit antigen presentation and stimulation of a
T cell clone of the Thl phenotype. McCabe and Lawrence (1991) were the first to show that this
was caused by the novel capacity of Pb to inhibit Thl stimulation while promoting presentation
to Th2 clones. Heo et al. (1996) provided both in vitro and in vivo results supporting this Pb
immunomodulation. Cytokine skewing accompanied the differential stimulation of Th cells.
Using naive splenic CD4+ T cells derived from D11.10 ovalbumin-transgenic mice, Heo
et al. (1998) developed T cell clones in vitro in the presence of Pb. The authors found the T cells
that developed from the naive precursors were significantly skewed toward the Th2 helper
phenotype and away from the Thl phenotype. If IL-4 was inhibited with the addition of
anti-IL-4 to the cultures or if the Thl- promoting cytokine IL-12 was added exogenously to the
culture, the effects of Pb could be largely overcome. This study provided firm evidence that Pb
can directly promote Th2 development among precursor Th(0) cells and impair development of
Thl cells. Among its effects, Pb enhanced adenyl cyclase activity and increased the levels of
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cAMP. The authors suggested that Pb may influence cell signaling in such as manner as to
promote the Th2 pathway.
Beyond the biasing of immune responses at the level of the T lymphocyte based on
Thl/Th2 balance, Pb has the capacity to bias usage of certain VP genes (VP5, VP?, and VP13)
among T lymphocyte clones in mice (Heo et al., 1997). This is of concern, as it suggests that
exposure to Pb may alter the T cell repertoire and skew its representation. Heo et al. (1997)
discussed the fact that many autoimmune diseases are characterized by a disproportionate usage
of certain VP genes. Different autoimmune conditions are associated with the differential
overabundant usage of a specific VP gene. They suggest that this feature of Pb-induced
T lymphocyte immunotoxicity may contribute to and enhance the risk of autoimmunity.
Lee and Dietert (2003) exposed the developing thymus of embryonic day 12 (E12)
chickens to Pb acetate (single injection of 400 jig) and evaluated the capacity of thymocytes
(ex vivo) from juvenile chickens to produce IFN-y. They found that embryonic exposure at
doses that impair juvenile delayed type hypersensitivity (DTH) also inhibit IFN-y production.
Similarly, IFN-y production was decreased when thymocytes from juvenile chickens were
exposed to Pb in vitro (0.45 |iM). However, in vitro exposure of thymic stroma to Pb did not
result in suppression of control thymocyte IFN-y production in co-cultures. There is a suggestion
that the balance of reproductive hormones in early life may influence the impact of Pb on
developing thymocytes (Hussain et al., 2005).

5.9.4.1 Delayed Type Hypersensitivity
The DTH assay is an in vivo assay requiring antigen-specific T lymphocytes to be primed,
expanded, and then recruited to a local site of antigen deposition. The most common application
of the DTH is the tuberculin assay for TB in humans. The assay has a long history of application
in immunotoxicology, and its utility within the national toxicology program assessment in the
mouse has been previously reported (Luster et al., 1992). The assay is known to depend largely
on Thl participation and is, therefore, an effective measure of Thl-dependent function.
However, there are at least two different portions of the response that are under somewhat
separate control. Priming and expansion of the antigen-specific T lymphocytes is largely Thl
dependent. However, recruitment of T lymphocytes to the periphery involves a variety of locally
produced chemotactic signals that may not be under the same regulation. In fact, Chen et al.

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(1999) showed that a commonly used chelator for Pb poisoning (succimer, meso-2,
3-dimercaptosuccinic acid [DMSA]) fails to restore Pb-induced suppression of DTK in rats,
because the chelator itself somehow interferes with the production of chemotactic factors
necessary for T lymphocyte recruitment. The DTH assay is also generally useful in questions of
possible developmental immunotoxicity, because of the natural skewing toward Th2 that occurs
during gestation through birth and the issue of effective Thl functional acquisition in the
newborn.
Lead-induced suppression of the DTH response is another hallmark of Pb-induced
immunotoxicity. At the time of the 1986 AQCD, the capacity of Pb to suppress DTH function
was already known from two studies conducted during the late 1970s. However, the association
of the function with Thl help had not been established. Muller et al. (1977) were among the first
to demonstrate Pb-induced suppression of DTH. Using mice, these investigators administered
Pb acetate i.p. for 30 days prior to assessment of primary and secondary DTH responses against
SRBCs. Both primary and secondary responses were severely depressed following exposure to
Pb, even at the lowest dose tested (0.025 mg). Faith et al. (1979) exposed developing Sprague-
Dawley rats to Pb acetate in the drinking water (lowest dose at 25 ppm) first via the dams during
gestation and through weaning and then with direct exposure of the offspring until 6 weeks of
age. In this case, the purified protein derivative (PPD) of tuberculin was used as the antigen
compared against the saline injection control. Rats administered the lowest dose of Pb evaluated
(producing a BLL of 29.3 |ig/dL) had a significantly reduced DTH response. Laschi-Loquerie
et al. (1984) measured the contact hypersensitivity reaction against picryl chloride in mice that
had received 0.5 mg/Kg Pb via s.c. administration. Lead administration was given from 3-6 days
in duration at varying times relative to the sensitization period. These investigators reported that
Pb suppressed the DTH type of response regardless of the window (before or during
sensitization) in which it had been administered.
More recently, Miller et al. (1998) found that female F344 rats gestationally exposed to
250 ppm of Pb acetate in drinking water had a persistently reduced DTH reaction against KLH
protein. Chen et al. (1999), Bunn et al. (2001a,b,c) and Chen et al. (2004) had similar findings in
studies that included both the F344 and CD strains of rats. In the last study conducted in F344
rats, a BLL of 6.75 |ig/dL at 4 weeks of age, postgestational exposure to Pb acetate (250 ppm in
drinking water) was associated with depressed DTH against KLH in the 13-week-old adult
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female offspring (Chen et al., 2004). McCabe et al. (1999) were among the first to draw
attention to the relationship between Pb-induced suppression of DTK and the prior observations
of Pb-induced Th skewing. These authors gave varying doses of Pb acetate in drinking water
(32,128, 512, 2048 ppm) to female BALB/c mice for 3 weeks prior to measuring the DTH
against SRBCs. They found that the 512 ppm dose producing a BLL of 87 |ig/dL significantly
impaired the DTH response. Antigen routes proved to be important as Pb depressed DTH when
an i.v. primed with SRBCs was used, but not when SRBCs were administered i.p. Timing of Pb
administration was found to be important relative to the capacity to depress the DTH response.
Lee et al. (2001) showed that Pb acetate (200 jig) administered in ovo to chicken embryos at
9 days of incubation failed to depress juvenile DTH against bovine serum albumin (BSA), but
when the same dose of Pb was administered 3 days later producing the same BLL, juvenile DTH
was severely reduced. Using the latter model, embryonic administration of exogenous thymulin
was found to partially restore juvenile DTH function following embryonic exposure to Pb (Lee
and Dietert, 2003).
With regard to developmental sensitivity of the DTH response to Pb-induced
immunosuppression, parallel findings were obtained in the developing rat (CD strain females)
(Bunn et al., 2001c) in agreement with those found in the chicken. Administration of 500 ppm
Pb acetate during gestational days 3 to 9 or 15 to 21 produced no DTH effect compared with
DTH suppression in the corresponding adult offspring. As shown in Figure 5-17, the sensitivity
of the DTH response to Pb appears to develop sometime between days 9 and 15 of rat embryonic
development. Apparently, the status of the developing thymus may be a consideration in the
capacity of Pb to impact the subsequent DTH response, as discussed further in Section 5.9.10.
It should be noted that in several studies, Pb-induced suppression of the DTH response
was associated with reduced capacity to produce the Thl cytokine, IFN-y (Chen et al., 1999;
Lee etal., 2001).

5.9.4.2 Other T-Dependent Cell-Mediated Immune Changes
The in vitro response of T lymphocyte populations to various mitogens (e.g.,
Concavavalin A [ConA], Phytohemagglutinin A [PHA]) has been used as a surrogate measure of
antigen-driven T lymphocyte stimulation. The impact of Pb on these parameters is presented in
Section 5.9.5. Another T cell response altered by exposure to Pb is the mixed lymphocyte
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Rat
Conception
0 3
18
Birth
21
13-14
1 post
Chicken
Start of
incubation
0 3
Day of Embryonic Development
Figure 5-17. Windows during prenatal development (days postconception for rat) or
embryonic development (days postincubation initiation for chicken) during
which sensitivity of DTH to lead emerges.
response (MLR). This in vitro assay is a measure for the responsiveness of T cells to the
presentation of allogeneic major histocompatibility complex (MHC) molecules by antigen
presenting cells. The in vivo correlate of the MLR is usually considered to be the graft versus
host (GvH) reaction. Several investigators have reported Pb alteration of the MLR as
summarized in Table AX5-9.4.
McCabe et al. (2001) demonstrated that Pb at very low physiological concentrations
(0.1 jiM or approximately the equivalent of 10 |ig/dL) in vitro significantly enhanced the
proliferation and expansion of murine alloreactive CD4+ T lymphocytes in the MLR reaction.
In fact, the resulting population was found to have a high density of CD4 molecules on
the cell surface, making them phenotypically similar to memory T lymphocytes. The authors
hypothesized that Pb-induced creation of an exaggerated pool of memory-type T lymphocytes
(possessing a lower threshold required for subsequent activation) would be problematic for the
host. In a study using Lewis strain rats, Razani-Boroujerdi et al. (1999) also found evidence for
Pb-induced stimulation of the in vitro MLR response. In this case, both the alloreactive mixtures
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of cells as well as syngeneic mixtures were elevated in proliferation when cultured in the
presence of Pb acetate (e.g., 50 ppm or approximately 131 jiM). When concentrations of Pb
were significantly higher (200 ppm or greater), proliferation was inhibited in these cultures.
Figure 5-17 illustrates the developmental appearance of initial sensitivity for Pb-induced
suppression of the DTH function. The mid-embryonic developmental window is the time during
which the capacity of Pb to impair later-life DTH responses first emerges. Earlier pulsed
exposure to Pb fails to impair juvenile and/or adult DTH despite the continuing of presence of Pb
in the embryo. However, during the second half of embryonic development the embryo becomes
remarkably sensitive to Pb-induced suppression of DTH. Both the rat and chicken are similar in
this window of emerging Thl-dependent functional sensitivity. Thymus-related developmental
events are indicated along with the emergence of DTH functional sensitivity to Pb. Information
was derived from Gobel (1996), Vicente et al. (1998), Dunon et al. (1998), Dietert et al., (2000),
Bunn et al. (2001c), Lee et al. (2001) and Holsapple et al. (2003).

5.9.5 Lymphocyte Activation and Responses
Many of the broader functional ramifications of Pb exposure on lymphocytes are
discussed under Sections 5-9.3 and 5-9.4. However, the capacity of Pb to directly alter lymphoid
responses is a significant component of Pb-induced immunotoxicity and is summarized within
the present section. Lymphoid responses are usually assessed in terms of proliferation and
activation (functional changes). One of the recent endpoints reflecting functional status is the
production of cytokines. These both autoregulate the producing cells and significantly impact
the activity of other immune and nonimmune cells carrying the appropriate receptors. The
spectrum and levels of cytokines produced by a population of immune cells tends to reflect their
capacity to regulate the host immune response.

5.9.5.1 Activation by Mitogens
The capacity of certain plant- and bacterially derived products to stimulate lymphoid
populations to enter the cell cycle and undergo mitogenesis has been used for decades to assess
the potential capacity of lymphocytes to receive proliferation signals and expand their
population. Among the mitogens employed within the Pb exposure studies are the T lymphocyte
subpopulation mitogens, PHA and Con A; the dual T and B cell mitogen, pokeweed mitogen
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(PWM); the B lymphocyte mitogen derived from gram-negative bacteria, lipopolysaccharide
(LPS), and the B cell mitogen, Staphylococcus aureus enterotoxin (SE). It should be noted that
these mitogens do not necessarily stimulate all T lymphocytes or all lymphocytes but, instead,
stimulate selected populations of the cells. The mitogens react with a large array of cell surface
molecules producing cross-linking and appropriate signal transduction to initiate mitogenesis.
In the case of the plant-derived mitogens, lectins, numerous glycoproteins and glycolipids
carrying the correct carbohydrate residues serve as the cell surface binding sites for cross-
linking. Mitogen stimulation in vitro has been used as a surrogate for antigen-driven stimulation
and proliferation of antigen-specific T and B cell clones. However, it should be noted that while
the assays have been used for decades, there are now more specific assays utilizing more
functionally relevant cell surface receptors to assess lymphoid activation potential.
The 1986 AQCD has an extensive review of mitogenic responses of lymphocytes
following both in vivo and in vitro treatment by Pb. The results at that time showed no clear
pattern. At low to moderate levels, Pb was potentially co-mitogenic for some cells and at very
high concentrations could suppress proliferation. Little has changed in conclusions for this
assessment measure since the 1986 AQCD. The most significant findings from the mitogenic
studies are that at doses encountered physiologically Pb is not a potent cytotoxic agent for most
immune cells. At low concentrations, it can marginally stimulate lymphoid mitogenesis.
However, as one examines more refined subpopulations of lymphocytes than what were able to
be identified prior to 1986 (e.g., Thl versus Th2 clone of T lymphocytes), it becomes clear that
Pb can promote expansion of some lymphoid populations while suppressing others.
Annex Table AX5-9.5 for this section summarizes results of Pb effects on mitogen-
stimulated proliferation of lymphoid populations.

5.9.5.2 Activation via Other Receptors
In recent years, lymphoid activation and population expansion has been measured using
the triggering of specific T and B cell surface receptors (e.g., CDS on T cells) as well as
antigen-driven proliferation of T cell clones known to be specific for the antigen in question.
The latter has provided the opportunity to simulate in vivo lymphoid activation and antigen-
driven proliferation by using receptors in vitro, which are more physiologically relevant than
those activated by plant lectins. Because Pb does not cause profound population loss across the
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entire population of T or B lymphocytes, these more refined and functionally-relevant assay
systems have enabled a much clearer picture to emerge concerning Pb-induced changes in
lymphoid population than was available for the 1986 AQCD report.
Smith and Lawrence (1988) and McCabe and Lawrence (1991) utilized antigen-specific
mouse T clones. They found that Pb directly promoted antigen presentation and stimulation of
the T cell clones when these clones were Th2 cells. However, when the Thl clones were used,
Pb suppressed the antigen-specific presentation signal. In the McCabe and Lawrence study,
direct comparisons were made between Thl and Th2 clones specific for mouse allogeneic MHC
molecules. These studies provided the first clear picture of the differential effects of Pb on Thl
versus Th2 cells. Several studies since these have verified this major effect of Pb (Heo et al.,
1996; 1997, 1998). Many of these later studies utilized the transgenic mouse strain (DO 11.10
OVA-tg) that carries T cells specific for a peptide fragment of ovalbumin. These enabled the
same comparisons to be made with the presentation of a soluble protein antigen as the
stimulating signal. Heo et al. (1998) showed that Pb not only selectively stimulates Th2 cells
and suppresses Thl cells but that it preferentially causes precursor ThO cells to mature into Th2,
rather Thl cells, as well. Additionally, the T cell clones in the presence of Pb are skewed in
terms of their usage of VP genes (as reflected in their cell surface receptors) (Heo et al., 1997).
This is of particular concern relative to the risk of autoimmunity. More recently, McCabe et al.
(2001) examined Pb exposure in the context of the allogeneic MLR against allogeneic MHC
molecules. In vitro exposure to Pb (as low as 1.0 jiM) enhanced the primary MLR response, but
not the secondary MLR response and not the mitogenic response using PHA. Significantly, the
T cell clones that emerged from the primary MLR were in greater proportion than normal and
were of the specialized phenotype CD4-plus high density (CD4+hlgh). Because these fit the
phenotype of memory cells, it is likely that an overabundance of memory cells was produced
during the primary response, where the antigen may be of lesser biological significance than in a
secondary response. The authors discussed the fact that Pb may cause T cells to respond under
conditions of low antigen concentration, which could waste valuable and limited resources by
generating T memory cell clones when they are not needed (against unimportant antigens) or
even increase the risk of autoimmune responses by altering the threshold requirements for
stimulation. The putative mechanisms suggested for differential effects of Pb on Th cells are
presented in Section 5.9.9.
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5.9.5.3 Cytokine Production
At the time of the 1986 AQCD, immune cytokines were essentially absent from the
information available for consideration. Only the antiviral interferons (a,P) had been examined
among studies available for that report. Therefore, one of the most important effects of Pb on the
immune system, i.e., Pb-induced cytokine production was not known at that time.
Most studies since 1986 have shown that Pb exposure at low to moderate levels causes a
significant shift in the production of Thl versus Th2 cytokines with the bias toward the latter.
Hence, production of IFN is decreased and IL-12 is inadequate for effective host resistance.
In contrast, production of IL-4, IL-6, and, frequently, interlukin-10 (IL-10) is elevated.
Table 5-7 illustrates the studies reporting shifts in cytokine production induced by Pb. (Please
note that TNF-a production is considered in the macrophage section, Section 5.9.6). These shifts
in cytokine production are remarkably consistent, occur even at low levels of exposure, and are
reported following both in vivo and in vitro exposure to Pb. Furthermore, the effects are
persistent even when exposure to Pb was restricted to early development and cytokine
assessment was performed in the subsequent juvenile or adult (Miller et al., 1998; Bunn et al.,
200Ic; Lee et al., 2001; Chen et al., 2004).
The only exceptions to Pb-induced biasing in favor of Th2 occur in the reports by Goebel
et al. (2000) and Mishra et al. (2003). In the latter case, the authors attributed this difference (in
humans) to the very high Pb levels considered in the study. In the prior case, Goebel et al.
(2000) saw a local bias to Thl in the intestinal tract of a specialized autoimmune diabetes-prone
strain of mice (NOD) but not in normal mice. Initially, the Pb-induced cytokine skewing favored
Th2 (after 1 day), but this shifted to Thl with more prolonged Pb exposure (after 10 days). Loss
of oral tolerance accompanied this long-term shift. These results suggest that in most cases,
Pb-induced skewing would favor Th2. But with some genotypes or additional disease
conditions, an imbalance may occur in the direction of a gut-associated Thl environment,
increasing risk for loss of oral tolerance and the potential for increased food allergies.
One ramification for the capacity of Pb to promote Th2 cells is the impact of elevated
IL-4 on IgE. It seems clear that Pb-induced overproduction of IgE (seen in virtually all animal
models examined as well as humans) is directly linked with the overproduction of IL-4.
Excessive IL-4 and the resulting IgE production increase the risk for IgE-mediated atopy and
asthma.
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Table 5-7. Studies Reporting Lead-Induced Shifts in Thl versus Th2 Cytokines
to
to
Species
Rat
Human
Chicken
Mouse
Rat
Chicken
Mouse

Mouse

Rat
Strain/Gender
F344 Females
Males
Cornell K females
Balb/c
CD females
Cornell K females
Balb/c male

NOD
Autoimmune strain
adult
C57 Bl/6 females
NOD autoimmune
strain females
F344 females
Age
Embryo-fetal
Adults
Embryonic
Neonatal/
Juvenile
Fetal
Embryonic
Adults

Adult

Embryo-
fetal
Cytokine Alterations
tIL-4
4 IFN-y splenic lymphocytes
t IFN-y
PHA stimulated peripheral
blood lymphocytes
4IFN-y
stimulated thymocytes
tJL-6
serum during infection
tIL-10
4IFN-y
tIL-6
serum during infection in
certain groups
4 IFN-y, no change long
term
4TGF-P intestinal levels
No effect on gut balance in
normal mice
4TGF-P in autoimmune mice
4IFN-y
tIL-10
In vivo/
Ex vivo
Yes
Yes
Yes
Yes
Yes
Yes
Yes

Yes

Yes
Lead Dose/
Concentration
250 ppm in water
to dams
Not available
400 |^g
0.5 mM in water to
dams and their
pups
550ppminwaterto
dams
50 ^g
2mM

Oral 10 mM and
ovalbumin antigen

0.5 mg/kg injection
and oral ovalbumin

250 ppm to dams
Duration of
Exposure
2 weeks prior and 3rd
week of gestation for
dam
Not available
Single injection E12
4 weeks (3 via dams)
6 days via gestation
of dam
Single injection
8 weeks

10 days

6 injections over
2 weeks

2 weeks before and
3rd week of gestation
References
Chen et al.
(2004)
Mishra et al.
(2003)
Lee and
Dietert (2003)
Dyatlov and
Lawrence
(2002)
Bunn et al.
(200 Ic)
Lee et al.
(2001)
Kim and
Lawrence
(2000)
Goebel et al.
(2000)

Goebel et al.
(1999)

Chen et al.
(1999)
-------
Table 5-7 (cont'd). Studies Reporting Lead-Induced Shifts in Thl versus Th2 Cytokines

(^
to
to
ON

Species Strain/Gender
Mouse DO11.10 ova-tg,
ova mice and RAG
knockouts
Rat F344 females
Mouse Balb/c ByJ females

Mouse Balb/c and
DO11. 10 ova-tg
mice
Mouse Balb/c ByJ female
or male
Mouse Balb/c ByJ female
or male

Age
Adult

Embryo-
fetal
Adult

Adult

Adult
Adult

Cytokine Alterations
4IFN-Y

4IFN-Y
4IFN-Y
tIL-6
4IFN-Y
4IFN-Y/IL-4
ratio
4IFN-Y
tIL-4
4IFN-Y
tIL-4
In vivo/ Lead Dose/
Ex vivo Concentration
No 25 nM

Yes 500 ppm to dams
Yes 2mM

Yes 50 |j,g each
injection (s.c.)
3 per week
Yes 50 ng each
injection (s.c.)
3 per week
No 10 |jM - 50 |jM

Duration of
Exposure
3 days

2 weeks before and
3rd week of gestation
3 weeks

2 weeks

2 weeks
2 days

References
Heo et al.
(1998)

Miller et al.
(1998)
Kishikawa
etal. (1997)
Heo et al.
(1997)
Heo et al.
(1996)
Heo et al.
(1996)
-------
Additionally, Kishikawa et al. (1997) demonstrated that administration of the potent
Thl-promoting cytokine, IL-12, to Pb-exposed mice can restore the balance of Thl (IFN-y)
versus Th2 cytokines (e.g., IL-6), reduce corticosterone levels, and enhance host resistance in
Listeria-infected mice. This observation supports the critical role of Thl/Th2 balance in overall
risk to host resistance against disease presented by Pb disruption of that balance

5.9.6 Macrophage Function
Macrophages represent a diverse population of cells that play critical roles in both host
defense and tissue homeostasis. Macrophage subpopulations provide a front line of defense
against bacteria, parasites, viruses, and tumors via the innate immune response. Additionally,
they are important in tissue repair and remodeling as well as in the removal of senescent cells.
Some forms of macrophages are efficient in the processing of antigens and the presentation of
antigen fragments to T lymphocytes. Additionally, macrophages can regulate lymphoid activity
through the secretion of a variety of cytokines and through the production of various
immunomodulatory metabolites (e.g., NO, ROIs) and the products of the cyclooxygenase and
lipoxygenase pathways.
Because macrophages can be found residing in most tissues, Pb-induced modulation of
macrophage functional capacity has the potential to alter overall organ function. Macrophages
originate in the bone marrow from pluripotent stem cells that give rise to both the monocyte-
macrophage lineage as well as polymorphonuclear leukocyte populations. Bone marrow-derived
macrophages mature under the influence of various cytokine growth factors to become the full
array of mature cell subpopulations. Various investigators have examined Pb effects on the
maturation of macrophages in vitro as well as on the functional capacity on fully mature cells
both in vitro and in vivo. Blood monocytes represent a functional, yet not fully specialized, form
of macrophage. As a result, the influence of environmental toxicants on monocytes may not be
fully predictive of the effects of the same toxicants on splenic or alveolar macrophages, glial
cells, or Kupffer cells.
Because macrophages give rise to several specialized populations, e.g., Kupffer cells in
the liver, glial cells in the brain, and various skin macrophage populations, it is important to
realize that different specialized macrophage populations are likely to have somewhat different
sensitivities to Pb, as well as potentially different responses following exposure. Not too
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surprisingly, blood monocytes may not always be an appropriate model to accurately predict the
outcome of Pb-induced immunotoxicity for alveolar macrophages following inhalation exposure.
The 1986 Lead AQCD identified macrophages as a significant target for Pb-induced
immunotoxicity. Research since the mid-1980s has served to underscore this point. The
understanding of Pb-induced alterations in macrophage function has increased significantly since
the 1996 AQCD report. The following sections describe the reported immunotoxic effects of Pb
on macrophages. It should be noted that for a number of endpoints, such as Pb-induced
alterations in the production of NO, ROIs and TNF-a, there is a general consensus among a
majority of immunotoxicology studies and agreement with the effects described for the
cardiovascular system (see Chapter 5.5).

5.9.6.1 Nitric Oxide (NO) Production
Nitric oxide is a short-lived metabolite produced in large quantities by macrophages
during cellular activation. The enzyme responsible is an inducible form of nitric oxide synthase
(iNOS), which, utilizing a bioptrin cofactor, converts the amino acid arginine into NO and
citrulline. A competing alternative pathway utilizing arginine leads to the production of
polyamines, which themselves are immunomodulatory for lymphocytes. Nitric oxide is critical
in the defense against certain infectious agents, including various bacteria.
Among the most sensitive immunomodulatory effects of Pb exposure is the capacity to
impair NO production by macrophages (Table AX5-9.6). Several research groups have shown
that in vitro as well as in vivo exposure to Pb results in significantly reduced production of NO
(Tian and Lawrence, 1995, 1996; Chen et al., 1997b; Lee et al., 2001; Pineda-Z aval eta et al.,
2004 [also reviewed in Singh et al., 2003]). Similar results were obtained in human, mouse, rat
and chicken. Depression of NO production capacity usually occurs shortly after exposure to
lead. However, the long-term effects of Pb on NO production following very early life exposure
are less clear (Miller et al., 1998; Chen et al., 1999; Bunn et al., 2001a).
Tian and Lawrence (1996) have hypothesized that because very low Pb concentrations
(in vitro equivalents to 10 |ig/dL) can impair NO production, impaired NO production may be
responsible for reduced host resistance to Listeria seen among Pb-exposed rodents as well as for
Pb-induced hypertension among humans (Pirkle et al., 1985). Indeed, impaired NO production
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by macrophages seems to be one of the more sensitive endpoints for immediate Pb-induced
immunotoxicity.

Other Functional Alterations
TNF-a Production
Early studies identified the fact that Pb exposure could predispose animals for a
dramatically increased sensitivity to bacterially derived endotoxin (Trejo et al., 1972; Filkins and
Buchanan, 1973; Schlick and Friedberg, 1981).
It is now known that the increased sensitivity to endotoxin is linked to the capacity of Pb
to increase production of TNF-a among macrophages (Dentener et al., 1989; Zelikoff et al.,
1993; Guo et al., 1996; Miller et al., 1998; Chen et al., 1999; Krocova et al., 2000; Flohe et al.,
2002). Studies in mouse, rat, rabbit, and human provide a clear indication that one effect of Pb
on macrophages is to boost production of the proinflammatory cytokine TNF-a. While most
studies examined the immediate effects of Pb exposure on TNF-a production, studies by Miller
et al. (1998) and Chen et al. (1999, (2004) showed that the effects of early gestational exposure
to Pb on macrophages could persist well into later life, including adulthood. Also, Chen et al.
(1999) showed that chelation of Pb with succimer in developing female rats in utero could
eliminate the persistent effect of elevated TNF-a production in the adult offspring. Flohe et al.
(2002) found evidence that Pb-induced elevation in TNF-a production is sensitive to both PKC
signaling as well as to protein production. While the production of TNF-a can be elevated
following exposure to Pb, the expression of the receptor for TNF-a (TNF-R) was also increased
during the in vitro exposure of human blood monocytes to Pb-chloride (Guo et al., 1996).
Therefore, the combined effect of elevated cytokine production by macrophages as well as
increased receptor expression would be expected to contribute to problematic inflammatory
responses.

Production of Other Proinflammatory Cytokines
Several studies have indicated that macrophage production of cytokines (or that levels of
cytokines known to be produced primarily by macrophage populations) is altered after exposure
to Pb. These vary somewhat, depending upon the exposure protocol and the source of
macrophages examined. In addition to the previously discussed elevation of TNF-a by Pb, the
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most significant and consistent Pb-induced effects seem to involve elevated production of the
other major proinflammatory cytokines, interleukin-lp (IL-1(3) and IL-6. Increased production
of IL-6 following exposure to Pb has been reported by Dyatlov and Lawrence (2002), Flohe et al.
(2002), Kim and Lawrence (2000), Krocova et al. (2000), Kishikawa and Lawrence (1998) and
Kishikawa et al. (1997). Because IL-6 is a proinflammatory cytokine, its increased production
following Pb exposure has the potential to influence many different tissues. Dyatlov et al.
(1998a,b) provided evidence that Pb, IL-6 and LPS can combine to exert a significant impact on
the permeability of the blood brain barrier as well as the properties of brain neurons and
endothelial cells. Lead-induced elevation of IL-1(3 production has been reported by Dyatlov and
Lawrence (2002). It is probable that enhanced co-production of IL-1P and IL-6 would increase
the likelihood of local tissue inflammation.

Production of Reactive Oxygen Intermediates (ROIs)
Reactive oxygen intermediates (ROIs) are important metabolites in the capacity of
macrophages and other inflammatory cells to kill invading bacteria and to attack cancer cells.
However, increased overall production or inappropriate triggering of ROI release by
macrophages can be a major contributor to tissue damage and the oxidation of cell surface lipids
as well as DNA. The latter is one mechanism through which improperly regulated macrophages
can actually increase the incidence of cancer. Results from many studies suggest that exposure
of macrophages to Pb can increase the release of superoxide anion and/or hydrogen peroxide at
least shortly after exposure. Key studies are summarized in Table AX5-9.6.
In a recent study on environmentally exposed children in Mexico, Pineada-Zavaleta et al.
(2004) reported that production of superoxide anion by directly activated (interferon-
gamma + LPS) monocytes was directly correlated with blood Pb level. This was in contrast with
the effect of arsenic, which had a negative association. In other studies involving low levels of
exposures, Zelikoff et al. (1993) demonstrated that rabbits exposed to Pb via inhalation had
pulmonary macrophages that produced elevated levels of both H2O2 and superoxide anion upon
stimulation in vitro. In an in vitro study, Shabani and Rabani (2000) reported that Pb nitrate
exposure produced a dose dependent increase in superoxide anion by rat alveolar macrophages.
Baykov et al. (1996) fed BALB/c mice dietary Pb and found that peritoneal macrophages had an
increased spontaneous release of H2O2.
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Other studies have reported no effects of Pb on superoxide anion production when a long
recovery period was included following in vivo exposure (Miller et al., 1998) as well as negative
effects of Pb on oxidative metabolism by certain macrophages or macrophage cell lines
(Castranova et al., 1980; Hilbertz et al., 1986; Chen et al., 1997b). These somewhat different
results suggest that the subpopulations of macrophages examined (e.g., alveolar versus splenic
versus peritoneal) and the timeframe of assessment relative to exposure may be important factors
in the effect of Pb on ROI production.
The biological importance of increased ROI production by Pb-exposed macrophages
should not be underestimated. Fernandez-Cabezudo et al. (2003) showed that the potent
antioxidant, vitamin E, could protect TO strain mice against some Pb-induced
immunosuppressive alterations. Hence, macrophage-associated oxidative damage following
exposure to Pb may be a mitigating factor in nonlymphoid organ Pb-induced pathologies.

Arachidonic Acid Content and Prostaglandin Production
Archidonic acid (AA) is a major surface component of many cells, including
macrophages, and is the precursor of cyclooxygenase and lipoxygenase metabolites. As a
result, the specific AA content of membranes and the capacity of macrophages to produce
immunomodulatory metabolites from AA are important to overall health of the individual. One
of the findings since 1986 concerning Pb-induced modulation of macrophage function is the
impact of Pb on PGE2 production. One study (Knowles and Donaldson, 1990) reported that diets
supplemented with Pb at 500 ppm and fed to chicks produced an increase in the percentage of
AA included in cell membranes. Such an increase would be expected to raise the risk of overall
inflammation.
Several groups have reported that Pb exposure increases macrophage production of the
immunosuppressive metabolite PGE2. Lee and Battles (1994) reported that mouse macrophages
exposed to Pb (10 jiM) in vitro had elevated basal PGE2 production, but under some stimulatory
conditions, had decreased production of PGE2. When Knowles and Donaldson (1997) fed Pb to
turkey poults in the diet at a level of 100 ppm, macrophage production of prostaglandin F2
(PGF2), PGE2 and thromboxane production were all significantly elevated versus the control.
Flohe et al. (2002) showed that exposure of mouse bone marrow-derived macrophages to Pb
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chloride resulted in increased production of PGE2that correlated with increased mRNA
production for the necessary enzyme, prostaglandin H synthase type-2.

Tissue Homeostasis
In an important observation reflecting the impact of Pb-induced immunotoxicity on
nonlymphoid tissues, Pace et al. (2005) showed that neonatal exposure of mice to Pb acetate via
drinking water (0.1 ppm for 6 weeks, both through maternal nursing and direct) produced a
significant reduction in the testicular macrophage population. This correlated with increased
estradiol levels in the testis and reduced male reproductive performance. The authors
hypothesized that Pb-induced alteration among testicular macrophages is linked to an impaired
tissue environment that likely includes increased oxidative stress, apoptotic somatic cells, and
reduced fertility of males.

Colony Formation and Population Distribution
The ability of bone marrow-derived macrophages (BMDM) to form colonies in response
to certain growth factors (e.g., colony stimulating factor-1 [CSF-1]) is a property related to the
growth and differentiation of subsequent macrophage populations. Kowolenko et al. (1991)
found that exposure to CBA/J female mice to Pb acetate (0.4 mM in drinking water for 2 weeks)
reduced colony formation of macrophages in response to CSF-1. Infection of the mice with
Listeria only exacerbated this effect of Pb. The same authors (Kowolenko et al., 1989) had
previously demonstrated that when BMDM were cultured in vitro with Pb chloride (0.1 |iM),
colony formation was significantly impaired. These combined results suggest that exposure to
Pb can impair the generation of macrophage populations as well as modulate the functional
spectrum of fully matured macrophages. Bunn et al. (200la) reported that gestational exposure
of CD rats to 50 ppm Pb acetate via the drinking water of the dams resulted in female adult
offspring with a significantly decreased percentage (58% reduced) of circulating monocytes.
A 100-ppm dose of Pb acetate produced a significant reduction (74% reduced) in the absolute
numbers of monocytes as well. The blood Pb level at birth associated with the decreased
percentage of macrophages in the adult offspring was 8.2 |ig/dL. In general agreement, Lee
et al. (2002) reported a significant decrease in the absolute numbers of circulating monocytes
and polymorphonuclear leukocytes (PMNs) in juvenile female chickens exposed ii
1 in ovo on
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embryonic day (E) 12 to 200 jig Pb acetate. The corresponding blood Pb level at hatching was
11.0 |ig/dL. However, in this case, the Pb-induced reduction in monocytes and PMNs was only
seen in concert with an airway viral infection (viral stressor) and not in the resting uninfected
animal.

Antigen Presentation and Lymphoid Stimulation
Exposure to Pb influences the interaction between macrophages and T lymphocytes, and
as a result, the capacity of macrophages to support T lymphocyte proliferation and activation can
be altered as well. Kowolenko et al. (1988) found that mouse macrophages exposed to Pb
(both in vivo and in vitro) can induce an increased proliferative response of T lymphocytes in
co-culture but that antigen-specific stimulation of primed T cells is significantly reduced.
Lead-suppressed antigen presentation capabilities of mouse macrophages were also reported by
both Smith and Lawrence (1988) and Blakley and Archer (1981).

Chemotaxis
Chemotactic activity of macrophages is an important function required for the directed
migration of macrophages to sites of infection and tumor growth. However, it is a functional
capacity that has not been systematically examined within the lead-immune literature. Using
female Moen-Chase guinea pigs, Kiremidjian-Schumacher et al. (1981) showed that Pb chloride
exposure of peritoneal macrophages in vitro (10-6 jiM) inhibited the electrophoretic mobility of
the cells.

Phagocytosis and Clearance of Particles
Phagocytosis of targets and removal/clearance of dead cells and particles are major
functions of macrophages. However, phagocytosis can involve a variety of different cell surface
receptors on macrophages, depending upon both the nature of the target encountered and the
subpopulation of macrophages examined. In general, phagocytic capacity of macrophages seems
to be relatively insensitive to Pb-induced immunomodulation compared with the effects on NO
and TNF-a production.
However, differences in outcome in phagocytosis evaluations are likely to be based on the
differences in the source of macrophages used and their relative activation state at the time of
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assessment. A few studies have described significant effects on phagocytosis, but these have
usually relied upon phagocytosis mediated through the Fc receptor on macrophages. Because
cell adherence to surfaces may be influenced negatively by Pb (Sengupta and Bishali, 2002),
impairment of phagocytosis may also involve some lack in efficiency with macrophage
anchoring to substrates. De Guise et al. (2000) reported no effect on bovine macrophage
phagocytosis of latex beads by Pb at in vitro treatment concentration of 10~4 M. This was in
contrast with suppressive effects of both cadmium and mercury. Using Sephadex-elicited
peritoneal macrophages derived from young turkeys fed 100 ppm Pb in the diet, Knowles and
Donaldson (1997) found a 50% reduction in the percentage of phagocytic macrophages using
SRBC targets. The activity per phagocytic macrophage was also reduced.
Kowolenko et al. (1988) studied the effect of Pb acetate at 10 mM in the drinking water of
CBA/J mice. They reported no effect on phagocytosis ofListeria monocytogenes targets, yet
they found an overall decreased resistance to Listeria. When the same investigators exposed
peritoneal and splenic macrophages to Pb in vitro (100 |iM), they also found no significant effect
of Pb on phagocytic activity. Zhou et al. (1985) reported that New Zealand white rabbit-derived
alveolar macrophages exposed to Pb in vitro at 10"5 M concentration were significantly impaired
in the phagocytosis of opsonized chicken erythrocytes (Fc receptor-mediated phagocytosis).
Trejo et al. (1972) reported that a single i.v. injection of Pb (5 mg/rat) into male Sprague Dawley
(SD) strain rats produced an inhibition in the phagocytic capacity of Kupffer cells.
Several studies have reported a decreased clearance capacity of the reticuloendothelial
system following in vivo exposure to Pb. Filkins and Buchanan (1973) found that injection of
5 mg of Pb acetate i.v. into male Holtzman strain rats produced reduced carbon clearance.
Similarly, Trejo et al. (1972) reported that a single i.v. injection of Pb (2.5 mg) into male SD
strain rats significantly reduced clearance of colloidal carbon.
In contrast, Schlick and Friedberg (1981) found that 20 |ig/kg Pb acetate in a single i.p.
injection of NMRI strain mice significantly increased the clearance of India ink. Ironically, oral
administration of Pb for 10, but not 30, days of 10 |ig/kg resulted in an increase in clearance
activity. Difference in route of Pb administration may be a factor in the different results
obtained.
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Induction of Heat Shock Proteins
One study (Miller and Qureshi, 1992), using a macrophage cell line, reported that
exposure of macrophages (MQ-NCSU) in culture to Pb acetate (1000 jiM) induced the same set
of four heat shock proteins as when the macrophages were subjected to thermal stress. This
result fits the hypothesis that Pb produces a profound immunomodulatory effect in macrophages
that has similarities with the exposure of macrophages to certain pathogens.

Apoptosis
Significant differences exist in the literature concerning the potential role of Pb in the
apoptosis of macrophages. The difference may be based on the exposure methodologies (in vivo
versus in vitro) as well as the source of macrophages utilized. De la Fuente et al. (2002) found
that human monocytes exposed to Pb in vitro at high concentrations did not undergo apoptosis.
This was in direct contrast with the apoptosis-promoting effects of cadmium in the same
assessment protocol. In contrast, Shabani and Rabibani (2000) exposed rat alveolar macrophage
to Pb nitrate in vitro and found that 60 jiM concentration produced a significant increase (2x) in
DNA fragmentation after 3 to 24 h in culture.

5.9.7 Granulocytes and Natural Killer (NK) Cells
Other cell types important in innate immunity, as well as in immunoregulation, are the
lymphoid population of natural killer cells and granulocytes, including PMNs (i.e., neutrophils).
Neither population appears to be a major target for Pb-induced immunotoxicity, although both
may be influenced indirectly via immune cell-cell interactions as well as by changes in cytokine
production. Among the two, neutrophils may be the more sensitive cell type based on assays
conducted to date. For neutrophils, several groups have reported alteration in chemotactic
activity following exposure to Pb. Queiroz et al. (1993) found impaired migration ability of
neutrophils from battery workers occupationally exposed to Pb. Likewise, Valentino et al.
(1991) had a similar observation among male occupationally exposed workers. Lead exposure of
young SD strain rats can increase the population of neutrophils (Villagra et al., 1997), although,
as the authors indicated, this does not necessarily afford enhanced host protection against
disease. Baginski and Grube (1991) reported that human neutrophils exposed to Pb had
increased killing capacity, probably via increased release of ROIs despite having reduced
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phagocytic capacity. This would fit the same general profile as the Pb effects on macrophages.
Therefore, neutrophils may contribute to Pb-induced tissue inflammation and damage via
increased ROI release. Yet, their effectiveness in protection against disease challenge may be no
greater following exposure to Pb, because some impairment in chemotaxis and phagocytosis has
been reported as well.
Yucesoy et al. (1997) reported that either Pb exposure or simultaneous exposure to Pb and
cadmium in human workers did not impair NK cytotoxicity activity. This finding was supported
by studies using in vivo exposure to Pb in rats (Kimber et al., 1986) and mice (Neilan et al.,
1983). Therefore, it would appear that NK cells are not a prime target associated with Pb-
induced immunotoxicity, although more subtle effects may certainly exist within the cell type.
Eosinophils represent an important granulocytic cell type in type 2 associated
inflammatory and allergic reactions. However, few studies have examined Pb exposure and
eosinophil activity. Villagra et al. (1997) reported that exposure of female juvenile SD rats to Pb
[four alternate-day s.c. injections of 172 mg/g body wt Pb acetate] increased the degranulation of
eosinophils (in animals given estrogen 1 day later). Such a response would be expected to
contribute to increased inflammation.

5.9.8 Hypersensitivity and Autoimmunity
At the time of preparation of the 1986 AQCD, little was known about the potential for Pb
to influence the risk of allergic and autoimmune diseases. However, since the early 1990s, a
significant number of studies have all pointed toward the fact that Pb causes a profound
dysregulation of the immune system. It skews the balance of responses in directions that reduce
certain host defenses against infectious diseases while enhancing the risk of allergic and
autoimmune disease. Lead exposure at low to moderate levels appears to alter T lymphocyte
responses in such a way as to increase the risk of atopy, asthma, and some forms of
autoimmunity. Increased IgE production following exposure to Pb is among the most frequently
reported immune alterations. Elevated IgE levels would be an associated risk factor for atopy
and allergic disease. Several investigators have discussed the fact that Pb is a likely risk factor
associated with the increased incidence of childhood allergic asthma (Miller et al., 1998; Heo
et al., 1998; Snyder et al., 2000; McCabe et al., 2001; Dietert et al., 2004; Trasande and
Thurston, 2005) as well as later life allergic disease (Heo et al., 2004; Carey et al., 2006).
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Joseph et al. (2005) observed no association for childhood blood Pb concentration and risk of
asthma among an African-American population. However, results on other populations from this
study, including those involving Caucasian children with blood Pb levels above 5 |ig/dL, led the
authors to call for further studies into the possible linkage of early life Pb exposure and risk of
asthma (Joseph et al., 2005).
As described by McCabe et al. (1991) and discussed by Dietert et al. (2004), Pb-induced
immunotoxicity is novel in that profound cellular toxicity is not evident following exposure at
low to moderate exposure concentrations. In fact, antibody responses overall are usually
unaffected or may be increased depending upon the class/isotype measured. However, the
functional responses mounted following Pb exposure do not reflect the normal immune balance
that would otherwise occur. This dysregulation can alter the risk of certain autoimmune diseases
based on several observations. Holladay (1999) has considered the importance of the timing of
exposure and the fact that early life exposure may establish the immune profile that then
contributes to later disease including autoimmunity.
Hudson et al. (2003) reported that Pb exposure can exacerbate systemic lupus
erythmatosus (SLE) in lupus-prone strains of mice. In contrast with the effect of mercury, these
authors found that for lupus, Pb exposure would not induce this autoimmune condition in
genetically resistant mice but would increase severity of the disease in genetically prone animals.
The authors noted some gender effects within certain strains (e.g., NZM88). Using early in ovo
exposure to Pb (10 jig/egg), Bunn et al. (2000) found that Pb acetate-exposed male chicks could
be induced to produce autoantibodies against thyroglobulin, which were not present in acetate-
exposed controls. No Pb-induced alteration was observed in females that were predisposed to
mount anti-thyroglobulin responses. The gender effect is intriguing in that autoimmune
thyroiditis in genetically predisposed strains is always more severe in females than in males.
Two lines of evidence suggest that the capacity of Pb to influence the risk of
autoimmunity is not always associated with simply a strict shift from Thl to Th2 responses.
Hudson et al. (2003) discussed the fact that lupus is not purely a Th2-mediated disease, but rather
seems to occur under conditions associated with skewing in either direction. McCabe et al.
(2001) found that Pb can increase the stimulation of alloantigen reactive T cells (where
macrophage processing of antigen is required) but not enhancement of T cell clonotypic
responses against either mitogens or superantigens (where processing is not required).
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This suggests that the role of Pb in influencing risk of autoimmune disease goes beyond a simple
consideration of Thl/Th2 balance. In fact, Goebel et al. (2000), studying mucosal immunity,
reported that administration of Pb-chloride to NOD strain mice produced a gut cytokine
microenvironment that was skewed toward Th2 over the short run, but later was shifted toward
Thl with increased production of IFN-y. This shift to Thl was accompanied by a loss of
tolerance and capacity to mount an immune response against a diet-associated protein (chicken
ovalbumin). The authors proposed that reduction of the capacity for oral tolerance would
predispose an individual toward autoimmune disease. The findings of Carey et al. (2006) had
similar implications. These authors reported that exposure of mice to Pb chloride increased
activation of neo-antigen-specific T cells, thereby increasing the risk of autoimmunity.
Finally, Waterman et al. (1994) and El-Fawal et al. (1999) have described the production
of autoantibodies against neural proteins in both battery workers and rats exposed to low levels
of Pb via drinking water. These authors have suggested that exposure to Pb may precipitate the
autoimmunity by altering antigen immunogenicity and/or the capacity of the immune system to
respond to certain antigens. This, in turn, may contribute to the eventual Pb-associated
neurological disease.

5.9.9 Mechanism of Lead-Based Immunomodulation
In the 1986 AQCD, there was little direct information available about the immune system
regarding the molecular mechanism(s) of Pb-induced immunotoxicity. Binding to thiol groups
and altering cell surface receptors were indicated as possible factors in altered immune function.
Since that time, some additional information has been generated through a variety of studies on
human and animal immune cells. However, a clear or simple explanation remains to be
determined. Table 5-8 lists studies on the immune system that have contributed to a better
understanding of potential mechanisms or have forwarded potential hypotheses with some
supporting data.
At the level of cell-cell interactions, it seems clear that Pb alters metabolism and cytokine
production by macrophages and antigen presenting cells. It also reduces their capacity to
respond to growth factors such as CSF-1 (Kowolenko et al., 1989). Pace et al. (2005) discussed
the hypothesis that reduced populations of functionally altered macrophages (because of
Pb-induced unresponsiveness to CSF-land over production of ROIs) in tissues can produce
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Table 5-8. Suggested Mechanisms of Lead-Induced Immunotoxicity
Species
Mouse
Mouse
Chicken

Mouse
Mouse
Rat

Mouse

Human

Mouse

Mouse
Strain/Gender
Balb/c
TO strain males
Cornell K Strain

Balb/c females
C57 Bl/6
females
C 57B1/6
PC-12 cells

DO11. 10 ova-
mice

DOll.lOova-tg
mice

CBA/J females

Swiss Females
Suggested
Endpoints
CSF-1
Responsiveness of
Macrophages
Vitamin E
protection against
lead-induced
splenomegaly
Thymulin partial
reversal of Th
skewing
Lead disruption of
antigen
processing and
presentation
signals
PKC activation
NF-KB activation
AP-1 induction
C-Jun kinase
induction
Adenylcyclase
activation with
elevated cAMP
levels
V(3 gene usage

NF-KB activation
in CD4+ cells

t Immunogenicity
of neural proteins

tTNF-a
production
Associated
Functional
Alteration
ITesticular
macrophages
^Fertility
tPutative ROI
associated
splenomegaly
ILead-induced
DTH
suppression
tAlloreactive
CD4+ hlgh cells
tRisk of
autoimmunity
tTNF-a, tIL-6
tPGE2
tROI

tTh skewing

tRisk of
autoimmunity

tRisk of
autoimmunity
and
hypersensitivity
t Autoimmune
mediated
neurological
damage

t Sensitivity to
endotoxin
Lowest
Effective
Dose
0.1 ppm
Img/kg
400 jig

0.5 |jM in
vitro
20|jM in
vitro
1 |jMin
vitro

2.5 |jM in
vitro

50 ng
2x/week
s.c.
IpM

Lead-
altered
proteins
used as
antigens
5mg
Duration
6 weeks
2 weeks
Single in
ovo
injection
4 days
4.5 hrs
5-120 min

15 mins-
6 hrs

8 weeks

30 min

3 injections
of lead-
modified
neural
proteins
Single i.p.
injection
References
Pace et al.
(2005)
Fernandez-
Cabezudo
et al. (2003)
Lee and
Dietert,
(2003)
McCabe
etal. (2001)
Flohe et al.
(2002)
Ramesh
etal. (1999)

Heo et al.
(1998)

Heo et al.
(1997)

Pyatt et al.
(1996)

Waterman,
etal. (1994)

Dentener
etal. (1989)
5-239
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nonimmune problems. The model they used is the homeostatic presence of testicular
macrophages and the likelihood that Pb-induced macrophage immunotoxicity contributes
directly to Pb-associated reduction in male fertility.
Also, Pb is known to selectively alter cell signaling to CD4+ T cell subpopulations,
promoting proliferation in some but not others. The outcome is enhanced tissue inflammation,
reduced CMI, and increased production of atopy-inducing antibodies. Risk of autoimmune
reactions is increased in some models of Pb-induced immunotoxicity. For example, Heo et al.
(1997) reported that Pb-exposed murine T lymphocytes are biased in expression of VP genes.
This is potentially problematic as this phenotype is common among a variety of human and
animal model autoimmune conditions. A variety of exogenous factors has been reported to
partially ameliorate Pb immunotoxic effects. Lead chelation in Pb-exposed dams corrected some
Pb-induced immunotoxic problems in the rat female offspring, but it left the animals with some
DMSA-induced immune alterations (Chen et al., 1999). Other exogenously administered factors
that have been reported to partially restore Pb-suppressed immune function are vitamin E
(Fernandez-Carbezudo et al., 2003) and thymulin (Lee and Dietert, 2003).
At the subcellular level, the bases for immunotoxic changes remain speculative. McCabe
et al. (2001) suggested that altered antigen processing and subsequent cell signaling to T cells
may be an explanation for the capacity of Pb to selectively increase CD4+ (high density) cells.
Certainly, Pb appears to alter signal transduction. It appears to elevate expression of the nuclear
transcription factor NF-KB (Pyatt et al., 1996; Ramesh et al., 1999) as well as to increase
expression of AP-1 and cJun (Ramesh et al., 1999). Flohe et al. (2002) found evidence that Pb
can elevate the activation of PKC. The authors speculated that this might be involved in
Pb-induced increases in TNF-a production. Additionally, Heo et al. (1998) reported that Pb
increases adenyl cyclase activity among T lymphocytes, generating elevated cAMP levels.
The authors hypothesized that this effect, in conjunction with differences in cell signaling
pathways for promoting Thl versus Th2 cells, may be involved in the capacity of Pb to skew
ThO helper cells toward Th2.

5.9.10 Age-Based Differences in Sensitivity
With the literature available at the time of the 1986 AQCD, it was virtually impossible to
evaluate age-based differences in susceptibility to Pb-induced immunotoxicity. However, in
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recent years, this has become a major topic of study for many toxicants, including Pb (Dietert
and Piependrink, 2006). Several studies have added to the available data assessing the
developmental immunotoxicity of Pb (reviewed in Barnett [1996], Dietert et al. [2000, 2004],
Lee and Dietert [2003]). Several patterns have emerged from exposure data using animals of
different ages.
First, it seems clear that blood Pb levels at or near birth of below 10 |ig/dL can be
associated with juvenile and/or adult immunotoxicity. Several studies reported effects at blood
Pb levels in the range of 5 to 8 |ig/dL. These low levels would seem to place the immune system
on par with the nervous system in terms of potential sensitivity to Pb. Table 5-9 shows examples
of studies in which low blood Pb levels were linked with immunotoxicity.
Table 5-9. Immunomodulation Associated with Low Blood Lead Levels
in Animals
Species
Mouse

Rat

Chicken

Chicken
Blood lead
Oig/dL)
-5.0

8.2

6.75

8.0

8.2

11.0

7.0
Age at
Measurement
1 week

1 day

4 weeks

1 day

1 day
Immune Parameter(s)
tlgE,
4 Splenic
T Cell Populations
Imonocytes

4DTH,
4IFN-Y,
tIL-4
tTNF-a
tRel. Spleen weight
^circulating lymphocytes post
infection
4DTH and 4TLC, monocytes,
PMNs post infection
tautoantibody production
Age at
Assessment
2 weeks

13 weeks

5 weeks

10 weeks
Reference
Snyder et al.
(2000)

Bunn et al.
(200 la)
Chen et al. (2004)

Lee et al. (2002)

Lee etal. (2001)

Bunn et al. (2000)
A second finding is that the immunotoxic effects induced by Pb are persistent long after
blood levels and potential body burdens of Pb are significantly reduced. Miller et al. (1998),
Chen et al. (1999), Snyder et al. (2000), and Lee et al. (2001) all emphasize this latter point.
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In fact, in most of these studies immunotoxic alterations were present when Pb levels in
exposed animals were not distinguishable from control levels. This should provide a cautionary
note regarding studies in humans. Data from adult exposures provide little insight into the
potential persistence following adult exposure to Pb. However, rather than the developing
immune system being more regenerative postexposure and able to withstand immunotoxic insult,
it appears that the non-dispersed developing immune system is a particularly susceptible target to
many immunotoxicants (Dietert et al., 2002).
A third, and somewhat surprising, finding concerning early exposure to Pb is that
qualitative differences in the spectrum of immune alterations can exist, depending upon the
developmental window of exposure. Figure 5-17 illustrates this point. Early embryonic
exposure of rats and chickens to Pb failed to alter juvenile DTH responses, despite significant
effects on macrophage function. However, exposure to Pb after the mid-embryonic point of
embryonic development readily suppressed subsequent DTH. As shown in Figure 5-17, the
development window in which sensitivity to DTH suppression emerges is quite similar in the
two species. This observation suggests that both quantitative (LOAELs) and qualitative (range
of immune alterations) differences in sensitivity to Pb can exist across different age groups.
Additionally, some studies in animals have noted gender differences in the effects of Pb
following exposure (Bunn et al., 2000, 2001a,b, c; Hudson et al., 2003). Gender differences
have also extended to results in humans as per Pb-induced immune and inflammatory alterations
(Karmaus et al., 2005; Fortoul et al., 2005). It seems feasible that, even in the embryo, hormonal
differences among females and males may impact some outcomes of low-level Pb exposure.
Table 5-10 shows comparisons of the lowest reported blood Pb levels at different ages
associated with the same immunotoxic endpoint. From these limited comparisons, it would
appear that different ages of rodents (e.g., embryonic versus adult) differ in dose sensitivity for
Pb-induced immunotoxicity somewhere in the range of 3 to 12-fold. Clearly, additional direct
comparisons would help to refine this estimate.
A fourth observation from the early exposure studies is that exposure to even very low
levels of Pb can predispose the immune system for unanticipated postnatal responses when the
system is stressed. This general phenomenon is called latency. Lee et al. (2002) provided an
example of this following the single in ovo exposure of embryonic day 5 chick embryos to low
levels of Pb (10 jig; blood lead level 1 day post hatch of 8.2 jig/dL). The leukocyte profiles of
5-242
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Table 5-10. Comparisons of Age-Based Sensitivity to Lead-Induced Immunotoxicity
Species
Mouse

Rat
Mouse

Rat
Altered Endpoint Embryo - fetal*
tlgE ~5|^g/dL

iDTH 34 |^g/dL
(persistent effect assessed
13 weeks post-exposure)
4DTH —

tTNF - a 8 ng/dL
(persistent effect assessed
13 weeks post-exposure)
Neonatal* Adult*
12 |^g/dL 38 |^g/dL

— >112|^g/dL
(measured at birth
for persistent
effect)
29 |^g/dL 87 |^g/dL

— >112|^g/dL
(measured at birth
for persistent
effect)
References
Snyder et al.
(2000)
Heoetal. (1996)
Miller et al.
(1998)
Bunn et al.
(200 Ib)
Faith etal. (1979)
McCabe et al.
(1999)
Miller et al.
(1998)
Chen et al. (2004)
*Lowest blood lead concentration reported with effect.

the animals appeared to be completely normal. However, when these animals were exposed to a
respiratory virus, their pattern of leukocyte mobilization was completely aberrant from controls.
Therefore, some immunotoxic alterations following early exposure to low levels of Pb may only
be evident during periods of postnatal stress.
Several studies have reported the positive association of blood Pb levels in children
with elevated serum IgE (Karmaus et al., 2005; Sun et al., 2003; Lutz et al., 1999). These
observations are supported by the animal data in rats and mice (Miller et al., 1998; Snyder et al.,
2000) and suggest that Pb-induced risk of atopy and asthma may be a particular health issue.
Trasande et al. (2005) recently discussed the fact that, despite progress in reducing the
deposition of Pb in the environment, Pb continues to be a concern relative to asthma and
children's health.

5.9.11 Summary
The immune system appears to be one of the more sensitive systems to the toxic effects
of Pb. The 1986 AQCD provided an excellent summary of the studies that had been conducted
prior to that date. But knowledge of fundamental immunology has progressed greatly during the
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past 20 years. Not surprisingly, the large number of studies conducted since the mid-1980s have

provided a much clearer understanding of the immune-associated problems that can arise from

problematic exposure to Pb. Studies across humans and a variety of animal models are in

general agreement concerning both the nature of the immunotoxicity induced by Pb as well as

the exposure conditions that are required to produce immunomodulation. Figure 5-18

summarizes the basic immunotoxic changes induced by Pb that result in Th skewing, impaired

macrophage function, and increased risk of inflammation-associated tissue damage.
Key Effects of Lead on the Immune System
Pb
Inflammatory Tissue

Macrophages. Macrophages and Antigen
Presenting Cells
T ceils
B cells
Membrane
Arachidonic
Acid
Ac b¥ ah on
Host Proteins
Effects on Antigen
Processing
Presentation
Reduced Ceil-
mediated
Immunity
I
Modified Neural Antigens
Auto-antibodies and
Possible Tissue
Damage
Increased Tissue Inflammation

Reduced Ceil-mediated Immunity
IgE I & IncreasecS Risk of
Atopy and Aston a
Figure 5-18. This figure shows the fundamental alterations to the immune system and to
immunological response and recognition induced by exposure to lead. The
functional shifts are disproportionate compared to the relatively modest
changes among leukocytes with low to moderate exposure to lead.
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• Lead is unlike many immunotoxicants in that, at low to moderate levels of exposure, it does
not produce overt cellular cytotoxicity or lymphoid organ pathology. However, it can induce
profound functional alterations that influence risk of disease. Lead preferentially targets
macrophages and T lymphocytes, although effects have been reported in B cells and
neutrophils as well.

• There are three major hallmarks of Pb-induced immunotoxicity. First, Pb can dramatically
suppress the Thl-dependent DTK response, as well as production of associated
Thl cytokines. Second, Pb can dramatically elevate production of IgE while increasing
production of Th2 cytokines, such as IL-4. Third, and perhaps most sensitive, is the
modulation of macrophages by Pb into a hyperinflammatory phenotype. After exposure to
Pb, macrophages significantly increase production of the proinflammatory cytokines TNF-a
and IL-6 (and in some studies IL-1). Many studies also reported elevated release of ROIs
and prostaglandins. Ironically, production of one of the most important host defense factors,
NO, is consistently and severely suppressed by Pb exposure. This package of Pb-induced
changes among macrophages makes them more prone to promote tissue destruction but
actually less capable of killing bacteria or possibly presenting antigens to T lymphocytes.
The Pb-induced shift in phenotype explains the capacity of inhaled Pb to promote bronchial
inflammation while bacterial resistance is severely depressed.

• Lead-induced skewing of Th activity (biasing responses toward Th2) across a population
argues for expectation of a greater risk of atopy, asthma, and some forms of autoimmunity.
Concomitantly, resistance to some infectious diseases could be reduced. This predicted
change of risk might help explain some recent trends in the incidence of diseases, such as the
epidemic rise in allergy and some forms of asthma in the United States at current blood Pb
levels well below 10 |ig/dL.

• Sensitivity of the immune system to Pb appears to differ across life stages. Studies in rats
and mice suggest that the gestation period is the most sensitive life stage, followed by the
early neonatal stage. But even during embryonic, fetal, and early neonatal development,
critical windows of vulnerability are likely to exist. Compared to adults, the increased dose
sensitivity of the embryo-fetus would appear to fall in the range of 3-10x depending upon the
immune endpoint considered. Some studies have found evidence for gender differences in
the impact of Pb on the immune system, particularly with early life exposures. Potential
gender differences in immunotoxic outcome may be important in the evaluation of those
populations at greatest risk.

• Recent studies have suggested that exposure of embryos to Pb producing neonatal blood Pb
levels below 10 |ig/dL can also produce later-life immunotoxicity (see Table 5-9).
Furthermore, immunotoxicity persists long after any evidence of prior embryonic Pb
exposure. This latter observation from several laboratories may have implications for the
design of human epidemiological studies.
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5.10 EFFECTS OF LEAD ON OTHER ORGAN SYSTEMS
In the 1986 Lead AQCD, discussion of other organ systems included the cardiovascular,
hepatic, gastrointestinal (GI), and endocrine systems. Due to our increased understanding on the
effects of Pb on cardiovascular and renal systems and their contribution to potential health
effects of Pb, separate sections (5.5, 5.7) were dedicated earlier in this chapter to detailed
discussions on these aspects. Similarly, with our increased understanding on the effects of Pb on
endocrine functions and its inherent role with respect to neurotoxicological, reproductive, and
developmental effects, literature reviewed for Pb endocrine effects was discussed in the
respective sections. This section focuses on the discussion of Pb effects on the hepatic and
GI systems.

5.10.1 Effects of Lead on the Hepatic System
The liver is a highly active metabolic tissue. Apart from its roles in fatty acid metabolism
and limited heme synthesis function, the liver also has a major role in guarding other systems
from the toxic effects of xenobiotic compounds using a huge complement of detoxification
machinery referred to as phase I and phase II enzyme systems. Limited studies on experimental
animals reported in the 1986 AQCD indicated that Pb induced effects in the hepatic system.
Laboratory animals, especially rats, exposed to Pb nitrate have exhibited increased liver cell
proliferation, DNA synthesis, cholesterol synthesis, and glucose -6-phosphate dehydrogenase
(G6PD) activity indicative of Pb-induced hyperplasia. Further, the literature reviewed in the
1986 AQCD reported alterations in the levels of drug metabolizing enzymes in experimental
animals given large doses of Pb. The evidence for such effects in humans was less consistent.
The 1986 document also concluded that the effects on the liver occurred only at high exposure
levels. The majority of studies on the effects of Pb on the hepatic system in experimental
animals that are reviewed in this document report functional and biochemical changes in the
liver, clearly pointing to metabolic perturbations in liver. For ease in understanding and
integration of these functional changes, the discussion is divided into the following four
subsections: hepatic drug metabolism, lipid and glycogen metabolism and lipid peroxidation,
and heme synthesis.
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5.10.1.1 Hepatic Drug Metabolism
Approximately 75% of the hepatic blood comes directly from the gastrointestinal viscera,
with the majority of drugs or xenobiotics absorbed coming directly to the liver in concentrated
form. The liver is equipped with a huge complement of drug metabolizing enzymes that detoxify
many of the xenobiotics but also activate the toxicity of others. Oxidation and conjugation of
xenobiotics have historically been referred to as phase I and phase II reactions. The phase I
enzymes include cytochrome P450 (CYP450) heme-containing monoxygenases, flavin-
containing monoxygenases, and epoxide hydrolases. The phase II enzymes include glutathione
(GSH) S-transferases (GST), UDP-glucuronyl transferases (UGT), N-acetyltransferases (NAT),
and sulfotransferases (SULT). Xenobiotic metabolism by these two complements of enzyme
systems are essential for catabolizing and eliminating drugs; however, this process can also
produce activated toxicants and carcinogens. A limited number of these CYP450s are involved
in the biosynthetic pathways of steroid and bile acid production. It has been increasingly
recognized that, under certain circumstances, CYP P450s can produce ROS that result in
oxidative stress and cell death.
Liver is an active tissue. In addition to xenobiotic metabolism, it also participates in
gluconeogenesis, fatty acid metabolism, and cholesterol biosynthesis. Research concerning the
effects of Pb on the hepatic system in the past 15 years has provided some preliminary
indications of Pb-induced alterations in many of the hepatic functions described above.
The following discussion presents, as much as possible, the effects of Pb on individual enzymes,
but due to the multifarious interactions of many of these metabolic enzymes, there may be places
such separation was not possible.

Phase I Enzyme
Earlier studies on the toxic effects of Pb on hepatic drug metabolizing enzymes
demonstrated that acute exposure to Pb acetate decreased rat hepatic CYP450s with increased
levels of urinary 5-aminolevulinic acid (ALA). Co-treatment with phenobarbitol, a CYP450
inducer, was shown to reverse the decrease CYP450 levels, suggesting a Pb acetate-mediated
inhibition of heme synthetic enzymes. Decreased activities of estradiol-17 beta enzyme
observed in rat liver treated with triethyl Pb chloride (Odenbro and Arhenius, 1984) suggest that
both Pb and organo-Pb compounds are capable of inhibiting CYP450 activities. Roomi et al.
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(1986) also observed decreased levels of hepatic microsomal CYP450s and decreased
aminopyrene-N-demethylase activity on exposure to a single dose of Pb nitrate (5-10 mmol/kg
body wt). This decrease in phase I enzymes was followed by increased levels of phase II
components such as GSH, GST, and DT diaphorase, suggesting that Pb nitrate and other Pb
compounds can induce biochemical properties characteristic of hepatocyte nodules. Subchronic
(2-3 months) exposure to Pb acetate (5-50 mg/kg body wt) had been found to induce CYP450s
and cytochrome b5 in rat liver and kidney (Nehru and Kaushal, 1992). As described earlier,
multiple isoforms of CYP450s exist in the liver.
To identify the inhibitory effect of acute Pb exposure on specific isoform(s), Degawa
et al. (1994) exposed male F344 rats to Pb nitrate (20,100 jimol/kg body wt) and evaluated liver
CYP450s 24 h postexposure. Lead nitrate exposure preferentially inhibited cytochrome
P4501A2 enzyme activity in liver microsomal preparations as assayed for mutagenic conversion
of substrates 2-amino-6-methyl-dipyridol [1,2-a; 3',2-d] imidazole and 3-amino-l-methyl-5H-
pyridol [4,3,-b] indole. Lead nitrate exposure also inhibited the induction of cytochrome
P4501A2 by the inducers 3-methylcholanthrene and 2-methoxy-4-aminoazobenzene at both the
protein and mRNA levels. The authors further concluded that the specific inhibition of P4501A2
by Pb nitrate observed may have been due to inhibition of heme synthesis, as Pb nitrate was not
found to inhibit P4501A2 activity in vitro. Additional studies carried out by the same group
using various metal ions (e.g., Pb, Ni, Co, and Cd) found that the specific inhibition of P4501A2
was unique to Pb nitrate (Degawa et al., 1994, 1995). Degawa et al. (1996) also investigated the
effect of Pb nitrate-mediated inhibition of CYP1A gene activity in rat liver by specific inducers
and reported that Pb nitrate inhibited the induction of CYP1A mRNA by aromatic amines, but
not by aryl hydrocarbons, suggesting the role of other cellular factors in the transcriptional
activation of CYP1A genes. Lead nitrate has been reported to induce the production of TNF-a in
rat liver (Shinozuka et al., 1994), a cytokine implicated in the suppression of constitutive
expression of CYP1A2 mRNA in rat hepatocytes. Based on these findings, Degawa et al. (1996)
concluded that the inhibition of constitutive and aromatic amine-induced expression of CYP1A2
in rat liver caused by Pb nitrate may occur at least in part by TNF-a-associated mechanisms.
Lead nitrate (0.33 mg/kg body wt) pretreatment-mediated protection conferred against carbon
tetrachloride (0.3 mL/kg)-induced hepatotoxicity as reported by Calabrese et al. (1995) may be
due to the inhibition of CYP450 activities in liver by Pb.
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Jover et al. (1996) investigated the effect of heme deficiency on Pb-induced hepatic P450
function and transcription. These authors concluded that the decrease in hepatic P450 resulting
from Pb intoxication was mediated by two different mechanisms. One mechanism is involved
inhibitory effects on P450 by Pb at the transcriptional level; the second was heme- dependent, as
Pb-mediated inhibition of heme synthesis decreased the heme saturation of P450 and the
apo-P450 ratio.
The effect of heavy metals (Cd, Co, Cu, Ni, Pb, and Zn) on 3-methylcholanthrene-
induction of cytochrome P4501A and the activity of ethoxyresorufin-O-deethylase (EROD)
were investigated in fish hepatoma cells (PLHC-1) by Briischweiler et al. (1996). The authors
reported that all the heavy metals tested had more pronounced effects on EROD activity
compared to controls. The inhibitory potency of Pb was reported to be very low compared to
cadmium or cobalt. A single treatment of Pb acetate induced hepatic DT diaphorase activity
(Sugiura et al., 1993). This induction of hepatic DT diaphorase by Pb acetate has been reported
to be decreased with concomitant administration of Dil, a calcium antagonist. Based on these
observations, Arizono et al. (1996) suggested that DT diaphorase induction by Pb acetate may
occur de novo via protein synthesis mediated by increased cellular calcium. The potential
interaction of metals, including Pb, on the induction of CYP1A1 and CYP1A2 by polycyclic
aromatic hydrocarbons (PAHs) in human hepatocyte cultures was investigated by Vakharia et al.
(2001). Lead nitrate, like other metals such as Cd, Hg, and As, decreased the extent of CYP1A1
and CYP1A2 induction by five different PAHs. The authors concluded from these studies that
Pb (5 jiM) diminished the induction of CYP1A1 and CYP1A2 in human hepatocytes by
ultimately decreasing the levels of CYP1A1 protein that was normally attainable through PAH
induction. Korashy and El-Kadi (2004) also investigated similar interactions of metals with
aryl hydrocarbon receptor (AHR)-regulated gene expression and enzyme activities in wild-type
murine hepatoma cells (Hepa Iclc?) and AHR-deficient cells (C12). These studies indicated
that metals alone (including Pb) did not significantly alter CYP1 Al proteins or activity, or
change AHR ligand-induced enzyme activity. There was no change in mRNA levels. Lead, in
the presence or absence of AHR ligand, increased the activity of NAD(P)H:quinone
oxidoreductase and its mRNA levels.
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Phase II Enzymes
A single injection of Pb nitrate (5-10 jiM/100 g body wt) was found to increase GST
activity levels (Roomi et al., 1986). Additional studies by the same group identified induction of
a specific form GST-P by Pb nitrate in rat liver (Roomi et al., 1987). Because a single injection
of Pb nitrate decreased phase I and increased phase II hepatic enzymes, these investigators
concluded that Pb nitrate treatment initiated a biochemical phenotype similar to carcinogen-
induced hepatocyte nodules. Immunohistochemical analysis by the same group reported that Pb
nitrate administration resulted in the appearance of GST-P in most of the hepatocytes, an enzyme
that is otherwise undetectable in normal rat liver (Columbano et al., 1988; Roomi et al., 1987).
On the other hand, Nakagawa (1991) reported inhibition of GST on acute exposure to Pb and
that the inhibition of GST followed a reduction in liver GSH levels. Nakagawa (1991)
concluded that the depletion of GSH was not necessarily a critical factor in inhibiting GST.
Planas-Bohne and Elizdale (1992) found that acute exposure to Pb nitrate (100 jimol/kg)
caused a significant increase in liver and kidney GST activity. Gel electrophoresis analysis to
evaluate the contribution of various GST isoforms indicated that enhancement of liver GST
activity was predominantly due to induction of GST isoform 7-7 in liver compared to all
isoforms in kidney. Liver GST-P isoform was reported to be induced by both Pb acetate and Pb
nitrate (Boyce and Mantle, 1993; Koo et al., 1994). This transient induction of GST-P has been
regulated at transcription, post-transcription, and post-translational levels. Suzuki et al. (1996)
utilized a transgenic approach to investigate the transcript!onal regulation of GST-P induced by
Pb and identified glutathione S-transferase P enhancer I (GPEI), an enhancer (whose core
consists of two AP-1 site-like sequences) located at the 5' flanking region of this gene. The
authors demonstrated that GPEI is an essential element in the activation of the GST-P by Pb and
that the trans activating factor AP-1 is likely to be involved, at least in part, in the transcriptional
activation of the GST-P gene by Pb via the GPEI sequence.
Daggett et al. (1997, 1998) investigated the effect of inorganic and organic Pb on liver
GST expression and other phase II detoxifying enzymes in rat liver and kidney. Triethyl Pb
chloride (TEL) injection (10 mg/kg body wt) decreased liver GST activity, as well as levels of
various other GST isoforms (Daggett et al., 1997), in contrast to significant induction of kidney
GST activity, suggesting that a single compound, TEL, had opposite effects on the expression of
GST isozymes and indicated the complexity of GST regulation. Similarly, this group also
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reported that a single injection of Pb acetate (114 mg/kg body wt) reduced GSH levels, increased
production of malondialdehyde (MDA), and did not change the expression of various GST
isoforms analyzed, except GST-pl on repeated injection (Daggett et al., 1998). Similar to
studies with TEL, Pb acetate also increased the expression of GST enzyme activity and
expression of various isoforms without changing GSH and MDA levels, suggesting that
oxidative stress may not be mediating the toxicity in kidney. On the other hand, TEL exposure
was found to decrease microsomal estradiol metabolism (Odenbro and Rafter, 1988).
The suppression of GST expression reported by Daggett et al. (1997, 1998) is in contrast to the
induction of GST reported by various other groups discussed earlier. Other GSH-dependent
enzymes (i.e., GSH peroxidase, GSH reductase) have been found to be suppressed with a
simultaneous increase in oxidized GSH (GSSG) and a reduction in GSH/GSSG ratio (Sandhir
and Gill, 1995). More detailed information on these and related studies is summarized in
Table AX5-10.1.

5.10.1.2 Biochemical and Molecular Perturbations in Lead-Induced Liver Tissue Injury
Oskarsson and Hellstrom-Lindahl et al. (1989) studied the cellular transport of Pb (203Pb),
in rat hepatocytes using dithiocarbamate (DTC). Cells treated with Pb acetate and Pb-DTC
lipophylic complex demonstrated increased cytosolic Pb levels compared to Pb alone. This was
further evaluated by measuring levels of ALAD. Cells treated with Pb-DTC complex showed
rapid and stronger inhibition of ALAD compared to Pb acetate, suggesting that this inhibition
was due to increased mobilization of Pb into cells treated with Pb-DTC complex. Another report
by the same group, Hellstrom-Lindahl and Oskarsson (1990), suggested that the increased
inhibition of ALAD was due to the release of Pb from the Pb-DTC complex by decomposition.
Using the mouse strain with a duplication of the ALAD gene (DBA), Claudio et al. (1997)
reported increased accumulation of Pb in this strain by many fold as compared to mice with a
single copy of the ALAD gene (C57).
A single injection of Pb nitrate was reported to cause hepatic hyperplasia correlating with
hepatic de novo synthesis of cholesterol along with alterations in glucose and lipid metabolism
leading to altered serum lipid profiles (Dessi et al., 1984; Pani et al., 1984). Mobilization of
hepatic glycogen and altered gluconeogenic enzymes, including differential expression of G6PD,
has been reported following Pb exposure (Batetta et al., 1990; Hacker et al., 1990). Chronic Pb
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intoxication has also been reported to inhibit gluconeogenic enzymes, alterations that were
implicated in Pb bio-transformation rather than liver cell proliferation in Wistar rats (Calabrese
and Baldwin, 1992). Although these studies point out to newer directions regarding Pb effects
on hepatic carbohydrate metabolism, due to lack of information on blood Pb levels, they have
limited value for extrapolation to human exposure scenarios and associated health effect
assessment. Increased levels of serum lipid peroxide (LPO) were also observed in rats given SC
injection of Pb acetate, supporting similar increased levels of serum LPO in humans exposed to
Pb (Ito et al., 1985). These initial studies suggest that alterations in liver intermediary
metabolism occur on exposure to Pb with a role for Pb-induced LPO in hepatotoxicity and
potential involvement of oxidative stress in Pb toxicity. Limited studies on the hepatic lipid
provide peroxidation blood Pb levels in the range of 18 to 35 |ig/dL.
Dessi et al. (1990 ) investigated the role of fasting on Pb-induced hepatic hyperplasia by
monitoring the activities of enzymes involved in cholesterol synthesis and the hexose
monophosphate shunt and reported that stimulation of these enzymes, even in Pb acetate-treated
fasting rats, supported the role of new endogenous synthesis of cholesterol and gluconeogenic
mechanisms in Pb-induced hepatic cell proliferation. Chronic exposure to Pb was found to
increase the arachidonate/linoleic acid ratio in liver and serum (Donaldson and Leeming, 1984;
Donaldson et al., 1985) along with the GSG concentration (McGowan and Donaldson, 1987).
As GSH and arachidonate are precursors for peptido-leukotrienes, Donaldson's group
investigated the potential effects of dietary Pb on levels of fatty acids, peptido-leukotrienes, and
arachidonate/linoleic ratios in chicken fed with diets low in calcium and methionine. These
investigations found similar increases in arachidonate/linoelic acid ratio and in GSH levels
without bearing on peptido-leukotriene levels. The authors also found the influence of a low
calcium and methionine diet on Pb-induced serum fatty acid profiles (Knowles and Donaldson,
1990).
Chronic sublethal Pb exposure (5 ppm Pb nitrate for 30 days) has been found to alter liver
lipid profiles in blood and liver tissue of the fresh water fish Anabas testudineus (Tulasi et al.,
1992). These authors reported significant increases in liver total lipids, cholesterol, and free fatty
acids. Tandon et al. (1994) reported that iron deficiency enhanced the accumulation of Pb in
liver and kidney and also increased liver calcium levels. Induced expression of metallothionein
(MT) in renal and intestine was also observed in iron deficiency. Han et al. (1996) investigated
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the effect of Pb burden on weight loss using an energy restriction diet regimen on rats with prior
Pb exposure. The authors reported that rats on a substantial weight loss regimen (40% of normal
calories) exhibited a significant increase in the quantity and concentration of liver Pb and a
decrease in the concentration of other metals (e.g., Ca, Cu, Mg, Zn). The authors concluded that
weight loss can increase the liver concentration of Pb, even in the absence of continued
exposure. Combined exposure to Pb (70 mg/kg) and Cd (20 mg/kg) in Buffalo rats for 7 weeks
was found to alter liver levels of Zn and Cu, with less accumulation of Pb and Cd, compared to
individuals exposure to either Pb or Cd alone (Skoczyriska et al., 1993). These authors also
reported that a combined exposure regimen interfered with serum lipid profiles (Skoczyriska and
Smolik, 1994).
Liu et al. (1997) utilized rat primary hepatocyte cultures to explore the protective effect of
Zn-induced expression of metallothionein (MT) in Pb toxicity. These authors found that, in the
control cells without prior Zn exposure, most of the Pb was found bound to high-molecular
weight proteins in the cytosol, while in the Zn pretreated cells, a majority of Pb bound to MT,
indicating a MT-mediated protection against Pb toxicity to hepatocytes. More details about these
and related studies are summarized in Table AX5-10.2.

5.10.1.3 Effects of Lead Exposure on Hepatic Cholesterol Metabolism
Lead nitrate-induced hyperplasia or liver cell proliferation involves simultaneous increase
in both liver and serum total cholesterol levels. Recent studies have reported various molecular
events associated with this process. Induction of gene expression for CYP51 (Lanosterol 14a-
demethylase), an essential enzyme for cholesterol biosynthesis, was reported in Pb nitrate-
induced liver hyperplasia, although other cytochrome P450 enzymes involved in drug
metabolism have been reported as being suppressed, as discussed in earlier sections. This gene
has various regulatory elements and its constitutive expression in liver is mediated by sterol
regulatory element (SRE) and by the SRE binding proteins-la, 2, and Ic. Kojima et al. (2002)
reported that Pb nitrate induced the expression of CYP51 in the livers of both immature
(4-week-old) and mature (7-week-old) rats and that this induction appeared to be mediated by the
upregulation of SRE binding protein-2. However, this increased synthesis of cholesterol
observed in rat liver was not mediated by endogenous feedback regulation by sterols, as no
decrease in serum total cholesterol was observed. To understand the molecular mechanisms
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involved in the Pb nitrate-mediated development of hepatic hypercholesterolemia, Kojima et al.
(2004) investigated the expression of various enzymes involved in cholesterol homeostasis,
including some of the associated transcription factors in male rats exposed to Pb nitrate
(100 jimol/kg body wt). The authors reported that Pb nitrate exposure caused a significant
increase in liver and serum total cholesterol levels at 3 to 72 h and 12 to 72 h, respectively.
The enzymes involved in cholesterol biosynthesis viz. (i.e., 3-hydroxy-3methyglutaryl-CoA
reductase, farnesyl diphosphate synthase, squalene synthase, CYP51) were all activated (3-24 h),
while the enzymes involved in cholesterol catabolism such as 7a-hydroxylase were remarkably
suppressed 3 to 72 h. Figure 5-19 shows the involvement of Pb at various stages of the
cholesterol synthesis pathway. The induction of the cytokines interleukin-loc and TNF-a in rat
liver prior to the induction of the genes for these synthesis enzymes suggested that Pb nitrate-
induced cholesterol synthesis is independent of sterol homeostasis regulation. Following
gestational and lactational exposure to Pb acetate (0.05 mg/kg body wt), Pillai and Gupta (2005)
reported that the activities of the hepatic steroid metabolizing enzyme 17-p-hydroxy steroid
reductase, UDP glucouronyl transferase, and CYP450 levels decreased in rat pups on PND21.
Alterations in the hepatic system of neonates and pups (at PND12 and PND21) after
gestational and lactational exposure to Pb acetate (300 mg/L) have been reported by Corpas et al.
(2002a). The authors found significant reductions in the liver weight of pups and in hepatic
glycogen that correlated with increased blood glucose levels. The authors also reported
reductions in liver protein, lipid levels, and alkaline and acid phosphatase activities but did not
find any gross structural alterations in liver tissue. These and other studies are summarized in
Table AX5-10.3.

5.10.1.4 Effect of Lead on Hepatic Oxidative Stress
Although several mechanisms have been proposed to explain Pb toxicity, no mechanism
has been defined explicitly. Recent literature on Pb toxicity suggests oxidative stress as one of
the important mechanisms of toxic effects of Pb in liver, kidneys, brain, and other organs.
Schematic representation of the various mechanisms by which Pb induces lipid peroxidation is
shown Figure 5-20. Lead toxicity to the liver has been found to be associated with significant
accumulation of Pb in the liver. This results in the accentuation of lipid peroxidation with
concomitant inhibition of antioxidant enzymes (i.e., SOD, catalase, GSH peroxidase, GSH
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Acetyl-CoA
HMG-CoA

Jl HMGR HMGR-3-hydroxy-3 methyl glutaryl- f Pb
^ CoA Reductase
Mevalonate

I I FPPS Farnesyl pyrophosphate synthase T Pb

Farnesyl pyrophosphate

4J. SQS Squalene synthase T Pb

Squalene

I I Squaline oxidosferol

Lanosterol

J I cyp51 Lanosterol 14-a-demethylase T Pb

Cholesterol

I I cyp7A1 Cholesterol 7-a-hydroxyiase t Pb

Bile acids
Figure 5-19. Flow diagram indicating the lead effects on the cholesterol synthesis
pathway.
Pb-lnduced Oxidative Stress
I
Cell membrane: Genome: Antioxidant system:
• Interaction with PUFA • Accumulated ALA acts as * Inhibition of functional
• FA chain elongation alkylating agent and forms sylfhydryl groups on enzymes:
• Altered membrane enzymes DNA adducts SOD, GSH peroxidase,
• Altered solute transport • Interference with Zn-binding GSH reductase
• Altered signal transduction proteins in transcription • Depletion of GSH
complex
Figure 5-20. Schematic diagram illustrating the mode of lead-induced lipid peroxidation.
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reductase) and a simultaneous increase in GSSG with a reduction in GSH/GSSG ratio (Sandhir
and Gill, 1995; Aykin-Burns et al., 2003). However, Furono et al. (1996) studied the potential of
various redox-active metals to induce LPO in normal and alpha-linolenic acid-loaded rat
hepatocytes and suggested that Pb ions were not capable of inducing lipid peroxidation in such
hepatocytes.
The currently approved clinical intervention method is to give chelating agents that form a
soluble complex with Pb and remove the same from Pb-burdened tissues. The details of these
studies are provided in Annex Table AX5-10.4.

5.10.1.5 Lead-Induced Liver Hyperplasia: Mediators and Molecular Mechanisms
The biochemical and molecular events associated with Pb-induced hyperplasia has been
accumulating in the scientific literature. Lead nitrate, a known mitogen, is also considered to be
a carcinogen that induces liver cell proliferation in rats without any accompanying liver cell
necrosis. It has been recognized that this proliferation is a transient process and that apoptosis
plays a major role in the regression of Pb nitrate-induced hepatic hyperplasia (Nakajima et al.,
1995). Columbano et al. (1996) studied the cell proliferation and regression phases by apoptosis
in Wistar male rat liver by monitoring the incorporation of tritiated thymidine as a marker for
increased DNA synthesis. These studies demonstrated the production of Pb-induced
proliferation 3 days after a single injection of Pb nitrate with complete regression of hyperplasia
seen after 15 days. The authors suggested that the apoptosis process observed in the regression
phase also involved newly initiated hepatocytes. On the other hand, Dini et al. (1999) reported
the regressive or involutive phase as beginning 5 days post single injection of Pb nitrate.
Apostoli et al. (2000) evaluated the proliferative effects of various Pb salts (i.e., Pb acetate, Pb
chloride, Pb monoxide, Pb sulfate) using liver-derived REL cells. These authors reported that all
the Pb compounds tested showed dose- and time-dependent effects on the proliferation of REL
cells. Unlike other tumor promoters, Pb compounds did not exhibit effects on cell junctional
coupling. Liver hyperplasia induced by Pb nitrate has been shown to demonstrate sexual
dimorphism in all phases of the proliferation as well as in apoptosis (Tessitore et al., 1995).
Biochemical changes associated with liver hyperplasia in the intermediary metabolic pathways
were discussed in earlier sections of this chapter; the present discussion focuses on other
molecular characteristics of this process. As the numerous molecular networks involved in both
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the proliferation and apoptosis processes have many common mediators and pathways, it is very
difficult to provide a discussion without an overlap.
DNA hypomethylation has been recognized to play a major role in the proliferation of
cells in regenerating and in hepatic pre-malignant lesions when compared to normal non-dividing
liver cells. A single dose of Pb nitrate (75 |iM/kg body wt) has been found to cause extensive
hypomethylation in rat liver (Kanduc et al., 1991). Additional investigations from the same
group reported that this hypomethylation status of liver DNA by Pb nitrate changed significantly
with age and exhibited liver cell specificity (Kanduc and Frisco, 1992).
Investigations of cell cycle-dependent expression of proto-oncogenes in Pb nitrate
(10 jiM/100 g body wt)-induced liver cell proliferation by Coni et al. (1989) showed that peak
DNA synthesis occurred at 36 h after a single injection of Pb nitrate. In addition to DNA
synthesis, induced expression of c-fos, c-myc, and c-Ha-ras oncogenes was also observed in rat
liver tissue. Additional studies by the same group reported that Pb nitrate-induced liver
hyperplasia involved an increased expression of c-jun in the absence of c-fos expression (Coni
et al., 1993). The induced expression of c-myc persisted up to 40 h post Pb nitrate exposure.
Lead nitrate-induced liver proliferation and DNA synthesis, as monitored by 5-bromo-
2-deoxyuridine immunohistochemistry, led to DNA labeling in a few hepatocytes (Rijhsinghani
et al., 1993). The observed DNA synthesis appeared to be due to the increased activity and
expression of DNA polymerase-a observed at 8 h postexposure to a single injection of Pb nitrate
(Menegazzi et al., 1992). Along with DNA synthesis, poly (ADP-ribose) polymerase was also
induced by Pb nitrate (Menegazzi et al., 1990). Differential activation of various PKC isoforms,
downregulation of PKC-a, and marked activation of PKC-e in Pb nitrate-mediated liver
hyperplasia suggested the involvement of these PKC enzymes in DNA synthesis and related
signal transduction pathways (Tessitore et al., 1994; Liu et al., 1997).
Coni et al. (1992) reported the proliferation of normal and pre-neoplastic hepatic cells
treated with the plasma derived from male Wistar rats treated with a single injection of Pb
nitrate; this was the first report on the secretion of biological cell proliferation signals in the liver
after Pb nitrate treatment. These authors reported that DNA synthesis was detected as early as
30 min and persisted up to 5 days after Pb nitrate exposure. This observation has opened up the
inquiry into the involvement of various growth factors and other biological mediators in hepatic
hyperplasia. Shinozuka et al. (1994) investigated the expression of various growth factors (i.e.,
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hepatocyte growth factor, TGF-a, TGF-P) in rat liver after a single injection of Pb nitrate
(100 |iM/kg body wt) and reported the involvement of these growth factors in liver cell
proliferation. Additional studies by this group to observe LPS sensitivity in rats given Pb nitrate
reported that animals given a single injection of LPS up to 100 jig survived, whereas in the
presence of Pb nitrate, they tolerated only 6 jig of LPS, indicating that Pb nitrate may sensitize
the animals for LPS toxicity.
Earlier studies by Honchel et al. (1991) reported that coexposure of rats to Pb acetate
(15 mg/kg) and LPS or TNF showed markedly increased serum levels for various liver injury
parameters. They concluded that Pb may potentiate liver toxicity by LPS via a TNF-mediated
pathway. The role of TNF-a in Pb nitrate-induced liver cell proliferation was further
investigated by (Ledda-Columbano et al., 1994) who demonstrated the inhibition of Pb
nitrate-induced cell proliferation by pretreatment with dexamethasone, an inhibitor of TNF-a
expression. Additional studies by the same group evaluated the liver cell specificity in Pb
nitrate-induced cell proliferation (Shinozuka et al., 1996). They monitored the incorporation of
5-bromo-2-deoxyuridine by immunohistochemical analysis on rat liver as induced by Pb nitrate
and TNF-a and observed 5-bromo-2-deoxyuridine incorporation in hepatocytes and non-
parenchymal cells (i.e., Kupffer cells, endothelial cells, periportal nondescript cells), confirming
that Pb-induced liver cell proliferation was mediated by TNF-a. Kubo et al. (1996) used various
TNF-a inhibitors to further confirm the role of TNF-a in Pb nitrate-induced hepatocyte
proliferation. Menegazzi et al. (1997) reported that Pb nitrate-induced proliferation involved the
induction of iNOS along with TNF-a and that appeared to be mediated by a strong, prolonged
activation of NFtcB but not activator protein-1 (AP-1). Nemoto et al. (2000) investigated the
potential role of neurotrophins and their receptors in Pb nitrate-induced hepatic hyperplasia. The
expression profile of TNF-a, neurotrophins (i.e., nerve growth factor, brain-derived neurotrophic
factor neurotrophin-3 and (their receptors), tyrosine kinase receptor (Trk) and neurotrophin
receptor (p75NTR) were investigated in liver tissue after a single injection of Pb nitrate
(100 jiM/kg body wt). The Pb nitrate-induced increased expression of TNF-a preceded the
expression of the neurotrophins and their receptors. Based on these results, the author's
suggested that neurotrophins and neurotrophin receptors are involved in mediating mitogenic
signals related to hepatic hyperplasia.
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The regression phase of Pb-induced liver hyperplasia appears to be mediated by OS.
As discussed earlier, this process involves LPO and other cytokine mediators, including TNF-a.
Sieg and Billings (1997) reported that Pb potentiated cytokine-induced OS, producing a
significant decline in intracellular ATP concentration in mouse hepatocyte culture studies.
The authors suggested that cytotoxic interaction between Pb and cytokines (e.g., TNF-a and
IFN) may be mediated by oxidative DNA damage resulting from OS. The potential role OS
along with TNF-a has been implicated in the apoptosis of hepatocytes by Milosevic and Maier
(2000). Using freshly isolated cultures of hepatocytes and Kupffer cells and their co-culture
system exposed to Pb acetate (2-50 jiM) and LPS (0.1-1000 ng/mL), the authors reported that, in
the co-culture system, the Pb-LPS-induced release of TNF-a from the Kupffer cells, increased
nitric oxide levels by 6-fold and downregulated the acute phase protein, albumin, in hepatocytes.
From these observations the authors concluded that Pb-induced Kupffer cell-derived signals
promoted the toxicity of Pb in hepatocytes, resulting in hepatocyte death by proteolysis.
The importance of the Kupffer cells role in Pb nitrate-induced heptatocyte apoptosis was further
demonstrated (Pagliara et al., 2003a,b). These authors reported that in vivo hepatic apoptosis
including oxidative response induced by Pb nitrate, was prevented by pretreatment with
gadolinium chloride, a Kupffer cell toxicant that specifically suppresses Kupffer cell activity.
When treated hepatocytes were exposed in vitro to Pb nitrate, hepatocyte apoptosis was not
observed. On the other hand, hepatocyte apoptosis was evident when the hepatocytes were
incubated with culture medium derived from Kupffer cells that had been exposed to Pb nitrate.
Based on these studies, the authors concluded that heptocyte apoptosis was potentiated by
soluble factors secreted by Pb-exposed Kupffer cells. The role of activated Kupffer cells,
macrophages, and TNF-a in chemical-induced hepatotoxicity is presented schematically in
Figure 5-21.
Dini et al. (1993) investigated the expression of asialoglycoprotein receptors on the
surface of hepatocytes and galactose-specific receptors of non-parenchymal cells during the
apoptic phase of Pb-induced hepatic hyperplasia. A significant increase in asialoglycoprotein
receptor expression in hepatocytes coincided with massive apoptosis. Later studies from this
group demonstrated that sinusoidal liver cells predominantly phagocytosed the Pb nitrate-
induced apoptic hepatic cells and concluded that this process appeared to be mediated by the cell
surface carbohydrate receptors (i.e., mannose and galactose receptors) (Ruzittu et al., 1999).
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Chemical Exposure
c
Tissue Necrosis
Macrophage Activation
Secretion of Proinflammatory
Cytokines
TNF-a
IL-1
( Apoptosis j (
Inflammation J
( Proliferation j
Reactive Oxygen Species
Nitric Oxide
Generalized Liver
Damage
Figure 5-21. Hypothesis of chemical-induced liver injury generated primarily on the basis
of different types of inhibitors.
Pretreatment of rats with gadolinium chloride, a kupffer cell toxicant, was also found to
abolish the altered expression of galactose receptors (Pagliara et al., 2003b).
The role of glucocorticoid-mediated signal transduction in the hepatotoxicity of Pb was
evaluated by Heiman and Tonner (1995), using H4-IIE-C3 hepatoma cells (HTC). Acute
exposure of cells to Pb (300 nMT1 or 10 jiM) was found to inhibit processes involved
inglucocorticoid-mediated enzyme induction (e.g., tyrosine aminotransferase activity) in a
dose-dependent manner both at the transcriptional and translational level, without altering
glucocorticoid receptor binding characteristics. Tonner and Heiman (1997) also reported
Pb-induced hepatotoxicity by glucocorticoid-mediated signaling and its involvement in the
interference with calcium-mediated events as well as the differential modulation and
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translocation of protein kinase isoforms a and P into the nucleus. More information on these
and other related studies is summarized in Table AX5-10.5.

5.10.1.6 Effects of Lead on Liver Heme Synthesis
Effects of Pb on heme metabolism have been extensively investigated in major target
tissues such as liver and erythrocytes. Section 5.2 described Pb effects on heme synthesis, with
particular relevance to erythrocytes. The effects of Pb on heme synthesis in the liver and the role
of chelation therapy in this process are discussed in this section.
Fifteen percent of heme is produced in the liver. Heme metabolism in the liver is an
essential component of various cytochrome P450s that participate in cellular redox reactions and
xenobiotic detoxification pathways in the liver tissue and, hence, heme plays a vital role in liver
function (Jover et al., 1996). Due to the important and critical role of heme in liver function,
Pb-induced effects on hepatic heme metabolism are discussed below.
Initial studies on the effects of Pb nitrate on hepatic heme biosynthesis were reported by
Lake and Gerschenson (1978) using the rat liver cell line (RLC-GAI). The effects of various
organic metal compounds on ALAD activity have been studied by Bondy (1986). The authors
reported that triethyl Pb chloride has the same potency as Pb nitrate in inhibiting ALAD both
in vitro and in vivo, with liver and blood ALAD exhibiting similar sensitivities to Pb
compounds. By measuring the conversion of ALA into heme, these authors showed that heme
biosynthesis was inhibited by Pb in a dose dependent manner. Using a lipophilic complex of Pb
acetate + DTC to increase the cellular uptake of Pb, Osksarsson et al. (1989) demonstrated the
inhibition of ALAD activity in primary rat hepatocytes cultures. Lead-acetate has been reported
to inhibit ALAD activity in rabbit liver tissue without any effect on delta-aminolevulinic acid
synthase (ALAS) activity (Zareba and Chemielnicka, 1992). Exposure to Pb (500 ppm) in
drinking water did not inhibit hepatic ALAS, but did inhibit ALAD activity in mice (Tomokuni
et al., 1991). Exposure to Pb acetate (20 mg/kg body wt for 3 days) has been reported to
decrease hepatic ALAD and uroporphyrinogen activity (Satija and Vij, 1995). These authors
also reported that IP injection of zinc (5 mg/kg body wt for 3 days) conferred protection against
Pb acetate effects in liver tissue.
Effects of Pb on hepatic porphyrins, intermediate metabolites of heme metabolism, were
investigated by few researchers. Quintanilla-Vega et al. (1995) reported that 3T3-hepatocyte
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cultures, when incubated with a micromolar concentration of Pb acetate increased cellular
porphyrin content and excretion. This increased porphyrin production may have been due to
an accumulation of protoporphyrin and coproporphyrin, as in coproporphyrinuria, a well-
characterized sign of Pb intoxication (Ichiba and Tomokuni, 1987; Zareba and Chemielnicka,
1992). Dietary supplementation of selenium and monensin increased Pb-induced accumulation
of prophyrins in chicken liver (Khan and Szarek, 1994). Species-specific differences in the
effects of Pb on protoporphyrins were reported by Jacobs et al. (1998). These authors
investigated the effect of Pb on zinc protoporphyrin synthesis in cultured chick and rat
hepatocytes and observed decreased levels of protoporphyrin in rat hepatocytes, but no effect on
chick hepatocytes. Santos et al. (1999) also reported Pb-induced derangements (including
porphyrin metabolism) in rat liver heme metabolism, but these effects were far less severe than
those observed in erythrocytes. Their investigations on the effect of chronic alcoholism on Pb
effects in hepatic heme metabolism suggested no potentiation by alcohol.
Transferrin (TF) is the major iron-transport protein in serum and other biological fluids.
Transferrin can also has the capacity to transport other metals. Lead was found to inhibit TF
endocytosis and transport of iron across the cell membrane of rabbit reticulocytes (Qian and
Morgan, 1990). The effect of Pb on TF gene expression was investigated by Adrian et al. (1993)
using a transgenic mouse with the human TF gene. They found that Pb suppressed the
expression of TF transgene in mouse liver at the transcriptional level; however, the same dose of
Pb did not inhibit mouse endogenous hepatic TF gene expression. Lead exposure was also found
to inhibit recombinant TF expression in human hepatoma hepG2 cells. Other studies by the
same group found that Pb exposure suppressed the expression of endogenous TF in HepG2 cells
(Barnum-Huckins et al., 1997). These authors further suggested that Pb effects on hepatic TF
levels may also interfere with iron metabolism in humans. (See Annex Table AX5-10.6 for more
information on these and related studies.)

5.10.2 Gastrointestinal System and Lead Absorption
Lead enters the body by many routes, but primarily via the GI tract. The intestinal
epithelium serves as one of the body's primary interfaces with the outside world. The
transporting epithelia in the small intestine are characterized by layers of anatomically and
biochemically polarized cells that are connected to each other by tight junctions and resting on a
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basement membrane. Classically, the intestinal epithelium is thought of primarily as a barrier,
but it also is a highly reactive barrier. Even modest perturbations in its functions may lead to
diarrhea, constipation, malnutrition, dehydration, and infectious diseases (i.e., ulcerative colitis,
collectively referred as chronic intestinal inflammatory diseases) (Gewirtz et al., 2002).
Abdominal colic and constipation are symptoms of Pb poisoning, but its mechanism is not fully
understood. Studies have been carried out in the past decade to increase our understanding of the
fundamental mechanism(s) in order to extrapolate the experimental observations to human
health effects.
One key factor required to access Pb-related risks is the understanding and quantification
of bioavailability. Detailed discussions on the bioavailability of Pb, methodological approaches
for bioavailability measurements, bioavailability and speciation, etc. are discussed in detail in
Section 4.2. This section is primarily focused on the gastrointestinal absorption of Pb with
relevance to animal studies and in vitro test systems of intestinal origin.
The intestinal absorption of Pb is influenced by a variety of factors, including the
chemical and physical forms of the element, age at intake, and various nutritional factors.
Gastrointestinal absorption of Pb is thought to occur primarily in the duodenum. In the isolated
rat intestine, absorption, and, in particular, serosal Pb transfer activity (net transfer of Pb from
the small intestine lumen across the epithelium and into the serosal space) is highest in the
duodenum. The mechanisms of absorption may involve active transport and/or diffusion through
the intestinal epithelial cells. Both saturable and non-saturable pathways of absorption have been
inferred from the studies in different animal models, although the understanding of the former is
slightly more robust (Diamond et al., 1998).
Transport of Pb as a complex with proteins via endocytosis or as a complex with amino
acids are postulated as possible mechanisms. Direct evidence for transport of an organic Pb
complex has not been provided, but it seems possible.
In the cell, Pb interacts with a variety of intracellular ligands, including calcium-binding
proteins and high-affinity Pb-binding proteins. Transfer across the cell or basolateral membrane
(or both) involves a mechanism(s) that may be sensitive to vitamin D and iron status. Alternate
transport mechanisms via a Ca2+-Na+ exchanger, independent of regulation by vitamin D, are
also possible.
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5.10.2.1 Lead and In Vitro Cytotoxicity in Intestinal Cells
In vitro cytotoxicity of metal salts for 48 h was determined in the intestinal epithelial cell
line 1-407 by Keogh et al. (1994). The investigations identified rank order cytotoxicity in terms
of LC50 values: HgCl2 (32 |iM) > CdCl2 (53 |iM) > CuCl2 (156 |iM) > Ti2SO4 (377 |iM) > Pb
(NC>3)2 (1.99 mM). Further studies using a noncytotoxic concentration of butathione
sulphoxamine pretreatment for GSH depletion revealed that the cytotoxicity of Pb was
unaffected by GSH depletion (see Table AX5-10.7).

5.10.2.2 Alterations in Intestinal Physiology and Ultrastructure
Karmakar et al. (1986) investigated the pathologic alterations that occur in the intestine,
liver, and kidney of Pb-intoxicated rats upon short-term exposure to sublethal doses of Pb
(44 mg/Kg body wt) and reported degeneration of intestinal mucosal epithelium leading to
potential malabsorption.
The effect of low-concentration Pb acetate (0.1%) on the jejunal ultrastructure was studied
by Tomczok et al. (1988) in young male rats. The studies revealed that the villi of jejunum of
rats exposed to Pb for 30 days had a rough appearance on the surface, which could be associated
with a distortion of glycocalyx layer. Areas of extensive degenerative lesions were also
observed on the surface of most villi on the 60th day of exposure. All intestinal epithelial cells
exhibited various degrees of glycocalyx disturbance, indicating that pronounced toxic effects of
Pb were related to modifications of the biochemical properties of the surface coat of the cells.
These authors also reported the appearance of goblet cells and of Pb deposition along the goblet
cell membrane in blocks of tissue along the border between duodenum and jejunum. Continued
treatment up to 60 days resulted in mucus droplets in the cytoplasm of goblet cells, along with
deposition of silver salts indicative of Pb in these cells. These results demonstrated the
significance of goblet cells in Pb detoxification.
In another study on the ultrastructure of rat jejunum exposed to Pb acetate (100 mg/kg
body wt/day), Tomczok et al. (1991) found that 30-day treatment resulted in numerous small,
rough-membraned vesicles and dilated golgi complexes in the cytoplasm. Continued treatment
for 60 days resulted in vacuolated cytoplasm associated with the golgi complexes, rough-
membraned vesicles, and dilated cisternae. Also, the surface of the intestinal epithelial cell
microvilli showed evidence of Pb deposition, as evidenced by Timm sulfide silver reaction sites.
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5.10.2.3 Intestinal Uptake and Transport
Infants are a particularly susceptible population for Pb toxicity, possibly due to the
immaturity of the digestive tract, feeding pattern, or source of Pb. To investigate these aspects,
Henning's group (Beach and Henning, 1988; Henning and Cooper, 1988) carried out a series of
experiments using suckling rat pups and reported that Pb in rat and bovine milk and infant milk
formula was primarily associated with casein micelles. Casein-bound Pb may be the most
common form of Pb presented to the small intestine (Beach and Henning, 1988). Other studies
by this group investigated potential differences in the mechanisms when Pb was presented in
ionic or milk-bound form, using 203Pb as a tracer. These studies clearly showed that when 203Pb
was administered intragastrically as a soluble salt, it was primarily accumulated in the
duodenum, regardless of dose or vehicle used. In contrast, substantial accumulation of 203Pb
was found in the ileal tissue following Pb administration in milk. These studies clearly indicated
strikingly different patterns in the intestinal accumulation of ionic and milk-bound Pb and
suggest a greater toxicity for Pb in drinking water compared to Pb ingestion via milk (Henning
and Cooper, 1988).
Dekaney et al. (1997) investigated the uptake and transport of Pb using intestinal
epithelial cells (IEC-6). The authors observed that Pb accumulation in Pb-exposed (5-10 jiM)
IEC-6 cells was time- and dose-dependent up to 1 h and that reduction of the incubation
temperature significantly reduced the total cellular Pb content of IEC-6 cells. Simultaneous
exposure to Zn resulted in decreased cellular Pb content compared to cells exposed to Pb only.
Exposure of cells to ouabin or sodium azide has been found to increase Pb accumulation in the
cells compared to cells treated with Pb (5 jiM) alone. These studies clearly demonstrate that Pb
transport in IEC-6 cells is time- and temperature-dependent, involves the presence of sulfydryl
groups, and competes with the uptake of Zn.
Lead speciation and transport across intestinal epithelium in artificial human digestive
fluid (chyme), both in vivo and in vitro, in Caco-2 cells were evaluated by Oomen et al. (2003).
In vivo studies indicated that in chyme, Pb-phosphate and Pb-bile complexes are important
fractions. The metal ions dissociated from these complexes can subsequently be transported
across the intestinal epithelium or they may traverse the intestinal membrane. In vitro studies, on
the transport of bioaccessible Pb across the intestinal epithelium in Caco-2 cells exposed to
diluted artificial chyme for 24 h, indicated that 3% of the Pb was transported across the cell
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monolayer. Lead associated with cells in a linear relationship to the total amount of Pb in the
system. Bile levels were not found to affect the fraction of Pb associated with the cells. The free
Pb2+ concentration in chyme was negligible. Extrapolating these results to the in vivo situation,
the authors concluded that Pb species other than the free metal ion may have contributed to the
Pb flux towards the cells, possibly involving the dissociation of labile Pb species, such as
Pb-phosphate and Pb-labile complexes and the subsequent transport of the released free metal
ions toward the intestinal membrane.

5.10.2.4 Alterations in Gastrointestinal Motility/Gastrointestinal Transit and Function
The effect of Pb on contractility of rat duodenum was determined in vivo in rats given an
oral dose of Pb acetate (44 mg/kg per day, Pb as 53 mM/L for 4 weeks) to investigate the
possible mechanisms associated with Pb-induced abdominal colic and constipation (Karmakar
and Anand, 1989). Deodenal motility and the amplitude of contractility of rat duodenum were
decreased significantly in the Pb-exposed rats, leading the authors to conclude that there was a
fundamental change in the contractility of the intestinal tract due to Pb intoxication.
Chronic Pb ingestion through drinking water (2-5 mg/mL, Pb acetate for 55 days) caused
a 20-fold increase in urinary excretion of D-ALA and an increase in blood Pb level (80 |ig/dL),
without any perturbations in propulsive motility of guinea pig colon (Rizzi et al., 1989). On the
other hand, Lawler et al. (1991) observed no changes in gastric contractions during ingestion in
red-tailed hawks exposed to Pb acetate (0.82 or 1.64 mg/kg body wt for 3 weeks). This low level
of exposure has also been found to have no bearing on the regular passing of pellets of
undigested material. Shraideh (1999) studied the effect of triethyl Pb chloride on the rhythmic
and peristalitic contractile activity of ileum isolated from Swiss mice. These authors observed
no significant effect below 40 jiM of TEL, while higher concentrations (40-120 jiM) caused
changes in contraction rhythm. These studies also reported that TEL above 120 jiM induced
irreversible changes in the ileal contractile activity. These and related studies are summarized in
Table AX5-10.8.

5.10.2.5 Lead, Calcium, and Vitamin D Interactions in the Intestine
The complex biological interactions between Pb and calcium have been recognized and
demonstrated in virtually every type of tissue. Studies of high-affinity Pb binding to intracellular
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calcium receptors and transport proteins, as well as the involvement of Pb in calcium-activated
and calcium-regulated processes, have added to our understanding of the effects of Pb on
biological processes at the cellular level. The intestinal absorption of Pb is influenced by a
variety of factors, including chemical and physical forms of the element, age at intake, and
various nutritional factors. Work dating back to the 1940s established that the deposition of Pb
in bone and soft tissue significantly increases under conditions of dietary calcium and
phosphorus deprivation or by the administration of vitamin D to rachitic animals. Later, in the
1970s, it was demonstrated that dietary calcium status was a major contributing factor
determining relative susceptibility to Pb intoxication.
Fullmer's group (Fullmer and Rosen, 1990; Fullmer, 1991, 1992, 1997) carried out a
series of studies to investigate the potential interaction between calcium and Pb in the ingestion
and intestinal absorption of Pb. Various parameters, such as absorption kinetics for Ca and Pb,
activity of alkaline phosphatase, expression of the clabindin D gene, and the potential role of
endocrine function in this interaction (as assessed by cholecalciferol and its active hormonal
form, 1,25-dihydroxycholecalciferol levels) were investigated. Fullmer and Rosen (1990)
observed that chicks fed with low (0.5%) and adequate (1.2%) dietary calcium and exposed to Pb
(0-0.8%) exhibited differential effects on intestinal Ca absorption depending on their dietary Ca
status. In the chicks fed a low-calcium diet, Pb inhibited intestinal Ca absorption and calbindin
D and alkaline phosphatase synthesis in a dose-dependent fashion. On the other hand, chicks fed
the normal diet, showed no inhibition of Ca absorption. Based on these results, the authors
postulated that Pb-induced alterations in intestinal Ca absorption may involve cholecalciferol and
the endocrine system. In an extension of this study using young growing chicks, Fullmer (1991)
observed similar results in 2-week Pb-exposed, but not in 1-week exposed, chicks.
As dietary Ca deficiency is associated with a marked increase in the body burden of Pb
and in the susceptibility to Pb toxicity during chronic ingestion, Fullmer (1992) examined the
effects of vitamin D supplementation on intestinal Pb and Ca absorption. When vitamin D-
deficient chicks received physiologic amounts of vitamin D (O.lmg/day), intestinal 203Pb and
47Ca absorption rates were elevated by 4- and 8-fold, respectively. Along with this, calbindin D
and alkaline phosphatase activities were also significantly elevated. Ingestion of even the
highest level of Pb (0.8 %) during the repletion phase had no effect on intestinal Ca absorption.
To further understand the Pb-Ca interactions and the potential involvement of vitamin D on
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intestinal absorption, Fullmer (1997) evaluated serum levels of 1,25-dihydroxyvitamin D. Lead
ingestion and Ca deficiency alone, or in combination, generally increased serum 1,25-
dihydroxyvitamin D levels over most of the ranges of Pb or Ca studied. However, in severe Ca
deficiency, Pb ingestion resulted in marked decreases in serum 1,25-dihydroxyvitamin D,
intestinal Ca absorption, and calbindin D mRNA. From these studies using response surface
models, Fullmer (1997) concluded that the interactions between Pb and Ca were mediated via
changes in circulating 1,25-dihydroxy vitamin D hormone, rather than via direct effects on the
intestine.
Similar to Ca deficiency, iron deficiency has also been found to increase intestinal
absorption of Pb, as indicated by increased blood and kidney Pb levels in iron-deficient rats
exposed to dietary Pb; but the mechanistic details are not known (Crowe and Morgan, 1996).
These and other related studies are summarized in Table AX5-10.9.

5.10.2.6 Lead and Intestinal Enzymes
Differential effects of Pb on intestinal brush border enzyme activity profiles were reported
by Gupta et al. (1994). Across a concentration range of 0.5-6.0 mM, Pb acetate was found to
significantly inhibit Ca-Mg-ATPase, g-glutamyl transpeptidase, and acetylcholinesterase
activities in a dose-dependent manner without effects on alkaline phosphatase.
Cremin et al. (2001) investigated the effects of oral succimer on the intestinal absorption
of Pb in infant rhesus monkeys. These studies indicated that chelation therapy with DMSA for
two successive 19-day periods significantly decreased GI absorption of Pb and increased urinary
excretion of endogenous lead (see Table AX5-10.9).

5.10.3 Summary
Extensive in vivo and in vitro experimental evidence has accumulated over the past
20 years and increased our understanding of the potential toxic effects of Pb in the hepatic
system. These studies ranged from simple biochemical studies to molecular characterizations of
the induction of drug-metabolizing enzymes, liver hyperplasia, and the protective effects of
chelation therapy.
• Rat liver microsomal cytochrome P-450 levels were found to decrease with a single dose
exposure of Pb nitrate. Inhibition of both constitutive and induced expression of
microsomal P450 Al and A2 activity occurred. Simultaneous induction of the activities
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of phase II drug metabolizing enzymes with decreased phase I enzymes with single
exposure to Pb nitrate suggests biochemical properties similar to hepatic nodules.

• Newer studies examined the induction of GST-P at both transcript onal and translational
levels using in vitro systems and indicated a role for Pb nitrate and Pb acetate in the
induction process. On the other hand, triethyl Pb compounds have been found to
suppress the activity of various GST isoforms.

• Studies on Pb-induced liver hyperplasia demonstrated de novo synthesis of cholesterol,
alterations in the gluconeogenic mechanism, as well as DNA hypomethylation and
subsequent changes in the expression of protooncogenes.

• Lead-induced alterations in cholesterol metabolism appear to be mediated by the
induction of several enzymes related to cholesterol metabolism and the decrease of
7 a-hydroxylase, a cholesterol catabolizing enzyme. This regulation of cholesterol
homeostasis is modulated by changes in cytokine expression and related signaling.

• Studies using an inhibitor to block TNF-a have clearly demonstrated TNF-a as one of the
major mitogenic signals that mediate Pb nitrate-induced liver hyperplasia. Lead-induced
hyperplasia also appears to be modulated by neurotrophins and their receptors.

• In vitro co-culture systems with Kupffer cells and hepatocytes suggested liver cell
apoptosis is mediated by Kupffer cell-derived signals and Pb-induced oxidative stress.

• Newer experimental evidence suggests that Pb-induced alterations in liver heme
metabolism involves perturbations in ALAD activity, and porphyrin metabolism,
alterations in Transferrin gene expression, and associated changes in iron metabolism.

• Limited experimental evidence on the role of weight loss on liver Pb burden in exposed
animals indicate that liver Pb content increases even in the absence of prolonged
continued exposure.

• Gastrointestinal absorption of Pb is influenced by a variety of factors, including chemical
and physical forms of the element, age at intake, and various nutritional factors. The
degeneration of intestinal mucosal epithelium leading to potential malabsorption and
alterations in the jejunal ultrastructure (possibly associated with distortion of glycocalyx
layer) have been reported in the intestine of Pb-exposed rats.

• Lead in rat and bovine milk and, also, infant milk formula was demonstrated to be
primarily associated with caseine micelles.

• Tracer studies using 203Pb indicated that intragastric administration of Pb as a soluble salt
resulted in Pb primarily accumulating in the duodenum, regardless of dose or vehicle
used, whereas Pb from milk was found to be taken up by ileal tissue. Studies also
suggested Pb ingestion through water was more toxic than ingestion through milk.

• Lead induced decreases in duodenal motility and amplitude of contractility of the
intestinal tract have been reported for rats.

• Nutritional studies using various levels of Pb, Ca, and vitamin D in the diet indicate
competition of Pb with Ca absorption. Supplementation with vitamin D has been
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reported to enhance intestinal absorption of Ca and lead. Physiological amounts of
vitamin D administered to vitamin D-deficient rats resulted in elevated Pb and Ca levels.
In the case of severe Ca deficiency, Pb ingestion results in a marked decrease in serum
1,25-dihydroxy vitamin D.
Overall, our understanding of Pb effects on hepatic and gastrointestinal systems using in
vitro cell culture models and in vivo animal models has increased greatly compared to the 1986
AQCD. Significant insights have emerged regarding the role of Pb in hepatic cholesterol
synthesis, the role of inflammation in Pb-induced hepatotoxicity, and the contribution of newer
chelation therapy in the amelioration of Pb-induced oxidative burden. Similarly, our knowledge
has greatly enhanced as to the absorption, transport, and toxicity of Pb in the gastrointestinal
tract.
5.11 LEAD-BINDING PROTEINS
Lead-binding proteins that are constitutively expressed within the cells and bind Pb can be
classified into two types of protein. The first type of Pb-binding proteins are inducible, i.e., their
concentration increases after exposure to Pb. The second type of Pb-binding proteins have
binding sites that are saturable by Pb, but no discernible increase in protein content occurs after
exposure to Pb. The second type is, perhaps, most pertinent to enzymes that can be inhibited
byPb.
The history of research on Pb-binding proteins dates back to 1936, when the presence of
intranuclear inclusion bodies in the liver and kidney as manifestations of Pb poisoning was first
described (Blackman, 1936). Later, detailed studies of the composition of renal tubular
intranuclear Pb inclusion bodies and consequent alterations in mitochondrial structure and
function followed.

5.11.1 Lead-Binding Proteins within Intranuclear Inclusion Bodies
in Kidney
Goyer (1968) examined the renal tubules of rats fed 1% Pb acetate for up to 20 weeks,
and found that dense, deeply staining intranuclear inclusions were located in the straight portion
of the proximal tubules, accompanied by swollen, globular or ovoid, closely packed
mitochondria with many marginated, irregular, or vesicular cristae. Accompanying these
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mitochondrial changes was the presence of generalized aminoaciduria. Goyer et al. (1968) also
isolated mitochondria from Pb-exposed and control rats and demonstrated that mitochondria
from the Pb-exposed rats showed reduced rates of respiration and oxidative phosphorylation.
Lead within the kidneys in Pb-poisoned rats was found to be concentrated in the nuclei
and, within nuclei, in the nuclear inclusion body (Goyer et al., 1970a,b). Choie and Richter
(1972) showed that rapid induction of inclusion bodies by injections of Pb salts in the rat resulted
in cytoplasmic inclusions, suggesting that they were precursors to the intranuclear inclusions.
This was further confirmed by McLachlin et al. (1980) who showed in tissue culture studies of
rat kidney cells incubated with lead that the cytoplasmic inclusion bodies preceded and
disappeared shortly after the appearance of nuclear inclusion bodies.
Lead-containing nuclear inclusions were also found in organs other than the kidney,
including liver and glial cells of the central nervous system (Goyer and Rhyne, 1973). Moore
et al. (1973) dissolved the rat renal intranuclear inclusions in strong denaturing agents and found
that the protein in the inclusions is acidic, with high levels of aspartic acid, glutamic acid,
glycine, and cystine. Moore and Goyer (1974) later characterized the protein as a 27.5 kDa
protein, which migrates as a single band on acrylamide gel electrophoresis. Repeated
intraperitoneal injections of CaNa2EDTA resulted in the disappearance of the inclusion bodies in
Pb-exposed rats, together with a marked decrease in kidney Pb levels (Goyer et al., 1978).
Shelton and co-workers have also explored the composition of Pb-binding proteins in the
nuclear inclusion proteins of Pb-exposed rat kidneys. Shelton and Egle (1982) first described a
32 kDa protein with an isoelectric point of 6.3, which was isolated from the kidneys of rats
treated with 1% Pb acetate in rat chow or 0.75% Pb acetate in drinking water for 13-17 weeks.
In contrast to Goyer and co-workers, they used two-dimensional gel electrophoresis to isolate the
protein from the nuclear inclusion bodies and demonstrated that it was present in Pb-exposed,
but not control, kidneys (hence, inducible). This protein has been termed p32/6.3. Inhibitor
studies with cycloheximide and actinomycin D (McLachlin et al., 1980; Choie et al., 1975) had
indicated earlier that protein synthesis was required for induction of the nuclear and cytoplasmic
inclusion bodies.
Egle and Shelton (1986) unexpectedly found that p32/6.3, now characterized by a
monoclonal antibody, was constitutively present in the cerebral cortex, both in neurons and
astrocytes. The protein was concentrated in the insoluble nuclear protein, findings similar to the
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Pb-exposed kidney. Brain p32/6.3 was detected in rat, mouse, dog, man, and chicken. In rat
brain, adult levels were achieved in 1 to 2 weeks after birth, whereas only trace amounts were
found at 3 days. Brain p32/6.3 increased between postnatal days 10 to 12 in the guinea pig and
days 15 to 21 in the rat, suggesting that the increase may be related in part to exposure to the
external environment (Shelton et al., 1993). When neuroblastoma cells were cultured after 1-day
and 3-day exposure to Pb, the abundance of p32/6.3 increased. Simultaneous incubation with Pb
and cycloheximide or actinomycin D showed an increase in p32/6.3, suggesting that Pb
selectively retards the degradation of the brain protein (Klann and Shelton, 1989). The amino
acid composition of partially purified p32/6.3 revealed a high percentage of glycine, aspartic and
glutamic acid (Shelton et al., 1990). Thus, the inducible protein, p32/6.3, can be extracted from
nuclear inclusion bodies from the Pb-exposed rat kidney, and a similar or identical protein from
adult rat brain. Whether the brain protein is constitutive or inducible by exposure to
environmental Pb has yet to be determined. Selvin-Testa et al. (1991) and Harry et al. (1996)
reported that developing rat brain astrocytes exposed to Pb developed an elevation in glial
fibrillary acidic protein (GFAP), a developmentally-regulated protein. Harry et al. (1996)
consider that the elevated levels of GFAP mRNA during the second postnatal week after Pb
exposure may reflect the demand on astrocytes to sequester Pb.
Oskarsson and Fowler (1985) examined the influence of pretreatment with Pb by a single
IP injection of Pb acetate (50 mg Pb per kg) 1, 3, and 6 days before injecting 203Pb. Rats were
sacrificed 24 h later and the kidneys were examined both microscopically and for the distribution
of 203Pb. At 3 days, rat kidneys displayed fibrillar cytoplasmic inclusions, but at 6 days, these
inclusions were less prominent and intranuclear inclusions were observed. 203Pb uptake at 6 days
was maximal in the purified nuclear fraction and in the nuclear inclusion bodies (7 x and
20 x control, respectively).

5.11.2 Cytoplasmic Lead-binding Proteins in Kidney and Brain
The remaining studies of non-Pb-stimulated cytoplasmic kidney and brain Pb-binding
proteins have been provided by Fowler and associates.
The first study (Oskarsson et al., 1982) reported on the Pb-binding proteins in kidney
postmitochondrial cytosolic fractions. Binding of 203Pb was found in two protein fractions of
control kidneys with molecular weights of 11.5 and 63 kDa. Binding was markedly decreased
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after Pb pretreatment. The use of cadmium to stimulate metallothionein synthesis did not
increase 203Pb binding to the 11.5 kDa protein. The two binding proteins were also present in
brain, but not in liver or lung. Subsequently, Mistry et al. (1985) demonstrated three Pb-binding
proteins (11.5, 63, and >200 kDa) in rat kidney cytosol, which had binding characteristics of
high affinity, low capacity with respective Kd values of 13, 40, and 123 nM. The 11.5 kDa and,
possibly, the 63 kDa proteins were capable of translocating Pb into the nucleus as shown by
uptake of 203Pb into nuclei incubated with tagged cytosolic proteins. Goering and Fowler (1984)
showed that the 11.5 kDa protein, but not the 63 kDa protein was capable of reversing
Pb-induced ALAD inhibition in liver homogenates. This effect was mediated both by chelation
of Pb by the Pb-binding protein and by donation of zinc to ALAD (Goering and Fowler, 1985).
Various divalent metal ions influence the binding of Pb to the rat kidney cytosolic binding
proteins, with an order of displacement of Cd2+>Zn2+>Pb2+. Ca2+ had no effect, while Fe2+ had a
cooperative effect (Mistry et al., 1986). These observations may account for the previously
demonstrated effect of concomitant Pb and cadmium administration in reducing total kidney Pb
(Mahaffey et al., 1981) and preventing the development of intranuclear inclusion bodies
(Mahaffey and Fowler, 1977).
Later studies by Fowler and Duval (1991) identified the rat renal Pb-binding protein as a
cleavage product of oc2-microglobulin, with a Kd of 10"8 M Pb. There are two forms of the
protein in the kidney, differentiated by the cleavage of the first 9-N terminal residues from the
higher-molecular weight form. Other studies by Smith et al. (1998) found two Pb-binding
proteins in environmentally exposed human kidneys, identified as acyl-CoA binding protein
(ACBP) or diazepam binding inhibitor (molecular weight 9 kDa) and thymosin (34 (molecular
weight 5 kDa). These polypeptides have a high affinity for Pb (Kd~14 nM).
In rat brain, Goering et al. (1986) and DuVal and Fowler (1989) explored the effects of
environmental Pb on Pb-binding proteins and the ability of rat brain Pb-binding proteins to
diminish the inhibition of hepatic ALAD by Pb (liver does not contain the Pb-binding protein).
In the first study, a brain protein of 12 kDa was described, in comparison to the kidney
Pb-binding protein of 9 kDa. Both competition of Pb binding between the brain Pb-binding
protein and ALAD and donation of zinc by the brain protein (shown by 65Zn uptake) were found
to account for the decreased ALAD inhibition. In the second study the rat brain Pb-binding
protein was described as having a molecular weight of 23 kDa, with significant levels of
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glutamic acid, aspartic acid, and cysteine. Polyclonal antibody to rat renal Pb-binding proteins
showed a lack of reactivity with the brain protein, indicating that the proteins are
immunologically distinct.
Fowler et al. (1993) examined monkey kidney and brain from non-Pb-treated animals and
isolated Pb-binding proteins that also had a relatively high content of aspartic and glutamic
amino acid residues and were similar in size to the rat Pb-binding proteins. Polyclonal
antibodies to oc-2 microglobulin and metallothionein did not cross-react with either monkey
kidney or brain proteins. Quintanilla-Vega et al. (1995) isolated a thymosin 04 and a second, as
yet unidentified, protein with a molecular weight of 20 kDa and a pi of 5.9 from brains of
environmentally Pb-exposed humans.
Lead also has been reported to bind to p32/6.3, a low abundance, highly conserved
nuclear matrix protein that becomes a prominent component of Pb-induced intranuclear inclusion
bodies (Klann and Shelton, 1989). Expression of this protein increases significantly during
ontogeny, and was proposed to be useful as an indicator of neuronal maturation (Klann and
Shelton, 1990). Expression also increases markedly in the presence of acute Pb2+ exposure in
vitro, suggesting that Pb2+ either structurally alters the protein or inhibits a protease for which
p32/6.3 is a substrate (Shelton etal, 1993).
Recently an astroglial glucose-regulated protein (GRP78) has been identified that acts as a
molecular chaperone in endoplasmic reticulum (Qian et a/., 2000, 2005a). Intracellular levels of
this protein are increased in cultured astroglia during a 1- week exposure to Pb2+. GRP78
depletion significantly increased the sensitivity of cultured glioma cells to Pb2+, as indicated by
the generation of reactive oxygen species. This suggests that GRP78 is a component of the
intracellular tolerance mechanism that handles high intracellular Pb accumulation through a
direct interaction. Thus, it appears that Pb2+ directly targets the protein and induces its
compartmentalized redistribution, enabling it to play a protective role in Pb neurotoxicity.
The generation of reactive oxygen species also has been reported to occur via Pb2+ binding to
astroglial copper-transporting ATPase, resulting in disruption of copper homeostasis (Qian etal.,
2005b).
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5.11.3 Lead-binding Proteins in Erythrocytes
Intra-erythrocytic Pb binding was initially attributed primarily to hemoglobin, molecular
weight 64 kDa (Barltrop and Smith, 1972; Raghavan and Gonick, 1977; Ong and Lee, 1980;
Lolin and O'Gorman, 1988), but more recent studies have ascribed the major Pb binding to
ALAD, molecular weight 240-280 kDa. In contrast to this protein, several studies have focused
on an inducible low molecular weight protein in workers chronically exposed to Pb and which
seems to have a protective effect. The first recognition of this protein was by Raghavan and
Gonick (1977) who found an -10 kDa protein in Pb workers but not in controls, following
Sephadex G-75 fractionation (Figure 5-22). Upon subsequent SDS-polyacrylamide gel
electrophoresis, the protein split into two bands, only the uppermost of which contained Pb
(Figure 5-23).
20r
1.0
0«-
LV/J-
o to
«SXX» 13JXX)
MW MW
-I 300
- 200
-noo
80 100 120 140
ELUTION VOLUME (ml)
160
Figure 5-22. Sephadex G-75 gel filtration of RBC hemolysate from lead-exposed
individual. Ultraviolet absorption and radioactivity of 210Pb are plotted
against elution volume. The column was calibrated with ovalbumin (mol wt
45,000) and ribonuclease (mol wt 13,700). Also indicated is the locus of
hemoglobin (Hb). Hemolysates from normal control individuals showed no
UV absorption or radioactivity in the volume eluting between 130 and
155 mL.
Source: Raghavan and Gonick (1977) with permission.
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t
B

Apparent
Molecular
Weight
+A 0,000
Figure 5-23. SDS-polyacrylamide gel electrophoresis of RBC hemolysates from normal
control (A) and lead-exposed individuals (B), and of low-mol-wt. lead-
binding protein (C) stained with coomassie blue.
Source: Raghavan and Gonick (1977) with permission.
Raghavan et al. (1980) then went on to fractionate the erythrocyte Pb into a hemoglobin
fraction, a 10 kDa fraction, free Pb, and a "residual Pb" fraction thought to be composed of
membrane Pb and a high-molecular weight fraction. Lead workers manifesting toxicity at both
high blood Pb and relatively low blood Pb levels showed high levels of residual Pb, attributed in
the workers with toxicity at low blood leads to a very low quantity of the 10 kDa fraction. In a
follow-up study, Raghavan et al. (1981) reported elevated levels of Pb in the high molecular
weight fraction (pre-hemoglobin) and in the membrane fraction in workers with toxicity at both
high and low blood Pb levels. Again, those with toxicity at low blood Pb had low levels of the
Pb bound to the 10 kDa protein. Membrane Pb was found to correlate inversely with membrane
NaK-ATPase: no correlation was seen with total blood Pb.
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Gonick et al. (1985) partially purified the 10 kDa protein by HPLC using a protein 1-125
column followed by isoelectric focusing on a sucrose gradient column. Three protein peaks
resulted: one of 30 kDa, and two of 10 kDa. Only one of the latter peaks contained Pb. This
peak had a pi of 5.3 and a molecular weight, determined by SDS-PAGE, of 12 kDa. The
majority of Pb was found in this peak, which also contained calcium, zinc, and cadmium. Amino
acid analysis showed a very high percentage of glycine (44%) and lower quantities of histidine,
aspartic acid, and leucine.
Ong and Lee (1980) studied the distribution of 203Pb in components of normal human
blood. Ninety-four percent of 203Pb was incorporated into the erythrocyte and 6% remained
inthe plasma. SDS-PAGE of plasma showed that 90% was present in the albumin fraction.
Within the erythrocyte membrane, the most important binding site was the high molecular
weight fraction, about 130-230 kDa. Within the erythrocytic cytoplasm, the protein band
associated with 203Pb had a molecular weight of 67 kDa as shown by the elution characteristics
on G-75 chromatography. This was thought to be hemoglobin.
Lolin and O'Gorman (1988) and Church et al. (1993 a,b), following the same procedure
as Raghavan and Gonick (1977), confirmed the findings of a low molecular weight protein in the
erythrocytes of Pb workers, but not found in control patients. Lolin and O'Gorman (1988)
quantitated the protein, which ranged from 8.2 to 52.2 mg/L RBC in Pb workers, but found none
in controls, again implying it to be an inducible protein. They found that the low molecular
weight protein first appeared when the blood Pb concentration exceeded 39 |ig/dL. A positive
correlation was seen between the amount of the intra-erythocytic low molecular weight protein
and dithiothreitol-activated ALAD activity but not the non-activated activity. Church et al.
(1993a,b) also confirmed the findings of Raghavan and Gonick (1977). In 1993a, they described
two patients with high blood Pb levels: an asymptomatic worker with a blood Pb of 180 |ig/dL,
and a symptomatic worker with a blood Pb of 161 |ig/dL. In the first patient, -67% of the
erythrocyte Pb was bound to a low molecular weight protein of-6-7 kDa. In the second patient,
the protein only contained 22% of the total erythrocytic Pb. Church et al. (1993b) found that a
sample of the low molecular weight protein purified from Pb workers, which they termed protein
M, had characteristics of metallothionein, such as a molecular weight of 6.5 kDa, a pi between
4.7 and 4.9, and a greater UV absorbance at 254 nm than at 280 nm. Amino acid composition
showed 33% cysteine but no aromatic amino acids. This composition differed from that of the
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low molecular weight protein described by Gonick et al. (1985), which had a molecule weight of
12 kDa, a pi of 5.3, and amino acid analysis that showed no cysteine. This discrepancy might be
explained by a combined Pb and cadmium exposure in the Church et al. (1993b) study, which
may have produced a Pb-thionein.
Xie et al. (1998) used a Biogel A column instead of Sephadex G-75 to separate
Pb-binding proteins from erythrocyte hemolysates from a control patient and from Pb-exposed
workers. They clearly showed that the major Pb-binding was associated with a large molecular
weight protein, consistent with ALAD, in both the controls and Pb workers. When they added
increasing amounts of Pb to the blood of the control patient, a second low molecular weight
protein peak occurred, in which Pb binding was larger than the ALAD peak (Figure 5-24). This
second peak was also seen in a chronically Pb-exposed worker (Figure 5-25) and was estimated
to be less than 30 kDa in molecular weight. Thus these results are consistent with the
aforementioned studies.

5.11.4 Lead-binding Proteins in Rat Liver
Sabbioni and Marafante (1976) explored the distribution of 203Pb in rat whole tissue as
well as in subcellular liver fractions. By far the largest quantity of Pb recovered was in the
kidney, with lesser amounts in liver, spleen, and blood. Upon subcellular fractionation of the
liver, the majority of 203Pb was found in the nuclei, and most of the Pb was detected in the
nuclear membrane fraction, bound exclusively to membrane proteins. The intranuclear Pb was
associated with histone fractions. As reported by Oskarsson et al. (1982), Pb binding proteins
were not found in the cytoplasm of the liver.

5.11.5 Lead-binding Proteins in Intestine
Fullmer et al. (1985) showed in the chick and cow that, although Pb does not directly
stimulate Pb-binding proteins in the intestine, Pb can displace calcium from calcium-binding
proteins; and, thus, calcium-binding proteins may play a role in intestinal Pb transport. Purified
calcium-binding protein from chick and cow, as well as calmodulin, troponin C, and
oncomodulin were dialyzed against added labeled and unlabeled Pb or calcium. Results
disclosed high affinity binding sites, with greater affinity for Pb than for calcium. Similar results
5-278
-------
11
16
21
26
Pb Cone.
ALAD Activity
Se Cone.
Protein
31
36
0.8
0.6
(0
«
D
O
0.4 -f
0.2
30
S 25
O

•§ 20
. 15
s>

I 10

I
o 5

0
11 16 21 26
Fraction number
31
36
0.8
Of,
.6
®-

8
O
0.4
0.2
Figure 5-24. Chromatographic profiles of protein, ALAD activity and lead in human
erythrocytes incubated with 5% glucose solution containing lead acetate.
Blood was incubated (a) without lead (b) 10 uM lead (final concentrations).

Source: Adapted from Xie et al. (1998).
were obtained with calmodulin, troponin C, and oncomodulin, all members of the troponin C

superfamily of calcium-binding proteins.
5-279
-------
20
O 15
10
20
8 15
10
-Q
Q.
20
O
00

I
O 15
c
a
•-^

Dl 10
I
Pb Cone.
ALAD Cone.

Se Cone.
Protein
11
11
16
21
16
21
26
31
11 16 21 26 31

Fraction number
36
36
36
to
in
D
0.6 O

0.4 1

0.2 <
ID
10
D

0.6 O

£•

0.4 1
0.2
0.8 5T
in
in
O

0.6 Q.
0.2 <
Figure 5-25. Chromatic profiles of protein, ALAD activity, lead, and Se in the

erythrocytes of lead-exposed workers, (a) control, (b) subacute exposure, (c)

chronic exposure.

Source: Xie et al. (1998) with permission.
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5.11.6 Lead-binding Protein in Lung
Singh et al. (1999) described intracellular Pb-inclusion bodies in normal human lung
small airway epithelial cells cultured with either Pb chromate particles or sodium chromate.
Cells exposed to both forms of chromate underwent dose-dependent apoptosis. Lead-inclusion
bodies were found in nucleus and cytoplasm of Pb chromate, but not sodium chromate, treated
cells. Lead, but not chromium, was detected in the inclusion bodies by energy-dispersive X-ray
analysis. The protein within the inclusion bodies has not been analyzed.

5.11.7 Relationship of Lead-binding Protein to Metallothionein
Similarities of Pb-binding protein to metallothionein have been discussed earlier. Maitani
et al. (1986) commented that hepatic zinc-metallothionein could be induced by intravenous and
intraperitoneal injections of Pb into mice, but not by subcutaneous injection. Ikebuchi et al.
(1986) found that a sublethal dose of Pb acetate injected intraperitoneally into rats induced the
synthesis of a Pb-metallothionein in addition to zinc-metallothionein. The Pb-metallothionein
contained 28% half-cysteine and cross-reacted with an antibody against rat zinc-thionein II.
Goering and Fowler (1987 a,b) demonstrated that pretreatment of rats with zinc 48 and
24 h prior to injection of 203Pb resulted in both zinc and Pb co-eluting with a zinc-thionein
fraction on Sephadex G-75 filtration. In addition, both purified zinc-thionein-I and II bound
203Pb in vitro. Gel filtration of incubates containing liver ALAD and 203Pb demonstrated that the
presence of zinc-thionein alters the cytosolic binding pattern of Pb, with less binding to ALAD.
Zinc-thionein also donates zinc to activate ALAD. Goering and Fowler (1987b) found that
pretreatment of rats with either cadmium or zinc affected liver ALAD activity when incubated
with Pb. Liver and kidney zinc-thioneins, and to a lesser extent, cadmium, zinc-thionein
decreased the free pool of Pb available to interact with ALAD, resulting in attenuated ALAD
inhibition. Liu et al. (1991) further showed that zinc-induced metallothionein in primary
hepatocyte cultures protects against Pb-induced cytotoxicity, as assessed by enzyme leakage and
loss of intracellular potassium.
Qu et al. (2002) and Waalkes et al. (2004) have shown that metallothionein-null
phenotypic mice are more susceptible to Pb injury over a 20-week period than wild type mice.
Unlike the wild type mice, Pb-treated metallothionein-null mice showed nephromegaly and
significantly decreased renal function after exposure to Pb. The metallothionein-null mice
5-281
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accumulated less renal Pb than wild type and formed no Pb-inclusion bodies. When the
observations were extended to 104 weeks, renal proliferative lesions (adenoma and cystic tubular
atypical hyperplasia) were more common and severe in metallothionein-null than in wild type
mice. A metastatic renal cell carcinoma occurred in a metallothionein-null mouse, whereas none
occurred in wild type mice. Such studies lend credence to the view that metallothinein, or a
closely related gene, is involved in the formation of Pb-binding proteins in the kidney.

5.11.8 Is ALAD an Inducible Enzyme and Is It the Principal Lead-binding
Protein in the Erythrocyte?
The enzyme ALAD has been found to be the most sensitive indicator of Pb exposure and
toxicity (Granick et al., 1973, Buchet et al., 1976). In the 1980s, two articles were presented
appearing to show that ALAD is inducible after Pb exposure in humans. By comparing a
nonexposed control population of Pb workers and assaying ALAD by means of immunoassay or
as "restored" ALAD activity (i.e., incubation with heat, zinc and dithiothreitol) both articles
indicated that the amount of ALAD, as contrasted to ALAD activity, was increased by Pb
exposure (Fujita et al., 1982; Boudene et al., 1984). Similar findings were reported for the rat
(Fujita et al., 1981). Subsequent studies have focused on the effect of ALAD polymorphism on
the susceptibility to Pb intoxication. ALAD is a zinc-containing enzyme, which catalyzes the
second step of heme synthesis, i.e., catalyzes the condensation of two delta-aminolevulinic acid
molecules into one molecule of porphobilinogen (Boudene et al., 1984). It is a polymorphic
protein with three isoforms: ALAD-1, ALAD 1-2, and ALAD 2-2. Several studies have shown
that, with the same exposure to Pb, individuals with the ALAD-2 gene have higher blood Pb
levels (Astrin et al., 1987; Wetmur, 1994; Wetmur et al., 1991; Smith et al., 1995a; Bergdahl
et al., 1997; Perez-Bravo et al., 2004; Kim et al., 2004). Initially, it was thought that these
individuals might be more susceptible to Pb poisoning (Wetmur et al., 1991), but it is now
appreciated that the ALAD-2 gene offers protection against Pb poisoning by binding Pb more
securely (Kelada et al., 2001). In support of this statement, it can be cited that individuals with
the ALAD 1-2/2-2 genotypes, in comparison to those with the ALAD 1-1 genotype, have not
only higher blood Pb but also decreased plasma levulinic acid (Schwartz et al., 1997a), lower
zinc protopophyrin (Kim et al., 2004), lower cortical bone Pb (Smith et al., 1995b), and lower
amounts of DMSA-chelatable Pb (Schwartz et al., 1997b, 2000).
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The significance of erythrocyte ALAD binding to Pb was initially confirmed by a study
by Bergdahl et al. (1997), in which an FPLC Superdex 200 HR 10/30 chromatographic column
coupled to ICP-MS (for determination of Pb) was used to examine erythrocytes from Pb workers
and controls. They found the principal Pb-binding protein peak to be 240 kDa, rather than the
presumed hemoglobin peak reported by Barltrop and Smith (1972) and Raghavan and Gonick
(1977), using Sephadex G-75 chromatography. This was shown to be ALAD by binding to
specific ALAD antibodies. Two additional smaller Pb-binding peaks of 45 kDA and 10 kDa
were also seen, but not identified. Bergdahl et al. (1997) attributed the discrepancies in the
studies to the fact that Sephadex G-75 separates proteins in the range of 3 to 80 kDa, making the
separation of hemoglobin (molecular weight 64 kDa) from ALAD (molecular weight 240-
280 kDa) very difficult. In addition, the earlier studies had utilized binding of 203Pb or 210Pb to
identify the binding proteins, a technique which may have skewed the findings if ALAD were
already saturated. ALAD binding capacity for Pb has been measured at 85 |ig/dL in erythrocytes
or 40 |ig/dL in whole blood (Bergdahl et al., 1998), which would permit a greater degree of
binding to the low molecular weight component when blood Pb exceeded 40 |ig/dL. Bergdahl
et al. (1998) have speculated that the low molecular weight component might be acyl-CoA-
binding protein, identical to the kidney Pb-binding protein described by Smith et al. (1995b).
Goering and Fowler (1987) had reported earlier that the presence of low molecular weight high
affinity (Kd 10"8M) Pb-binding proteins in kidney and brain served as protection against ALAD
inhibition in those organs, whereas the absence of the low molecular weight proteins in liver
contributed to the greater sensitivity to ALAD inhibition in that organ.
A summary of key findings on Pb-binding protein is presented below, whereas more
detailed summarization of pertinent individual studies is provided in Table AX5-11.1.

5.11.9 Summary
• Nuclear inclusion bodies stimulated by Pb have been extensively investigated. The
nuclear inclusion body within the kidney and brain of rats contains a relatively insoluble
protein, tentatively identified as a 32 kDa protein with an isoelectric point of 6.3. The
nuclear inclusion body is preceded by the development and subsequent disappearance of a
cytoplasmic inclusion body. Whether the proteins within these two inclusion bodies are
similar or the same remains to be determined.
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• There appears to be a consensus that the enzyme, ALAD (a 280 kDa protein), is inducible
and is the major Pb-binding protein within the erythrocyte. ALAD polymorphism
influences the degree of Pb-binding, as the ALAD-2 phenotype binds more Pb in a
nontoxic fashion than ALAD-1. What is more confusing is the nature and importance of
the low molecular weight erythrocytic Pb-binding protein. There is no doubt that it
appears in Pb-exposed workers but not in controls and that its molecular weight is
-10 kDa. The in vitro addition of Pb to erythrocytes of controls results in progressively
increasing Pb binding to a low molecular weight protein peak migrating in the same
position as the low molecular weight protein from Pb workers. This confirms the fact that
once the binding capacity of ALAD is saturated, Pb shifts to the low molecular weight
protein. The nature of the low-molecular weight protein is also questionable; it has been
variously identified as a 12 kDa protein with a high percentage of glycine plus histidine,
aspartic acid, and leucine and as a 6.5 kDa molecule with a large percentage of cysteine
and a greater UV absorbance at 254 than 280 nm. The latter findings suggest that the
protein might be a metallothionein.

• Metallothionein is a protein that is mildly inducible by Pb but to a much greater degree by
zinc and cadmium. What is more significant is that Pb binds to pre-formed
metallothionein, stimulated by zinc or cadmium, so that under these conditions a
Pb-thionein forms. Thus, if concomitant Pb and cadmium exposure occurred in Pb
workers that could account for the finding of a metallothionein-like protein in those
workers.

• The possible role of metallothionein as a renal Pb-binding protein assumes greater
importance because of the work showing that metallothionein-null mice failed to respond
to Pb exposure by developing intranuclear Pb inclusion bodies or greatly increased Pb
content of the kidneys.

• Extensive studies of cytoplasmic Pb-binding proteins in non-Pb-treated rats, human, and
monkeys have been reported. The Pb-binding protein in rat kidney has been identified as
a cleavage product of a-2 microglobulin. The low molecular weight Pb-binding proteins
in human kidney have been identified as thymosin 04 (molecular weight 5 kDa) and acyl-
CoA binding protein (molecular weight 9 kDa). In human brain, the Pb-binding proteins
were thymosin 04 and an unidentified protein of 23 kDa. Antibodies to a-2 microglobulin
and metallothionein did not cross-react with monkey kidney or brain Pb-binding proteins,
suggesting species differences. Whether the low molecular weight human kidney and
brain Pb-binding proteins are similar or identical to the low molecular weight Pb-binding
proteins in erythrocytes is at present unknown. Perhaps some clarification would be
provided were subsequent investigators to contrast normal with Pb-exposed rats and to
measure the resting and inducible Pb-binding protein levels in kidney, brain, and
erythrocyte.
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6. EPIDEMIOLOGIC STUDIES OF HUMAN
HEALTH EFFECTS ASSOCIATED WITH
LEAD EXPOSURE

6.1 INTRODUCTION
This chapter assesses epidemiologic information regarding the biological effects of lead
(Pb) exposure, with emphasis on (1) qualitative characterization of Pb-induced effects and
(2) delineation of concentration-response relationships for key health effects of most concern.
Epidemiologic studies linking Pb exposure to health effects were earlier assessed in the 1986 Air
Quality Criteria for Lead (U.S. Environmental Protection Agency, 1986a), an associated 1986
Addendum (U.S. Environmental Protection Agency, 1986b), and a 1990 Supplement (U.S.
Environmental Protection Agency, 1990). Environmental exposures to Pb result from human
contact with multimedia exposure pathways (e.g., air, food, water, surface dust), as discussed
extensively in Chapters 3 and 4 of this document. In this chapter, while recognizing the
multimedia nature of Pb exposure of the general population, Pb exposure is generally indexed by
tissue Pb concentrations measured in biomarkers such as blood and bone. Many earlier studies
reported Pb effects on child development (psychometric intelligence), blood pressure and related
cardiovascular endpoints, heme biosynthesis, kidney, and reproduction and development.
Numerous more recent epidemiologic studies discussed in this chapter have further evaluated
these relationships to Pb exposure, thereby providing an expanded basis for assessment of health
effects associated with exposure to Pb at concentrations currently encountered by the general
U.S. population.
Special emphasis is placed here on discussion of the effects of Pb exposure in children.
Children are particularly at risk due to sources of exposure, mode of entry, rate of absorption and
retention, and partitioning of Pb in soft and hard tissues. The greater sensitivity of children to Pb
toxicity, their inability to recognize symptoms, and their dependence on parents and healthcare
professionals make them an especially vulnerable population requiring special consideration in
developing criteria and standards for Pb.
As discussed elsewhere in this document (Chapter 5), extensive experimental evidence
also links Pb exposure with health effects in laboratory animals. Thus, many of the reported
epidemiologic associations of Pb health effects have considerable biological credibility.
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Accordingly, the new epidemiologic studies of Pb assessed here are best considered in
combination with information from the other chapters on Pb exposure and on toxicological
effects of Pb in animals. The epidemiologic studies constitute important information on
associations between health effects and exposures of human populations to "real world" Pb
concentrations and also help to identify susceptible subgroups and associated risk factors.

6.1.1 Approach to Identifying Lead Epidemiologic Studies
Numerous Pb epidemiologic papers have been published since completion of the 1986
Lead AQCD/Addendum, and 1990 Supplement. A systematic approach has been employed to
identify relevant new epidemiologic studies for consideration in this chapter. In general, an
ongoing literature search has been used in conjunction with other strategies to identify Pb
epidemiologic literature pertinent to developing criteria for the National Ambient Air Quality
Standards (NAAQS) for Pb. A publication base was established using Medline, Pascal, BIOSIS,
and Embase, and a set of search terms aimed at identifying pertinent literature.
While the above search regime accessed much of the pertinent literature, additional
approaches augmented such traditional search methods. For example, a Federal Register Notice
was issued requesting information and published papers from the public at large. Also, non-EPA
chapter authors, expert in this field, identified literature on their own; and EPA staff also
identified publications as part of their assessment and interpretation of the literature. Lastly,
additional potentially relevant publications have been identified and included as a result of
external review of this draft document by the public and CAS AC. The principal criteria used for
selecting literature for the present assessment is to focus mainly on those identified studies that
evaluate relationships between health outcome and Pb exposure at concentrations in the range of
those currently encountered in the United States. New studies published or accepted for
publication through December 2005, as identified using the approaches above, have generally
been included in this Lead Air Quality Criteria Document (Lead AQCD), and additional efforts
have been made to identify and assess a few more recent but important studies.

6.1.2 Approach to Assessing Epidemiologic Evidence
Epidemiologic studies have evaluated Pb effects on a wide range of health endpoints that
include, but are not limited to: neurotoxic effects (e.g., psychometric intelligence, behavioral
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disturbances, and neurodevelopmental deficits), renal effects, cardiovascular effects,
reproductive and developmental effects, genotoxic and carcinogenic effects, and immune effects.
The epidemiologic strategies most commonly used in Pb health studies are: (1) cross-sectional
studies that examine the exposure and health outcome at a single point in time; and/or
(2) prospective longitudinal cohort studies that follow a group of individuals over time.
Both of these are types of observational, rather than experimental, studies.
An overall approach useful for assessing epidemiologic evidence was stated in the 2004
PM AQCD (U.S. Environmental Protection Agency, 2004) and is summarized here. That is, the
critical assessment of epidemiologic evidence presented in this chapter is conceptually based
upon consideration of salient aspects of the evidence of associations so as to reach fundamental
judgments as to the likely causal significance of the observed associations (see Hill, 1965).
The general evaluation of the strength of the epidemiologic evidence reflects consideration not
only of the magnitude and precision of reported Pb effect estimates and their statistical
significance, but also of the robustness of the effects associations. Statistical significance
corresponds to the allowable rate of error (Type I error) in the decision framework constructed
from assuming that a simple null hypothesis of no association is true. It is a conditional
probability; for statistical significance, typically there is a less than 0.05 chance of rejecting the
null hypothesis given that it is true. Robustness of the associations is defined as stability in the
effect estimates after considering a number of factors, including alternative models and model
specifications, potential confounding by copollutants, as well as issues related to the
consequences of measurement error.
Consideration of the consistency of the effects associations, as discussed in the following
sections, involves looking across the results obtained by various investigators in different
populations, locations, and times. Relevant factors are known to exhibit much variation across
studies, e.g., (1) presence and levels of other toxicants or pollutants of concern and (2) relevant
demographic factors related to sensitive subpopulations. Thus, consideration of consistency is
appropriately understood as an evaluation of the similarity or general concordance of results,
rather than an expectation of finding quantitative results within a very narrow range.
Looking beyond the epidemiologic evidence, evaluation of the biological plausibility of
the Pb-health effects associations observed in epidemiologic studies reflects consideration of
both exposure-related factors and dosimetric/toxicologic evidence relevant to identification of
-------
potential biological mechanisms underlying the various health outcomes. These broader aspects

of the assessment are only touched upon in this chapter but are more fully addressed and

integrated in Chapter 8 (Integrative Synthesis) discussions.

In assessing the relative scientific quality of epidemiologic studies reviewed here and to

assist in interpreting their findings, the following considerations were taken into account:

(1) To what extent are the biological markers used of adequate quality and sufficiently
representative to serve as credible exposure indicators, well-reflecting interpersonal
differences in exposure for specified averaging times?

(2) Were the study populations well defined and adequately selected so as to allow
for meaningful comparisons between study groups or meaningful temporal analyses
of health effects results?

(3) Were the health endpoint measurements meaningful and reliable, including clear
definition of diagnostic criteria utilized and consistency in obtaining dependent
variable measurements?

(4) Were the statistical analyses used appropriate, as well as being properly performed
and interpreted?

(5) Were likely important covariates (e.g., potential confounders or effect modifiers)
adequately controlled for or taken into account in the study design and statistical
analyses?

(6) Were the reported findings internally consistent, biologically plausible, and coherent
in terms of consistency with other known facts?

These guidelines provide benchmarks for judging the relative quality of various studies

and in assessing the overall body of epidemiologic evidence. Detailed critical analysis of all

epidemiologic studies on Pb health effects, especially in relation to all of the above questions, is

beyond the scope of this document.
6.1.3 Considerations in the Interpretation of Epidemiologic Studies of
Lead Health Effects
Prior to assessing results from recent Pb epidemiologic studies, issues and questions
arising from study designs and analysis methods used in the evaluation of Pb health effects are
first briefly discussed here. Study design can restrict the health effect parameters that can be
estimated. Separate considerations need to be made for acute versus chronic effect studies, as
well as individual versus aggregate-level analyses. Issues include measurement error, the
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functional form of relationships (especially at low Pb exposure levels) and the potential for
confounding. Aspects of these issues are briefly noted below and are then considered as studies
are reviewed in the following sections on specific health effect endpoints. These epidemiologic
considerations are further examined in Section 6.10 and Chapter 8 (Integrative Synthesis).
Measurement error is an important factor to consider, both for measurement of the health
effect outcome and the representativeness of the biomarkers of exposure (principally blood and
bone Pb) used in most key epidemiologic studies. For health outcome measures, the reliability
and validity of the measurements need to be assessed. In addition, the appropriateness of the
outcome measure for studying the hypothesis of interest needs to be determined. The critical
issues of outcome measurement and classification are, to some extent, endpoint-specific and are
therefore discussed further in ensuing individual sections.
Exposure misclassification can result in a notable reduction of statistical power in studies,
especially in those that focus on the lower end of the exposure range. Limitations of blood Pb as
an exposure index include use of a single blood Pb concentration to represent Pb body burden.
Also of concern is the most relevant blood sample collection time point in evaluating possible
associations with health outcomes (e.g., blood Pb values at ~2 years of age when peak Pb
exposure is expected versus those at later time point(s) concurrent with measurement of effect).
Another consideration is that similar blood Pb levels in two individuals do not necessarily reflect
similar body burdens. An added complication is that the relationship between Pb intake and
blood Pb concentration is curvilinear. Bone Pb determinations are typically considered a
measure of longer-term Pb exposure; but, the X-ray fluorescence (XRF) method typically used to
assess Pb levels in bone also has limitations, including the relatively high minimum detection
limit. The type of bone measured to determine Pb exposure is another important aspect.
The relationship between a measurement of a health outcome endpoint and an estimate of
Pb exposure based on a biomarker is an important concept. Modeling this relationship provides
a numerical slope that quantifies the relationship between Pb exposure and health outcome.
These models must address differences in the relationship at different concentration ranges of
exposure and present the functional form that best describes such data. Various models, both
linear and nonlinear, have been considered to examine Pb exposure-health effect relationships.
This is especially important at low Pb exposures. For example, a curvilinear relationship has
been reported for neurodevelopmental and cardiovascular outcomes at low Pb exposure levels.
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Depending on the subjects being examined for Pb exposure effects, various other factors
can lead to confounding of the relationship being considered. Potential confounding factors
largely depend on the health outcome of interest and the study population. Some potential
confounding factors in children, for whom the major health concerns include neurological and
developmental deficiencies, include: socioeconomic status (SES); nutritional status; quality of
home environment (e.g., HOME score); parental education; parental IQ; and birth weight, as a
few examples. For adults, factors that may confound the association between Pb exposure and
cardiovascular health outcomes include: age; diet; alcohol use; smoking; and potential for
copollutant exposures, such as to cadmium (Cd). For adult neurotoxic effects, potential
confounders include age, education, depressive symptoms, medications, alcohol use and
smoking. Control for potential confounding factors can be attempted at the study design phase
and/or during statistical analysis.

6.1.4 Approach to Presenting Lead Epidemiologic Evidence
In the main body of this chapter, each section starts by concisely highlighting important
points derived from the 1986 Lead AQCD/Addendum and the 1990 Supplement. Particular
emphasis is focused on studies and analyses that provide pertinent information of importance for
the critical assessment of health risks from Pb exposure. Not all studies are accorded equal
weight in the overall interpretive assessment of evidence regarding Pb-associated health effects.
Among well-conducted studies with adequate control for confounding, increasing scientific
weight is accorded in proportion to the precision of their effect estimates. To ensure a thorough
appraisal of the evidence, more detailed information on key features (including study design,
analysis, biomarkers of exposure, and health outcome results) of important new studies are
summarized in tables in the Annex for this Chapter 6 (Annex AX6).
In the main body text discussion, emphasis is placed on (1) new studies employing
standardized methodological analyses for evaluating Pb effects across different study populations
and providing overall effect estimates based on combined analyses of information pooled across
different cohort groups; (2) meta-analyses of individual studies conducted in various study
populations; (3) studies assessing Pb health effects at current relevant levels of exposure
(e.g., blood Pb levels <10 |ig/dL); and (4) studies conducted in the United States. Multiple
cohort studies are of particular interest and value due to their evaluation of a wider range of
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Pb exposures and large numbers of observations, thus generally providing more precise effect
estimates than most smaller scale studies of single cohorts. Furthermore, multiple cohort studies
have the potential to provide especially valuable evidence regarding relative homogeneity and/or
heterogeneity of Pb health effects relationships between different study populations. Also of
particular interest in recent years are those health effects observed at the lower range of Pb
exposure, as typically assessed using blood Pb levels. The potential impacts of the underlying
health status of populations and cultural differences in the case of intelligence testing (one of the
major health outcomes in children) also need to be accounted for in the assessment; thus, U.S.
studies are emphasized over non-U.S. studies. In accordance with the emphasis placed on the
Pb epidemiologic studies in this chapter, Chapter 6 Annex tables are organized by emphasis on
multiple cohort studies and U.S. studies.
In the ensuing sections, epidemiologic studies of the neurotoxic effects of Pb exposure in
children are discussed first, in Section 6.2. The neurotoxic effects of Pb on adults are then
assessed in Section 6.3. This is followed by discussion of the renal and cardiovascular effects of
Pb in Sections 6.4 and 6.5. Section 6.6 next discusses reproductive and developmental effects of
Pb, Section 6.7 discusses genotoxic and carcinogenic effects of Pb, and Section 6.8 discusses Pb
effects on the immune system. Lead effects on other organ systems (including the
hematopoietic, endocrine, hepatic, and gastrointestinal systems) are assessed in Section 6.9.
The effects of Pb on bone and teeth, as well as on ocular health, are also discussed in Section 6.9.
Finally, Section 6.10 discusses methodologic considerations and summarizes key epidemiologic
evidence for Pb-related health effects.
6.2 NEUROTOXIC EFFECTS OF LEAD IN CHILDREN
This section assesses epidemiologic evidence for neurotoxic effects of Pb exposure in
children. First presented are studies of neurocognitive effects of Pb in children, with a focus on
several prospective studies examining neurocognitive ability. Other topics include measures of
academic achievement, cognitive abilities, disturbances in behavior, mood, and social conduct,
measures of brain anatomical development and activity, gene-environmental interaction, and
reversibility of neurodevelopmental deficits. The neurotoxic effects of environmental and
occupational Pb exposure of adults are then discussed in Section 6.3.
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6.2.1 Summary of Key Findings on Neurotoxic Effects of Lead in Children
from 1986 Lead AQCD and Addendum, and 1990 Supplement
The 1986 Lead AQCD stated that children were particularly susceptible to Pb-induced
neural damage. In particular, human infants and toddlers below 3 years of age were considered
to be at special risk due to possible in utero exposure, increased opportunity for exposure
because of normal mouthing behavior of Pb-containing objects, and increased rates of Pb
absorption due to factors such as iron and calcium deficiencies.
Effective blood Pb levels for producing encephalopathy or death in children were noted in
the 1986 Lead AQCD as starting at 80 to 100 |ig/dL. Various types of neural dysfunction were
stated as being evident at lower blood Pb levels. Behavioral (e.g., reaction time, psychomotor
performance) and electrophysiological (e.g., altered electrophysiological patterns, evoked
potential measures, and peripheral nerve conduction velocities) effects were observed at blood
Pb levels as low as 15 to 30 |ig/dL and possibly lower. A concentration-response relationship
between blood Pb levels and IQ also was observed; a 1-2 point difference in IQ was generally
seen with blood Pb levels in the 15 to 30 |ig/dL range. However, Schroeder and Hawk (1987)
found a highly significant linear relationship between a measure of IQ and blood Pb levels over
the range of 6 to 47 |ig/dL among a cohort of all African-American children of low SES,
suggesting that IQ effects might be detected even at blood Pb levels below 15 to 30 |ig/dL.
The 1986 Addendum also discussed the newly published results of several prospective
cohort studies on the developmental effects of Pb in children. These studies improved upon
previous studies by utilizing longitudinal study design that followed children from the prenatal
stage, larger numbers of subjects, and better analytic techniques to more accurately measure
blood Pb levels. The four prospective studies (conducted in Boston, MA; Cincinnati, OH;
Cleveland, OH; and Port Pirie, Australia) reported significant associations between prenatal and
postnatal blood Pb levels and neurobehavioral deficits, after adjusting for various potential
confounding factors, such as maternal IQ and HOME (Home Observation for Measurement of
Environment) scores (Bellinger et al., 1984; Dietrich et al., 1986; Ernhart et al., 1985, 1986;
McMichael et al., 1986; Vimpani et al., 1985; Wolf et al., 1985). In these studies, the observed
maternal and cord blood Pb levels were fairly low, with mean levels of-10 |ig/dL. These results
led to the conclusion in the 1986 Addendum that neurobehavioral deficits, including declines in
Bayley Mental Development Index (MDI) scores and other assessments of neurobehavioral
-------
function, are associated with prenatal blood Pb exposure levels on the order of 10 to 15 |ig/dL
and possibly even lower, as indexed by maternal or cord blood Pb concentrations.
The 1990 Supplement updated evidence from the above-mentioned longitudinal cohort
studies and summarized results from other more recent prospective cohort studies conducted in
Glasgow, Scotland; Kosovo, Yugoslavia; Mexico City; and Sydney, Australia. Results from
several other international cross-sectional studies also were discussed. The collective evidence
from the various prospective cohort and cross-sectional studies reaffirmed the conclusions from
the 1986 Addendum that neurobehavioral effects were related to blood Pb levels of 10 to
15 |ig/dL and possibly lower. Further analyses of the Boston data indicated that deficits in MDI
could be detected in relation to cord blood Pb levels of 6 to 7 |ig/dL in children within the lower
strata for SES (Bellinger et al., 1988). In the Port Pirie study, the relationship between postnatal
blood Pb levels and MDI at two years of age provided little evidence of a threshold effect (Wigg
et al., 1988). Restricting the analysis to children with blood Pb levels below 25 |ig/dL yielded an
even stronger association between integrated postnatal blood Pb and McCarthy General
Cognitive Index (GCI) scores in the Port Pirie study (McMichael et al., 1988).
Impaired neurobehavioral development was associated with blood Pb measures in
pregnant women, umbilical cords, and infants up to at least 2 years of age; thus, no distinction
could be made as to whether this level of concern applied to only fetuses or infants or preschool-
age children. The issue of the persistence of the neurobehavioral effects from low-level Pb
exposure also was considered. Although the Boston and Cincinnati studies provided limited
evidence suggesting that the effects of prenatal Pb exposure on neurobehavioral development
were not persistent, the evidence available to support this conclusion was inadequate.

6.2.2 Introduction to Neurotoxic Effects of Lead in Children
Several major developments have occurred in Pb research on child neurodevelopment
following the 1986 Lead AQCD/Addendum and the 1990 Supplement. First, there has been an
attempt to broaden outcome assessments beyond neurocognitive deficits. The earlier emphasis
on neurocognitive measures (e.g., MDI, GCI, IQ) in previous studies is understandable from the
perspectives of the strong psychometric properties of most of these rigorously standardized
measures as well as the immediate pubic health concerns. Examples of other outcomes used to
assess neurodevelopment include the number of errors on tests of visual-motor integration, the
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time required to complete a task assessing manual dexterity, the number of errors and false
alarms on a continuous performance test, and the efficiency of short term memory. Additional
neurodevelopment outcomes include those which elucidate brain-behavior relationships or the
potential real life consequences of early exposure to Pb, such as academic and vocational failure
and maladjustment to the daily demands of living in a complex society. Thus, epidemiologic
studies of Pb neurotoxicity have been expanded to adopt measures of academic achievement,
specific cognitive abilities, behavior and mood, sensory acuities, neuromotor function, and direct
measures of brain anatomical development and activity. Another development has been the
initiation of nutritional and pharmacological intervention studies to assess the impact of
treatment on reducing blood Pb levels and preventing or moderating the degree of harm to the
central nervous systems of young children. Also, in addition to blood and tooth Pb, bone Pb has
emerged as a reliable biomarker of Pb exposure. The technology for the assessment of Pb in
cortical (tibial) and trabecular (patellar) bone using K-shell X-ray fluorescence (XRF) has
advanced to the point where it could be applied as a reliable and valid index of cumulative Pb
dose in neuroepidemiologic studies (Aro et al., 1994).
In recent years, more studies have investigated the impact of blood Pb levels below
10 |ig/dL on the developing brain. Average blood Pb levels in U.S. children ages one to five
years decreased from -15 |ig/dL to ~3 |ig/dL between 1976-1980 and 1991-1994, allowing
newer studies to examine the effects of low level Pb exposure on the neurodevelopment of
children (Centers for Disease Control and Prevention [CDC], 2000; Pirkle et al., 1998).
At the time of the last previous criteria review, it was recognized that estimating a
threshold for toxic effects of Pb on the central nervous system entailed a number of difficulties.
There is the critical question of reversibility or the persistence of Pb effects identified in infants
and preschoolers into school age and later. A given effect observed at younger ages may not
persist due to functional compensation or a return to a normal neuromaturational trajectory
(Dietrich et al., 1990). On the other hand, insults to the human brain may persist, making it
difficult to determine whether any measured insult is the result of current or past exposures.
An observed effect concurrent with a measured blood Pb concentration may be the result of
exposure in the child's earlier life in the womb or infancy. Another problem is that it is
sometimes difficult to distinguish between neurobehavioral effects due to Pb and effects owing
to the many social, economic, urban-ecological, nutritional, and other medical factors that are
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known to have important effects on neurobehavioral development. Equally important is the high
probability that the concentration-response relationship and even the neurobehavioral lesion
associated with childhood Pb exposure may vary as a function of these cofactors (Bellinger,
1995).
In the following sections, prospective cohort studies and cross-sectional studies of
neurocognitive ability published since the 1990 Supplement are first discussed. Then, studies
examining the effect of Pb on a variety of neurodevelopmental outcomes, including academic
achievement; specific cognitive abilities; disturbances in behavior, mood, and social conduct;
sensory acuities; neuromotor function; and brain anatomical development and acuity, are
discussed. This is followed by discussion of several issues involved in understanding Pb
neurotoxicity in children, including gene-environment interactions, reversibility of Pb effects,
times of vulnerability, and potential threshold levels for effects.

6.2.3 Neurocognitive Ability
6.2.3.1 Prospective Longitudinal Cohort Studies of Neurocognitive Ability
Several prospective longitudinal cohort studies were initiated in the 1980s because it
became widely recognized that the cross-sectional study design was inadequate to address a
number of research issues (U.S. Environmental Protection Agency, 1986a; World Health
Organization [WHO], 1977). These longitudinal studies were characterized by serial measures
of dose (blood Pb levels) spanning (in most cases) the prenatal and postnatal periods of central
nervous system development, thus helping to clarify the temporal association between exposure
and insult. Also, developmental assessments that extended into the school-age period were
planned to determine if early Pb-associated neurobehavioral impairments were persistent. It was
also determined that assessment of potential confounding factors should be comprehensive and
include measures of perinatal health, nutrition, maternal consumption of other neurotoxicants
during pregnancy, parental intelligence, and direct observations of parenting behavior. These
studies were also characterized by very careful attention to biostatistical issues and strategies
(Bellinger, 1995; Ernhart, 1995).
At the time of the 1990 Supplement, studies were underway or planned in the United
States, Australia, Scotland, the former Yugoslavia, and Mexico. These cohorts differed in the
source and degree of Pb exposure and in other important aspects, notably ethnicity and SES.
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Nevertheless, the early results from several of these studies have been largely responsible for the
emergence of the current perspective that blood Pb concentrations as low as 10 |ig/dL, or
perhaps even lower, may pose a risk for neurodevelopmental toxicity (Davis and Svendsgaard,
1987; U.S. Environmental Protection Agency, 1990). Most of the prospective studies underway
in 1990 continued to follow their subjects into the later preschool and school age years with age-
appropriate measures of intelligence. Continued follow-up of these cohorts was important due to
the following: (1) greater reliability and precision of measurements attained with assessments of
older children; (2) high predictability of adult intellectual functioning from measures of IQ in the
older child; and (3) examination of potential effects of Pb on important abilities that cannot be
easily tapped during infancy such as executive functions and higher order reasoning (McCall,
1979).
A unique aspect of this research was that most investigators agreed during the formative
stages of their projects to develop somewhat similar assessment protocols (Bornschein and
Rabinowitz, 1985). This has facilitated comparison of results across studies and allowed for
sophisticated meta- and pooled-analyses of these data (e.g., Pocock et al., 1994; Schwartz, 1994;
WHO, 1995; Lanphear et al., 2005; Rothenberg and Rothenberg, 2005).
In the following sections, further updates on the individual prospective cohort studies are
presented in chronological order of study initiation. The prospective cohort studies reviewed are
summarized in Annex Table AX6-2.1. Results of the meta- and pooled-analyses are presented
later in this section.

6.2.3.1.1 Boston Study
In the 1986 Addendum, the most advanced investigation at that time was the Boston
Prospective Study (Bellinger et al., 1984). The subjects were 216 middle-to upper-middle-class
Boston children, 90% of whom had cord blood Pb levels below 16 |ig/dL (maximum 25 |ig/dL).
The children in this cohort were generally of high SES standing. While this might limit
generalization of results to a wide population, this study enhanced the ability to isolate the effect
of low level Pb exposure on cognitive function, as there were no associations between cord blood
Pb level and several indicators of social disadvantage (e.g., receipt of public assistance, lower
educational achievement, unmarried) in this highly selected subsample. Cord-blood Pb levels in
the "high" group (mean 14.6 |ig/dL) were associated with lower covariate-adjusted scores on the
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Mental Development Index (MDI) of the Bayley Scales of Infant Development (BSID) at
6 months of age. It was concluded that although lower level Pb exposure in utero may result in
delays in early sensorimotor development, the Boston results did not allow estimation of the
persistence of these effects nor the public health significance of the findings.
In the 1990 Supplement, particular attention was focused on the Boston study, which was
among the more mature in terms of follow-up (Bellinger et al., 1987, 1991). With respect to the
effects of cord blood Pb concentrations on MDI assessed longitudinally from 6 to 24 months, the
Pb-associated deficits were evident across the entire range of blood Pb levels starting at
10 |ig/dL, which reinforced the previous designation of 10 to 15 |ig/dL as a blood Pb of concern
for early neurodevelopmental deficits. At ~5 years of age, significant associations of McCarthy
GCI with the cord blood Pb level (effect estimates not provided) and concurrent blood Pb level
(-2.26 points [95% CI: -6.0, 1.4] per one unit increase in In blood Pb) were not observed, but
the blood Pb level at 2 years of age (mean 6.8 |ig/dL [SD 6.3]) was significantly associated with
lower scores (-2.95 points [95% CI: -5.7, -0.2]). Boston investigators also examined the
relationship between Pb measured in shed deciduous teeth obtained from 102 children in their
cohort (mean 2.8 ppm [SD 1.7]) and GCI at 5 years of age. Prior to covariate-adjustment, there
was a very strong and significant relationship amounting to a decrement of 10.04 points
(95% CI: 2.6, 17.4) in GCI for each unit increase in In dentine Pb. However, in the
multivariable analysis, the effect estimate for the tooth Pb coefficient decreased to -2.51 points
(95% CI: -10.2, 5.2) per unit increase in In dentine Pb.
Since the 1990 Supplement, the Boston investigators reexamined 148 of their subjects at
10 years of age with the Wechsler Intelligence Scale for Children-Revised (WISC-R) and other
neurobehavioral assessments (Bellinger et al., 1992). They examined the association of WISC-R
scores at 10 years of age with blood Pb concentrations in the cord blood and at 6 months,
12 months, 18 months, 24 months, 57 months, and 10 years. Only blood Pb levels at 24 months
were significantly associated with full scale and verbal IQ and marginally associated with
performance IQ, after adjusting for HOME score, maternal age, birth weight, and maternal IQ.
The integrated average blood Pb level in this cohort over the first 2 years was 7.0 |ig/dL
(range 4-14 jig/dL). An increase of 10 |ig/dL in blood Pb level at age 2 was associated with a
decrement of 5.8 points (95% CI: 1.8, 9.9) in full scale IQ. These findings indicated that
children's performance was much more strongly associated with blood Pb levels at age 2 than
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with blood Pb levels at other ages. It is unclear whether this reflects (1) a special vulnerability of
the nervous system during this period; (2) the typical peaking of blood Pb levels in the second
year; or (3) random chance.
A reanalysis involving the total Boston cohort that employed nonparametric smoothing
revealed that the inverse association persisted at blood Pb levels below 5 |ig/dL (Schwartz,
1994). Bellinger and Needleman (2003) reanalyzed data on 48 children whose measured blood
Pb concentrations never exceeded 10 |ig/dL. Reduction in full scale IQ at 10 years was
significantly associated with blood Pb levels at 2 years of age following covariate adjustment.
A larger deficit of 15.6 points (95% CI not presented) per 10 |ig/dL increase in blood Pb levels
was observed in this cohort, compared to the 5.8 point deficit observed in the entire cohort.
These findings indicated that the inverse slope might be steeper at blood Pb levels below
10 |ig/dL.

6.2.3.1.2 Cincinnati Study
Interim results on a partial sample of 185 subjects from a cohort of 305 were available
from the Cincinnati prospective study in the 1986 Addendum and the 1990 Supplement (Dietrich
et al., 1986, 1987a). The Cincinnati study investigators reported an inverse relationship between
prenatal maternal blood Pb levels (mean 8.3 |ig/dL) and 6 month Bayley MDI. This effect was
mediated, in part, through Pb-associated reductions in birth weight and gestational maturity.
A more complete analysis of the full Cincinnati cohort confirmed these interim findings
(Dietrich et al., 1987b).
Further updates of the Cincinnati study appeared after the 1990 Supplement. For one,
the Kaufman Assessment Battery for Children (KABC) was administered to -260 children at
4 and 5 years of age (Dietrich et al., 1991; 1992). The principal findings at 4 years were that
higher neonatal blood Pb concentrations were associated with poorer performance on all KABC
subscales. However, this relationship was confined to children from the poorer families. After
full covariate adjustment, few statistically significant relationships remained. At 5 years of age,
postnatal blood Pb levels were associated with performance on all subscales of the KABC;
however, few statistically significant relationships remained after adjustment for covariates.
Nevertheless, it is of interest that, at both 4 and 5 years, the KABC subscale that assessed
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visual-spatial skills was among those that remained the most highly associated with various
indices of postnatal exposure following covariate adjustment.
At ~7 years, 253 children in the Cincinnati cohort were administered the WISC-R
(Dietrich et al., 1993a). In this cohort, -35% had at least one blood Pb concentration >25 |ig/dL,
whereas 95% exceeded 10 |ig/dL sometime during the first 5 years of life. Postnatal blood Pb
concentrations were inversely associated with full scale and performance IQ, after adjusting for
HOME score, maternal IQ, birth weight, birth length, child gender, and cigarette consumption
during pregnancy. Figure 6-1 presents the unadjusted and adjusted concentration-response
relationship between lifetime average blood Pb concentrations and performance IQ. After
covariate adjustment, a statistically significant relationship was observed between postnatal
blood Pb levels at 5 and 6 years of age and full scale IQ. Postnatal blood Pb levels at nearly all
ages (including the integrated average blood Pb level) were inversely associated with
performance IQ. Due to the high intercorrelation among blood Pb measures taken at different
time points, it was not practical to examine exposures during any given year for evidence of a
sensitive neurodevelopmental period. Concurrent blood Pb levels were strongly associated with
full scale IQ (-3.3 points [95% CI: -6.0, -0.6] for each 10 jig/dL increase in blood Pb level)
and performance IQ (-5.2 points [95% CI: -8.1, -2.3]). A 10 |ig/dL increase in lifetime
average blood Pb concentration was associated with a 2.6 point (95% CI: 0.2, 5.0) decline in
performance IQ.
At 15 to 17 years of age, the Cincinnati subjects were administered a comprehensive
neuropsychological battery (Ris et al., 2004). Variables derived from the Cincinnati
neuropsychological battery were subjected to a principal components factor analysis that yielded
five factors, including a learning/IQ factor that had high loadings for the Vocabulary and Block
Design subtests from the WISC-III as well as the Reading, Spelling, and Arithmetic subscales of
the Wide Range Achievement Test-Revised (WRAT-R). Prenatal, Average Childhood, and
78 month blood Pb levels were used in a series of multiple regression analyses. After covariate-
adjustment, there was a trend towards significance for higher blood Pb concentrations in later
childhood (e.g., 78 months) to be associated with lower learning/IQ factor scores, but this was
largely observed in subjects from the lower end of the SES scale in the sample. This finding is
consistent with previous reports that children in the lower social strata may be more vulnerable
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100
OJ
g 95
03
o
t
CD
Q- 90
.2
10-15 >15-20 >20
Average Lifetime Blood Lead (ng/dL)
Figure 6-1.
Unadjusted and adjusted relationships between average lifetime blood lead
concentrations and Wechsler Scale performance IQ. Mean + SD lifetime
average blood lead concentrations within each category were as follows:
0-10 ug/dL, 7.7 ± 1.4 ug/dL (n = 68); >10-15 ug/dL, 12.3 ± 1.4 ug/dL (n = 89);
>15-20 ug/dL, 17.1 + 1.2 ug/dL (n = 53); and >20 ug/dL, 26.3 + 5.0 ug/dL
(n =
Source: Dietrich et al. (1993a).

to general effects on cognitive development and learning (Bellinger, 2000; Winneke and

Kraemer, 1984).

6.2.3.1.3 Cleveland Study

Early results of the Cleveland prospective study also were reviewed in the 1986

Addendum and 1990 Supplement. By selection, about half of the mothers had histories of

alcohol abuse as measured by the Michigan Alcoholism Screening Test. The other women were

matched controls. Through repeated interviews during pregnancy, detailed information
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regarding maternal alcohol use, smoking, and the use of marijuana and other illicit drugs was
obtained. While alcohol use was correlated with maternal blood Pb levels and smoking was
correlated with both cord and maternal blood Pb levels, no correlations were observed between
use of marijuana or other illicit drugs and any blood Pb marker. The initial cohort included
389 infants with a mean cord blood Pb level of 5.8 |ig/dL (maximum 14.7). In addition to size,
minor morphological anomalies, and 1- and 5-minute Apgar performances, infants were
evaluated on the Brazelton Neonatal Behavioral Assessment Scale (NBAS) and part of the
Graham-Rosenblith Behavioral Examination for Newborns (G-R). Of the 17 neonatal outcomes
examined, the neurological soft signs assessed by G-R were associated with cord blood Pb levels
ranging from 3 to 15 |ig/dL following covariate adjustment (Ernhart et al., 1986). A follow-up
study observed a significant effect for the neurological soft signs measure on Bayley MDI scores
at 12 months (Wolf et al., 1985).
In 285 children from the original cohort, maternal and cord blood Pb levels, as well as
postnatal blood Pb levels at 6 months, 2 years, and 3 years were examined in relation to Bayley
MDI, Psychomotor Index (PDI), and Kent Infant Development Scale (KID) at 6 months, MDI at
1 year and 2 years, and Standford-Binet IQ (S-B IQ) at 3 years of age (Ernhart et al., 1987,
1988). The increment in variance of the IQ that can be attributed to the blood Pb level was
presented as the effect estimate. Most blood Pb indices (maternal, cord, and postnatal up to
3 years) were negatively correlated with the various neurodevelopmental outcomes. However,
only maternal blood Pb level at delivery (mean 6.5 |ig/dL [maximum 11.8]) was found to
contribute to the variance of MDI, PDI, and KID scores at 6 months after adjustment for various
covariates, including HOME score, maternal IQ, parent education, race, medical problems,
maternal alcohol use in pregnancy, Michigan Alcoholism Screening Test score, maternal use of
marijuana, and several categories of psychosocial trauma scale.
Language development also was assessed in the Cleveland cohort at 1, 2, and 3 years of
age. Once again, correlations between blood Pb measures and speech/language outcomes were
generally negative, but none of the relationships remained significant after covariate adjustment
(Ernhart and Greene, 1990).
The relationship between blood Pb levels at 2 years of age with the MDI at 2 years, S-B
IQ at 3 years, and Wechsler Preschool and Primary Scale of Intelligence (WPPSI) test at 4 years
and 10 months of age was further examined (Greene et al., 1992). The effect estimates ranged
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from -11.2 to -14.6 point declines (95% CI not provided) in IQ scores for the three tests with an
increase in 2-year blood Pb from 10 to 25 |ig/dL. After adjusting for the various covariates,
effect estimates decreased to -0.36 to -1.79 point declines for the same incremental change in
blood Pb levels.
The associations between dentine Pb and IQ scores were also examined in this cohort
(Greene and Ernhart, 1993). In 164 children, shed deciduous incisors were collected between
ages 5 and 7 years. Circumpulpal dentine Pb levels were found to be significantly associated
with full scale, verbal and performance IQ, assessed using the WPPSI test, at 4 years and
10 months, after adjustment for various covariates except for HOME score. After additional
adjusting for HOME score, the effect estimates for all three IQ measures diminished, but
remained statistically significant for verbal IQ (p = 0.01) and marginally significant for full scale
IQ (p = 0.06). An increase in dentine Pb from the 10th percentile to the 90th percentile level
(13.5 |ig/g to 129.4 |ig/g) was associated with a 6.0 point (95% CI: 1.4, 10.6) decrease in verbal
IQ and a 4.5 point (95% CI: -0.2, 9.2) decrease in full scale IQ. Sensitivity analyses indicated
that the estimated Pb effect was smaller in magnitude when measurement error was ignored.
These findings using dentine Pb provide stronger evidence of inverse associations between Pb
exposure and IQ scores compared to the previous analyses of this cohort, which indicated that
blood Pb levels were generally not associated with cognitive outcomes after covariate
adjustment.

6.2.3.1.4 Port Pine, Australia Study
Preliminary results from the Port Pirie, Australia study also were described in the 1986
Addendum (Vimpani et al., 1985). Lower Bayley MDI scores at 2 years from 592 children were
significantly associated with higher integrated postnatal blood Pb levels (-20% of the sample
had blood Pb levels >30 |ig/dL at the time of assessment), but not with maternal prenatal,
delivery, or cord blood Pb levels. Results of this interim analysis were interpreted with caution,
since important covariates such as maternal IQ and HOME scores were not available for the
entire cohort at the time of the analyses.
The Port Pirie cohort study had reported results out to 4 years when the 1990 Supplement
was released (McMichael et al., 1988). Following adjustment for covariates, Pb concentrations
at most postnatal sampling points as well as an integrated average for the 4-year postnatal period
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were significantly and inversely associated with scores on the McCarthy Scales of Children's
Abilities. The GCI scores declined by -4.5 points (95% CI: 0.2, 8.8) for a doubling in blood
Pb levels. Similar deficits occurred in the perceptual-performance and memory scores.
The integrated postnatal blood Pb levels among the 537 children in this cohort were among the
highest of the prospective studies (geometric mean 19 jig/dL). However, further analyses
indicated that the effects observed did not depend on children with the more extreme levels of
exposure. The concentration-response relationship between blood Pb and GCI was stronger
among children with blood Pb levels <25 |ig/dL than it was overall.
Of all of the prospective studies of Pb and child development, the Port Pirie cohort study
was probably among the best positioned to reliably detect effects of low level Pb exposure into
later childhood owing to its wide range of exposure, large sample size, and lack of extremes in
terms of sample social advantage or disadvantage. The WISC-RIQ test was administered to
494 children between 7 and 8 years of age (Baghurst et al., 1992). IQ scores were examined in
relation to In-transformed blood Pb concentration. Following adjustment for covariates, there
was little association with pre- and perinatal Pb exposure assessments. However, significant
decrements in full scale and verbal IQ were found to be associated with postnatal blood Pb
levels. The estimated effect size was a loss of 3.3 points (95% CI: 0.2, 6.5) in full scale IQ and
4.0 points (95% CI: 0.7, 7.2) in verbal IQ in association with a doubling of the integrated
postnatal blood Pb concentration up to three years. In light of the Cincinnati findings, it is of
interest that the Block Design subtest of the WISC-R (a measure of visual-spatial abilities),
exhibited the strongest association with Pb exposure. Port Pirie investigators also collected
deciduous central upper incisors from 262 children in their cohort (McMichael et al., 1994).
After covariate adjustment, a significant inverse association was observed between tooth Pb
concentration and WISC-R full scale IQ at 7 years of age. The adjusted estimated decline in full
scale IQ across the tooth Pb range from 3 to 22 jig/g (range for 90% of population) was
5.1 points (90% CI: 0.2, 10.0). Once again, the Block Design subtest was among the most
highly sensitive.
Port Pirie children were assessed again at 11 to 13 years of age to examine the persistence
of relationships between environmental Pb exposure and impacts on intelligence (Tong et al.,
1996). At that age, Port Pirie investigators were able to recall 375 children for IQ assessments.
At 11 to 13 years of age, the geometric mean lifetime average blood Pb concentration was
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14.1 |ig/dL. WISC-R scores were significantly and inversely associated with integrated lifetime
average blood Pb concentrations out to 11 to 13 years. Later blood Pb concentrations after
3 years of age were more predictive of lower IQ. Mean full scale IQ declined by 3.0 points
(95% CI: 0.1, 5.9) for a doubling of lifetime average blood Pb concentrations. The authors
could find no clear evidence of a threshold level in their data.

6.2.3.1.5 Sydney, Australia Study
Unlike Port Pirie, the reports on the Sydney cohort study were consistently negative with
respect to the effects of Pb exposure on neurodevelopment (Cooney et al., 1989a,b; McBride
et al., 1989). In the 298 mothers and infants sampled, geometric mean blood Pb levels at
delivery were 9.1 |ig/dL and 8.1 |ig/dL, respectively, with less than 2% in excess of 15 |ig/dL.
Mean postnatal blood Pb levels peaked at 16.4 |ig/dL when children reached 18 months and then
declined to 10.1 |ig/dL at 48 months. No significant, inverse relationships were reported
between prenatal or postnatal blood Pb concentrations and neurodevelopmental assessments
conducted from 6 months through 4 years of age. The McCarthy Scales of Children's Abilities
was administered to 207 children at 4 years of age, but no associations with blood Pb levels were
observed prior to or following covariate-adjustment. As in the case of the Cleveland study, the
authors noted that the HOME score was a strong contributor to the neurodevelopmental
assessments at all ages. As stated in the 1990 Supplement, this raises the questions of whether
Pb exposure might have covaried with HOME scores. If so, adjusting for HOME scores would
reduce the statistical power by which to detect postnatal blood Pb effects on the neurocognitive
measures. It is also noteworthy that the interpretation of the Sydney findings has been
complicated by concerns about possible contamination of capillary blood Pb samples collected
during the early phases of the investigation (Cooney et al., 1989b).
The Sydney prospective study further assessed 175 subjects that remained in the study at
7 years of age (Cooney et al., 1991). Geometric mean blood Pb concentrations peaked at 2 years
of age (15.2 jig/dL). The geometric mean blood Pb level at 7 years of age was 7.7 |ig/dL.
The WISC-R and other neurobehavioral assessments were administered. The adjusted
correlations between postnatal blood Pb levels and WISC-R scores were consistently negative
but nonsignificant at the p < 0.05 level. The r value (units = SD of IQ per SD of blood Pb) for
the correlation between full scale IQ and concurrent blood Pb at age 7 years was -0.06 (95% CI:
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-0.20, 0.09). Sufficient data are not presented in this study to convert the correlation
coefficients to a slope estimate. The correlation coefficient is not significantly different from
that found by Bellinger et al. (1992) for 57-month-old children (-0.07 [95% CI: -0.23, 0.08]),
or by Lanphear et al. (2005) for children aged 4.8 to 10 years (-0.20 [95% CI: -0.28, -0.12]).
All correlation coefficients are for full scale IQ and concurrent blood Pb levels.
Results from this follow-up study were consistent with the investigators earlier reports of
no association between blood Pb levels <15 |ig/dL and developmental deficits among the Sydney
cohort children. However, the authors noted that their study was not designed to examine small
deficits associated with blood Pb levels at this magnitude. They reported that the size of their
cohort did not provide sufficient power to detect effects less than 5%. Cooney et al. concluded
that results from their study indicate that if developmental deficits do occur at blood Pb levels
below 25 |ig/dL, the effect size is likely to be less than 5%.

6.2.3.1.6 Mexico City Study (A)
Preliminary results of the Mexico City cohort prospective study (Rothenberg et al., 1989)
were presented in the 1990 Supplement. Blood Pb levels from 42 mother-infant pairs were
measured at 36 weeks of pregnancy (mean 15.0 |ig/dL) and delivery (mean 15.4 |ig/dL), and
in the cord blood (mean 13.8 |ig/dL). The Brazelton NBAS was administered to infants at
48 hours, 15 days, and 30 days after birth. None of the Pb measures were associated with the
NBAS outcomes; however, several differential Pb measures (i.e., maternal blood Pb at 36 weeks
of pregnancy minus cord blood Pb) were found to be associated with several outcome variables.
Increases in the blood Pb of the mother during the last month of pregnancy or a cord blood Pb
level higher than the mother's blood Pb level were associated with adverse changes in
Regulation of States, Autonomic Regulation, and Gestation Age.
Schnaas et al. (2000) further examined the effect of postnatal blood Pb level on cognitive
development in 112 children with complete data from the Mexico City study. Lead was
measured in blood every 6 months from 6 to 54 months. Intellectual status was assessed with the
McCarthy GCI. The purpose of the study was to estimate the magnitude of the effect of
postnatal blood Pb level on the GCI and to determine how the effect varies with the time
between blood Pb measurements and the neurocognitive assessments. The geometric mean
blood Pb level between 24 to 36 months was 9.7 |ig/dL (range 3.0 to 42.7). A number of
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significant interactions were observed between blood Pb levels and age of assessment.
The greatest effect was found at 48 months, with a decrease of 4.0 points (95% CI not presented)
in adjusted GCI score being observed for a doubling of the 24 to 36 month blood Pb level.
The authors concluded that 4 to 5 years of age (when children are entering school) appears to be
a critical period for the manifestation of earlier postnatal blood Pb level effects.
In another study by Schnaas et al. (2006), the effect of prenatal Pb exposure on child
development was examined. Of the 321 infant comprising the original cohort, 150 with
complete data were included in the analyses. They used general linear models with random
intercepts and slopes to analyze the pattern of Pb effects on full scale WISC-RIQ evaluated from
6 to 10 years of age. The geometric mean blood Pb levels during pregnancy, ages 1 to 5 years,
and ages 6 to 10 years were 8.0 |ig/dL (range 1-33), 9.8 |ig/dL (range 2.8-36.4), and 6.2 |ig/dL
(range 2.2-18.6), respectively. The effect of log-transformed blood Pb levels from various time
points on full scale IQ was examined. Only third trimester prenatal blood Pb levels were found
to be significantly associated with IQ at age 6 through 10 years, after adjusting for potential
confounders. A 3.44 point (95% CI: 1.28, 5.61) deficit in full scale IQ was observed for each
natural log increment in blood Pb. In their discussion, however, the authors note that given the
modest sample size and relatively low power of this study, they do not claim that Pb exposure
from other developmental periods has no effect on child IQ.

6.2.3.1.7 Kosovo, Yugoslavia Study
The neurodevelopment results of a large birth cohort study of 577 children in two towns
in Kosovo Province, Yugoslavia, were not available at the time of the 1990 Supplement.
The study took place in Titova Mitrovica, near the site of a longstanding Pb smelter, refinery,
and battery plant, and in Pristina, a less exposed community 25 miles to the south. A unique
characteristic of this cohort was the high prevalence of anemia secondary to iron deficiency
(34% with hemoglobin concentrations <10.5 |ig/dL at 2 years of age). The investigators began
providing iron-fortified multivitamin supplements to the entire cohort when the children were
between 18 to 38 months of age (Wasserman et al., 1994).
Like Port Pirie, this was one of the more highly exposed cohorts. Blood Pb levels were
obtained during the second trimester, from the umbilical cord at delivery, and postnatally at
6-month intervals to 90 months. At birth, geometric mean cord blood Pb levels were nearly
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21 |ig/dL in the smelter area (Wasserman et al., 1992). At age 2 years, geometric mean blood Pb
concentrations were 35.5 |ig/dL and 8.4 |ig/dL among infants from Titova Mitrovica and
Pristina, respectively.
Neurocognitive measures of mental abilities were administered at 2, 4, 7, and 10 to
13 years of age. Relationships between these neurocognitive outcomes and log-transformed
blood Pb levels were assessed. A doubling of blood Pb levels at 2 years of age was associated
with a covariate-adjusted decline of 1.6 points (95% CI: 0.2, 3.0) in Bayley MDI. Statistically
nonsignificant decrements in MDI were associated with blood Pb levels measured at all other
time points. Iron deficiency anemia also was an independent predictor of lower MDI
(Wasserman et al., 1992). When examined at 4 years of age, the geometric mean blood Pb
concentration of children from the smelter area was 39.9 |ig/dL, whereas the geometric mean for
children in the "unexposed" area was 9.6 |ig/dL (Wasserman et al., 1994). Children were
administered the McCarthy Scales of Children's Abilities. Higher prenatal and cord blood Pb
concentrations were associated with lower GCI scores. Following covariate-adjustment, children
of mothers with prenatal blood Pb levels >20 |ig/dL scored a full standard deviation below
children in the lowest exposure group (<5 |ig/dL prenatal blood Pb). A statistically significant
association also was observed between nearly every blood Pb measurement (at 6-month intervals
since birth) and GCI. At 4 years of age, a doubling of blood Pb levels was associated with a
reduction of 2.8 points (95% CI: 1.4, 4.3) on the GCI. The Perceptual-Performance subscale of
the McCarthy was found to be most sensitive to Pb exposure.
When 301 children were examined at 7 years of age with the WISC-III, significant
associations were observed between postnatal blood Pb concentrations and IQ, with consistently
stronger associations between performance IQ and later blood Pb measures (Factor-Litvak,
1999). The adjusted intellectual loss associated with a doubling in lifetime average blood Pb was
2.7 points (95% CI: 1.7, 3.7) in full scale IQ, 2.8 points (95% CI: 1.7, 4.0) in performance IQ,
and 2.1 points (95% CI: 1.1, 3.2) in verbal IQ. By 7 years, measures of iron status were no
longer significantly associated with IQ.
At age 10 to 12 years, 290 subjects with complete data on exposure and covariate factors
were again assessed with the WISC-III (Wasserman et al., 2003). However, in addition to well-
characterized exposure histories based on serial blood Pb assessments, tibial bone Pb was also
measured using 109Cd based K-shell XRF (Todd et al., 2001) on a representative subsample of
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167 subjects from both communities. Blood Pb and bone Pb measurements were highly
correlated in Titova Mitrovica, but not in Pristina. Following covariate-adjustment, average
lifetime blood Pb level was significantly and negatively related to all components of WISC-III
IQ. A doubling of average blood Pb concentration was associated with a decrease in full scale,
performance, and verbal IQ of 1.6 points (95% CI: 0.4, 2.8), 1.5 points (95% CI: 0.3, 2.8), and
1.5 points (95% CI: 0.3, 2.6), respectively. The relationships between bone Pb and IQ scores
were stronger than those for blood Pb, at least in the more highly exposed smelter community.
For each doubling of tibial bone Pb concentrations, full scale, performance, and verbal IQ
decreased by an estimated 5.5, 6.2, and 4.1 points, respectively. The authors also reported that
significant associations between tibial Pb concentrations and IQ scores persisted despite
inclusion of blood Pb into the model. The inference drawn from these findings was that
associations between bone Pb and IQ outcomes may be stronger than those between blood Pb
measures and IQ.

6.2.3.1.8 Shanghai, China Study
A prospective study of low-level prenatal and postnatal exposure was initiated in 1993 by
Shen et al. (1998) in Shanghai, China. Pregnant women were recruited from a maternal and
child health care facility in the community. Lead levels were determined on 348 cord blood
samples. The geometric mean cord blood Pb level was 9.2 |ig/dL (range 1.6-17.5); 40.8% of the
infants had cord blood Pb levels > 10 |ig/dL. Infants were further selected for study on the basis
of their cord blood Pb concentrations—the low Pb group (n = 64) had levels <30th percentile
while the high Pb group (n = 69) had levels >70th percentile. Mean cord blood Pb
concentrations in the high Pb group and low Pb group were 13.4 |ig/dL (SD 2.0) and 5.3 |ig/dL
(SD 1.4), respectively. At 3, 6, and 12 months, infants were administered the Chinese version of
the BSID. Capillary blood samples were collected at each visit to ascertain levels of postnatal
exposure. Mean blood Pb at 1 year of age was 14.9 |ig/dL (SD 8.7) in the high Pb group and
14.4 |ig/dL (SD 7.7) in the low Pb group. Postnatal blood Pb levels were not significantly
different in the high and low Pb groups.
At all three ages, the Bayley MDI, but not PDI, was associated with cord blood Pb
groupings following adjustment for covariates, which included a wide range of perinatal,
demographic, social, and environmental factors. Postnatal blood Pb concentrations were not
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associated with any Bayley measures. Differences in mean MDI between cord blood Pb groups
were 3.4 points at 3 months (p = 0.02), 6.3 points at 6 months (p = 0.03), and 5.2 points at
12 months (p = 0.03). The early results of this prospective study are generally in accord with
similar investigations in Boston, Cincinnati, and Cleveland. The authors concluded that the
adverse effects of prenatal Pb exposure on early neurobehavioral development are readily
discernible and stable over the first year of life.

6.2.3.1.9 Rochester Study
The Rochester prospective study, initiated in 1994, examined the relationship between
blood Pb levels and IQ at 3 and 5 years of age in 172, predominantly African-American, lower
SES children (Canfield et al., 2003a). Participants were enrolled when children were 5 to
7 months of age in what was originally a study of Pb dust control methods (Lanphear et al.,
1999). Blood Pb concentrations were assessed at 6-month intervals until 2 years and annually
thereafter. No data were available on prenatal exposure. The measure of IQ was the abbreviated
Stanford-Binet Intelligence Scale-4th Edition (SBIS-4). Potential confounders assessed included
gender, birth weight, iron status, HOME scores, maternal IQ, SES, and tobacco use during
pregnancy.
Blood Pb concentrations in the Rochester cohort were quite low for an urban population,
as this study was conducted after public health measures to reduce blood Pb levels in children
were already having a dramatic impact in the U.S. population. Blood Pb levels peaked at 2 years
of age (mean 9.7 |ig/dL). The mean lifetime average blood Pb concentration was 7.7 |ig/dL at
the age of 3 years and 7.4 |ig/dL at the age of 5 years. At 5 years of age, 56% of the children had
a peak blood Pb concentration <10 |ig/dL. Following adjustment for covariates, there were
significant inverse associations with full scale IQ at both 3 and 5 years of age for all blood Pb
variables, including lifetime average up to age of behavioral assessment.
The effect of Pb on IQ was estimated in all children using lifetime average, peak,
concurrent, and average in infancy (6-24 months) blood Pb levels. Lead effects on IQ for the
subgroup of children whose peak Pb concentration never exceeded 10 |ig/dL also was estimated.
The effect estimates were larger in the subsample of children with peak blood Pb concentrations
<10 |ig/dL compared to those for all children. For example, the overall estimate including all
children indicated that an increase in the lifetime average blood Pb concentration of 1 |ig/dL was
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associated with a decrease of 0.46 points (95% CI: 0.15, 0.76) in IQ. In comparison, a 1 |ig/dL
increase in lifetime average Pb concentration was associated with a decline of 1.37 points
(95% CI: 0.17, 2.56) in children with peak blood Pb level <10 |ig/dL. In an accompanying
editorial on the Canfield et al. (2003a) study, Rogan and Ware (2003) noted that the steepness in
the concentration-response relationship below 10 |ig/dL might have been influenced by
10 children with blood Pb concentrations at or below 5 |ig/dL and IQs above 115. However,
they added that it was unlikely that the associations reported by Canfield et al. were solely due to
these values. Regression diagnostics performed by Canfield et al. identified only one potential
outlier (a child who had a low IQ and low Pb concentration); however, this value was retained in
all analyses, as it did not pass the discordancy test.
In the Rochester study, the relationship between children's IQ score and their blood Pb
level was found to be nonlinear. A semiparametric analysis indicated a decline of IQ of
7.4 points for a lifetime average blood Pb concentration of up to 10 |ig/dL, whereas for levels
between 10 to 30 |ig/dL a more gradual decrease of-2.5 IQ points was estimated. The authors
concluded that the most important aspect of their findings was that effects associated with blood
Pb levels <10 |ig/dL observed in previous cross-sectional studies (e.g., Chiodo et al., 2004;
Fulton et al., 1987; Lanphear et al., 2000; see Section 6.2.3.2) were confirmed by this rigorous
prospective longitudinal investigation.

6.2.3.1.10 Mexico City Study (B)
In another prospective cohort study conducted in Mexico City, Gomaa et al. (2002)
examined prenatal and postnatal Pb exposure effects on the neurodevelopment of 197 children
aged 2 years. The study cohort was recruited from 3 maternity hospitals in Mexico City that
served a low- to moderate-income population. Lead was measured in the umbilical cord and
maternal venous blood samples at the time of delivery. Maternal body burden was measured by
obtaining cortical (tibial) and trabecular (patellar) bone Pb measurements using K-shell XRF
within 4 weeks of delivery. At 2 years of age, the Bayley MDI and PDI were administered.
The major objective of this study was to compare Pb levels in umbilical cord blood and maternal
bone as independent predictors of infant mental development. Mean blood Pb concentrations in
the cord blood, at 12 months of age, and at 24 months at age were 6.7 |ig/dL (SD 3.4), 7.2 |ig/dL
(SD 2.8), and 8.4 |ig/dL (SD 4.6), respectively. Mean maternal patella and tibia bone Pb levels
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were 17.8 |ig/g (range 10 |ig/dL, the coefficients of concurrent blood Pb for both the
24-month MDI and PDI were significantly steeper for children whose blood Pb never exceeded
10 |ig/dL (p = 0.01). In children with blood Pb levels <10 |ig/dL, a statistically significant
decrement of 1.04 points (p < 0.01) was observed per 1 |ig/dL increase in 24-month blood Pb
compared to a 0.07 point increase (p = 0.84) in children with blood Pb levels > 10 |ig/dL.
In addition, a steeper inverse slope was observed over the blood Pb range up to 5 |ig/dL
(-1.71 points per 1 |ig/dL increase in blood Pb, p = 0.01) compared to the range between 5 and
10 |ig/dL (-0.94 points, p = 0.12); however, these slopes were not significantly different
(p = 0.34). In conclusion, a major finding of this prospective study was that a significant inverse
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relationship between blood Pb concentration and neurodevelopment was observed among
children whose blood Pb levels did not exceed 10 |ig/dL at any age.

6.2.3.1.11 Pooled-Analyses of Prospective Longitudinal Cohort Studies
Investigators have collectively analyzed the results of multiple independent studies using
the methods of meta- and pooled data analyses. A powerful approach involves pooling the raw
data from several high quality studies to examine concentration-response relationships in a large
sample of children with diverse sociodemographic backgrounds and levels of exposure.
The studies reviewed here are summarized in Annex Table AX6-2.2.
Lanphear et al. (2005) reported on a pooled analysis of seven prospective studies that
were initiated prior to 1995. The analysis involved 1,333 children with complete data on
confounding factors that were essential in the multivariable analyses. The participating sites
included Boston, MA; Cincinnati, OH; Cleveland, OH; Rochester, NY; Mexico City; Port Pirie,
Australia; and Kosovo, Yugoslavia. A prospective cohort study conducted in Sydney, Australia
was not included because the authors were unable to contact the investigators (Cooney et al.,
1989b, 1991). The sample size of 175 for children at age 7 years in the Sydney cohort and the
wide confidence intervals of the effect estimates, as implied by the lack of significant
associations, indicate that the nonavailability of this study was unlikely to have influenced the
results of the pooled analysis by Lanphear et al.
The primary outcome measure was full scale IQ measured at school age (mean age at IQ
testing was 6.9 years). All children were assessed with an age-appropriate version of the
Wechsler scales. Four measures of Pb exposure were examined: concurrent blood Pb (blood Pb
level closest in time to the IQ test), maximum blood Pb level (peak blood Pb measured at any
time prior to the IQ test), average lifetime blood Pb (mean blood Pb from 6 months to the
concurrent blood Pb test), and early childhood blood Pb (defined as the mean blood Pb from 6 to
24 months). A pooled analysis of the relationship between cord blood Pb levels and IQ also was
conducted in the subsample for which cord blood Pb tests were available.
Multivariate regression models were developed adjusting the effect of blood Pb for site as
well as assessing ten common covariates potentially acting as confounders of the relationship
between Pb and cognitive development, including HOME scores, birth weight, maternal
education and IQ, and prenatal substance abuse. A thorough statistical analytic strategy was
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employed to determine the linearity or nonlinearity of the relationship between blood Pb levels
and full-scale IQ. Regression diagnostics also were performed to ascertain whether Pb
coefficients were affected by collinearity or influential observations. The fit of all four measures
of postnatal blood Pb levels was compared using the magnitude of the model R2. The blood Pb
measure with the largest R2 (adjusted for the same covariates) was nominated a priori as the
preferred blood Pb index relating Pb exposure to IQ in subsequent inspections of the
relationships. The primary analysis was done using a fixed-effects model, although a mixed
model treating sites as random effects was also examined. The authors further investigated the
impact of any one site on the overall model by estimating the blood Pb coefficient in seven
identical models, each omitting data from one of the seven cohort studies. Similar models were
fitted for verbal and performance IQ as well.
The median lifetime average blood Pb level was 12.4 |ig/dL (5th-95th percentile 4.1-34.8)
with about 18% of the children having peak blood Pb levels <10 |ig/dL. The 5th to 95th
percentile concurrent blood Pb levels ranged from 2.4 to 30 |ig/dL. The mean IQ of all children
was 93.2 (SD 19.2) but this varied greatly between studies. All four measures of postnatal Pb
exposure were highly correlated. However, the concurrent blood Pb level exhibited the strongest
relationship with IQ, as assessed by R2. Nevertheless, the results of the regression analyses for
all blood Pb measures were very similar. Multivariable analysis resulted in a six-term model
including log of concurrent blood Pb, study site, maternal IQ, HOME Inventory, birth weight,
and maternal education.
Various models, including the linear model, cubic spline function, the log-linear model,
and the piece-wise linear model, were investigated in this analysis. The shape of the dose-
response relationship was determined to be nonlinear; the log-linear model was found to be a
better fit for the data. Using the log-linear models, the authors estimated a decrement of
1.9 points (95% CI: 1.2, 2.6) in full scale IQ for a doubling of concurrent blood Pb. However,
the IQ point decrements associated with an increase in blood Pb from <1 to 10 |ig/dL compared
to 10 to 20 ng/dL were 6.2 points (95% CI: 3.8, 8.6) versus 1.9 points (95% CI: 1.2, 2.6).
As shown in Figure 6-2, the individual effect estimates for the seven studies used in the
pooled analysis also generally indicate steeper slopes in studies with lower blood Pb levels
compared to those with higher blood Pb. The issue of greater effects observed at lower blood Pb
levels will be discussed in detail in Section 6.2.13.
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120
110
g
« 100
RJ
u
CO
^ 90
80
70
Boston
Mexico
Port Pirie
^Cleveland
Cincinnati
Rochester
Yugoslavia
10 20 30 40 50
Concurrent Blood Lead (|jg/dl_)
60
Figure 6-2. Linear models for the 7 cohort studies in the pooled analysis, adjusted for
maternal IQ, HOME score, maternal education, and birth weight. The
range of data shown for each study represents the 5th to 95th percentile of
the concurrent blood lead level at the time of IQ testing.
Source: Lanphear et al. (2005).

Ernhart (2006) expressed the concern that one study site was driving the results and that
the HOME score was not always measured with the IQ test. Other limitations were also
mentioned, such as the use of capillary finger stick for the early blood Pb tests rather than venous
blood Pb samples. Lanphear et al. (2006) noted that though they agree that using an early
measure of the HOME inventory in the Rochester cohort was a potential limitation, excluding
this cohort from the pooled analysis changed the coefficient by <3%. Sensitivity analyses
reported in Lanphear et al. (2005) indicated that no single study was responsible for the
estimated relationship between Pb and deficits in IQ, thus diminishing concerns about unique
attributes or potential limitations for any specific sites.
In summary, the log-linear model in Lanphear et al. estimated a decline of 6.2 points in
full scale IQ for an increase in concurrent blood Pb levels from <1 to 10 |ig/dL. This effect
estimate was comparable to the 7.4 point decrement in IQ for an increase in lifetime mean blood
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Pb levels up to 10 |ig/dL observed in the Rochester study (Canfield et al., 2003a), as well as
other studies reviewed above.

6.2.3.2 Cross-sectional Studies of Neurocognitive Ability
Among the cross-sectional studies reviewed in the 1986 Lead AQCD and the 1990
Supplement, the most thorough and methodologically rigorous were those of Needleman et al.
(1979) and Fulton et al. (1987). Needleman et al. (1979) measured Pb in the dentin of deciduous
teeth in elementary school children from two Boston area communities. After statistical
adjustment for a number of potential confounding factors, children in the higher tooth Pb group
performed significantly less well on full scale and verbal IQ. Differences in full scale IQ
between the high and low tooth Pb groups were on the order of 4.5 points.
The general population study by Fulton et al. (1987) studied 501 children aged 6 to
9 years in Edinburgh, Scotland who were at risk for Pb exposure owing to a plumbosolvent water
supply and a large number of houses with Pb plumbing. Blood Pb levels averaged 11.5 |ig/dL
(range 3 to 34 |ig/dL). Following covariate adjustment, there were statistically significant
relationships between concurrent blood Pb levels and total scores on the British Ability Scale and
the Quantitative and Reading subscales. Data showed a clear concentration-response
relationship, with no evidence of a threshold.
More recent cross-sectional studies of neurocognitive ability are summarized in Annex
Table AX6-2.3, and key studies are discussed in this section. Lanphear et al. (2000) examined
the relationship between blood Pb concentrations and cognitive deficits in a nationally
representative sample of 4,853 children aged 6 to 16 years children who participated in the third
National Health and Nutrition Examination Survey (NHANES III). The purpose of the study
was to examine the relationship between low blood Pb concentrations (especially those
<10 |ig/dL) and two subtests of the WISC-R, Block Design (a measure of visual-spatial skills)
and Digit Span (a measure of short-term and working memory). Academic achievement tests
also were administered but are discussed in a later section (see Section 6.2.4). A number of
potential confounders were assessed and included in multivariable analyses, including gender,
racial/ethnic background, child's serum ferritin level, serum cotinine level, region of country,
marital status and education level of primary caregiver, and a poverty index ratio (the ratio of
total family income, as reported by the adult informant, to the federal poverty level for the year
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of the interview). Other potential confounders, such as in utero and postnatal exposure to
tobacco smoke, birth weight, and admission to the neonatal intensive care unit, were only
available for children between 6 and 11 years of age. Therefore, the authors conducted a
secondary analysis of the data on these children to verify that inclusion of these potentially
important variables did not alter the findings of the main analysis using the larger sample.
The geometric mean blood Pb level for children in the study sample was 1.9 |ig/dL
(SE 0.1). Only 2.1% of the NHANES III sample in this analysis had blood Pb concentrations
> 10 |ig/dL. In multivariate analyses, a significant covariate-adjusted relationship was found
between blood Pb level and scores on both WISC-R subtest for all children as well as among
those children with blood Pb levels <10 |ig/dL. Blood Pb concentration also was significantly
associated with Block Design when the multivariate analysis was restricted to children with
blood Pb levels <7.5 |ig/dL. For a 1 |ig/dL increase in blood Pb level, Block Design scores
declined by 0.10 points (SE 0.04) for all children, 0.13 points (SE 0.06) for children with blood
Pb levels <10 jig/dL, and 0.11 points (SE 0.06) for children with blood Pb levels <7.5 jig/dL.
The authors concluded that deficits in intellectual functioning were associated with blood Pb
levels <10 ng/dL; however, it is not clear whether the cognitive deficits observed were due to Pb
exposure that occurred during early childhood or were a function of concurrent exposure.
Nevertheless, concurrent blood Pb levels likely reflected both ongoing exposure and preexisting
body burden. It should be noted that while a large number of potential confounding factors were
controlled in these analyses, no data on maternal IQ or direct observations of caretaking quality
in the home were available. The study did, however, control for the poverty index ratio and
education level of the primary caregiver which may have served as surrogates for maternal IQ or
the HOME score.
Chiodo et al. (2004) studied the relationship between blood Pb concentrations and IQ,
assessed using WISC-III at 7.5 years of age in a sample of 237 African-American inner-city
children from Detroit, MI. This cohort was derived from a larger study of the effects of prenatal
alcohol exposure on child development. However, -83% of children for whom blood Pb levels
were obtained had either low or no gestational exposure to alcohol. Blood Pb levels were low,
with a mean of 5.4 |ig/dL (SD 3.3, range 1-25). Following adjustment for a wide range of
covariates (including drug and alcohol exposure, HOME scores, SES status, and perinatal health
among others), there was a statistically significant association between blood Pb concentrations
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and full scale, verbal and performance IQ, with the strongest relationship observed for
performance IQ. Significant effects of Pb on full scale and performance IQ were still evident at
blood Pb <7.5 |ig/dL. Nonparametric smoothing analyses confirmed that these effects were
linear in nature.
Kordas et al. (2004, 2006) examined the relationship between Pb exposure and various
indices of psychometric intelligence in a cohort of 602 first grade children attending public
schools in Torreon, a highly industrialized city in northern Mexico. The mean blood Pb
concentration was 11.5 |ig/dL (SD 6.1). Approximately half of the children had blood Pb levels
<10 |ig/dL, and only 20% of the subjects had blood Pb levels >15 |ig/dL. Subjects were
administered Spanish versions of the Peabody Picture Vocabulary Test-Revised (PPVT-R), the
Cognitive Abilities Test, and subtests of the WISC-R (Coding, Digit Span, and Arithmetic
subtests). Letter and Number Sequencing tests (adapted from the Trail Making Test, Trails A)
also were administered. After adjustment for sociodemographic variables, anemia, iron status,
and growth, higher blood Pb levels were significantly associated with poorer performance on the
PPVT, WISC-R Coding, and Number and Letter Sequencing. Segmented linear regressions
revealed steeper slopes at lower blood Pb levels. Significant Pb effects were observed only for
the segments defined by a concurrent blood Pb concentration <10 to 14 |ig/dL. The authors
acknowledged that a major limitation of their study is the lack of earlier measures of Pb exposure
and nutritional status, and information on potentially confounding variables such as parental
intelligence and quality of caretaking in the home.
Walkowiak et al. (1998) conducted a cross-sectional study examining relationships
between low-level Pb and mercury (Hg) exposure and various measures of neurocognitive and
neuromotor functioning in 384 children aged 6 years in three German cities. Blood Pb was
measured at the time of testing and Hg burden was estimated from urine samples. As their
measure of IQ, two subtests of the German WISC, Vocabulary and Block Design were
administered. These subtests were treated separately as well as a summed index, which served
as a surrogate for full scale IQ. Blood Pb concentrations were low (geometric mean 4.3 |ig/dL
[95th percentile 8.9]). Following covariate-adjustment, Vocabulary and the combined index, but
not Block Design, exhibited negative associations with blood Pb of statistical or borderline
statistical significance; but no associations were observed for Hg. The authors concluded that
these findings roughly correspond with those of other studies that find Pb effects exposure on
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measures of intelligence at blood Pb levels <10 |ig/dL. However, they also cautioned that some
important covariates and potential confounding variables were not measured, including parental
IQ and home environment (e.g., HOME score).
The relationship between blood Pb levels and intelligence also was examined in
533 Saudi Arabian school girls aged 6 to 12 years (Al-Saleh et al., 2001). The Beery-Visual-
Motor Integration (Beery-VMI) test and Test of Non-Verbal Intelligence were used as measures
of intelligence. The mean blood Pb level was 8.11 |ig/dL (SD 3.50), with 69% having a blood
Pb level <10 |ig/dL. After adjusting for various factors, including family income and parental
education, a significant inverse relationship was observed for blood Pb and Beery-VMI.
However, unlike various other studies, most notably the seven study pooled analysis by
Lanphear et al. (2005), that observed larger effects at lower blood Pb levels, a significant
association between blood Pb and Beery-VMI was not observed when data were restricted to
those with blood Pb levels <10 |ig/dL.
The cross-sectional studies examining the effect of Pb on neurocognitive abilities varied
widely in study location, population, age of testing, and outcomes measured. These studies
found that blood or tooth Pb levels were significantly associated with declines in intelligence and
other neurocognitive outcomes. In general, these associations were consistently observed in
studies with mean blood Pb levels <10 ng/dL.

6.2.3.3 Meta-analyses of Studies of Neurocognitive Abilities
Several meta-analyses of studies investigating associations between Pb exposure and
neurocognitive abilities included results from both prospective cohort studies and cross-sectional
studies. The studies reviewed here are summarized in Annex Table AX6-2.2. Needleman and
Gatsonis (1990) conducted a meta-analysis of 12 studies that used multiple regression techniques
to assess the relationship between Pb levels in tissues (blood or teeth) while adjusting for
potentially confounding variables. Studies were weighted based on sample sizes, which ranged
from 75 to 724 children. The authors divided studies into two groups according to the type of
tissue analyzed for Pb (blood or teeth). Joint p-values and average effect sizes as measured by
partial correlation coefficients were calculated using two different methods by Fisher and by
Mosteller and Bush (Rosenthal, 1984). The joint p-values for the blood Pb studies were <0.0001
for both methods, whereas joint p-values of <0.0006 and <0.004 were obtained for tooth Pb
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studies. The partial correlations ranged from -0.27 to -0.0003. Sensitivity analyses revealed
that no single study was responsible for the significance of the final findings. The authors
concluded that the hypothesis that Pb lowers children's IQ at relatively low Pb exposure dose
was strongly supported by their quantitative analysis.
Another meta-analysis conducted by Schwartz (1994) took a different approach. Only
studies relating blood Pb to IQ were chosen for quantitative review, since the concentration of Pb
in the bloodstream is the main index of Pb exposure typically used as the basis for public health
policy. Three longitudinal and four cross-sectional studies relating blood Pb to IQ were
examined. Furthermore, while the work of Needleman and Gatsonis (1990) essentially involved
combining partial correlations, the measure of effect used in the Schwartz analysis was the
predicted change in full scale IQ as blood Pb increased from 10 to 20 |ig/dL. For the prospective
longitudinal studies, blood Pb levels at 2 years of age or average blood Pb levels up to 3 years of
age were used in the analysis. This approach by Schwartz may be related to the belief at the time
of the analysis that blood Pb levels during the first 3 years of life were the most critical in
determining the severity of neurodevelopmental toxicity. The exclusion of blood Pb levels from
other time points may be at issue, given that it now appears that later blood Pb levels may be
more predictive of mental deficits (Baghurst et al., 1992; Canfield et al., 2003a; Chen et al.,
2005; Dietrich et al., 1993a; Factor-Litvak et al., 1993). Studies were weighted by the inverse of
the variances using a random-effects modeling procedure. The estimated decrease in IQ for an
increase in blood Pb from 10 to 20 |ig/dL was 2.6 points (95% CI: 1.8, 3.4). Sensitivity
analyses indicated that the results were not determined by any individual study. Effect estimates
were similar for longitudinal and cross-sectional studies. In another analysis, studies with mean
blood Pb concentrations <15 |ig/dL and >15 |ig/dL had estimated effect sizes of -3.23 points
(95% CI: -5.70, -0.76) and -2.32 points (95% CI: -3.10, -1.54), respectively. When the
study with the lowest mean blood Pb level was examined in greater detail using nonparametric
smoothing, no evidence of a threshold was found down to a blood Pb of 1 |ig/dL.
Pocock et al. (1994) conducted a review of the epidemiologic evidence for Pb effects on
IQ that included a meta-analysis. For the meta-analysis, the fixed-effect method described by
Thompson and Pocock (1992) was used. Five prospective and 14 cross-sectional studies (with
both tooth and blood Pb measures) were included. For consistency, only blood Pb levels at or
around 2 years of age were considered for the prospective studies. Their overall conclusion was
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that a doubling of blood Pb levels from 10 to 20 |ig/dL or of tooth Pb from 5 to 10 jig/g was
associated with an average estimated IQ deficit of about 1 to 2 points.
Other earlier meta-analyses of Pb-IQ studies have been published but are not reviewed
here, because later work greatly extended these efforts and included more studies, rendering
those analyses outdated (Needleman and Bellinger, 1988; Schwartz, 1985; Thacker et al., 1992).
The meta-analyses of studies investigating the effect of Pb on neurocognitive ability all
consistently observed significant associations between blood or tooth Pb levels and decrements
in IQ. Also, the Schwartz (1994) analysis found no evidence of a threshold at blood Pb levels
below 10 |ig/dL.

6.2.4 Measures of Academic Achievement
Relatively little data are available on the relationship between Pb exposure and objective
measures of academic achievement. A few earlier studies reported an inverse relationship
between Pb exposure and reading skills (Fergusson et al., 1988a; Fulton et al., 1987; Yule et al.,
1981). Since the 1990 Supplement, more studies have focused on the practical consequences of
childhood Pb exposure by including measures of academic performance in their batteries.
Studies reviewed in this section are summarized in Annex Table AX6-2.4.
Using NHANES III data, Lanphear et al. (2000) examined the relationship between blood
Pb levels and a standardized measure of academic achievement in 4,853 children aged 6 to
16 years. This cohort was previously described in Section 6.2.3.2. Subjects were administered
the Arithmetic and Reading subtests of the Wide Range Achievement Test-Revised (WRAT-R).
The WRAT-R Arithmetic subtest includes oral and written problems ranging in level from
simple addition to calculus, while the Reading subtest assesses letter recognition and word
reading skills. The geometric mean blood Pb concentration was 1.9 |ig/dL. Only 2.1% of the
subjects had blood Pb levels > 10 |ig/dL. Multiple linear regression revealed a 0.70 point (95%
CI: 0.37, 1.03) decrement in arithmetic scores and a 0.99 point (95% CI: 0.62, 1.36) decrement
in Reading scores for each 1 |ig/dL increase in concurrent blood Pb concentration (p < 0.001).
In the next phase of the analysis, the adjusted relationship between performance on WRAT
subtests and blood Pb concentration for children with blood Pb levels <10 |ig/dL, <7.5 |ig/dL,
<5 |ig/dL, or <2.5 |ig/dL were carried out. Statistically significant inverse relationships between
blood Pb levels and performance for both Reading and Arithmetic subtests were found for
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children with blood Pb concentrations <5 |ig/dL. Secondary analysis limited to younger children
with data on all covariates did not alter findings from the main analysis. The authors concluded
that results of these analyses suggest that deficits in academic skills are associated with blood Pb
levels <5 |ig/dL. The potential limitations of this study, including the lack of information on
previous blood Pb levels, maternal IQ, and caretaking quality in the home, were discussed in
Section 6.2.3.2.
Needleman et al. (1990) reexamined the Chelsea and Somerville, MA cohort of first and
second graders recruited in the 1970s (Needleman et al., 1979). Of the original 270 children,
132 were recalled. Relationships between concentration of Pb in shed deciduous teeth and
neurobehavioral deficits had persisted into late adolescence. Subjects with dentin Pb levels
>20 ppm were at higher risk of dropping out of high school (adjusted odds ratio of 7.4, [95% CI:
1.4, 40.7]) and of having a reading disability (adjusted odds ratio of 5.8 [95% CI: 1.7, 19.7]).
Higher dentin Pb levels were also significantly associated with lower class standing, increased
absenteeism, and lower vocabulary and grammatical reasoning scores on the Neurobehavioral
Evaluation System (NES). The authors concluded that undue exposure to Pb had enduring and
important effects on objective parameters of success in real life.
Bellinger et al. (1992) administered a battery of neuropsychological tests to 148 Boston
Lead Study cohort children at age 10 years. The short-form of the Kaufman Test of Educational
Achievement (KTEA) was administered in addition to IQ studies. The KTEA assesses
reading, math, and spelling skills. The primary outcome was the Battery Composite Score.
As previously indicated, exposures in this cohort were low (with a peak mean blood Pb at
18 months of only 7.8 |ig/dL [SD 5.7]), and the cohort consisted of high-SES White children
from intact families with college-educated parents. Average KTEA scores in this cohort were
about one standard deviation above the population mean. Nevertheless, postnatal blood Pb
levels measured at virtually all ages were significantly associated with lower KTEA Battery
Composite Scores. However, after covariate-adjustment, including full scale IQ in the model,
only blood Pb levels at 24 months of age were significantly predictive of lower academic
achievement. Over the range of ~0 to 25 |ig/dL, Battery Composite scores declined by
-8.9 points (95% CI: 4.2, 13.6) for each 10 |ig/dL increase in 24-month blood Pb. The specific
subscales of the KTEA that were most significantly associated with Pb were Spelling and Math.
Within the Math subscale, Pb appeared to be more strongly associated with performance on the
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advanced quantitative Concepts/Applications items than on computation. The associations
between these early measures of low level Pb exposure and achievement were significant even
after adjustment for IQ, suggesting that Pb-sensitive neuropsychological processing and learning
factors not reflected in indices of global intelligence may contribute to reduced performance on
academic tasks.
Leviton et al. (1993) evaluated the relationship between pre- and postnatal Pb exposure
and academic problems in -2,000 children born in one Boston hospital between 1979 and 1980
using the Boston Teacher Questionnaire (BTQ). A teacher provided an assessment of each
child's academic functioning when the child reached the age of 8 years. Mean umbilical cord
blood Pb was 6.8 |ig/dL and mean tooth (dentin) Pb concentration was 2.8 jig/g. There was
limited information on covariate factors. Still, after adjustment for potential confounding
variables, elevated dentin Pb levels were associated with statistically significant reading and
spelling difficulties as assessed by the BTQ among girls. The authors concluded that their
findings supported the case for Pb-associated learning problems at levels prevalent in the general
population. However, they added that the inability to assess child-rearing quality in this
questionnaire study conducted by mail limits the inferences that can be drawn from the findings.
Rabinowitz et al. (1992) examined the relationship between tooth Pb concentrations and
scores on BTQ clusters in 493 Taiwanese children in first through third grade. Mean Pb levels in
incisors were 4.6 jig/g (SD 3.5). Factors associated with Pb and BTQ scores included
13 variables measuring perinatal, familial, and economic parameters. Prior to adjustment for
covariates, girls in this sample with higher exposures to Pb showed a borderline significant trend
for reading difficulties, whereas boys displayed significantly increased difficulties with respect to
activity levels and task attentiveness. In multiple logistic regression models, tooth Pb terms
failed to achieve statistical significance. The authors concluded that Pb levels found in the teeth
of children in their Taiwanese sample were not associated with learning problems or syndromes
as assessed by the BTQ.
Fergusson et al. (1993) examined the relationship between dentin Pb levels in shed
deciduous teeth at 6 to 8 years and measures of academic attainment and classroom performance
in a birth cohort of over 1,200 New Zealand children enrolled in the Christchurch Health and
Development Study when they reached 12 to 13 years of age. This study was an extension of
earlier work in these children indicating a relationship between low Pb levels and deficits in
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academic skills around the age of 8 years (Fergusson et al., 1988a). Average dentine Pb levels in
the cohort were 6.2 jig/g (SD 6.2). Measures of academic performance included word
recognition from the Burt Reading Test, reading comprehension from the Progressive
Achievement Test, a general measure of scholastic skills based on children's scores on the Test
of Scholastic Abilities, and teacher ratings of classroom performance in reading, written
expression, and mathematics. Following adjustment for a wide range of covariates (including
residence in potentially Pb-hazardous housing), dentin Pb levels were significantly associated
with virtually every formal index of academic skills and teacher ratings of classroom
performance. Statistical evaluations included a multivariate analysis of all 12 regression
equations simultaneously using LISREL modeling methods. This conservative analysis clearly
showed that the probability of observing these results under the null hypotheses that Pb was
unrelated to all covariate-adjusted test outcomes was extremely small. In an adjunct analysis,
Fergusson and Horwood (1993) examined low-level Pb exposure effects on the growth of word
recognition in this cohort from 8 to 12 years of age, using growth curve modeling methods.
After adjustment for potential confounding variables, children with dentin Pb levels >8 |ig/g
displayed significantly slower growth in word recognition abilities with no evidence of catch up.
The authors concluded that these results were consistent with their earlier analyses and suggest
that early exposure to very low levels of Pb result in small but detectable and enduring deficits in
children's cognitive abilities.
Academic achievement in relationship to Pb was reexamined in the New Zealand cohort
when subjects reached 18 years of age (Fergusson et al., 1997). The sample at 18 years consisted
of 881 subjects, or -70% of the original cohort. Measures of educational achievement included
the Burt Reading Test, number of years of secondary education, mean number of School
Certificate passes (based on results of national examinations), and leaving school without formal
qualifications (analogous to failure to graduate from high school in the United States). As in
previous analyses, a wide range of potentially confounding sociohereditary factors were
measured and controlled for in multivariable analyses, which included both linear and logistic
regressions. Prior to and following covariate adjustment, there were statistically significant
concentration-response relationships between dentin Pb concentrations and lower reading test
scores, having a reading level of less than 12 years, failing to complete 3 years of high school,
leaving school without qualifications, and mean number of School Certificates subjects passed.
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The authors concluded that their results are consistent with the view that there is a relationship
between early low-level Pb exposure and later educational outcomes. The late results of the New
Zealand studies confirm the Needleman et al. (1990) findings in a cohort with lower levels of
environmental Pb exposure.
Wang et al. (2002a) examined the relationship between blood Pb levels and class ranking
in 934 third graders living in an urban industrial area of Taiwan. The outcome variables were
grades for Chinese (reading and writing), Mathematics, History and Society, and Natural
Science. To avoid the impact of teacher's bias in grading criteria, the authors converted the
children's grades into class rankings. A limited number of potentially confounding factors were
measured, including maternal education and father's SES. Mean blood Pb level was 5.5 |ig/dL
(SD 1.89). In multiple regression analyses adjusting for gender, maternal education, and father's
SES, blood Pb was significantly associated with lower class ranking in all academic subjects.
The major shortcoming of this cross-sectional study is the lack of control for potentially
important confounding factors such as parental intelligence. However, the strength and
consistency of the reported relationships suggest that relatively low level Pb exposure may play
a role in lowering academic performance.
Al-Saleh et al. (2001) studied the association between blood Pb levels and academic
achievement in 533 girls aged 6 to 12 years in Riyadh, Saudi Arabia. At the time of this study
leaded gasoline was still in wide use. The measure of academic achievement was based on the
class ranking of each student as assessed by the teacher. A large number of confounding
variables were considered, including growth parameters and various assessments of SES, health
status, geographical location and family structure. The mean blood Pb in the cohort was
8.11 |ig/dL (SD 3.5). Following covariate adjustment, there was a statistically significant
relationship between higher blood Pb levels and lower class rank percentile subscales. When
multiple regression models were fitted to a subset of students with blood Pb levels <10 |ig/dL,
class rank percentile continued to show a statistically significant association with blood Pb
levels.
Kordas et al. (2006) examined the relationship between blood Pb levels at 7 years of age,
and math achievement and vocabulary in 594 second graders living near a metal foundry in
Torreon, Mexico. The mean blood Pb level was 11.4 |ig/dL (SD 6.1). Following adjustment for
covariates measuring other well-documented predictors of cognitive functioning as well as
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concurrent arsenic exposure, blood Pb concentrations were statistically significantly related to
poorer math and vocabulary scores. Furthermore, in segmented regression analyses, the slopes
for the association of blood Pb with vocabulary and math scores were both significantly steeper
below 10 |ig/dL than above.
The results of these studies strongly suggest that Pb exposure can affect the academic
performance of children. Consistent associations also were observed in cohorts of children with
mean blood Pb levels below 10 |ig/dL.

6.2.5 Measures of Specific Cognitive Abilities
Outcomes of specific cognitive abilities, in particular, the domains of Attention and
Executive Functions, Language, Memory and Learning, and Visuospatial Processing have
been examined in some detail in recent studies. These studies are summarized in Annex
TableAX6-2.5.
In the aggregate, studies suggest that Pb exposure impairs a child's ability to regulate
attention and to engage several related higher-order cognitive processes that have come to be
termed "executive functions." Executive functions refer to strategic planning, control of
impulses, organized search, flexibility of thought and action, and self-monitoring of one's own
behavior—activities that help the subject maintain an appropriate mental set in order to achieve
an immediate or future goal (Spreen et al., 1995). In some earlier studies, increased Pb exposure
was found to be associated with a higher frequency of negative ratings by teachers and/or parents
on behaviors such as inattentiveness, impulsivity, distractibility, and less persistence in assigned
tasks, as well as slow psychomotor responses and more errors on simple, serial, and choice
reaction time tasks (e.g., Hatzakis et al., 1989; Hunter et al., 1985; Needleman et al., 1979; Raab
et al., 1990; Winneke et al., 1990). The concept that Pb may impact executive functions in
particular is biologically plausible. The prefrontal cortex is highly innervated by projections of
neurons from the midbrain and has the highest concentration of dopamine of all cortical areas.
Dopamine plays a key role in cognitive abilities mediated by the prefrontal cortex. It has been
known for some time that the dopamine system is particularly sensitive to Pb, based upon studies
of rodents and nonhuman primates (Cory-Slechta, 1995).
Bellinger et al. (1994a) examined a portion of the original Chelsea and Somerville cohorts
at 19 to 20 years of age. The main neurobehavioral outcomes used were scores on a battery of
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attentional measures assembled by Mirsky (1987). Higher tooth Pb concentrations were
significantly associated with poorer scores on the Focus-Execute and Shift factors of the battery,
leading the authors to conclude that early Pb exposure may be associated with poorer
performance on executive/regulatory functions thought to depend on frontal or prefrontal brain
regions.
Stiles and Bellinger (1993) administered a neuropsychological battery of tests to
10-year-old children in the Boston Lead Study cohort. A large number of assessments were
made and, as the authors acknowledge, the number of significant associations was about equal to
those that would be expected by chance. However, as in previous studies, tasks that assess
attentional behaviors and executive functions tended to be among those for which Pb was a
significant predictor of performance. For example, higher blood Pb concentrations at 2 years
were significantly associated with (a) lower scores on the Freedom from Distractibility factor of
the Wechsler scales and (b) an increase in the percentage of preservative errors on the Wisconsin
Card Sorting Test and the California Verbal Learning Test. At 2 years of age, 90% of the
children had blood Pb levels <13 |ig/dL.
Canfield et al. (2003b) conducted a comprehensive evaluation of effects of low-level Pb
exposure on executive functioning and learning in children from the Rochester Lead Study
cohort at 48 and 54 months of age. The mean blood Pb level at 48 months was 6.49 |ig/dL
(range 1.7-20.8), with 80% of the children having a blood Pb <10 |ig/dL. The authors used the
Shape School Task (Espy, 1997), which requires only knowing simple shape and primary color
names. However, embedded in the tasks are protocols requiring inhibition, attention switching,
and a combination of inhibition and switching mental sets. Following covariate-adjustment,
blood Pb level at 48 months was negatively associated with children's focused attention while
performing the tasks, efficiency at naming colors, and inhibition of automatic responding.
Children with higher blood Pb levels also completed fewer phases of the task and knew fewer
color and shape names.
Canfield et al. (2004) also administered portions of the Cambridge Neuropsychological
Testing Automated Battery (CANTAB) to 174 Rochester cohort children at -66 months of age.
Children were tested with the Working Memory and Planning CANTAB assessment protocols to
assess mnemonic and executive functions. Blood Pb levels ranged from 0 to 20 |ig/dL in this
cohort. Following covariate adjustment, children with higher blood Pb levels showed impaired
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performance on tests of spatial working memory, spatial memory span, cognitive and cognitive
flexibility, and planning as indexed by tests of intradimensional and extradimensional shifts and
an analog of the Tower of London task.
Ris et al. (2004) administered an extensive neuropsychological battery to 15-17 year old
subjects from the Cincinnati Lead Study cohort. Besides executive functions assessed by the
Wisconsin Card Sorting Test and the Rey-Osterrieth Complex Figure, other domains examined
included attention, memory, achievement, verbal abilities, visuoconstructional skills, and
fine-motor coordination. About 30% of the subjects had blood Pb levels >25 |ig/dL during the
first 5 years of life; and 80% had at least one blood Pb > 15 |ig/dL. A factor analysis of scores
selected a priori revealed five factors that included Attention. A strong "executive functions"
factor did not emerge. After covariate-adjustment, the strongest associations between Pb
exposure and performance were found for factor scores derived from the Attention component,
which included high loadings on variables from the Conners Continuous Performance Test.
However, this relationship was restricted to males as indicated by a strong Pb-by-gender
interaction. This observed gender interaction suggests that neuromechanisms sub-serving
attention were affected by Pb in this cohort for boys but not for girls. This is not surprising given
the heightened vulnerability of males for a wide range of developmental perturbations.
For example, a substantial gender difference in the incidence of Attention Deficit/Hyperactivity
Disorder (ADFID) is well established, and one could speculate that early exposure to Pb
exacerbates a latent potential for such problems.
Visual-spatial skills have also been also been explored in some depth by a few studies.
When investigations of Pb-exposed children have used global IQ measures and conducted
subscale analyses, it has been observed that Performance IQ or subtests contributing to the
performance IQ (i.e., Block Design) are frequently among the most strongly associated with
biological indices of Pb exposure (Baghurst et al., 1992; Chiodo et al., 2004; Dietrich et al.,
1993a; McMichael et al., 1988; Wasserman et al., 1994). Dietrich et al. (1991, 1992) have also
observed that integrated measures of Pb exposure over a child's lifetime are most consistently
associated with simultaneous processing abilities, cognitive functions closely associated with
visual-spatial integration skills and right cerebral functioning (Kaufman and Kaufman, 1983).
In addition, studies employing specific measures of visual-motor integration skills, such as the
Developmental Test of Visual Motor Integration (VMI), the Bender Visual-Motor Gestalt Test,
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and others, have found them to be among the most consistently associated with early Pb
exposure (Al-Saleh et al., 2001; Baghurst et al., 1995; Dietrich et al., 1993b; Wasserman et al.,
2000a; Winneke et al., 1990). In a follow-up of Cincinnati Lead Study cohort subjects at age
16 years, Ris et al. (2004) observed a significant association between prenatal maternal blood Pb
levels and deficits in visual-spatial and constructional skills as indexed by Visual-Constructional
factor scores. Variables with high loadings on this factor included scores on the WISC-III Block
Design subtests and selected variables from the Rey Osterrieth Complex Figure.
Kordas et al. (2006) administered an extensive battery of tests assessing specific abilities
to 594 first graders (mean blood Pb of 11.4 |ig/dL ) in Torreon, Mexico, the site of a metal
foundry. The battery included well validated assessments of mental distractibility, sequencing
skills, memory, visual spatial skills and stimulus discrimination. Following adjustment for
covariates in linear regression analyses, blood Pb remained significantly associated with
performance on the Sternberg Memory test. For the various tests, steeper slopes were generally
observed for blood Pb levels below 10 |ig/dL than above.
It is still unclear whether the domains of attention/executive functions or visual-motor
integration per se are specifically sensitive to Pb. This is because there is rarely a one-to-one
correspondence between performance on a focused neuropsychological test and an underlying
neuropsychological process. Thus, for example, a low score on the Berry VMI may reflect
singular or multiple neurobehavioral deficits, including difficulties with graphomotor control,
visual perception, behavioral monitoring (impulsivity), or planning (executive functions).

6.2.6 Disturbances in Behavior, Mood, and Social Conduct
Lead effects on behavior and mood of children has been an area of recent research.
Studies conducted prior to 1990 clearly pointed to behavioral problems as potential sequelae of
lower level Pb toxicity in children. Several early case control studies have linked Pb to
hyperactivity (David et al., 1972, 1976, 1979). Low levels of Pb in blood and/or teeth have been
associated with teacher ratings of hyperactive behavior, aggression, and attention problems
(e.g., Fergusson et al., 1988b; Hatzakis et al., 1985; Silva et al., 1988; Thomson et al., 1989;
Yule et al., 1984). In the seminal study by Needleman et al. (1979), children with higher Pb
concentrations in dentin were more likely to be rated unfavorably by teachers on the dimensions
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of hyperactivity, impulsivity, and frustration tolerance. New studies reviewed in this section are
summarized in Annex Table AX6-2.6.
While there is no compelling evidence that Pb exposure is directly related to ADHD,
elevated blood or tooth Pb levels have been linked to behavioral features of ADHD, including
distractibility, poor organization, lacking persistence in completing tasks, and daydreaming
(Bellinger and Rappaport, 2002). Bellinger et al. (1994b) studied the relationship between early
Pb exposure and problem behaviors in the classroom in a cohort of 1,782 children born at one
hospital in Boston. Umbilical cord blood Pb levels were low (mean 6.8 |ig/dL [SD 3.1]) as were
tooth Pb levels (mean 3.4 jig/g [SD 2.4]). Teachers filled out the Achenbach Child Behavior
Profile (ACBP), which yields both broad and narrow band scales indexing externalizing and
internalizing problems. Cord blood Pb levels were not associated with the prevalence or nature
of behavioral problems reported by teachers. However, tooth Pb level was significantly
associated with ACBP Total Problem Behavior Scores (TPBS). TPBS scores increased by
~2 points for each log unit increase in tooth Pb. Statistically significant tooth Pb-associated
increases in both externalizing and internalizing scores were also noted. Each log unit increase
in tooth Pb was associated with a 1.5 point increase in scores for these broadband scales
assessing under- and overcontrol of behavior. Only weak associations were seen between tooth
Pb concentrations and the tendency to score in the clinically significant range on these scales.
As the authors noted, it was somewhat surprising that Pb exposure was not more strongly related
to externalizing behavior problems than with internalizing behavior problems. This contradicted
several earlier investigations, including one by Sciarillo et al. (1992) (see Annex
Table AX6-2.6). It may be that more attention has been accorded undercontrolled behaviors,
because they are more readily visible and disruptive in settings such as the classroom.
Therefore, internalizing problems may be part of the full spectrum of behaviors in which the
developmental neurotoxicity of Pb is expressed in children. The authors also cautioned that
residual confounding could not be ruled out, because of the lack of covariate information on
parental psychopathology or direct observations of the family environment—a problem not
unique to this particular study. Nevertheless, these findings are in accord with other studies
which suggest that social and emotional dysfunction may be another expression of increased Pb
exposure during the early postnatal period.
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Fergusson et al. (1993) studied relationships between tooth Pb levels and inattention/
restlessness in the large national New Zealand study of over 1,000 children at 12 and 13 years of
age. Mothers and teachers were asked to respond to a series of items derived from the Rutter and
Conners parental and teacher questionnaires. The selected items related to the degree to which
the child was restless, inattentive, easily distracted, and lacking in concentration. At each age, an
index of the subject's propensity to inattentive and restless behavior was obtained by summing
the total reports of attention deficit behaviors made by both teacher and parent respondents.
Following adjustment for a wide range of sociodemographic and other covariate factors, a
statistically significant, concentration-response relationship was observed between tooth Pb
concentrations (range 1 to 12+ |ig/g) and the inattention/restlessness variable. The authors
concluded that their results were consistent with the view that early mildly elevated Pb levels
were associated with small but long-term deficits in attentional behaviors.
As part of the 11-year follow-up of the Dunedin Multidisciplinary Health and
Development Study, a longitudinal study of a birth cohort of children born in Dunedin's only
obstetric hospital, blood Pb levels were measured in 579 children at age 11 years old (Silva et al.,
1988). The study sample was over-representative of higher SES. The mean blood Pb level was
11.1 |ig/dL (SD 4.91), with a range from 4 to 50 |ig/dL (only two children had blood Pb levels
>30 |ig/dL). The correlations between blood Pb levels and WISC IQ variables were negative but
not statistically significant. However, blood Pb levels were found to be significantly associated
with increased behavioral problems as assessed by both parents and teachers, even after
controlling for various factors including SES, other disadvantageous factors, maternal cognitive
ability, and IQ.
Two prospective studies have also examined early Pb exposure relationships to behavioral
problems as assessed by the Achenbach system. Wasserman et al. (1998) studied the
relationship between Pb exposure and behavior in the Yugoslavian prospective study. The study
surveyed 379 children at 3 years of age with the parent report form of the Achenbach CBCL.
Following covariate adjustment, concurrent blood Pb levels were significantly associated with
scores on the Destructive Behaviors CBCL subscale, although the variance accounted for by
Pb was small compared to sociodemographic factors. As blood Pb increased from 10 to
20 |ig/dL, CBCL subscale scores increased by -0.5 points. The authors concluded that while
statistically significant, the contribution of Pb to social behavioral problems in this cohort was
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small compared to the effects of correlated social factors. Burns et al. (1999) examined the
relationship between Pb exposure and children's emotional and behavioral problems at ages
11 to 13 years in the Port Pirie, Australia, cohort study. After adjusting for many confounding
variables, including HOME scores, maternal psychopathology and the child's IQ, regression
models showed that, for an increase in average lifetime blood Pb levels from 10 to 30 |ig/dL, the
externalizing behavior problem score increased by 3.5 points (95% CI: 1.6, 5.4) in boys but only
by 1.8 points (95% CI: -0.1, 11.1) in girls. In contrast, internalizing behavior problems were
predicted to increase by 2.1 points (95% CI: 0.0, 4.2) in girls, but by only 0.8 points (95% CI:
-0.9, 2.4) in boys.
Recently, the potential role of Pb in delinquent and criminal behavior has been addressed
by several investigations. Previous studies linking attention deficits, aggressive and disruptive
behaviors, and poor self-regulation with Pb exposure have raised the prospect that early exposure
may result in an increased likelihood of engaging in antisocial behaviors in later life.
Denno (1990) surveyed 987 Philadelphia African-American youths enrolled in the
Collaborative Perinatal Project. Data were available from birth through 22 years of age.
The analysis initially considered over 100 predictors of violent and chronic delinquent behavior.
Repeat offenders presented consistent features such as low maternal education, prolonged male-
provider unemployment, frequent moves, and higher Pb intoxication (although the level of Pb
intoxication was not indicated in Denno's report). In male subjects, a history of Pb poisoning
was among the most significant predictors of delinquency and adult criminality.
Needleman et al. (1996) examined relationships between Pb exposure and several
measures of behavioral disturbance and delinquent behavior in subjects from the Pittsburgh
Youth Study. The Pittsburgh Youth Study is a prospective study of the developmental course of
delinquency (Loeber et al., 1991). The population consisted of 850 boys who were prescreened
with an instrument that measured serious and potentially indictable behaviors extracted from the
teachers' and parents' CBCL. Subjects who scored above the 30th percentile on the risk score
and an approximately equal number of subjects randomly selected from the remainder of the
distribution formed the sample (n = 503). Body burden of Pb was measured in the tibia by
K-shell XRF. Measures of antisocial behavior were administered at 7 and 11 years of age, and
included the Self Reported Antisocial Behavior scale (SRA), the Self Report of Delinquent
Behavior (SRD), and the parents' and teachers' versions of the CBCL. Outcome data were
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adjusted for a number of covariates, including mother's IQ, SES, childhood medical problems,
and quality of child rearing. Parents of subjects with higher Pb levels in bone reported
significantly more somatic complaints, more delinquent and aggressive behavior, and higher
internalizing and externalizing scores. Teachers reported significant increases in scores on
somatic complaints, anxious/depressed, social problems, attention problems, delinquent
behavior, aggressive behavior, and internalizing and externalizing problems in the higher Pb
subjects. At 11 years, the SRD scores of subjects were also significantly related to bone Pb
levels. More of the high Pb subjects had CBCL scores in the clinical range for the CBCL
subscales assessing attention problems, aggression, and delinquency. Odds ratios for these
outcomes ranged from 1.5 (95% CI: 0.45, 4.9) for parental reports of aggression to 19.5
(95% CI: 8.9, 41.6) for attention problems. The authors concluded that Pb exposure was
associated with an increased risk for antisocial and delinquent behavior.
Dietrich et al. (2001) reported on the relationship between early Pb exposure and juvenile
delinquency in 195 subjects from the Cincinnati Lead Study. As previously noted, this is an
inner-city cohort of urban children exposed to relatively high levels of Pb by virtue of their
residence in older, deteriorated housing units. Relationships were evaluated between prenatal
(maternal) and postnatal exposure to Pb (through serial blood Pb determinations) and measures
of antisocial and delinquent behaviors (self- and parental reports) examined when the subjects
were 16 or 17 years old. Parents were administered a questionnaire developed specifically for
the study, while the subjects were given the SRD. A wide range of candidate covariates and
confounders were examined, but the only ones predicting antisocial or delinquent behavior were
birth weight, HOME scores, SES, and parental IQ. In multiple linear regression analyses,
prenatal exposure was significantly associated with a covariate-adjusted increase in the
frequency of parent-reported delinquent and antisocial acts, whereas prenatal and postnatal Pb
exposure was significantly associated with a covariate-adjusted increase in frequency of self-
reported delinquent and antisocial behaviors, including marijuana use. In order to clarify
concentration-response relationships, blood Pb indices were transformed to categorical variables
and least-square means were calculated from an analysis of covariance procedure. Subjects in
the highest prenatal blood Pb category (>10 |ig/dL) engaged in 2.3 more delinquent acts over the
preceding 12 months than subjects in the lowest category (<5 |ig/dL). Using average childhood
blood Pb levels, subjects in the medium (16-20 |ig/dL) and highest (>20 |ig/dL) category
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engaged in -1.5 more delinquent acts compared to the lowest category (< 10 |ig/dL). Subjects in
the highest 78-month blood Pb category (>15 |ig/dL) engaged in 4.5 more delinquent acts than
subjects in the lowest category (<5 jig/dL). The authors concluded that Pb might play a
measurable role in the epigenesis of behavioral problems in inner-city children independent of
other social and biomedical cofactors assessed in the study.
Needleman et al. (2002) conducted a case-control study that examined Pb levels in bone
of 194 adjudicated delinquents and 146 non-delinquent community control subjects recruited
from high schools in the city of Pittsburgh and environs of Allegheny County, PA. Since many
delinquents are not arrested or adjudicated, care was taken to ensure that unidentified delinquents
did not populate the control group. Potential control subjects were excluded from the analyses if
found to have a Juvenile Court record or an SRD score above the 90th percentile. Tibial bone Pb
was measured by K-shell XRF. Covariates included race, parental education and occupation,
presence of two parental figures in the home, number of children in the home, and neighborhood
crime rate. Logistic regression analyses were used to model the association between bone Pb
concentration and delinquent status. Cases had significantly higher average tibia Pb levels than
controls (11.0 |ig/g [SD 32.7] versus 1.5 |ig/g [SD 32.1]). Stratified analyses showed this for
both White and African-American subjects. Following adjustment for covariates, adjudicated
delinquents were four times more likely to have bone Pb concentration >25 |ig/g than controls
(odds ratio of 4.0 [95% CI: 1.4, 11.1]). The effect of Pb on delinquency was found to be
substantial in this study. Bone Pb level was the second strongest factor in the logistic regression
models, exceeded only by race. In models stratified by race, bone Pb was exceeded as a risk
factor only by single parent status. The authors concluded that elevated body Pb burdens were
associated with elevated risk for adjudicated delinquency.
The extension of Pb effects into delinquent and criminal behavior is significant for both
the individual and society as a whole. Specific biological mechanisms that may underlie Pb
effects on aggression, impulsivity, and poor self-regulation are not clearly understood. Lead
impacts a large number of sites and processes in the brain involved in impulse control (Lidsky
and Schneider, 2003). However, Needleman et al. (2002) proposed another pathway.
In addition to direct impacts on brain development and neuronal function, Pb exposure may
increase risk of delinquency through a separate, indirect route: impaired cognitive abilities and
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academic performance. That is, students who have difficulties in school and fail to achieve
academic goals are more likely to become lawbreakers.

6.2.7 Sensory Acuities
In comparison to cognitive outcomes, there has been relatively less interest in Pb effects
on sensory functions. However, there are clear indications that Pb exposure during the
developmental period has an impact on complex aspects of visual and auditory acuities. Much of
this work has been carried out in animal models (Otto and Fox, 1993). Epidemiologic studies
have typically assessed hearing thresholds and features of auditory processing in Pb-exposed
children. Studies reviewed in this section are summarized in Annex Table AX6-2.7.
Schwartz and Otto (1987) observed significant Pb-associated elevations in pure-tone
hearing thresholds at various frequencies within the range of human speech among over 4,500
4 to 19-year-old subjects in NHANES II. In a later study, this finding was replicated in a sample
of over 3,000 subjects aged 6 to 19 years in the Hispanic Health and Nutrition Examination
Survey (HHANES) (Schwartz and Otto, 1991). An increase in blood Pb from 6 to 18 jig/dL was
associated with a 2 db loss in hearing at all frequencies, and an additional 15% of children had
hearing thresholds that were below the standard at 2,000 Hz. These relationships continued at
blood Pb levels below 10 |ig/dL.
Dietrich et al. (1992) assessed the relationship between scores on a test of central auditory
processing (SCAN) and prenatal/postnatal blood Pb concentrations in 215 children 5 years of age
drawn from the Cincinnati Lead Study. Higher prenatal, neonatal, and postnatal (up to
concurrent) blood Pb concentrations were associated with more incorrect identification of
common monosyllabic words presented under conditions of filtering (muffling). Other variables
associated with impaired central auditory processing included results for pure-tone audiometry
testing, social class, HOME scores, birth weight, gestational age, a measure of obstetrical
complications, and consumption of alcohol during pregnancy. Following adjustment for these
covariates, neonatal and postnatal blood Pb levels remained significantly associated with
impaired performance on the Filtered Word subtest, more prominently in the right ear. In the
right ear, the Filtered Word subtest score decreased by 0.7 points (p < 0.05; 95% CI not
presented) for a 10 |ig/dL increase in lifetime average blood Pb levels.
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Osman et al. (1999) examined the relationship between concurrent blood Pb levels and
hearing loss in 155 children 4 to 14 years of age living in an industrial region of Poland. Blood
Pb levels ranged from 1.9 to 28 |ig/dL (median 7.2 |ig/dL). Hearing thresholds increased
significantly with higher blood Pb levels at all frequencies (500-8,000 Hz). This relationship
remained statistically significant when restricted to children with blood Pb levels <10 |ig/dL.
A limited number of epidemiologic studies provide supportive evidence of a relationship
between Pb exposure and auditory processing decrements. Lead-related deficits in hearing and
auditory processing may be one plausible mechanism by which an increased Pb burden might
impede a child's learning (Bellinger, 1995).

6.2.8 Neuromotor Function
Relatively few studies have focused on neuromotor deficits as an outcome of early Pb
exposure. However, those that have examined motor functions in Pb-exposed children often
report positive findings. Studies reviewed here are summarized in Annex Table AX6-2.8.
In an early study, unsteadiness, clumsiness, and fine-motor dysfunctions were noted in a
group of mildly symptomatic Pb-poisoned children in Boston, with such effects persisting long
after medical treatment (Pueschel et al., 1972). A study of moderately exposed children living in
the vicinity of a longstanding Pb smelter in Greece found that children with blood Pb levels of
35 to 60 |ig/dL had significantly lower scores on both the Gross and Fine Motor Composite
scores from the Oseretsky scales when compared to control subjects (Benetou-Marantidou et al.,
1988).
Only two modern prospective studies of Pb have assessed motor development in a
comprehensive manner. Dietrich et al. (1993b) investigated possible Pb exposure associations
with motor developmental status in 245 children 6 years of age in the Cincinnati Lead Study
cohort. Following covariate adjustment, they found that postnatal Pb exposure was significantly
associated with poorer scores on measures of bilateral coordination, visual-motor control, upper-
limb speed and dexterity, and the fine motor composite from the Bruininks-Oseretsky scales.
Neonatal, but not prenatal, blood Pb concentrations also were significantly associated with
poorer scores on upper-limb speed and dexterity and the fine motor composite. The strongest
and most consistent relationships were observed with concurrent blood Pb levels (mean
10.1 |ig/dL [SD 5.6]). A 10 |ig/dL increase in concurrent blood Pb levels was associated with
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a 4.6 point (95% CI: 2.1, 7.1) decline in the fine motor composite score. In the same Cincinnati
cohort, postnatal Pb exposure was associated with greater postural instability as assessed by a
microprocessor-based strain gauge platform system (Bhattacharya et al., 1995). When assessed
at 16 years of age, 78-month postnatal blood Pb levels were significantly associated with poorer
fine-motor skills as indexed by covariate-adjusted factor scores derived from a factor analysis of
a comprehensive neuropsychological battery (Ris et al., 2004). The variables loading highly on
the fine-motor component came from the grooved pegboard and finger tapping tasks.
Some results of the Cincinnati Lead Study were replicated by Wasserman et al. (2000a)
in the Yugoslavian Prospective Study. The Bruininks-Oseretsky Test of Motor Proficiency was
adapted for use in their population residing in two towns in Kosovo Province. The main measure
of exposure was the log of the lifetime average blood Pb concentration through 54 months of
age. After covariate-adjustment, average childhood blood Pb levels were associated with poorer
fine motor and visual motor function, but were unrelated to gross motor function.
A recent study by Despres et al. (2005) of multiple exposures including Pb, Hg, and
poly chlorinated biphenyls (PCB's) found that only blood Pb levels measured at the time of
assessment were associated with neuromotor functions in 110 preschool Inuit children residing in
Canada. The mean blood Pb level was 5.0 |ig/dL (range 0.8-27.1 |ig/dL). Blood Pb levels were
significantly associated with increased reaction time, sway oscillations, alternating arm
movements, and action tremor. Only 10% of the children had blood Pb levels >10 |ig/dL. After
eliminating these children from the analyses, results remained significant for reaction time, sway
oscillations, and alternating arm movements. These findings indicated that neuromotor effects of
Pb occurred at blood Pb concentrations <10 |ig/dL.

6.2.9 Brain Anatomical Development and Activity
Electrophysiological evaluations have been conducted on Pb-exposed children in order to
obtain a more direct measure of the toxicant's impact on the nervous system. Studies reviewed
in this section are summarized in Annex Table AX6-2.9. Much of this work was conducted by
Otto and colleagues during the 1980s (e.g., Otto et al., 1985). These studies have demonstrated
Pb effects on neurosensory functioning (auditory and visual evoked potentials) within a broad
range of exposures (Otto and Fox, 1993).
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Rothenberg et al. (1994) reported that higher maternal blood Pb levels at 20 weeks of
pregnancy were associated with increased I-V and III-V interpeak intervals in the brainstem
auditory evoked response recorded in 1-month-old infants. Mean maternal blood Pb level at
20 weeks in this subsample from the Mexico City Prospective Study was only 7.7 |ig/dL, with a
range of 1 to 30.5 |ig/dL. Rothenberg et al. (2000) repeated these measurements with a larger
group of 5- to 7-year-old children (n = 133). In contrast to their previous findings, prenatal
blood Pb levels at 20 weeks were associated with decreased interpeak intervals. However, after
fitting a nonlinear model to their data, they observed that I-V and III-V interpeak intervals
decreased as blood Pb rose from 1 to 8 |ig/dL and increased as blood Pb rose from 8 to 30 |ig/dL.
The biphasic effect was only observed with maternal blood Pb levels at 20 weeks of pregnancy.
Increasing postnatal blood Pb at 12 and 48 months was related to decreased conduction intervals
for I-V and III-V interpeak intervals across the entire blood Pb range.
Magnetic Resonance Imaging (MRI) and Magnetic Resonance Spectroscopy (MRS) have
recently been applied in studies of Pb-exposed children. Trope et al. (1998) were the first to
apply MRI and MRS in an evaluation of a Pb-exposed subject (see Annex Table AX6-2.9 for a
description of the case study). Trope et al. (2001) performed identical MRI and MRS studies on
a sample of 16 subjects with a history of elevated blood Pb levels (23 to 65 |ig/dL) before five
years of age. Average age at time of evaluation was 8 years. These subjects were compared to
age-matched controls composed of siblings or cousins who had blood Pb levels that never
exceeded 10 |ig/dL. Although all of the participants had normal MRI examinations, the Pb-
exposed subjects exhibited a significant reduction in N-acetylaspartate: creatine and
phosphocreatine ratios in frontal gray matter compared to controls.
Meng et al. (2005) performed MRI and MRS studies on children with blood Pb levels
>27 |ig/dL (n = 6) and age- and gender-matched controls with blood Pb levels <10 |ig/dL
(n = 6). The average age at time of evaluation was approximately 11 years. Subjects came from
the Anhui province in China. Lead-exposed children had an average blood Pb concentration of
37.7 |ig/dL (SD 5.7), while controls averaged 5.4 |ig/dL (SD 1.5). MRS was used to measure
N-acetylaspartate, choline-containing compounds, and total creatine in the frontal lobes and
hippocampus in cases and controls. All children presented with normal MRI with no evidence
of structural abnormalities. However, peak values of N-acetylaspartate, choline, and creatine in
all four brain regions were reduced in Pb-exposed children relative to controls. The authors
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concluded that the reduced brain N-acetylaspartate levels they observed in cases may be related
to decreased neuronal density or neuronal loss. Furthermore, reduced choline signal may
indicate decreased cell membrane turnover or myelin alterations that can lead to central nervous
system hypertrophy, while lower creatine may indicate reduced neuronal cell viability.
Using functional MRI (fMRI), the influence of childhood Pb exposure on language
function was examined in a subsample of 48 young adults from the Cincinnati Lead Study
(Cecil et al. 2005; Yuan et al., 2006). At age 20-23 years, subjects performed an integrated verb
generation/finger tapping paradigm. Higher childhood average blood Pb levels were
significantly associated with reduced activation in Broca's area, a recognized region of speech
production in the left hemisphere. This association remained statistically significant after
adjustment for the subject's latest IQ assessment. Higher childhood blood Pb levels were also
associated with increased activation in the right temporal lobe, the homologue of Wernicke's
area (an area associated with speech production) in the left hemisphere. These results suggest
that elevated childhood Pb exposure strongly influences neural substrates of semantic language
function in normal language areas, with concomitant recruitment of contra-lateral regions
resulting in a striking, dose-dependent atypical organization of language function.

6.2.10 Gene-Environment Interactions in the Expression of Lead-associated
Neurodevelopmental Deficits
The discussion of gene-environment interactions with respect to Pb exposure
encompasses differential susceptibilities with respect to race, gender, and genetic polymorphisms
associated with Pb biodistribution, and neurotransmitter metabolisms and function. While the
differential effects of Pb on neurodevelopment have been studied to some extent in regard to race
and gender, little work has been reported with respect to specific genetic polymorphisms.
In the United States, African-American children are at increased risk for elevated blood
Pb levels compared to White children. For example, in the last two NHANES surveys,
African-American children were found to have significantly higher blood Pb levels than Whites,
even after adjusting for urban residential status and family income (Brody et al., 1994; Mahaffey
et al., 1982). However, reliable differences with respect to Pb effects on neurodevelopmental
morbidity as a function of race have not been reported with consistency.
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Most surveys find that boys have higher blood Pb levels than girls. The data are less clear
with respect to gender-related differences in Pb-associated neurodevelopmental morbidities.
At various assessments from birth to adolescence, a greater male vulnerability has been noted in
the Cincinnati Lead Study (e.g., Dietrich et al., 1987b; Ris et al., 2004). Data from a cross-
sectional study in England showed that the Pb-IQ deficit association was more pronounced in
boys at 6 years of age (Pocock et al., 1987). However, in a study of 764 children in Taiwan, it
was found that the relationship between Pb exposure and IQ scores was substantially stronger in
girls (Rabinowitz et al., 1991). In the Port Pirie cohort study, Pb effects on cognition were
significantly stronger in girls at ages 2, 4, 7, and 11-13 years (Baghurst et al., 1992; McMichael
et al., 1992; Tong et al., 2000).
At least two genetic polymorphisms have been identified that appear to influence the
absorption, retention and toxicokinetics of Pb in humans (Onalaja and Claudio, 2000).
The 8-aminolevulinic acid dehydratase (ALAD) gene has been the most studied but, as yet, the
consequences of the different alleles for susceptibility to the neurodevelopmental consequences
of Pb exposure are unclear. Individuals with the ALAD 1-2 or ALAD 2-2 polymorphism tend to
have higher blood Pb levels than those with ALAD 1-1. ALAD2 could increase vulnerability by
raising blood Pb levels or decrease it by maintaining Pb in a sequestered state in the bloodstream.
Only one pediatric study has examined this directly. Bellinger et al. (1994a) found that subjects
with the ALAD2 polymorphism tended to have lower dentin Pb levels than those with ALAD1.
This is consistent with the concept that increased affinity of the ALAD2 polymorphism inhibits
entry of Pb from the blood stream into other tissues. After adjustment for exposure level,
Bellinger et al. (1994a) found that adolescents with the ALAD2 polymorphism performed better
in the areas of attention and executive functioning assessed in their study when compared to
subjects with the ALAD1 polymorphism. However, because there were only 5 subjects with the
ALAD2 form, meaningful statistical comparisons could not be made.
The other gene that has been studied is the vitamin D receptor or VDR gene. This gene is
involved in calcium absorption through the gut. Research on Pb workers has shown that variant
VDR alleles modify Pb concentrations in bone and the rate of resorption and excretion of Pb
over time (Schwartz et al., 2000a). Haynes et al. (2003) examined the relationship between the
VDR Fokl polymorphism and blood Pb concentrations in 275 children enrolled in the Rochester
Longitudinal Study. It was hypothesized that children homozygous for the F allele—a marker
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for increased calcium absorption—would have higher blood Pb concentrations than
heterozygotes and children homozygous for the f allele, after adjusting for environmental sources
of Pb (floor dust Pb). A statistically significant interaction was found between floor dust Pb
loading and VDR Fokl genotypes on blood Pb concentration, with the FF genotypes having the
highest adjusted mean blood Pb concentrations at 2 years of age. Consistent with other reports,
Haynes et al. (2003) also found that African-American children were significantly more likely to
have the VDR FF genotype than were non-African American children. The ability of
African-American children to have increased calcium absorption may partially explain the higher
blood Pb concentrations observed in African-American children. Unfortunately, there have been
no studies to indicate which, if any, of the VDR polymorphisms are associated with increased
vulnerability to the neurodevelopmental toxicity of Pb.

6.2.11 Persistence of Lead-Related Neurodevelopmental Deficits Associated
with Prenatal and Postnatal Exposure
Neurodevelopmental effects of Pb have been shown to persist into later childhood and
adolescence when environmental conditions have not markedly changed (i.e., no reduction in Pb
exposure). The ramifications of Pb effects on neurodevelopment depend not only on the extent
of the initially observable effects in early childhood, but also on their enduring consequences for
cognition, attainment, and behavior over the lifetime of the individual. Recent studies examining
the persistence of Pb-related neurodevelopmental deficits are summarized in Annex
Table AX6-2.10. Key studies are further discussed in this section.
Since 1990, several studies attempted to eliminate or at least reduce Pb-associated
neurodevelopmental damage through nutritional and/or pharmacological interventions.
Optimism that such interventions might be effective was raised by a New York study published
in the early 1990s. Ruff et al. (1993) observed that among children 13 to 87 months old (with
blood Pb levels of 25-55 |ig/dL) who were given chelation with EDTA and therapeutic iron,
those with the greatest decline in blood Pb levels had improved cognitive test scores,
independent of whether they had been given iron or chelation therapy.
The Treatment of Lead-Exposed Children (TLC) study was originally designed to test the
hypothesis that children with moderate blood Pb levels who were given an oral chelating drug
(dimercaptosuccinic acid or "succimer") would have better scores than children given placebo on
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a wide range of tests measuring cognition, neuropsychological functions, and behavior at
36 months of follow-up (Rogan et al., 2001). The TLC study enrolled 780 children from four
clinical sites into a randomized, placebo-controlled, double-blind trial of up to three 26-day
courses of treatment with succimer. Most children lived in deteriorating inner-city housing, with
77% of the subjects being African-American. Succimer was effective in lowering the blood Pb
levels of subjects on active drug during the first 6 months of the trial. However, after 1 year,
differences in the blood Pb levels of succimer and placebo groups had virtually disappeared.
All data analyses were conducted on an intent-to-treat basis. At 36 months of follow-up, the
mean IQ score on the WPPSI-R of children given active drug was 1 point lower than that of
children administered placebo, and children given succimer evinced more behavioral problems
as rated by the primary caregiver on the Conners Parent Rating Scale. Children given succimer
scored marginally better on the Developmental Neuropsychological Assessment (NEPSY), a
battery of tests designed to measure neuropsychological deficits that can interfere with learning.
However, all of these differences were statistically nonsignificant.
Although results for the first wave of follow-up for TLC were consistently negative for
drug effects on cognition and behavior, they were not necessarily conclusive. Lead may affect
higher-level neurocognitive processes that are inaccessible, difficult to assess, or absent in the
preschool age child. In older children, scores on psychometric measures are more precise and
reliable, a wider and more differentiated range of abilities can be examined, and early academic
performance and social functioning outside the home environment can be evaluated. Therefore,
TLC followed the cohort into the first years of elementary education to determine whether these
later emerging neurodevelopmental functions were spared the effects of Pb in treated children
compared to placebo controls (Dietrich et al., 2004). While remaining within the limits of
hypothesis driven inference, a comprehensive battery of tests were administered to TLC subjects
at 7 and 7.5 years of age. These included assessments of cognition, learning, memory, global
intellectual attainment, attention/executive functions, psychiatric status, behavioral and academic
conduct, neurological functioning, and motor speed. However, treatment with succimer resulted
in no benefit in cognitive, behavioral, neurological, and neuromotor endpoints. Indeed, children
treated with succimer fared worse than children in the placebo group in several areas, including
linear growth, hospitalized and outpatient injury events in the first 3 years of follow-up, and
neuropsychological deficits as assessed by the Attention and Executive Functions core domain
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score from the NEPSY. The authors concluded that these latest follow-up data confirmed their
earlier finding of the TLC regimen of chelation therapy not being associated with
neurodevelopmental benefits in children with blood Pb levels of 20 to 44 |ig/dL. These results
furthermore, emphasize the importance of undertaking environmental control measures so as to
prevent Pb exposure in light of the apparent irreversibility of Pb-associated neurodevelopmental
deficits.
Liu et al. (2002) used the TLC succimer trial data set (Rogan et al., 2001) to examine the
question of reversibility. As reviewed above, intent-to-treat analyses revealed no benefits of
chelation on neurodevelopmental indices beyond 6 months of treatment. Thus, the scores on the
cognitive tests from the two treatment groups could be analyzed either within the treatment
groups or as a whole. Data from 741 children were available for analyses. Mean blood Pb levels
in TLC subjects were 26.2 |ig/dL at baseline, 20.2 |ig/dL at the 6-month follow up, and
12.2 |ig/dL at the 36-month follow-up. Mean declines in blood Pb levels were 6.0 |ig/dL from
baseline to 6-month follow-up, 14.1 |ig/dL from baseline to 36-month follow-up, and 8.0 |ig/dL
from 6- to 36-month follow-ups. Blood Pb levels declined more quickly in the first 6 months in
the succimer group than in the placebo group, but the mean blood Pb levels were very similar at
baseline and at the 36-month follow-up. Prior to examining changes in blood Pb levels in
relationship to changes in cognitive test scores, it was verified that baseline and later blood Pb
levels were indeed significantly associated with deficits on measures administered at specific
points in the study after adjustment for sociohereditary factors surveyed in the study, including
maternal IQ. Unlike in the New York study by Ruff et al. (1993), Liu et al. (2002) found no
overall effect of changing blood Pb level on changes in cognitive test score from baseline to
6 months. However, during the follow-up from baseline to 36 months and from 6 to 36 months,
falling blood Pb levels were significantly associated with increased cognitive test scores, but
only because of an association in the placebo group. Cognitive test scores increased by 2 points
overall and 4 points in the placebo group when blood Pb levels declined by 10 |ig/dL from
baseline to 36 months. There is a possibility that the succimer drug regimen blunted the
beneficial effect. Due to the inconsistency in the results, the data do not provide strong
supportive evidence that Pb-induced cognitive impairments are reversible.
In addition to pharmacological interventions, a few studies have attempted to remediate or
prevent Pb-associated neurodevelopmental deficits through nutritional supplementation. Again,
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recent studies attempting to reduce Pb absorption through mineral hypersupplementation have
been disappointing (Sargent et al., 1999). However, to date, there has been only one controlled
clinical trial involving Pb-exposed children where central nervous system outcomes have been
the focus of study. Kordas et al. (2005) and Rico et al. (2006) conducted a double-blind
nutritional supplementation trial among 602 first grade children in the city of Torreon in northern
Mexico. The city is located near a metal foundry that has been a source of Pb contamination in
the community. The average blood Pb concentration at baseline was 11.5 |ig/dL (SD 6.1) and
about half of the children had blood Pb levels >10 |ig/dL. Subjects received 30 mg ferrous
fumarate, 30 mg zinc oxide, both, or placebo daily for 6 months. In their first report, the
principal outcome assessment used at baseline and at follow-up was the parent and teacher forms
of the Conners Rating Scales, with no consistently significant treatment effects being found and
the authors concluding that this regimen of supplementation did not result in improvements in
ratings of behavior in Pb-exposed children over 6 months. In addition to behavior, the authors
also assessed cognitive functioning with 11 tests of memory, attention, visual-spatial abilities,
and learning. Again, no consistent or lasting differences in cognitive performance were found
among treatment groups, confirming the earlier conclusion that nutritional supplementation alone
is not effective in eliminating or reducing the impact of early Pb exposure on functional
neurodevelopment.
Children's blood Pb levels generally decline after they peak somewhere around 2 years of
age. However, the degree of decline is a function of a number of factors, including previously
acquired Pb body burden and sources of continuing exposure. Some observational studies have
examined the extent to which the rate of decline in blood Pb levels may be associated with
improvements in neurocognitive status. Tong et al. (1998) assessed the reversibility of Pb
cognitive effects in early childhood in the Port Pirie, Australia cohort study. A total of
375 children were followed to the age of 11 to 13 years. Average blood Pb levels decreased
from 21.2 |ig/dL at 2 years to 7.9 |ig/dL at 11-13 years. However, scores on standardized
measures of intellectual attainment administered at 2, 4, 7, and 11-13 years of age in children
whose blood Pb levels declined the most were not significantly improved over those obtained by
children with a more shallow decline in Pb body burden.
Collectively, these studies indicate that nutritional and pharmacological interventions,
without an effort to reduce environmental exposure to Pb, are relatively ineffective in improving
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neurological function in children. Primary prevention and preventing additional increases in
blood Pb levels among children whose blood Pb levels are high remain the most effective means
of dealing with Pb toxicity.

6.2.12 Periods of Enhanced Developmental Susceptibility to Central Nervous
System Effects of Environmental Lead
It has been difficult to identify discrete periods of development when the fetus or child is
particularly susceptible to Pb effects on neurodevelopment. When the prospective studies of Pb
and child development were underway, it was hoped that this methodological approach would be
revealing. However, these studies observed that age strongly predicted the period of peak
exposure (around 18-27 months when there is maximum hand-to-mouth activity), making it
difficult to distinguish whether greater neurotoxic effects resulted from increased Pb exposure or
enhanced susceptibility at a particular age. Furthermore, children with the highest blood Pb
levels tended to maintain their rank order relative to their lower exposed peers throughout these
studies (e.g., Dietrich et al., 1993a; McMichael et al., 1988), limiting the degree to which
investigators could identify any particular period of development as being critical.
From the perspective of human neurodevelopmental biology, one could argue that the first
3 years of life should represent a particularly vulnerable period. The maximal period of Pb
ingestion coincides with the same period of time when major events are occurring in the
development of the central nervous system, including some neurogenesis, rapid dendritic and
axonal outgrowth, synaptogenesis, synaptic pruning, and programmed apoptosis (see Figure 6-3).
The belief that the first 3 years of life represents a critical window of vulnerability is
evident in the Pb literature (Chen et al., 2005). Two major meta-analyses of the relationships
between childhood Pb exposure and IQ focused primarily on the strength of the association
between IQ at school age and blood Pb concentrations at 2 years of age or average blood Pb
levels up to 3 years of age (Pocock et al, 1994; Schwartz, 1994). Neither meta-analysis
considered the importance of concurrent blood Pb associations in older children. The focus on
these particular age groups implied that the interpretation most consistent with the overall results
was that peak blood Pb concentration, achieved somewhere between 1 and 3 years of age, was
most likely responsible for cognitive effects observed years later. These meta-analyses were
highly influenced by findings from the Boston prospective study where blood Pb levels at
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ii.i
<
o
l i i.
1 i

Newborn
1 month
6 months
2 years
Figure 6-3. Golgi-stained section of human cerebral cortex taken from equivalent areas of
the anterior portion of the middle frontal gyrus at different ages. Although
the packing density of cortical neurons does not appear to change, there is
a tremendous increase in the complexity of dendritic arborizations with
increasing age with maximal density occurring between 2 and 3 years of age.
Source: Nolle (1993).
2 years of age have been exclusively and consistently associated with lower IQ and academic
achievement (Bellinger et al., 1992).
This particular interpretation of the Pb literature has also influenced screening programs
(which focus on 1 and 2 year olds), clinical trials that recruit children during the first 3 years of
life, and current interpretation of the cross-sectional literature. For example, the report by
Lanphear et al. (2000) that school-age children enrolled in the NHANES III survey displayed a
significant inverse relationship between concurrent blood Pb concentrations and measures of IQ
and academic achievement at blood Pb concentrations <10 |ig/dL has been interpreted by some
to reflect the effects of higher blood Pb concentrations in children when they were between 1 and
3 years of age.
Recent epidemiologic studies have found other blood Pb indices, including concurrent
blood Pb levels or lifetime averages, to be stronger predictors of Pb-associated IQ effects than
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peak blood Pb concentration. Prospective studies of children with both high and low Pb
exposures found concurrent or lifetime average blood Pb levels to be more strongly associated
with school age IQ and other measures of neurodevelopment (Canfield et al., 2003a; Dietrich
et al., 1993a,b; long et al., 1996; Wasserman et al., 2000b). Lanphear et al. (2005) examined the
relationship of IQ to four blood Pb indices — early childhood, peak, lifetime average, and
concurrent blood Pb levels. All four blood measures were highly correlated, with correlation
coefficients ranging from 0.74 to 0.96. The results of the regression analyses for the different
blood Pb indices were very similar, but the concurrent blood Pb variable exhibited the strongest
relationship with IQ as measured by R2.
One recent study has attempted to directly address the question regarding periods of
enhanced susceptibility to Pb effects. Chen et al. (2005) sought: to clarify the strength of the
association between IQ and blood Pb at various time points, to examine whether cross-sectional
associations observed in school age children 84-90 months of age represent residual effects from
2 years of age or "new" effects emerging among these children, and to evaluate how change in
blood Pb over time is related to IQ at later ages. Chen et al. (2005) used data on 780 children
from the previously-described TLC multicenter clinical trial (Dietrich et al., 2004; Rogan et al.,
2001) to examine these relationships. Homogeneity between the two treatment groups was
verified. There were no statistical differences between succimer and placebo groups in either
blood Pb concentrations or cognitive scores at the time points under consideration. At baseline,
children were given the Bayley Scales of Infant Development. The children's full scale IQ at the
36-month follow-up was measured with the WPPSI-R. At the 60 month follow-up, IQ was
assessed with the WISC-III. All neurodevelopmental outcomes were adjusted for clinical center,
race, gender, language, parent's education, parent's employment, single parent family, age at
blood Pb concentration, and caregiver's IQ. Figure 6-4 displays the mean IQ at current and
subsequent ages by quartiles of blood Pb measured at 2, 5, and 7 years of age. The concurrent
blood Pb concentration always had the strongest association with IQ. As the children aged, the
relationship grew stronger. The peak blood Pb level from baseline to 7 years of age was not
associated with IQ at 7 years of age. Also, in models including both prior and concurrent blood
Pb concentrations, concurrent blood Pb was always more predictive of IQ. Adjustment for prior
IQ did not fundamentally change the strength of the association with concurrent blood Pb
concentration. Chen et al. (2005) found a stronger relationship between IQ at 7 years of age and
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95 -i
90 -
£

80 -
75
I
10
I
15
\
20
5 10 15 20 25 30
Blood Lead Concentration ((jg/dL)
35
~
40
-O- MDI at baseline by baseline blood lead concentration
-Q- IQ at age 5 by baseline blood lead concentration
-A" IQ at age 7 by baseline blood lead concentration
-X- IQ at age 5 by blood lead concentration at age 5
-#- IQ at age 7 by blood lead concentration at age 5
-O~ IQ at age 7 by blood lead concentration at age 7
Figure 6-4. Full scale IQ test scores by previous or concurrent blood lead concentration.
Each data point shows the mean IQ test scores of children measured at
baseline or at two follow-up times, grouped by quartiles of blood lead
concentration. The abscissa of each point is the middle value of each
blood lead concentration category.
Source: Chen et al. (2005).
blood Pb concentration at 7 years compared with blood Pb at 2 years of age. A similar
relationship was observed between IQ and blood Pb at 5 years of age. The strength of the cross-
sectional associations increase over time, despite lower blood Pb levels in older children. These
results support the idea that Pb exposure continues to be toxic to children as they reach school
age, and do not lend support to the interpretation that all the damage is done by the time the child
reaches 2 to 3 years of age. These findings also imply that cross-sectional associations seen in
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children, such as the study recently conducted by Lanphear et al. (2000) using data from
NHANES III, should not be dismissed. Chen et al. (2005) concluded that if concurrent blood Pb
remains important until school age for optimum cognitive development, and if 6 and 7 year olds
are as or more sensitive to Pb effects than 2 year olds, then the difficulties in preventing Pb
exposure are magnified but the potential benefits of prevention are greater.

6.2.13 Effect of Environmental Lead Exposure on Neurodevelopment at the
Lower Concentration Range
Over the past three decades, epidemiologic studies of Pb and child development have
demonstrated inverse associations between blood Pb concentrations and children's IQ and other
outcomes at successively lower Pb exposure levels. The 1986 Addendum and 1990 Supplement
concluded that neurobehavioral effects were related to blood Pb levels of 10 to 15 |ig/dL and
possibly lower. In response to these data, agencies such as the U.S. CDC and the WHO have
repeatedly lowered the definition of an elevated blood Pb concentration, which now stands at
10 |ig/dL (CDC, 1991; WHO, 1995). At the time when these policies were put in place, there
were too few studies of children with blood Pb levels consistently below 10 |ig/dL on which to
base an opinion as to effects at lower exposure levels. Since the removal of Pb from gasoline,
the median blood Pb concentration has dropped dramatically in U.S. children, permitting more
studies of this nature to be done in recent years. Also, the use of meta- and pooled analytic
strategies has enabled investigators to get a clearer picture of effects below 10 |ig/dL.
A recent review by the CDC of the epidemiologic literature on Pb effects on children's
health concluded that the overall weight of available evidence supported an inverse association
between blood Pb levels <10 |ig/dL and the cognitive function of children (CDC, 2005).
The CDC review further noted that a steeper slope in the dose-response curve appeared to be
seen at lower compared to higher blood Pb levels. However, the CDC review also recognized
several important limitations in the available evidence, most notably the small number of directly
relevant cohort studies (i.e., studies that specifically examined the effect of blood Pb levels
<10 |ig/dL) and the inherent limitations of cross-sectional studies (i.e., the lack of data regarding
blood Pb levels earlier in life). Since the CDC review, new publications have been added to the
database evaluating the effect of blood Pb levels <10 |ig/dL, which to some extent help address
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the potential limitation of the number of relevant cohort studies. These are discussed in the
following text.
Table 6-1 presents studies that examined the relationship between IQ and blood Pb level
in children with blood Pb concentrations <10 |ig/dL. The first group includes studies where all
or the majority of the subjects in the study had blood Pb levels <10 |ig/dL. The second group
includes studies where an analysis has been done on the subset of children whose blood Pb levels
were <10 |ig/dL. In both groups, the average slope from the 10th percentile blood Pb level to
10 |ig/dL was estimated. One additional study, Chiodo et al. (2004), examined the association
between IQ and blood Pb levels in children with blood Pb levels <10 ng/dL. However, a slope
estimate of change in full scale IQ per 1 ng/dL change in blood Pb could not be calculated from
the standardized regression coefficients presented in the publication.
The Rochester Prospective Study (n = 172) by Canfield et al. (2003a) is illustrative. This
study extended the relationship between deficits in IQ and blood Pb concentrations to levels well
below 10 ng/dL. Over half of the children in this study did not have a recorded blood Pb
concentration above 10 ng/dL. In covariate-adjusted linear models, each 1 ng/dL increase in
concurrent blood Pb levels was associated with a -1.8 point decline in IQ at 5 years of age using
only data from children with peak blood Pb levels <10 ng/dL compared to a -0.6 point decline in
IQ when data from all children were included. Nonlinear semiparametric smoothing revealed a
covariate-adjusted decline of more than 7 points up to 10 ng/dL of childhood average blood Pb,
whereas a more gradual decline of-2.5 points was associated with an increase in blood Pb from
10 to 20 ng/dL. In response to the Rochester findings, Bellinger and Needleman (2003)
reanalyzed data from the Boston Prospective Study focusing on children whose blood Pb levels
never exceeded 10 |ig/dL (n = 48). In their analyses, 10 year IQ was inversely related to blood
Pb levels at 24 months following adjustment for covariates. Nonparametric smoothing analyses
indicated that the inverse association persisted at blood Pb levels below 5 ng/dL.
Other recent studies demonstrating effects below 10 ng/dL include a prospective study
conducted in Mexico City by Tellez-Rojo et al. (2006). In a cohort of 294 children with blood
Pb levels never exceeding 10 |ig/dL, a statistically significant relationship was observed between
blood Pb concentrations and MDI assessed concurrently at 24 months of age. Furthermore, a
stronger effect of Pb on MDI was observed among infants with blood Pb levels <5 |ig/dL.
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ON
ON
Table 6-1. Summary of Studies with Quantitative Relationships of
IQ and Blood Lead for Blood Lead Levels Less than 10 ug/dL
Reference
Study Location
n
Blood Lead
Age of IQ Testing Measurement Model Used
Slope and 95% CI
(IQ points/jig/dL) for
Blood Lead <10 fig/dLa
Studies of Populations with Blood Leads <10 jig/dL
Al-Saleh etal. (200 l)b
Tellez-Rojo et al. (2006)

Bellinger et al. (1992;
reanalyzed Bellinger
and Needleman, 2003)
Canfield et al. (2003a)
Kordas et al. (2006)
Lanphear et al. (2005)d
Riyadh, Saudi Arabia
Mexico City, Mexico
Studies with

Boston, Massachusetts

Rochester, New York
Torreon, Mexico
International Pooled
Analysis
533
294
Analyses

101
293
244
6-12 years
24 months
Restricted to Subjects with

10 years

5 years
6-8 years
4 years 10 months
to 10 years
Concurrent
Concurrent
Blood Leads

24 months °

Concurrent
Concurrent
Concurrent
Log-linear
Linear
Log-linear
<10 ug/dL

Linear

Linear
Linear
Linear
Log-linear
-0.8 (-1.4,
-1.0 (-1.8,
-0.9 (-1.4,

-1.6 (-2.9,

-1.8 (-3.0,
-0.4 (-1.2,
-0.8 (-1.7,
-0.4 (-0.6,
-0.2)
-0.3)
-0.5)

-0.2)

-0.6)
0.4)
0.1)
-0.3)
a The slopes for blood lead levels <10 ug/dL were estimated from the 10th percentile to 10 ug/dL.
b In Al-Saleh et al. (2001), 69% (n = 368) of the children had blood lead levels <10 ug/dL. The estimated slope is based on the model for the entire sample
population.
0 The original analyses by Bellinger et al. (1992) included slope estimates for concurrent as well as other blood lead measurements, including 24-month blood
lead levels. Bellinger and Needleman (2003) presented reanalyses restricted to subjects with blood lead levels <10 ug/dL only for 24-month blood lead levels.
d The pooled analysis by Lanphear et al. (2005) included data from seven individual studies, including Bellinger et al. (1992) and Canfield et al. (2003a).
-------
The most compelling evidence for effects at blood Pb levels <10 |ig/dL comes from an
international pooled analysis of seven prospective cohort studies (n = 1,333) by Lanphear et al.
(2005) described earlier (Section 6.2.3.1.11). Although exposures in some cohorts were high, by
pooling data from these studies a substantial number (n = 244) of children with blood Pb levels
that never exceeded 10 |ig/dL were included in the analyses.
The slope for Pb effects on IQ was steeper at lower blood Pb levels as indicated by the
cubic spline function, the log-linear model, and the piece-wise linear model. Initially, the
authors attempted to fit a linear model, but the shape of the dose-response relationship was
determined to be nonlinear insofar as the quadratic and cubic terms for concurrent blood Pb were
statistically significant (p < 0.001, p = 0.003, respectively). As illustrated in Figure 6-5, the
shape of the spline function indicated that the steepest declines in IQ were at blood Pb
concentrations below 10 |ig/dL. Additional support for the notion of steeper slopes at lower
blood Pb levels was given in Figure 6-2 (see Section 6.2.3.1.11), which presented the individual
effect estimates for the seven studies used in the pooled analysis. The studies with the lowest
mean blood Pb concentrations had a steeper slope compared with the studies with higher mean
blood Pb concentrations.
The cubic spline is mainly used for descriptive purposes, but it can also be used to test for
departure from linearity and from other nested models. It is not advisable to use the cubic spline
as a concentration-response curve in a risk assessment because it is a complex model deliberately
overfitting the data for the sole purpose of producing a smoothed representation of the trends in
the data. The spline results agreed quite closely with the parametric log-linear model. Because
the restrictive cubic spline indicated that a log-linear model provided a good fit of the data, the
log of concurrent blood Pb was used in all subsequent analyses of the pooled data. Using a
log-linear model, the authors estimated a decrement of 1.9 points (95% CI: 1.2, 2.6) in full scale
IQ for a doubling of concurrent blood Pb. However, the IQ point decrements associated with an
increase in blood Pb from <1 to 10 |ig/dL compared to an increase from 10 to 20 |ig/dL were
6.2 points (95% CI: 3.8, 8.6) versus 1.9 points (95% CI: 1.2, 2.6), respectively. Figure 6-6
illustrates the log-linear model and adjusted mean IQ for the intervals <5, 5-10, 10-15, 15-20,
and >20 |ig/dL. The vertical separation of the confidence limits on the log-linear regression line
appear to be slightly wider at higher blood Pb levels. However, this separation increases
symmetrically in both directions for log of blood Pb levels that are further away from the
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g

CO
"5
105
100
95
90
85
Log-linear Model
5 Knot Spline
95% Cl
5 10 15 20 25 30 35 40
Concurrent Blood Lead (fjg/dL)
Figure 6-5. Restricted cubic splines and log-linear model for concurrent blood lead
concentration. The dotted lines are the 95% confidence intervals for the
restricted cubic splines.
Source: Lanphear et al. (2005).
geometric mean. The widening does not seem appreciable within the range of the observed
blood Pb levels. The 5th and 95th percentile concurrent blood Pb values were 2.4 |ig/dL and
33.1 |ig/dL, respectively. The mean IQs for each of the specified intervals are within the
confidence limits for the log linear model. These results support the conclusion that the observed
steeper slopes at lower blood Pb levels in the pooled analysis were not due to the use of a
log-linear model.
To further investigate whether the Pb-associated decrement was greater at lower blood Pb
concentrations, the investigators divided the data at two cutpoints a priori, a maximal blood Pb of
7.5 and 10 |ig/dL. Separate linear models were then fit to the data above and below the cutpoints
and the concurrent blood Pb coefficients were compared (see Figure 6-7 for the analysis using
the cutpoint of 10 |ig/dL). The coefficient for the 103 children with maximal blood Pb levels
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105
100
S 95
W
90
85
10 20 30
Concurrent Blood Lead (ug/dL)
40
Figure 6-6. Log-linear model (95% CI shaded) for concurrent blood lead concentration
adjusted for HOME score, maternal education, maternal IQ, and birth
weight. The mean IQ (95% CI) for the intervals <5, 5-10,10-15,15-20, and
>20 ug/dL are shown.

Source: Lanphear et al. (2005).
105
100
to
_
95
90
85
Log-linear Model
Peak Blood Lead <10 |jg/dL
Peak Blood Lead 210 pg/dL
10 20 30
Concurrent Blood Lead (pg/dL)
40
Figure 6-7. Log-linear model for concurrent blood lead concentration along with linear
models for concurrent blood lead levels among children with peak blood lead
levels above and below 10 ug/dL.

Source: Lanphear et al. (2005).
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<7.5 |ig/dL was significantly greater (p = 0.015) than the coefficient for the 1,230 children with a
maximal blood Pb >7.5 ng/dL (-2.94 points [95% CI: -5.16, -0.71] per 1 ng/dL increase in
blood Pb versus -0.16 points [95% CI: -0.24, -0.08]). The coefficient for the 244 children who
had a maximal blood Pb <10 |ig/dL also was greater than that for the 1,089 children who had a
maximal blood Pb >10 |ig/dL (-0.80 points [95% CI: -1.74, 0.14] versus -0.13 points [95% CI:
-0.23, -0.03]), although the difference was not statistically significant (p = 0.10). Thus, while
the pooled analysis used a log-linear model to quantify the Pb-associated decrements, the
nonlinear relationship observed in the analysis was clearly not due to the influence of the
log-linear model itself.
Rothenberg and Rothenberg (2005) reanalyzed the Lanphear et al. (2005) pooled study to
examine the form of the concentration-response function for the Pb exposure effect on child IQ.
This further analysis also focused on concurrent blood Pb levels. Rothenberg and Rothenberg
reported that a log-linear relationship between blood Pb and IQ was a significantly better fit
within the ranges of the blood Pb levels than was a linear relationship (p = 0.009), with little
evidence of residual confounding from included model variables. Once again, this is consistent
with a steeper slope at lower compared to higher levels of blood Pb.
For the entire pooled data set, the observed decline of 6.2 points in IQ for an increase in
blood Pb from 1 to 10 |ig/dL was comparable to the decrements for an increase in lifetime mean
blood Pb levels from <1 to 10 |ig/dL observed in the Rochester Longitudinal Study (Canfield et
al., 2003a). The pooled analysis of Lanphear et al. also demonstrated that deficits in IQ extended
to blood Pb levels <7.5 |ig/dL. Therefore, recent evidence indicates that Pb is associated with
neurocognitive deficits at blood Pb levels below 10 |ig/dL in children. The data also suggest that
there may be Pb effects associated with blood Pb <5 |ig/dL, but the evidence is less definitive.
A common observation among some of these low blood-Pb level studies is that of
nonlinear dose-response relationships between neurodevelopmental outcomes and blood Pb
concentrations. At first this may seem at odds with certain fundamental toxicological concepts.
However, there are a number of examples of nonlinear or supralinear dose-effect relationships in
toxicology. It is conceivable that the initial neurodevelopmental lesions at lower Pb levels may
be disrupting very different biological mechanisms than the more severe effects of high
exposures that result in symptomatic poisoning or frank mental retardation (Dietrich et al., 2001).
As Kordas et al. (2006) states, this might help explain why, within the range of exposures not
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producing overt clinical effects, an increase in blood Pb beyond a certain concentration might
cause less additional impairment in children's cognitive functions. It should be noted that the
observation of nonlinear or supralinear dose-effect relationships between blood Pb and
neurodevelopmental outcomes does not preclude the presence of a threshold, as blood or bone Pb
levels currently measured in populations without obvious Pb poisoning are still orders of
magnitude higher than those of pre-industrial humans (Patterson et al., 1991).

6.2.14 Selection and Validity of Neuropsychological Outcomes in Children
Considerable material has been written about methodologies for neurobehavioral
evaluation in studies of environmental chemicals and child development (Bellinger, 2002, 2003;
Dietrich et al., 2005). Much of the discussion has centered on the ability of neurobehavioral tests
to detect damage to the central nervous system as a result of in utero or early postnatal Pb
exposures. In other words, the sensitivity of these tests to toxicity has been in question.
The sensitivity of a neuropsychological or any other diagnostic test is defined as the proportion
with the abnormality that the test classifies as abnormal (true positives). In the selecting of
neurodevelopmental measures for use in studies of Pb or any other toxicant, it is clearly
advantageous to include tests that have the best prognostic value. This is particularly important
in the current context, because the neurobehavioral endpoints assessed in this document are
being incorporated into an assessment of risk (Bellinger, 2002). In addition, it is important to
select instruments that tap into neurodevelopmental domains that have been shown to be
sensitive to particular environmental toxicants. As evident in this assessment, numerous
neuropsychological instruments, tapping a wide range of domains have proven to be sensitive to
lower level Pb exposure. Certain domains such as attention, executive functions, visual-spatial
skills, fine-motor abilities, academic achievement (reading in particular), and externalizing
behaviors appear to be affected by Pb with some degree of consistency. However, the
identification of behavioral phenotypes for Pb has been a largely elusive goal. There are a
number of plausible reasons for this. The sample's SES level, pattern and timing of Pb
exposures; nutritional intake; general health; educational opportunities; and the particular
instruments that were employed in a given study probably play an important role in contributing
to between-study differences (Bellinger, 1995; Schantz, 1996). This may be one reason why the
broad net provided by global, multiple domain assessments of cognition such as IQ have proven
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to be the most consistently sensitive across studies of various design and sample characteristics.
These measures combine subscales that are representative of a broad number of underlying
cognitive functions; and they are thusly likely to pick up exposure-related deficits across cohorts
that differ in their functional expressions of toxicity (Dietrich et al., 2005).
The validity of neuropsychological tests as indices of neurodevelopment in Pb studies is
also of concern. In psychometrics, there are various types of validity. But the validity Pb
researchers are usually most concerned about is "construct validity." If a measure has construct
validity, it measures what it purports to measure. Most Pb researchers utilize assessments with
proven construct validity. This means the instruments utilized by the investigator have been
shown that they possess concurrent and predictive "criterion" validity (i.e., it relates to other
manifestations of the construct that the instrument is supposed to be measuring and predicts an
individual's performance in the future in specific abilities). It also means that the instrument
possesses good "convergent validity." That is, that the test returns similar results to other tests
that purport to measure the same or related constructs. Finally, the instrument should
demonstrate "discriminant validity." This means that the instrument is not measuring a construct
that it is not supposed to measure, but rather it discriminates.
Bellinger (2003) states that the general literature attests to robust observations between IQ
and important measures of life success, such as grades in school, years of education, job success,
social status, and income (Neisser et al., 1996; Salkever, 1995). Testing is difficult depending on
examined age, especially for infants who are in a period of rapid developmental change. Also,
the way an infant's cognitive function can be probed is restricted. The lack of continuity
between their response modalities and ones that can be exploited as a child gets older is also a
factor. Still neurobehavioral tests scores in infancy do possess strong concurrent validity.
There are many potential sources of invalidity which researchers take steps to avoid.
These include unreliability (an instrument that, all other things being equal, yields scores that are
unrepeatable and inconsistent) and bias (e.g., due to factors such as culture, gender). Most
modern standardized measures of development and cognitive attainment have taken steps to
reduce these sources of invalidity and must meet certain minimum requirements such as those
formulated by the American Educational Research Association, American Psychological
Association, and the National Council on Measurement in Education (American Educational
Research Association et al., 1999). One reason that global measures of IQ have been used so
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widely is because of their outstanding psychometric properties. The Wechsler series has
excellent reliability and validity (Groth-Marnat, 2003). For example, the average internal
consistency for the Wechsler children's scales across all age groups is 0.96. Test-retest
reliability is similarly very high. The underlying factor structure of these scales has also been
strongly confirmed. The validity of so-called experimental measures of learning and cognition is
sometimes less certain.
All measurement procedures have the potential for error; so, the goal of the researcher is
to minimize it. In elementary psychometric theory, any observed test score is made of the "true"
score plus measurement error. It is assumed that the measurement error in the outcome variable
is essentially random (the child's true score may not be reflected in the observed score because
of errors of administration, inconsistency of administration across examiners, the child's health,
or aspects of the testing environment that are not conducive to performance); thus, this
measurement error does not bias the estimated effect size for exposure but does reduce the power
to detect a significant effect. Therefore, efforts are made to minimize measurement error through
attention to training, establishing inter-examiner reliability, attention to child factors, site factors,
and vigilant monitoring of examiner performance throughout the course of a study (Dietrich
et al., 2005).

6.2.15 Confounding, Causal Inference, and Effect Modification of the
Neurotoxic Effects of Lead in Children
The major challenge to observational studies examining the impact of Pb on parameters of
child development has been the assessment and control for confounding factors. By definition, a
confounder is associated with both the exposure and the outcome and thusly has the potential to
influence the association between the exposure and the outcome. Confounding by various
factors can be controlled for in the design phase of the study or in the analytical phase. In the
realm of Pb research, there are a wide range of potential confounders, the foremost of which is
socioecomic status (SES). Socioeconomic status is measured rather crudely in most studies with
such indices as the Hollingshead Four-Factor Index of Social Position that incorporates
education and income of both parents. However, even these so-called blunt measures often
account for a great deal of the variance in neurodevelopmental outcomes. Given the crude nature
of these measures, to control for confounding by SES as well as rearing environment of the child,
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many recent Pb studies have incorporated more direct assessments such as the HOME scale,
parental intelligence, parental attitude assessments, and measures of parental substance abuse
and psychopathology. Given the relatively high correlation between indices of Pb exposure and
social environmental factors, the consistency among studies in finding effects following
adjustment for these confounding factors is remarkable. In the Boston Prospective Study,
confounding by SES was largely controlled for by study design (Bellinger et al., 1984). That is,
the study subjects were generally middle- to upper-middle-class children in intact families with
college-educated parents. Hence, the potential for confounding by SES in this study was
considerably less compared to other Pb health effect studies; and yet it reported similar and, at
times, even larger effects on neurodevelopmental outcomes. In addition, it is important to
consider the extensive supporting experimental animal evidence not compromised by the
possibility of confounding in examining Pb effects on health (Bellinger, 2004; Davis et al., 1990;
U.S. Environmental Protection Agency, 1986a, 1990).
Another problem in the analyses of data regarding Pb effect on child development is the
lack of critical consideration of which potential confounder in a particular model "owns" the
variance in neurodevelopmental performance. Thus, for example, in the case of social class, it is
assumed that if an effect of Pb is reduced to nonsignificance following adjustment for some
measure of SES standing, the assumption is that all of the variance belongs to the confounder.
However, in some instances this could be seen as an excessively conservative interpretation and
raises the specter of Type II error. Social class could be seen as either a confounder or a proxy
for exposure. In addition, Pb may be on the causal pathway of the association between social
class and IQ. Lower social class in urban children is closely linked to residence in older housing
in poor condition that, in turn, is associated with higher levels of environmental Pb (Clark et al.,
1985). If studies adjust for social class in the usual manner, the effects of the toxicant will be
underestimated (Bellinger, 2004). One extreme example of overcontrol of this nature can be
found in the New Zealand studies where investigators regularly "controlled" for residence in
older "weatherboard" housing (e.g., Fergusson et al., 1988a,b). However, it is worth noting that,
even in the models including this variable, Pb remained a significant predictor of intellectual and
academic under-attainment in the Christchurch Health Study. The proper way to address the
possibility that Pb may be on the causal pathway of the association between social class and IQ
is to use structural equation models, but this has not generally been done.
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In addition to being a confounder, social class and related variables have been shown to
be effect modifiers in many studies of Pb and child development (Bellinger, 2000; long et al.,
2000). Effect modification occurs when the magnitude of an association between an exposure
(Pb) and an outcome (neurobehavior) varies across strata of some other factor (Last, 2001).
The disadvantages that accompany poor education and underemployment have been found to
exacerbate Pb effects when carefully examined (Bellinger et al., 1989). Indeed, evaluating
potential effect modifiers should be considered an important part of an overall data analytic plan.
Most of the important confounding factors in Pb studies have been identified, and efforts
have been made to control them in studies conducted since the 1990 Supplement. Further
discussion on confounding is presented in Section 6.10.3. Invocation of the poorly measured
confounder as an explanation for positive findings is not substantiated in the database as a whole
when evaluating the impact of Pb on the health of U.S. children (Needleman, 1995). Of course,
it is often the case that following adjustment for factors such as social class, parental
neurocognitive function, and child rearing environment using covariates such as parental
education, income, and occupation, parental IQ, and HOME scores, the Pb coefficients are
substantially reduced in size and statistical significance (Dietrich et al., 1991). This has
sometimes led investigators to be quite cautious in interpreting their study results as being
positive (Wasserman et al., 1997). This is a reasonable way of appraising any single study, and
such extreme caution would certainly be warranted if forced to rely on a single study to confirm
the Pb effects hypothesis. Fortunately, there exists a large database of high quality studies on
which to base inferences regarding the relationship between Pb exposure and neurodevelopment.
In addition, Pb has been extensively studied in animal models at doses that closely approximate
the human situation. Experimental animal studies are not compromised by the possibility of
confounding by such factors as social class and correlated environmental factors. The enormous
experimental animal literature that proves that Pb at low levels causes neurobehavioral deficits
and provides insights into mechanisms must be considered when drawing causal inferences
(Bellinger, 2004; Davis et al., 1990; U.S. Environmental Protection Agency, 1986a, 1990).
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6.2.16 Summary of the Epidemiologic Evidence for the Neurotoxic Effects
of Lead in Children

Effects of Pb on neurobehavior have been reported with remarkable consistency across

numerous studies of various designs, populations studied, and developmental assessment

protocols. The negative impact of Pb on IQ and other neurobehavioral outcomes persist in most

recent studies following adjustment for numerous confounding factors including social class,

quality of caregiving, and parental intelligence. Moreover, these effects appear to persist into

adolescence and young adulthood in the absence of marked reductions in environmental

exposure toPb.

• An international pooled analysis of seven prospective studies and several meta-analyses
provide strong evidence that exposure to Pb at low dose has an effect on the intellectual
attainment of preschool and school age children. Recent studies examining the Pb
associations with intellectual attainment and academic performance in children with low
Pb exposures have consistently observed effects at blood Pb concentrations below
10 |ig/dL. The large international pooled analysis of 1,333 children estimated decline of
6.2 points (95% CI: 3.8, 8.6) in full scale IQ for an increase in concurrent blood Pb
levels from 1 to 10 |ig/dL.

• A common observation among some of these studies of low level Pb exposure is the
nonlinear dose-response relationships between blood Pb and neurodevelopmental
outcomes. At first this may seem at odds with certain fundamental toxicological
concepts. However, a number of examples of non- or supralinear dose-response
relationships exist in toxicology. It is conceivable that the initial neurodevelopmental
lesions at lower Pb levels may be disrupting very different biological mechanisms (e.g.,
early developmental processes in the central nervous system) than the more severe effects
of high exposures that result in symptomatic Pb poisoning and frank mental retardation.

• Studies examining aspects of academic achievement related to Pb exposure indicate the
association of deficits in academic skills and performance, which in turn lead to enduring
and important effects on objective parameters of success in real life.

• The effects of Pb on behavior and mood of children has been an important area of recent
research. These studies have demonstrated that the impact of Pb may extend into
increased risk for antisocial and delinquent behavior. This could be a consequence of
attentional problems and academic underachievement among children who have suffered
Pb exposures during their formative years.

• Several studies that have used methods of MRI and MRS to assess direct measures of
brain damage are also adding important and direct evidence of harm due to Pb exposure.
Reduced brain N-acetylaspartate levels observed may be related to decreased neuronal
density or neuronal loss.
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• It is not clear that only periods of peak blood Pb concentrations matter in terms of risks
for neurodevelopmental morbidity. One study attempts to address this question directly
and reports that concurrent blood Pb concentrations always had the strongest association
with IQ as measured at ages 2, 5, and 7 years, with a stronger relationship as the children
grow older.
• Attempts to reverse or limit Pb-associated neurodevelopmental morbidities with
pharmacological or nutritional intervention strategies have thus far been ineffective.
Epidemiologic studies are reporting effects at blood Pb levels for which there is no
effective means of medical or secondary environmental interventions to avoid
developmental morbidity, thus emphasizing the importance of taking primary protective
measures to substantially reduce and ultimately prevent exposure of young infants and
children to Pb.
6.3 NEUROTOXIC EFFECTS OF LEAD IN ADULTS
6.3.1 Summary of Key Findings on the Neurotoxic Effects of Lead in Adults
from the 1986 Lead AQCD
Lead intoxication in adults occurred primarily in occupational settings with historically
high Pb exposure levels. In more recent times, occupational Pb exposure has been reduced to
much lower levels and is often associated with no symptoms. The symptom constellation
associated with high levels of Pb exposure include impaired memory and attention span,
irritability, headache, muscular tremors, and hallucinations (Cantarow and Trumper, 1944) that
may progress to signs of frank encephalopathy (Smith et al., 1938). Symptoms of clinical Pb
intoxication begin with blood Pb >40 |ig/dL (Baker et al., 1979) accompanied by poorer
performance on cognitive and visuomotor tasks, reaction time, verbal learning, and reasoning
ability that reflect involvement of both the central nervous system and the peripheral nervous
system (Arnvig et al., 1980; Campara et al., 1984; Grandjean et al., 1978; Haenninen et al., 1978,
1979; Hogstedt et al., 1983; Mantere et al., 1982; Valciukas et al., 1978; Zimmermann-Tansella
et al., 1983). Impaired occulomotor function, measured by saccade accuracy and velocity,
depended upon the age group of the Pb-exposed worker (Baloh et al., 1979; Glickman et al.,
1984; Spiveyetal., 1980).
With regard to peripheral nerve function as measured by nerve conduction studies, the
28 studies reviewed by the U.S. EPA in the 1986 Lead AQCD found no consistent single nerve
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involved but, overall, the exposed group had slower conduction velocity at blood Pb levels as
low as 30 |ig/dL.
Studies reviewed in 1986 also found that amyotrophic lateral sclerosis (ALS) was
inconsistently associated with elevated Pb levels in the nervous system. Chelation for 1 year did
not did not alter elevated Pb levels in the tissue of patients with motor neuron disease.

6.3.2 Overview of Cognitive and Psychomotor Tests Used to Assess Adult
Lead Exposure
Examination of Pb effects on neurobehavioral performance in adults differs from that in
children, since the neurobehavioral tests in adults focus on loss of abilities previously present
rather than the lack of attainment of those abilities. Also, there is the contribution of cognitive
reserve acquired by years of education, self-education, on-the-job training, avocational, and
non-avocational activities that increases the ability to compensate for the effects of Pb exposure
on learning new information. However, cognitive reserve does not fully protect against the
neurotoxic effects of Pb. The concept of brain reserve capacity has many examples in neurologic
disease where neuropathology progresses in the absence of clinical expression—clinical
parkinsonism develops once -85% of nigrostriatal cells and dopamine are depleted; multi-infarct
dementia is expressed once an aggregate volume of infarction involves 50 to 100 cc of the brain;
the weakness and atrophy associated with poliomyelitis requires 80% loss of the anterior horn
cells; and Alzheimer disease usually requires a frequency for senile plaques and neurofibrillary
tangles of >60% in the hippocampus and cerebral cortex (Satz, 1993). Therefore, the goal is to
identify these conditions in a preclinical phase. For instance, it is now known that, in individuals
diagnosed with Mild Cognitive Impairment (only impairment of recent memory with
preservation of other cognitive domains) are at increased risk of developing Alzheimer disease at
rates of 12 to 15% per year compared to 1 to 2% in age-matched normal patients. Because of
this brain reserve capacity and the decreased exposure to Pb both environmentally and
occupationally, it is not expected to find clinical disease associated with exposure; however,
subclinical effects, even if reversible, are important to identify as they may be impacting brain
reserve capacity. It is known that diminished cognitive reserve increases the risk of decreased
cognitive performance associated with Pb exposure (Bleecker et al., 2002). Therefore, at this
time, it is more critical to study the contribution of Pb to performance after adjusting for
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potential confounders and to identify subpopulations that are at increased risk for the
neurobehavioral effects of Pb. A few studies have stratified outcome using clinical criteria and
found higher Pb levels to be associated with more severe clinical abnormalities (Bleecker et al.,
2003; Bleecker et al., 2005a).
Because alterations in mood may influence neuropsychological performance, many
neurobehavioral batteries use self-administered questionnaires to screen for mood. For example,
the Center for Epidemiologic Studies Depression Scale (CES-D) screens for depression.
Also, the Profile of Mood State (POMS) screens for six subscales, namely anger, confusion,
depression, fatigue, anxiety/tension, and vigor. The six mood scales of the POMS were
originally validated in a clinical psychiatric population; thus, the factor structure needed to be
validated in an occupational population. Factor analysis of the POMS in Pb smelter workers
found only two relevant factors: one composed of five scales (anger, confusion, depression,
fatigue, and tension) and the other contained vigor (Lindgren et al., 1999). This brings into
question the use of the six scales as separate outcome variables in the study of Pb exposure.
The Mini-Mental-State Examination (MMSE), a screening tool for cognitive impairment,
is a compilation of many cognitive domains, including orientation to time and place, registration,
and recall of three words, attention, language, and visual construction, with a total possible score
of 30 (Folstein et al., 1975). The MMSE is sensitive to age and education. In 194 healthy
subjects aged 40 to 89 years with 7 to 21 years of education, only 1% of the subjects obtained an
MMSE score of 24/30 and none below (Bleecker et al., 1988). MMSE errors are sensitive to age
effects, including delayed recall, spelling "WORLD" backwards, and repetition of "no ifs, ands,
or buts." With Pb exposure, examination of errors is important to compare with age-related
changes and to determine the biological plausibility of the effects of exposure, especially when
performing repeated measures of the test. This test is sometimes used to describe a population
and not as an outcome.
Neuropsychological batteries used to screen for Pb effects in adults usually include the
following domains (for a more complete description, see Lezak et al., 2004): attention/
concentration (Digit Span); conceptual and executive functioning (Stroop, Trails B);
visuoperceptive/visuoconstructive (Block Design); visuomotoric (Reaction Time, Pegboard Test,
Digit Symbol Substitution, Trails A); verbal memory (Rey Auditory Verbal Learning Test,
Logical Memory, Paired Associated Learning); and nonverbal memory (Rey-Osterreith Complex
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Figure, Benton Visual Retention). When analyzing possible associations between Pb exposure
and test performance, adjusting for potential confounders is critical. Potential confounders are
age, education (preferably a measure of verbal intelligence), depressive symptoms, medications,
alcohol use, and smoking. In some cases, age and education may serve as effect modifiers; for
example the association of Pb and poorer neurobehavioral outcome was greater in older workers
(Bleecker et al., 1997a) or in those with less cognitive reserve (Bleecker et al., 2002).
Adults with medical conditions requiring medications that have nervous system side
effects, history of severe head trauma, neurodegenerative disease and other neuropsychiatric
conditions that may have a global impact on nervous system performance should be omitted
from analyses for Pb effects.

6.3.3 Adult Environmental Lead Exposure Effects
6.3.3.1 Neurobehavioral Effects Associated with Environmental Lead Exposure
Exposure to chronic low levels of environmental Pb and its association with effects on the
nervous system were examined in several populations originally followed to study conditions
associated with aging: the VA Normative Aging Study (NAS) (Payton et al., 1998; Rhodes et
al., 2003; Weisskopf et al., 2004a; Wright et al., 2003); the Study of Osteoporotic Fractures
(Muldoon et al., 1996); the Kungsholmen Project on aging and dementia (Nordberg et al., 2000)
and the third National Health and Nutrition Evaluation Survey, NHANES III (Krieg et al., 2005).
Studies reviewed in this section are summarized in Annex Table AX6-3.1.
The VA Normative Aging Study (NAS) conducted at the VA Outpatient Clinic in Boston,
MA is a multidisciplinary longitudinal investigation of the aging process established in 1961 and
involving 2,280 men aged 21 to 80 years with no current or past chronic medical conditions.
Participants are evaluated every three years with self-administered questionnaires and Brief
Symptom Inventory (BSI) for psychiatric symptoms. By evaluating relationships of bone Pb
(tibia 21.9 |ig/g and patella 32.1 |ig/g) and blood Pb (6.3 |ig/dL) to psychiatric symptoms in
526 men (age 67 years), Rhodes et al. (2003) found education and mood symptoms for anxiety,
depression, and phobic anxiety potentially to be associated with bone Pb levels after adjusting
for age, age2, alcohol, education and employment variables.
Neuropsychological testing in NAS found response speed to be sensitive to low levels of
Pb, but it was not a consistent finding in all tests measuring the same domain upon examination
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of 141 healthy men with a mean age of 67 years and 14 years of education. The mean blood Pb
level was 6 |ig/dL, patella bone Pb was 32 jig/g bone mineral, and tibia bone Pb was 23 jig/g
bone mineral (Payton et al., 1998). Education was negatively correlated with bone and blood Pb
levels. The handling of multiple comparisons was not addressed.
Another analysis of the NAS (Wright et al., 2003) examined 736 men, mean age 68 years
with education level of 54% high school or less. The mean blood Pb was 5 |ig/dL, and mean
patella and tibia Pb levels were 30 and 22 jig/g bone mineral, respectively. The subjects had a
mean MMSE score of 27. Relation of MMSE scores <24 (n = 41) and blood Pb by logistic
regression estimated an odds ratio of 1.21 (95% CI: 1.07, 1.36). For patella and tibia Pb, odds
ratios of 1.21 (95% CI: 1.00, 1.03) and 1.02 (95% CI: 1.00, 1.04), respectively, were observed.
Risk of MMSE <24 (6% of the present population versus 1% of previously described healthy
aging study), when comparing the lowest and highest quartiles, was 2.1 (95% CI: 1.1, 4.1) for
patella Pb, 2.2 (95% CI: 1.1, 3.8) for tibia Pb, and 3.4 (95% CI: 1.6, 7.2) for blood Pb.
Interaction of age with patella Pb and blood Pb in predicting MMSE found steeper decreases in
MMSE scores relative to age in the higher quartiles of patella and blood Pb. Types of errors on
the MMSE were not included. Not addressed was the issue of how medical conditions and
medications that developed over the duration of the study were handled. Another publication on
this population found that blue-collar participants in NAS had significantly more high school
graduates as well as higher blood and bone Pb compared to white-collar participants, with
non-White blue-collar workers having the highest bone Pb levels (Elmarsafawy et al., 2002).
Weisskopf et al. (2004a) expanded the MMSE study in NAS by examining 466 men
(mean age 70 years), who had completed the MMSE twice with an interval of about 3.5 years.
Mean blood Pb was 4 |ig/dL, and mean patella and tibia bone Pb were 23 and 19 jig/g bone
mineral, respectively. A one-interquartile range (20 |ig/g bone mineral) higher patella Pb
concentration was associated with a MMSE score change of -0.24. This association between
patella Pb and change in MMSE score had a steeper inverse association at lower Pb levels.
Baseline mean MMSE score was 27 and mean change in MMSE score of -0.24 was equivalent
to aging 5 years on baseline MMSE scope. Five years of aging in a healthy population is not
associated with any change in MMSE score (Bleecker et al., 1988). To address the biological
plausibility of change in the MMSE over 3.5 years, errors by functional domain need to be
identified to rule out the possibility of random errors with repeat performance.
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Muldoon et al. (1996) studied participants in the Study of Osteoporotic Fractures for any
association between nonoccupational Pb exposure and cognitive function. The Study of
Osteoporotic Fractures began in 1986 and included women over age 65 years living in four
different communities - Baltimore, MD; Portland, OR; Minneapolis, MN; and the Monongahela
Valley outside of Pittsburgh, PA. A sample of 325 women from rural sites with a mean age of
71 years (mean blood Pb 4.5 |ig/dL) and 205 women from urban sites with a mean age of
69 years (mean blood Pb 5.4 |ig/dL) were examined. The urban group was more educated and
had higher use of cigarettes and alcohol. Performance examined by blood Pb groups adjusting
for age, education, smoking, and alcohol use found no significant differences in the urban group.
However, in the rural group, individuals with blood Pb >7 |ig/dL had significantly poorer
performance when compared to those with blood Pb <4 |ig/dL for Trails B, Digit Symbol, and
Reaction Time. Response time across blood Pb groups increased for the rural group and
decreased or remained the same for the urban group. Mean MMSE for the whole population
was 25, with poorer performance in the rural group. MMSE scores as low as 15 were reported to
be compatible with significant cognitive deterioration, as seen in Alzheimer's disease. Even
though the neuropsychological battery was simple, 9 participants were unable to perform some
of the tests including 3 on the MMSE. Such severe impairments were not found among those
with higher occupational Pb exposures.
In the Kungsholmen Project on aging and dementia in Stockholm, Sweden, no
relationship was found between blood Pb and MMSE (Nordberg et al., 2000). The study
population included 762 participants, with a mean age of 88 years. The mean blood Pb in this
group was 3.7 |ig/dL, and the mean MMSE was 25. In contrast to the other populations
examined, this study cohort was more homogenous, comprised entirely of elderly Swedes.
Their likelihood of prior exposure to elevated Pb levels was low.
NHANES III administered 3 computerized neurobehavioral tests (simple reaction time,
symbol-digit substitution, and serial digit learning) to 5,662 adults aged 20 to 59 years with a
mean blood Pb of 3.30 |ig/dL (range 0.7 to 41.8 |ig/dL). No relationship between blood Pb and
performance was found after adjusting for sex, age, education, family income, race/ethnicity,
computer or video game familiarity, alcohol use, test language, and survey phase. Eleven adults
with blood Pb levels between 25 and 42 |ig/dL were analyzed separately, but no statistically
significant relationship was found after adjusting for the covariates (Krieg et al., 2005).
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6.3.3.2 Summary of Adult Environmental Lead Exposure Effects
The evidence relating environmental Pb exposure with impaired cognitive performance in
the elderly is somwhat mixed. That is, several studies with adequate power using tests sensitive
to the effects of Pb found no association between cognitive performance and blood Pb levels.
On the other hand, those studies that evaluated both blood and bone Pb generally found
significant associations between neurocognitive deficits and bone Pb, but not blood Pb. This
suggests that long-term cumulative exposure, more than current exposure, may contribute to
neurotoxic effects in adults.

6.3.4 Adult Occupational Lead Exposure Effects
Occupational studies on neurotoxic effects of Pb are summarized in Annex
Tables AX6-3.2 through AX6-3.8. Key conclusions from these studies are briefly presented in
this section.
Several studies observed a greater likelihood of neurological symptoms such as difficulty
with concentration, irritability, and muscle pain in workers with elevated blood Pb levels (mean
blood Pb levels >25 ng/dL) (Lucchini et al., 2000; Maizlish et al., 1995). The study by Lucchini
et al. (2000) suggested a threshold for neurological symptoms at a blood Pb of 12 |ig/dL.
However, other studies with higher blood Pb levels (mean blood Pb levels >30 |ig/dL) found no
associations with symptoms related to the nervous system (Chia et al., 1997; Osterberg et al.,
1997).
Schwartz et al. (200la) estimated that performances on psychomotor, motor speed, and
dexterity begin to show a decline at a blood Pb threshold of 18 |ig/dL. Some studies with mean
current blood Pb levels <30 |ig/dL found no significant associations with neurobehavioral
performance, whereas cumulative blood Pb indices reflecting past high exposure were found to
be a predictor of poorer performance (Bleecker et al., 1997a, 2005a; Lindgren et al., 1996).
Recent occupational Pb exposure studies consistently found peripheral sensory nerve
impairment as opposed to the classic motor neuropathy described historically with high Pb
exposure. Sensory nerve conduction studies, most commonly of the median nerve, were related
to long-term exposure, lifetime integrated blood index, and duration of exposure or Pb body
burden (Chia et al., 1996a,b; Kovala et al., 1997; Yokoyama et al., 1998). A possible threshold
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for this effect on the sensory nerves was observed at a blood Pb of 28 to 30 |ig/dL (Bleecker
et al., 2005b, Chuang et al., 2000).
Visual evoked potentials (VEPs) and brainstem auditory evoked potential (BAEPs)
measure speed of conduction in the visual and auditory pathways. A detailed study by Abbate
et al. (1995) found blood Pb to be associated with prolonged VEPs with a threshold effect at
17 to 20 |ig/dL. Four studies examining BAEPs and Pb exposure consistently found prolonged
interpeak latencies in the brainstem auditory pathway more strongly associated with cumulative
or weighted average blood Pb levels (Bleecker et al., 2003; Discalzi et al., 1992, 1993; Holdstein
etal., 1986).
Postural sway is a complex task that requires the integration of visual, vestibular, and
peripheral sensory inputs, as well as motor output. Various blood Pb indices have been
associated with postural sway (Chia et al., 1994a, 1996c; Dick et al., 1999; Ratzon et al., 2000;
Yokoyama et al., 1997). A benchmark dose level (the 95% lower confidence limit of Pb
concentration resulting in an increased probability of an abnormal endpoint) for postural sway
was calculated to be a current blood Pb level of 14 |ig/dL by Iwata et al. (2005).
Parasympathetic and sympathetic integrity was compromised in Pb-exposed workers
beginning at blood Pb levels >20 |ig/dL and possibly lower (Ishida et al., 1996; Niu et al., 2000;
Teruya et al., 1991). Quantitative electroencephalographs found increased beta activity
associated with a mean blood Pb level of 29 |ig/dL (Niu et al., 2000).
Several other studies examined the neurotoxic effects of occupational exposure to
organolead (e.g., trimethyl Pb, tetraethyl Pb) (Balbus et al., 1997, 1998; Schwartz et al., 1993,
2000b, 2001b; Stewart et al., 1999). Exposure to tetraethyl Pb was associated with poorer
performance in many cognitive domains, but most often in manual dexterity and verbal
memory/learning.

6.3.5 Amyotrophic Lateral Sclerosis and Other Neurological Outcomes
Associated with Lead in Adults
Studies reviewed in this section are summarized in Annex Table AX6-3.9. The 1986
Lead AQCD concluded that the evidence for an association of Pb and ALS or motor neuron
disease was inconsistent. The subsequent publications remain mixed, but more studies have
reported an association. Using 109 cases of ALS and 256 controls matched for age, gender,
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and region of residence, Kamel et al. (2002) examined the relation of Pb to ALS, using blood Pb
and bone Pb levels. Ranges of exposure were <1 to 14 |ig/dL for blood Pb, -4 to 107 jig/g for
patella Pb, and -7 to 61 jig/g for tibia Pb. History of occupational Pb exposure increased the
risk of ALS (adjusted odds ratio of 1.9 [95% CI: 1.1, 3.3]). Elevations both in blood Pb and in
patella and tibia bone Pb were found in ALS cases, though the precision of these measurements
was questioned. In summary, this study found Pb exposure from historical questionnaire data
and biological markers to be associated with ALS. The same data were used to determine the
associations of ALS with polymorphism in ALAD and VDR and the influence of genotype in the
previously discussed associations of ALS with Pb (Kamel et al., 2003). The ALAD2 allele was
associated with a 2-fold increased risk of ALS after adjustment for age, gender, region,
education, and physical activity. Additionally adjusting for blood Pb strengthened the
association of ALAD2 and ALS risk. This was not found for bone Pb or occupational history of
Pb exposure. VDR was not associated with Pb or ALS risk.
A study from the Mayo Clinic examined risk factors for sporadic ALS in 45 male ALS
patient-patient control pairs (Armon et al., 1991). When lifetime exposure to Pb exceeded
200 hours, the relative risk for ALS was 5.5 (95% CI: 1.44, 21.0). Overall, men with ALS had
worked more at blue-collar jobs with significantly more time welding or soldering than controls
(p < 0.01). The association between Pb exposure and development of ALS was supported, as
these authors had the same findings in a previous pilot study of another patient population
(Roelofs-Iverson et al., 1984).
Another study of risk factors for ALS in 103 patients found increased odds ratio for
manual occupation (2.6 [95% CI: 1.1, 6.3]) and occupational exposure to Pb (5.7 [95% CI: 1.6,
30]) (Chancellor et al., 1993). A Swedish study of 92 cases of motor neuron disease (includes
ALS, progressive bulbar palsy, and progressive muscular atrophy) found a Mantel-Haenszel
odds ratio for welding equal to 3.7 (95% CI: 1.1, 13.0) (Gunnarsson et al., 1992).
Guidetti et al. (1996) performed a retrospective incidence, prevalence, and mortality
survey in northern Italy. The area studied had documented Pb pollution for years. Based upon
79 cases, incidence and prevalence rates of ALS were comparable to the surrounding area.
A subsequent publication by this group found that mean blood Pb levels in cases of sporadic
ALS and controls were not significantly different (mean blood Pb of 13 |ig/dL versus 11 |ig/dL)
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(Vinceti et al., 1997). Blood Pb was associated with disability due to ALS but no support was
found for involvement of Pb in the etiology of sporadic ALS.
Louis et al. (2003) examined the relationship between blood Pb and essential tremor (ET)
in 100 cases with ET (mean blood Pb 3 |ig/dL) and 143 controls (mean blood Pb 2 jig/dL).
Ten cases and 7 controls had bone Pb levels measured that were significantly correlated with
blood Pb, suggesting that higher blood Pb may have occurred in the past. Logistic regression
adjusting for age and current cigarette smoking found an association between blood Pb and ET.
An odds ratio of 1.19 (95% CI: 1.03, 1.37) was estimated. Blood Pb was higher in the 39 ET
cases with no family history. Both current and lifetime prevalence of occupational Pb exposure
was the same in ET cases and controls. In a second publication (Louis et al., 2005), 63 ET cases
(mean blood Pb of 4 |ig/dL) and 101 controls (mean blood Pb of 3 |ig/dL) who were similar in
age, education, gender, and ethnicity were examined for interaction of blood Pb and ALAD gene
polymorphisms and increased odds of ET. Of the 63 ET cases, 18 (29%) had an ALAD2 allele
compared to 17 (17%) of the 101 controls (odds ratio of 1.98 [95% CI: 0.93, 4.21]). When log
blood Pb was examined by presence of ALAD2 allele in ET, log blood Pb was highest in ET
cases with the ALAD2 allele, intermediate in ET cases without an ALAD2 allele, and lowest in
controls (test for trend, P = 0.10; p = 0.001). When the ALAD2 allele was present, blood Pb was
significantly associated with odds of ET (80.29 [95% CI: 3.08, 2,096.36]). This increased odds
of ET with an ALAD2 allele was 30 times greater than in individuals with only ALAD1 alleles.
In the highest log blood Pb tertile, ALAD2 allele was present in 22% of ET cases and 5% of
controls. It was proposed that increased blood Pb along with the ALAD2 allele could affect the
cerebellum and, thereby, increase the risk of tremor.
Graves et al. (1991) performed a meta-analysis on 11 case-control studies of Alzheimer's
disease for occupational exposure to solvents and Pb. Four studies had data for Pb exposure,
with a pooled analysis of relative risks for occupational Pb of 0.71 (95% CI: 0.36, 1.41).
The exposure frequencies were 16 of 261 (6%) for the cases and 28 of 337 (8%) for the controls.
These nonsignificant results were further confirmed by measuring Pb concentration in the brain
of cases with diffuse neurofibrillary tangles with calcification (DNTC), Alzheimer's disease, and
non-demented controls. The Pb concentration was significantly higher in DNTC compared to
Alzheimer's disease and non-demented controls (Haraguchi et al., 2001).
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In summary, more studies are reporting an association with past exposure to Pb, usually in

the occupational setting, and the motor neuron disease ALS. There appears to be a 2-fold

increased risk for ALS when the ALAD2 allele is present. The odds of ET in individuals

with the ALAD2 allele were 30 times greater compared to those with only ALAD1 alleles.

No increased risk of Alzheimer's disease was related to Pb exposure.
6.3.6 Summary of the Epidemiologic Evidence for the Neurotoxic Effects
of Lead in Adults

Neurobehavioral tests in adults focus on loss of abilities previously present. Cognitive

reserve acquired by years of education and life activities increases the ability to compensate for

the effects of Pb exposure on learning new information. However, it should be noted that

cognitive reserve is not fully protective against the neurotoxic effects of Pb. Several new

publications evaluate effects associated with environmental Pb exposure and other information is

related to effects associated with occupational Pb exposure.

• In the limited literature examining environmental Pb exposure, there appears to exist
mixed evidence regarding associations between Pb and impaired cognitive performance
in adults. Studies using concurrent blood Pb levels as the marker for Pb exposure found
no association between cognitive performance and Pb exposure. However, significant
associations were observed in relation to bone Pb concentrations, suggesting that long-
term cumulative exposure may be crucial in contributing to neurocognitive deficits in
adults.

• Chronic occupational Pb exposure was found to be associated with peripheral sensory
nerve impairment, visuomotor and memory impairment, prolonged VEPs and BAEPs,
and postural sway abnormalities. A possible threshold at blood Pb levels > 14 |ig/dL was
observed for these neurotoxic effects.

• Past occupational exposure to Pb increased the risk of developing ALS and motor neuron
disease in 4 studies. This risk was increased 2-fold by the presence of the ALAD2 allele.
Essential tremor in two well-done studies was associated with low blood Pb levels (mean
3 jig/dL). The odds of developing ET with the ALAD2 allele increased 30-fold
compared to those individuals with only an ALAD1 allele.
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6.4 RENAL EFFECTS OF LEAD
6.4.1 Summary of Key Findings on the Renal Effects of Lead from the
1986 Lead AQCD
Chronic Pb nephropathy is a disease characterized by tubulointerstitial nephritis, which
can ultimately result in small, fibrotic kidneys. It occurs in individuals who sustain chronic high-
level Pb exposure. In these individuals, Pb exposure is the primary cause of renal failure.
The pathophysiologic characteristics of Pb nephropathy and the populations at increased risk for
this diagnosis were the foci of the human research portion of Section 12.5, entitled "Effects of
Lead on the Kidney," in the 1986 Lead AQCD. The 1986 document clearly identified several
high-risk groups for this diagnosis, including children in the Queensland, Australia Pb poisoning
epidemic, moonshine alcohol drinkers, and Pb workers in poorly controlled settings. The section
concluded that data in the latter group indicated an increased risk for Pb nephropathy associated
with blood Pb levels ranging from 40 to >100 |ig/dL, with adverse renal effects possibly
occurring at levels as low as 30 |ig/dL.
The 1986 Lead AQCD noted that research at that time was not sufficient to address some
of the most critical questions relating to the impact of Pb exposure on the kidney. The last
paragraph of the renal section begins with "Among the questions remaining to be answered more
definitively about the effects of Pb on the kidneys is the lowest blood Pb level at which renal
effects occurs." The last sentence reads "Conversely, the most difficult question of all may well
be to determine the contribution of low levels of Pb exposure to renal disease of non-Pb
etiologies." Advances in the research conducted since that document was written allow a much
more informed discussion of exactly those critical issues. As discussed below, recent research
indicates that Pb nephropathy is merely the tip of the iceberg in terms of the contribution that Pb
makes to renal dysfunction overall. Research increasingly indicates that Pb, at much lower doses
than those causing Pb nephropathy, acts as a cofactor with other more established renal risks to
increase the risk for renal dysfunction and the rate of subsequent decline. The populations at risk
for renal dysfunction (diabetics and hypertensives) are increasing worldwide, particularly in
countries where obesity is epidemic. Pb exposure is declining in many industrialized countries,
although less so among high-risk minority populations. The extent of the public health impact of
Pb on the kidney depends on the balance of these two factors.
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6.4.2 Renal Outcome Definitions
The renal literature can be confusing, because several of the clinical renal measures are
inversely related. Therefore, the pertinent outcomes are briefly reviewed below. The glomerular
filtration rate (GFR) is considered to be the best measure of renal function. GFR is assessed by
urinary clearance of exogenous (e.g., 125I-iothalamate) or endogenous (e.g., blood urea nitrogen
[BUN] and serum creatinine) compounds. Creatinine is used most commonly. Therefore,
increases in BUN or serum creatinine or decreases in renal clearance of creatinine or other
markers are all consistent with decreased renal function. Serum creatinine and its reciprocal
have been the most frequently used measures of renal function in the Pb-kidney literature.
However, creatinine is not an ideal GFR marker, because it is influenced by factors such as
muscle mass, diet, gender, age, and tubular secretion. Measurement or calculation of creatinine
clearance takes some of these variables into account. Measured creatinine clearance utilizes
timed urine collections, traditionally over a 24-h period, making compliance difficult. Therefore,
equations to estimate creatinine clearance have gained popularity. The Cockcroft-Gault equation
(Cockcroft and Gault, 1976) has been used most commonly. Recently, several equations to
estimate actual GFR were studied in the Modification of Diet in Renal Disease (MDRD) Study
(Levey et al., 1999). The abbreviated MDRD equation (GFR in mL/min/1.73m2 = 186 x
creatinine L154 x age °'203 x (0.742 if female) x (1.212 if African-American); Stevens and Levey
[2005a]) estimates GFR more accurately than the Cockcroft-Gault equation in patients with renal
insufficiency (Levey et al., 2003). Despite their promise, however, the MDRD equations are
relatively new and their use in studies of renal effects of Pb exposure has been limited to date.
Cystatin C is another recent addition to the tools used to assess GFR (Stevens and Levey,
2005b). This is a 13,000 Dalton, non-glycosylated basic protein, which is generated by all
nucleated cells and filtered, reabsorbed, and catabolized, but not secreted, in the kidney.
Very little appears in the urine. The majority of studies done to date indicate that serum cystatin
C is a better marker for GFR than serum creatinine (Stevens and Levey, 2005b).
Most of the renal outcome measures discussed above were developed for use in the
clinical setting. Unfortunately, they are insensitive for early renal damage, as evidenced by the
fact that serum creatinine remains normal after kidney donation. Therefore, in the last two
decades, the utility of renal early biological effect (EBE) markers as indicators of preclinical
renal damage has been of interest. These can be categorized as markers of function (i.e., low
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molecular weight proteins that should be reabsorbed in the proximal tubules such as
(32-microglobulin and retinol-binding protein [RBP]); biochemical alteration (i.e., urinary
eicosanoids such as prostaglandin E2, prostaglandin F2aipha, 6-keto-prostaglandin FI aipha, and
thromboxane B2); and cytotoxicity (e.g., N-acetyl-(3-D-glucosaminidase [NAG]) (Cardenas et al.,
1993). Elevated levels may indicate an increased risk for subsequent renal dysfunction.
However, with the exception of microalbuminuria in diabetes and (32-microglobulin in Cd
exposure, most are research tools only, and their prognostic value remains controversial. Asian
and European nephrotoxicant researchers have used them more frequently than have U.S. renal
researchers. Prospective studies of most of these markers in nephrotoxicant-exposed populations
are quite limited to date.

6.4.3 Lead Exposure Measure Definitions
Although these definitions are reviewed in detail elsewhere in this Lead AQCD, a brief
discussion is included here due to the number of key studies in this section that measured bone or
chelatable Pb dose. Inorganic Pb is a cumulative toxicant that is stored in bone. Blood Pb is a
relatively short-term measure (half-life of 30 days [Hu et al., 1998]) that reflects exposure from
current exogenous sources and the release of Pb from internal Pb stores. Bone is an internal
source of Pb as well as a repository (Hu et al., 1998). As such, bone Pb measures provide an
index not only of cumulative Pb exposure but also the potential for ongoing internal exposure,
as well. Lead in trabecular bone (commonly measured in the patella or calcaneus) is more
bioavailable than Pb in cortical bone (measured in the mid-tibia) and has a shorter half-life
(Gerhardsson, et al., 1993; Hu et al., 1998). An additional Pb measure, chelatable Pb, is thought
to represent a bioavailable pool of Pb from blood, soft tissue, and bone. Two chelation agents,
either calcium disodium ethylenediaminetetraacetic acid (EDTA) or dimercaptosuccinic acid
(DMSA; succimer) have mainly been used for this purpose, although DMSA is newer and, thus,
used less frequently to date.

6.4.4 Lead Nephrotoxicity in Adults
6.4.4.1 General Population Studies
Over the past two decades, several studies have examined the effect of Pb exposure on
renal function in general populations. This is a new category of Pb-renal research. No high
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quality examples (by current standards) were available for review in the 1986 Lead AQCD.
The studies discussed below provide critical evidence that the adverse effects of Pb on the
kidney occur at much lower doses than previously appreciated. Traditional renal function
measures, such as serum creatinine, BUN and creatinine clearance, are emphasized below, since
much more is known regarding the clinical relevance of these measures than for the renal EBE
markers. General population studies of the renal effects of Pb are further summarized in
Annex Table AX6-4.1.

6.4.4.1.1 Cadmibel Study
In the first large environmental study that adjusted for multiple renal risk factors, Staessen
et al. (1992) evaluated 965 men and 1,016 women in the Belgian Cadmibel study. Lead dose
was indexed by blood Pb and zinc protoporphyrin. Renal outcome measures included (a) serum
creatinine and p2-microglobulin and (b) 24-h measured and calculated (Cockcroft and Gault,
1976) creatinine clearances. Mean blood Pb was 11.4 |ig/dL (range 2.3-72.5) and 7.5 |ig/dL
(range 1.7-60.3) in men and women, respectively. After adjustment, log transformed blood Pb
and zinc protoporphyrin, in separate models, were negatively associated with measured
creatinine clearance. A 10-fold increase in blood Pb was associated with a decrease in creatinine
clearance of 10 and 13 mL/min in men and women, respectively. Both Pb measures were also
negatively associated with estimated creatinine clearance. This landmark study raised concern
that the Pb dose threshold for adverse renal effects in the general population might be much
lower than had been previously appreciated based on occupational exposure data.

6.4.4.1.2 Normative Aging Study
Research in the Normative Aging Study population reached similar conclusions. Four
studies assessing the renal impact of Pb exposure in this population have thus far been published.
Participants in this study were originally recruited in the 1960s in the Greater Boston area.
Inclusion criteria included male gender, age 21 to 80 years, and absence of chronic medical
conditions. Payton et al. (1994) analyzed data from a periodic follow-up evaluation performed
between 1988 and 1991 in 744 participants. Lead dose was indexed by blood Pb; renal outcome
measures included serum creatinine and 24-h measured and calculated (Cockcroft and Gault,
1976) creatinine clearances. Mean blood Pb concentration and measured creatinine clearance
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were 8.1 ng/dL (SD 3.9) and 88.2 mL/min (SD 22.0), respectively. After adjustment, In blood
Pb was negatively associated with In measured creatinine clearance (P = -0.04 [95% CI:
-0.079, -0.001]). Borderline statistically significant associations (p < 0.1) between blood Pb
and serum creatinine and estimated creatinine clearance were also observed. Kim et al. (1996)
studied 459 men whose blood Pb levels from past periodic examinations, conducted every 3 to
5 years during 1979-1994, were measured from stored samples. Participants were randomly
selected to be representative of the entire Normative Aging Study population in terms of age and
follow-up. Renal status was assessed with serum creatinine. Data from 4 to 5 evaluations were
available for the majority of participants. Relations were evaluated cross-sectionally
(associations between blood Pb and concurrent serum creatinine) as well as longitudinally
(associations between blood Pb and change in serum creatinine over the subsequent follow-up
period). Mean age, blood Pb level, and serum creatinine, at baseline, were 56.9 years (SD 8.3),
9.9 ng/dL (SD 6.1), and 1.2 mg/dL (SD 0.2), respectively. With random-effects modeling, a
significant positive association between In-transformed blood Pb and concurrent serum
creatinine was observed. This association was stronger when models were confined to
participants with lower peak blood Pb levels, i.e., the P coefficient was largest in the
141 participants whose highest blood Pb level was < 10 ng/dL (P = 0.06 [95% CI: 0.023,
0.097]). In the longitudinal analysis, In-transformed blood Pb was associated with change in
serum creatinine over the subsequent follow-up period in the 428 participants whose highest
blood Pb level was <25 jig/dL (P = 0.027 [95% CI: 0.0, 0.054]). Similar to the cross-sectional
analysis, the P coefficient in the participants whose highest blood Pb level was < 10 |ig/dL was
larger; however, in the longitudinal analysis, the standard error also increased such that the
p-value was not significant.
Cortical and trabecular bone Pb measurements were obtained in evaluations performed
between 1991 and 1995 in 709 participants in the Normative Aging Study (Wu et al., 2003a).
Lead dose was assessed with blood, tibia, and patella Pb concentrations. Renal outcome
measures included serum creatinine and estimated creatinine clearance. Mean blood, tibia and
patella Pb levels were 6.2 ng/dL (SD 4.1), 22.0 |ig/g bone mineral (SD 13.4), and 32.1 |ig/g bone
mineral (SD 19.5), respectively. After adjustment, analyses in the 670 participants from whom
these data were available, revealed a significant inverse association between patella Pb and
creatinine clearance (P = -0.069 [SE not provided]). A borderline significant (p = 0.08) inverse
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association between tibia Pb and creatinine clearance was also observed. None of the Pb
measures were significantly associated with serum creatinine.
Tsaih et al. (2004) reported associations between baseline Pb dose and change in serum
creatinine in 448 men. Lead dose was assessed in terms of blood, tibia, and patella Pb. Serum
creatinine was measured at baseline and at follow-up, an average of 6 years later. Six percent
and 26% of subjects had diabetes and hypertension, at baseline, respectively. Mean blood Pb
levels and serum creatinine decreased significantly over the follow-up period in the group. Lead
dose was not associated with change in creatinine in all participants. However, a significant
interaction was found between blood and tibia Pb and diabetes on change in serum creatinine.
For In blood Pb, P = 0.076 (95% CI: 0.031, 0.121) in diabetics compared to P = 0.006 (95% CI:
-0.004, 0.016) in non-diabetics. A similar relationship was observed for tibia Pb. An interaction
was also observed between tibia Pb and hypertension, although it is possible that many of the
26 diabetics were also included in the hypertensive group and were influential there as well.

6.4.4.1.3 NHANES III
Muntner et al. (2003) analyzed associations between blood Pb and renal outcomes in
15,211 adult subjects enrolled in the NHANES III study, conducted from 1988 through 1994.
Dichotomous renal outcome measures analyzed included elevated serum creatinine and chronic
kidney disease (GFR < 60mL/min/1.73 m2). Due to an interaction between blood Pb and
hypertension, the population was stratified. Mean blood Pb level was 4.21 |ig/dL in the
4,813 hypertensives and 3.30 |ig/dL in normotensives. The prevalence of elevated serum
creatinine in hypertensives and nonhypertensives was 11.5% and 1.8%, respectively, but the
prevalence of chronic kidney disease was similar. The odds ratios for both renal outcomes
increased by quartile of blood Pb among the hypertensive subjects but not among those without
hypertension. Among those with hypertension, after adjustment for age, race and gender, the
odds ratios for elevated creatinine in quartiles 2, 3, and 4 compared to the lowest quartile of
blood Pb, were 1.56 (95% CI: 1.04, 2.35), 1.68 (95% CI: 1.24, 2.26), and 2.07 (95% CI: 1.26,
3.40), respectively. The odds ratios were the same following additional adjustment. The authors
noted that the "associations were strong, dose-dependent and consistent before and after
comprehensive adjustment." They also noted that in nonhypertensives, higher blood Pb was
associated with a higher prevalence of chronic kidney disease in diabetics. This study is notable
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for sample size, comprehensive adjustment for other renal risk factors, and the fact that this study
population is representative of the U.S. non-institutionalized, civilian population.

6.4.4.1.4 Women's Health in the Lund Area Study
In a study of 820 women (age 53 to 64 years) in Sweden, significant negative associations
were observed between blood Pb and both GFR (estimated from serum cystatin C) and creatinine
clearance (estimated by the Cockcroft-Gault equation [Cockcroft and Gault, 1976]) (Akesson
et al., 2005). Mean blood Pb was only 2.2 |ig/dL; the association was apparent over the entire
dose range (Akesson, 2006). This study has the additional advantage of blood and urinary Cd
assessment.

6.4.4.1.5 Summary of Lead-Related Nephrotoxicity in the General Population
General population studies constitute one of the two most important types of research on
the renal effects of Pb during the past two decades. Overall, a number of strengths are present in
this body of literature. These include study design with longitudinal data in some studies; large
populations in both Europe and the United States; comprehensive assessment of Pb dose,
including the use of bone Pb as a measure of cumulative Pb body burden in some studies; and
statistical approaches that utilize a range of exposure and outcome measures, while adjusting for
numerous renal risk factors. Associations between Pb dose and worse renal function were
observed in most of the general population studies.

Threshold for Lead-Related Nephrotoxicity
Increased risk for nephrotoxicity has been observed at the lowest Pb dose levels studied to
date. Specifically, blood Pb ranged from 2.5 to 3.8 |ig/dL in the first significant category in
Muntner et al. (2003), and associations between blood Pb as a continuous variable and worse
renal function have been reported at a mean of 2.2 |ig/dL (Akesson et al., 2005). An association
between cumulative Pb dose (mean tibia Pb of 21.5 jig/g bone mineral) and longitudinal decline
in renal function has been observed as well, although data on any threshold for this effect were
not reported (Tsaih et al., 2004). The data available to date are not sufficient to determine
whether nephrotoxicity is related more to current blood Pb levels, higher levels from past Pb
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exposures, or both. However, Kim et al. (1996) noted associations in participants whose peak
blood Pb levels were < 10 |ig/dL as far back as 1979.

Alternative Explanations for Observed Associations
Potential residual confounding as a possible explanation for associations between Pb dose
and adverse health effect outcomes is always a consideration. One general population study
provided data useful to address this concern in the Pb-renal literature. For both renal outcomes
assessed, Muntner et al. (2003) observed that the odds ratios in hypertensives initially adjusted
for age, race, and gender, increased further after additional adjustment for diabetes, systolic
blood pressure, smoking status, history of cardiovascular disease, body mass index, alcohol
consumption, household income, education level, marital status, and health insurance.
In contrast, after adjustment, regression coefficients decreased in Wu et al. (2003b). However,
the analyses were performed in slightly different populations, making interpretation of the
adjustment differences less certain. Further, as noted in the Agency for Toxic Substances and
Disease Registry's Draft Toxicological Profile For Lead ([2005] Atlanta, GA: U.S. Department
of Health and Human Services), since increased blood pressure is associated with Pb dose in
general populations, adjustment for hypertension or blood pressure, although extremely common
in Pb-renal studies, risks underestimating the actual slope of the association between Pb dose and
renal dysfunction. Overall, one of the strengths of the Pb-renal general population literature is
the number of factors adjusted for. Thus, residual confounding is an unlikely explanation for
observed associations.
Reverse causality has also been considered as a possible explanation for associations
between lower blood Pb levels (e.g., <10 |ig/dL) and worse renal function (Staessen et al., 1992).
Reverse causality attributes increased Pb dose to reduced Pb excretion as a consequence of renal
insufficiency. The temporal relation between Pb dose and renal function decline is a critical
factor in determining causality. This can be assessed in longitudinal observations of participants
with mean blood Pb levels in this lower dose range. Two analyses of longitudinal data from the
Normative Aging Study population have been published to date (Kim et al., 1996; Tsaih et al.,
2004). Lead dose predicted subsequent decline in renal function over follow-up periods ranging
from 3 to 6 years. This was observed even after adjustment for renal function at the beginning of
the follow-up period. Longitudinal studies in patients with renal insufficiency have reported
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similar findings. Both blood and EDTA-chelatable Pb levels at baseline were significantly
associated with decline in estimated GFR over a 4 year follow-up period in 121 patients, even
after adjustment for a wide range of covariates, including baseline renal function (Yu et al.,
2004) (discussed in Section 6.4.4.3). The same was true in a larger study of 202 chronic renal
insufficiency patients over a 2-year follow-up period (Lin et al., 2003). Notably, in both studies,
EDTA-chelatable Pb levels were <600 |ig/72 h in all participants, with means well below this
traditional cut-point. The PheeCad study (the 1990-95 follow-up to the Cadmibel study) appears
to have collected relevant data, but the Pb data were not reported in the publication (Hotz et al.,
1999).
Biologically, reverse causality should be most prominent in populations with renal
insufficiency for a prolonged period of time. However, Kim et al. (1996) observed that blood Pb
was positively associated over the entire serum creatinine range, most of which was normal in
this general population study and where a substantial decrease in Pb excretion was unlikely.
Further, in reverse causality, urinary excretion of Pb should decrease as renal function declines.
Urine Pb is not a commonly used Pb dose biomarker, so data from the lower Pb exposure studies
are generally not available to assess this. However, higher urine Pb was associated with lower
estimated creatinine clearance in Swedish women (Akesson, 2006). Finally, the positive impact
of Pb chelation on renal function (discussed in Section 6.4.4.3) may provide evidence against
reverse causality. However, the possibility of a direct beneficial effect of the chelating agent on
renal function cannot be excluded as an explanatory factor (Gonick et al., 1996). In summary,
several lines of evidence suggest that reverse causality is not likely to be a major explanatory
factor accounting for observed associations between Pb dose and renal dysfunction.

Consistency of the Magnitude of Associations
Slopes of the associations between blood Pb and creatinine clearance in the general
population studies that provide data relevant for such a comparison are shown in Figure 6-8.
Since these studies generally had mean blood Pb levels less than 10 |ig/dL, slopes of the reported
relations were estimated at a blood Pb level of 5 |ig/dL. Measured or estimated creatinine
clearance data were used from those studies that reported relations for those outcomes.
For studies that only reported data for serum creatinine, the slope at a blood Pb of 5 |ig/dL
was estimated and then the slope was converted to a creatinine clearance slope using the
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Tsaih et al. (2004), follow-up

Kim etal. (1996)
Tsaih et a!. (2004), baseline

Payton etal. (1994)
Akesson et al. (2005)
Staessen et al. (1992), men

Staessen et al. (1992), women
-4.0 -2.0 0.0 2.0 4.0
Slope for Creatinine Clearance per Blood Lead (ml_/min)/(fjg/dl_)
Figure 6-8. Creatinine clearance versus blood lead slope at a blood lead of 5 ug/dL.

Cockcroft-Gault equation (Cockcroft and Gault, 1976). Publication bias may impact the data
available for this figure. No significant associations between blood Pb and renal function were
observed in two of the general population studies; beta coefficients were not reported (Wu et al.,
2003a; de Burbure et al., 2003). However, since Wu et al. (2003a) observed a significant
association between patella Pb and creatinine clearance, the study is consistent with results in the
majority of the other general population studies. Lastly, a third study (Pocock et al., 1984)
reported only that the correlation coefficient between crude blood Pb and serum creatinine was
0.0. Furthermore, publications derived from evaluation of the Normative Aging Study
population outnumber those from other populations. Slopes ranged from 0.2 to -1.8 mL/min
change in creatinine clearance per |ig/dL increase in blood Pb.

Clinical Relevance
It is now clear that chronic kidney disease (CKD) at earlier stages than those requiring
actual renal dialysis or transplantation represents a risk factor for cardiac disease and other
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causes of mortality and morbidity (Levey et al., 2003). The clinical relevance of the Pb effect
can be estimated from the study by Akesson et al. (2005), in which the 5th and 95th percentile
values for blood Pb were reported. An increase in blood Pb from the 5th to the 95th percentile
(3.5 |ig/dL) has the same adverse impact on glomerular filtration as an increase of 4.7 years in
age or 7 kg/m2 in body mass index, both of which are known renal risk factors. In populations at
high risk for Pb exposure, a 10-fold increase in blood Pb (e.g., from 1 to 10 |ig/dL) would result
in an 16.2 mL/min decrease in estimated creatinine clearance or a 22.5% decrease from the mean
(Akesson et al., 2005). Sixteen and 9% declines due to a 10-fold increase in blood Pb were
predicted based on data for women (Staessen et al., 1992) and men (Payton et al., 1994),
respectively. Although Pb exposure is higher in rapidly industrializing countries, high risk
populations remain in the United States. In populations with lower blood Pb levels, a downward
shift in renal function of the entire population due to Pb may not result in CKD in identifiable
individuals; however, that segment of the population with the lowest renal reserve may be at
increased risk for CKD when Pb is combined with another renal risk factor. The potential public
health importance of population shifts is discussed by the American Thoracic Society (2000) and
Rose and Day (1990). Data in both general and patient populations support this concept for Pb
exposure. Of note, the above estimates are for general populations. Effect estimates for
susceptible populations, such as those with diabetes, hypertension, or chronic renal insufficiency
from non-Pb related causes, are likely to be higher.

At-risk Populations
Susceptible populations include those with other risk factors for renal disease, including
hypertension, diabetes, and renal disease from other causes. Lead-exposed populations also at
increased risk for obesity, diabetes, and hypertension represent groups likely to be the most
impacted by Pb exposure. Frequently, both Pb and other risk factors are present in the same
lower SES status groups.
In conclusion, the general population literature on the adverse renal effects of Pb benefits
from a number of strengths. The consistent associations observed in the majority of these studies
provide strong evidence indicating that Pb is a contributor to renal dysfunction in susceptible
populations at much lower Pb exposure levels than those previously identified based on data
available at the time of the 1986 Lead AQCD.
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6.4.4.2 Occupational Studies
The vast majority of studies in the Pb-renal literature were conducted in the occupational
setting. This was especially true prior to the 1986 Lead AQCD, but is still also currently the
case. Occupational studies of the renal effects of Pb are presented in Annex Table AX6-4.2.
In contrast to the general population research discussed above, research on the adverse renal
effects of occupational Pb exposure is much less consistent. This is puzzling, since most dose-
response relations are thought to be linear. Therefore, biologically, notably elevated Pb doses
(as indexed by 30-50 |ig/dL blood Pb levels) should be nephrotoxic if lower doses are. Several
explanations for this seeming inconsistency are possible. Some are unique to the occupational
literature, such as smaller sample sizes. In addition, employed workers are typically healthier
and younger than the general population—resulting in the healthy worker bias. This is a
particular problem as susceptible risk groups are identified. Survivor bias in cross-sectional
studies is also a concern, since workers whose renal function has declined are generally removed
from exposure, particularly if they are followed in a medical surveillance program. Few studies
have included former workers. Also, statistical analyses have been more limited in occupational
studies. Analyses for some outcomes were limited to comparisons between exposed workers and
controls whose Pb levels were in the range associated with adverse renal outcomes in
environmental work. Use of multiple linear regression has generally involved more limited
adjustment for covariates than in most of the environmental studies. Many of these limitations
result in bias towards the null, which increases the risk that true associations may not be
detected.
Other limitations are pertinent for research on the adverse renal effects of Pb exposure in
any population. These factors are likely to have a greater impact on the validity of studies in
which one or more of the biases discussed above are also present. These include the insensitivity
of the clinical renal outcomes and the lack of uniformly accepted early markers of renal damage
in Pb exposure. Limited Pb exposure assessment may also be a factor. Finally, Pb appears to be
able to induce an element of hyperfiltration in some settings. Hyperfiltration is a process initially
observed in diabetes but is also implicated in other settings, including hypertension and obesity
(Nenov et al., 2000). In this process, initial supranormal renal function is paradoxically
associated with increased risk for subsequent renal dysfunction. Several occupational studies
have reported statistically significant higher mean creatinine clearance in Pb-exposed workers
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compared to controls and/or positive associations between higher Pb dose and lower BUN,
serum creatinine and/or higher creatinine clearance (Roels et al., 1994; Weaver et al., 2003a,
2005a; Hsiao et al., 2001). Hu (1991) has also reported increased mean creatinine clearance in
22 adults who were Pb poisoned as children, as compared to matched controls (discussed in
Section 6.4.5.1), and a recent study reported higher blood Pb to be associated with lower serum
creatinine and cystatin C in a study of 800 European children (discussed in Section 6.4.5.3).
Longitudinal data for Pb-exposed rodents (discussed in Section 5.7.4.2) are critical in relating
this process to Pb. However, in that work, despite similar initial hyperfiltration, subsequent renal
dysfunction was much more severe in the high-dose Pb-exposed rodents compared to the low-
dose animals. This suggests that hyperfiltration may be one, but not the only, mechanism
underlying adverse renal effects of Pb. Whether hyperfiltration contributes to pathology in
humans is unclear; longitudinal studies are needed. Regardless, the issue for risk assessment is
that significant findings could be obscured if opposite direction associations are present in
different segments of the study population and interaction models are not performed to address
this.
In the work of Weaver et al. (2003a), no associations were observed when the entire
population was studied by several models; however, when interaction models using age as the
effect modifier were evaluated, significant associations in opposite directions were observed.
This is illustrated in Figure 6-9. This is a valid concern for risk assessment, since the factors
involved in these inverse associations in Pb-exposed populations are not well defined at present.
Weaver and colleagues have used age as the effect modifier; however, other factors, such as Pb
job duration, may be important as well.
In conclusion, a number of limiting factors are observed in the body of research on
occupational Pb exposure and adverse renal outcomes. Most of these factors increase the risk
that true associations will be missed (bias towards the null). Moreover, Pb appears to have a
paradoxical effect on the kidney that further increases this possibility. As a result, the more
consistent body of literature in general populations at current Pb exposure conditions provide an
appropriate data base for assessing potential renal effects.
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Worse
Renal
Function
Potential
Hyperfiltration
Older
Lead dose
Figure 6-9. Effect on associations between lead dose and renal function depending on
whether effect modification (age in this example) is assessed.
6.4.4.3 Patient Population Studies
Studies in various patient populations have also contributed to the body of knowledge
concerning adverse renal impacts of Pb exposure (summarized in Annex Table AX6-4.3).
Populations studied include those with chronic renal insufficiency (CRT), end-stage renal disease
(ESRD), gout, and hypertension, since these diseases are thought to be increased by high-level
Pb exposure, particularly when two or more coexist in the same patient. Early research focused
on patients with potential Pb nephropathy; and Pb body burdens of interest, assessed with EDTA
chelation, were above 600 to 650 jig/72 h. These studies suggested that chelation might be
beneficial in Pb nephropathy (Morgan, 1975; Wedeen et al., 1979).
Recurring concerns in this work are, first, whether Pb body burden is higher in all patients
with renal insufficiency or failure due to decreased Pb excretion (reverse causality); and, second,
whether EDTA-chelatable Pb levels, when measured over a 72-h period in patients with CRT, can
be equated to those in participants with normal renal function measured over 24 h. It is possible
that, due to decreased excretion of EDTA in renal insufficiency, more Pb per dose is ultimately
chelated.
Chelation also may have a direct beneficial effect on kidney function, regardless of Pb
exposure, since DMSA has been reported to prevent renal damage in a non-Pb-exposed rat
model of nephrosclerosis (Gonick et al., 1996). If so, the benefits of chelation do not appear to
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occur via reversal of structural damage (Khalil-Manesh et al., 1992); improved hemodynamics
from reduction of reactive oxidant species may be a mechanism (Gonick et al., 1996).
In one of the key studies, Yu et al. (2004) followed 121 patients over a 4-year period.
Eligibility required well-controlled CRI. Importantly, serum creatinine between 1.5 and
3.9 mg/dL and EDTA-chelatable Pb <600 jig/72 h were required at baseline. Patients with
potentially unstable renal disease were excluded (i.e., due to systemic diseases such as diabetes).
Mean age of the study population was 57 years. Mean blood Pb and EDTA-chelatable Pb levels
were 4.2 jig/dL and 99.1 jig/72 h, respectively. In a Cox multivariate regression analysis,
chelatable Pb was significantly associated with overall risk for the primary endpoint (doubling of
serum creatinine over the 4-year study period or need for hemodialysis). The hazard ratio for
each 1 jig chelatable Pb was 1.01 (95% CI: 1.00, 1.01; p = 0.002). Of the many traditional renal
risk factors adjusted for in these models, only the diagnosis of chronic interstitial nephritis was
significantly associated with an increase in GFR. Associations between baseline chelatable Pb or
blood Pb level and change in GFR (estimated by an MDRD equation [Levey et al., 1999]) were
modeled separately using GEE. Based on these models, a 10 jig higher chelatable Pb level or
1 jig/dL higher blood Pb level reduced the GFR by 1.3 and 4.0 mL/min, respectively, during the
4-year study period. This work supports results observed for general populations by suggesting
that Pb is nephrotoxic in susceptible populations at lower levels than currently appreciated.

6.4.4.4 Mortality Studies
As summarized in Steenland et al. (1992), mortality studies have consistently shown
excess mortality from chronic kidney disease in Pb workers. This increased risk has been most
apparent in workers exposed in earlier time periods, becoming nonsignificant in later calendar
time periods in a number of studies. Steenland et al. (1992) reported similar results in a study of
1990 former Pb smelter workers. This cohort was made up of predominantly White men who
had worked in a Pb-exposed department for at least 1 year between 1940 and 1965. Mean (SD)
blood Pb, measured in 1976 in 173 members of this cohort, was 56.3 jig/dL (12.9). There were
8 deaths from chronic kidney disease. Compared to the U.S. White male population, the
standardized mortality ratio was 1.26 (95% CI: 0.54, 2.49). The standardized mortality ratio
increased with duration of exposure from 0.79 in Pb workers exposed 1 to 5 years to 2.79 in
workers exposed for >20 years, although the standardized mortality ratios did not reach
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statustical significance (CI not reported). Lead exposure in U.S. industries has declined over the
years, and this has been hypothesized as an explanation for the reduction in mortality from renal
disease observed in this type of study. However, that fact that improved treatments for chronic
renal disease have led to a decrease in mortality from end-stage renal disease (U.S. Renal Data
System, 2004) may also be an important factor. The mortality studies by Steenland et al. (1992)
and others are described further in Annex Table AX6-4.4.

6.4.5 Lead Nephrotoxicity in Children
6.4.5.1 Studies in Adults Following Childhood Lead Poisoning
Henderson clearly established an increased risk for Pb nephropathy in adult survivors of
untreated childhood Pb poisoning (Henderson, 1955). Lead nephropathy was responsible for
substantial mortality in the Queensland, Australia population. However, as noted in the 1986
Lead AQCD, other studies of adults who survived childhood Pb poisoning have not reported this
degree of renal pathology. Studies published since 1986 are presented in Annex Table AX6-4.5
and also have not observed the degree of renal pathology noted in the Queensland work.
Chelation when Pb poisoning was diagnosed may be an explanatory factor in some of these
studies.
A study comparing 21 adults, who had experienced childhood Pb poisoning between 1930
and 1942, to age-, sex-, race-, and neighborhood-matched controls found no significant
differences in blood Pb level, serum creatinine, or BUN (Hu, 1991). Mean measured creatinine
clearance was unexpectedly higher in the previously Pb-poisoned group compared to controls
(112.8 versus 88.8 mL/min/1.73 m2 [p < 0.01]). The mean in the Pb-exposed group was also
higher than the predicted value of 94.2 mL/min/1.73 m2 from the nomogram of Rowe et al.
(1976). One survivor, who was identified but not included in the study, had been diagnosed with
chronic interstitial nephritis on renal biopsy. Her blood Pb was 30 |ig/dL, and her presentation
was thus consistent with actual Pb nephropathy. Strengths of this study included clear criteria
for Pb poisoning and assessment of clinical renal function that included both measured and
estimated creatinine clearances. However, the study was limited by small size and the fact that
the number enrolled was a very small subset of the initially identified cohort of 192. At least
43 (22.4%) of the 192 were confirmed to be deceased. That group had evidence of higher initial
Pb exposure, which raises concern regarding survivor bias in the study group. More importantly,
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the higher mean creatinine clearance in the Pb-exposed group provides further evidence for
Pb-related hyperfiltration. Again, as discussed in the occupational study section, this may
hamper attempts to detect associations between Pb dose and adverse renal effects.

6.4.5.2 Lead Body Burden in Children with Chronic Renal Disease
Scharer et al. (1991) reported higher Pb content in deciduous teeth in 22 German children,
age 5 to 14 years, with varying degrees of renal insufficiency compared to a control group of
20 siblings or neighbors and a group of 16 children without known Pb exposure. Mean dental Pb
content was 2.8, 1.7, and 1.4 |ig/g, in the three groups, respectively. Lead levels in teeth were
significantly higher in both the patient and sibling/neighbor control groups compared to the
unexposed control group. Mean blood Pb in the renal patients was only 2.9 |ig/dL (range
1.1-10.1 |ig/dL). Lead in teeth was not correlated with duration of renal impairment.
The authors attributed elevated Pb levels to both exposure and accumulation from decreased
renal excretion.

6.4.5.3 Environmental Studies in Children
The insensitivity of the clinical renal outcome measures for early renal damage is a
particular problem in children who do not have many of the other renal risk factors, such as
hypertension and diabetes, that older adults do. As a result, recent studies in children have
favored early biological effect (EBE) markers over clinical renal measures. However, data to
determine the predictive value of such biomarkers for subsequent renal function decline in Pb
exposed populations are extremely limited. Coratelli et al. (1988) reported a decline in urinary
NAG in association with a 1 month period of decreased occupational exposure in 20 adult Pb
battery factory workers followed over a 1 year period. Clinical renal function measures were not
studied however. Sarasua et al. (2003) studied 526 adults and children, a mean of 4.5 years after
an initial evaluation of renal function including measurement of urinary albumin, NAG, RBP,
and alanine aminopeptidase. These participants were drawn from three populations exposed to
volatile organic compounds and explosives via groundwater and controls. Follow-up was
performed to determine if the EBE markers remained elevated and whether the presence of
elevated EBE markers at baseline was associated with abnormalities in serum creatinine, serum
cystatin C, 24 h creatinine clearance, and urine osmolality at follow-up. Among children who
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had elevated EBE markers at baseline, renal EBE markers remained elevated in 38%. However,
none remained elevated in the 32 who had completed adolescence by the time of the follow-up.
The authors noted the potential for puberty related biomarker changes. Also, abnormalities in
the clinical measures were rare at follow-up.
The environmental studies in children generally focused on children living near industrial
sources and controls. These studies are summarized in Annex Table AX6-4.5. Three studies
that included analysis of clinical renal outcomes are of note. Pels et al. (1998) found no
difference in mean serum creatinine between 62 exposed and 50 control children; correlations, if
assessed were not reported. Staessen et al. (2001) studied 200 17-year-old Belgian children.
The two exposed groups were recruited from industrialized suburbs, whereas the control group
was recruited from a rural area. Mean blood Pb levels were 1.5, 1.8, and 2.7 |ig/dL in controls,
and exposed groups one and two, respectively. Although blood Pb levels were low, after
adjustment for sex and smoking status, blood Pb was positively associated with both serum
cystatin-C and urinary p2-microglobulin. Blood Cd was not associated with either outcome.
In contrast, De Burbure et al. (2006) observed associations between higher blood Pb and lower
serum creatinine and cystatin C in models with 300-600 European children (depending on
outcome). The authors considered this to be suggestive of hyperfiltration. Additional research in
children, including longitudinal follow-up, is needed.

6.4.6 Mechanisms for Lead Nephrotoxicity
Individuals who have been heavily exposed to Pb are at increased risk for both gout and
renal disease (Shadick et al. 2000; Batuman 1993). Lead is thought to increase serum uric acid
(urate) by decreasing its renal excretion (Emmerson, 1965; Ball and Sorensen, 1969; Emmerson
and Ravenscroft, 1975). As discussed above, research during the past decade indicates that Pb is
nephrotoxic at lower levels than previously recognized. The same is true for uric acid (Johnson
et al., 2003). Therefore, it is possible that one mechanism for Pb-related nephrotoxicity, even at
current lower levels of Pb exposure, is via increasing serum uric acid.
In order to address this question, Weaver et al. (2005a) analyzed data from 803 current
and former Pb workers to determine whether Pb dose was associated with uric acid and whether
previously reported associations between Pb dose and renal outcomes (Weaver et al., 2003a)
were altered after adjustment for uric acid. Outcomes included uric acid, blood urea nitrogen,
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serum creatinine, measured and calculated creatinine clearances, and urinary NAG and RBP.
Mean uric acid, tibia Pb, and blood Pb levels were 4.8 mg/dL (SD 1.2), 37.2 jig/g bone mineral
(SD 40.4), and 32.0 |ig/dL (SD 15.0), respectively. None of the Pb measures (tibia, blood, and
DMSA-chelatable Pb) were associated with uric acid, after adjustment for age, gender, body
mass index, and alcohol use. However, when effect modification by age on these relations was
examined, both blood and tibia Pb were significantly associated in participants in the oldest age
tertile ((3 = 0.0111 [95% CI: 0.003, 0.019] and (3 = 0.0036 [95% CI: 0.0001, 0.007]) for blood
and tibia Pb, respectively). These models were further adjusted for blood pressure and renal
function. Hypertension and renal dysfunction are known to increase uric acid. However, they
are also risks associated with Pb exposure. Therefore, adjustment for these variables in models
of associations between Pb dose and uric acid likely results in overcontrol. On the other hand,
since non-Pb-related factors contribute to both renal dysfunction and elevated blood pressure,
lack of adjustment likely results in residual confounding. Therefore, as expected, associations
between Pb dose and uric acid decreased after adjustment for systolic blood pressure and serum
creatinine, although blood Pb remained borderline significantly associated ((3 = 0.0071 [95% CI:
-0.001, 0.015]). However, when the population was restricted to the oldest tertile of workers
with serum creatinine greater than the median (0.86 mg/dL), likely the highest risk segment of
the population, blood Pb remained significantly associated with uric acid even after adjustment
for systolic blood pressure and serum creatinine ((3 = 0.0156). Next, in models of renal function
in all workers, uric acid was significantly associated with all renal outcomes except NAG.
Finally, in the oldest tertile of workers, after adjustment for uric acid, associations between Pb
dose and NAG were unchanged, but fewer of the previously significant (p < 0.05) associations
noted between Pb dose and the clinical renal outcomes in Weaver et al. (2003a) remained
significant.
Data from the Normative Aging Study indicate that Pb dose, at levels lower than those
known to increase the risk for gout or in the study of Weaver et al. (2005a), is associated with
increased uric acid (Shadick et al., 2000). In 777 participants, mean blood, patella, and tibia
Pb levels were 5.9 |ig/dL, 30.2 jig/g bone mineral, and 20.8 jig/g bone mineral, respectively.
A significant association between patella Pb and uric acid ((3 = 0.007 [95% CI: 0.001, 0.013];
p = 0.02) was found, after adjustment for age, BMI, diastolic blood pressure, alcohol ingestion,
and serum creatinine. Borderline significant associations between tibia (p = 0.06) and blood
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Pb (p = 0.1) and uric acid were also observed. Notably these associations were significant even
after adjustment for blood pressure and renal function, providing further evidence that low-level
Pb exposure increases uric acid.
These data suggest that older workers comprise a susceptible population for increased uric
acid due to occupational Pb exposure. Uric acid may be one mechanism for Pb-related
nephrotoxicity. However, this is not the only mechanism, since in Weaver et al. (2005a), the
association between blood Pb and serum creatinine remained significant even after adjustment
for uric acid. These mechanistic relations have more than just theoretical importance. Clinically
relevant therapies may be possible since EDTA chelation has been reported to improve both
renal function and urate clearance in patients with renal insufficiency and gout, even when
EDTA-chelatable Pb body burdens were low (Lin et al., 2001b).

6.4.7 Susceptible Populations for Lead Nephrotoxicity
6.4.7.1 Chronic Medical Diseases
The general population studies by Tsaih et al. (2004) and Muntner et al. (2003) (discussed
in Section 6.4.4.1) indicate that patient populations with diabetes and hypertension are at
increased risk for adverse renal effects of Pb. Lin et al. (2001a, 2002) indicate that patients with
CRT and gout are also at increased risk. In these settings, Pb appears to acts as a cofactor with
other renal risk factors to cause early onset of renal insufficiency and/or a steeper rate of renal
function decline. It is likely that the presence of larger high risk populations within general
populations is an important factor in the lower Pb dose thresholds noted for the adverse effects of
Pb on the kidney in environmental compared to occupational research.

6.4.7.2 Age
Weaver et al. (2003a, 2005a,b) found older age to be a risk factor for adverse renal effects
in Korean Pb workers. This is consistent with research in general populations (Lindeman et al.,
1985) and is biologically plausible, since most renal risk factors increase with age. Gonick and
Behari (2002) have summarized the data regarding the potential contribution of Pb exposure to
essential hypertension; similar issues may be involved with the renal dysfunction observed in
aging.
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6.4.7.3 Genetic Polymorphisms
Research in the last two decades suggests that several genetic polymorphisms affect Pb
toxicokinetics (i.e., modify the relation between Pb exposure and dose). Of those potentially
relevant to the kidney, data on the gene that encodes for ALAD are the most important in this
regard. The ALAD enzyme is a principal Pb-binding protein; the isozymes in those with the
ALAD2 allele are more electronegative and bind a greater proportion of blood Pb than does the
protein in individuals with the ALAD 1-1 genotype (Bergdahl et al., 1997). Research to date
indicates that individuals with the ALAD2 allele generally have higher blood Pb levels than
those with the ALAD 1-1 genotype, although this may not be the case at lower levels of Pb
exposure (i.e., mean blood Pb levels <10 |ig/dL) (Kelada et al., 2001). Participants with the
ALAD2 allele have been found to have lower bone Pb levels in some studies (Hu et al., 2001;
Kamel et al., 2003); other toxicokinetic differences have also been reported (Fleming et al.,
1998; Hu et al., 2001; Schwartz et al., 1997; Smith et al., 1995). Overall, these data suggest that
tighter binding of Pb by the isozymes of the ALAD2 allele decreases Pb sequestration in bone.
In contrast, data to determine whether the ALAD polymorphism impacts the renal toxicity
of Pb are still quite limited. The only environmentally exposed population in which this has
been addressed is the Normative Aging Study. Wu et al. (2003a) (discussed in detail in Section
6.4.4.1.2) analyzed data to determine whether the ALAD genetic polymorphism modified
associations between Pb dose and uric acid, serum creatinine, and estimated creatinine clearance.
A total of 114 (16%) of the study group were either homozygous or heterozygous for the variant
ALAD2 allele. None of the three outcomes were significantly different by genotype. However,
effect modification by genotype on the association between tibia Pb and serum creatinine was
observed; the P coefficient (and slope) was greater in the group with the variant allele ((3 = 0.002
[SE not provided]; p = 0.03). Effect modification of borderline significance (p < 0.1) for
relationships between patella or tibia Pb and uric acid was observed; this was significant in
participants whose patella Pb levels were above 15 jig/g bone mineral ((3 = 0.016 [SE not
provided]; p = 0.04). Similar to the serum creatinine model, patella Pb was associated with
higher uric acid in those with the variant allele. Genotype did not modify Pb associations in
models of estimated creatinine clearance.
The impact of the ALAD polymorphism on renal outcomes has been studied in four
occupationally-exposed populations to date. The two that assessed both associations and effect
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modification by genotype are discussed here. Weaver et al. (2003b) analyzed data from 798 Pb
workers. A total of 79 (9.9%) participants were heterozygous for the ALAD2 allele (none was
homozygous). After adjustment, participants with the ALAD2 allele had lower mean serum
creatinine and higher calculated creatinine clearance. Effect modification by ALAD on
associations between blood Pb and/or DMSA-chelatable Pb and three of six renal outcomes was
observed. Among those with the ALAD 1-2 genotype, higher Pb measures were associated with
lower BUN and serum creatinine and higher calculated creatinine clearance. Among older
workers (age > median of 40.6 years), ALAD genotype modified associations between Pb dose
and uric acid levels. Higher Pb dose was significantly associated with higher uric acid in
workers with the ALAD 1-1 genotype; associations were in the opposite direction in participants
with the variant ALAD 1-2 genotype (Weaver et al., 2005c).
Ye and colleagues (2003) assessed effect modification by ALAD on associations between
blood Pb with urinary NAG and albumin in a study of 216 Pb workers. Geometric mean blood
Pb was 37.8 |ig/dL in 14 workers with the ALAD 1-2 genotype and 32.4 |ig/dL in workers with
the ALAD 1-1 genotype. After adjustment for age, NAG was borderline statistically higher in
those with the variant allele whose blood Pb levels were >40 |ig/dL. In all Pb workers, after
adjustment for age, gender, smoking, and alcohol ingestion, a statistically significant positive
association between blood Pb and creatinine adjusted NAG was observed in the workers with the
ALAD 1-2 genotype but not in Pb workers with the ALAD 1-1 genotype (the groups were
analyzed separately rather than in an interaction model).
Thus, two of the three studies reported steeper slopes for one or more associations
between Pb dose and adverse renal function in participants with the ALAD2 allele compared to
those with the ALAD 1-1 genotype, which suggests that the variant ALAD gene confers
additional risk for adverse renal outcomes in Pb-exposed populations. If the associations of
Weaver et al., (2003b) represent Pb-induced hyperfiltration, their results could be consistent with
increased risk from the variant allele as well. Ultimately, analysis of longitudinal data in the
Korean Pb worker population will be needed to understand these complex relationships.
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6.4.8 Confounding of the Renal Effects of Lead by Other Potential
Risk Factors
Studies selected for discussion in Section 6.4 above have generally controlled for at least
the most basic risk factors known to affect renal function, such as age, gender, and body mass
index (or weight and height separately). Some have controlled for many other potentially
important risk factors. In addition, exposure to other nephrotoxicants must be considered.
Notably, although these are listed under confounders, some may be effect modifiers as well.

6.4.8.1 Cadmium
Similar to Pb, cadmium (Cd) is an ubiquitous nephrotoxi cant that accumulates in the
body. Environmental exposure to Cd in the United States occurs primarily through food and
smoking (Agency for Toxic Substances and Disease Registry, 1993). Cadmium in food is a
result of soil pollution from a variety of human activities such as phosphate fertilizer use,
industrial releases from smelting, and fuel combustion. An analysis of NHANES III data,
collected in a representative sample of the U.S. population from 1988-1994, indicates that mean
urinary Cd is 0.48 |ig/g creatinine, and 97.7% of the population has a level <2.0 |ig/g creatinine
(Paschal et al., 2000). Also similar to Pb, Cd causes proximal tubule pathology and is a known
risk factor for chronic renal insufficiency (CRI).
Existing data indicate that Cd, at exposure levels common in the United States, confounds
associations between Pb exposure and at least one renal outcome, NAG. Roels et al. (1994)
reported higher mean NAG in their Pb-exposed group; however, NAG was correlated with
urinary Cd but not blood or tibia Pb, despite mean urinary Cd being only 1.04 and 0.53 |ig/g
creatinine in workers and controls, respectively. Cardenas et al. (1993) reported a similar
finding. Bernard et al. (1995a) found an association between urinary Cd and the NAG-B
isoenzyme (released with breakdown of proximal tubular cells) in 49 Cd workers and 20 age-
matched controls. In multiple linear regression, urinary Cd, but not Pb, was associated with
NAG-B after adjustment for age. The association was significant even in the 44 participants
with Cd levels <2 jig/g creatinine. However, NAG-A (released by exocytosis) was correlated
with urinary Pb (the only Pb measure), but not Cd. Roels et al. (1995) reviewed data pertinent to
the potential for Cd confounding of associations between Pb and NAG. In more recent work,
Weaver et al. (2003a) measured urinary Cd in a subset of 191 of the 803 workers in their study
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(mean urinary Cd was 1.1 |ig/g creatinine). Higher urinary Cd levels were associated with
higher NAG. Of the Pb measures obtained, only tibia Pb was significantly associated with NAG
in the Cd subset. When urinary Cd and tibia Pb were entered as covariates in the same model,
both remained associated with NAG (p < 0.05). However, in comparing the effects, a 0.5 jig/g
creatinine increase in Cd had the same effect on NAG as a 66.9 jig/g bone mineral increase in
tibia Pb. When compared by ranges of exposure in this population, environmental level Cd dose
had a larger impact on NAG than did occupational Pb dose.
Cadmium exposure may well confound relations between Pb exposure and other renal
outcomes as well, but available data are too limited to draw firm conclusions. Positive
associations between urinary Cd, which is thought to be the best measure of cumulative Cd
exposure in the absence of Cd-related renal damage, and low molecular weight (LMW)
proteinuria are well established in the occupational setting. LMW proteinuria, most commonly
assessed by (32-microglobulin, is generally progressive at Cd levels >1500 jig/g creatinine in
workers with substantial body burdens (one or more historical urinary Cd >20 |ig/g creatinine)
but may also be progressive at lower levels (Roels et al., 1997; Bernard, 2004). More
importantly, clinical renal function also declines as evidenced by decreasing GFR in Cd-exposed
workers followed longitudinally after removal from exposure due to LMW proteinuria (Roels
etal., 1989; 1997).
In contrast to the clear evidence that Cd is a renal toxicant at occupational levels of
exposure, the renal risk from lower level Cd exposure remains uncertain. Most studies of
environmental Cd exposure are cross-sectional and have assessed EBE markers, rather than
clinical renal outcomes (Alfven et al., 2002; Jarup et al., 2000; Noonan et al., 2002; Olsson et al.,
2002). The Cadmibel study, a general population study of exposed residents from both
Cd-polluted and unpolluted areas (discussed in Section 6.4.4.1.1), found correlations between
urinary Cd and several urinary EBE markers (NAG, RBP, p2-microglobulin, calcium, and amino
acids) (Buchet et al., 1990). In those models, after adjustment for urinary Cd and other
covariates, blood Pb was significant in models of p2-microglobulin and amino acids but not
NAG. However, in this same population, blood Pb was inversely associated with creatinine
clearance, whereas urinary and blood Cd were not (Staessen et al., 1992). A 5-year follow-up
was conducted to determine the significance of the EBE abnormalities (Hotz et al., 1999). In this
study, models of renal function (two dichotomized outcomes: a 20% decline in creatinine
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clearance and a 20% increase in albumin excretion) in relation to quartiles of urinary Cd and the
EBE markers at baseline were analyzed by likelihood ratios. Baseline variables did not predict
adverse renal outcomes. However, 25% of the original population was lost to follow-up;
available data indicated that their baseline renal function was worse than those who participated
in the follow-up study. This may have biased the study towards the null.
Three recent publications suggest that low-level Cd exposure is associated with adverse
clinical renal outcomes. Elevated urine Cd levels were associated with decreased calculated
creatinine clearance and with prevalent microalbuminuria after adjustment for age, sex, race,
smoking, and use of diuretics in an analysis of 16,094 participants in the NHANES III study
(Young et al., 2004). Also, Hellstrom et al. (2001) reported increased rates of renal dialysis
and transplantation in residents of Cd-polluted areas in Sweden. Compared to the "no exposure
group" (domicile >10 km from a battery plant), age-standardized rate ratios were 1.4 (95% CI:
0.8, 2.0) in the low-exposure group (domicile 2 to 10 km) and 1.9 (95% CI: 1.3, 2.5) in the
moderate-exposure group (domicile <2 km). Exposure categorization was based on
environmental monitoring in the study areas. Cadmium dose was not directly measured,
although occupationally exposed participants were considered in a separate group. The third
study, by Akesson et al. (2005), also assessed Pb exposure as a covariate, an important approach
given the Cadmibel results (Staessen et al., 1992). Blood and urinary Cd were associated with
worse GFR and creatinine clearance. The association for blood Cd and decreased creatinine
clearance remained statistically significant even in non-smokers, suggesting that a public health
remedy, in addition to smoking cessation, may be of value.
In conclusion, Cd clearly confounds associations between Pb dose and NAG. Given the
similarities in both nephrotoxicants, Cd may confound and/or modify associations between Pb
and other renal outcomes. However, data regarding the concentration-response relationship
between environmental Cd and the kidney are too limited to assess the potential for this at
present. Future studies assessing both Pb and Cd are needed.

6.4.9 Summary of the Epidemiologic Evidence for the Renal Effects of Lead
During the past two decades, the quality of research on the renal impacts of Pb exposure
has advanced dramatically. As a result, a much more accurate assessment of the adverse renal
impact of Pb exposure can now be made. General population studies are the most important
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advance in this regard. These studies provide strong evidence that renal effects occur at much

lower blood Pb levels than previously recognized. These effects are clinically relevant in U.S.

subpopulations who continue to have higher Pb exposure than the general population. At levels

of exposure in the general U.S. population overall, Pb combined with other risk factors, such as

diabetes, hypertension, or chronic renal insufficiency from non-Pb related causes, can result in

clinically relevant effects. Notably, the size of such susceptible populations is increasing in the

United States due to obesity.

• The majority of studies in general adult and patient populations published during the past
two decades have observed associations between Pb dose and worse renal function.
Other explanations, such as residual confounding or reverse causality, are less likely.
The renal effects of Pb on children are difficult to assess, as most of these studies only
measured early biological effect markers which have unknown clinical significance.

• The magnitude of the effect of Pb on renal function ranged from 0.2 to -1.8 mL/min
change in creatinine clearance per |ig/dL increase in blood Pb in general population
studies. The size of the effect was relatively consistent across the studies, although only
five provided data useful for this determination (three were at different time points in the
Normative Aging Study population) and a form of publication bias may be present in
studies that provided no data and only reported that associations were not significant.
One patient population (individuals with CRI) study reported a similar effect of blood Pb
longitudinally on yearly decline in GFR.

• The cumulative effect of higher blood Pb levels from past exposure may be a factor in
nephrotoxicity observed at current blood Pb levels. However, one study found
associations between blood Pb and concurrent serum creatinine in participants whose
peak blood Pb levels were < 10 |ig/dL.

• The threshold for Pb-related nephrotoxicity cannot be determined based on current data.
However, associations with clinically relevant renal outcomes have been observed in
populations with mean blood Pb levels as low as 2.2 |ig/dL.

• Research in the occupational setting is far less consistent. However, a notable finding
from several of these studies is the observation of inverse associations (higher Pb dose
with lower BUN, serum creatinine, and/or higher creatinine clearance). This may
indicate Pb-related hyperfiltration and may have mechanistic implications.
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6.5 CARDIOVASCULAR EFFECTS OF LEAD
6.5.1 Summary of Key Findings of the Cardiovascular Effects of Lead from
the 1986 Lead AQCD and Addendum, and 1990 Supplement
The greater part of the evidence reviewed up to 1990 included analyses of the largest
datasets available at the time: (1) the National Health and Nutrition Evaluation Survey II
(NHANES II), studying the U.S. population between 1976 and 1980; and (2) the British
Regional Heart Study (BRHS), studying men aged 40-59 years from 24 British towns. Analyses
of the Welsh Heart Programme, a regional Welsh study, and the Caerphilly Collaborative Heart
Disease Study, a cohort study of men aged 45-59 years living in one town in Wales, as well as
smaller population and occupational exposure studies in the United States, Canada, and Europe,
provided further supporting evidence. These studies set enduring design and analysis standards
by example for evaluating cardiovascular effects associated with blood Pb levels in samples from
diverse populations.
In general, the reviewed studies used multiple linear regression modeling of blood
pressure and multiple logistic regression modeling of hypertension, cardiovascular mortality, and
other cardiovascular disease, allowing adjustment of the blood Pb effect on outcome by other
factors known or suspected to be related to the exposure and outcome under study. The most
commonly considered potential confounding factors were age, body mass index (BMI), alcohol
use, and cigarette smoking.
These studies were almost exclusively cross-sectional, measuring cardiovascular outcome,
blood Pb, and control variables once, although one Canadian occupational study and one Danish
birth-year cohort study used a longitudinal design. Some studies presented analyses stratified by
sex or age, by both sex and age, or by race. Other analyses only reported results for one
particular stratum. Separate analyses of datasets partitioned by stratified variables always reduce
sample size available for statistical models and may thereby reduce power to detect real effects.
Evaluated as a whole, the earlier available blood pressure studies supported a small but
significant association between increasing blood Pb concentrations and increasing blood pressure
in study groups. The effect was more consistent across studies in middle-aged men than in other
groups, ranging from a 1.5 to 3.0 mm Hg increase in systolic blood pressure for each doubling of
blood Pb from the mean blood Pb level, and from a 1.0 to 2.0 mm Hg increase in diastolic blood
pressure for each blood Pb doubling, across a wide range of blood Pb concentrations down to
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7 |ig/dL. Most studies using multiple regression analyses stratified by sex did not find
significant associations between blood pressure and blood Pb in females, though one reanalysis
of the NHANES II dataset did report a statistically significant relationship between diastolic
blood pressure and blood Pb in women aged 20 to 74 years. In studies reporting the use of
different blood Pb-blood pressure concentration-response relationships, log blood-Pb terms had
lower probability values than linear blood-Pb terms, suggesting that increases in blood pressure
with fixed increases in blood Pb might be greater at lower than higher blood Pb concentrations.
Three studies of groups with occupational exposure reported mixed results. One study
found significant excess mortality due to cardiovascular disease during the 1946-1965 period in a
case-control study in the United Kingdom, but not during 1966-1985. A study of U.S. battery
and Pb production workers from 1947-1980 found significant excess mortality due to "other
hypertensive disease" (codes 444-447 in the ICD 1955 classification system), but not due to
hypertensive diseases outside those classifications. No excess mortality due to hypertension was
found in a study of U.S. smelter workers between 1940 and 1965.
The BRHS study did not find significant associations between blood Pb and ischemic
heart disease and stroke, though low power to detect such an effect should be noted. However,
electrocardiogram abnormalities associated with left ventricular hypertrophy were found to be
related to blood Pb in a subset of the NHANES II data, confirming an earlier study finding
significant associations between blood Pb and ischemic changes in Pb workers.
Noninvasive measurement of bone Pb concentration using XRF techniques was still
maturing during the literature review period covered by the 1986 Lead AQCD/Addendum and
1990 Supplement. No cardiovascular studies had yet been reported using bone Pb as a marker
for Pb exposure. The previous Lead AQCD concluded that there was a small but statistically
significant relationship between blood-Pb level and adverse cardiovascular outcome. Future
research needs noted were large sample sizes, identification of susceptible populations, more
precise quantitative estimation of effect size, and better definition of the dose-response
relationship.
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6.5.2 Effects of Lead on Blood Pressure and Hypertension
6.5.2.1 Introduction
Blood Pb concentration has remained the most widely used Pb exposure index in blood
pressure/hypertension epidemiologic studies from 1990 to present. Obtaining the sample is
relatively noninvasive and quick, pertinent measurement techniques are well standardized and
inexpensive, there is wide access to external quality assurance programs, and existing regulation
and medical decision-making are based on blood Pb levels. If exogenous Pb exposure were the
only determinant for blood Pb concentration, it could be fair to state that a single blood Pb
measurement reflects exposure to Pb during the 30-90 day period preceding the measurement.
However, blood Pb concentration represents a combination of recent exposure to external
sources and the influence of internal sources, principally bone Pb. As detailed in Chapter 4,
bone is a long-term storage depot for much of the Pb absorbed by the body from external sources
and, by weight, can represent over 95% of the total Pb body burden in middle-aged persons,
especially where current external exposures are low. Bone Pb has residence times of years to
decades. Also, bones constantly absorb Pb from and release Pb back into the circulatory system.
Consequently, blood Pb concentration is not only determined by current and recent past external
Pb exposure but is also influenced by existing bone Pb concentration to a degree determined by
current external exposure, accumulated past exposure Pb stored in bones, and the physiological
state of the bones due to aging, disease, pregnancy, and lactation, among other factors. Studies
using only blood Pb concentration as an exposure index cannot determine the relative
contributions of current exogenous exposure and endogenous exposure to blood Pb. Thus, they
are unable to assess what part of measured blood Pb effect on the circulatory system is due to
possibly higher long duration past exposure and what part is due to the possibly immediate toxic
effects of currently circulating Pb. They are, instead, assessing a combined effect of past and
present exposure in a proportion that will differ among subjects according to their past and
present exposure, health history, and age.
The newly developed in vivo technique of XRF measurement of bone Pb concentration
has been used in a handful of studies to better assess the role of past Pb exposure on blood
pressure and hypertension in essentially cross-sectional studies. Bone Pb concentration provides
a record of cumulative past exposure due to the long residence times of Pb in bones, though the
specific temporal pattern of past exposure cannot be readily determined from the measurement.
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Primarily cortical bones (such as tibia) have residence times measured in decades, whereas
primarily trabecular bones (such as calcaneus and patella) have Pb residence times measured in
years to decades, reflecting different metabolic rates of the two bone types. As there is continual
interchange of Pb in bone and Pb in blood, studies combining the measurement and modeling of
both bone and blood Pb have the best chance of dissecting out the roles of past and present Pb
exposure on blood pressure and hypertension.
Elevated blood pressure can be evaluated as a continuous measure (mm Hg) or as a
dichotomized measure (hypertension). The definition of hypertension involves a categorical cut
point of mm Hg above which one is hypertensive and below normotensive. Kannel (2000a,b)
notes that this number has dropped over time for systolic/diastolic pressure and further notes a
continuous graded influence of blood pressure on health even within what is regarded as the
normotensive range. The hypertension definition is to some extent arbitrary, as the cut point has
changed over time. However defined for any given study, regardless of medical definition for
the year of the study, hypertension classification offers a different perspective than blood
pressure per se. Hypertension has a different clinical relevance than blood pressure changes
themselves. The disease condition as an outcome and a change in mm Hg in relation to exposure
both offer the opportunity for insight into the clinical relevance of the relationships. Biomarkers
like bone Pb and blood Pb also help to distinguish acute and chronic exposure effects.
Blood pressure is an inherently variable measure. Even when measured with indwelling
catheters, blood pressure varies on a minute to minute interval in the same individual. Extrinsic
sources of blood pressure variability include measurement technique, the tester, and conditions
under which the measurements are taken. All the sources of measurement error are additive, but
the expected total error will be symmetrically distributed about some true blood pressure, i.e.,
unbiased. Under conditions where the size of the expected Pb effect is in the same range as the
total measurement error, large studies with high power are required for detecting real effects
where they exist. These factors favor studies with large numbers of subjects. Using blood Pb as
a surrogate for brain Pb biases the blood Pb regression coefficient towards zero. This is an
example of classical measurement error. Smaller blood pressure studies may fail to detect Pb
effects simply because of low power. As noted in the 1986 AQCD/Addendum and the 1990
Supplement, stratification of data sets always reduces power.
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The growing field of toxicogenetics now includes Pb exposure epidemiology. The several
studies combining subject evaluation of polymorphisms of genes thought to play a role in either
the origin of cardiovascular disease, the toxicokinetics of Pb, or both are also reviewed here.

6.5.2.2 Blood Pressure and Hypertension Studies Using Blood Lead as Exposure Index
Table 6-2 lists studies showing estimates of the relationship between systolic blood
pressure and blood Pb level. The table focuses on the key studies with low blood Pb in the
United States to include studies of the general population (NHANES), the Boston NAS, and the
international meta-analysis by Nawrot et al. (2002). Studies were included if the population was
not all occupationally exposed or limited to women during pregnancy, or children. Effects were
included only if they were based on the entire study group or secondly, the subgroup had more
than 500 people. Other studies are discussed in the text and presented in Annex Table AX6-5.1.

6.5.2.2.1 NHANES Studies
NHANES contributed the largest datasets analyzed in this review. As the surveys are also
representative of the U.S. population, their results may be more readily applied to the general
U.S. population than smaller cohort or occupational studies. The several papers using this
dataset sometimes come to different conclusions, depending on the statistical techniques used in
the analyses, including logarithmic or linear specification of the Pb variable, stratification of
analyses according to sex or ethnic groups or use of interaction terms to define these groups,
use of survey-design corrected models, choice of covariates in the models, and different age
ranges analyzed.

NHANES II (1976-1980)
In one NHANES Il-based study, males and females (number unreported but less than
9,000 combined) aged 20 to 74 years were studied with separate stepwise multiple regression
models adjusted for sampling design (Schwartz, 1991). Mean blood Pb levels and ranges were
not reported. Covariates common to both male and female models were age and age2, BMI, race,
family history, cholesterol, zinc, tricep fold, and natural log Pb. Models for men also included
height and cigarette smoking. Natural log blood Pb was significantly associated with diastolic
blood pressure (systolic not reported) in males, with a 2.03 mm Hg diastolic (95% CI: 0.67,
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Table 6-2. Summary of Studies with Quantitative Relationships of
Systolic Blood Pressure and Blood Lead
Reference
Vupputuri et al.
(2003)

Den Hond et al.
(2002)

Nash et al.
(2003)
Sorel et al.
(1991)

Cheng et al.
(2001)
Proctor et al.
(1996)
Nawrot et al.
(2002)
Study Location and Gender
NHANES* III
White males
White females
Black males
Black females
NHANES III 1988-94
White males
White females
Black males
Black females
NHANES II
Females
NHANES II
Males
NHANES II
Females
Boston Normative Aging Study
Males (about 97% White)
Boston Normative Aging Study
Males
3 1 U.S. and European Studies
(includes occupationally exposed)
n
5,360
5,188
2,104
2,300

4,685
5,138
1,761
2,197
1,786
2,044
2,056
519
798
>58,490
Blood Lead
Arithmetic Mean
(25th and 75th
Percentiles)
4.4(1.0,4.9)
3.0(0.5,2.9)
5.4(1.2,6.0)
3.4(1.0,4.0)

3.6(2.3,5.3)
2.1 (1.3,3.4)
4.2 (2.7, 6.5)
2.3 (1.4,3.9)
2.9 (range 0.5-3 1.1)
Black: 20.1
White: 16.8
Black: 13.2
White: 12.1
5.9 (3.4, 7.4)
6.5(3.8,8.1)
Meta-analysis
Estimated Slope
(95% CI), mm Hg per
Change in Blood Lead
from 5 to 10 jig/dL
0.43 (-0.36, 1.26)
0.52 (-0.74, 1.77)
1.24(0.29,2.18)
2.35(0.71,4.00)

0.3 (-0.2, 0.7)
0.1 (-0.4, 0.5)
0.9 (0.04, 1.8)
1.2 (0.4, 2.0)
1.60(0.05,3.15)
0.40 (-0.15, 0.95)
0.20 (-1.40, 1.00)
-0.16 (-1.67, 1.35)
0.59 (-0.76, 1.87)
1.0 (0.5, 1.4)
NHANES—United States population sample.
3.39) increase for every doubling of blood Pb and, for females, a 1.14 mm Hg increase (95% CI:
0.13, 2.08). Interactions between blood Pb and sex and between blood Pb and race in a
combined model were insignificant (not shown). The conclusion from these interaction terms is
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that the association between blood Pb and diastolic blood pressure was not significantly different
between men and women or between races. Stepwise modeling may inflate statistical Type I
error.
The other NHANES II-based study focused on Black-White differences in blood pressure
related to blood Pb (Sorel et al., 1991). There were 473 Blacks and 3,627 Whites in the study,
each divided nearly evenly by sex, aged 18 to 74 years. Blood Pb means and ranges were not
given. As is usual in U.S.-based studies, race/ethnicity was based on self-report. Survey design-
adjusted multiple regression models were stratified on sex and included age, BMI, and linear
blood Pb as covariates. Effects of race and poverty index were assessed by including their terms
in models with and without blood Pb and determining change in race or poverty coefficients by
comparing confidence intervals. Each 1 |ig/dL increase in linear blood Pb significantly predicted
increased systolic blood pressure for both males (0.13 mm Hg/|ig/dL) and females (0.08 mm
Hg/jig/dL), but not diastolic blood pressure. The differences in Black and White (race variable)
blood pressure coefficients did not significantly change when Pb was in or out of the model,
either for subjects below the poverty index or above the poverty index. Race did not appear to
significantly modify the relationship between blood Pb and systolic blood pressure. There were
reporting inconsistencies in the female-stratified models, in which the coefficients and 95% CI
did not correspond.

NHANES III (1988-1994)
A study using the NHANES III dataset from all adults 20 years of age and up examined
the effect of natural log blood Pb on systolic and diastolic blood pressure (Den Hond et al.,
2002). Multiple regression analyses for each blood pressure measurement were stratified by sex
and race, yielding four models for each blood pressure measurement. The mean Pb blood levels
were 3.6 |ig/dL for White males (n = 4,685), 2.1 |ig/dL for White females (n = 5,138), 4.2 |ig/dL
for Black males (n = 1,761), and 2.3 |ig/dL for Black females (n = 2,197). The overall blood Pb
range was < 0.8 to > 20.0 |ig/dL. One group of covariates (age, age-squared, BMI, hematocrit,
smoking, alcohol consumption, and an indicator variable for use of antihypertensive
medications) were first entered as a block regardless of significance in each model. Next,
another group of variables (coffee consumption, dietary calcium, dietary sodium/potassium ratio,
total serum protein, total serum calcium, diabetes, and poverty index) was entered stepwise into
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the model without Pb, and the variable was retained only if it was statistically significant
(p < 0.05). Then log-transformed blood Pb was forced into each model. The model building
procedure resulted in eight distinct models, each with their own unique mix of covariates.
Adjustment of results by survey sample weights and design was not reported. Only Blacks had
significant Pb-systolic blood pressure associations; each doubling in blood Pb was associated
with a 0.90 mm Hg (95% CI: 0.04, 1.8) and 1.20 mm Hg (95% CI: 0.4, 2.0) increase in males
and females respectively. The association of Pb-diastolic blood pressure was also significant for
Black females (0.50 mm Hg [95% CI: 0.01, 1.1]). Interestingly, increasing blood Pb was
associated with significantly decreased diastolic blood pressure in White males (-0.6 mm Hg
[95% CI: -0.9, -0.3]). The authors did not comment on their finding that the significant total
serum calcium covariate in these two groups had opposite signs too (White male serum calcium
P = 6.50 mm Hg/mmol/L, Black female serum calcium P = -5.58 mm Hg/mmol/L). Though the
authors offered no formal test of the difference between the two serum calcium coefficients,
since both were significantly different than the null hypothesis coefficient of 0 and different in
sign, it could be concluded that those coefficients were significantly different between the two
groups. As the authors do not present the serum calcium coefficients before forcing Pb into the
models, it is not certain that blood Pb in the model was associated with the significant sign
difference of the calcium coefficients or if the calcium coefficients had opposite signs between
the two groups without Pb in the model. As each model had a different set of covariates, the
presence or absence of one of the other covariates could have produced the same results. Still,
this pattern of results may indicate significant confounding between serum calcium and blood Pb
associations with blood pressure. Though the study suggested differences between Blacks and
Whites in response to Pb, no statistical tests were performed of differences between Pb
coefficients based on race. In addition, the Black-White effect differences associated with blood
Pb may be due to possible confounding in some or all of the models.
Limiting the study sample from NHANES III to women aged 40 to 59 years, another
group of researchers addressed the relationship between blood Pb and both blood pressure
(n = 1,786) and hypertension (n = 2,165) over a blood-Pb range of 0.5 to 31.1 |ig/dL (mean
2.9 |ig/dL) (Nash et al., 2003). Blood pressure models excluded women who reported being
under treatment for hypertension. Separate blood pressure multiple regression models were
presented for diastolic and systolic blood pressure, each with and without stratification for
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dichotomous premenopausal/postmenopausal status. One block of covariates was entered
without regard to statistical significance (age, race/ethnicity, BMI, and serum creatinine).
Another block of covariates (education, poverty income ratio, alcohol use, and cigarette smoking
status) was entered second but only retained if variables were significantly associated with blood
pressure. Finally, linear blood Pb was forced in last. Logistic regression models for
hypertension used the same covariate entry scheme with and without stratification on the
menopause variable, but using a blood Pb quartile exposure variable. Despite the stated
procedure for covariate selection, all models used the same set of covariates: linear (or quartile)
Pb, age, race/ethnicity, alcohol use, cigarette smoking status, BMI, and serum creatinine.
All models were adjusted for survey weights and design. Linear Pb was significantly associated
with systolic blood pressure only in the entire study sample; each 1 |ig/dL increase in blood Pb
was associated with a 0.32 mm Hg (95% CI: 0.01, 0.63) increase in blood pressure.
No associations were observed in the menopause-stratified analyses. Linear Pb was also
significantly associated with diastolic blood pressure in the entire study sample (0.25 mm Hg
[95% CI: 0.07, 0.43]). Odd ratios of diastolic hypertension (>90 mm Hg) in logistic regression
models were significantly related to blood Pb, with an odds ratio of 4.26 (95% CI: 1.36, 12.99)
comparing the 1st quartile blood Pb group (0.5-1.6 |ig/dL) to the 4th quartile blood Pb group
(4.0-31.1 |ig/dL) in all women not taking antihypertensive medications. Further stratification
produced occasional significant odds ratios for either diastolic or systolic hypertension. There
were some differences in table and text reporting of results and an inconsistency between the
SE and the p-values.
Another study using the NHANES III database was notable for its formal testing of race
and sex differences in Pb effect by interactions terms (Vupputuri et al., 2003). The study used
5,360 White men (mean blood Pb 4.4 |ig/dL), 2,104 Black men (mean blood Pb 5.4 |ig/dL),
5,188 White women (mean blood Pb 3.0 |ig/dL), and 2,300 Black women (mean blood Pb
3.4 |ig/dL). Blood Pb ranges were not given. Multiple linear and logistic regression models of
blood pressure and hypertension (systolic > 140 mm Hg, diastolic >90 mm Hg, and/or taking
antihypertensive medication), respectively, were adjusted for age, high school education, BMI,
alcohol, leisure-time physical activity, and dietary intake of sodium, potassium, and total energy.
The models used linear blood Pb, except for one set of hypertension models with a cut point for
"high" Pb exposure at >5 |ig/dL. Subjects taking antihypertensive medication (n = 2,496) were
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not included in linear regression models of blood pressure. Neither age nor blood Pb range were
reported, nor was the technique of selecting and entering covariates in multiple regression
models. Only coefficients for linear Pb effect for each model were reported. Significant
interactions in multivariate models were found between Pb and race and between Pb and sex,
though these analyses were not shown. Only Black men and women had significant linear
Pb-blood pressure effects in adjusted systolic (0.25 mm Hg [95% CI: 0.06, 0.44] for Black men
and 0.47 mm Hg [95% CI: 0.14, 0.80] for Black women with each 1 |ig/dL increase in blood Pb)
and diastolic blood pressure (0.19 mm Hg [95% CI: 0.02, 0.36] for Black men and 0.32 mm Hg
[95% CI: 0.11, 0.54] for Black women). Linear blood Pb association with hypertension was
significant only in women. The odds ratios were 1.09 (95% CI: 1.04, 1.13) for White women
and 1.10 (95% CI: 1.06, 1.16) for Black women for each 1 |ig/dL increase in blood Pb.
The authors presented insufficient detail to evaluate this pattern of results.

6.5.2.2.2 Other U.S. Cohort Studies
The Boston-based Normative Aging Study, part of a longitudinal study of male veterans,
examined the effects of blood Pb on blood pressure in 798 men, aged 45-93 years old, with blood
Pb between 0.5 and 35.0 |ig/dL (Proctor et al., 1996). Using multiple regression modeling with
forced entry of natural log Pb and other covariates (age, age2, BMI, dietary calcium, exercise,
smoking, alcohol, heart rate, and hematocrit), the authors found a significant increase of only
diastolic blood pressure (0.83 mm Hg [95% CI: 0.08, 1.52]) for each doubling of blood Pb.
Though the relationship between blood Pb and systolic blood pressure was positive, it was not
significant. Nearly half the blood Pb measures were derived from frozen red blood cells
collected previously (up to several years earlier) and corrected for hematocrit determined at the
time blood pressure was measured. Possible errors in correction of these samples and the
non-contemporaneous nature of the resulting blood Pb concentrations may have compromised
the results.
Cheng et al. (2001), using the same Normative Aging Study data and stepwise multiple
regression, found a near-zero association between systolic blood pressure and linear blood Pb
(-0.03 mm Hg for each |ig/dL increase in blood Pb) in 519 men aged 48 to 93. The subjects
selected for this analysis were all free of hypertension (systolic > 160 mm Hg or diastolic
> 95 mm Hg). Differences in subject selection procedures and modeling techniques may have
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accounted for the different results between Cheng et al. and Proctor et al. They also reported on
incidence of hypertension developing between 1991 and 1997 using Cox proportional hazards
models. Controlling for age, age2, BMI, and family history of hypertension, linear blood Pb was
not significantly associated with risk of developing hypertension (systolic > 140 mm Hg or
diastolic > 90 mm Hg) in subjects normotensive at the start of the period (rate ratio of 0.98 [95%
CI: 0.91, 1.06]) for each 1 |ig/dL increase in blood Pb.
Gerr et al. (2002) similarly reported near-zero linear blood Pb effects on blood pressure on
a combined group of 19-29 year old males and females (n = 502), half of whom had lived around
active Pb smelters as children, using forced entry of all covariates. Mean blood Pb was
2.2 |ig/dL with a range from <1 to >7 |ig/dL. Among the covariates forced into the model was
tibia Pb concentration, expected to be significantly correlated with blood Pb. This may have
reduced or confounded the effects of blood Pb.
Korrick et al. (1999) examined linear and natural log blood Pb, mean 3 |ig/dL (range <1 to
14 |ig/dL), effects on hypertension, defined as self-reported or physician hypertension diagnosis
(i.e., systolic or diastolic blood pressure > 140/90 mm Hg), in 284 middle-aged women from the
Nurse Health Study based in Boston. The association of hypertension and blood Pb was not
significant. The study had low power (n = 284).
Rothenberg et al. (1999) tested a group of 1,527 women, aged 15 to 42 years, in their third
trimester of pregnancy, with blood Pb ranging from 0.5 to 40.4 |ig/dL. They stratified testing
into immigrant (n = 1,188) and nonimmigrant (n = 439) groups. They used forced entry of all
covariates in multiple regression models, including natural log Pb, age, BMI, coffee, iron
supplement, and job stress, and found Pb-related significant increases in systolic (1.18 mm Hg
[95% CI: 0.45, 1.91] for each doubling of blood Pb) and diastolic (1.02 mm Hg [95% CI: 0.37,
1.34]) blood pressure only in immigrants. The small size of the nonimmigrant group may have
reduced power to detect significant effects. In a follow-up study of 668 women returning for
postpartum testing (Rothenberg, et al., 2002a), using multiple regression models with forced
entry of natural log blood Pb, tibia and calcaneus Pb, age, BMI, parity, smoking, immigrant
status, and education, the authors found significant decreases in systolic (-1.05 mm Hg [95% CI:
-1.96, -0.14]) and diastolic (-1.16 mm Hg [95% CI: -1.98, -0.35]) blood pressure associated
with doubling in blood Pb in the postpartum women. This subgroup of women had no
significant blood Pb effects in the third trimester. Although the covariate pattern was different
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from the larger prenatal study (Rothenberg et al., 1999), thorough testing of possible
confounding, especially with the bone Pb measures, revealed no significant change in blood Pb
effects. This study finding is similar to that reported by Den Hond et al. (2002) for White males.
No significant effect of blood Pb on prenatal or postpartum hypertension (> 140/90 mm Hg) was
found.
Morris et al. (1990) recruited a group of 105 women and 145 men, aged 18-80 years, from
a clinic specializing in nondrug hypertension treatment. Blood Pb ranged from 5 to 40.5 |ig/dL.
Multiple regression was performed with forced entry of natural log Pb, age, BMI, dietary
calcium, "other nutrients," serum ionized calcium, and erythrocyte protoporphyrin. Only men
were found to have blood Pb-related significant increases in systolic (3.17 mm Hg [95% CI:
-2.13, 8.48] for each doubling of blood Pb) and diastolic (1.32 mm Hg [95% CI: -2.12, 4.75])
blood pressure. Small study size limits conclusions based on nonsignificant findings in women.
Dietary calcium is associated with reduced blood Pb in many studies and could be considered a
confounder with blood Pb. Erythrocyte protoporphyrin is a biomarker of Pb exposure and
correlates with blood Pb over at least part of the blood range in study subjects. There were at
least two variables collinear with blood Pb, a high proportion of covariates to subjects, and
possible subject selection bias.

6.5.2.2.3 European Cohort Studies
The Glostrup Population Study (Copenhagen) evaluted data for 1,009 men and women
(all born in 1936) longitudinally studied from 1976 to 1987 (M011er and Kristensen, 1992).
Blood Pb levels ranged from 2 to 62 |ig/dL, depending on the year and sex stratum studied, with
mean concentration dropping by -40% over the study period. Multiple regression analyses were
used, with forced entry of natural log Pb, BMI, tobacco use, and physical activity. Strongest
associations between a doubling of blood Pb and blood pressure were found early in the study
period. In 1976, a doubling of blood Pb was associated with a 3.42 mm Hg (95% CI: 1.25, 5.58)
increase in systolic blood pressure and a 2.95 mm Hg (95% CI: 1.08, 4.83) increase in diastolic
blood pressure in women. For men in 1981, a doubling of blood Pb was associated with an
increase of 1.89 mm Hg (95% CI: 0.00, 3.78) in systolic blood pressure and 1.14 mm Hg (95%
CI: -0.37, 2.65) in diastolic blood pressure. No formal longitudinal analyses were performed,
only analyses stratified by year and sex and analyses relating change in Pb and other covariates
6-125
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to change in blood pressure from one study period to the next. As the relative risk of mortality
was associated with increasing blood Pb over the study period (see below), the observed general
reduction in the Pb-associated blood pressure increase over the study period may have been
related to Pb-associated mortality.
The Europe New Risk Factor Project in Rome collected data from 1,319 males aged 55 to
75 years with blood Pb between 4.0 and 44.2 |ig/dL (Menditto et al., 1994). They reported
significantly increased systolic (4.71 mm Hg [95% CI: 2.81, 6.61]) and diastolic (1.25 mm Hg
[95% CI: 0.33, 2.16]) blood pressure associated with a doubling of blood Pb.
The Cadmibel studies from Belgium specifically selected part of their study group from
those living near nonferrous smelters. Staessen et al. (1993) reported on 827 men and
821 women, aged 20 to 88 years, with blood Pb ranging from 2.7 to 84.9 |ig/dL for men and
1.3 to 42.4 |ig/dL for women. They forced natural log blood Pb into stepwise multiple
regression models stratified by sex. Covariates available for selection were age, age2, BMI,
pulse rate, log gamma-glutamlytranspeptidase, serum total calcium, log serum creatinine, urinary
potassium, smoking, alcohol, contraceptive use, and menopause. Near-zero nonsignificant
relationships were found between blood Pb and blood pressure for systolic blood pressure for
women and diastolic blood pressure for men and women. They reported a significant decrease in
men's systolic blood pressure with increasing blood Pb (-1.1 mm Hg for a doubling of blood
Pb), similar to the relationship found by Den Hond et al. (2002) for White men and by
Rothenberg et al. (2002a) for postpartum women. Stepwise regression results in different
covariate patterns for each stratum and capitalizes on chance significance due to multiple testing.
In a follow-up of the Cadmibel study, the PheeCad study evaluated 359 men and
369 women, aged 20 to 82 years (Staessen et al., 1996a). Fifty-nine percent of the men had
occupational Pb exposure. They were measured twice, at baseline and at follow-up about 5 years
later. Men's mean blood Pb at baseline and follow-up was 11.4 |ig/dL (range 5.6-28.8) and
7.7 |ig/dL (range 3.7 to 20.1). Women's mean blood Pb at baseline and follow-up was 6.6 |ig/dL
(range 3.3-24.50 and 4.8 |ig/dL (range 1.7-11.8). Multiple regression models were stratified on
sex and, in women, on menopausal status. Time-integrated blood pressure measurements were
used. Each doubling of log blood Pb was significantly associated with a 5.19 mm Hg (95% CI:
1.05, 9.34) increase in diastolic blood pressure in 187 pre- and perimenopausal women. None of
the other strata showed significant blood Pb-related effects. Using 24-h ambulatory blood
6-126
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pressure readings during the follow-up showed significant associations between natural log
blood Pb and diastolic blood pressure in the group of all 345 women (2.42 mm Hg [95% CI:
0.00, 4.84]). There were no significant Pb effects on systolic blood pressure in women or all
blood pressure in men. Change in blood pressure and change in covariates between baseline and
follow-up were used to assess the effect of change of blood Pb in longitudinal analyses, similar
to M011er and Kristensen (1992) above. No significant effects of change in blood Pb on change
in blood pressure were found. Due to stratification and resulting small groups, there may have
been reduced power to detect significant Pb effects.
The Health Survey for England 1995 examined a representative sample of the English
population living in private households and provided up to 2,563 men and 2,763 women with a
mean age of 47.6 years in a study of blood Pb-blood pressure relationships (Bost et al., 1999).
Precise blood Pb ranges were not given, but were at least from less than 1.5 |ig/dL to greater than
8.5 |ig/dL, with geometric means of 2.6 |ig/dL (females) and 3.7 |ig/dL (males). The study used
stepwise multiple regression modeling of diastolic and systolic blood pressure stratified by sex,
with and without adjustment for alcohol, and with and without subjects on antihypertensive
medications. Candidate covariates, selected from a larger pool, included age, alcohol use (heavy
drinkers versus all other drinkers and nondrinkers), SES (manual classes versus non-manual
classes), location of residence in country (northern resident versus non-northern resident),
smoking, and common log blood Pb. As nonsignificant variables did not remain in the models,
each model contained a unique mix of covariates. A doubling in blood Pb in men was associated
with an increase in diastolic blood pressure of 1.07 mm Hg (95% CI: 0.37, 1.78) when alcohol
consumption was not in the model and 0.88 mm Hg (95% CI: 0.13, 1.63) when alcohol
consumption was in the model. Women had a significant response to Pb only for diastolic blood
pressure in the model without adjustment for alcohol and with subjects using antihypertensive
medication. There were no significant Pb effects on systolic blood pressure in any model.
The authors provided no statistical justification for stratified modeling nor did they test for
significant differences in Pb coefficients as a result of the stratification.
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6.5.2.2.4 Occupational Studies
U.S. Occupational Studies
Glenn et al. (2003) was one of the few studies to use a prospective design and was the
only study using statistical techniques designed for repeated measures. They studied 496 male
workers from New Jersey with former organolead exposure. Using generalized estimating
equations (GEE) with baseline linear blood Pb, age, BMI, smoking, education, antihypertensive
medication, measurement technician, and number of years to follow-up measurement of blood
pressure (range 10 months-3.5 years), they found every 1 |ig/dL increase in baseline blood Pb to
be associated with 1.13 mm Hg/year (95% CI: 0.25, 2.02) increase in blood pressure over the
observation period.
Schwartz et al. (2000c) reported significant blood Pb associations with 543 male former
organolead workers. Stepwise backward multiple regression showed an increase of 2.3 mm Hg
in systolic blood pressure for each doubling in blood Pb. The association with diastolic blood
pressure was not significant.
Sharp et al. (1990) studied 132 Black bus drivers (blood Pb range 3.1-20.9 (ig/dL) and
117 non-Black bus drivers (blood Pb range 2.0 to 14.7 |ig/dL) in San Francisco, aged 30 to
60 years. They used natural log blood Pb in multiple regression models and found for each
doubling of blood Pb an increase of 5.22 mm Hg (95% CI: 0.60, 9.84) in systolic blood pressure
among Blacks, 3.27 mm Hg (95% CI: 0.10, 6.44) in diastolic blood pressure among Blacks, and
-3.96 mm Hg (95% CI: -8.32, 0.42) in systolic blood pressure among non-Blacks.
Sokas et al. (1997) reported a possible race interaction (p = 0.09) on systolic blood
pressure with linear blood Pb in 264 construction workers aged 18-79 years. Each 1 |ig/dL
increase in blood Pb increased systolic blood pressure in Blacks by 0.86 mm Hg more than in
Whites. Neither the Black or White Pb coefficients were significant.

European Occupational Studies
Maheswaran et al. (1993) reported on 809 male factory workers with blood Pb levels
between <21 to >50 |ig/dL from Birmingham, England. Unfortunately, the inclusion of other
factors strongly related to blood Pb, including an additional direct measure of Pb exposure (years
working in factory) in addition to linear blood Pb and inclusion of zinc protoporphyrin, may have
biased the blood Pb effect and resulted in nonsignificant Pb effects on blood pressure.
6-128
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Telisman et al. (2004) also reported nonsignificant effects of natural log blood Pb on
blood pressure in 115 male industrial workers with blood Pb levels between 9.9 and 69.9 |ig/dL,
but included erythrocyte protoporphyrin in models, a variable correlated with blood Pb over
much of the observed blood Pb range. Coefficients were not given, as Pb did not enter into
stepwise regression models. The study had very low power.

Asian Occupational Studies
Male and female factory workers (n = 798) from Chonan, Korea (blood Pb between
17.8 and 64.8 |ig/dL) were studied principally for the effects of genotype of ALAD and vitamin
D receptor on cardiovascular response to Pb (Lee et al., 2001). These aspects are covered more
thoroughly below. As part of their work, the authors developed multiple regression models
examining the effect of linear blood Pb on blood pressure with forced entry of age and age2,
BMI, sex, antihypertensive medication, lifetime alcohol, and ALAD and vitamin D genotypes.
A marginally significant effect of blood Pb on systolic blood pressure (diastolic blood pressure
not modeled) was noted, with a 10 |ig/dL increase in blood Pb associated with a 0.7 mm Hg
(95% CI: -0.04, 1.4) increase in blood pressure.
Nomiyama et al. (2002) used a combined group of 193 female crystal glass workers and
nonexposed controls, aged 16 to 58 years, with blood Pb between 3.8 and 99.4 |ig/dL.
The authors used a stepwise multiple regression with a novel technique to reduce collinearity
among covariates. From a large group of covariates, they selected covariates eligible to enter the
regression from a factor analysis. Although the stepwise entry of these variables resulted in
different models for systolic and diastolic blood pressure, both models included linear blood Pb,
age, urine protein, and plasma triglycerides. The diastolic model additionally included family
hypertension and low density lipoprotein. Each 10 |ig/dL increase in blood Pb was significantly
associated with a 1.26 mm Hg (95% CI: 0.58, 1.94) increase in systolic blood pressure and a
1.05 mm Hg (95% CI: 0.52, 1.57) in diastolic blood pressure. In alternative models with
ordered categories of blood Pb, systolic blood pressure was 7.5 mm Hg (95% CI: 3.0, 12.0) and
diastolic blood pressure was 6.3 mm Hg (95% CI: 3.4, 9.1) higher in workers with blood Pb
>60 |ig/dL than in controls with <11.4 |ig/dL. Models did not control for BMI.
Wu et al. (1996) examined the effect of ordered blood Pb category on blood pressure of
112 male (aged 18-67 years) and 110 female (aged 18-71 years) Pb battery factory workers in
6-129
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multiple regression models. Blood Pb ranged from 8.3 to 95.4 |ig/dL. Nonsignificant blood Pb
effects were found possibly due to the inclusion of two additional Pb exposure measurements,
ambient air Pb and work history, likely leading to substantial collinearity with blood Pb.

6.5.2.2.5 Meta-Analyses of Blood Lead-Blood Pressure Studies
The most recent meta-analysis of the blood Pb-blood pressure literature analyzed
31 studies from a large pool of studies published up to 2001 (Nawrot et al., 2002). Two other
meta-analyses (Schwartz, 1995; Staessen et al., 1994) that were also published during this
reporting period covered many of the earlier papers cited in Nawrot et al. (2002) and derived
similar coefficients for the Pb effect; so, they are not reviewed here. The Nawrot et al. (2002)
meta-analysis authors selected studies with 50 or more subjects, with subjects 10 years of age
and up, with blood pressure and blood Pb measurement techniques presented in sufficient detail
to estimate effect sizes, and with preference given to papers with models adjusting for age, BMI,
and "additional factors of proven importance." Where possible, studies with stratified analyses
based on sex and race were entered in the meta-analysis as separate subgroups. Studies were
weighted by the number of subjects to arrive at estimates and CIs for Pb effect on diastolic and
systolic blood pressure. Nearly half the studies reported Pb effects from linear Pb terms, the
remainder from log-transformed Pb. To include both types of studies in the analyses, the authors
reported effect sizes based on doubling the mean blood Pb concentration. For models using
logarithmic blood Pb, this doubling has the same effect anywhere in the range of blood Pb in the
study. For models using linear blood Pb, the doubling effect was referenced from the mean
blood Pb reported. Figures 6-10 and 6-11 depict the effect estimates for systolic and diastolic
blood pressure, respectively, included in the meta-analysis from Nawrot et al. (2002). Ninety-
five percent CIs overlapped for males and females and for Blacks and Whites, suggesting to
Nawrot et al. no significant differences in Pb effect by gender or race. The results from the
various studies are generally consistent, with a large majority indicating positive effects. In the
group of studies as a whole, the combined meta-analysis coefficients for each doubling of blood
Pb were highly significant for both systolic (1.0 mm Hg [95% CI: 0.5, 1.4]) and diastolic
(0.6 mm Hg [95% CI: 0.4, 0.8]) blood pressure. The meta-analysis provides strong evidence for
an association between increased blood Pb and increased blood pressure over a wide range of
populations.
6-130
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Study
Population
M
M
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M, W
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Reference

Reimerand Tittelbach (1989)
Weiss etal. (1986)
deKortetal. (1987)
Morris el al. (1990)
Lockett and Arbuckle (1987)
Sharp etal. (1990)
Sharp etal. (1990)
Staessen et al. (1990)
Morris etal. (1990)
Sokasetal. (1997)
Apostoli etal. (1990)
Apostoli etal. (1990)
Nerietal. (1988)
Staessen et al. (1996a)
Staessen et al. (1996a)
Gartside(1988)
Staessen et al. (1990)
Gartside(1988)
Parkinson etal. (1987)
Orssaud et al. (1985)
Moller and Kristensen (1992)
Rothenberg et al. (1999)
Grandjean et al. (1989)
Schwartz et al. (2000b)
Grandjean et al. (1989)
Kromhout et al. (1985)
Proctor etal. (1996)
Maheswaran et al. (1993)
Elwoodetal. (1988a,b)
Elwoodetal. (1988a,b)
Elwoodetal. (1988a,b)
Rothenberg et al. (1999)
Menditto et al. (1994)
Chuetal. (1999)
Chuetal. (1999)
Henseetal. (1993)
Henseetal. (1993)
Den Hond et al. (2002)
Nerietal. (1988)
Den Hond et al. (2002)
Bostetal. (1999)
Gartside(1988)
Bostetal. (1999)
Gartside(1988)
Rabinowitz et al. (1987)
Den Hond et al. (2002)
Den Hond et al. (2002)
Pococketal. (1984)

All
Systolic Blood Pressure

Figure 6-10. Change in the systolic pressure (effect estimate in mm Hg) associated with
a doubling of the blood lead concentration. Studies arranged vertically by
increasing study size.
Study key:
Source: Nawrot et al. (2002).
C = Caerphilly Study, HP = Welsh Heart Program, P = PheeCad Study,
W = Whites, B = Blacks, NI = nonimmigrants, I = immigrants,
FW = foundry workers, CS = civil servants.
6-131
-------

Study
Reference Population
Rffiaimar fiinrJ TrMc»IHcij"h /"tQftQ^ **
KSajnsF 3F1Q P luGioacn \ ryoyj
yVoieo da* a1 MQftK^
WcISS 61 3J. 1, [yoOjl
PVa K'Art ttt at f 1 CkP7Y
Lf@ r\oin si EH. ^ i yo i j
Morris etai, (1990)
Lockett and Arbuckte (1987)
Sharp etal, (1990)
Sharp et al. (1990)
Staessenetal, (1990)
yonisetal. (1990)
Krofnhout 6t 3l, (1985)
Sokasetal. (1997)
Apostoti et ai. (1990)
Apostoti et ai, (1990)
Nerieial. (1988)
Staessen et al, (1996a)
*^tja*»ft«£
J ti-1 1 fl

145
•1 ft "5
!;>*:
1 f*n i A

254
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£. 1 !
288
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546
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1136 (•
1188
1319
1329
1471 i
1661
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1760 I
2193 h
2187
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2827
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)

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II 1 1 I 1 I
-10 -5 0 5 10 15 20 25
Diastolic Blood Pressure

Figure 6-11. Change in the diastolic pressure (effect estimate in mm Hg) associated with
a doubling of the blood lead concentration. Studies arranged vertically by
increasing study size.
Study key:
Source: Nawrot et al. (2002).
C = Caerphilly Study, HP = Welsh Heart Program, P = PheeCad Study,
W = Whites, B = Blacks, NI = nonimmigrants, I = immigrants,
FW = foundry workers, CS = civil servants.
6-132
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Figure 6-12 compared the effect estimates obtained from studies using log-linear versus linear
models. Several of the studies from Nawrot et al. (2002) are included. The studies shown in
Figure 6-12 were selected to include those studies with subjects having contemporary blood Pb
with contemporary blood pressure, published 1990 to present. Effects for the entire study
population are presented unless only effects in subsamples are reported. Other selection criteria
used are detailed in the legend of Figure 6-12. Results from these individual studies also
generally appear to agree with the results of the meta-analysis by Nawrot et al. that increased
blood Pb levels are significantly associated with increased systolic and diastolic blood pressure.
A random effects meta-analysis was preformed to examine the use of log-linear and linear blood
Pb models in blood pressure studies. A significant blood Pb effect on systolic blood pressure
was observed for both the log-linear (p = 0.05) and linear models (p < 0.001). Heterogeneity
was significant for the log-linear model (p = 0.0002), but not the linear model (p = 0.319).
The log-linear and linear effects were 0.62 mm Hg (95% CI: 0.12, 1.11) and 0.55 mm Hg
(95% CI: 0.33, 0.772) per 5 |ig/dL respectively, for systolic blood pressure. The difference
between these effect estimates using linear or log linear models is non-significant. A meta-
regression analysis was done using the geometric mean of the blood Pb in each study.
Geometric mean blood Pb was insignificant, indicating that the heterogeneity found is not due to
the slopes varying with the mean level of blood Pb. These meta-analyses suggest there may be
some differences between the studies, but overall there is an effect of blood Pb on systolic blood
pressure. Furthermore, the meta-analyses results suggest that studies not detecting an effect may
be due to small sample sizes or other factors affecting precision of estimation of the exposure-
effect relationship.

6.5.2.3 Blood Pressure and Hypertension Studies Using Bone Lead as Exposure Index
Since the 1990 Supplement, several studies have examined the association between Pb
and blood pressure or hypertension using bone Pb levels as the exposure index. The key studies
are discussed here, and additional studies are summarized in Annex Table AX6-5.1.
Korrick et al. (1999) used a case-control design to study relationships between
hypertension in women and three measures of Pb exposure: blood Pb, tibia (cortical bone) Pb,
and patella (trabecular bone) Pb. The final study sample consisted of 89 hypertension cases and
6-133
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Natural Log Lead in Model
Linear Lead in Model
'-'hflp G| 3.1 l13DO!r noiiBMi * §
Sharp et al.( 1990): BM
Rosenberg et al. ( 1 999): nonIMM i-
Utoller etal. (1992): M ^
Msileretal.(1992):F
Prwtor etal. (1996): M i~
Staessen et al. (1993): F H-
Slaessen et al. (1993): M |_^n
Roltienberg e! al. (1999): IMM
Den Hond etal. (2002): BM
Den Hond et al. (2002): BF
Den Hond etal. (2002): WM i
Den Hond et si. (2002): WM f

, -« ,
« — i
— i 1
i « j
-n
t — i

HH
HH
I-»H
4H
>
Hen see! a!. (1994): M
Nomiyama el al. (2002): F
Geiretal. (2002): M/T
Cheng etal. (2001): M
Schwartz et al. (2000c): M
Hense etal. (1994): F
Lee etal. (2001): M/F
Maheswaran etal. (1993): M
Nash et al. (2003): F
Soreietal. (1991): M
Vupputurietal. (2003): BM
Vupputuri et al. (2003): BF
Vupputyri et al. (2003): WF
Vupputun et al. (2003): WM
-10 -5 0 5 10
Change in Systolic Blood Pressure
-10 -5 0 5 10
Change in Systolic Blood Pressure
Sharp el al. (1990): nonBM
Sharp el al. (1990): BM
Rottienterg ei al. (1999): nonIMM
Mailer el al. (1992): M
Muter el al. (1992): F
Proctor eial. (1996): M
Slaessen el al. (1993): F
Slaessen eial, (1993): M
Rotnenberg et al. (1999): IMM
Den Hond et al. (2002): BM
Den Hond eial. (2002): BF
Bast eial. (1999): M/F
Schwartz (1991): F
Den Hond et ai. (2002): WM
Schwartz (1991): M
Den Hond et ai. (2062): WM
Hense etal. (1994): M
Nomiyama et al. (2002): F
Gert ei al. (2002): M/F
Schwartz etal.(2000c):M
Hense etal, (1994): F
Maheswaran et al. (1993): M
Nash et al. (2003): F
Soreietal. (1991). M
Vuppuluri et ai, (2003): 8M
Vupputun et al. (2003): BF
Vuppuiuri et al. (2003): WF
Vupputuri et al. (2003): WM
-10 -5 0 5 10
Change in Diastolic Blood Pressure
-10 -5 0 5 10
Change in Diastolic Blood Pressure
Figure 6-12. Effect of doubling mean blood lead on estimate of blood pressure change with 95%
CIs. In studies using linear blood lead terms the effect size was calculated using
blood lead doubling from 5 to 10 ug/dL. Studies not reporting sufficient
information to present coefficients and CIs were not included. Studies arranged
vertically by increasing study size. Where multiple models from the same study
were presented, such as repeated measures over time or adding a confounding
variable, only the effect estimate from the first model is shown. When the same
study was multiply published with subsamples, only the effect estimate from largest
study is shown.
Study key: B = Blacks, W = Whites, M = males, F = females, IMM = immigrants,
nonIMM = nonimmigrants.
6-134
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195 controls, excluding those with history of hypertension, cardiovascular disease, renal disease,
diabetes, or malignancy, use of antihypertensive medications, BMI >29, and incomplete data,
aged from 47 to 74 years. Cases were selected through a randomization procedure that produced
approximately equal numbers of cases for each of three blood pressure categories, hypertensive
(> 140 mm Hg or 90 mm Hg), high normal (> 121/75 mm Hg up to hypertension limit), and low
normal (<121/75 mm Hg). As many as four controls were matched to cases by 5 year age
grouping. Though they did not match cases and controls on other potential confounding
variables, they included these variables in their models. The dependent variable was constructed
by placing blood pressure measurements into the three groups. The mean blood Pb level was
3.1 |ig/dL; the mean tibia and patella Pb levels were 13.3 jig/g and 17.3 jig/g, respectively.
An ordered logistic regression with proportional odds assumptions was used to asses linear blood
Pb, patella and tibia bone Pb effects on odds of hypertension, controlling for age, BMI, dietary
calcium, alcohol use, dietary sodium, smoking, and family hypertension. They presented results
from four models with the same covariates determined a priori, but with each Pb variable tested
separately. Only patella Pb concentration significantly (p = 0.03) predicted increased odds for
hypertension, but the effect was small. Each 10 jig/g increase in patella Pb was associated with
an odds ratio of 1.28 (95% CI: 1.03, 1.60). Separate analyses testing interactions of alcohol use,
age, and menopausal status showed no significant interaction with patella Pb, though the small
sample size had little power to detect significant interaction effects. Model diagnostics were
given for justifying the use of proportional odds ordinal regression, but none were given
justifying use of a linear blood Pb term in the models.
Rothenberg et al. (2002a) investigated associations between both hypertension and blood
pressure with blood Pb, tibia Pb, and calcaneus Pb in 668 women, aged 15 to 44 years, in the
third trimester of pregnancy and during a 3-month postpartum period using a cohort design and
multiple logistic and multiple linear regression modeling. Subject exclusion criteria were blood
Pb > than 5 geometric SDs from the geometric mean, documented renal disease, cardiovascular
disease, diabetes, use of stimulant drugs, and extreme postnatal obesity (BMI >40). Geometric
mean prenatal and postnatal blood Pb levels were 1.9 |ig/dL and 2.3 |ig/dL, respectively. Mean
tibia and calcaneus Pb levels were 8.0 jig/g and 10.7 |ig/g, respectively. Variables in all models
were selected a priori and retained in the models regardless of significance level. Control
variables were education, smoking status, immigrant status, parity, age, and BMI in all models.
6-135
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Prenatal models also controlled for postpartum hypertension in lieu of family history of
hypertension. None of the subjects used antihypertensive medications during the study.
All three Pb variables were simultaneously tested in all models. Third trimester blood Pb ranged
from 0.4 to 30.0 |ig/dL, postpartum blood Pb ranged from 0.2 to 25.4 |ig/dL. Calcaneus Pb
ranged from -30.6 to 49.9 |ig/g and tibia Pb ranged from -33.7 to 42.5 |ig/g. Only calcaneus Pb
was significantly associated with an increase in hypertension (either > 140 mm Hg systolic or
>90 mm Hg diastolic) during pregnancy, with an odds ratio of 1.86 (95% CI: 1.04, 3.32) for
each 10 jig/g increase of calcaneus Pb. No association between calcaneus Pb and hypertension
was found postpartum. The authors found the same pattern of trabecular Pb concentration
association with blood pressure during but not after pregnancy in normotensive women.
A 10 |ig/g increase in calcaneus Pb was associated with -0.75 mm Hg (95% CI: 0.04, 1.46)
increase in systolic and -0.58 mm Hg (95% CI: 0.01, 1.16) increase in diastolic blood pressure
in the third trimester. Thorough diagnostic testing was performed for all models. Only linear
age terms were used in the models without exploration of age2 terms. The authors did not use
the repeated measures nature of the design in their analyses; instead they analyzed third trimester
pregnancy data and postpartum data separately. They did not statistically test differences in
coefficients from the same variables in the two parts of the study.
Two studies examined a subset of subjects participating in the Normative Aging Study.
Hu et al. (1996) used a cross-sectional design of 590 men with median age in the mid-60s
(range 48-92 years). Blood Pb ranged from 1 to 28 |ig/dL, tibia Pb from <1 to 96 |ig/g, and
patella Pb from 1 to 142 |ig/g. Logistic regression models were initially constructed by adding
age, race, BMI, family history of hypertension, smoking, alcohol use, and dietary sodium and
calcium. Testing linear blood Pb, tibia Pb, and patella Pb one by one against hypertension status
(systolic >160 mm Hg, diastolic >96 mm Hg, or taking antihypertensive medication), they found
no significant relationships with any of the Pb variables, each entered separately. Only when
they used backward elimination of nonsignificant variables did they find a significant odds ratio
of 1.50 (95% CI: 1.09, 2.10) for each doubling of tibia Pb from the mean (20.8 |ig/g) for
hypertension.
Later, Cheng et al. (2001) followed up the same group, constructing a multiple linear
regression model for systolic blood pressure (diastolic blood pressure was not mentioned in
model descriptions) in subjects not hypertensive at baseline measurement. They used a fixed set
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of control variables, including age and age terms, BMI, family history of hypertension, and
alcohol and calcium intake, selected by univariate and bivariate testing of a larger set.
After entering linear blood Pb, tibia Pb, and patella bone Pb separately into the models, they
reported a significant association only with tibia Pb (1.60 mm Hg [95% CI: 0.00, 4.44] increase
in systolic blood pressure for each doubling of tibia Pb from the mean). Several years later (not
specified in methods but no more than 6 years), the group of subjects that was originally not
classified as having definite hypertension was retested for presence of definite hypertension
(> 160/95 mm Hg). Each Pb measure was separately entered into a Cox's proportional hazards
model of incident definite hypertension. Only patella Pb showed a significant increase in the rate
ratio in subjects with no history of definite hypertension, 1.14 (95% CI: 1.02, 1.28) for each
10 |ig/g increase in patella Pb. Similar results were obtained when the borderline hypertensive
group (> 140/90 mm Hg) was combined with the definite hypertension group in patella Pb.
A rate ratio of 1.23 [95% CI: 1.03, 1.48]) was estimated. Use of linear Pb terms may have
affected the ability of the studies to detect significant blood Pb effects.
A pair of studies, using the same group of male workers (age range 42 to 74 years)
previously exposed to organic and inorganic Pb at an industrial plant in the United States,
investigated the role of blood Pb and bone Pb on blood pressure. Blood Pb ranged between
1 and 20 |ig/dL, and tibia Pb from -1.6 to 52 |ig/g. The study by Schwartz et al. (2000c)
controlled for age, BMI, current smoking, and current use of antihypertensive medication in
backward elimination linear multiple regression models for blood Pb, tibia Pb, and
DMSA-chelatable Pb, forcing each Pb term into separate models. Only blood Pb was a
significant predictor of blood pressure. In multiple logistic regression models, only blood Pb in
workers <58 years of age was significant in predicting hypertension (>160/96 mm Hg).
Although this study used linear blood Pb in one model, it used another model with both linear
and squared-blood Pb. Both Pb terms were significant in the respective models.
In a follow-up study (Glenn et al., 2003) with most of the same subjects from the first
study, subsequent measurements of blood pressure occurred at intervals of 4-12 months for
10.2 months to 3.5 years. The study was notable not only for its prospective nature but in the use
of statistical models adjusting for repeated measurements. Models were constructed by adding to
a base model containing age at start of study, race, BMI, and indicator variables for technician.
Lead variables were always forced in the models, but it is not clear if they were each tested
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separately. Other potential confounder variables were added stepwise to the model if they met a
probability criterion. Both increasing linear blood Pb and tibia Pb were significantly associated
with increasing systolic blood pressure times the number of years of follow-up blood
measurement, but not with change in diastolic blood pressure. Each 10 jig/g increase in tibia Pb
was associated with a 0.78 mm Hg, year (95% CI: 0.24, 1.31) increase in systolic blood pressure
for workers followed for the longest time.
Gerr et al. (2002) tested the effect of blood Pb and tibia Pb only in young adults (age
19-29 years), both males and females, on blood pressure. Half the subjects had grown up around
an active Pb smelter. Multiple linear regression models always used age, sex, height, BMI,
current smoking status, frequency of alcohol consumption, current use of birth-control
medication, hemoglobin level, serum albumin, and income, regardless of significance levels.
Both blood Pb (as a linear term) and bone Pb (a four category ordinal variable from <1 jig/g to
>10 |ig/g) were tested together. Tibia Pb concentration in the highest group was associated with
a significant increase in both systolic (4.26 mm Hg) and diastolic (2.80 mm Hg) blood pressure
when compared to the lowest tibia Pb group.

6.5.3 Other Cardiovascular Outcomes
Cardiovascular morbidity studies reviewed in this section are further summarized in
Annex Table AX6-5.2. Cardiovascular mortality studies are presented in Annex Table AX6-5.3.

6.5.3.1 Ischemic Heart Disease
A community-based case-referent study taken from the Stockholm Heart Epidemiology
Program compared survivors of first-time myocardial infarction with matched referents based on
sex, age, year of study enrollment, and hospital catchment area (Gustavsson et al., 2001).
The authors assessed Pb exposure by a three category ordinal scale based on Pb levels in
airborne dust. In the comparison of unexposed to >0-0.03 mg/m3 (mean 0.01 mg/m3) and
unexposed to >0.04 mg/m3 (mean 0.10 mg/m3), the relative risk was 0.88 (95% CI: 0.69, 1.12)
and 1.03 (95% CI: 0.64, 1.65), respectively.
In a reanalysis of the NHANES II dataset, the influence of linear blood Pb on the
diagnosis of left ventricular hypertrophy (LVH), based on examination of electrocardiograms
and body habitus data in -9,900 subjects (exact number not given) aged 25 to 74 years, was
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tested in a survey-adjusted stepwise logistic regression model (Schwartz, 1991). The final model
adjusted LVH by age, race, and sex. The odds ratio for LVH was 1.33 (95% CI: 1.20, 1.47) for
each 10 |ig/dL increase in blood Pb over an unreported blood Pb range. The author reported no
significant interactions between blood Pb and race or between blood Pb and sex, though the
article noted that the number of cases of LVH was small. The linear Pb effect had greater
significance than the natural log Pb effect, the reverse of the relationship between the two Pb
specifications usually seen when blood pressure is the outcome variable.
In another study of electrocardiograms in 775 men (mean age 68 years, range 48-93) from
the Normative Aging Study, patella and tibia Pb concentrations were significantly associated
with increased heart rate-corrected QT and QRS intervals in men under 65 years but not over
65 years old in multiple regression stepwise analysis (Cheng et al., 1998). Only tibia Pb
concentration was significantly associated with an increased odds ratio of intraventricular
conduction deficit (2.23 [95% CI: 1.28, 3.90]) for every 10 |ig/g increase in tibia Pb), but only in
men under 65 years. In contrast, both tibia and patella Pb concentration were significantly
associated with atrioventricular conduction deficit (odds ratio of 1.22 [95% CI: 1.02, 1.47] and
1.14 [95% CI: 1.00, 1.29] for each 10 jig/g increase in tibia and patella Pb, respectively), but
only for men >65 years old. None of the Pb measurements were significantly associated with
arrhythmia. Linear blood Pb terms were not significantly associated with any of the above
outcomes. Though the authors reported examining both saturated models (models with all
considered control and confounding variables, significant or not) and stepwise models, only
stepwise models were presented or discussed with each Pb term forced into separate models.
Thus, each model had an individual mix of control/confounding variables, though age was
common to all models. Despite using age as a control/confounding variable in all models, the
article offered no statistical justification for the age-stratified analysis.
A group of male and female battery factory workers (n = 108) working for at least
10 years and who were hired from 1960 to 1983 had blood Pb levels during 1970 to 1994 that
ranged from 5 to 93 |ig/dL (Tepper et al., 2001). Using a fixed covariate multiple logistic
regression model, including age, BMI, sex, and family history of hypertension, the authors found
a nonsignificant odds ratios for risk of hypertension (>165/96 mm Hg or self-reported use of
hypertension medications) comparing the first tertile (138-504 jig/dL'year) cumulative blood Pb
index with the third tertile (747-1447 jig/dL'year) index. Echocardiogram left ventricular mass
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was not significantly related to cumulative blood Pb index or time-weighted average blood Pb.
The study had very low power to detect significant effects.
The discrepancy in blood Pb results between the two electrocardiogram studies by
Schwartz (1991) and Cheng et al. (1998) could well be explained by population differences.
Though both used large datasets, the age range of the NHANES II subject pool was between
25 and 74 years and used both men and women, whereas the age range for the Normative
Aging study was 48 to 93 years and used only men. Furthermore, the Cheng et al. study had
775 subjects whereas the Schwartz had a much larger, though unspecified number. The Tepper
et al. (2001) study had the least number of subjects (n = 108), which may have resulted in not
detecting significant effects on a different measure of LVH. Still, both electrocardiogram studies
reported a significant Pb effect, and the study with bone Pb (Cheng et al., 1998) is particularly
interesting, not only for its older sample but because the bone Pb exposure measure reflected
accumulated past exposure, which blood Pb only partly reflects. The two studies are in
agreement that Pb exposure, either past or present, is significantly associated with ischemic heart
disease.

6.5.3.2 Cardiovascular/Circulatory Mortality
A recent follow-up of the NHANES II cohort provided mortality data used to associate
past blood Pb concentration with increased circulatory mortality in the U.S. population (Lustberg
and Silbergeld, 2002). Blood Pb concentration as measured during 1976 to 1980 was divided
into three categories (<10 |ig/dL, 10-19 |ig/dL, and 20-29 |ig/dL) after eliminating 109 subjects
with blood Pb >30 |ig/dL, leaving 4,190 subjects 30-74 years of age in the mortality sample
followed to the end of 1992. During the follow-up period, 929 subjects died of all causes.
ICD-9 codes 390-459 (circulatory) accounted for 424 deaths. Proportional hazards models using
a priori selected potential confounding variables (age, sex, race, education, income, smoking,
BMI, exercise, and location) were used to calculate risk ratios of cardiovascular mortality for the
two higher Pb categories compared to a <10 |ig/dL reference. The 20-29 |ig/dL category showed
significantly elevated relative risk of 1.39 (95% CI: 1.01, 1.91) for cardiovascular mortality.
Although the NHANES II analysis using data from 1976 to 1980 suggested an increased
risk of mortality at blood Pb levels above 20 |ig/dL, blood Pb levels have dramatically decreased
since the late 1970s. More recent data from NHANES have found that the geometric mean
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blood Pb levels decreased from 12.8 ng/dL in 1976-1980 to 2.8 ng/dL in 1988-1991 (Annest
et al., 1983) and 2.3 jig/dL in 1991-1994 (CDC, 1997). NHANES III data (1988-1994) were
used to further analyze risk of mortality in adults (age 40 years) at lower blood Pb levels
(Schober et al., 2006). A total of 9,757 subjects were followed for a median of 8.55 years during
which there were 2,515 deaths. An increased risk of cardiovascular mortality was associated
with blood Pb levels of 5-9 |ig/dL and 10 |ig/dL compared to 5 |ig/dL. The relative risk was
1.20 [95% CI: 0.93, 1.55] for 5-9 jig/dL and 1.55 [95% CI: 1.16, 2.07] for 10 |ig/dL, and the
test for trend was statistically significant. Increased risks of all cause and cancer mortality also
were observed at blood Pb levels of 5-9 |ig/dL compared to <5 |ig/dL (relative risk of 1.24
[95% CI: 1.05, 1.48] for all cause mortality and 1.44 [95% CI: 1.12, 1.86] for cancer mortality).
The authors noted that an important limitation of this study was that exposure classification was
based on one blood Pb level measurement taken at baseline. Older individuals were more likely
to have notably higher past peak and cumulative Pb exposure, and their blood Pb levels might
have been disproportionately influenced by release of Pb from bone stores compared to younger
individuals.
Another longitudinal study combined fatal and nonfatal coronary heart disease (ICD-8
codes 410-414) and cardiovascular disease (ICD-8 codes 410-414 and 430-435) categories from
a Danish 1936 birth cohort (n = 1,052) followed from 1976 to 1990 (M011er and Kristensen,
1992). During the study period, 54 cases of cardiovascular disease with 19 deaths were reported.
Log-transformed blood Pb was used in a Cox proportional hazards model, controlling for a priori
selected variables of tobacco use, cholesterol, physical activity, sex, systolic blood pressure, and
alcohol. Two other models were also examined, those leaving out alcohol or both alcohol and
systolic blood pressure. None of the adjusted models showed significant risk hazard for
combined fatal and nonfatal cardiovascular disease, though blood Pb was significantly associated
with outcome in all models except the one containing both alcohol and systolic blood pressure
for "total mortality" risk hazard. This article is notable for its detailed discussion of using
confounding variables, such as hemoglobin and alcohol use, in multivariate models of Pb-
cardiovascular associations. However, small sample size and low death rate may have
contributed to the nonsignificant results.
An occupational study, using 1,990 male workers who worked at least 1 day between
1940 and 1965 in an active Pb smelter in the United States (mean length of employment at
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smelter 13.8 years; mean estimated length of Pb exposure 9.9 years), failed to show an
association between Pb and standardized mortality ratios compared to the U.S. population
reference group up to 1988 (Steenland et al., 1992). Neither mortality from ischemic heart
disease (ICD-9 410-414), hypertension with heart disease (ICD-9 402 and 404), hypertension
with no heart disease (ICD-9 401, 403, and 405), nor cerebrovascular disease (ICD-9 430-438)
were significantly higher in the study group than in the U.S. population when examined in their
totality or stratified by "high Pb exposure" (>0.2 mg/m3 Pb in air, surveyed in 1975) or "duration
of exposure." Imprecise estimation of Pb exposure may have contributed to the nonsignificant
results.
A study of 664 male workers in a Swedish Pb smelter from 1942-1987 evaluated
standardized mortality ratios for cardiovascular disease compared to the county population
mortality figures from 1969-1989 (Gerhardsson et al., 1995). Blood Pb measurements were
available from the workers since 1969 (mean 62.1 |ig/dL) and dropped steadily from that date to
1985 (mean 33.1 jig/dL). The consecutive blood Pb measurements in the subjects allowed
construction of a cumulative blood Pb index. Standardized mortality ratios were significantly
elevated in the group for all cardiovascular diseases (ICD-8 390-458) and for ischemic heart
disease (ICD-8 410-414), 1.46 (95% CI: 1.05, 2.02) and 1.72 (95% CI: 1.20, 2.42),
respectively. However, there were no indications of a concentration-response relationship when
analyses were stratified by cumulative blood Pb index, peak blood Pb, or other exposure indices.
In a study of 1,261 male newspaper linotype operators working in 1961 and followed until
1984, 38% had died from all causes (Michaels et al., 1991). Compared to the New York City
population reference group, there was a marginally significant increased standardized mortality
ratio in the printers of 1.35 (95% CI: 0.98, 1.82) for cerebrovascular disease (ICD-8 430-438),
which became highly significant in those with 30 or more years exposure (1.68 [95% CI: 1.18,
2.31]; 37 of the total 43 deaths due to cerebrovascular disease). Atherosclerotic heart disease
(ICD-8 410-414) mortality in printers was significantly below that expected from the general
population, with a standardized mortality ratio of 0.63 (95% CI: 0.59, 0.73).
Mortality studies need to follow large groups over extended periods to achieve adequate
statistical power. When large groups with well characterized exposure are followed for long
periods, results of mortality studies assess the effects of long cumulative Pb exposure. Without
detailed exposure histories stretching over decades, it is nearly impossible to determine if past
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peak Pb exposure, time-integrated Pb exposure or average Pb exposure plays the critical role in
producing greater then expected mortality. Noting that both population and occupational Pb
exposures were at greater than current levels when the mortality studies reported here were
begun, one can expect a 20 to 30 year lapse before one could assess any effects of current
population and occupational Pb exposure on cardiovascular morbidity and mortality.

6.5.3.3 Other Cardiovascular Effects
Peripheral arterial disease (PAD), flow-limiting atherosclerosis in lower limb muscular
arteries, was studied using Phase 1 (1999-2000) of the NHANES IV, the most recent NHANES
dataset (Navas-Acien et al., 2004). PAD was categorized as a ratio of brachial artery (arm)
systolic blood pressure to posterior tibial artery (ankle) systolic blood pressure < 0.90, with
139 subjects being classified as having PAD; there were 1,986 subjects without the disease.
Blood Pb was classified by quartile, with the 1st quartile containing subjects with blood Pb
<1.4 |ig/dL and the 4th quartile containing subjects with blood Pb >2.9 |ig/dL. Age ranged from
40 to >70 years. Three sets of covariates were tested in separate models. The first set, common
to all models, included age, sex, race, and education. The second set included the first set and
added BMI, alcohol intake, hypertension, diabetes, hypercholesterolemia, and glomerular
filtration rate. The third set added self-reported smoking status and serum cotinine. Compared
to first quartile blood Pb, 4th quartile blood Pb subjects had significant odds ratios for PAD of
3.78 (95% CI: 1.08, 13.19) and 4.07 (95% CI: 1.21, 13.73) for the first two models. The odds
ratio of 2.88 (0.87, 9.47) for the third model was not statistically significant. However, the
increasing odds ratio trend from 1st through 4th quartile was significant for all 3 models
(p < 0.02).
The associations of umbilical cord blood Pb levels with pregnancy hypertension and
blood pressure during labor were assessed in 3,851 women whose babies were delivered at the
Boston Hospital for Women (Rabinowitz et al., 1987). The mean cord blood Pb level was
6.9 |ig/dL (SD 3.3, range 0-35). Blood Pb concentrations were log transformed after adding one
because some values were zero and the distribution was skewed. In a multivariate model
adjusting for hematocrit levels, diabetes, ponderal index, race, tobacco use, and birth weight,
cord blood Pb levels were significantly associated with systolic blood pressure during labor.
A 10 |ig/dL increase in cord blood Pb was associated with an increase of about 3 mm Hg in
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systolic blood pressure. Significant associations were also observed between cord blood Pb
levels and pregnancy hypertension. Compared to a cord blood Pb level of 0.7 |ig/dL, the relative
risk of pregnancy hypertension was 1.7 (95% CI: 1.3, 2.1) at a blood Pb level of 6.3 |ig/dL and
2.2 (95% CI: 1.5, 2.9) at 15 |ig/dL. Cord blood Pb levels were not found to be associated with
preexisting hypertension or preeclampsia.

6.5.4 Lead and Cardiovascular Function in Children
Despite the potential importance of identifying the effects of Pb on cardiovascular
function in children, only three studies addressed the issue. These studies are summarized in
Annex Table AX6-5.4 and reviewed here. Factor-Litvak et al. (1996) studied the association of
blood Pb on blood pressure in 260 children at age 5.5 years from a prospective study of Pb on
child development from two cities in Serbia. They used multiple linear regression of
contemporary linear blood Pb (range 4.1-76.4 |ig/dL) on systolic and diastolic blood pressure,
apparently stepwise as diastolic and systolic models had different covariates. Systolic models
controlled for height, BMI, gender, ethnic group, and birth order, while diastolic models
controlled for waist circumference, ethnic group, and birth order. Additional models further
adjusting for maternal blood pressure, maternal hemoglobin, and for town were also presented,
but not discussed here. Every 1 |ig/dL increase in blood Pb was associated with 0.054 mm Hg
and 0.042 mm Hg increase of systolic and diastolic blood pressure, respectively. Though no
diagnostics were reported, the authors did try combined linear and quadratic blood Pb terms in
alternative models and found the quadratic term to be nonsignificant. These marginally
significant results may partially obscure early indications of altered cardiovascular health in
young children exposed to Pb due to small sample size and use of linear Pb and stepwise
regression in models.
Gump et al. (2005) studied blood Pb effects on both resting and induced-stress
cardiovascular function in 122 children 9.5 years old from Oswego, NY. The effect of linear
cord blood Pb (mean 3.0 |ig/dL [SD 1.8]; range not given) and contemporary blood Pb
(range 1.5-13.1 |ig/dL) on blood pressure and other cardiovascular functions were tested in
stepwise multiple regression models. The response to stress was evaluated by taking the
difference between baseline and post-stress scores on the cardiovascular evaluations. Diastolic
blood pressure change from baseline to stress condition increased 0.07 mm Hg for every 1 |ig/dL
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increase in contemporary blood Pb. Change in total peripheral resistance between baseline and
stress conditions was 0.09 dyn-s/cm3 for every 1 |ig/dL increase in contemporary blood Pb.
Authors reported testing blood Pb with linear, quadratic, and cubic terms that "did not add
significantly to the prediction of these cardiovascular effects." Nonetheless, the scatterplot of the
Pb effect on peripheral resistance change shows a notable nonlinear effect, as does an alternative
analysis of the same outcome using blood Pb quartiles, in which a significant effect was seen
between the group with 1.5 to 2.8 |ig/dL and all higher blood Pb groups, as might be expected
from a log-linear like blood Pb-peripheral resistance dose-response curve. The stepwise
modeling technique and small sample size likely combined to over-fitting of the models.
For instance, the total vascular resistance model had 12 covariates. The model might be difficult
to replicate with an independent sample. Due to the study location, investigators had reason to
believe that the children also had significant exposure to Hg and pesticide residues from eating
contaminated lake fish.
A randomized succimer chelation trial of 780 12- to 33-month-old children with baseline
blood Pb from 20 to 44 |ig/dL resulted in significantly lower blood Pb in the treated group for
only the first 9 to 10 months of the 60-month follow up period (Chen et al., 2006). No difference
in blood pressure was noted between succimer and placebo groups during treatment. However,
longitudinal mixed models showed systolic blood pressure to be 1.09 mm Hg (95% CI: 0.27,
1.90) higher in the succimer treated group than in the placebo treated group from month 12-60
of follow up and near zero coefficient for diastolic blood pressure. Cross-sectional regression
analyses of blood pressure and linear blood Pb adjusted for clinical center, treatment group, race,
sex, parents' education, single parent, age at test, height and BMI during follow up revealed
near-zero Pb coefficients for systolic and diastolic blood pressure. The authors note the short
duration of significant blood Pb difference between the groups, the overall downward trend in
blood Pb with age in both groups, and the relatively short duration of elevated blood Pb in both
groups as possible factors for the non-significant results. The lack of positive chelation effect
observed in this study mirrors the results of chelation trials in the same children that showed no
benefit in IQ and neurobehavioral performance. No diagnostics were mentioned, though non-
parametric fitting of blood pressure and blood Pb with non-adjusted data were shown.
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6.5.5 Potential Confounding of the Cardiovascular Effects of Lead
6.5.5.1 Confounding by Copollutants
High on the list of other metals that might be associated with cardiovascular disease is Cd,
through its known effects on kidney function. If blood Pb and blood Cd strongly covary in a
sample by sharing a common source (e.g., when the study sample is drawn from a population
living near a nonferrous smelter emitting both metals), including simultaneous blood Pb and Cd
measurements in the same model would likely show a significant reduction in both coefficients
when compared to either metal alone. If, however, blood Cd and Pb do not covary in the sample,
their coefficients in the model together would be similar to when tested separately. In a study of
PAD (Navas-Acien et al., 2004) discussed in Section 6.5.3.3, investigators not only tested both
Pb and Cd in separate models but also tested them simultaneously. In addition, they tested
possible interactions between Pb and Cd, and between the two metals and sex, race-ethnicity,
smoking status, renal function, and C-reactive protein. The correlation coefficient between
natural log Pb and natural log Cd was 0.32 (p < 0.001), highly significant, though leaving 90% of
the variance between them unexplained. When blood Pb and blood Cd were in the same model
together, they both had significant trends of increasing odds ratios with increasing quartile of
each metal. However, the nonsignificant point estimate of the odds ratio for blood Pb comparing
the 1st and 4th quartile decreased when Cd was also included in the model (odds ratio of 2.88
versus 2.52). The odds ratio for Cd comparing the 1st and 4th quartile showed a similar decrease
when modeled with Pb (odds ratio of 2.82 versus 2.42), but both point estimates remained
significant. Thus, though point estimates of both Pb and Cd were approximately the same
whether tested alone or together, the larger variance associated with the Pb coefficients rendered
them nonsignificant. Part of the difference in variance between the two metals could be
explained by noting that the reference group (lowest quartile) for Pb contained a little over half
the number of subjects (n = 472; 18 cases, 454 noncases) than the reference group for Cd
(n = 856; 27 cases, 829 noncases). The odds ratios for PAD with smoking status dropped from
4.13 (95% CI: 1.87, 9.12) to 3.38 (95% CI: 1.56, 7.35) when Pb was added to the model, but
both odds ratios remained highly significant and the difference was not statistically tested.
The failure to find a significant interaction between the two metals and between smoking status
and both metals suggests that none of the odds ratio changes discussed above were significant.
The same pattern of results was found when using cotinine blood levels instead of self-reported
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smoking habit. Adding Cd alone or Cd and Pb together resulted in nonsignificant odds ratios for
both indices of smoking.
The Belgian Cadmibel studies also were ideally situated to test possible interactions
between blood Pb and Cd, but the technique of stepwise addition of variables to the multiple
regression models of blood pressure did not allow retention of both metal variables together in
the same model (Staessen et al., 1996b). From the lack of both Cd and Pb in any one model, it
can be inferred that, if both variables had been forced into the model together, they both would
have had nonsignificant coefficients.

6.5.5.2 Confounding by Smoking Status
Most studies reviewed in this section have controlled for tobacco use, where it often
appears related to lower blood pressure. The majority of reviewed studies including smoking as
a covariate never present the coefficients of smoking or related covariates. Only the Navas-
Acien et al. (2004) study discussed in the previous section systematically addressed the issues
related to possible confounding or effect modification with tobacco use.

6.5.5.3 Confounding by Alcohol Consumption
Possible confounding by alcohol use, generally associated with increased blood pressure,
was discussed in the 1990 Supplement (Grandjean et al., 1989). Alcohol, especially in Europe,
contained substantial Pb during much of the 20th century. This can be seen in the MONICA
Augsberg, Germany cohort study (Hense et al., 1994). The study group was stratified by sex and
then, only in men, by rural-urban location. Within each strata, the blood Pb range differed by
alcohol use. In women, for example, the 10th and 90th percentile values of blood Pb (as
estimated from graphs) were -3.5 and 8.5 |ig/dL for self-reported abstainers, 4.5 and 10.5 |ig/dL
in those drinking from 1 to 39 g/day, and 6.0 to 14.0 |ig/dL in those drinking 40 plus g/day.
Despite the finding that only women in the highest alcohol-use group had a significant Pb effect,
it cannot be determined if the increase in Pb coefficient is significant, because the three
coefficients associated with use of alcohol strata were not tested for differences among
themselves; they were only tested for their significance from the null hypothesis of 0. Another
study was based on subjects from the New Risk Factors Survey from the area around Rome,
intended to determine confounding effects of a number of social and biochemical variables on
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the blood Pb-blood pressure relationship (Menditto et al., 1994). Alcohol consumption, as well
as BMI, heart rate, non-HDL cholesterol, and HDL cholesterol, triglycerides, cigarettes
smoked/day, and skinfold thickness were all examined. A doubling of blood Pb was associated
with an increase of 4.71 mm Hg in systolic and 1.25 mm Hg in diastolic blood pressure.
As covariates were successively added to the model, the systolic coefficient was 4.6 (+BMI),
4.9 (+age), 5.1 (+heart rate), 4.3 (+high density lipids), 4.2 (+triglycerides), 3.9 (+glucose),
4.4 (+cigarettes/day), 4.1 (+skinfold), and 3.9 (+non-high density lipids). Similar changes were
found upon adding covariates to the diastolic model. Alcohol never entered the models, but was
significantly and positively associated blood Pb in bivariate testing. Unfortunately, neither
standard errors or confidence intervals were given and the significance of the changes in the Pb
coefficient could not be determined.
Alcohol as a true confounding variable is likely limited to studies in areas where alcohol
contributes significantly to blood Pb. In a study of 249 bus drivers in San Francisco, CA, natural
log Pb coefficients against blood pressure changed less than 10% when alcohol use was included
as a covariate (Sharp et al., 1990). Blood Pb according to alcohol use was not reported. Still
another study based on a U.S. population found a significant increase in blood Pb of a mixed
group of males and females according to alcohol use, ranging from mean blood Pb of 7.3 |ig/dL
in nonusers to 9.2 |ig/dL in those reporting more than 2 ounces/day over 3 days (Morris et al.,
1990), with no report of significant effects of alcohol on blood pressure.

6.5.5.4 Confounding by Dietary Calcium Intake
The main thrust of the previously reported Morris et al. (1990) study was to examine the
effects of dietary calcium on the effect of Pb on blood pressure in 78 males and 64 females, 18 to
80 years old, many of whom were hypertensive (undisclosed number), though those using
medications for hypertension discontinued their use 1 month before testing started. Subjects
were excluded if they had "secondary hypertension." The investigators measured serum calcium
and assessed dietary calcium intake, among other variables. There were no changes in blood Pb
or blood pressure noted as a result of dietary calcium supplementation.
Proctor et al. (1996), using the Normative Aging Study, examined possible modification
of the effect of natural log blood Pb (blood Pb range 0.5-35 |ig/dL) on blood pressure in
798 men, aged 45 to 93 years, by dietary calcium intake assessed by food questionnaire.
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The study used multiple regression models with a fixed set of covariates, including age and age2,
BMI, adjusted dietary calcium, exercise, smoking, alcohol use, sitting heart rate, and hematocrit.
Increased blood Pb was significantly associated with diastolic blood pressure and systolic blood
pressure. Only systolic blood pressure significantly decreased with increased dietary calcium
(0.004 mm Hg decrease for every 1 mg/day increase of dietary calcium). The authors formed
dichotomized calcium intake (cut point at 800 mg/day) and blood Pb (cut point at 15 |ig/dL)
variables to test the interaction between blood Pb and calcium on blood pressure. They did not
find a significant interaction; nor did they show the interaction coefficients.
A study of a subset of the Cadmibel Study with 827 males and 821 females, age 20 to
88 years, selected from areas known to represent a wide range of Cd exposure, specifically
studied total serum calcium interactions with blood Pb on blood pressure (Staessen et al., 1993).
Stepwise regression models, selecting from log blood Pb, age and age2, BMI, pulse rate, log
serum gamma-glutamyltranspeptidase, serum calcium, log serum creatinine, urinary potassium,
smoking, alcohol intake, contraceptive pill use (females only), and a menopause indicator
variable (females only), were stratified by sex for systolic and diastolic blood pressure.
The stepwise procedure resulted in models each with a different mix of covariates. Increased
serum calcium was significantly associated with increased systolic blood pressure in both males
and females. Every increase of one log unit of blood Pb was associated with nonsignificant
changes in blood pressure in women, but with a significant decrease in systolic blood pressure in
men (systolic log blood Pb P = - 5.2). A separate set of models were constructed with an
interaction term between serum calcium and log blood Pb (details not shown). In women only,
both main effects of Pb and calcium and the interaction effect were significant (no coefficients
presented). At the 25th percentile of serum calcium (2.31 jimol/L), a doubling of blood Pb was
associated with a 1.0 mm Hg increase in systolic blood pressure. At the 75th percentile of serum
calcium (2.42 jimol/L) a doubling of blood Pb was associated with a 1.5 mm Hg increase in
systolic blood pressure. Furthermore, serum calcium may itself be confounded with age in
women, as women showed a sharp rise in serum calcium in their sixth decade of life, coincident
with menopause, whereas the trend for serum calcium in men was steadily downward for each
subsequent decade of age. The authors did not test an interaction term including calcium and age
or calcium and menopausal status. Thus, the significant interaction effect between calcium and
Pb on blood pressure may be a result of differences due to menopause.
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6.5.5.5 Summary of Potential Confounding of the Lead Effect on Cardiovascular Health
The effects of Cd exposure, smoking, alcohol use, dietary and serum calcium levels have
all been formally tested in a few studies, without significant effects as confounders of the Pb
effect. Failure to find a significant confounding effect with Pb, however, does not argue to
maintain these variables uncritically in models of blood pressure. If alcohol contains Pb,
increased alcohol use will lead to increased blood Pb. In this case, both variables in the model
will be collinear and this tends to distort estimated coefficients and standard errors of their effect
on cardiovascular outcome. Tobacco use may influence Pb levels much more in occupational
studies than in community exposure studies, especially if smoking in the factory is allowed.
Frequent hand to mouth behavior will increase Pb exposure and, consequently, raise blood Pb
concentrations. Serum calcium may statistically modify the Pb effect differentially by gender
due to menopause in women. Menopause also affects Pb turnover. If serum calcium, blood Pb,
and blood pressure are all statistically related, serum calcium should not be used in blood
Pb-blood pressure/hypertension studies.
Epidemiologic studies cannot by themselves determine cause and effect relationships
between Pb and cardiovascular disease. However, toxicological studies that observe similar
phenomena in experimental animals give biological plausibility to the epidemiological results.
In addition they may suggest mechanisms by which Pb might cause the observed epidemiologic
effects. Chapter 5.5 details a series of results that run parallel to and give biological plausibility
to the results in humans detailed in this section. In intact animals, elevated blood pressure
develops only in response to continued exposure to Pb. If duration of Pb exposure is key to
Pb-induced hypertension, as suggested by the consistently observed elevation of blood pressure
and increased risk for hypertension associated with increased bone Pb (long term exposure) and
the difficulty of detecting Pb effects on blood pressure in children, these animal studies argue
that the Pb effects observed in humans are not the result of statistical artifact or confounding.

6.5.6 Gene-lead Interactions
Sodium-potassium adenosine triphosphatase a2 (ATP1A2) polymorphism was
characterized in 220 workers formerly exposed to a mix of organic and inorganic Pb in the
United States, noted above in other references (Glenn et al., 2001). The ATP1A2 (3') one
kilobase probe produced two homozygous (4.3/4.3 and 10.5/10.5) and one heterozygous
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(4.3/10.5) genotypes and two homozygous (8.0/8.0 and 3.3/3.3) and one heterozygous (8.0/3.3)
genotypes for the 2.5 kilobase ATP1A2 (5') probe. Of the 209 subjects with data on both
polymorphisms, 43.5% were doubly homozygous for 8.0/8.0 and 4.3/4.3, 34.4% were
homozygous for 8.0/8.0 and heterozygous for 4.3/10.5, 11.5% were heterozygous for 8.0/3.3 and
homozygous for 4.3/4.3, 5.3%. Also, 5.3% were doubly homozygous for 8.0/8.0-10.5/10.5, and
4.8% were doubly heterozygous for the two genotypes. Although only 13 African-American
workers participated, prevalence of the 10.5 kilobase allele in the ATP1A2 (3') genotype was
statistically higher for them than for other races. Prevalence of hypertension (> 160/96 mm Hg or
use of hypertension medication) was significantly higher in those with the 10.5/10.5 genotype
than in others. When controlling for age, BMI, lifetime number of alcoholic drinks, the
10.5/10.5 genotype was associated with an odds ratio of 7.7 (95% CI: 1.9, 31.4) for hypertension
as compared to the 4.3/4.3 homozygous genotype, but there were no effects of either blood Pb,
tibia Pb, or their interaction with ATP1A2 (3') genotype.
A multiple linear regression model for linear blood Pb and systolic blood pressure,
controlling for age, use of hypertensive medication, current smoking, quartiles of lifetime
alcohol consumption, and season, showed a significant main effect for 10.5/10.5 homozygous
contrasted against combined 4.3/4.3 and 4.3/10.5 groups, associated with a 25.5 mm Hg
reduction in blood pressure, primarily due to the limited blood Pb range for the homozygous
group (maximum blood Pb of the 10.5/10.5 group 9 |ig/dL; maximum blood Pb of the contrast
group = 20 |ig/dL). But the interaction between linear blood Pb and the 10.5/10.5 condition
resulted in a significant increase of the blood Pb effect on blood pressure by 5.6 mm Hg for
every 1 |ig/dL blood Pb compared to the blood Pb effect in the other genotypes. The authors
stated, but did not show analysis or coefficients, that the ATP1A3 (3') polymorphism also
significantly interacted with tibia Pb and systolic blood pressure. There were no significant
relationships using the ATP1A2 (5') gene. Thus, the ATP1A2 (3') polymorphism appears to
directly influence both prevalence of hypertension and the effect of Pb on blood pressure, though
the small group (n = 9 with all measures) with the important 10.5/10.5 homozygous pattern
would argue for enlarging this important study.
Another research group focused on polymorphisms of two genes suspected to be involved
in Pb toxicokinetics, VDR and ALAD (Lee et al., 2001). Polymorphism of both genes is well
studied and prevalence appears to be associated with race or ethnic background. Nearly
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800 Korean workers aged 18 to 65 years (79.4% males) from Pb-using businesses were classified
according to ALAD polymorphism (1-1 [homozygous] versus 1-2 [heterozygous]) and VDR
polymorphism (bb [predominant homozygous] versus Bb plus BB [infrequent polymorphisms]).
The homozygous ALAD 1-1 polymorphism was found in 90.1% of the group and the
homozygous bb one was found in 88.8% of the group. When compared to a smaller group of
non-Pb-exposed workers, blood Pb concentration (mean exposed 32.0 |ig/dL [range 4-86] mean
nonexposed 5.3 |ig/dL [range 2-10] and tibia Pb concentration mean exposed 37.2 jig/g
[range -7 to 338]; and mean nonexposed 5.8 |ig/dL [range -11 to 27]) were much higher.
The study used stepwise multiple regression models, selecting covariates remaining significant in
the models from among a large set of potential control and confounding variables. Potential
confounders were also allowed to remain in the models if "there were substantive changes in the
coefficients of predictor variables" with their addition. Systolic models controlled for age and
age2, sex, BMI, antihypertensive medication use, and cumulative lifetime alcohol use.
Depending on the presence or absence of linear blood Pb, tibia Pb, and DMSA-chelatable Pb in
the models, and the gene-age interactions tested, blood urea was added to the model. Diastolic
models controlled for age, sex, BMI, cumulative alcohol consumption, and linear blood Pb.
Hypertension (systolic >160 mm Hg or diastolic >96 mm Hg) logistic multiple regression
models controlled for age, sex, BMI, tibia Pb, and current alcohol use. Among the exposed
workers bb VDR genotypes had significantly lower DMSA-chelatable blood Pb and lower
diastolic and systolic blood pressure than the combined Bb and BB genotypes. The only
significant interaction reported between predictor variables and gene polymorphism on blood
pressure was with the VDR polymorphism bb allele, which had a less pronounced increase in
systolic blood pressure with age than subjects with the B allele. There were only marginally
significant associations of systolic blood pressure with tibia Pb and linear blood Pb. There were
no significant associations in models of diastolic blood pressure with linear blood Pb, DMSA-
chelatable blood Pb, or tibia Pb. Tibia Pb was significantly associated with hypertension (odds
ratio of 1.05 [95% CI: 1.00, 1.12] for each 10 |ig/dL increase in tibia Pb). Workers with VDR B
allele had significantly higher prevalence of hypertension (odds ratio = 2.1 [95% CI: 1.0, 4.4])
than workers with the bb genotype, but no other Pb variable or interaction with VDR status was
reported significant. Though VDR status was significantly related to blood pressure and
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prevalence of hypertension, there were no significant effects of ALAD polymorphism on blood
pressure or hypertension or of VDR interactions with any Pb exposure variable.
Lustberg et al. (2004) studied the same Korean Pb workers (n = 793) to examine
relationships between the Q894_x894 polymorphism in the gene regulating endothelial nitric oxide
synthase (eNOS) and blood Pb effects on blood pressure and hypertension. Nitric oxide
metabolism has been suggested both as a mechanism for altered blood pressure and for
moderating the effects of Pb on blood pressure, though there is experimental support for and
against both hypotheses. After classifying subjects as homogenous for the GG type (85%),
heterogeneous for both types (TG) (14%), or homogenous for TT (1%), the TG and TT types
were combined into a single group (TG/TT). Diastolic and systolic multiple regression models
were constructed with a fixed set of covariates, including smoking, alcohol consumption, age,
sex, BMI, and education. Logistic regression models used blood pressure criteria of either
> 140 mm Hg diastolic blood pressure, >90 mm Hg systolic blood pressure, or self-report of
using antihypertensive medications. There was no effect of genotype on diastolic or systolic
blood pressure or on hypertension prevalence in multiple regression models, nor any significant
interaction of Pb exposure indices with gene status.
Because interaction testing in statistical models optimally requires balanced groups for
uncomplicated interpretation, further gene-Pb interaction exploration should use studies with
nearly equal numbers of heterogeneous and homogenous groups. Also, because adequate power
for testing significant interactions requires large groups, subsequent studies should draw subjects
from the general population. In addition to enlarging the potential subject pool, population
studies may more easily avoid the selection biases often found in occupational studies.

6.5.7 Summary of the Epidemiologic Evidence for the Cardiovascular
Effects of Lead
The combined blood Pb studies using blood pressure/hypertension as an outcome
continue to support the conclusions of the 1990 Supplement that there is a positive association
between blood Pb and increased blood pressure. The occasional finding of significant negative
associations of blood Pb with blood pressure (e.g., the Cadmibel study, one NHANES III study,
the postpartum phase of the Los Angeles pregnancy study) have not been adequately explained
and require further confirmation and study
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The most promising developments in this field since the 1990 Supplement have been the
use of bone Pb as a long-term cumulative Pb exposure index and the introduction of genetic
analysis into the studies as potential Pb effect modifiers. With one exception, all studies using
bone Pb have found a consistently positive and significant effect on blood pressure and/or
hypertension. The ability to estimate past exposure in cross-sectional studies is a significant
advance. The results of the bone Pb studies to date highlight the important role of accumulated
Pb exposure in the development of cardiovascular problems.
Animal toxicologic studies have found that cell, tissue, and organ response to Pb is
immediate and may provide clues to the mechanisms by which Pb contributes to cardiovascular
disease in humans. Lead interference in calcium-dependent processes, including ionic transport
systems and signaling pathways important in vascular reactivity may only represent the first step
in the cascade of Pb-induced physiological events that culminates in cardiovascular disease.
Lead alteration of endothelial cell response to vascular damage, inducement of smooth muscle
cell hyperplasia, alteration of hormonal and transmitter systems regulating vascular reactivity,
and its clear role as promoter of oxidative stress suggest mechanisms that could explain the
Pb-associated increase in blood pressure, hypertension, and cardiovascular disease noted in this
section.
• Studies support the relationship between increased Pb exposure and increased adverse
cardiovascular outcome, including increased blood pressure, increased incidence of
hypertension, and cardiovascular morbidity and mortality. For blood Pb and blood
pressure, every doubling of blood Pb is associated with a -1.0 mm Hg increased systolic
and -0.6 mm Hg increased diastolic blood pressure for blood Pb between 1 and
>40 |ig/dL.
• Cumulative past Pb exposure, measured by bone Pb, may be more important than present
exposure in assessing cardiovascular effects of Pb exposure. Over the range of bone Pb
concentration of < 1.0 jig/g to 96 |ig/g, every 10 jig/g increase in bone Pb was associated
with increased odds ratio of hypertension between 1.28 and 1.86, depending upon the
study. Two studies measured averaged increased systolic blood pressure of-0.75 mm
Hg for every 10 |ig/g increase in bone Pb concentration over a range of <1 to 52 |ig/g.
• Although females often show lower Pb coefficients than males, and Blacks higher Pb
coefficients than Whites, where these differences have been formally tested, they are
usually not statistically significant. The tendencies may well arise in the differential Pb
exposure in these strata, lower in women than in men, higher in Blacks than in Whites.
The same sex and race differential is found with blood pressure.
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Though genotyping has not yet produced results predicting differential cardiovascular
response to Pb, this field has potential to identify individuals at higher risk of adverse Pb
effects.
6.6 REPRODUCTIVE AND DEVELOPMENTAL EFFECTS OF LEAD
6.6.1 Summary of Key Findings of the Reproductive and Developmental
Effects of Lead from the 1986 Lead AQCD
Lead has been implicated as a risk factor for reproductive outcomes for over a century
Rom, 1976; Oliver, 1911). As early as 1860, increased rates of stillbirths and spontaneous
abortions were found in women with occupational Pb exposure (usually in the ceramics industry)
and in women with husbands employed in the Pb industry, compared to unexposed women
(Rom, 1976). Other early investigations found increased rates of physically and mentally
"retarded" offspring among these same groups. In 1910, these findings resulted in the first
Pb-related occupational regulation; the British Committee on Occupational Health recommended
that women not be employed in the Pb industry (Oliver, 1911). These observations, however,
were based on exposure levels far above those considered acceptable today, and current research
now focuses on substantially lower exposure levels.
The 1986 Lead AQCD provided evidence that Pb, at high exposure levels, exerted
significant adverse health effects on male reproductive functions. Several studies observed
aberrations in both sperm count and morphology in men occupationally exposed to relatively
high levels of Pb (blood Pb levels of 40-50 jig/dL). However, the effects of Pb on female
reproductive function and fetal growth were suggestive but equivocal, perhaps due to the small
sample sizes and inadequate controlling for potential confounding factors.
This section provides a critical review of the literature regarding the associations between
exposure to environmental Pb and reproductive outcomes. First, the evidence for the placental
transfer of Pb is reviewed; this is key to providing a basis and mechanism for fetal exposure.
Second, associations between Pb exposure and various outcomes are reviewed. Outcomes of
interest are reproductive function (fertility), spontaneous abortion, fetal growth, preterm delivery,
and congenital anomalies. Each section below begins with a summary of the literature up to
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1986, the year of the last Lead AQCD. Then, key studies are reviewed and each section ends
with a conclusion based on the evidence provided.

6.6.2 Placental Transfer of Lead
In 1969, Barltrop (1969) demonstrated that Pb crosses the placenta beginning as early as
gestational week 12 and that the transfer rate then increased to term. Pb accumulations were
found in the bones, livers, blood, hearts, kidneys, and brains of stillborn and spontaneously
aborted fetuses. These observations were replicated by numerous investigators. For example,
Casey and Robinson (1978) found Pb accumulations in the livers, kidneys, and brains of stillborn
fetuses. Lead accumulations were also found in the livers, brains and kidneys of first trimester-
aborted fetuses (Chaube et al., 1972), suggesting placental transfer of Pb earlier than 12 weeks of
gestation. Newer findings, published since 1986, are summarized in Annex Table AX6-6.1 and
are assessed below.
Placental transfer of Pb is confirmed by correlations of maternal blood Pb, umbilical cord
blood Pb, and placental Pb concentrations in a variety of settings. Umbilical cord blood reflects
fetal blood. Early studies, prior to 1986, found correlation coefficients between maternal and
umbilical cord blood Pb ranging from 0.5 to 0.8 (all highly statistically significant). More recent
studies also found significant correlations between maternal and fetal blood Pb. For example, a
prospective study in Kosovo, Yugoslavia recruited 1,502 women at mid-pregnancy in two
towns—one with high exposure due to the presence of a Pb smelter, refinery, and battery plant,
and one with relatively low Pb exposure. The correlation between maternal blood Pb (either at
delivery or at mid-pregnancy) and cord blood Pb ranged from 0.8 to 0.9 (Graziano et al., 1990).
Among women with substantially lower levels of exposure (e.g., blood Pb 1.9 |ig/dL) the
correlation between maternal and cord blood Pb was 0.79 (Harville et al., 2005).
Chuang et al. (2001) propose that while maternal and cord whole blood Pb are highly
related, fetal exposure may be even more influenced by maternal plasma Pb. Using data from a
cohort of 615 women in Mexico City recruited in 1994-1995, these investigators used structural
equation modeling to estimate the associations of cord blood Pb with whole blood Pb, bone Pb
(cortical and trabecular), and the latent variable, plasma Pb. They found that maternal plasma Pb
had a stronger association with cord blood Pb compared to maternal whole blood Pb.
The greatest contributors to plasma Pb were bone Pb and airborne Pb. However, with declining
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exogenous Pb exposure, these investigators note that the measurement of plasma and bone Pb
may become increasingly important in assessing fetal exposure.
These data provide little doubt of fetal exposure to Pb via placental transport. Further, it
appears that Pb crosses the placenta throughout pregnancy, leading to continual exposure of the
fetus. Indeed, there is evidence to suggest that maternal blood Pb levels during the later half of
gestation increase (Gulson et al., 2004; Hertz-Picciotto et al., 2000; Rothenberg et al., 1994;
Sowers et al., 2002). The magnitude of the increase ranges from 14 to 40%, possibly due to the
different starting blood Pb in each study (Bellinger, 2005). The increase in blood Pb in the later
half of pregnancy may result from physiologic changes in maternal homeostasis during
pregnancy and, in particular, to mobilization of Pb stores from other body organs (Bellinger,
2005). Indirect evidence for such mobilization comes from the increased rate of bone turnover
during the later half of gestation, prompted by the increased fetal need for calcium (Moline et al.,
2000). Thus, both the epidemiological evidence and the biological plausibility of the
associations support the role of maternal-fetal transfer of Pb.
Additionally, in populations with greater Pb burdens, the fetus may be at even greater
increased risk for exposure and possible adverse effects of exposure. Among the variables
associated with Pb exposure in pregnant (and nonpregnant) women are: smoking and alcohol
consumption (Graziano et al., 1990; Rhainds and Levallois, 1997), pica (Rothenberg et al.,
1999), use of ethnic remedies and cosmetics (Al-Ashban et al., 2004; CDC, 1993), and food
preparation in inappropriately Pb-glazed pottery (Azcona-Cruz et al., 2000; Rothenberg et al.,
2000). There is some evidence that low calcium intake is also associated with higher blood Pb
(Gulson et al., 2004; Hernandez-Avila et al., 2003; Hertz-Picciotto et al., 2000). Finally, the
location where the mother resides (or resided as a child) may increase blood Pb (Graziano et al.,
1990). Blood Pb levels are elevated among U.S. immigrants, especially those who migrated
from countries where Pb is still used as a gasoline additive (CDC, 2000); indeed, blood Pb levels
are inversely associated with the number of years since migration (CDC, 2000; Klitzman et al.,
2002; Rothenberg et al., 1999).
In conclusion, the epidemiologic evidence indicates that Pb freely crosses the placenta,
resulting in continued fetal exposure throughout pregnancy. Indeed, the evidence is strong that
exposure increases during the later half of pregnancy. Exposure to the fetus is more pronounced
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in high-risk populations, especially those who migrated from countries still using Pb as a
gasoline additive.

6.6.3 Effects of Lead on Reproductive Function
6.6.3.1 Effects on Male Reproductive Function
Male reproductive function is measured using the reproductive history of the male
(i.e., number of pregnancies fathered), time to pregnancy and direct measures of semen quality
(usually sperm count, motility and morphology). Most studies relating Pb exposure to male
reproductive function are based on data collected in the occupational setting linked to population
birth registries and on studies directly collecting questionnaire exposure and outcome data.

6.6.3.1.1 Sperm Count, Motility and Morphology
Recent publications which purport a decline in sperm concentration, motility, and
morphology seek the explanation in the rising use of man-made chemical endocrine disrupters
(Auger et al., 1995; Fisch et al., 1997; Farrow, 1994; Gyllenborg et al., 1999; Kavlock et al.,
1996; Keiding et al., 1994; Keiding and Skakkebaek, 1996; Lerchl, 1995; Olsen et al., 1995;
Sherins, 1995). Several studies from the 1970s and early 1980s suggest aberrations in both
sperm count and morphology in men exposed to relatively high levels of Pb. Results from these
studies as well as more recent studies are summarized in Annex Table AX6-6.2. In the earliest
study, Lancranjan et al. (1975) found decreased sperm counts and an increased prevalence of
morphologically abnormal sperm among workers heavily exposed to Pb (mean blood Pb
74.5 |ig/dL) as well as those moderately exposed (mean blood Pb 52.8 jig/dL). These findings
have been corroborated by results of studies in the United States (Cullen et al., 1984) and Italy
(Assennato et al., 1986) which describe similar effects in workers with blood Pb levels above
60 |ig/dL.
More recently, corroborating data was described in a comprehensive review by Apostoli
et al. (1998). In studies of men with blood Pb levels above 40 |ig/dL, decreases in sperm count
and concentration, motility and morphologic aberrations were found. Chowdhury et al. (1986)
found a significant decrease in sperm count and motility and an increase in the number of sperm
with abnormal morphology in 10 men with occupational Pb exposure; the average blood Pb in
the exposed group was 42.5 |ig/dL compared to 14.8 |ig/dL in the unexposed. Similar results
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were found in a group of 30 Pb-exposed factory workers compared to controls (Lerda, 1992).
In a large study of male Pb smelter workers, Alexander et al. (1996a) found a decreasing trend of
sperm concentrations with increasing Pb exposure. In this cohort, 152 workers provided blood
specimens and 119 also provided semen samples. Geometric mean sperm concentrations were
79.1, 56.5, 62.7, and 44.4 million cells/mL for blood Pb levels of <15, 15-24, 25-39, and
>40 |ig/dL, respectively. Long-term body Pb burden was estimated from current blood Pb
concentrations and historical blood Pb monitoring data. Using this measure of long-term Pb
body burden, a similar trend was found for sperm concentration, total sperm count, and total
motile sperm count. No associations were found for sperm morphology or serum concentrations
of reproductive hormones. A study of traffic police in Peru, where leaded gasoline is still in use,
found decreases in sperm morphology, concentration, motility and viability among men with
blood Pb >40 |ig/dL compared to men with blood Pb <40 |ig/dL.
Using data from an international study of 503 men employed in the Pb industry,
Bonde et al. (2002) considered the lowest adverse effect level associated with perturbed semen
parameters. Median sperm concentration was reduced by 49% in men with blood Pb >50 |ig/dL;
regression analysis indicated a threshold value of 44 |ig/dL. These investigators conclude that
adverse effects on sperm quality were unlikely at blood Pb levels <45 |ig/dL.
In a population of couples undergoing either artificial insemination or in vitro fertilization,
Benoff et al. (2003a,b) found higher concentrations of Pb in seminal fluid in the male partner
among couples who did not conceive, compared to those who did conceive. While not directly
measuring the adverse effects of Pb on sperm per se, these data suggest a possible mechanism for
the transfer of Pb from paternal exposure to the fetal environment. Hernandez-Ochoa et al.
(2005) also provide evidence that Pb concentrations in seminal fluid may be a better indicator of
exposure than blood Pb. Mean blood Pb in this sample was lower than in most other studies,
9.3 |ig/dL. Decreases in sperm concentration, motility, morphology, and viability were
correlated with seminal fluid Pb or Pb in spermatozoa, but not with blood Pb.
Overall, the available evidence suggests a small association between exposure to Pb,
usually in the workplace, and perturbed semen quality. It appears that sperm count and
morphology (% normal forms) may be decreased at exposures >45 |ig/dL. Future research
should focus on studies of men exposed to lower levels of Pb, as exposures in the very high
range are associated primarily with occupational exposure. These studies should also account for
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variables known to be associated with semen quality and which may also be associated with
exposure, e.g., social class, other environmental exposures such as heat and vibration, and
lifestyle variables such as cigarette smoking and alcohol use.

6.6.3.1.2 Time to Pregnancy
Time to pregnancy represents a sensitive measure of fecundity. Time to pregnancy is
important because it measures the end effect of perturbed reproductive function. While it is
important and necessary to understand the associations between prenatal exposures and
endocrine abnormalities and semen characteristics, they represent possible antecedents to the
occurrence of pregnancy. Previous reports demonstrate good validity and reliability for reports
of time to pregnancy in both males and females and when time to recall has been both long and
short (Weinberg et al., 1993, 1994).
One advantage to the use of this parameter, as compared to just an infertility measure, is
that it does not require categorization of men into fertile and infertile groups. Among couples
that succeed in establishing pregnancy, there is considerable variability in the time between
discontinuation of contraception and conception (Weinberg et al., 1994). With the possible
exception of cigarette smoking and age, little is known regarding such intercouple variability.
Delays in time to pregnancy may be indicative of a range of reproductive abnormalities of both
partners, including impaired gametogenesis, hormonal disruptions, and very early unrecognized
pregnancy loss. Time to pregnancy has the menstrual cycle as its natural unit and is thus
measured in integer units of menstrual cycles.
Usually, time to the most recent pregnancy is taken as the outcome (Baird et al., 1986).
The measure of exposure in these studies usually is the fecundity density ratio, which is similar
to an incidence density ratio. Fecundity density ratios can be interpreted as the risk of pregnancy
among the exposed during an interval, compared to the risk of pregnancy among the unexposed
during the same interval. In such studies, the intervals of interest are menstrual cycles.
Fecundity density ratios less than one indicate reduced fecundity (i.e., longer time to pregnancy)
among the exposed compared to the unexposed, while those greater than one indicate enhanced
fecundity (i.e., shorter time to pregnancy) in the exposed. Usually fecundity density ratios are
calculated using discrete time Cox proportional hazards regression models.
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Several recent studies evaluate time to pregnancy when the male partner is occupationally
exposed to Pb. These studies are summarized in Annex Table AX6-6.3 and reviewed here.
The Asclepios Project, a large European collaborative cross-sectional study, evaluated time to
pregnancy in 1,108 men of whom 638 were exposed to Pb (Joffe et al., 2003). The reference
group consisted of Pb workers for whom exposure did not coincide with time of pregnancy.
The investigators only included pregnancies which resulted in live births. Fecundity density
ratios were 1.12 (95% CI: 0.84, 1.49), 0.96 (95% CI: 0.77, 1.19), 0.88 (95% CI: 0.70, 1.10) and
0.93 (95% CI: 0.76, 1.15) for blood Pb levels <20, 20-29, 30-39, and >40 |ig/dL, respectively.
These results indicate that no association was found between blood Pb and delayed time to
pregnancy. Similar results were found when duration of exposure or cumulative exposure was
used as the exposure metric.
A separate report was published in the Italian group of men included in the Asclepios
project (Apostoli et al., 2000). Blood Pb at the time closest to conception was used as the
measure of exposure. Lead-exposed men (n = 251) who had experienced at least one completed
pregnancy were compared to nonexposed men (n = 45). Contrary to what was expected, time to
pregnancy was significantly shorter among couples in which the male partner was exposed to Pb
compared to those in which the male partner was not exposed. In secondary analyses, time to
pregnancy was longer among men with the highest blood Pb (i.e., >40 |ig/dL). Limiting
the analysis solely to exposed men, time to pregnancy was longer among men with higher
blood Pb levels.
Among 502 couples identified by Sallmen et al. (2000) from the Finnish Institute of
Occupational Health in which the male partner was exposed to Pb, time to pregnancy was
reduced among those with blood Pb >10 |ig/dL compared to those with blood Pb < 10 |ig/dL.
However, when blood Pb was stratified, no concentration-response relationship was found.
Fecundity density ratios were 0.92 (95% CI: 0.73, 1.16), 0.89 (95% CI: 0.66, 1.20), 0.58
(95% CI: 0.33, 0.96) and 0.83 (95% CI: 0.50, 1.32) for exposures of 10-20, 21-30, 31-40, and
>40 |ig/dL, respectively. In this study, blood Pb concentrations close to the time of conception
were available on 62% of men, while in 38% it was estimated using blood Pb levels obtained at
other points or based on job descriptions.
Among 280 pregnancies in 133 couples in which the male partner was employed in a
battery plant, 127 were conceived during exposure while the remainder conceived prior to
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exposure (Shiau et al., 2004). Time to pregnancy increased with increasing blood Pb, especially
when blood Pb levels were >30 |ig/dL. Fecundity density ratios were 0.50 (95% CI: 0.34, 0.74)
and 0.38 (95% CI: 0.26, 0.56) for blood Pb levels of 30-39 and >39 |ig/dL, respectively.
In 41 couples, one pregnancy occurred prior to exposure and one during exposure—time to
pregnancy during exposure was significantly longer. Of note, this is the only study to estimate
decreases in time to pregnancy when blood Pb was below 40 |ig/dL; time to pregnancy increased
by 0.15 months for each 1 |ig/dL increase in blood Pb between 10 and 40 |ig/dL.

6.6.3.1.3 Reproductive History
Since the 1986 Lead AQCD, several studies examining the association of Pb with
reproductive history have been published. Results from these studies are presented in Annex
Table AX6-6.4. Population-based birth registries in the Scandinavian countries provide data on
medically diagnosed pregnancies. These registries provide a basis for linking occupational data
on Pb exposure obtained by place and duration of employment or by direct measures of blood Pb
relative to the timing of marriage or conception. Using a roster of men employed in three battery
plants in Denmark, Bonde and Kolstad (1997) matched all births to the 1,349 employees when
they were age 20-49 years. A control group of 9,656 men who were not employed in a Pb
industry was chosen. No associations were found between employment or, among those
employed in the Pb industry, duration of employment in the Pb industry and birth rate.
A similar study in Finland (Sallmen et al., 2000) examined the association between
conception and blood Pb among men monitored for occupational exposure at the Finnish
Institute of Occupational Health (n = 2,111). Men were categorized as probably exposed and
possibly exposed based on their measured blood Pb in relation to the time of marriage.
A nonexposed group of 681 men with blood Pb < 10 |ig/dL was similarly evaluated. Among men
in the probable exposure group, the risk of failing to achieve a pregnancy increased with
increased blood Pb in a monotonic concentration-response fashion. Compared to the
nonexposed, the risk ranged from 1.3 to 1.9 for blood Pb levels 10-20 |ig/dL and >50 |ig/dL,
respectively.
Lin et al. (1996) linked records from the Heavy Metal Registry in New York State to birth
certificates from the New York State Office of Vital Statistics for the period 1981 to 1992.
Exposure was defined as having at least one blood Pb measurement above 25 |ig/dL and
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identified 4,256 men. A reference group of 5,148 men was frequency matched for age and
residence. The exposed group had fewer births than expected, and was especially pronounced
among men employed in the Pb industry for over 5 years.
Among 365 men occupationally exposed to metals, Gennart et al. (1992) identified
74 exposed continuously for more than 1 year and with at least one blood Pb measurement
>20 |ig/dL. Compared to a reference group with no occupational exposure, the probability of at
least one live birth was significantly reduced. Fertility decreased with increasing duration of
exposure but no concentration-response relationship with blood Pb was found (possibly due to
the small sample size of exposed men).
A study of men exposed to Pb in a French battery plant (Coste et al., 1991) reported no
effect on fertility. However, this study did not adequately control for potentially confounding
variables, particularly those related to the women. Further, nonexposed workers were defined as
those with no recorded blood Pb values, which likely resulted in exposure misclassification.
One potential mechanism to explain the associations between Pb exposure and male
reproductive outcomes may be through an effect of Pb on circulating pituitary and testicular
hormones. Several studies have evaluated this hypothesis in groups of workers (Braunstein
et al., 1978; Cullen et al., 1984; Erfurth et al., 2001; Ng et al., 1991; Rodamilans et al., 1988).
Further discussions on the effect of Pb on male reproductive hormones are presented in
Section 6.9.3.3.1. In general these studies find perturbations in concentrations of follicle
stimulating hormone, luteinizing hormone, and testosterone. Although many of these studies
were limited by small sample sizes, lack of control groups, and mixtures of exposures, taken
together, they provide evidence for this possible mechanism.

6.6.3.1.4 Genotoxicity and Chromosomal Aberrations
The potential genotoxicity and ability to induce chromosomal aberrations speak to the
mechanisms by which Pb is a potential reproductive toxin. Two possible mechanisms by which
Pb may affect reproduction are through affinity with proteins and ability to mimic the actions of
calcium (Silbergeld et al., 2000).
Data from occupational studies regarding the effects of Pb on chromosomes are
contradictory; however, the bulk of evidence suggests that there may indeed be a genotoxic
effect. Early studies in occupational groups find associations between Pb exposure and increased
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frequency of sister chromatid exchanges (Grandjean et al., 1983; Huang et al., 1988; Leal-Garza
et al., 1986; Maki-Paakkanen et al., 1981). Similar results were found in a group of
environmentally-exposed children with blood Pb levels ranging from 30 to 63 |ig/dL (Dalpra
et al., 1983). Increased frequencies of chromosomal aberrations, particularly chromatid
aberrations, were found in battery plant workers and were correlated with increased blood Pb
(Huang et al., 1988). A more marked increase was found when blood Pb were above 50 |ig/dL.
Other occupational studies find similar associations (Al-Hakkak et al., 1986; Forni et al., 1976,
1980; Nordenson et al., 1978; Schwanitz et al., 1970). Other studies find no evidence of
chromosomal aberrations when blood Pb ranged from 38 to 120 |ig/dL (Bauchinger et al., 1977;
Maki-Paakkanen et al., 1981; O'Riordan and Evans, 1974; Schmid et al., 1972; Schwanitz et al.,
1975). More recently, two studies in battery plant workers (mean blood Pb 40.1 |ig/dL) and
controls (mean blood Pb 9.8 |ig/dL) found an increase in high-frequency cells and sister
chromatid exchanges among the workers, indicating the cytogenetic toxicity of Pb (Duydu et al.,
2001, 2005). An increase in sister chromatid exchanges, although not statistically significant,
was also found in individuals exposed to Pb and/or alcohol and tobacco (Rajah and Ahuja, 1995,
1996). In the Lithuanian populations exposed to either environmental or occupational Pb, a
higher incidence of sister chromatid exchanges and chromosomal aberrations was found
(Lazutka et al., 1999), although these populations were also exposed to other potentially
genotoxic substances. Recent data also indicates that Pb may inhibit DNA repair responses
among Pb-exposed workers (Karakaya et al., 2005).
Occupational exposure to Pb, particularly when blood Pb was high (i.e., >40 jig/dL), was
associated with increased mitotic activity in peripheral lymphocytes and with an increased rate of
abnormal mitosis (Forni et al., 1976; Minozzo et al., 2004; Sarto et al., 1978; Schwanitz et al.,
1970). Again, to the extent these changes influence the production of gametes, this is a potential
mechanism explaining associations between Pb exposure and decreased male fecundity.

6.6.3.1.5 Issues Concerning Studies of Male Fecundity Related to Lead Exposure
In examining studies of fecundity and fertility, several issues relating to interpretation and
bias must be addressed. Infertility usually is defined as 12 months of continuous unprotected
intercourse without pregnancy. Fecundity represents both a characteristic of the individuals and
a characteristic of a couple, meaning that both partners must be biologically able to procreate.
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Thus, one possible explanation for observations of reduced fecundity related to occupational Pb
exposure in the male partner is the exposure he "takes home" via transport of dust on clothing
and shoes, ultimately resulting in an effect related to the female partner. Other possible
interpretations need to account for measurement error, especially related to the outcomes of
reproductive history and time to pregnancy, bias in the selection of subjects for study, and the
control for potentially confounding variables.
Both reproductive history and time to pregnancy are subject to errors of recall and rely on
the veracity of the subject. Several studies have evaluated recall and veracity of the male partner
using the female partner as the "gold standard." In general, these find good reliability between
the male and female (Weinberg et al., 1993, 1994). Nevertheless, it is possible, at least for
studies using men as the sole informant, that the number of pregnancies a man has fathered is
underreported. If reporting is nondifferential with regard to Pb exposure, then associations will
generally be biased towards the null value; however, since characteristics such as social
circumstances, ethnicity, and age may affect both exposure and reporting, it is difficult to
evaluate the role of bias.
It was not clear from many of the studies that men with medical conditions which affect
fecundity/fertility were excluded. Further, several prescription and over-the-counter medications
also affect fecundity as does a history of surgery in the genital area (e.g., varicocele). To the
extent that these conditions are related to the absence of employment in Pb-industries, then the
results may be subject to a type of "healthy worker" effect. Because it is unclear whether many
of these studies asked about these conditions, this cannot be ruled out as a possible source
of bias.
In retrospective studies it is often useful to use the outcome of the most recent pregnancy
in the primary analysis. The reason for this is to reduce any possible recall bias. This type of
bias may also be an issue in studies which use occupational registry data, i.e., men may have
fathered an additional pregnancy after employment in the industry ceased.
Variables considered potential confounders in studies of fertility and fecundity include
sociodemographic characteristics (e.g., age, ethnicity, education, occupation); prenatal and recent
lifestyle variables such as cigarette smoking, alcohol use, and medication use; exposures through
occupation and hobby, and recent medication use. Also important in these studies is control for
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factors which may affect the partner's fertility, e.g., cigarette smoking. Many of the studies
reviewed did not carefully measure or adjust for confounding variables.
The issues presented above potentially limit the interpretation of results from studies
examining the association of Pb exposure with male fecundity and fertility. Nevertheless, most
studies find small associations between Pb exposure at high levels (i.e., >45 |ig/dL) and slightly
reduced male fecundity or fertility.

6.6.3.2 Effects on Female Reproductive Function
Few data directly address the effects of Pb exposure on fecundity in the female. A recent
retrospective study of time to pregnancy among wives of Pb workers provides limited support
that Pb exposure is associated with increased time to pregnancy. Fecundity density ratios were
0.92 (95% CI: 0.72, 0.16), 0.89 (95% CI: 0.66, 1.20), 0.58 (95% CI: 0.33, 0.96), and 0.83
(95% CI: 0.50, 1.32) for blood Pb in the male partners of 10-20, 21-30, 31-38 and >39 |ig/dL
compared to <10 |ig/dL, respectively. Note however, that exposure here is measured in the male
partners and not the females.
Time to pregnancy was evaluated in 121 women biologically monitored for Pb exposure
at the Finnish Institute of Occupational Health between 1973 and 1983 (Sallmen et al., 1995).
Fecundity did not differ with level of exposure (defined as <10 |ig/dL, 10-19 |ig/dL and
>20 |ig/dL), but among women with blood Pb between 29 and 50 |ig/dL, there was a suggestion
of reduced fecundity (longer time to pregnancy). However, only a small number of subjects
(n = 8) were exposed in this range.
In the limited number of studies, there is little evidence regarding the associations
between Pb exposure and fertility in the female to draw any conclusions at this time.

6.6.4 Spontaneous Abortion
6.6.4.1 Spontaneous Abortion and Maternal Exposure to Lead
Historical observations suggest increased rates of spontaneous abortion among
Pb-exposed women, particularly those employed in cottage industries (Rom, 1976). Two early
studies in a smelter town in Sweden (Nordstrom et al., 1978a, 1979) suggest elevated rates of
spontaneous abortion among female employees at the smelter and among female residents living
in close proximity to the smelter. Neither of these studies used biological markers of
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Pb exposure. Moreover, the Swedish smelter study included other exposures such as arsenic,
zinc, and Cd; thus the conclusions for these analyses should be tempered.
In contrast, a prospective study in and around a smelter town in Port Pirie, Australia
(McMichael et al., 1986) did not find an association between blood Pb concentration and
spontaneous abortion. However, it was likely that complete ascertainment of spontaneous
abortions was not obtained (Rowland and Wilcox, 1987), since most women were recruited for
this study after the first trimester of pregnancy. A retrospective cohort study in two towns in the
former Yugoslavia (Murphy et al., 1990) showed no associations between Pb exposure and
spontaneous abortion in the first reported pregnancy. One of these towns was a smelter town
with relatively high Pb exposure (at recruitment during mid-pregnancy, the mean blood Pb
concentration was 17.1 |ig/dL, while in the control town the mean blood Pb was 5.1 |ig/dL).
A similar study in Poland (Laudanski et al., 1991) evaluated the association between Pb-exposed
and nonexposed areas for their reproductive histories. Among women in the exposed areas,
11% reported having at least one prior spontaneous abortion, compared to 19.5% of women in
the unexposed areas.
Two studies in Finland (Lindbohm et al., 1991a; Taskinen, 1988) used hospital registry
data to ascertain women with either spontaneous abortions or livebirths. Either maternal job
histories (Taskinen, 1988) or both maternal and paternal job histories were obtained from a
registry of occupational blood Pb measurements. Neither study found evidence of an association
between maternal Pb exposure and spontaneous abortion. In the Lindbohm et al. (1991a) study,
maternal exposure was extrapolated from the occupation of the father.
In Bulgaria, pregnant women residing in or near Pb smelting areas or petrochemical plants
were prospectively followed for pregnancy outcomes (Tabacova and Balabaeva, 1993).
The investigators compared blood Pb in those women with spontaneous abortions and those
without. Blood Pb concentrations in cases were significantly higher than in controls (mean
blood Pb 7.1 |ig/dL versus 5.2 |ig/dL, respectively). However, this study did not fully describe
the selection of women nor the definition for cases.
Women employed by the U.S. Forest Service and exposed to Pb-based paint (to mark
trees for clearing) were studied using self-reported questionnaires (Driscoll, 1998). Adjustment
was made for potential confounders and generalized estimating equations were used to adjust for
multiple pregnancies per woman. Significant associations were found for three types of paint
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containing Pb pigment (odds ratios of 4.3 [95% CI: 2.0, 9.3], 2.0 [95% CI: 1.2, 3.3] and
1.8 [95% CI: 1.2, 2.6]). While these findings are intriguing, the response rate was only 59%
(with no evaluation of selection bias) and the paint also contained solvents thought to be
associated with spontaneous abortions.
Borja-Aburto et al. (1999) examined the association between blood Pb concentrations and
spontaneous abortions in a nested case-control study using incidence density methods and
matching for age, calendar time of study entry, public versus private clinic, and gestational age at
study entry. They ascertained 668 women during the first trimester of pregnancy in Mexico
City. After contacting women biweekly to update pregnancy status, they found 35 cases (6.4%)
of spontaneous abortion among women not lost to follow up. An odds ratio of 1.8 (95% CI:
1.1,3.1) per 5 |ig/dL increase in blood Pb was observed after adjustment for spermicide use,
active and passive smoking, use of alcohol and coffee, maternal age, education, income, physical
activity, hair dye use, use of video display terminals, and medical conditions. Mean blood Pb in
cases (12.0 |ig/dL, range 3.1-29 |ig/dL) was slightly higher than in controls (10.1 |ig/dL, range
1.3-26 |ig/dL). Further, after categorizing blood Pb into 5 |ig/dL intervals, a concentration-
response relationship was evident.
More recently, a small study of 57 female workers in a battery plant in China and
62 controls found that 6 spontaneous abortions occurred in the exposed group, compared to none
in the controls (Tang and Zhu, 2003). A long-term follow-up of survivors of acute plumbism
(Hu, 1991) found increased risk of spontaneous abortions or stillbirths (odds ratio of 1.6
[95% CI: 0.6, 4.0]). Although the study was based on small numbers, the data suggest a
persistent association between childhood exposure and outcomes later in life.
A review of eight studies (Borja-Aburto et al., 1999; Driscoll, 1998; Laudanski et al.,
1991; Lindbohm et al., 1991a; McMichael et al., 1986; Murphy et al., 1990; Tabacova and
Balabaeva, 1993; Taskinen, 1988) evaluating maternal exposure to Pb (blood Pb >30 |ig/dL) and
spontaneous abortion concluded that there was little evidence that Pb exposure at these relatively
high levels was associated with an increased risk in spontaneous abortions (Hertz-Picciotto,
2000). However, Hertz-Picciotto also concluded that methodological difficulties in most of these
studies (i.e., small sample sizes, inadequate ascertainment of outcome, and possible residual
confounding) limited the confidence in these findings. Further, she noted that exposure in many
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of these studies was either measured in an ecologic fashion or biological measures were
available, but they were not ascertained during a biologically meaningful period.
Collectively, there is little evidence to support an association between Pb exposure in the
female and spontaneous abortion. The only well-designed study which found an association is
that of Borja-Aburto et al. (1999); however, these results need to be confirmed in other
populations. Studies of spontaneous abortion need be done carefully to avoid possible bias due
to recall, use of pregnancies other than the first, and confounding. Retrospective studies, for
example, should take full pregnancy histories, including probing for spontaneous abortions
versus induced abortions versus stillbirths. In some cultures, for example, induced abortions are
frowned upon and women may report spontaneous abortions instead. Additionally, some women
may confuse a stillbirth with spontaneous abortion, especially if she is unable to adequately date
her pregnancy using date of last menstrual period. Although use of the most recent pregnancy
may curtail problems of recall, other concerns dictate that the first pregnancy be used in studies
of spontaneous abortion because the risk of subsequent spontaneous abortion depends on the
history of spontaneous abortion. Finally, while few variables are known confounders of this
relationship, the following should be controlled: maternal age, education and other SES
indicators, cigarette smoking, and alcohol use. Several studies of spontaneous abortion did not
properly adjust for these potentially confounding variables.
One final concern regards the type of spontaneous abortion. Very early spontaneous
abortions, i.e., before a clinical pregnancy is diagnosed, may be missed; assuming, however, that
both exposed and unexposed women have the same rates of early spontaneous abortions, this
would bias the association towards the null. Indeed, this may be true, as many very early
spontaneous abortions may be chromosomally abnormal and probably not attributable to Pb
exposure.

6.6.4.2 Spontaneous Abortion and Paternal Exposure to Lead
Three studies evaluated paternal exposure to Pb and spontaneous abortion. Lindbohm
et al. (1991a), using national databases to identify pregnancy outcomes among 99,186 births in
Finland, found no association between paternal employment in jobs with Pb exposure and
spontaneous abortion (odds ratio of 0.9 [95% CI: 0.9, 1.0]). In a follow up case-control study
(Lindbohm et al., 1991b), they ascertained paternal exposure status during the period of
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spermatogenesis in 213 cases of spontaneous abortion and 500 controls. Exposure was
ascertained using blood Pb concentrations measured during spermatogenesis for 6% of men;
for the remaining 94%, exposure was estimated using a regression model where the independent
variables were blood Pb levels measured either prior to or after the period of spermatogenesis.
Blood Pb (either measured or estimated) was not associated with spontaneous abortion.
However, when the analysis was restricted to men with measured blood Pb, blood Pb levels
>30 |ig/dL were associated with an increased odds of spontaneous abortion (odds ratio of
3.8 [95% CI: 1.2, 2.0]); but, this result was based only on 12 cases and 6 controls.
The third study (Alexander et al., 1996b) found no association between men employed in
a Pb smelter and spontaneous abortion. For men with "moderate" exposure jobs the estimated
odds ratio was 0.8 (95% CI: 0.5, 1.5) and for those with "high" exposure jobs, the estimated
odds ratio was 1.4 (95% CI: 0.7, 2.5). Further when blood Pb 1 year prior to the pregnancy was
used as the exposure measure, no increased odds of spontaneous abortion were found. These
results, however, are based on a low participation rate in eligible workers (37%) and should be
interpreted with caution. Overall, the available studies provide little evidence for an association
between Pb exposure in the male and spontaneous abortions.

6.6.5 Fetal Growth
The results of epidemiologic studies regarding the association between Pb exposure and
birth weight are inconsistent. Cross-sectional studies (Clark, 1977; Gershanik et al., 1974;
Moore et al., 1982; Rajegowda et al., 1972) did not find significant correlations between blood
Pb and birth weight, nor did a study using placental Pb as the exposure variable (Wibberley
et al., 1977). A case-control study (Bogden et al., 1978) comparing 25 low birth weight babies
(1,500-2,500 grams) to 25 controls (>2,500 grams) matched on maternal age, race and social
class found a small, nonsignificant difference in maternal and cord blood Pb levels. Mean
maternal blood Pb concentrations were 16.2 + 4.5 |ig/dL and 15.3 + 5.2 |ig/dL and mean cord
blood Pb levels were 13.8 + 4.4 |ig/dL and 13.1 + 4.3 |ig/dL in cases and controls, respectively.
A further study (Huel et al., 1981) found no differences in maternal and fetal hair Pb levels
between infants born small-for-gestational-age compared to those of normal birth weight.
In 1984, Needleman et al. (1984) reported on a cross-sectional study of 5,183 births of at
least 20 weeks gestation in Boston, MA. No associations were found between the proportion of
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births under 2,500 grams and cord blood Pb. Exposure levels in this study were relatively low
for the time; cord blood Pb ranged from <1 to 35 |ig/dL. A reanalysis of these data found no
relationship between cord blood Pb and birth weight when birth weight was considered as a
continuous variable (Bellinger, et al., 1991). However, when birth weight was categorized as
low birth weight (<2,500 grams), small for gestational age (<10thpercentile for gestational age),
or intrauterine growth retarded (>2 standard deviations below the mean for gestational age),
relative risks of 1.6 (95% CI: 1.0, 2.6), 1.2 (95% CI: 0.8, 1.6) and 1.9 (95% CI: 1.0, 3.4),
respectively, were found for each 10 |ig/dL increase in cord blood Pb levels. Increased relative
risks also were found for cord blood Pb levels > 15 |ig/dL, compared to cord blood Pb levels
<15 |ig/dL; however, only 83 of the 5,183 women had exposures in the high range, resulting in
imprecise estimates. These data suggest that Pb-related modest reductions in birth weight are
perhaps plausible when birth weight is expressed as a function of gestational age.
The prospective study of Pb exposure in and around Port Pirie, Australia (McMichael
et al., 1986) followed 749 pregnancies of at least 20 weeks duration. Mean maternal blood Pb
levels at mid-pregnancy were 10.1 |ig/dL and 7.0 |ig/dL for women residing in Port Pirie and the
surrounding communities, respectively. After excluding 9 sets of twins and 10 cases for which
the maternal last menstrual period could not be ascertained, no relationship was found between
either cord blood Pb or maternal blood Pb measured at mid-pregnancy or at delivery and birth
weight in a multivariate regression model controlling for known determinants of birth weight.
A prospective study in two towns in Kosovo, Yugoslavia evaluated relationships between
birth weight (adjusted for gestational age using last menstrual period) and (a) maternal blood Pb
at mid-pregnancy and delivery and (b) cord blood Pb (Factor-Litvak et al., 1991). The towns
were vastly different in exposure patterns, as one was the site of a Pb smelter, refinery and
battery plant (n = 401, mean mid-pregnancy blood Pb 19.0 |ig/dL) and one was relatively
unexposed (n = 506, mean mid-pregnancy blood Pb 5.6 |ig/dL). No associations were found
between any of the biomarkers of Pb exposure and birth weight in either crude analyses or
analyses adjusted for potentially confounding variables.
While the aforementioned studies generally found no association between environmental
Pb exposure and birth weight, three other studies have shown large reductions in birth weight
related to Pb exposure. These studies, however, have questionable study designs. Nordstrom
et al. (1978b, 1979) in a series of ecologic analyses known as the Swedish Smelter Study, found
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significant reductions in birth weight between the offspring of women either working at or living
in close proximity to the smelter. The 125 gram deficit in birth weight among the offspring of
women living closest to the smelter was confined to those with parity of three or more, an
observation which does not appear to be biologically plausible. Moreover, the ecological nature
of the study did not allow for individual measurements of blood Pb or for control of potentially
confounding variables. Hence, while suggestive, these data do not provide strong evidence for a
causal association between Pb exposure and birth weight.
In a cross-sectional study of 100 "normal" singleton births, a negative correlation was
found between placental Pb concentration and birth weight (Ward et al., 1987). Mean placental
Pb levels in 21 infants weighing less than 3,000 grams were 2.35 + 0.9 |ig/g compared to
1.12 + 0.4 |ig/g in 10 infants weighting more than 4,000 grams. This study has several
limitations. First, no statistical adjustment was made for multiple comparisons (many exposures
were studied). Second, potentially confounding variables were not controlled. Third, only 31 of
the 100 infants, representing the extremes of the birth weight distribution, were studied. Hence,
this study also does not provide strong evidence for an association.
In Cincinnati, OH, the association between Pb exposure and birth weight was examined in
offspring of a cohort of young (mean maternal age 22.7 years) inner city women, 85% being
African-American and 86% being on public assistance, with a mean IQ of 75 (Dietrich et al.,
1987a). The mean gestational period of the neonates, as determined by physical examination,
was 39.5 weeks. A decrement in birth weight of 172 grams was associated with an increase in
blood Pb from 10 to 30 |ig/dL. Lead exposure in this group was relatively low with a mean
blood Pb of 8.0 + 3.7 |ig/dL. In a sample of women from this cohort, the interaction between
blood Pb and maternal age was significantly associated with birth weight; the effect varied from
a decrease of 64 grams for 18 year old mothers to 660 grams for 30 year old mothers, as blood
Pb rose from 10 to 30 |ig/dL (Bornschein et al., 1989). Although the Cincinnati study is highly
suggestive of an effect (especially an effect which varies by maternal age) three factors should be
considered in the interpretation of their findings. First, length of gestation was estimated by
examining the neurological and physical maturity of the neonate (Ballard et al., 1979); other
investigators find assessment of gestational age using this scale overestimates gestational age in
preterm infants (Constantine et al., 1987; Kramer et al., 1988; Shukla et al., 1987; Spinnato et al.,
1984). Second, it is possible that the association between Pb and birth weight differs by maternal
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characteristics such as race, ethnicity, and SES; however, no study has provided a population
sufficiently heterogeneous to examine this possible source of difference. Finally, it is possible
that confounding by unmeasured maternal lifestyle characteristics may account for the reported
association.
A hospital-based study of cord blood Pb and pregnancy outcomes in Quebec, Canada,
between June 1993 and January 1995 found a slight increase in cord blood Pb levels among
infants with birth weight <2,500 grams (Rhainds et al., 1999). For those infants with birth
weight <2,500 grams, the geometric mean blood Pb was 1.8 |ig/dL (95% CI: 1.6, 2.9) compared
to 1.6 |ig/dL (95% CI: 1.5, 1.7), 1.6 |ig/dL (95% CI: 1.5, 1.7), and 1.5 |ig/dL (95% CI: 1.5, 1.6)
among those with birth weights 2,500-2,990, 3,000-3,499, and >3,500 grams, respectively.
Although suggestive, the study did not control for potentially confounding variables. Also mean
levels of measured cord blood levels of mercury and organochlorine compounds were higher as
well in infants who weighed <2,500 g.
More recently, Irgens et al. (1998) using data from the Norwegian birth registry found that
women occupationally exposed to Pb (none/low compared to moderate/high) were more likely to
deliver a low birth weight infant (odds ratio of 1.3 [95% CI: 1.1,1.6]). No association was
found for paternal occupational Pb exposure. Parental occupational exposure to Pb was not
associated with low birth weight in the Baltimore-Washington Infant Study database (Min et al.,
1996), although subgroup analysis suggested that high paternal exposure may be associated with
small-for-gestational-age infants (odds ratio of 2.9 [95% CI: 0.9, 9.2]). Similar findings were
reported by Lin et al. (1998) who compared offspring of Pb-exposed workers with those of bus
drivers. No associations were reported between Pb exposure and low birth weight except among
the group of men with blood Pb levels >25 |ig/dL for over 5 years (relative risk of 3.4 [95% CI:
1.4,8.4]).
Using bone Pb as the metric of exposure, Gonzalez-Cossio et al. (1997) found
associations between tibia bone Pb (but not patella bone Pb or umbilical cord blood Pb) and
reduced birth weight. Bone Pb was measured one month after delivery. Infants with tibia bone
Pb in the highest quartile (> 15.15 jig Pb/g bone mineral) were, on average, 156 g lighter than
those in the lowest quartile (<4.50 jig Pb/g bone mineral). Further analyses of these data
(Hernandez-Avila et al., 2002) found an association between infants in the highest quintile of
tibia bone Pb and shorter birth length (odds ratio of 1.8 [95% CI: 1.1, 3.2]).
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Two studies have evaluated relationships between Pb exposure and head circumference.
In one study (Hernandez-Avila et al., 2002), among 233 women in Mexico City, high maternal
patella bone Pb was associated with increased risk of a low head circumference score at delivery
(1.02 per jig Pb/g bone mineral [95% CI: 1.01, 1.04]). Similar findings were reported by
Rothenberg et al. (1999) who found a reduction in six-month head circumference of 1.9 cm (95%
CI: 0.9, 3.0) as maternal blood Pb rose from 1 to 35 |ig/dL. This study, however was plagued by
multiple comparisons, as head circumference was measured nine times and prenatal blood Pb six
times—with only one statistically significant result being found.
Potential confounders need to be adjusted for to properly assess the relationship between
Pb exposure and fetal growth. Factors consistently associated with fetal growth include gender,
ethnic origin, maternal body build (i.e., pre-pregnancy weight, height), parity, SES, gestational
weight gain and nutritional intake during pregnancy, maternal illness, and cigarette smoking
(Kramer, 1987). Factors with less established associations include alcohol consumption (Kline
et al., 1987; Kramer, 1987) and street drug use (Kline et al., 1987; Kramer, 1987; Zuckerman
et al., 1989). To the extent that these factors are associated with blood Pb as well as with fetal
growth, they must be accounted for in the analysis.
A further problem pertains to the measure of exposure used in most of these studies.
Blood Pb concentration reflects relatively recent (i.e., in the past 90 days) exposure; thus it does
not reflect exposure over the mother's lifetime. Indeed, there is some evidence suggesting that
bone Pb, a measure of cumulative exposure, may be mobilized during pregnancy (Silbergeld,
1991). A single blood Pb measure will not reflect such mobilization, particularly if mobilization
is not constant over the course of pregnancy. Thus, in all studies, excepting that of Gonzalez -
Cossio et al. (1997), exposure may be misclassified. The effect of such misclassification will be
to strengthen the findings of studies which support the null hypothesis. For those studies which
find an association between blood Pb concentration and fetal growth, the inference would be to
higher exposure levels. It is difficult to examine the extent of this misclassification as no studies
have sufficient numbers of serial blood Pb measures to estimate the variation during pregnancy.
Studies to date are inconsistent regarding the association between Pb exposure and birth
weight. Several large prospective studies found no association (Factor-Litvak et al., 1991;
McMichael et al., 1986), while at least one (Bornschein et al., 1989) did find an association in
specific subgroups of women. However, there is limited evidence (Bellinger et al., 1991) for an
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association between Pb exposure and low birth weight (i.e., <2,500 g), small for gestational age
(i.e., <10th percentile for gestational age), and intrauterine growth retardation (i.e., >2 standard
deviations below the mean for gestational age). These prospective studies were all well-
conducted, adequately measured exposure and outcome, and controlled for potential confounding
variables. They did, however, take place in very different populations, suggesting that the
association between Pb and fetal growth may depend on the population being studied.
The Yugoslavia study (Factor-Litvak et al., 1991) took place in two towns in Kosovo,
Yugoslavia, which were divergent on exposure and somewhat comparable on other variables.
The Port Pirie study took place in a middle class area of Australia (McMichael et al., 1986).
The Boston study (Bellinger et al., 1991) took place across a range of social strata in Boston; the
exposure in the highest social group was attributable to renovation of older housing stock.
Finally, in the Cincinnati study (Bornschein et al., 1989), the study sample was comprised of
lower social class African Americans and the mean IQ of the mothers was 75. It is possible that
in this latter study, there was some unmeasured variable which accounts for the observed
interaction. Thus, the evidence suggests at most a small effect of Pb exposure on birth weight
and possibly a small association between Pb exposure and several dichotomized measures of
fetal growth.

6.6.6 Preterm Delivery
Early evidence regarding an association between environmental Pb exposure and preterm
delivery was inconsistent. In 1976, Fahim et al. found a preterm delivery rate of 13% in
254 pregnant women living near a Pb mining community in Missouri, compared to 3% in
249 women living in a control location. These investigators also found higher concentrations of
Pb in amniotic membrane, but not higher placental or cord Pb in preterm compared to term
deliveries, regardless of the women's residential locale. This observation prompted other studies
of Pb and preterm delivery.
Of the cross-sectional studies, the three which show no association employed cord blood
Pb as the exposure measure and restricted gestational age (Angell and Lavery, 1982; Bellinger
et al., 1991; Needleman et al., 1984; Rajegowda et al., 1972). In contrast, three other studies
used different exposure markers (placental Pb, maternal and cord blood Pb, and maternal and
fetal hair Pb) and found statistically significant associations (Huel et al., 1981; Moore et al.,
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1982; Ward et al., 1987). Other studies evaluated pregnancy outcomes in relation to maternal
delivery blood Pb (McMichael et al., 1986; Rahman and Hakeem, 2003).
Of the prospective studies, the Cincinnati study (Bornschein et al., 1989) found no
association between both maternal blood Pb at mid-pregnancy or maternal blood Pb during the
neonatal period (10 days post delivery) and preterm delivery. However, gestational age was
estimated by examining the neurological and physical maturity of the neonates (which tends to
overestimate gestational age) and not actual dates. In Port Pirie, Australia (McMichael et al.,
1986), a concentration-response relationship between maternal delivery blood Pb and preterm
delivery was reported. Odds ratios ranged from 2.1 to 4.4 in women with blood Pb of 7.7 to
10.6 |ig/dL and >13.5 |ig/dL, respectively, compared to those with blood Pb <7.7 |ig/dL. Savitz
et al. (1990) used data from the National Natality Survey and found an odds ratio of 2.3 (95% CI:
0.7, 7.0) between maternal occupational exposure to Pb and preterm delivery; however, the
estimated odds ratio was based on only 7 cases. In the Yugoslavia study (Factor-Litvak et al.,
1991), no associations were found between cord blood Pb or blood Pb measured at mid
pregnancy or delivery and either preterm delivery (defined as delivery <37 completed weeks) or
gestational age. A registry study in Norway (Irgens et al., 1998) which linked births between
1970 and 1993 to census-based occupation records found a slightly increased odds of preterm
delivery among moderate/high Pb-exposed women, compared to those with no or low exposure
(odds ratio of 1.13 [95% CI: 0.98, 1.29]). Paternal exposure was not found to increase the risk
of preterm birth.
An ecologic study in Canada (Philion et al., 1997) examined 30 years of birth records,
corresponding to 9,329 births in a smelter city and a control city. Outcome variables were
intrauterine growth retardation defined as small for gestational age. The odds ratio for
intrauterine growth retardation in the smelter city compared to the control city was 0.83.
Further analysis, stratifying time into 5-year intervals also revealed no associations.
A case control study in Mexico City (Torres-Sanchez et al., 1999) evaluated 161 preterm
births and 459 full term births. Cord blood Pb was significantly higher in the preterm group
(9.8 + 2.0 |ig/dL) compared to the full term group (8.4 + 2.2 |ig/dL) only among primiparous
women.
Using data from the Baltimore-Washington Infant Study database, Min et al. (1996)
found a small association between paternal occupational exposure in the high range and preterm
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delivery with appropriate weight for gestational age (odds ratio of 2.1 [95% CI: 0.7, 6.5]) and
preterm delivery with small for gestational age (odds ratio of 2.4 [95% CI: 1.9, 3.1]). Similar
findings were reported by Lin et al. (1998). Comparing the offspring of Pb-exposed workers
with those of bus drivers, they found an elevated relative risk for preterm delivery (3.0 [95% CI:
1.6, 6.8]) only among men with blood Pb >25 |ig/dL for over 5 years.
In contrast to fetal growth, few factors are consistently related to preterm delivery; thus in
both developed and developing countries the majority of preterm deliveries remain unexplained
(Kramer 1987; Van Den Berg and Oechsli, 1984). Factors which are inconsistently associated
with preterm delivery include maternal age, SES, pre-pregnant weight, prior history of preterm
delivery or spontaneous abortion, and cigarette smoking (Kline et al., 1987; Kramer, 1987).
Thus, these factors must be evaluated as potentially confounding factors in studies of Pb
exposure and preterm delivery.
For preterm delivery, or reduced length of gestation, the evidence for an association with
Pb exposure is contradictory. Several of the prospective studies found no evidence of an
association (Bornschein et al., 1989; Factor-Litvak et al., 1991), whereas one found a
concentration-response relationship (McMichael et al., 1986). Further, two well-done registry
studies (Irgens et al., 1998; Savitz et al., 1990) found some evidence of an association, albeit the
number of exposed cases was small. It seems unlikely that the association between Pb exposure
and preterm delivery is large, but, more research is clearly necessary.

6.6.7 Congenital Abnormalities
Needleman et al. (1984) found an association between cord blood Pb and minor
congenital anomalies among 4,354 infants born in a single hospital in Boston, MA. All data
were obtained from hospital records, not from direct examination of the infants. The most
common anomalies were hemangiomas, lymphangiomas, minor skin problems (tags and
papillae), and undescended testicles. Blood Pb levels were not found to be associated with
individual anomalies.
More recently, a number of studies have considered parental Pb related to occupational
exposure and risk of congenital anomalies in the offspring. In Finland, Sallmen et al. (1992)
evaluated the associations between congenital malformations and paternal exposure during the
time of spermatogenesis. The overall estimated unadjusted odds ratio for men with blood Pb
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levels >20 |ig/dL was 2.4 (95% CI: 0.9, 6.5). Due to small sample sizes, the investigators could
only adjust for one potentially confounding factor at a time; this resulted in odds ratios ranging
from 1.9 to 3.2. Of note is the lack of consistency of malformations among the five men with the
highest blood Pb. The malformation observed included congenital heart disease, oral cleft, club
foot, polydactyly, and anomalies of the adrenal gland. The breadth of these anomalies suggests
either that Pb affects physical development throughout gestation or that this association
represents a chance finding. Among 2,021 pregnancies, Alexander et al. (1996b) found slightly
elevated odds ratios for congenital defects among men in the Pb smelting industry with moderate
exposure (odds ratio of 1.9 [95% CI: 0.6, 6.3]) and high exposure (odds ratio of 2.7 [95% CI:
0.7, 9.6]). These estimates are based on 30 birth defects and 12 stillbirths. No analyses were
presented which considered individual birth defects. In Norway, neither maternal (odds ratio of
1.25 [95% CI: 0.8, 1.9]) nor paternal (odds ratio of 0.94 [95% CI: 0.8, 1.1]) occupational Pb
exposure was associated with serious birth defects (Irgens et al., 1998). Similar results were
reported by Kristensen et al. (1993) between paternal Pb exposure and birth defects, with the
exception of a fourfold increase in the risk of cleft lip among male offspring.
The risk of parental Pb exposure and neural tube defects was evaluated in a case-control
study of 88,449 births (363 neural tube defects) over a 25-year period in Fylde, England (Bound
et al., 1997). Women living in areas in which the water Pb concentration was >10 |ig/L were
more likely to deliver a child with a neural tube defect. The association was consistent for
anencephaly (n = 169) and spina bifida/cranium bifidum (n = 195), even after adjusting for social
class. These authors posit that the association could be a direct effect of Pb on neural tube
closure or an indirect effect, the latter meaning a reduction in uptake of zinc (due to Pb exposure)
leading to a reduction in folate uptake. Irgens et al. (1998) partially confirmed these effects on
neural tube defects in mothers occupationally-exposed to Pb (relative risk of 2.87 [95% CI: 1.05,
6.38]), but not for paternal Pb exposure.
The association between total anomalous pulmonary venous return and parental Pb
exposure during pregnancy (self reported, obtained from industrial hygiene measures, or from a
job exposure matrix) was examined in the Baltimore-Washington Infant Study (Jackson et al.,
2004). In this case-control study, maternal periconceptional (i.e., 3 months prior to conception
through the first trimester) exposure to Pb resulted in an estimated odds ratio of 1.57 (95% CI:
0.64, 3.47). For Pb-exposed men, the estimated odds ratio was 1.83 (95% CI: 1.00, 3.42).
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Findings from this study support a possible association between paternal Pb exposure and total
anomalous pulmonary venous return.
Taken together, the evidence suggests few associations between periconceptional or
prenatal exposure to Pb and congenital anomalies. There is a suggestion of small associations
with high levels of exposure, but many of those studies relied on occupational histories rather
than on actual measures of blood Pb levels.

6.6.8 Summary of the Epidemiologic Evidence for the Reproductive and
Developmental Effects of Lead
Overall, since the 1986 Lead AQCD, a substantial body of work has evaluated the
associations between Pb exposure and reproductive outcomes. It is now clear that Pb clearly
crosses the placenta during all trimesters and maternal exposure results in fetal exposure.
For many other outcomes, the observed associations are relatively small, especially at the levels
of exposure that are currently of interest.
Further, there may be populations with increased fetal susceptibility, including
populations with high rates of smoking and alcohol use, those using ethnic remedies and
cosmetics, and those who use Pb glazed pottery. Low levels of calcium intake may also increase
fetal exposure.
• The available evidence suggests small associations between exposure to Pb and male
reproductive outcomes. These include perturbed semen quality and increased time to
pregnancy. These associations appear at blood Pb levels >45 |ig/dL, as most studies only
considered exposure in the occupational setting. More research is needed regarding
possible male reproductive effects at exposure levels in the lower (and currently more
relevant) range. There are no adequate data to evaluate associations between Pb exposure
and female fertility.
• With one exception, there is no evidence to suggest an association between either
maternal or paternal Pb exposure and increased risk of spontaneous abortions. One study
in Mexico where the mean maternal blood Pb levels were in the moderate range (i.e.,
10-12 |ig/dL) suggests an association.
• To date, the evidence suggests at most a small association of Pb exposure with birth
weight, fetal growth, preterm delivery, and congenital anomalies. The reviewed studies
occurred in very different populations, and the small associations may reflect some
unmeasured or unknown confounding variable.
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6.7 GENOTOXIC AND CARCINOGENIC EFFECTS OF LEAD
6.7.1 Summary of Key Findings from the 1986 Lead AQCD
The 1986 EPA Lead AQCD reviewed five epidemiologic studies of occupationally
exposed workers (Cooper and Gaffey, 1975; Davies, 1984; Selevan et al., 1985; Sheffet et al.,
1982; McMichael and Johnson, 1982). These workers were exposed to inorganic Pb compounds
such as Pb oxides and Pb sulfides. The EPA noted that Cooper and Gaffey reported a significant
increase in lung and gastrointestinal cancer among battery and smelter workers in the United
States (standardized mortality ratios of 1.50 and 1.48 respectively among smelter workers, and
1.32 and 1.23 among battery workers). Further, much of this exposure was by inhalation and
ingestion of Pb oxides, which are relatively insoluble, adding some plausibility to the occurrence
of cancer at these two sites. Sheffet et al. (1982) found a nonsignificant excess of stomach
cancer among U.S. Pb chromate pigment workers. However, Davies (1984) did not find any
cancer excess among U.K. Pb chromate pigment workers. The EPA noted that Selevan et al.
(1985) found a significant excess of kidney cancer among U.S. Pb smelter workers based on
6 cases. This finding was judged striking because it mimicked the findings of kidney cancer in
animals. The EPA judged that the McMichael and Johnson (1982) study of Pb-poisoned workers
was not particularly informative because the non-poisoned workers may have had substantial Pb
exposure and no details were given on how Pb poisoning was determined. In summary the EPA
felt the evidence was insufficient, stating that "little can now be reliably concluded from
available epidemiologic studies."
The studies by Cooper and Gaffey (1975) and Selevan et al. (1985), which are both
important because they are large occupational cohorts with documented high exposure, have
been updated and are further reviewed below. A cohort study of U.K. battery workers (Malcolm
and Barnett, 1982) is also reviewed below.
EPA in 1986 also presented data on human cytogenetic studies, reproducing data from
an earlier 1980 International Agency for Research on Cancer (IARC) monograph for metals and
metallic compounds (IARC, 1980). For Pb, 10 chromosomal aberration studies were judged to
be "positive" and 6 such studies were judged to be "negative." On the whole, the EPA
considered that "under certain conditions Pb compounds are capable of inducing chromosomal
aberrations in vivo and in tissue cultures." The EPA also reviewed more limited data from two
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human studies of sister chromatid exchange (Dalpra et al., 1983; Grandjean et al., 1983), one of
which was positive and one negative.

6.7.2 Summary of Key Findings by the International Agency for Research
on Cancer and the National Toxicology Program
IARC reviewed inorganic and organic Pb compounds in its monograph number 87 in
February of 2004 (IARC, 2005), and concluded that inorganic Pb compounds were probable
human carcinogens (Group IIA). The IARC classification of inorganic Pb compounds as
probable human carcinogens was based on limited evidence in humans and sufficient evidence in
animals. Regarding organic Pb compounds (e.g., tetraethyl Pb), IARC concluded that there was
insufficient information to make any judgment.
Regarding the human studies, IARC based its evaluation largely on six occupational
cohort studies of highly-exposed workers, which were felt to be particularly informative (battery
workers in the United States and the United Kingdom, smelter workers in Italy, Sweden, and the
United States). The IARC assessment focused on four cancer sites: lung, stomach, kidney, and
brain. IARC noted that lung showed a significant elevation in one study (Lundstrom et al., 1997)
and nonsignificant elevations in a number of others. However, the significant elevation of lung
cancer in Lundstrom et al. appeared to be inextricably associated with arsenic in addition to Pb
exposure (Englyst et al., 2001). IARC concluded that the strongest epidemiologic evidence for
Pb carcinogenicity was for stomach cancer, noting that four cohort studies showed a consistent
30-50% excess of stomach cancer versus external referent populations. IARC noted that
confounding by ethnicity, diet, Helicobacter pylori infections, or SES could have played a role in
the stomach cancer excesses. Finally, IARC noted that while one cohort study showed a 2-fold
excess of renal cancer (Steenland et al., 1992), the other studies showed no excess. Similarly,
there were no consistent excesses of brain cancer, although one study did find a significant
positive dose-response between glioma and blood Pb levels, based on small numbers (Anttila
etal., 1996).
The National Toxicology Program (NTP) in 2003 evaluated the carcinogenicity of Pb and
Pb compounds. A summary of its evaluation can be found in NTP's Report on Carcinogens
(NTP, 2004), and the detailed evaluation is also available (NTP, 2003). NTP, like IARC,
concluded that "Pb and Pb compounds are reasonably anticipated to be human carcinogens based
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on limited evidence from studies in humans and sufficient evidence from studies in experimental
animals." The NTP considered that "the strongest epidemiologic evidence was for lung and
stomach cancer, which are consistently but weakly associated with occupations and industries
entailing Pb exposure and with indices of individual Pb exposure, including job history and
biological monitoring of occupationally exposed and general populations. However, most
studies of Pb exposure and cancer reviewed had limitations, including poor exposure assessment
and failure to control for confounders (other factors that could increase the risk of cancer,
including lifestyle factors and concurrent occupational exposure to other carcinogens), and did
not demonstrate relationships between the amount of exposure (concentration or duration, for
example) and the magnitude of cancer risk." NTP, like IARC, also relied heavily on
occupational cohort studies in its evaluation of the epidemiologic evidence. NTP (2003) noted
that "the mechanisms by which Pb causes cancer are not understood. Lead compounds do not
appear to cause genetic damage directly, but may do so through several indirect mechanisms,
including inhibition of DNA synthesis and repair, oxidative damage, and interaction with DNA-
binding proteins and tumor-suppressor proteins."
Both the IARC and NTP evaluations of human evidence relied primarily on occupational
studies of highly exposed workers, in which limited evidence of stomach and to some extent lung
carcinogenicity was found. There are seven such studies with relatively large populations
(Anttila et al., 1995; Carta et al., 2005; Fanning, 1988; Gerhardsson et al., 1995; Lundstrom
et al., 1997; Steenland et al., 1992; Wong and Harris, 2000). A further study (Ades and
Kazantzis, 1988) also addresses Pb exposure in a large occupational cohort, albeit compromised
by the strong correlation between arsenic and Pb exposure in the cohort. It should be noted that
the blood Pb levels among these workers were generally three to five times higher than blood Pb
levels in the two studies of the general U.S. population (Jemal et al., 2002; Lustberg and
Silbergeld, 2002; both based on NHANES II) with environmental exposures. For example, mean
blood levels in two studies of U.S. Pb smelter workers averaged 56 |ig/dL in Steenland et al.
(1990) in 1976 and 80 |ig/dL in Cooper et al. (1985) during the period 1947-1972, while the
U.S. population enrolled in NHANES II in late 1976-1980 averaged 14 |ig/dL. General
population blood Pb levels have decreased markedly since the 1970s in many industrial countries
with the banning of leaded gasoline. U.S. general population levels in the early 1990s thus
averaged 3 |ig/dL according to NHANES III (ATSDR, 1999; see Lead Toxicological Profile,
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page 409). Regarding the occupational studies, while exposure is well documented, detailed
exposure-response data are generally not available, precluding quantitative inference about likely
effects in low exposure groups based on these studies. The high exposure occupational cohorts
are the most informative for deciding whether Pb is likely to cause cancer, simply because high
doses are more likely to show detectable effects than low doses, if effects exist. If Pb does cause
cancer, and assuming there is no threshold below which exposure does not cause cancer, current
low level exposures among the general public may produce some level of Pb-related cancers due
to the potential exposure of a large number of people.

6.7.3 Genotoxicity of Lead
The NTP reviewed in some detail the genotoxicity studies over the period 1970-2002.
These studies are cross-sectional studies, mostly of occupationally exposed workers compared to
a control population. Usually blood Pb levels are available to document exposure. Outcomes
consisted of chromosomal aberrations (CA), sister chromatid exchange (SCE), micronuclei
formation (MN), and studies of DNA damage (often via the comet assay) and/or measures of the
mitotic activity. Of these outcomes, only CAs have been shown to have a positive relationship to
subsequent cancer (Hagmar et al., 2004, Rossner et al., 2005). SCEs are generally considered a
marker of exposure to environmental agents which affect DNA, but do not necessarily predict
cancer risk. MN and DNA damage are thought to indicate genotoxicity with unknown effect on
cancer risk. The informativeness of these outcomes regarding possible human carcinogenicity of
Pb are thus clearly secondary to direct information on cancer risk from epidemiologic studies.
Since the NTP review, there have been three additional cytogenetic studies which are
informative regarding Pb (Palus et al., 2003, Minozzo et al., 2004, and Fracasso et al., 2002),
as well as one mutation study (Van Larebeke et al., 2004). As detailed in Annex Table AX6-7.1,
all four of these studies (two of DNA damage, one of MN, and one of a specific mutation
frequency) were positive in significantly linking Pb exposure to the outcome. Treatment of
potential confounding factors varied across studies, but there was no indication that more
extensive adjustment for such factors was associated with weaker relationships between Pb
exposure and genotoxic endpoints. Potential coexposure to other potentially genotoxic metals
remains an issue, although Palus et al. (2003) found as much or more evidence of genotoxicity
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for each major endpoint examined among heavily Pb-exposed workers as among those expected
to have the heaviest exposure to Cd.
The results of the four most recent studies as well as those reviewed by the NTP are
summarized in Table 6-3. Of eleven studies of chromosomal aberrations (CA), six were judged
to show a positive relationship between CA and Pb, four were judged negative, and one was
neither clearly positive nor negative. In general, these studies were done in the 1970s and 1980s;
only one dates from the 1990s. There were nine studies of sister chromatid exchange. Of these,
four were judged positive, three negative, and two could not be judged clearly one way or the
other. It is notable that the positive studies were generally the most recent. There were four MN
studies, all of which were judged positive. Finally, there were nine studies of DNA damage
and/or mitotic activity. These varied in the specific outcome, although many used a comet assay
to measure oxidative damage to DNA. Eight of these nine studies were judged positive in the
sense that increased DNA damage or mitotic activity was related to Pb exposure, while one was
judged negative.

Table 6-3. Results of Epidemiologic Studies on the Genotoxicity of Lead Exposure"
Studied Outcome
Chromosomal Aberrations
(CA)
Sister Chromatid Exchange
(SCE)
Micronucleus Formation
(MN)
DNA Damage/Mitosis
Gene Mutation

Positive
6
4
5
10
1
Results
Mixed
1
2
0
0
0

Negative
4
3
0
1
0
Results summarize the overall findings of epidemiologic studies addressing the potential genotoxic effects of
lead exposure. Some studies addressed multiple aspects of genotoxicity; for these studies, their results for each
of the listed categories of genotoxic outcomes are presented separately.
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While the overall the evidence from cytogenetic studies is mixed, more recent studies
which were focused on DNA damage or mitotic activity have tended to be largely positive.
However, it is not known whether these outcomes predict subsequent cancer risk.

6.7.4 Meta-analyses of Lead and Cancer
There have been two published meta-analyses of the carcinogenicity of Pb and Pb
compounds. Their major findings are summarized in Table 6-4. Steenland and Boffeta (2000)
relied on eight occupational cohort studies of highly-exposed workers (seven cohort studies, one
nested case-control), all of which had documentation of exposure levels. Meta-analyses were
conducted for lung, stomach, kidney, and brain cancer. The combined lung cancer relative risk
was 1.30 (95% CI: 1.15, 1.46), based on 675 lung cancer deaths. However, the authors noted
that the lung cancer findings were not consistent across studies, and were influenced highly by
one study (Lundstrom et al., 1997) in which confounding by arsenic was likely. Exclusion of
this study dropped the combined lung cancer relative risk to 1.14 (95% CI: 1.04, 1.73).
The strongest positive evidence was for stomach cancer (relative risk 1.34 [95% CI: 1.14, 1.57],
181 observed deaths). There was little positive evidence for renal cancer (relative risk 1.01
[95% CI: 0.72, 1.42], 40 deaths), or brain cancer (relative risk 1.06 [95% CI: 0.81, 1.40]).
All meta-analyses used fixed effects models, given that no evidence of heterogeneity was found
across studies (as long as Lundstrom et al.'s lung cancer results were excluded).
Table 6-4. Results of Meta-Analyses Addressing the Association Between
Lead Exposure and Cancer
Risk Estimate (95% CI) for Indicated Outcome
[Number of Studies Utilized in Estimate]
Meta-Analysis
Fu and Boffetta
(1995)
Fu and Boffetta
(1995)
Steenland and Boffetta
(2000)
Lung Cancer
1.24(1.16, 1.33)
[n=15]
1.42(1.05, 1.92)
[Battery/smelter only]
1.30(1.15,1.46)
[n = 8 - cohort only]
Stomach Cancer
1.33(1.18, 1.49)
[n=10]
1.50(1.23, 1.83)
[Battery /smelter only]
1.34(1.14, 1.57)
[n = 8 - cohort only]
Renal Cancer
1.19(0.96, 1.48)
[n=5]
1.26 (0.70, 2.26)
[Battery /smelter only]
1.01 (0.72, 1.42)
[n = 7 - cohort only]
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Fu and Boffetta (1995) conducted an earlier meta-analysis in which they reviewed
16 cohort and 7 case-control studies. Different numbers of studies were used for meta-analyses
of different outcomes, dependent on whether that outcome was reported separately, among other
factors. Twelve occupational studies were used in a meta-analysis of lung cancer, resulting in a
combined relative risk of 1.29 (95% CI: 1.10, 1.50) (random effects model, reflecting significant
heterogeneity of lung cancer results across studies). Meta-analyses using fixed effects (no
significant heterogeneity between studies) resulted in relative risks of 1.33 (95% CI: 1.18, 1.49)
for stomach cancer (10 studies), of 1.19 (95% CI: 0.96, 1.48) for kidney cancer (5 studies), and
1.41 (95% CI: 1.16, 1.71) for bladder cancer (5 studies). No meta-analysis was conducted for
brain cancer. Restricting analyses for stomach, lung, and kidney cancer to those studies with the
highest occupational exposure to Pb (3 to 5 studies of battery and smelter workers) resulted in
slightly higher relative risks. The authors concluded that "the findings from the workers with
heavy exposure to Pb provided some evidence to support the hypothesis of an association
between stomach and lung cancer and exposure to Pb. The main limitation of the present
analysis is that the excess risks do not take account of potential confounders, because little
information was available for other occupational exposures, smoking, and dietary habits.
The excess risk of stomach cancer may also be explained, at least in part, by nonoccupational
factors. For bladder and kidney cancers, the excess risks are only suggestive of a true effect
because of possible publication bias."

6.7.5 Review of Specific Studies on the Carcinogenicity of Lead Since the
1986 Lead AQCD
6.7.5.1 Introduction
Many epidemiologic studies of Pb exposure and cancer have been conducted in the past
two decades. The most relevant studies focus on exposure via occupational sources, wherein the
most intense exposure to Pb can be expected to occur. This exposure predominantly involves
inorganic Pb species. Relevant studies are discussed below, beginning with the most key
occupational and general population studies, followed by a brief summary of other relevant
studies.
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6.7.5.2 Key Studies of Occupational Populations in the United States
The strongest evidence in the key occupational studies linking Pb exposure to human
cancers is that for cancers of the lung and those of the stomach. Of seven large occupational
cohort studies available (Ades and Kazantzis, 1988; Anttila et al., 1995; Carta et al., 2005;
Gerhardsson et al., 1995; Lundstrom et al., 1997; Steenland et al., 1992; Wong and Harris,
2000), all showed results consistent with an increase in lung cancer risk among Pb-exposed
workers, and in four of these studies the association was statistically significant. Details of these
studies are summarized in Annex Table AX6-7.2. A few of these studies are discussed below.
Steenland et al. (1992) followed up 1,990 male U.S. Pb smelter workers, employed from
1940 to 1965, through 1988. Standardized mortality ratios indicated an excess of lung, stomach,
kidney, and bladder cancer that did not reach statistical significance. Focusing on workers
classified as highly Pb-exposed based on air monitoring records yielded a significant excess for
kidney cancer (standardized mortality ratio of 2.39 [95% CI: 1.03, 4.71]), although it did not
appear to increase with duration of exposure. Estimates for the other cancers (standardized
mortality ratio of 1.11 [95% CI: 0.82, 1.47] for lung; 1.28 [95% CI: 0.61, 2.34] for stomach;
1.33 [95% CI: 0.48, 2.90] for bladder) showed little change with restriction to the high-exposure
group. While neither arsenic nor Cd exposure could be controlled for, 1975 NIOSH monitoring
data indicated less intense exposure to airborne Cd or arsenic than to Pb. Lead averaged
3.1 mg/m3 and arsenic 14 |ig/m3, compared to current OSHA standards of 0.05 mg/m3 for Pb and
10 |ig/m3 for arsenic. No data on workers' smoking status were available.
Wong and Harris (2002) extended follow-up on the battery and smelter worker cohort
previously reported on by Cooper et al., 1985 through 1995, an additional 15 years. With the
additional follow-up, standardized mortality ratios for lung, tracheal, or bronchial cancer
decreased to 1.14 (95% CI: 0.99, 1.30) for battery workers but showed little change for smelter
workers at 1.22 (95% CI: 1.00, 1.47). An elevated standardized mortality ratio for stomach
cancer (1.53 [95% CI: 1.12, 2.05]) persisted among battery workers, with a lesser elevation
among smelter workers (1.33 [95% CI: 0.75, 2.20]). Among other cancers, only thyroid cancer
among all workers combined showed a significantly elevated standardized mortality ratio
(3.08 [95% CI: 1.33, 6.07]). As with earlier analyses based on this cohort, concomitant
exposures to other compounds could not be controlled for, but as these were likely to be most
intense among Pb production workers, whose standardized mortality ratios were similar to or
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lower than those for battery workers, any bias resulting from such exposure probably was
minimal. No data were available to assess the possible role of smoking, diet, or other potential
nonoccupational risk factors in the results.
Anttila et al. (1995) linked 20,700 Finnish workers whose blood Pb was monitored during
1973 to 1983 by the Finnish Institute of Occupational Health to the Finnish Cancer Registry.
Exposure was subdivided according to highest peak blood Pb measured: low (0 to 0.9 |imol/L
[0 to 18.6 jig/dL]), moderate (1.0 to 1.9 |imol/L [20.7 to 39.4 jig/dL]), and high (2.0 to
7.8 |imol/L [41.4 to 161.6 |ig/dL]). The total cohort showed no elevation in cancer mortality
based on standardized mortality ratio analyses. Among male workers with moderate exposure,
however, incidence of total respiratory cancer and lung cancer both were elevated (standardized
incidence ratio of 1.4 [95% CI: 1.0, 1.9] for both). Risks of total digestive, stomach, bladder,
and nervous-system cancer also were modestly elevated. Risks of mortality for all cancer for
both men and women (relative risk of 1.4 [95% CI: 1.1,1.8]) and lung or tracheal cancer
(relative risk of 2.0 [95% CI: 1.2, 3.2]) were even stronger when a person-year analysis was
applied to compare workers with moderate Pb exposure to those with low exposure. Risks did
not increase in the highest exposure group, although the power of analyses specific for this group
were limited by its relatively small size (e.g., lung or tracheal cancer deaths among men in the
low-, moderate-, and high-exposure groups numbered 25, 34, and 11, respectively, for the
person-year-based analyses).
In summary, occupational exposure to Pb was associated with elevated risks of cancers of
the lung and stomach. However, the modest elevation of lung cancer risk seen in most relevant
studies is in the range of possible confounding due to smoking or other occupational exposures,
particularly arsenic, which precludes the evidence from these studies being seen as conclusive.
In particular, one occupational study with the highest lung cancer risk (standardized incidence
ratio of 5.1 [95% CI: 2.0, 10.5] with a latency period of 15 years among workers with the
highest Pb exposure) (Lundstrom et al., 1997) has been subsequently shown to be highly
confounded by arsenic, and without this study, the combined evidence for a lung cancer
elevation across studies is considerably reduced (e.g., the estimated relative risk falls from
1.30 to 1.14). A moderate elevation of stomach cancer is also found in most studies of
occupationally-exposed populations with applicable data on this outcome. As with lung cancer,
it is possible that other risk factors such as intake of smoked meats or H. pylori infection could
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have contributed to the observed associations, but the observed elevation (combined risk
estimate of 1.33 or 1.34) coupled with the known effect of diet makes it unlikely that the
elevation in stomach cancer is entirely due to confounding by diet. Data for other sites such as
kidney, brain, and bladder show some indications of an excess, but the results across studies are
not consistent and are based on small numbers.

6.7.5.3 Key Studies of the General Population
There are two key general population cohort studies in which Pb exposure was assessed
via blood Pb levels (see Annex Table AX6-7.3 for additional details). Jemal et al. (2002)
conducted the first biomarker-based general population cohort study of Pb exposure and cancer.
The study employed the subsample of 3,592 White U.S. participants in NHANES II (1976 to
1980) who had undergone blood Pb level determinations at time of entry. Deaths among this
population were enumerated through 1992 by linkage to the National Death Index (NDI) and
Social Security Administration Death Master File. Median blood Pb levels in this population
were 12 |ig/dL. Adjusted for age, smoking, drinking, region, year, and gender, risk of mortality
from any cancer rose across quartiles of blood Pb level, but this trend was not statistically
significant. The trend across quartiles was not consistent in gender-specific analyses, although
relative risks were elevated for the highest quartile of blood Pb level in both men and women
(relative risk 2.0 for men and 1.6 for women). The relative risk for lung cancer based on
comparison of subjects with blood Pb levels above or below the median was 1.5 in the combined
population, with higher risk observed among women than men. The highest relative risks were
observed for cancer of the esophagus (3.7 [95% CI: 0.2, 89]), pancreas (3.6 [95% CI: 0.6,
19.8]), and stomach (2.4 [95% CI: 0.3, 19.1]); no elevations were noted for cancers of other
sites. Total cancer mortality was also addressed through a spline regression (Figure 6-13). The
mortality curves were visually suggestive of an upward trend at low blood Pb levels
(<20 jig/dL), but no statistically significant dose-response pattern was present except for
analyses restricted to women.
The lack of statistically significant results reflects the small number of deaths during
follow-up, which limited the study's power; of the nine major sites examined, the number of
deaths ranged between 5 and 16 for all sites except the lung. Further, only 4 and 16 deaths
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100
10
RJ
N
(0
I
0}
(3)
0.1
0.01
0.001
100
10
0.1
0.01
B
0 10 20 30 40 50 60 70 80 0 10 20 30 40 50 60
Blood Lead Level (pg/dL) Blood Lead Level (pg/dL)
Figure 6-13. Five-knot cubic spline regression models of total cancer mortality and blood
lead level by gender, based on analyses of the NHANES II cohort. Relative
risk of all cancer mortality for different blood lead levels compared with
referent blood lead level of 8 ug/dL (the 12.4th percentile) among White men
(A) and White women (B) in the United States (NHANES II). The solid line
shows the fitted 5-knot spline relationship; the dashed lines are the point wise
upper and lower 95% confidence limits.

Source: Jemal et al. (2002).
occurred among men and women, respectively, with blood Pb levels <9.8 |ig/dL, precluding

assessment of potential effects within that range. Detailed exposure-response analyses were

restricted to all cancers combined, although potential effects could have been strongly target-

organ specific. In addition, the use of quartile cut points based on the distribution of Pb

concentrations estimated for the total U.S. population resulted in relatively small numbers in the

referent group (lowest exposure quartile) for males and in the high-exposure quartile for females.

Use of a biomarker provided an objective measure of Pb exposure. Nevertheless, reliance on a
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single blood Pb measurement produces less reliable estimates than would be obtained through
multiple measurements and precludes addressing temporal changes in Pb exposure over the
follow-up period. Lack of control for exposure to occupational carcinogens other than Pb and
potential residual confounding by duration and intensity of tobacco smoking also could have
biased the results, especially for men.
Lustberg and Silbergeld (2002) carried out another biomarker-based general population
study based on the same NHANES II mortality cohort used by Jemal et al. (2002). This study
did not exclude non-Whites, however, and employed more extensive adjustment for potential
confounding factors than the Jemal et al. (2002) analyses (i.e., education, body mass index, and
exercise were included in the regression models, although alcohol intake was not). In addition,
persons with blood Pb levels >30 |ig/dL were excluded in order to restrict comparisons to levels
below the OSHA standard for Pb exposure. Persons with levels below 10 |ig/dL served as the
referent group. Survival analyses adjusted for potential confounders found a relative risk for
cancer mortality of 1.5 (95% CI: 0.9, 2.5) for those with blood Pb levels of 10 to 19 |ig/dL and
1.7 (95% CI: 1.0, 2.8) for those with levels of 20 to 29 |ig/dL. Separate analyses of lung-cancer
and non-lung-cancer deaths yielded estimates of increased risk for moderate- or high-exposure
groups, compared with the referent population, both for lung cancer and non-lung cancer.
However, none of the estimates reached the p < 0.05 level of statistical significance, and the
results for non-lung cancers showed no evidence of an exposure-response relationship.
As with Jemal et al. (2002), the use of a biomarker for exposure and the prospective
design of the study are strengths. Its attempts to control for potential confounders were more
extensive, and its choice of cut points for the referent category yielded more males in the referent
group, although that group still included less than 20% of the study population. However, it is
notable that blood Pb levels rose significantly with smoking level. The models included terms
for former smoking, current light smoking, and current heavy smoking (>1 pack per day). Still,
some degree of residual confounding due to smoking might have remained, which could have
contributed to the estimated risk of lung cancer for the highest exposure category (relative risk of
2.2 [95% CI: 0.8, 6.1]). Such residual confounding would have had less effect on the results for
non-lung cancer. As noted regarding the other NHANES-based study, however, mortality due to
cancers of other sites was too uncommon to allow for reliable site-specific comparisons. In the
Lustberg and Silbergeld analysis, all cause and cardiovascular mortality increased monotonically
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with blood Pb level, which might indicate residual confounding from SES or smoking affecting
both heart disease and cancer.

6.7.5.4 Other Lead Studies
There are a variety of other epidemiologic studies of Pb exposure, which are less
important than the key studies above but which offer some information. Studies reviewed in this
section are summarized in Annex Table AX6-7.4. The weaknesses in these studies largely stem
from potential confounding by other metals in the cohort studies and likely misclassification of
Pb exposure in the case-control studies.
Some studies have examined the potential link between parental Pb exposure and
childhood cancer. These are briefly described in Section 5.6.2.1, but are not further detailed
here. Lack of direct measures of child exposure, the fact that many of the same interpretational
problems (e.g., potential coexposures) noted for occupational studies as a whole, and low
statistical power due to the rarity of the outcome under study render the available evidence less
relevant than that from direct study of exposed occupational and general populations.

6.7.6 Confounding of Occupational Lead Studies Due to Other Occupational
Exposures: Arsenic, Cadmium
A number of studies of Pb workers come from smelters, where exposures to other metals
are common. Of particular concern are other lung carcinogens, not only especially arsenic
(workers exposed to high levels of arsenic historically have had a lung cancer relative risk of 3 to
4, see Steenland et al. 1996), but also Cd. Glass workers are also of limited use for inference
about Pb effects, as they are also typically exposed to Cd, arsenic, chromium, and nickel, all of
which are lung carcinogens (e.g., see Wingren and Axelson, 1993).
In some smelters, measurements have been taken which indicate clearly that exposures to
these other carcinogens was minimal and the main suspect is Pb (e.g., Steenland et al., 1992).
In others, however, one is unable to disentangle the effects of arsenic and Pb (Ades and Kazantis,
1988, Lundstrom et al., 1997). As a result, these studies cannot yield strong evidence regarding
the possible relation between lung cancer and Pb specifically. The study by Lundstrom et al.,
1997 is particularly important in this regard, because it had a high relative risk of 2.8 (95% CI:
2.0, 3.8), and had an important effect in raising the overall result when included in meta-analyses
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(e.g., Steenland and Boffetta [2000], where exclusion of the Lundstrom et al. study lowered the
estimated combined lung cancer relative risk from 1.30 to 1.14). A subsequent publication by
Englyst et al. (2001) indicated that the smelter workers studied by Lundstrom et al. (1997) likely
had significant exposure to arsenic, and the authors concluded that it was impossible to separate
the effects of Pb and arsenic.

6.7.7 Confounding of Lead Studies: Smoking and Other Factors
The most informative studies of Pb carcinogenicity are those comparing highly exposed
workers to general populations. In these comparisons one must consider typical differences
between worker populations and the general populations, in particular differences due to
smoking and diet. Smoking can be a major confounder for lung cancer, while diet or SES can
be a confounder, albeit weaker, for stomach cancer.
Regarding smoking, it has been shown both theoretically and empirically that
confounding due to smoking differences between workers and the general population will
typically account for an observed relative risk of-1.1 to 1.2, with a possible maximum of
about 1.4 (Axelson and Steenland, 1988; Siemiatycki et al., 1988). Furthermore, most
occupational cohort studies are retrospective and have little information on smoking, making it
impossible to control directly for potential confounding by this strong risk factor. As noted
above, the lung cancer relative risk in the meta-analysis of Steenland and Boffetta (2000), after
excluding the Lundstrom et al. study, was 1.14 (95% CI: 1.04, 1.73), based on seven
occupational cohort studies, six of which used a non-worker external referent population, and
none of which controlled for smoking as a confounder. This relatively small excess relative risk
could plausibly be due to confounding by smoking. Unfortunately the occupational cohort
studies were usually not followed by nested-case control studies of lung cancer, which could
have controlled for smoking and, furthermore, they usually did not involve internal exposure-
response analyses, wherein confounding by smoking is usually minimal. An exception was the
lung cancer case-control study conducted by Anttila et al. (1995) within a large cohort of Finnish
workers with known blood Pb levels. In this case-control study smoking-adjusted lung cancer
odds ratios were increased among workers with higher estimated cumulative blood Pb or higher
peak blood Pb exposure compared to workers with the lowest exposure, and the authors noted
that smoking actually appeared to be a "weak negative confounder" for the high peak blood Pb
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group. Also, in one large population-based case-control study with extensive information on
other cancer risk factors, there remained an elevated odds ratio for lung cancer with substantial
Pb exposure after controlling for smoking (Siemiatycki et al., 1991). Hence, there is some
evidence that confounding by smoking does not likely explain the modest excess lung cancer risk
seen in many studies.
Diet high in salt or smoked meats, Helicobacter pylori infection, and SES are possible
confounders for stomach cancer. Those of highest SES compared to those of lower SES have
been shown to have a relative risk of about 3.0 (Tomatis, 1990). None of the occupational cohort
studies, in which again stomach cancer in workers was compared to the general population,
controlled for these potential confounders. However, these potential confounding factors are
much less powerful risk factors in respect to stomach cancer than smoking is with respect to lung
cancer, and hence are unlikely to account for relative risks higher than perhaps 1.1 or at most 1.2.
Given that the occupational cohort studies had a combined relative risk of 1.34 (95% CI: 1.14,
1.57) in the meta-analysis of Steenland et al. (2002) and 1.33 (95% CI: 1.18, 1.49) in that of Fu
and Boffetta (1995), it seems unlikely that confounding by these factors can fully account for the
excess stomach cancer risk observed in the occupational studies.

6.7.8 Summary of Epidemiologic Evidence for the Genotoxic and
Carcinogenic Effects of Lead
The availability of studies of cancer in Pb-exposed populations was limited at the time of
the 1986 Lead AQCD. The number and range of studies have since notably expanded, including
extended follow-ups of previous cohorts, new cohort and case-control studies, and analyses
addressing not only cancer but genotoxicity. The newly available epidemiologic data greatly
enhance the knowledge base regarding Pb carcinogenicity, with key findings and conclusions
emerging as follows.
• Studies of genotoxicity consistently link Pb-exposed populations with DNA damage and
micronuclei formation, although less consistently with chromosomal aberrations, a more
established indicator of cancer risk.
• The epidemiologic data reviewed above from key high Pb exposure occupational studies
suggest a relationship between Pb exposure and cancers of the lung and the stomach, as
supported by two meta-analyses. Clear conclusions are limited by potential confounders,
e.g., other occupational exposures (arsenic, Cd), smoking, and dietary habits.
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• Two general population cohort studies (NHANES II) have been conducted. Low Pb
exposure was assessed via blood levels. These studies show internal dose-response trends
but suffer at times from small numbers for site-specific analyses or lack of site-specific
analyses altogether, and from possible residual confounding by smoking and SES.

• Overall, the above findings provide only very limited evidence suggestive of Pb exposure
associations with carcinogenic or genotoxic effects in humans. On the other hand, animal
studies (see Chapter 5) provide reproducible results in several laboratories and multiple rat
strains (with some evidence of multiple tumor sites), along with short-term studies
indicating that Pb affects gene expression. However, although the animal studies clearly
demonstrate Pb-related carcinogenic effects in response to dietary and subcutaneous
exposures to several soluble Pb salts, they may be of limited relevance here because
human exposures of most concern here are to inhaled Pb oxides.

• Nevertheless, the most recent IARC (2005) review concluded that inorganic Pb
compounds were probable human carcinogens (Group IIA), based on limited evidence in
humans and sufficient evidence in animals. This is consistent with the U.S. National
Toxicology Program's Carcinogen Review Committee Report, which recommends that Pb
and Pb compounds be considered as "reasonably anticipated humans carcinogens."
Similarly, although the human evidence is inadequate according to EPA's Guidelines for
Carcinogen Risk Assessment (U.S. Environmental Protection Agency, 2005), Pb is likely
classifiable under those guidelines as a probable human carcinogen based on the available
animal data.
6.8 EFFECTS OF LEAD ON THE IMMUNE SYSTEM

6.8.1 Summary of Key Findings of the Effects of Lead on the Immune
System from the 1986 Lead AQCD

The 1986 Lead AQCD concluded that studies conducted in laboratory animal models
provided evidence for immunosuppressive effects of Pb; however, evidence for such effects in
humans was lacking. Since then, the epidemiological study of immunological effects of Pb has
progressed considerably. The currently available epidemiologic and clinical observations are
consistent with the greater body of evidence derived from studies in experimental animals
indicating that Pb can suppress cellular and humor immunity and decrease host resistance to
infection agents and tumor cells (see Section 5.9). Findings from the epidemiologic studies
suggest that Pb exposure (as reflected in blood Pb concentration) may be associated with effects
on cellular and humoral immunity. These effects include changes in serum immunoglobulin
levels (e.g., elevated serum IgE); perturbation of peripheral lymphocyte phenotype profiles,
including decreases in peripheral blood T-cell abundance and changes in T-cell:B-cell abundance
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ratios; suppression of lymphocyte activation; and suppression of neutrophil chemotaxis and
phagocytosis.
Available studies of associations between Pb exposure and immunological outcomes are
summarized in Annex Tables AX6-8.1 and AX6-8.2. In general, while the studies provide
support for associations between Pb exposure and immunological outcomes, the studies have
numerous limitations that complicate the assessment of the strength of reported associations and
potential causation. Furthermore, the health consequences of outcomes that have been associated
with Pb exposure are uncertain. All studies have been cross-sectional in design and most
included relatively small cohorts. The studies implemented varying degrees of quantitative
analysis of potential covariables and confounders. In most studies, a detailed analysis of
covariables and confounding was lacking, and many of the reports offered no analysis of
covariables or confounding. Covariables that were considered (but not consistently) in
multivariate analyses or controlled by stratification included age, sex, race, smoking habits,
alcohol consumption, and illness and/or medications that might affect the immune system.
Studies that offer the strongest designs are discussed in greater detail below.

6.8.2 Host Resistance, Hypersensitivity, and Autoimmunity
Associations between Pb exposure and host resistance have not been rigorously examined
in humans. Two analyses of illness surveys in children (Rabinowitz et al., 1990) and Pb workers
(Ewers et al., 1982) have been reported. Both studies relied on personal surveys for assessment
of illness and neither study considered covariates or confounders in the analyses. In the
Rabinowitz et al. (1990) study, the highest relative risks (blood Pb concentration > 10 |ig/dL
compared to <10 |ig/dL) were: other respiratory tract illnesses, 1.5 (95% CI: 1, 2.3); severe ear
infections, 1.2 (95% CI: 1, 1.4); illnesses other than cold or influenza, 1.3 (95% CI: 1.0, 1.5).
Ewers et al. (1982) reported mean frequency of self-reported colds and influenza per year of
employment in Pb workers (blood Pb range 21-85 |ig/dL) compared to a reference group (range
6-21 |ig/dL). Mean frequency of 2 to 4 illnesses per year was higher among the Pb workers
28.8% versus 16.1%); however, a statistical analysis of the data was not reported. Collectively,
these studies do not provide convincing evidence for a strong association between Pb exposure
and altered disease resistance in humans.
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Two studies have also been reported that have examined possible associations between Pb
exposure (e.g., blood Pb concentration) and asthma. In the Rabinowitz et al. (1990) study
described above, the relative risk of asthma (blood Pb <5 |ig/dL compared to > 5 or > 10 |ig/dL)
were not significant in Caucasian or African-American cohorts. In the Caucasian cohort, the
hazard ratios were 1.4 (95% CI: 0.7, 2.9) for >5 |ig/dL and 1.1 (95% CI: 0.2, 8.4) for
> 10 |ig/dL. In the African-American cohort, the corresponding hazard ratios were 1.0 (95% CI:
0.8, 1.3) for >5 |ig/dL and 0.9 (95% CI: 0.5, 1.4) for > 10 |ig/dL. Hazard ratios for asthma
incidence in African-Americans compared to Caucasians (<5 |ig/dL) were 1.6 (95% CI: 1.4, 2.0)
for <5 |ig/dL, 1.4 (95% CI: 1.2, 1.6) for >5 |ig/dL, and 2.1 (95% CI: 1.2, 3.6) for > 10 |ig/dL.
Thus, while there appeared to be an elevated risk of asthma in the African-American cohort,
relative to the Caucasian cohort, a significant effect of blood Pb level on risk in African-
Americans or Caucasians was not evident in this study. Covariates included in the analysis were
average annual income, birth weight and gender. Similar results were obtained when a more
stringent definition of asthma was applied to the subjects. Collectively, these studies do not
provide convincing evidence for a strong association between Pb exposure and asthma in
children.

6.8.3 Humoral Immunity
A characteristic immunological response to Pb exposure in animals is an increase in
production of IgE, immunoglobulin that has been associated with allergy and allergic airway
disease (see Section 5.9.3.2). Although epidemiologic literature is not conclusive regarding the
dose-response relationships for Pb effects on immunoglobulin production in humans, studies in
children have consistently found significant associations between increasing blood Pb level and
increasing serum IgE levels (Karmaus et al., 2005; Lutz et al., 1999; Sun et al., 2003)
(Table 6-5). These effects were evident at blood Pb values <10 |ig/dL. Increasing serum IgE
levels also have been observed with increasing blood Pb concentration (blood Pb >30 |ig/dL) in
association with occupational exposures to Pb (Heo et al., 2004). Outcomes for other
immunoglobulin indices in adults have been less consistent (Pinkerton et al., 1998; Sarasua et al.,
2000).
Possible associations between Pb exposure and biomarkers of humoral immunity in
children have been examined in several cross-sectional studies (Annesi-Maesano et al., 2003;
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Table 6-5. Summary of Results of Selected Studies of Associations Between Lead Exposure
and Serum Immunoglobulin Levels
oo
Blood Lead Oig/dL)
Study
Children
Annesi-Maesano et al. (2003)
Karmaus et al. (2005)
Lutzetal. (1999)
Sarasua et al. (2000)
Sun et al. (2003)
Adults
Heo et al. (2004)
Pinkerton et al. (1998)
Sarasua et al. (2000)
-, decrease; +, increase; o, no effect;
Subjects

Neonates
Children, 7-10 yr
Children, 9 mo-6 yr
Children, 6-35 mo
Children, 3-6 yr

Batter manufacture workers
Smelter workers
General population
NR, not reported; Ig, serum
"a Mean (SD)

374 67 (48)b
331 3
270 NR
372 7
73 NR

606 -22 (~10)e
229 39f
433 4.3
immunoglobulin level.
Range J§A ^ ^ ^

NR NR +c NR NR
l-5e o + oo
1-45 NR + NR NR
~2-16d + NR + +
-3^0 NR +

NR o + o o
<2-55 o NR o
~l-10d o NR o o

a Total number of subjects (including reference group)
b Infants cord blood (maternal blood lead mean was 96 ug/dL [SD 58])
0 In association with increasing neonatal hair lead
d 5th-95th percentile range
e Mean of age-group means and SDs
f Median
-------
Karmaus et al., 2005; Lutz et al., 1999; Reigart and Graher, 1976; Sarasua et al., 2000; Sun et al.,
2003; Wagnerova et al., 1986). Four studies warrant particular attention because they examined
a relatively low range of blood Pb concentrations and applied multivariate analyses to the data in
attempts to control for possible covariables (Karmaus et al., 2005; Lutz et al., 1999; Sarasua
et al., 2000; Sun et al., 2003). Three studies found significant associations between increasing
blood Pb concentration and serum IgE levels (Karmaus et al., 2005; Lutz et al., 1999; Sun et al.,
2003). The reported percent increase in serum IgE levels measured in these studies ranged from
-50 to 400%. The Lutz et al. (1999) study measured serum IgE and IgG (against Rubella) in
270 children (age range 9 months to 2 years; blood Pb range 1 to 45 |ig/dL). The observed blood
Pb-age-IgE relationship is shown in Figure 6-14. The highest IgE levels (mean 211 ITJ/mL
[SD 441], n = 17) were observed in children who had blood Pb concentrations in the range of
15 to 19 |ig/dL; by comparison, mean IgE levels were 52 ITJ/mL (SD 166) for subjects who had
blood Pb concentrations <10 |ig/dL (n = 174). The Karmaus et al. (2005) study measured serum
IgA, IgE, IgG, and IgM levels in 331 children (age range 7-10 years). Blood Pb levels were
lower in this study than in the Lutz et al. (1999) study (1 to 5 |ig/dL). A multivariate linear
regression analysis revealed a significant association between blood Pb (p < 0.05) and serum IgE
(but not IgA, IgG, or IgM). The change in serum IgE level may appear not to be monotonic with
increasing blood Pb concentration (Figure 6-15). However, the two lowest means are not
significantly different so that apparent non-monotonicity of the effect/exposure relationship does
not have statistical support. The highest IgE levels (adjusted mean 59 IU/L) were observed in
the children who had blood Pb concentrations ranging from 2.8 to 3.4 |ig/dL (n = 86) and
>3.4 |ig/dL (n = 82). Sun et al. (2003) measured serum IgE, IgG, and IgM levels in children,
ages 3 to 6 years (blood Pb range 2.6-44 |ig/dL, n = 73). A nonparametric comparison of
immunoglobulin levels between low (<10 |ig/dL) and high (> 10 |ig/dL) blood Pb strata revealed
significantly higher IgE levels and significantly lower IgG and IgM levels in the high blood Pb
stratum.
The study by Annesi-Maesano et al. (2003) provides futher suggestive evidence for an
association between Pb exposure and increasing IgE levels. The study included 374 mother-
infant pairs who had relatively high mean blood Pb levels (maternal mean 96 |ig/dL [SD 58];
infant cord 67 |ig/dL [SD 48]). Serum IgE level was significantly associated with increasing
infant hair Pb (p < 0.001), but not with cord blood Pb or placental Pb level. The association
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1800
Figure 6-14. Relationship between blood lead concentration, age, and serum IgE level in
children. Spearmen partial correlation between blood lead and serum IgE is
0.22 (p = 0.0004, n = 221).
Source: Lutzetal. (1999).
between IgE and hair Pb levels was evident in a subset of mother-infant pairs, in which mothers
were classified as nonallergenic, and was unrelated to maternal smoking (i.e., urinary cotinine).
The ATSDR Multisite Lead and Cadmium Exposure Study (ATSDR, 1995) is one of the
largeststudies to assess humoral immune status in association with Pb exposures; however, it did
not include an assessment of IgE. The study included a cross-sectional analysis of serum IgA,
IgG, and IgM levels in 1,561 subjects (age range 6 months to 75 years) who resided in areas
impacted by Pb mining and/or smelting operations and in 480 demographically-matched controls
(Sarasua et al., 2000). A multivariate linear regression analysis of immunoglobulin levels and
blood Pb concentration (exposed and control groups combined) revealed associations between
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70

60-

j" 50-
-x:
— 40-
LU
_O)
£ 30-
2
0)
W 20-

10-

0
<2.2 (n = 82) 2.2-2.8 (n = 81) 2.8-3.4 (n = 86)
Blood Lead Category (jjg/dL)
>3.4 (n = 82)
Figure 6-15. Relationship between blood lead concentration and serum IgE level in
children. Mean serum IgE levels (standard deviations not reported) are
adjusted for age, number of infections in the previous 12 months, exposure
to passive smoke in the previous 12 months, and serum lipids (sum of
cholesterol and triglycerides). Means of serum IgE levels in blood lead
categories were significantly different (F-test p = 0.03).
Source: Karmaus et al. (2005).
increasing blood Pb and increasing serum IgA, IgG, and IgM levels in subjects 6 to 35 months of
age (blood Pb 5th-95th percentile range 1.7-16 |ig/dL, Figure 6-16).
Possible associations between Pb exposure and biomarkers of humoral immunity also
have been examined in several cross-sectional studies of Pb workers (Alomran and Shleamoon,
1988; Anetor and Adeniyi, 1998; Ayatollahi, 2002; Coscia et al., 1987; Ewers et al., 1982;
Heo et al., 2004; Kimber et al., 1986; Pinkerton et al., 1998; Undeger et al., 1996).
Outcomes from these studies, with respect to humoral immune parameters, measured as serum
and/or salivary immunoglobulin levels, are mixed. Some studies finding positive associations
with blood Pb (Heo et al., 2004), negative associations (Anetor and Adeniyi, 1998; Ewers et al.,
1982; Pinkerton et al., 1998), or no (or mixed) effects (Alomran and Shleamoon, 1988; Kimber
et al., 1986; Queiroz et al., 1994b; Sarasua et al., 2000; Undeger et al., 1996). Based on study
6-201
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1.3

E
B 1.21
0)
c 1.1 -
ro
£
o
30 |ig/dL)
compared to lower strata (<10 or 10-29 |ig/dL) for the age strata 30-39 years, >40 years, and for
all ages combined.
Although the Pinkerton et al. (1998) study did not assess IgE outcomes, it offers the
strongest study design of the three for assessment of other immunoglobulin classes. Even though
it is a relatively small cross-sectional study, it considered immune illnesses and immune
6-202
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2000
1800-
-. 1600-
_l
3) 1400-
11 120°"
- 1000-
800-
600-
400-
200-
0
E
3
0
CO
Figure 6-17.
<10 10-29 >30
Blood Lead Category (M9/dL)

Relationship between blood lead concentration and serum IgE level in
lead workers. Mean serum IgE levels in high blood lead category were
significantly higher for all ages (shown), and within age categories
>40 years and 30-39 years, but not within age category <30 years.
Source: Heo et al. (2004).
suppressant drugs in the construction of the cohorts and examined a relatively large number of
potential covariates in the data analysis. Serum immunoglobulin levels were measured in male
smelter (n = 145) workers and hardware workers (n = 84). Excluded (by blind evaluation) from
the study cohorts were individuals who had "serious" illnesses of the immune system, who were
taking immune suppressant drugs, or who had chemical exposures (other than to Pb) that might
affect immune function. Median blood Pb concentrations were 39 |ig/dL (range 15-55 |ig/dL) in
the Pb workers and <2 |ig/dL (range <2-12 |ig/dL) in the reference group. Covariate-adjusted
(logistic regression) geometric mean serum IgA, IgG, and IgM, and salivary IgA levels in the Pb
workers were not significantly different from the reference group; however, the adjusted
regression coefficient for serum IgG and time-integrated (but not current) blood Pb level was
negative and significant.
The Sarasua et al. (2000) study, described above for its assessment of children, also
included a cross-sectional analysis of serum IgA, IgG, and IgM levels in adults (age 16-75 years,
6-203
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n = 433; blood Pb 5th-9th percentile range 1 to 10 |ig/dL) and found no significant associations
between blood Pb and serum immunoglobulin levels (serum IgE outcomes were not assessed).
Also germane to the evidence for effects of Pb on humoral immunity in humans are the
results of a clinical study in which serum immunoglobulin levels were repeatedly measured in a
Pb smelter worker who underwent CaEDTA chelation therapy three times per week for a period
of 10 weeks (Sata et al., 1998). Serum IgA, IgG, and IgM were significantly higher when
assessed 24 h after each CaEDTA treatment compared to assessments made prior to treatment.
Furthermore, serum IgG levels were significantly negatively correlated with blood Pb level
during the treatment period. Before-treatment and after-treatment blood Pb concentration means
were 45.1 jig/dL (SD 16.0) and 31.0 jig/dL (SD 9.8), respectively.

6.8.4 Cell-Mediated Immunity
Studies conducted in animals and in vitro experimental models indicate that Pb
preferentially targets macrophages and T lymphocytes (see Section 5.9.4). However, the
prominent effects are largely on immune system function, rather than overt cytotoxicity to
lymphoid tissues. Lead suppresses Thl-dependent responses (e.g., delayed type
hypersensitivity) and production of Thl cytokines; and stimulates macrophages into a
hyperinflammatory state. These types of functional changes have not been rigorously evaluated
in human epidemiological studies, which have relied, for the most part, on changes in
lymphocyte abundance as the main outcomes for assessing status of cellular immune systems.
Lead-induced functional changes in immune responses may not be reflected in changes in
lymphocyte abundance and, correspondingly, specific functional changes may not be readily
discerned from observed changes in lymphocyte abundance. Studies of children have found
significant associations between increasing blood Pb level and decreases in T-cell abundance,
with corresponding increases in B-cell abundance (Karmaus et al., 2005; Sarasua et al., 2000;
Zhao et al., 2004). These effects have been observed in children whose blood Pb concentrations
were <10 |ig/dL (Karmaus et al., 2005; Sarasua et al., 2000), although not all studies (e.g., Lutz
et al., 1999) have found such associations at higher blood Pb levels (e.g., 10-45 jig/dL). Studies
of occupational Pb exposures have also found associations between increasing blood Pb levels
and changes (increases or decreases) in T-cell abundance (Fischbein et al., 1993; Pinkerton et al.,
1998; Sata et al., 1997). Effects were observed in association with blood Pb concentrations
6-204
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<25 |ig/dL (Fischbein et al., 1993) and in populations whose blood Pb levels ranged from ~7 to
55 |ig/dL (Pinkerton et al., 1998; Sata et al., 1997). Outcomes from these studies are
qualitatively summarized in Table 6-6 and are discussed in greater detail below.
Several cross-sectional studies have examined possible associations between Pb exposure
and biomarkers of cellular immunity in children (Karmaus et al., 2005; Lutz et al., 1999; Sarasua
et al., 2000; Zhao et al., 2004). Three studies (Karmaus et al., 2005; Sarasua et al., 2000; Zhao
et al., 2004) found significant associations between increasing Pb exposure and decreases in
T-cell abundance (Table 6-6). The largest study (Sarasua et al., 2000) examined abundance of
total lymphocytes, T-cells (CD3+), B-cells (CD20+), NK cells, and CD4+ and CD8+ T-cell
phenotypes in infants, children, and adolescents. Associations between increasing blood Pb
concentration and increasing B-cell abundance (% and number), and decreasing T-cell
abundance (%) were found for children 6-35 months of age (n = 312), after adjustment for age,
sex, and study site (of four mining/smelting sites). Comparison of adjusted means for outcomes
across blood Pb strata revealed that the differences were significant for the > 15 |ig/dL stratum
only, compared to the <5 |ig/dL stratum. The Karmaus et al. (2005) study examined children in
the age range of 7 to 10 years (n = 331) with blood Pb levels <5 |ig/dL. In addition to age and
sex, regression models relating outcomes to blood Pb concentration included exposure to
environmental tobacco smoke and infections in the previous year as covariates. Similar to the
Sarasua et al. (2000) study, Karmaus et al.(2005) found significant associations between blood
Pb and decreased T-cell abundance (CD3+, CD3+CD8+) and increased B-cell (CD19+) abundance
(for the blood Pb quartile 2.2 to 2.8 |ig/dL) (Figure 6-18). Zhao et al. (2004) examined
lymphocyte phenotype abundance in children in the age range 3 to 6 years (n = 73) and found
significantly lower % abundance of T-cell phenotypes CD3+CD4+, CD4+CD8+ and significantly
higher abundance of D3+CD8+ cells in children whose blood Pb concentrations were > 10 |ig/dL
compared to <10 |ig/dL. Lutz et al. (1999) found no significant associations between blood Pb
concentration and age-adjusted T-cell (CD3+) or B-cell (CD 19+) abundance or abundance of
various other lymphocyte phenotypes (i.e., CD2+, CD25+, CD28+, CD71+) in children whose
blood Pb concentrations were 10 to 14, 15 to 19, or 20 to 45 |ig/dL compared to <10 |ig/dL.
A larger set of studies have evaluated potential associations between Pb exposure and
biomarkers of cellular immunity in adults (Basaran and Undeger, 2000; Cohen et al., 1989;
Coscia et al., 1987; Fischbein et al., 1993; Kuo et al., 2001; Mishra et al., 2003; Pinkerton et al.,
6-205
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Table 6-6. Summary of Results of Selected Studies of Associations Between Lead Exposure
and Serum Lymphocyte Abundances
Oi
to
o
Oi
Blood Lead (jig/dL)
Study
Children
Karmaus et al. (2005)
Lutzetal. (1999)
Sarasua et al. (2000)
Zhao et al. (2004)
Adults
Fischbein et al. (1993)
Pinkerton et al. (1998)
Sarasua et al. (2000)
Sataetal. (1997)
Subjects

Children, 7-10 yr
Children, 9 mo-6 yr
Children, 6-35 mo
Children, 3-6 yr

Firearms instructors
Smelter workers
General population
Lead stearate workers
na

331
270
372
73

87
229
433
99
Mean (SD)

3
NR
7
NR

3 1(4?
39k
4.3
19
PTib
Range

1-51
1-45 o
-2-161
-3-40 o

NR
<2-55 o
-1-101 o
7-50 o
TC PTI d PTI C
H AC AHC

o - NR
NR NR NR
o o NR
+

o NR
000
000
o + NR
T f
AM

0
NR
NR
NR

NR
+
NR
-
NKg

0
NR
0
NR

0
0
0
NR
Bh

-
0
+
0

+
+
0
0
-, decrease; +, increase; o, no effect; NR, not reported.

a Total number of subjects (including reference group)
b T-cells (CD3+)
c T-helper cells (CD4+)
d Cytotoxic T-cells (CD8+)
e CD4+CD8+
f T-memory cells (CD45RO+, CD45RA+)
g Natural killer cells (e.g., CD16+, CD56+)
h B-cells (e.g., CD19+, CD20+)
1 5th-95th percentile range
1 High exposure group
k Median
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1.12

1.08-

1.04-

1.00-

0,96-

0,92-

0.88-

0.84-
0.80
—S£— Natural killer cells

—9— T-cells

—4— T-helper cells
-Memory T-helper cells

• Cytotoxic T-cells

• B-cells
<2.2(n = 82) 2.2-2.8 (n = 81) 2.8-3.4 (n = 86)

Blood Lead Category (jjg/dL)
>3.4 (n = 82)
Figure 6-18. Relationship between blood lead concentration and T- and B-cell abundances
in children. Shown are relative changes in covariate-adjusted absolute cell
numbers (cells/uL) compared to the lowest blood lead group; adjusted for
age, number of infections in the previous 12 months, exposure to passive
smoke in the previous 12 months, and serum lipids (sum of cholesterol and
triglycerides). Abundances for T-cells, cytotoxic T-cells, and B-cells in the
2.2-2.8 ug/dL group were significantly different (p < 0.05) from the
<2.2 ug/dL group. Receptor phenotypes assayed were: T-cells, CD3+; T-
helper cells, CD3+CD4+; cytotoxic T-cells, CD3+CD8+; memory T-helper
cells, CD4+CD45RO+; natural killer cells, CD16+CD56+; B-cells,
CD3+CD5+CD19+.

Source: Karmaus et al. (2005).
1998; Sarasua et al., 2000; Sata et al., 1998, 1997; Yucesoy et al., 1997b; Undeger et al., 1996).

Four studies warrant particular attention because they implemented relatively stronger study

designs (i.e., cohort criteria, size, treatment of covariates): Fischbein et al., 1993; Pinkerton

et al., 1998; Sarasua et al., 2000; Sata et al., 1998). With one exception (Sarasua et al., 2000), all

were studies of relatively small occupational cohorts. The Sarasua et al. (2000) study included a

cross-sectional analysis of abundance of total lymphocytes, B-cells, NK cells, and CD4+ and

CD8+ T-cell phenotypes in individuals (n = 433), age 16 to 75 years. Associations were not
6-207
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found between blood Pb concentration and either B-cell or T-cell abundance, after adjusting for
age, sex, and study site (of four mining/smelting sites). The study did detect significant
associations among these variables in infants and children (see above discussion of cellular
immunity outcomes in children). However, all three occupational studies found significant
associations between increasing blood Pb concentrations and changes in abundance of
circulating T-cells with either no effect or an increasing B-cell abundance (Fischbein et al., 1993;
Pinkerton et al., 1998; Sata et al., 1997). The strengths of the Pinkerton et al. (1998) study have
been described previously with respect to outcome measures for humoral immunity. The study
included male smelter workers (n = 145, mean blood Pb 39 |ig/dL; range 15-55) and hardware
workers (n = 84, mean <2 |ig/dL, range <2-12). Covariate-adjusted significant outcomes were
an increase in B-cell (CD19+) abundance (% and number) and increases in CD4+CD45RA+ cell
abundance (%, number) in association with increasing blood Pb concentration. Covariate-
adjusted mean levels of monocytes (%), and T-cells (% D4+CD8+, CD8+CD56+) were lower in
Pb workers compared to the reference group.
The Fischbein et al. (1993) study examined a small group of firearms instructors (n = 51)
and age-matched reference subjects (n = 36). Fifteen of the instructors had blood Pb levels
>25 |ig/dL (mean 31.4, SD 4.3), the mean of the remaining 21 subjects was 4.6 |ig/dL (SD 4.6).
Mean blood Pb concentration of the reference group was reported as <10 |ig/dL. Increasing
blood Pb concentration was significantly associated with decreasing covariate-adjusted T-cell
(CD4+) abundance (Figure 6-19). Covariate-adjusted T-cell (CD3+% and number, CD4+% and
number, CD4+CD8+ number) abundance was significantly lower and B-cell (CD20+ cells % and
number) abundance was higher in the instructors than in the reference group.
The Sata et al. (1998) study included male Pb stearate manufacture workers (n = 71) and
a nonexposed reference group (n = 28). Mean blood Pb concentration was 19 |ig/dL
(range 7-50) in the Pb workers (reference group blood Pb concentration not reported).
Categorical covariate-adjusted Pb exposure classification (exposed, not exposed) was
significantly associated with lower T-cell (CD3+CD45RO+) number. Lead workers, relative to
the reference group, had significantly lower covariate-adjusted mean CD3+CD45RO+ number
and higher CD8+ cells (%).
The above observations of decreasing T-cell abundance in association with Pb exposure,
as assessed from blood Pb concentrations, is supported by results of several smaller cross-
6-208
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1.8
« 1.6 H
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3
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O
CO
c
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O
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1.2-
1.0-
0.8-
"75 0.6-
0.4
-B-cells
-Cytotoxic T-cells
-T-helper cells
Reference (<10)
(n = 36)
-Natural killer cells
- T-cells
***
Low Exposure (<25)
(n = 36)
High Exposure (>25)
(n = 15)
Exposure Category (Blood Lead,
Figure 6-19. Relationship between lead exposure and T- and B-cell abundances in
firearms instructors. Shown are relative changes in absolute cell numbers
compared to the reference group. Comparisons of exposed relative to the
reference group are shown as: * for p < 0.05; ** for p < 0.01; and *** for
p < 0.002. Receptor phenotypes assayed were: T-cells, CD3+; T-helper cells,
CD4+; cytotoxic T-cells, CD8+; natural killer cells, CD16+; B-cells, CD20+.
The CD4+/CD8+ ratio (not shown) was significantly lower in both the low
exposure (1.38 [SD 0.5], p < 0.002) and higher exposure group (0.95
[SD 0.5], p < 0.002), compared to the reference group (1.95 [SD 0.66]).

Source: Fischbein et al. (1993).
sectional studies, including Basaran and Undeger (2000), Coscia et al. (1987), and Undeger et al.

(1996), as well as a clinical study in which T-cell and NK cell abundance was found to increase
after CaEDTA chelation therapy of a Pb smelter worker (Sata et al., 1997). Lower serum levels

of the cytokines that function in the regulation of cellular immune responses, including IL-lp

and IFN-y, in Pb workers compared to nonexposed subjects have also been observed (Yiicesoy

etal., 1997a).
6-209
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6.8.5 Lymphocyte Function
Studies conducted in animal models have found mixed effects of Pb on mitogen-induced
lymphocyte activation, promoting expansion of some types of lymphoid populations, while
suppressing others (see Section 5.9.5). Lead promotes the activation of Th2-type lymphocytes
and suppresses Thl type lymphocytes; it also shifts the balance in the production of cytokines,
decreasing Thl cytokines (e.g., IFN, IL-12) and increasing production of Th2 cytokines
(e.g., IL-4, IL-6, IL-10). The above findings are somewhat echoed in the overall findings from
epidemiologic studies, with mixed outcomes when proliferation of peripheral lymphocytes was
the outcome measured, whereas, Pb preferentially stimulated Th2 cytokine production and
suppressed Thl cytokine production when human peripheral lymphocytes were exposed to Pb
in vitro.
Several studies (all of adults) have examined associations between Pb exposure in adults
and lymphocyte activation, assessed as a proliferative response to mitogens and/or antigens
(Alomran and Shleamoon, 1988; Cohen et al., 1989; Fischbein et al., 1993; Kimber et al., 1986;
Mishra et al., 2003; Pinkerton et al., 1998; Queiroz et al., 1994b). Results of these have been
mixed. Three studies found no significant associations between blood Pb concentrations in Pb
workers and lymphocyte proliferative response to activating agents (Kimber et al., 1986;
Pinkerton et al., 1998; Queiroz et al., 1994b). Four studies found decreasing proliferative
response with increasing blood Pb concentration (Alomran and Shleamoon, 1988; Cohen et al.,
1989; Fischbein et al., 1993; Mishra et al., 2003). The Alomran and Shleamoon (1988), Cohen
et al. (1989), Mishra et al. (2003), and Queiroz et al. (1994b) studies, which found significant Pb
associations, included subjects who had relatively high blood Pb levels (>60 |ig/dL) compared to
the Kimber et al. (1986) and Pinkerton et al. (1998) studies. The inclusion of subjects with
higher Pb concentrations may have contributed to the differences in outcomes.
As noted in the previous section, the Fischbein et al. (1993) and Pinkerton et al. (1998)
studies are particularly noteworthy because of the strengths of the cohort selection and the data
analyses which attempted to account for potential confounders. Also, these are the only reported
studies that examined antigen-specific lymphocyte activation in humans. Mean blood Pb levels
in the two studies were similar: 31 |ig/dL (SD 4) in the Fischbein et al. (1993) study and
39 |ig/dL (range 15-55) in the Pinkerton et al. (1998) study. Both studies found no significant
associations between blood Pb concentration and antigen-specific lymphocyte proliferation,
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assessed in the Pinkerton et al. (1998) study with tetanus toxoid as the antigen and in the
Fischbein et al. (1993) study with staphylococcus aureus as the antigen. However, the Fischbein
et al. (1993) study also measured mitogen-induced lymphocyte proliferation (induced with PHA
or PWM) and found a significantly lower proliferative response to the mitogens in association
with Pb exposure. This study also found a significant association between increasing blood Pb
concentration and decreasing proliferative response in mixed lymphocyte cultures (i.e.,
proliferative response of lymphocytes from exposed subjects when incubated with inactivated
lymphocytes from a reference subject).
Inorganic Pb has been shown by in vitro studies to perturb several aspects of lymphocyte
function when introduced into primary isolates of human blood monocytes. Activated
lymphocytes show altered lysosomal enzyme secretion and altered expression and secretion of
cytokines (Bairati et al., 1997; Guo et al., 1996a; Hemdan et al., 2005). Lymphocytes activated
with Salmonella enteritidis or with monoclonal antibodies of CD3, CD28 and CD40, and
exposed to inorganic Pb had suppressed expression of T-helper cell type TH-! cytokines,
interferon (IFN-y), interleukin (IL-1P), and tumor necrosis factor (TNF-a), whereas activation
by CD antibodies increased secretion of TH-2 cytokines, IL-5, IL-6, and IL-10 (Hemdan et al.
2005). Inorganic Pb also activates transcription factor NK-KP in CD4+ cells (Pyatt et al., 1996),
an important regulator of T-cell activation, and increases expression of MHC class II surface
antigens (HLA-DR), an important surface antigen in the CD4+ response to exogenous antigens
(Guo et al., 1996b). Lead increases antibody production in cultured human B-cells (McCabe and
Lawrence, 1991). These observations suggest that Pb may perturb cellular immune function
through a variety of mechanisms.

6.8.6 Phagocyte (Macrophage and Neutrophil) Function
Animal studies and in vitro models have shown that Pb can modulate macrophages into a
hyperinflammatory phenotype, with increased production of proinflammatory cytokines TNF-a
and IL-6, increased release of reactive oxygen intermediates and prostaglandins, and, conversely,
depressed production of nitric oxide (see Section 5.9.6). Epidemiologic studies have found
associations between blood Pb concentrations and modified activation of macrophages in
children whose blood Pb levels ranged from 4 to 50 |ig/dL (Pineda-Zavaleta et al., 2004).
Consistent with the above experimental observations, outcomes have included decreased
6-211
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stimulated nitric oxide release and increased superoxide anion production. In addition, studies
have observed suppressed PMNL chemotaxis in association with occupational exposures that
resulted in blood Pb concentrations of 12 to 90 |ig/dL (Bergeret et al., 1990; Queiroz et al.,
1994a, 1993).
Pineda-Zavaleta et al. (2004) examined mitogen (PHA)- and cytokine (INFy)-induced
activation of blood monocytes collected from 65 children (age range 6 to 11 years) who resided
near an active Pb smelter. Mean blood Pb concentrations of subjects at three schools located
8,100 meters, 1,750 meters, and 650 meters from the smelter were: 7.0 |ig/dL (range 3-25),
21 |ig/dL (range 11-49), and 30 |ig/dL (range 10-48), respectively. Endpoints measured
included nitric oxide and superoxide anion production, a response generally attributed to
activated macrophages. Increasing blood Pb concentration was significantly associated with
decreasing PHA-induced nitric oxide production and increasing INFy-induced superoxide anion
production. The mitogen, PHA, activates macrophages indirectly through activation of
lymphocytes, whereas INFy, a cytokine released from CD44 (Tnl) cells, directly activates
macrophages. Thus, one interpretation of this outcome is that Pb suppressed T-cell mediated
macrophage activation and stimulated cytokine-induced macrophage activation.
Possible associations between occupational Pb exposure and PMNL chemotaxis and
phagocytic activity have been explored in several small cross-sectional studies. Consistent
findings are significantly reduced chemotactic response and phagocytic activity (i.e., respiratory
burst, luminal uptake) in Pb workers compared to reference groups. The largest study is that of
Queiroz et al. (1993, 1994a) which evaluated PMNL function in several (possibly overlapping)
cohorts of Pb battery manufacture workers (n = 60). Blood Pb concentrations in the study
groups ranged from 12 to 90 |ig/dL. PMNL chemotaxis and lytic activity were significantly
lower in the Pb workers compared to the reference group. Bergeret et al. (1990) assessed PMNL
chemotaxis and phagocytosis in a group of battery smelting workers (n = 34) and in a group of
reference subjects (n = 34) matched to the Pb worker group by age, sex, ethnic origin, smoking
and alcohol consumption habits, and intake of antibiotics and NS AIDs. Mean blood Pb levels
were 71 |ig/dL (SD 18) in the Pb workers and 9 |ig/dL (SD 4) in the reference group.
Significantly lower PMNL chemotactic response to FMLP and phagocytic response in opsonized
zymosan were significantly lower in the Pb workers than in the reference group. Lead
6-212
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introduced into primary cultures of human PMNLs suppressed chemotaxis and phagocytosis
(Governa et al., 1987).

6.8.7 Summary of the Epidemiologic Evidence for the Effects of Lead
on the Immune System
Studies conducted in animals and in vitro experimental models have shown that Pb can
alter immune system function (see Section 5.9). Lead appears to target, preferentially,
macrophages and T lymphocytes; although, effects on B cells and neutrophils have also been
reported. The prominent effects are largely on immune system function, rather than overt
cytotoxicity to lymphoid tissues. Lead suppresses Thl-dependent responses (e.g., delayed type
hypersensitivity) and production of Thl cytokines and shifts the Thl/Th2 balance towards
Th2 responses; increases the production of IgE and Th2 cytokines (e.g., 11-4); and stimulates
macrophages into a hyperinflammatory state. These types of functional changes have not been
rigorously evaluated in human epidemiologic studies, which have relied, for the most part, on
changes in lymphocyte abundance or circulating immunoglobulin levels, as the main outcomes
for assessing status of cellular immune systems. The above outcomes may be relatively
insensitive to for detecting disturbances in humoral or cellular immune function. Few studies
have attempted to examine associations between Pb exposure and integrated immune function
(e.g., host resistance, hypersensitivity, autoimmunity) and current epidemiologic evidence for
associations between Pb exposure and compromised immune function in humans, reflected in
risk of asthma or infections, is not compelling, but these studies may not have been adequate to
address this.
• Several epidemiological studies have examined possible associations between Pb
exposures and various indices of humoral and cellular immune status. Findings from
these studies suggest that Pb exposure (as reflected in blood Pb concentration) may be
associated with changes in serum immunoglobulin levels; perturbation of peripheral
lymphocyte phenotype profiles, including decreases in peripheral blood T-cell abundance
and changes in T-cell :B-cell abundance ratios; modulation of lymphocyte activation
(increased stimulated lymphocyte release of reactive oxygen intermediates and
suppressed production of nitric oxide; increased production of Th2 cytokines and
suppression of Thl cytokines); and suppression of neutrophil chemotaxis and
phagocytosis. Observations of increased circulating levels of IgE, increased release of
reactive oxygen intermediates and suppressed production of nitric oxide in peripheral
lymphocytes are of particular interest in that such effects have been consistently observed
in studies conducted in animals and in vitro model.
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• Studies in children have consistently found significant associations between increasing
blood Pb concentration and increasing serum IgE. These effects have been observed at
blood Pb concentrations <10 |ig/dL. Findings of studies of adults have been mixed with
significant associations between blood Pb (>30 |ig/dL) and serum immunoglobulin levels
and no association in a study group in which blood Pb concentrations were <10 |ig/dL.

• Studies in children have also found significant associations between increasing blood Pb
concentration and decreases in T-cell abundance, with corresponding increases in B-cell
abundance. These effects have been observed in children whose blood Pb concentrations
were <10 |ig/dL, although not all studies have found such associations at higher blood Pb
concentrations (e.g., in the 10 to 45 |ig/dL range).

• Studies of occupational Pb exposures have also found associations between increasing
blood Pb concentration and decreasing T-cell abundance. Effects were observed in
association with blood Pb concentrations <25 |ig/dL and in populations whose blood Pb
concentrations ranged from ~7 to 55 |ig/dL.

• Studies of lymphocyte and phagocyte (i.e., macrophage, neutrophil) function have found
associations between blood Pb concentrations and modulation of the activation of
lymphocytes and macrophages in children whose blood Pb concentrations ranged from
4 to 50 |ig/dL, suppressed PMNL chemotaxis in association with occupational exposures
that resulted in blood Pb concentrations of 12 to 90 |ig/dL, and suppressed mitogen-
induced activation of peripheral lymphocytes in adults in association with occupational
exposures that resulted in blood Pb levels that ranged from 15 to 55 |ig/dL. Consistent
with observations made in animal models, Pb exposures in vitro suppressed production of
Thl cytokines and stimulated production of Th2 cytokines in isolates of peripheral
lymphocytes.
6.9 EFFECTS OF LEAD ON OTHER ORGAN SYSTEMS

6.9.1 Biochemical Effects of Lead

6.9.1.1 Summary of Key Findings of the Biochemical Effects of Lead from the
1986 Lead AQCD

The 1986 Lead AQCD provided an extensive discussion of the effects of Pb on heme
biosynthesis and on quantitative relationships between exposure and effects in humans. Lead
interferes with heme synthesis by inhibiting the enzymes ALAD and ferrochelatase. As a
consequence, heme biosynthesis decreases, relieving the rate-limiting enzyme of the heme
synthesis pathway, 5-aminolevulinic synthetase (ALAS), from negative feedback inhibition by
heme (Figure 6-20). The outcomes of decreased activity of ALAD and ferrochelatase, and
increased activity of ALAS are increased urinary excretion of coproporphyrin (CP) and
6-214
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MITOCHONDRION
MITOCHONDRIA!. MEMBRANE
GLYCINE
SUCCINYL-CoA
HEME
ALASYNTHETASE
(INCREASE)

_ Pb (DIRECTLY OR
BY DEREPRESS1ON)
FERRO-
CHELATASE --
IRON + PROTOPORPHYRIN
Pb ,
AMINOLEVULIN1C ACID
(ALA)
ALA
DEHYDRASE
(DECREASE)
Pb
PORPHOBiLINOGEN
IRON
CORPROPORPHYRIN
(INCREASE)
Figure 6-20. Effects of lead on heme biosynthesis.
Source: Derived from EPA (1986).
5-aminolevulinic acid (ALA), increased level of ALA in blood plasma, and increased erythrocyte
protoporphyrin (EP) levels.
Associations between Pb exposure and blood ALAD activity and EP levels, and urinary
ALA and CP excretion have been studied extensively in adults and children, and quantitative
relationships between exposure and effect are well understood. Much of this information was
available prior to completion of the 1986 Lead AQCD and is discussed in detail in that criteria
document (e.g., Alessio et al., 1976; Hernberg et al., 1970; Lilis et al., 1978; Piomelli et al.,
1982; Roels et al., 1979; Selander and Cramer, 1970; Valentine et al., 1982). Numerous studies
published since the 1986 AQCD provide additional support for the Pb concentration-response
relationships in humans described in the 1986 AQCD. The most pertinent new studies are
summarized in Annex Tables AX6-9.1 and AX6-9.2. The studies that provide the strongest basis
6-215
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for empirically-derived expressions relating blood Pb concentration, blood ALAD activity,
urinary ALA, and EP are listed in Table 6-7 and are discussed below.
Since completion of the 1986 Lead AQCD, a literature has developed on the effects of Pb
on serum and blood lipids, including cholesterol levels and indications of oxidative stress, in the
form of lipid peroxides, depletion of erythrocyte reduced glutathione (GSH), and production of
reactive oxygen species (ROS). These studies also are summarized in Annex Tables AX6-9.1
and AX6-9.2, and key findings are discussed below.

6.9.1.2 Heme Biosynthesis
6.9.1.2.1 ALAD Inhibition
Numerous studies published since the 1986 AQCD have explored associations between
Pb exposure and inhibition of ALAD activity, as assessed from measurements of blood ALAD
activity (Gurer-Orhan et al., 2004; Kim et al., 2002; Lee et al., 2000; Makino et al., 1997; Roels
and Lauwerys, 1987; Schuhmacher et al., 1997), or urinary ALA excretion (Gennart et al., 1992;
Oishi et al., 1996; Schuhmacher et al., 1997; Wildt et al., 1987; Soldin et al., 2003). Quantitative
estimates derived from the larger, more recent studies are presented in Table 6-7. Blood Pb
concentration is inversely correlated with the log of blood ALAD activity and log of urinary
ALA and quantitative estimates of the change in blood. ALAD activity per unit change in blood
Pb concentration are consistent across studies (observed blood Pb range 5 to 150 |ig/dL).
Halving of blood ALAD activity occurs with an increase in blood Pb concentration of-20 |ig/dL
in both children (Roels and Lauwerys, 1987) and adults (Morita et al., 1997). These estimates
are consistent with earlier studies of adults (e.g., Hernberg et al., 1970) and children (e.g.,
Alessio et al., 1976, 1977), discussed in the 1986 AQCD. Greater variability is apparent
in estimates of the change in urinary ALA per unit change in blood Pb concentration (Table 6-7).
This may be related, in part, to gender-heterogeneity in the relationship. Roels and Lauwerys
(1987) estimated that urinary ALA doubles in association with a 20 |ig/dL increase in blood Pb
concentration in females and 50 |ig/dL in males. In a much larger study (Oishi et al., 1996), an
analysis that combined data from males (n = 253) and females (n = 165) found that a doubling of
urinary ALA occurred in association with a 13.7 |ig/dL increase in blood Pb concentration.
Urinary ALA excretion increases as a linear function of plasma ALA concentration (Oishi et al.,
1996); thus, the gender heterogeneity for the blood Pb-urinary ALA relationship may derive
6-216
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Table 6-7. Blood Lead-Response Relationships for Heme Synthesis Biomarkers in Adults and Children
Study
Age
Blood Lead
Oig/dL)
Range
Regression Equation (r)
Blood Lead Change
Oig/dL)
Predicted to Halve or
Double Effect Biomarker
to
ALAD Activity Decrease

Roels and Lauwerys (1987)

Alessioetal. (1976, 1977)

Hernbergetal. (1970)

Moritaetal. (1997)

Urinary ALA Increase

Roels and Lauwerys (1987)

Alessioetal. (1976, 1977)

Gennartetal. (1992)

Oishietal. (1996)

Selander and Cramer (1970)

Roels and Lauwerys (1987)

Roels and Lauwerys (1987)
143
169
158
58
37
316
183
418
150
39
36
10-13 yr
Adult (M)
Adult (M, F)
Adult (M)
10-13 yr
Adult (M)
Adult (M, F)
Adult (M, F)
Adult (M, F)
Adult (M)
Adult (F)
5-41
15-150
5-95
2-82
20^1
10-150
4-75
10-99
6-90
10-60
7-53
log[ALAD] = 1.864-0.015[blood Pb] (r= 0.87)

log[ALAD] = 3.73-0.031 [blood Pb] (r= 0.87)

log[ALAD] = 2.274-0.018[blood Pb] (r = 0.90)

log[ALAD] = 1.8535-0.00971[bloodPb] (r = 0.76)

log[ALAU] = 0.94+0.11 [blood Pb] (r= 0.54)

log[ALAU] = 1.25+0.014[bloodPb] (r= 0.62)

log[ALAU] = 0.37+0.008[bloodPb] (r= 0.64)

log[ALAU] = -0.387+0.022[bloodPb] (r = 0.71)

log[ALAU] = -1.0985+0.157[bloodPb] (r = 0.74)

log[ALAU] = 0.37+0.006[blood Pb] (r = 0.41)

log[ALAU] = 0.15+0.015[bloodPb] (r= 0.72)
20.1

22.4

16.1

20.1

20.9

49.5

37.6

13.7

19.2

50.2

20.1
-------
Septembel
^
to
o
o
ON

ON
to
oo

Table 6-7. (cont'd). Blood Lead-Response Relationships for Heme Synthesis Biomarkers in Adults and Children

Study
EP Increase
Marcus and Schwartz
(1987)
Piomelli et al. (1982)
Roels and Lauwerys (1987)
Soldin et al. (2003)

Alessioetal. (1976, 1977)
Alessioetal. (1976, 1977)
Gennart etal. (1992)
Roels and Lauwerys (1987)
Roels and Lauwerys (1987)
Wildt etal. (1987)
Wildt etal. (1987)

1,677

2,002
51
4,908

95
93
183
39
36
851
139

Age

2-6

2-12
10-13
0-17

Adult (M)
Adult (F)
Adult (M)
Adult (M)
Adult (F)
Adult (M)
Adult (F)
Blood Lead
Oig/dL)
Range

6-65

2-98
15-U
<1-103

10-90
10-70
4-75
10-60
7-53
10-80
10-80

Regression Equation (r)

Nonlinear kinetic model

log[EP] = 1.099+0.016[bloodPb] (r= 0.509)
log[EP] = 1.321+0.025[bloodPb] (r= 0.73)
EP=-0.0015[bloodPb]3+0.1854[bloodPb]2-
2.7554[bloodPb]+30.911 (r = 0.999)
log[EP] = 0.94+0.01 17[bloodPb]
log[EP] = 1.60+0.0 143 [blood Pb]
log[EP] = 0.06+0.019[bloodPb] (r= 0.87)
log[EP] = 1.41+0.014[bloodPb] (r= 0.74)
log[EP] = 1.23+0.027[bloodPb] (r = 0.81)
log[EP] = 1.21+0.0148[bloodPb] (r = 0.72)
log[EP] = 1.48+0.0113[bloodPb] (r= 0.56)
Blood Lead Change
Oig/dL)
Predicted to Halve or
Double Effect Biomarker

20 -40a

18.8
12.0
20.6

25.7
21.1
15.8
21.1
11.1
20.3
20.6
ALA, 5-aminolevulinic acid; ALAD, 5-aminolevulinic acid dehydratase; ALAU, urinary 5-aminolevulinic acid; EP, erythrocyte protoporphyrin; F, female;
M, male.

""Approximately 20 ug/dL at low transferrin saturation (<31%), approximately 40 ug/dL at higher transferrin saturation (>31%).
-------
from a gender difference in the effect of Pb on plasma ALA concentration or from differences in
renal plasma clearance of ALA.

6.9.1.2.2 ALAD Polymorphism
ALAD is a polymorphic enzyme with two alleles (ALAD1 and ALAD2) and three
genotypes: ALAD 1-1, ALAD 1-2, and ALAD 2-2 (Battistuzzi et al., 1981). The corresponding
phenotypes appear to have nearly identical catalytic properties (Battistuzzi et al., 1981).
The predominant genotype is ALAD 1-1 which has a prevalence of-90% (Astrin et al., 1987;
Battistuzzi et al., 1981; Hsieh et al., 2000; Shen et al., 2001). A significantly higher percentage
(p = 0.03) of erythrocyte Pb was bound to ALAD in carriers of the ALAD2 allele (84%)
compared to carriers of the ALAD1 allele (81%); however, no differences were evident in the
distribution of Pb between erythrocytes and plasma (Bergdahl et al., 1997), and there is no
evidence that the ALAD genotype confers different sensitivity to inhibition of heme biosynthesis
(Hsieh et al., 2000; Perez-Bravo et al., 2004; Schwartz et al., 1997; Suzen et al., 2003).

6.9.1.2.3 Ferrochelatase Inhibition
Lead inhibition of ferrochelatase results in an accumulation of protoporphyrin IX in
erythrocytes (EP, also referred to as zinc protoporphyrin, or ZPP). Numerous studies have
examined relationships between blood Pb concentration and EP levels in adults and children.
Quantitative estimates based on the most pertinent studies are presented in Table 6-7. Results
across these studies are similar (observed blood Pb range: <1 to 103 |ig/dL). In both children
and adults (males and females), a doubling of EP levels occurs in association with an increase in
blood Pb concentration of-20 |ig/dL (Marcus and Schwartz 1987; Piomelli et al., 1982; Soldin
et al., 2003; Wildt et al., 1987). However, the relationship between blood Pb concentration and
EP level is not linear (Marcus and Schwartz, 1987; Soldin et al., 2003). The slope of the blood
Pb concentration range over which a change (threshold) in EP occurs is relatively small and
appears to extend to -20 |ig/dL in iron replete children but decreases with increasing iron
deficiency (Marcus and Schwartz, 1987). A pronounced gender difference in the relationship
between EP and blood Pb concentration was observed by Roels and Lauwerys (1987) which was
not observed in the much larger study of Wildt et al. (1987).
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Sakai et al. (2000) examined the relationship between ALAD genotypes and disturbances
in heme biosynthetic pathway upon Pb exposure in 192 males occupationally exposed to Pb and
125 controls. ALAD1 homozygotes had significantly higher EP levels compared to ALAD2
carriers at blood Pb values >20 |ig/dL, suggesting that they might be more susceptible to
disturbances in heme metabolism caused by Pb exposure.
Inhibition of ferrochelatase also gives rise to an increase in urinary coproporphyrin, with a
similar relationship to blood Pb concentration; a doubling of urinary EP occurs in association
with an increase in urinary coproporphyrin of-20 |ig/dL (Alessio et al., 1976).

6.9.1.3 Effects on Blood Lipids: Cholesterol
Associations between occupational exposure to Pb and changes in blood lipid
composition have been observed. These include increased levels of lipid peroxides in blood
and/or serum (Ito et al., 1985; Jiun and Hsien, 1994; Sugawara et al., 1991) and increased serum
levels of total and HDL cholesterol (Kristal-Boneh et al., 1999). Increased levels of glucose-
6-phosphate dehydrogenase (G6PD) in erythrocytes have also been observed in Pb workers
(Cocco et al., 1995; Gurer-Orhan et al., 2004).
Kristal-Boneh et al. (1999) measured serum total, HDL, and LDL cholesterol, and
triglycerides in a group of male battery manufacture workers. Covariate-adjusted serum total-
cholesterol and HDL cholesterol levels were 6% and 12% higher, respectively, in Pb workers
(n = 56, mean blood Pb 42 |ig/dL, SD 15) compared to reference group (mean blood Pb:
2.7 |ig/dL). Increasing blood Pb concentration was significantly associated with increasing
covariate-adjusted total cholesterol and HDL cholesterol. A similar outcome was found in a
larger study (Ito et al., 1985) of male steel workers (n = 712, blood Pb range 5-62 |ig/dL).
When stratified by age, total and HDL cholesterol levels in serum were 3.6% and 7.5% higher,
respectively, in Pb workers in the age range 40 to 49 years, compared to corresponding strata of
the office workers (n = 155). Although a smaller study, the Kristal-Boneh et al. (1999) study
considered a larger set of potential covariables (e.g., dietary fat, cholesterol, and calcium intakes,
sport activities, alcohol consumption, cigarette smoking).
Oxidative changes in blood lipids (e.g., increased levels of lipid peroxides and
malondialdehyde levels) as well as decreased levels of erythrocyte superoxide dismutase (SOD),
catalase, G6PD, and GSH peroxidase, indicative of increased oxidative stress, have been
6-220
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observed in Pb workers, in comparison to reference groups (Ito et al., 1985; Jiun and Hsien,
1994; Solliway et al., 1996; Sugawara et al., 1991). However, none of these studies have
developed concentration-response relationships that take into account potential confounders.
The largest study is that of (Ito et al., 1985), described above. When stratified by age, serum
lipoperoxide levels were 16% higher in the Pb workers in the age range 40 to 49 years, compared
to corresponding strata of the reference group. Serum lipoperoxide levels also appeared to
increase as blood Pb increased above 30 |ig/dL, while erythrocyte SOD appeared to decrease
with increasing blood Pb concentration (a statistical evaluation was not reported).
Evidence for increased oxidative stress (increased reactive oxygen species) in
lymphocytes of Pb workers has also been reported (Fracasso et al., 2002). Peripheral
lymphocytes collected from battery manufacture workers (n = 37, mean blood Pb: 40 |ig/dL)
exhibited increased DNA strand breaks, higher production of ROS and lower GSH levels
compared to a reference group of office workers (n = 29, mean blood Pb 4 jig/dL).
The covariate-adjusted odds ratios (exposed versus not exposed) were 1.069 (95% CI: 1.020,
1.120) for increased DNA strand breaks and 0.634 (95% CI: 0.488, 0.824) for lower GSH levels.

6.9.2 Effects of Lead on the Hematopoietic System
6.9.2.1 Summary of Key Findings of the Effects of Lead on the Hematopoietic System
from the 1986 Lead AQCD
The 1986 Lead AQCD concluded that Pb decreases heme production and shortens
erythrocyte survival; both effects contributing to Pb-induced anemia in children and adults,
which becomes evident in children at blood Pb concentrations >40 |ig/dL and, in adults,
>50 |ig/dL. The 1986 Lead AQCD also concluded that effects of Pb on blood hemoglobin level
extend below 50 |ig/dL, with effects detected in Pb workers at blood Pb concentrations
<25 |ig/dL (Baker et al., 1979; Grandjean, 1979). More recent epidemiologic studies,
summarized below, provide additional information on concentration-response relationships for
hematopoietic effects of Pb. The studies support the conclusion that clinical anemia can occur in
children in association with blood Pb levels >40 |ig/dL (Schwartz et al., 1990). The newer
studies suggest that perturbation of erythropoiesis, indicated by changes in serum erythropoietin,
occurs in association with blood Pb concentrations <40 |ig/dL and in the absence of detectable
changes in blood hemoglobin levels or hematocrit. Details regarding the design of these studies
6-221
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and outcomes are presented in Annex Tables AX6-9.3 and AX6-9.4. Outcomes of the most
pertinent studies are discussed below.

6.9.2.2 Blood Hemoglobin Levels
Several studies reported since the completion of the 1986 Lead AQCD have explored
associations between Pb exposure and blood hemoglobin levels in children and adults.
Consistent findings have been a lack of discernable depression of blood hemoglobin levels in
study populations whose mean blood Pb concentrations were <40 |ig/dL (Table 6-8). Of note is
the findings relating patella bone Pb to both blood hemoglobin levels and hematocrit.
The Kosovo prospective study of pregnancy outcomes is one of the largest epidemiologic
evaluations of associations between Pb exposure and blood hemoglobin levels in infants and
children (Graziano et al., 2004; Factor-Litvak et al., 1998, 1999). The study included pregnant
women (n = 1502) and their children (n = 311) who resided in one of two regions of Kosovo,
Yugoslavia; one was heavily impacted by Pb industries (high-Pb area), the other had relatively
little Pb contamination (low-Pb area). Mean blood Pb concentrations of children (measured at
birth and at intervals to 12 years of age) ranged from 30 to 40 |ig/dL in the high-Pb area and 6 to
9 |ig/dL in the low-Pb area. Mean blood hemoglobin levels in the low-Pb and high-Pb children,
measured at 4.5, 6.5, 9.5, and 12 years of age, were not significantly different. These findings
are consistent with those from a smaller cross-sectional study (n = 89; blood Pb range 2 to
84 |ig/dL, 84% <35 |ig/dL) that also found no association between blood Pb concentration and
blood hemoglobin levels (Liebelt et al., 1999). Results from these two studies suggest that, in
the absence of iron deficiency, Pb exposures that result in blood Pb levels <40 |ig/dL do not
produce detectable changes in blood hemoglobin levels in children.
Associations between Pb exposure and blood hemoglobin levels in adults have been
examined in numerous epidemiological studies (Froom et al., 1999; Gennart et al., 1992;
Horiguchi et al., 1991; Hu et al., 1994; Makino et al., 1997; Poulos et al., 1986; Romeo et al.,
1996; Solliway et al., 1996). The Graziano et al. (1990) and Makino et al. (1997) studies warrant
particular attention because of the design (longitudinal), relatively large size (>1000 subjects),
and relatively low blood Pb levels of the subjects (<40 |ig/dL). Both studies support the general
conclusion that blood hemoglobin levels are not depressed in association with blood Pb
concentrations <40 |ig/dL. In the Kosovo prospective study, no discernable effect of Pb on
6-222
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Table 6-8. Summary of Results of Selected Studies of Associations Between Lead Exposure and Blood Hemoglobin Levels
Blood Lead Oig/dL)
Study
Children
Graziano et al. (2004)
Liebelt etal. (1999)
Adults
Graziano etal. (1990)
-------
maternal blood hemoglobin levels was evident from a comparison of the high-Pb exposure
group (mean blood Pb 17 |ig/dL, range 7-43) with the low-Pb exposure group (mean blood Pb
5.1 |ig/dL, range 2-11). Makino et al. (1997) found a positive association between increasing
blood Pb concentration and increasing blood hemoglobin levels in a longitudinal survey of adult
males (n = 1,573) who worked in pigment or vinyl chloride stabilizer manufacture (mean blood
Pb 13 |ig/dL, range 1-39). A simple linear regression model predicted a 10 |ig/dL increase in
blood hemoglobin per 10 |ig/dL increase in blood Pb concentration (typical level 10-20 |ig/dL).
Two other cross-sectional studies are also notable, because of design considerations
and/or blood Pb concentration ranges of the subjects. Solliway et al. (1996) observed no
differences in mean blood hemoglobin levels in a comparison of adult male battery manufacture
workers (n = 34; mean blood Pb 41 |ig/dL, range 23-63) and a matched reference group (n = 56;
mean blood Pb 7 |ig/dL, range 1-13). Hu et al. (1994) conducted a cross-sectional assessment of
adult male carpentry workers (n = 119) whose blood Pb levels were <25 |ig/dL. Blood
hemoglobin was not significantly associated with blood Pb concentration. Of note, however,
was the finding that increasing patella bone Pb was significantly associated with decreasing
blood hemoglobin levels. Covariate-adjusted blood hemoglobin levels were predicted to
decrease by 1.1 g/dL per 37 jig/g increase (mean of first and fourth quartiles) in patella bone Pb.
Studies of Pb workers whose blood Pb levels were higher than in the studies noted above
have, in general, found lower blood hemoglobin levels in association with increasing blood Pb
concentrations; these include Gennart et al. (1992) with a blood Pb range of 40 to 70 |ig/dL,
Horiguchi et al. (1991) with a mean blood Pb level of 54 |ig/dL (SD 16), and Poulos et al. (1986)
with mean blood Pb range of 21 to 27 |ig/dL. In the latter study (Poulos et al., 1986), blood
hemoglobin levels decreased by 0.6 to 0.9 g/dL per 10 |ig/dL increase in blood Pb (simple linear
regression) in adult males. Analyses or adjustments for potential covariables were not reported
for these studies.

6.9.2.3 Erythrocyte Volume and Number
Schwartz et al. (1990) conducted a concentration-response analysis of data collected at the
Bunker Hill smelter site in Idaho in 1974, shortly after the failure of the smelter bag house
resulted in extensive contamination of the surrounding area with uncontrolled smelter emissions.
This analysis is unique in that it collected hematocrit measurements in children (n = 579, age
6-224
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range 1 to 5 years) who had relatively high blood Pb levels (range 11-164 |ig/dL, -40%
exceeded 40 |ig/dL). A logistic model relating blood Pb concentration and age to hematocrit
predicted a 10% decrease in hematocrit (from 39.5 to 35.5%) in association with blood Pb
concentrations of 85, 115, and 145 |ig/dL at ages 1, 3, and 5 years, respectively (Figure 6-21).
A 10% probability of anemia (hematocrit <35%) was predicted in association with a blood Pb
concentration of-20 |ig/dL at age 1 year, 50 |ig/dL at age 3 years, and 75 |ig/dL at age 5 years
(Figure 6-21).
Numerous studies of associations between Pb exposure and erythrocyte volume (e.g.,
hematocrit) or number have been reported in adults (Gennart et al., 1992; Horiguchi et al., 1991;
Hsiao et al., 2001; Hu et al., 1994; Makino et al., 1997; Osterode et al., 1999; Poulos et al., 1986;
Solliway et al., 1996). The Hu et al. (1994) and Makino et al. (1997) studies examined groups of
workers that had blood Pb concentrations that were relatively low, compared to other studies,
and found either no association or weak association between blood Pb concentration and
hematocrit and/or erythrocyte number. The Hu et al. (1994) cross-sectional study of carpentry
workers (n = 119, blood Pb concentration range 2-25 |ig/dL) found no association between
blood Pb concentration and hematocrit; however, increasing patella bone Pb was associated with
a significant decrease in hematocrit. Covariate-adjusted blood hematocrit was predicted to
decrease by 0.03% (95% CI: 0.01, 0.05) per 37 |ig/g increase (mean of first and fourth quartiles)
in patella bone Pb. The Makino et al. (1997) longitudinal study of pigment and vinyl chloride
stabilizer manufacture workers (n = 1,573; blood Pb range 1-39 |ig/dL) found a positive
association between blood Pb concentration and hematocrit, and erythrocyte count. A simple
linear regression model predicted an increase in hematocrit of 0.6 (typically 43) and an increase
in erythrocyte count of 0.07 x 106/mm3 (typically 4-7 x 106/mm3) per 10 |ig/dL increase in
blood Pb concentration.
Studies that included subjects who had higher blood Pb concentrations (i.e., >40 |ig/dL)
have, in general, found negative associations between blood Pb concentration and hematocrit
Gennart et al., 1992; Horiguchi et al., 1991; Poulos et al., 1986; Solliway et al., 1996), with two
exceptions, Hsiao et al. (2001) and Osterode et al. (1999). Hsiao et al. (2001) conducted an
11-year retrospective longitudinal analysis of blood Pb concentration, hematocrit, and
erythrocyte count in a group of battery manufacture workers (n = 30; mean blood Pb 30-
60 jig/dL). A repeated measures regression analysis (generalized estimation equation) yielded
6-225
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40
38
2-.
•"£ 36
u
I
i 34 H
32
30
20
Age 3 years
Age 5 years
40 60 80 100
Blood Lead (pg/dL)
120
140
.Q
O
0.8

0.7

0,6

0.5

0.4

0.3

0.2

0.1

0,0
Age 1 year

Age 2 years
Age 3 years

Age 4 years

Age 5 years
50 100 150
Blood Lead (jig/dL)
200
Figure 6-21. Relationship between blood lead and hematocrit in children. The top panel
shows central tendency predictions based on a logistic regression model
relating hematocrit and blood lead concentration, adjusted for age. The
regression coefficients relating hematocrit and blood lead were (p = 0.0133
[SE 0.0041], p = 0.0005). The bottom panel shows corresponding
concentration-response (hematocrit <35%) relationships.

Source: Schwartz etal. (1990).
6-226
-------
a significant association between increasing blood Pb concentration and increasing hematocrit
and erythrocyte count. Osterode et al. (1999) measured erythrocyte number and packed cell
volume in a group of Pb workers (n = 20) and an age-matched reference group (n = 20). Mean
blood Pb concentration was 45.5 |ig/dL (range 16-91) in the Pb workers and 4.1 |ig/dL (range 3-
14) in the reference group. Mean erythrocyte number and packed cell volume in the Pb workers
and reference group were not different.

6.9.2.4 Erythropoiesis
Several studies have found associations between Pb exposure and serum erythropoietin
levels in children (Graziano et al., 2004; Liebelt et al., 1999) and adults (Graziano et al., 1991;
Osterode et al., 1999; Romeo et al. 1996). A qualitative summary of outcomes from these
studies are provided in Table 6-9.
Two studies have examined possible association between Pb exposure and serum
erythropoietin levels in children. In the Kosovo prospective study (Factor-Litvak et al., 1998,
1999; Graziano et al., 2004) a significant association was evident between increasing blood Pb
concentration (3-70 |ig/dL) and increasing serum erythropoietin levels after adjustment for age
and blood hemoglobin levels (Figure 6-22). The association weakened with age; it was
significant at ages 4.5 and 6.5 years, but not at ages 9.5 or 12 years. A multivariate linear
regression model predicted a 36% increase in serum erythropoietin per 10 |ig/dL increase
(3-13 |ig/dL, hemoglobin 13 g/dL) in blood Pb at age 4.5 years and an 18% increase per
10 |ig/dL at age 6.5 years. These outcomes suggest that erythropoiesis is stimulated in children
in association with increasing blood Pb concentrations <40 |ig/dL and in the absence of
depressed blood hemoglobin levels.
A smaller cross-sectional study examined serum erythropoietin levels in a group of
children (n = 89), 1 to 6 years of age (Liebelt et al., 1999). The blood Pb level range in the study
group (2-84 |ig/dL) was similar to that in the Graziano et al. (2004) study and, consistent with
this study, Liebelt et al. (1999) found no association between blood Pb concentration and serum
hemoglobin levels. However, in contrast to the Graziano et al. (2004) study, blood hemoglobin-
adjusted serum erythropoietin levels decreased in association with an increase in blood Pb
concentration (0.3 mlU/mL decrease per 10 |ig/dL blood Pb increase). The Liebelt et al. (1999)
study did not include age as a covariate in the regression model, which was shown in the Kosovo
6-227
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Table 6-9. Summary of Results of Selected Studies of Associations Between Lead Exposure and Serum Erythropoietin

Oi
to
to
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10
£
LU
E
9 -
8-
7 '
6
8
o>
CO
4-
3 '
Age 12 years
\
Age 4.5 years
10
20
30 40
Blood Lead
50
60
70
Figure 6-22. Relationship between blood lead and serum erythropoietin in children.
Coefficients relating erythropoietin and blood lead were significant for
ages 4.5 (p = 0.21 [95% CI: 0.13, 0.30], p < 0.0001) and 6.5 years (p = 0.12
[95% CI: 0.03, 0.20], p < 0.001).
Source: Graziano etal. (2004).
prospective study to be a significant covariable in blood Pb-serum erythropoietin relationship
(Graziano et al., 2004); this may have contributed to the different outcome in the two studies.
Liebelt et al. (1999) studied a convenience sample from a Pb/primary care clinic (rather than a
prospectively selected cohort) that specifically excluded children who had symptoms of severe
iron deficiency, or were taking iron supplements or other bone marrow suppressing drugs. Iron
status of the children in the Graziano et al. (2004) study was not reported. However, serum
ferritin levels in the mothers, at mid-pregnancy, was not indicative of iron deficiency (Graziano
et al., 1990). Although the direction of the outcome measure was different in the two studies,
both studies (Graziano et al., 2004; Liebelt et al., 1999) found evidence for an effect of Pb
exposure on serum erythropoietin levels in the absence of significant Pb-associated changes in
blood hemoglobin levels.
6-229
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Three studies have found associations between Pb exposure and changes in erythropoiesis
biomarkers in adults. As part of the Kosovo prospective study, serum erythropoietin was
measured at mid-pregnancy and at term in a subset of women enrolled in the study (Graziano
et al., 1991). The high- and low-Pb cohorts were constructed from the six highest and lowest
mid-pregnancy blood Pb concentrations, within each of four blood hemoglobin strata, ranging
from 9.0 to 12.9 g/dL. Mean blood Pb concentrations in the strata ranged from 17 to 39 |ig/dL in
the high-Pb group and 2.4 to 3.6 |ig/dL in the low Pb group. Serum erythropoietin levels
significantly decreased in association with increasing blood Pb concentration, independently of
an effect of blood hemoglobin (Figure 6-23). Romeo et al. (1996) also found an association
between increasing blood Pb concentration and decreasing serum erythropoietin, in the absence
of discernable changes in blood hemoglobin levels, in a comparison of group male Pb workers
(n = 28, blood Pb range 30-92 |ig/dL) and a similar-aged reference group (n = 113, mean blood
Pb 10 |ig/dL [range 3-20]). Osterode et al. (1999) examined several measures of erythropoiesis
in a group of Pb workers (n = 20, mean age 46 years) and in an age-matched reference group
(n = 20). Mean blood Pb concentration was 45.5 |ig/dL (range 16-91) in the Pb workers and
4.1 |ig/dL (range 3-14) in the reference group. Mean blood hemoglobin levels in the Pb worker
and reference groups were not different. Lead workers with blood Pb concentrations >60 |ig/dL
had significantly lower circulating erythrocyte progenitor cells than the reference group. Also,
erythrocyte progenitor cell number was significantly negatively correlated with blood Pb and
urine Pb concentrations. Serum erythropoietin levels increased exponentially with decreasing
packed blood cell volume in the reference group, but not in the Pb workers (i.e., serum
erythropoietin level was not significantly correlated with packed cell volume in the Pb workers).
Thus, unlike the reference group (blood Pb concentration < 14 |ig/dL), Pb workers appeared to
have a suppressed erythropoietin response to declining blood cell volume.
Collectively, the above studies suggest that Pb exposure depresses serum erythropoietin
levels, in the absence of significant depression in blood hemoglobin levels. Lead-induced
nephrotoxicity may contribute to a suppression of erythropoietin levels in Pb-exposed
individuals. Although this cannot be entirely ruled out in these studies, both the Romeo et al.
(1996) and Osterode et al. (1999) studies excluded people who had a history of hematological or
kidney disease. Nevertheless, renal nephrotoxicity, including proximal tubular nephropathy,
6-230
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80
j"
| 70-

^ 60-
•i
2 50-
c

c 40^
O
o
.E 30-
Q. 20-
2
10-
til
(4.4)
(23.1)
- "Low" Blood Lead

- "High" Blood Lead

- Mean Blood Lead
(3,9)
(3.0)
(28.8)
(33.2)
_L
(3.0) (36.2)
J_ T
9.0-9.9 10,0-10,9 11.0-11.9 12.0-12.9

Hemoglobin Concentration Category (g/dL)

Figure 6-23. Association between blood lead concentration and serum erythropoietin
in pregnant women. Shown are combined data for mid-pregnancy and
delivery. Each bar represents the mean (±SD) of 12 subjects. ANOVA of the
data at mid-pregnancy and at delivery showed blood lead effects (p = 0.049, p
= 0.055, respectively) and blood hemoglobin effects (p = 0.0001, p = 0.009,
respectively), with no significant interaction between the two variables.

Source: Graziano etal. (1991).
could have been a confounder in these studies which included subjects with blood Pb levels

>40 |ig/dL.

6.9.2.5 Other Effects on Erythrocyte Metabolism and Physiology

6.9.2.5.1 Erythrocyte Nucleotide Metabolism

Lead inhibits erythrocyte pyrymidine-5 'nucleotidase (P5N) and adenine dinucleotide

synthetase (NADS). Inhibition of P5N leads to the accumulation of pyrimidine nucleotides in

the erythrocyte and hemolysis. Associations between increasing blood Pb concentration and

decreasing blood P5N and NADS activity have been observed in studies of Pb workers (Kim
6-231
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et al. 2002; Mohammed-Brahim et al., 1985; Morita et al., 1997). Mean blood Pb levels in these
study groups were >35 |ig/dL and ranged up to 80 |ig/dL. Inhibition of P5N has also been
observed in children whose blood Pb concentrations were >30 |ig/dL (Angle and Mclntire, 1978;
Angle et al., 1982; summarized in the 1986 Lead AQCD).

6.9.2.5.2 Erythrocyte Deformability
Horiguchi et al. (1991) compared the deformability of erythrocytes collected from adult
male secondary Pb refinery workers (n = 17, age range 24 to 58 years) with a reference group of
male subjects (n = 13, age range 22 to 44 years). Erythrocyte deformability was assessed as
microfilterability of erythrocytes under a negative (-10 cm H^O) pressure head. Erythrocytes
from the Pb workers showed significantly lower deformability compared to the reference group.
The mean blood Pb concentration in the Pb workers was 53.5 |ig/dL (SD 16.1); blood Pb values
for the reference group were not reported.

6.9.2.5.3 Erythrocyte Membrane Transport
Hajem et al. (1990) measured erythrocyte membrane activities of Na+-K+-ATPase, Na+-
K+-cotransport, Na+-Li+-antiport, and passive Na+ and K+ permeability in erythrocytes collected
from adult males (n = 122; geometric mean blood Pb 16 |ig/dL, range 8.0-33.0; geometric mean
hair Pb 5.3 |ig/g, range 0.9-60). Na+-K+-cotransport activity was negatively correlated with
blood Pb concentration, but not with hair Pb. Na+-K+-ATPase activity was negatively correlated
with hair Pb, but not with blood Pb.

6.9.3 Effects of Lead on the Endocrine System
6.9.3.1 Summary of Key Findings of the Effects of Lead on the Endocrine System from
the 1986 Lead AQCD
The 1986 Lead AQCD concluded that various endocrine processes may be affected by
Pb at relatively high exposure levels. These included effects on thyroid hormone levels (e.g.,
Refowitz, 1984; Robins et al., 1983), effects on male sex hormone levels (e.g., Braunstein et al.,
1978), and impairment of the production of 1,25-dihydroxy vitamin D (1,25-(OH)2D3) (e.g.,
Rosen et al., 1980). Effects on these endocrine systems were concluded to be apparent only at
blood Pb concentrations exceeding 30-40 |ig/dL. The 1986 Lead AQCD concluded that studies
6-232
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from which the effects of Pb on reproductive hormones in females could be assessed were
lacking.
More recent epidemiologic studies have examined possible associations between Pb
exposure (as reflected by blood and/or bone Pb levels) and various biomarkers of endocrine
function, including the thyroid, male reproductive, and calcitropic endocrine systems. These
studies have examined endocrine outcomes at lower blood Pb ranges and in the absence of overt
clinical Pb toxicity, and have more rigorously attempted to control for confounding factors.
Evidence for Pb effects on these systems, in association with blood Pb concentrations <30-
40 |ig/dL, remains absent. The strongest study designs have yielded no associations, or weak
associations, between Pb exposure and thyroid hormone status (Erfurth et al., 2001; Schumacher
et al., 1998; Tuppurainen et al., 1988; Zheng et al., 2001). Similarly, studies of the male
reproductive system that attempted to control for confounding effects of age, have yielded mixed
outcomes (Alexander et al., 1996a, 1998; Erfurth et al., 2001; Gustafson et al., 1989; McGregor
and Mason, 1990; Ng et al., 1991). Results of a more recent epidemiologic study of the
calcitropic endocrine system in children suggest that associations between serum vitamin D
status and blood Pb may not be present in calcium-replete children who have average lifetime
blood Pb concentrations <25 |ig/dL (Koo et al., 1991). In adults, exposures to Pb that result in
blood Pb concentrations >40-60 |ig/dL may increase, rather than decrease, circulating levels of
1,25-(OH)2D3 and PTH (Kristal-Boneh et al., 1999; Mason et al., 1990), possibly as a
compensatory response to increased urinary calcium losses secondary to impaired kidney
function. Details regarding the design of these studies and outcomes are presented in Annex
Tables AX6-9.5 and AX6-9.6. Outcomes of the most pertinent studies are summarized below.

6.9.3.2 Thyroid Endocrine Function
Several studies have examined possible associations between Pb exposure and thyroid
hormone status. Most of these have been studies of occupational exposures. The results of these
studies have been mixed; some studies have found significant associations with Pb exposure
(e.g., blood Pb concentration), but most studies have found none or relatively weak associations.
In studies that have controlled for the effects of age, outcomes also have been mixed, with the
strongest study designs finding none or weak associations between Pb biomarkers and thyroid
hormone status (Erfurth et al., 2001; Schumacher et al., 1998; Tuppurainen et al., 1988; Zheng
6-233
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et al., 2001). The strength of the association and, possibly, the direction of the effect (i.e.,
increase or decrease in hormone levels) may change with exposure duration or level (Robins
et al., 1983; Tuppurainen et al., 1988). The overall picture that emerges is that those studies that
have included subjects having blood Pb levels >100 |ig/dL have found depression of serum T3
and/or T4 levels, without a detectable increase in serum TSH. However, studies in which the
blood Pb distribution was dominated by levels well below 100 |ig/dL, have found either no
effects or subclinical increases in serum T3, T4, with no change in TSH levels. Outcomes from
the most pertinent studies are summarized qualitatively in Table 6-10 and are described in
greater detail below.
Siegel et al. (1989) measured serum total thyroxine (TT4) and free thyroxine (FT4)
in children ages 11 months to 7 years (n = 68) who were outpatients at a clinical care facility.
Mean blood Pb concentration in the study group was 25 |ig/dL (range 2-77). In a simple
(univariate) linear regression analysis, hormone levels were not significantly associated with
blood Pb concentration.
Zheng et al. (2001) measured concentrations of TT4 and transthyretin (TTR) in serum and
cerebral spinal fluid (CSF) of adult hospital patients (n = 82) admitted for evaluation of CSF
clinical chemistry (e.g., for head wounds, tumors, neurological symptoms). Mean blood Pb
concentration was 14.9 |ig/dL (SD 8.3). Age-adjusted serum TT4 and TTR, and CSF TT4 were
not significantly associated with blood Pb concentration; however, increasing CSF Pb level was
associated with decreasing CSF TTR levels (r = -0.30, p = 0.023).
Possible associations between Pb exposure and thyroid hormone status have been
examined in several studies of Pb workers (Dursun and Tutus, 1999; Erfurth et al., 2001; Gennart
et al., 1992; Gustafson et al., 1989; Horiguchi et al., 1987; Lopez et al., 2000; Refowitz, 1984;
Robins et al., 1983; Schumacher et al., 1998; Singh et al., 2000; Tuppurainen et al., 1988).
Of these, six warrant particular attention because the design and/or analysis attempted to control
for effects of age (Erfurth et al., 2001; Dursun and Tutus, 1999; Gustafson et al., 1989;
Schumacher et al., 1998; Tuppurainen et al., 1988; Robins et al., 1983). Outcomes of these
studies are summarized in Table 6-10. The largest studies were Erfurth et al. (2001),
Schumacher et al. (1998), and Tuppurainen et al. (1988).
Erfurth et al. (2001) was a cross-sectional study of secondary smelter workers (n = 62)
and a reference group of metal (not Pb) workers (n = 26). Excluded from the study were
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to
Table 6-10. Summary of Results of Selected Studies of Associations Between Lead Exposure and
Thyroid Hormone Levels
Blood Lead (jig/dL)
Study
Children
Siegel etal. (1989)
Adults
Dursun and Tutus (1999)
Erfurth etal. (2001)
Gustafson etal. (1989)
Robins etal. (1983)
Schumacher etal. (1998)
Tuppurainen et al. (1988)
Zheng etal. (2001)
Subjects

Children, 1 1 mo-7 yrs

Metal powder manufacture workers
Secondary smelter workers
Secondary smelter workers
Brass foundry workers
Primary smelter workers
Battery manufacture workers
General population
na

57
88
42
47
151
176
82
Mean (SD)

17.1 (9.0)
31. 1°
39.4(2.1)
NR
24.1
55.9 (23.8)
14.9(8.3)
Range

2-77 NR

1-36 +
4-93 o
NR o
16-127 NR
<15to>40% o
5-134
NR NR
T4 TSH

o NR

+ o/ob
0 0
+ 0
NR
0 0
0
o NR
-, decrease; +, increase; o, no effect; NR, not reported; T3, triiodothyronine; T4, thyroxine; TSH, thyroid stimulating hormone.

a Total number of subjects (including reference group)
b Basal/thyroid releasing hormone-stimulated
0 Median
-------
individuals with ongoing thyroid disease or who were taking thyroid hormone supplements or
other drugs that would interfere with thyroid hormone levels (e.g., beta-blockers). Median blood
Pb concentration in the Pb workers was 31 |ig/dL (range 8-93). Age-adjusted basal serum levels
of FT3, FT4, and TSH were not associated with blood, urine, or finger bone Pb levels. Thyroid
releasing hormone (TRH)-induced TSH secretion (area under serum TSH concentration-time
curve) was measured in an age-matched subset of the study group (9 Pb workers and
11 reference subjects) and was not significantly different in the two groups. The Schumacher
et al. (1998) study measured serum FT4, TT4, and TSH levels in a group of male workers
(n = 151) at the Trail British Columbia smelter complex. Excluded from the study were
individuals who had ongoing clinical thyroid disease. Mean blood Pb concentration in the study
group was 24 |ig/dL (15% >40 jig/dL). Covariate-adjusted (age, alcohol consumption) hormone
levels were not significantly associated with current blood Pb concentration or 10-year average
blood Pb concentrations. Prevalence of abnormal hormone values was also unrelated to blood
Pb concentration.
Tuppurainen et al. (1988) measured serum total triiodothyronine (TT3), FT4, TT4,
and TSH levels in a group of male battery manufacture workers (n = 176). Mean blood Pb
concentration was 56 |ig/dL (range 14-134). Although, hormone levels were not significantly
associated with blood Pb concentrations, increasing exposure (i.e., employment) duration was
significantly associated with decreasing FT4 (r2 = 0.071, p = 0.001) and TT4 (r2 = 0.059,
p = 0.021) levels. The r2 was not improved by including age or blood Pb as covariables.
Strength of the association was greater when the analysis was restricted to workers who had an
exposure duration >7.6 years (FT4: r2 = 0.33, p < 0.002; TT4: r2 = 0.21, p < 0.001). Consistent
with the results of the Tuppurainen et al. (1988) study, Robins et al. (1983) found a significant
association between increasing blood Pb concentration and decreasing FT4 (r2 = 0.085,
p = 0.048) in a group of brass foundry workers (n = 47). The blood Pb range in the subjects was
16-127 |ig/dL. When stratified by race (Black, White) the association was significant in the
Black stratum (r2 = 0.21, p = 0.03), but not in the White stratum (r2 = 0.05, p = 0.27).
The strength of association was not changed by including age in the regression model. Both
the Robins et al. (1983) and Tuppurainen et al. (1988) included subjects with blood Pb levels
>100 |ig/dL.
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Blood Pb concentrations were lower in the Dursun and Tutus (1999) and Gustafson et al.
(1989) studies than in the above studies, and both studies found significant associations between
Pb exposure and increasing serum TT4 levels. Dursun and Tutus (1999) measured serum FT3,
TT3, FT4, TT4, and TSH in a group of metal powder manufacture workers (n = 27) and a
reference group (n = 30). Mean blood Pb concentration in the workers was 17 |ig/dL
(range 9-36). A linear regression model that included age, blood Pb concentration, and exposure
duration, indicated a significant association between increasing exposure duration and increasing
serum TT4 levels (r2 = 0.3, p = 0.03). The Gustafson et al. (1989) study examined a group of
male secondary smelter workers (n = 21) and reference subjects, individually matched to the Pb
workers by age, sex, and work shift. Mean blood Pb concentration in the workers was 39 |ig/dL
(SD 2). Serum TT4 levels were significantly higher (p < 0.02) in the Pb workers compared to
the reference group. The difference strengthened when the analysis was restricted to the age
range <40 years (p = 0.01).

6.9.3.3 Reproductive Endocrine Function
6.9.3.3.1 Male Reproductive Endocrine Function
Low testosterone (TES) levels, blunted sex hormone secretion in response to
gonadotropin releasing hormone (GnRH), and defects in spermatogenesis have been observed in
humans exhibiting clinical neurological symptoms of Pb poisoning (Braunstein et al., 1978;
Cullen et al., 1984). However, the effects of lower exposure levels on reproductive endocrine
status are less clear. Possible associations between Pb exposure and changes in male
reproductive hormone levels have been examined in studies of Pb workers. Of these, five studies
attempted to control for effects of age, an important determinant of testosterone levels
(Alexander et al., 1998; Erfurth et al., 2001; Gustafson et al., 1989; McGregor and Mason, 1990;
Ng et al., 1991). The outcomes from these studies are qualitatively summarized in Table 6-11.
Blood Pb ranges in the latter studies were similar (4 to 90 |ig/dL), yet outcomes were mixed,
with observations of no change (Erfurth et al., 2001; Gustafson et al., 1989; McGregor and
Mason, 1990) or subclinical decrease (Alexander et al., 1996a, 1998; Ng et al., 1991) in serum
testosterone (TES) in association with Pb exposure. Mixed effects were observed for the effect
of Pb exposure on serum follicle stimulating hormone (FSH) and luteinizing hormone (LH),
6-237
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Table 6-11. Summary of Results of Selected Studies of Associations Between Lead Exposure and
Male Sex Hormone Levels in Adults

Oi
K>
OJ
oo
Study
Alexander et al. (1996a, 1998)
Erfurthetal. (2001)
Gustafson et al. (1989)
McGregor and Mason (1990)
Ngetal. (1991)
- Hprrpacp* 4- inrrpacp* n nn pfff
Subjects
Primary smelter workers
Secondary smelter workers
Secondary smelter workers
Lead workers
Battery manufacture workers
^rt* XTR nnt rpnnrtpH* F*sW fnllirlp ctii
Blood Lead (jig/dL)
"a Mean(SD) Range FSH LH PRL TES
152 NR 5-58 o o NR -b
88 31.1° 4-93 o/-^ o/od o/od od
42 39.4(2.1) NR - - o o
176 NR 17-77 + + NR o
171 35(13) 10-72 + + o -
Tnilntina hnrmnnp* TM liitpiniTina hnrmnnp* PRT nrnl^rtin* TR^l tpctnctprnnp
b
Total number of subjects (including reference group)
In association with increasing semen lead levels, not with blood lead
Median
Basal/gonadotropin releasing hormone-stimulated
Effect was evident in comparison between groups, but not in multivariate regression that adjusted for age
-------
increases (McGregor and Mason, 1990; Ng et al., 1991), decreases (Gustafson et al., 1989), and
with no change (Alexander et al., 1996a, 1998; Erfurth et al., 2001) in hormone levels observed.
The inconsistency in the direction of effects on TES and the two androgen regulating
pituitary hormones, FSH and LH, is particularly noteworthy, and suggest the possibility of
multiple effect of Pb on the hypothalamic-pituitary-gonad axis, consistent with observations that
have been made in some experimental animal studies. Erfurth et al. (2001) observed a
suppressed FSH response to GnRH in a group of Pb workers compared to an age matched
reference group; however, the magnitude of the response was not significantly associated with
Pb exposure indices in a multivariate regression analysis that accounted for age. In rats, Pb
exposure can suppress serum testosterone levels in the absence of a change in circulating levels
of GnRH or LH, even though levels of GnRH mRNA increase in the hypothalamus (Klein et al.,
1994; Ronis et al., 1996; Sokol et al. 2002). Thus, changes in GnRH production, at the
molecular level, do not necessarily translate to changes in hormone levels. This may be the
result of Pb inhibition of release of GnRH for nerve terminals in the median eminence (Bratton et
al., 1994; Sokol, 1987; Sokol et al., 1998, 2002).
Alexander et al. (1996a, 1998) examined serum FSH, LH, and TES in males (n = 152)
who worked at the Trail British Columbia smelter complex. Covariate-adjusted hormone levels
and prevalence of clinically abnormal values were unrelated (p > 0.05) to blood Pb level (range
5-58 |ig/dL); however, increasing semen Pb concentration (range 0.3-17 |ig/dL) was
significantly associated with decreasing semen testosterone levels (p = 0.004). Erfurth et al.
(2001) measured serum TES, sex hormone binding globulin (SHBG), and GnRH-stimulated
changes in serum FS, LH, and PRL in male secondary smelter workers (n = 62) and in
a reference group (n = 26). Mean blood Pb in the Pb workers was 31 |ig/dL (range 8-93).
Age-adjusted basal hormone levels were unrelated to blood, plasma, or urine Pb concentrations.
In an age-matched subset of the cohorts (n = 9 Pb workers, n = 11 reference), median GnRH-
stimulated serum FSH was significantly lower in Pb workers than in the reference group;
however, GnRH-stimulated LH, FSH, and PRL were not significantly associated with any of the
Pb measures in a multivariate regression analysis. Gustafson et al. (1989) measured serum FSH,
LH, and TES (total and free) in a group of male secondary smelter workers (n = 21) and in a
group of reference subjects individually matched to the Pb workers by age, sex, and work shift.
Mean blood Pb concentrations were 39 |ig/dL (SD 2) in the Pb workers and 5.0 |ig/dL (SD 0.2)
6-239
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in the reference group. Serum FSH levels were significantly lower (p = 0.009) in Pb workers
compared to reference group. When the analysis was restricted to the age range <40 years, Pb
workers had significantly lower FSH and LH compared to the reference group. McGregor and
Mason (1990) measured serum FSH, LH, TES, and SHBG in a group of male Pb workers
(n = 90) and in a reference group (n = 86). Blood Pb range in the Pb workers was 17-77 |ig/dL;
blood Pb concentrations in the reference subjects were <12 |ig/dL. Prevalences of abnormal
hormone levels in the Pb workers and reference group were not different; however, age-adjusted
serum FSH was significantly higher in Pb workers compared to reference group and increasing
FSH levels were significantly associated with increasing blood Pb concentrations. Increasing
serum LH was significantly associated with increasing exposure duration but not with blood Pb
concentration or age. Serum TES or SHBG levels were unrelated to blood Pb concentration or
exposure duration. Ng et al. (1991) measured serum FSH, LH, PRL, and TES in a group of male
battery manufacture workers (n = 122) and a reference group (n = 49). Mean blood Pb levels
were 35 |ig/dL (range 10-77) in the Pb workers and 8 |ig/dL (range 3-15) in the reference group.
When cohorts were stratified by age, serum FSH and LH levels were significantly higher in Pb
workers <40 years of age compared to corresponding age stratum of the reference group; serum
TES was significantly lower in Pb workers >40 years of age, compared to the same age stratum
in the reference group. Covariate-adjusted (age, tobacco smoking) serum TES levels were
significantly lower in Pb workers in the 10-year exposure duration stratum, compared to the
reference group. Covariate-adjusted serum FSH and LH were significantly higher in Pb workers
in the <10-year exposure duration stratum, compared to the reference group.

6.9.3.3.2 Female Reproductive Endocrine Function
Although delays in sexual maturation in humans have been associated with increases in
blood Pb concentrations (Selevan et al., 2003; Wu et al., 2003b), and Pb has been shown to alter
levels of female sex hormones and the menstrual cycle in nonhuman primates (Foster, 1992;
Franks et al., 1989; Laughlin et al., 1987), epidemiologic studies of interactions between Pb
exposure and reproductive endocrinology in females have not been reported. Lead introduced
into cultures of human ovarian granulosa cells suppresses progesterone production (Paksy et al.,
2001) and suppresses expression of aromatase and estrogen receptor P (Taupeau et al., 2003).
6-240
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6.9.3.4 Pituitary and Adrenal Endocrine Function
Several studies of possible associations between Pb exposure and levels of pituitary
hormones that regulate production and secretion of thyroid hormones (see Section 6.9.3.2) and
reproductive hormones (see Section 6.9.3.3) have been reported. In addition to the above
studies, Gustafson et al. (1989) found that serum cortisol levels were lower in a group of male
secondary smelter workers (n = 21) compared to a reference group individually matched to the
Pb workers by age, sex, and work shift. Mean blood Pb concentrations were 39 |ig/dL (SD 2) in
the workers and 5.0 |ig/dL (SD 0.2) in the reference group. Campbell et al. (1985) measured
various biomarkers of status of the renin-angiotensin-aldosterone system in male welders (n = 5)
and reference subjects (n = 8). Mean blood Pb concentration was 35 |ig/dL (range 8-62 jig/dL).
Significant positive correlations were observed between blood Pb concentration and plasma
aldosterone (r = 0.53, p < 0.002), which may have been, at least in part, secondary to a Pb effect
on plasma renin activity (r = -0.76, p < 0.001) and angiotensin I levels (r = 0.68, p < 0.002).
Saenger et al. (1984) found lower urinary levels of 6-p-OH-cortisol, but not cortisol, in children
who had elevated urinary Pb in an EDTA provocation test (>500 |ig/24 h), compared to children
who did not have elevated urinary Pb levels, or whose blood Pb levels were <30 |ig/dL.
The change in urinary excretion of 6-p-OH-cortisol in the absence of a change in cortisol levels
may reflect an effect of Pb on liver cytochrome P450 activity, rather than an effect on the adrenal
gland (see Section 6.9.4).

6.9.3.5 Calcitropic Endocrine Function
Children exposed to relatively high level of Pb >30 |ig/dL may exhibit depressed levels of
circulating 1,25-(OH)2D3 (Mahaffey et al., 1982; Rosen et al., 1980). These effects were not
detected in a study of calcium-replete children with average lifetime blood Pb levels below
25 |ig/dL (Koo et al., 1991). In adults, Pb exposures that result in blood Pb concentrations
>40-60 |ig/dL may increase, rather than decrease, circulating levels of 1,25-(OH)2D3 and PTH.
These studies also are summarized in Annex Tables AX6-9.5 and AX6-9.6. Outcomes from the
more pertinent studies are qualitatively summarized in Table 6-12 and are discussed in greater
detail below.
Epidemiologic studies of possible associations between Pb exposure and vitamin D status
in children have yielded mixed results. Mahaffey et al. (1982) and Rosen et al. (1980) observed
6-241
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o\
to
to
Table 6-12. Summary of Results of Selected Studies of Associations Between Lead Exposure
and Calcitropic Hormones
Study
Children
Koo etal. (1991)
Mahaffey etal. (1982)
Rosen etal. (1980)
Adults
Chalkley etal. (1998)
Kristal-Boneh et al. (1998)
Mason etal. (1990)
Subjects

Ages: 21,27, 33 mo
Ages: 1-16 yr
Ages: 1-5 yr

Smelter workers0
Battery manufacture workers
Lead workers
Blood Lead
"a Mean (SD)

105 9.7
177 NR
45 18, 47, 74b

19 47
140 43
138 NR
Oig/dL)
Range

5-24
12-120
10-120

21-76
1-77
15-95
PTH CAL 1,25D 25D

0000
oo-o
+ o - -

NR NR +c o
+ NR + NR
o NR + NR
-, decrease; +, increase; o, no effect; NR, not reported; PTH, parathyroid hormone; CAL, calcitonin; 1,25D, 1,25-dihydroxyvitaminD; 25D, 25-
hydroxy vitamin D.

a Total number of subjects (including reference group)
b Group means: low, moderate, high
0 Cadmium, lead, zinc smelter workers, effect on 1,24D in association with high blood cadmium and lead and high urinary cadmium
-------
lower 1,25-(OH)2D3 in association with increasing blood Pb concentration. Koo et al. (1991)
found no association between 1,25-(OH)2D3 and blood Pb concentration. The Koo et al. (1991)
study was a longitudinal analysis of a subset of a prospective study of pregnancy outcomes.
Serum calcium, magnesium, phosphorus, PTH, CAL, 25-OH-D3, 1,25-(OH)2D3, and bone
mineral content were measured in children (n = 105) at ages 21, 27, and 33 months. Mean
lifetime average blood Pb concentrations (based on quarterly assessments) was 9.7 |ig/dL (range
4.8-23.6). The range of highest values observed was 6 to 63 |ig/dL. A structural equation model
was developed that initially considered age, sex, race, sampling season, and dietary intake of
calcium, phosphorus, and vitamin D as covariables; the final model retained age, sex, race, and
sampling season. Decreasing blood Pb (In-transformed) was significantly associated with
covariate-adjusted decreasing serum phosphorus. No other covariate-adjusted outcomes were
significantly associated with blood Pb. The distribution of dietary calcium intakes was 4% for
<600 mg/day, 55% for 600-1200 mg/day, and 41% for >1200 mg/day. Intakes of phosphorous
were similar, suggesting that the subjects were nutritionally replete with respect to these two
nutrients.
The different outcomes in Koo et al. (1991) compared to the Mahaffey et al. (1982) and
Rosen et al. (1980) studies may reflect, in part, the lower blood Pb range in the subjects in
Koo et al. (1991) (range of lifetime average 5-24 |ig/dL, range of observed highest values 6-
63 |ig/dL) compared to the Mahaffey et al. (1982) and Rosen et al. (1980) studies (10-
120 jig/dL). Subjects in the Koo et al. (1991) study also had higher calcium intakes (4% with
<600 mg/day, 43% with >1200 mg/day) than in the Rosen et al. (1980) study (mean 580 mg/day
[SE 15] in high blood Pb group). Calcium intake (and/or related nutritional factors) may also
have been an uncontrolled confounder in the Rosen et al. (1980) study, as higher blood Pb
concentration appeared to be associated with lower calcium intakes (Sorrell et al., 1977).
Mahaffey et al. (1982) did not report calcium intakes. Thus, the effect of Pb exposure on vitamin
D status may be more pronounced at higher blood Pb concentrations (i.e., >60 |ig/dL) and in
combination with lower intakes of calcium (or other nutritional limitations).
Studies of Pb workers have found evidence for higher serum levels of 1,25-(OH)2D3 and
PTH in association with increasing blood Pb concentration (Chalkley et al., 1998; Kristal-Boneh
et al., 1998; Mason et al., 1990). The Chalkley et al. (1998) study was a small study (n = 19) of
subjects exposed to both Cd and Pb, and effects of Pb and Cd on 1,25-(OH)2D3 could not be
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isolated. The Kristal-Boneh et al. (1998) and Mason et al. (1990) studies included larger samples
of subjects whose exposure was primarily, but not exclusively, to Pb. Attempts were made to
control for effects of age and, in the Kristal-Boneh et al. (1998) study, other potential
covariables. Kristal-Boneh et al. (1998) measured serum calcium, magnesium, phosphorus,
PTH, 25-OH-D3, and 1,25-(OH)2D3 in a group of male battery manufacture workers (n = 56) and
a reference group (n = 90). Mean blood Pb concentrations were 43 |ig/dL (SD 14, range 1-77)
in the Pb worker group and 4.5 |ig/dL (SD 2.6, range 1.4-19) in the reference group. Serum
1,25-(OH)2D3 and PTH, but not 25-OH-D3, were significantly higher in Pb workers compared to
the reference group. Increasing blood Pb concentration (In-transformed) was significantly
associated with covariate-adjusted increasing serum PTH and 1,25-(OH)2D3 levels. No effects
on serum calcium were apparent. Occupational Pb exposure was also significantly associated
with increasing PTH and 1,25-(OH)2D3 level. Covariates retained in the multivariate model
were age, alcohol consumption, smoking; calcium intake, magnesium intake, and calorie intake.
Mason et al. (1990) measured serum calcium, phosphate, PTH, and 1,25-(OH)2D3 in male Pb
workers (n = 63) and in a reference group (n = 75) and found significantly higher prevalence of
elevated 1,25-(OH)2D3 (defined as >2 SD higher than reference mean) in Pb workers (13%)
compared to the reference group (1.3%). Serum levels of 1,25-(OH)2D3 were also significantly
higher in Pb workers compared to the reference group. After stratification of the Pb workers into
exposure categories (high exposure: blood Pb >40 |ig/dL and bone Pb >40 |ig/g; low exposure:
blood Pb <40 |ig/dL and bone Pb <40 |ig/g), serum 1,25-(OH)2D3 levels were significantly
higher in the high Pb group. Serum calcium levels were not different in the two groups.
Increasing blood Pb was significantly associated with increasing 1,25-(OH)2D3 levels (r2 =
0.206; with age and bone Pb included, r2 = 0.218). After excluding 12 subjects whose blood Pb
concentrations >60 |ig/dL, the regression coefficient was no longer significant (r2 = 0.162,
p = 0.26).

6.9.4 Effects of Lead on the Hepatic System
6.9.4.1 Summary of Key Findings of the Effects of Lead on the Hepatic System
from the 1986 Lead AQCD
The 1986 Lead AQCD noted that effects of Pb on liver function in humans had not been
extensively studied. Possible association between Pb exposures (blood Pb concentrations
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>70 |ig/dL) and nonspecific liver injury (i.e., increases in liver enzymes in serum) were noted
based on studies of workers (e.g., Cooper et al., 1973; Hammond et al., 1980). Also noted was
evidence for possible association of suppression of hepatic cytochrome P450 activity with high
blood Pb concentrations (>70 |ig/dL) (Meredith et al., 1977).
Few studies of hepatic effects of Pb on humans have been reported since the 1986 Lead
AQCD. Studies of hepatic enzyme levels in serum suggest that liver injury may be present in Pb
workers; however, associations specifically with Pb exposures are not evident (Al-Neamy et al.,
2001; Hsiao et al., 2001). Studies of urinary metabolites of cytochrome P450 phenotypes
CYP2A6 and CYP3 A4 suggest possible associations between Pb exposure and suppression of
hepatic enzyme activity. The effect on CYP2A6 activity was observed in children with high Pb
burdens (i.e., blood Pb concentration >40 |ig/dL, EDTA-provoked urinary Pb >500 |ig/dL).
The effect on CYP3A4 was observed in association with blood Pb ranges of-30-112 |ig/dL
(based on reported serum Pb concentrations). These studies are summarized in Annex
Table AX6-9.7 and the most pertinent findings are discussed below.

6.9.4.2 Nonspecific Hepatic Injury
Possible association between occupational Pb exposure and liver injury has been assessed
from measurements of serum enzymes (Al-Neamy et al., 2001; Hsiao et al., 2001). Al-Neamy
et al. (2001) found significantly higher serum activity of alkaline phosphatase (AP) and lactate
dehydrogenase (LDH), both within clinically normal ranges, in a group (n = 100) of male Pb
workers (e.g., gas pump attendants, garage workers, printing workers, construction workers),
compared to an age-matched reference group (n = 100). Serum levels of alanine
aminotransferase (ALT), aspartate aminotransferase (AST), and y-glutamyl transferase (y-GT)
were not different in the two groups. The mean blood Pb concentrations were 78 |ig/dL (SD 43)
in the Pb workers and 20 |ig/dL (SD 12) in the reference group. Hsiao et al. (2001) found no
association between blood Pb concentration and ALT activity, in a longitudinal study of a group
of battery manufactory workers (n = 30). Mean blood Pb concentrations ranged from 60 |ig/dL
(-range 25-100) at the start of the study (1989) to 30 |ig/dL (-range 10-60) in the final year of
the study (1999).
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6.9.4.3 Hepatic Cytochrome P450 Function
Studies conducted in animals have shown that Pb can decrease the activity of hepatic
cytochrome P450 and its induction by various inducing agent, through a mechanism that, at least
in part, involves a disruption of heme synthesis (see Section 5.10.1.1). Possible associations
between Pb exposure and cytochrome P450 activity have been studied in children and adults
(Saenger et al., 1984; Satarug et al., 2004). Although direct assay of hepatic cytochrome P450
levels is not feasible in epidemiological studies, changes in activities of P450 isozymes can be
detected from measurements of urinary metabolites of P450 substrates. Urinary excretion of
6-p-hydroxycortisol (6-p-OH-cortisol) derives primarily from oxidation of cortisol through the
hepatic cytochrome P450 phenotype CYP3A4. A lower urinary 6-p-OH-cortisol:cortisol ratio is
indicative of possible suppression of hepatic CYP3A4 activity. Saenger et al. (1984) found
significantly lower (-45% lower) urinary excretion of 6-p-OH-cortisol and lower urinary
6-p-OH-cortisol:cortisol ratio in 2 to 9 year-old children (n = 26) who qualified for chelation
(EDTA-provoked urinary Pb >500 |ig/24 h) than in children who did not qualify, and
significantly lower than in an age-matched reference group. Urinary 6-p-OH-cortisol:cortisol
ratio was significantly correlated with blood Pb (r = -0.514, p < 0.001), urinary Pb, and EDTA-
provoked urinary Pb (r = -0.593, p < 0.001). Mean blood Pb concentrations were 46 |ig/dL
(range 33-60), prior to chelation, and 42 |ig/dL (range 32-60) in the children who did not qualify
for chelation.
Satarug et al. (2004) measured urinary excretion of 7-hydroxy-coumarin (7-OH-
coumarin) following a single oral dose of coumarin to assess effects of Cd and Pb exposure on
cytochrome P450 phenotype CYP2A6. The rationale for this approach is that 7-hydroxylation of
coumarin occurs solely through the CYP2A6 pathway. Coumarin-induced urinary 7-OH-
coumarin was measured in a group (n = 118) selected from the general population in Bangkok,
Thailand. All subjects were nonsmokers. The study found a significant association between
increasing urinary Pb and decreasing covariate-adjusted urinary 7-OH-coumarin in males, but
not in females. Covariates retained included age and zinc excretion. A significant association, in
opposite direction, was found between urinary Cd and urinary 7-OH-coumarin. Mean urinary Pb
levels (blood Pb concentrations were not reported) were 1.3 jig/g creatinine (range 0.1-12) in
males, and 2.4 |ig/g creatinine (range 0.6-6.8) in females. Mean serum Pb concentrations were
4.2 |ig/L (range 1-28) in males and 3.0 |ig/L (range 1-12) in females. The range of 1 to 28 |ig/L
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serum would correspond to a blood Pb concentration range of-30 to 112 |ig/dL (U.S.
Environmental Protection Agency, 2003). These results are consistent with observations of
depressed excretion of metabolites of the CYP2A6 substrate, phenazone, in association with
overt clinical Pb toxicity in Pb workers (Fischbein et al., 1977; Meredith et al., 1977).

6.9.5 Effects of Lead on the Gastrointestinal System
6.9.5.1 Summary of Key Findings on the Effects of Lead on the Gastrointestinal
System from the 1986 Lead AQCD
The 1986 Lead AQCD described gastrointestinal colic (abdominal pain, constipation,
intestinal paralysis) as a consistent early symptom of Pb poisoning in humans and noted that
such GI symptoms may be present in association with blood Pb concentrations in the range of
30-80 |ig/dL. The 1986 Lead AQCD concluded that information was insufficient to establish
clear concentration (i.e., blood concentration)-response relationships in the general population in
association with environmental exposure. Subsequent to the 1986 AQCD several studies of
prevalence of symptoms of GI colic in Pb workers have been reported that provide evidence for
symptoms in association with blood Pb levels >50-80 |ig/dL (Awad el Karim et al., 1986;
Holness and Nethercott, 1988; Lee et al., 2000; Matte et al., 1989). These studies are
summarized in Annex Table AX6-9.8. Similar types of studies of children have not been
reported.

6.9.5.2 Gastrointestinal Colic
Lee et al. (2000) collected data on symptoms (self-reported questionnaire) in male Pb
workers (n = 95) who worked in secondary smelters, PVC-stabilizer manufacture facilities, or
battery manufacture facilities. A logistic regression model was applied to prevalence data for GI
symptoms (loss of appetite, constipation or diarrhea, abdominal pain). The covariate-adjusted
odds ratio for symptoms, in association with blood Pb concentration (> versus < the group
median, 45.7 |ig/dL), was not significant (1.8, [95% CI: 0.7, 4.5]). The corresponding odds ratio
for DMSA-provoked urinary Pb (>versus <260.5 |ig/4 h, the group median) was also not
significant (1.1, [95% CI: 0.4, 2.5]). However, the odds ratio for neuromuscular symptoms in
association with DMSA-provoked urinary Pb was significant (7.8, [95% CI: 2.8, 24.5]),
suggesting that neuromuscular symptoms may occur in association with exposures insufficient to
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result in detectable GI symptoms. Covariates retained in the final regression models were age,
tobacco smoking, and alcohol consumption.
Three other studies have attempted to quantify associations between Pb exposure and GI
symptoms in Pb workers (Awad el Karim et al., 1986; Holness and Nethercott, 1988; Matte
et al., 1989). Holness and Nethercott (1988) found a significantly (p < 0.05) higher prevalence
of symptoms in a group of demolition workers (n = 119) in association with a blood Pb range of
50 to 70 |ig/dL (n = 87), 37% for abdominal cramps and 42% for constipation, or >70 |ig/dL
(n = 19) 77% for abdominal cramps and 62% for constipation compared to a group of workers in
which the blood Pb values were <50 |ig/dL (n = 13), prevalences of 8% and 6%. Awad el Karim
et al. (1986) found higher prevalence of GI symptoms, for abdominal colic and constipation,
respectively, in male battery manufacture workers, 41.3% for abdominal colic and 41.4% for
constipation, compared to a reference group of workers, n = 40 prevalences of 7.5% and 10% for
abdominal colic and constipation, respectively. The blood Pb ranges were 55 to 81 |ig/dL in the
Pb workers and 7 to 33 |ig/dL in the reference group. Matte et al. (1989) did not find a
significant difference in prevalence of GI symptoms (decreased appetite, nausea, abdominal
pain) among a group of battery manufacture and repair workers (n = 63) when stratified by blood
Pb concentration (<60 |ig/dL, >60 jig/dL). The prevalence ratio (high/low blood Pb strata) for
abdominal pain was 1.5 (95% CI: 0.5, 4.6).
In a small study of environmentally-exposed adults, Bercovitz and Laufer (1991) found
that Pb levels in the dentine of patients with GI ulcers (n = 11), even long after recovery, were
significantly higher (mean Pb 75.02 |ig/g [SE 8.15]) than those in healthy subjects (mean Pb
25.62 |ig/g [SE 10.15]). Ten of the 11 peptic ulcer patients had a higher Pb level than the
healthy subjects. In these 10 patients, increased severity of the ulcer and longevity of suffering
was associated with increased tooth Pb levels. The authors suggested that increased Pb
absorption was associated with damage to the epithelial mucosal cells of the GI tract.

6.9.6 Effects of Lead on Bone and Teeth
6.9.6.1 Summary of Key Findings of the Effects of Lead on Bone and Teeth from
the 1986 Lead AQCD
The 1986 Lead AQCD did not discuss the effects of Pb on bone and teeth. Since the 1986
AQCD, an additional development in Pb epidemiology has been studies which explored possible
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associations between Pb exposure and risk of dental caries (Campbell et al., 2000; Dye et al.,
2002; Gemmel et al., 2002; Moss et al., 1999). Also, a limited number of studies also examined
the toxic effect of Pb on bone. These studies are summarized in Annex Table AX6-9.9.

6.9.6.2 Bone Toxicity
The number of papers dealing with direct toxicity of Pb on bone is limited. Most papers
are reviews (Hu et al., 1991; Puzas, 2000; Puzas et al., 1992; Rabinowitz, 1991; Silbergeld,
1991; Silbergeld et al., 1993; Vig and Hu, 2000) or are based on cellular studies (e.g., Pounds
et al., 1991) or laboratory animal evaluations.
Various authors have suggested that Pb is a potential risk factor for osteoporosis because
of the pivotal role of the skeleton in Pb toxicokinetics (Goyer et al., 1994). Bone cells
accumulate Pb actively and earlier ideas suggested that Pb was incorporated into the mineral
matrix of the bone (Wittmers et al., 1988). However, in an in vivo iliac bone biopsy using laser
microbeam mass analysis on a Pb-intoxicated adult female following chelation therapy, Flood
et al. (1988) found the extracellular Pb was concentrated in the superficial 3 to 6 jim of the
osteoid zone of bony trabeculae. Because Pb was absent from the deeper parts of the
mineralized matrix, the authors suggested that Pb binds more strongly to the organic matrix than
to bone mineral.
There is increasing evidence from cell culture experiments, animal studies, and from
measurements in humans that Pb may exert detrimental effects on bone mineral metabolism.
In humans, this evidence comes from several studies. Following on from the earlier observations
of Rosen et al. (1980) that 1,25 (OH)2 vitamin D levels are reduced in Pb-poisoned children,
Markowitz et al. (1988) found that osteocalcin levels were inversely related to Pb body burden in
moderately Pb-poisoned children. During chelation treatment for Pb, the osteocalcin levels were
shown to increase.
An inverse relationship between blood Pb and stature and chest circumference was
observed in children from the NHANES II study (Schwartz et al., 1986). There are several
explanations for the inverse correlation between blood Pb and growth in children. First, blood
Pb level may be a composite factor reflecting other genetic, ethnic, nutritional, environmental,
and sociocultural factors. Second, nutritional deficits that retard growth also enhance Pb
absorption. Finally, there may be a direct effect of low level Pb on growth in children. This
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condition was explained by Dowd et al. (1994) as resulting from the inhibition by Pb2+ of
binding of osteocalcin to hydroxyapatite. Effects similar to those described by Schwartz et al.
(1986) were reported by Angle and Kuntzelman (1989), Lauwers et al. (1986), and Shukla et al.
(1989).
Puzas et al. (1992) suggested Pb could upset the very sensitive interactive metabolic
activity of osteoblasts and chondrocytes and thereby affect bone growth. In a later review, Puzas
(2000) enlarged upon his earlier paper and described in more detail potential mechanisms of Pb
effects on growth plate cartilage metabolism and on osteoclasts and osteoblasts, especially
associated with osteoporosis.
Observational studies by Spencer et al. (1992, 1994) suggested a link between
occupational Pb exposure and Paget's disease in both males and females, but the authors
declined to advocate a causal effect. Later Spencer et al. (1995) found that 92% of a group of
48 patients with Paget's disease were exposed to Pb either from occupational or environmental
sources. Adachi et al. (1998) explored a possible association between Pb and bone disease from
XRF analyses of cortical and trabecular bone Pb content in 117 patients who attended a
metabolic bone disease clinic (n = 92) or were undergoing dialysis for renal failure (n = 25).
In patients suffering from Paget's disease, cortical bone Pb content was higher than it was in
controls, patients with osteoporosis, and patients on dialysis. Trabecular bone Pb content was
lowest in patients with Paget's disease or osteitis fibrosa. However, the authors could not
distinguish between two alternatives, the first being that increased bone turnover due to Paget's
disease releases Pb from trabecular bone that is then available for deposition into cortical bone,
or secondly, that increased Pb content in cortical bone may cause increased turnover with release
of Pb from trabecular bone.
In another facet of the Normative Aging Study, Shadick et al. (2000) investigated possible
associations between long-term Pb accumulation and hyperuricemia and gouty arthritis in
777 male subjects. They found a positive association between patella bone Pb and uric acid
levels (p = 0.022) but no association between bone or blood Pb and gout in this environmentally-
exposed group.
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6.9.6.3 Dental Health
Caries is considered an infectious disease arising from a multifactorial process involving
particular flora, dietary exposures, and a susceptible host (Schafer and Adair, 2000). Increased
caries risk has been detected in association with increasing blood Pb levels in populations with
mean blood Pb concentrations of-2-3 |ig/dL (Dye et al., 2002; Gemmel et al., 2002; Moss et al.,
1999).
Several studies have examined relationships between Pb exposure and the occurrence of
dental caries in children and adults. The two largest studies were analyses of data collected in
the NHANES III; both found significant associations between increasing caries prevalence and
increasing blood Pb in children/adolescent (Moss et al., 1999) and adult (Dye et al., 2002)
populations with geometric mean blood Pb levels of -2.5 |ig/dL. In the Moss et al. (1999) study,
the odds ratios for caries in association with a 5 |ig/dL increase in blood Pb concentration (i.e.,
from <2 |ig/dL) was 1.8 (95% CI: 1.3, 2.5). Outcomes of two smaller studies were mixed, with
one study finding no significant association between blood Pb and caries prevalence (Campbell
et al., 2000) and the other finding significant associations (Gemmel et al., 2002); the latter being
for children having a mean blood Pb concentration of 2.9 |ig/dL (maximum 13).
The Moss et al. (1999) NHANES III analysis included results from coronal caries
examinations on 24,901 subjects, stratified by age: 2 to 5 years (n = 3,547), 6 to 11 years
(n = 2,894), and > 12 years (n = 18,460). Specific outcomes assessed varied by age group:
for children 2 to 11 years old who had at least one deciduous tooth, the number of deciduous
teeth displaying decayed or filled surfaces (DPS); for subjects >6 years and who had at least one
permanent tooth, the number of permanent teeth displaying decayed or filled surfaces; and for
subjects > 12 years, the sum of decayed, missing, and filled surfaces on permanent teeth (DMFS).
In a multivariate linear regression model, increasing blood Pb concentration (log-transformed)
was significantly associated with covariate-adjusted increases in dfs in the 2 to 5 year age group
(P = 1.78 [SE 0.59], p = 0.004) and in the 6-11 year age group (P = 1.42 [SE 0.51], p = 0.007).
Log-transformed blood Pb also was associated with increases in DFS in the 6-11 years age
group (P = 0.48 [SE 0.22], p = 0.03) and in the > 12 years age group (P = 2.50 [SE 0.69],
p < 0.001), and increases in DMFS in the > 12 years age group (P = 5.48 [SE 1.44], p = 0.01).
The odds ratios (compared to 1st tertile, < 1.66 |ig/dL) for the binomial outcome, 0 or > 1 DMFS,
were 1.36 (95% CI: 1.01, 2.83) for the blood Pb range 1.66-3.52 |ig/dL, and 1.66 (95% CI:
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1.12, 2.48) for the range >3.52 |ig/dL. Corresponding population risks attributable to blood Pb
concentration were 9.6% and 13.5% in the blood Pb strata, respectively. An increase in blood Pb
of 5 |ig/dL was associated with an odds ratio of 1.8 (95% CI: 1.3, 2.5). Covariates included in
the models were age, gender, race/ethnicity, poverty income ratio, exposure to cigarette smoke,
geographic region, educational level of head of household, carbohydrate and calcium intakes,
and frequency of dental visits.
Gemmel et al. (2002) conducted a cross-sectional study of associations between blood Pb
concentration and dental caries in children, 6-10 years of age (n = 543), who resided either in an
urban (n = 290) or rural (n = 253) setting. Mean blood Pb concentrations were 2.9 |ig/dL
(SD 2.0, maximum 13 |ig/dL) in the urban group and 1.7 |ig/dL (SD 1.0, maximum 7 |ig/dL) in
the rural group. Increasing blood Pb concentration (In-transformed) was significantly associated
with covariate-adjusted number of caries (dfs + DPS) (In-transformed) in the urban group
(P = 0.22 [SE 0.08], p = 0.005), but not in the rural group (P = -0.15 [SE 0.09], p = 0.09). When
dfs counts were stratified by permanent or deciduous teeth, the blood Pb association in the urban
group was significant for deciduous teeth (P = 0.28 [SE 0.09], p = 0.002), but not for permanent
teeth (P = 0.02 [SE 0.07], p = 0.8). Covariates retained in the linear regression model were age,
sex, ethnicity, family income, education of female guardian, maternal smoking, frequency of
tooth brushing, firmness of toothbrush bristles, and frequency of chewing gum.
Campbell et al. (2000) was a retrospective cohort study in which dfs were assessed in
children 7 to 12 years of age (n = 248) from Rochester, NY. Mean blood Pb concentration,
measured at ages 18 and 37 months of age, was 10.7 |ig/dL (range 18.0-36.8). The covariate-
adjusted odds ratios for caries associated with a blood Pb concentration >10 |ig/dL compared to
< 10 |ig/dL were 0.95 |ig/dL (95% CI: 0.43, 2.09) for permanent teeth and 1.77 |ig/dL (95% CI:
0.97, 3.24) for deciduous teeth. Covariates retained in the logistic model were age, grade in
school, number of tooth surfaces at risk. Other covariates examined in the models, all of which
had no significant effect on the outcome, were gender, race/ethnicity, SES, parental education,
residence in community supplied with fluoridated drinking water, and various dental hygiene
variables. This study did not demonstrate that Pb exposure >10 |ig/dL as a toddler was a strong
predictor of caries among school-age children, but the authors noted that this might be due to
limited statistical power.
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Dye et al. (2002) analyzed data collected in NHANES III on indices of periodontal bone
loss. The analysis was confined to subjects 20 to 69 years of age (n = 10,033). The geometric
mean blood Pb concentration of the study group was 2.5 |ig/dL (SE 0.08), with 2.4% of the
group having blood Pb levels >10 |ig/dL. Increasing log-transformed blood Pb was significantly
associated with increasing prevalence of covariate-adjusted dental furcation (P = 0.13 [SE 0.05],
p = 0.005). Dental furcation is indicative of severe periodontal disease. Covariates retained in
the linear regression model were age, sex, race/ethnicity, education, smoking, and age of home.
Smoking status was a significant interaction term when included in the model (P = 0.10
[SE 0.05], p = 0.034). When stratified by smoking status, the association between dental
furcation and blood Pb concentration was significant for current smokers (P = 0.21 [SE 0.07],
p = 0.004) and former smokers (P = 0.17 [SE 0.07], p = 0.015), but not for nonsmokers
(P = -0.02 [SE 0.07], p = 0.747).
Some studies examined the relationship between tooth Pb levels and dental caries.
In their compilation of metal concentrations in 1,200 deciduous teeth from a Norwegian
population, Tvinnereim et al. (2000) found that carious teeth had higher Pb concentrations than
noncarious teeth. Gil et al. (1994) measured Pb concentrations from 220 whole deciduous and
permanent teeth from Coruna, Spain. The geometric mean Pb level was 10.36 jig/g of tooth.
There was a significant increase in teeth Pb levels with advancing age. Permanent teeth showed
higher mean Pb values (13.09 |ig/g [SEM 1.07]) than deciduous teeth (3.96 |ig/g [SEM 1.07]).
The authors reported a possible relationship between increased Pb content and periodontal
pathology but did not observe any relationship between Pb concentrations and caries.

6.9.7 Effects of Lead on Ocular Health
6.9.7.1 Summary of Key Findings of the Effects of Lead on Ocular Health from the
1986 Lead AQCD
The 1986 Lead AQCD did not address Pb effects on ocular health in humans. Various
disturbances of the visual system have been observed in association with overt clinical Pb
poisoning, including retinal stippling and edema, cataracts, ocular muscle paralysis, and impaired
vision (see Otto and Fox, 1993 for review). Two longitudinal studies completed since 1986
provide evidence for (a) possible associations between Pb exposure and visual evoked retinal
responses in children of mothers whose blood Pb levels in mid-pregnancy were in the 10 to
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32 |ig/dL range (Rothenberg et al., 2002b); and (b) a possible association between Pb exposure
and risk of cataracts in males whose tibia bone Pb levels were in the 31 to 126 jig/g range
(Schaumberg et al., 2004). These studies are summarized in Annex Table AX6-9.10.

6.9.7.2 Ocular Effects
In the Mexico City prospective Pb study, Rothenberg et al. (2002b) measured
flash-evoked electroretinograms (ERG) in a subset of the study group (n = 45) at ages
7-10 years. As part of the prospective study, blood Pb concentrations had been measured during
pregnancy and in the children, at birth and every 6 months, thereafter. Increasing maternal blood
Pb, measured at 12 weeks of gestation, was significantly associated with increasing ERG a-wave
and b-wave amplitude, with significant increases in a-wave in the second maternal blood Pb
tertile (range 6.0-10.0 jig/dL), and a-wave and b-wave in the third maternal blood Pb tertile
(range 10.5-32.5 |ig/dL), compared to the first blood Pb tertile (range 2.0-5.5 |ig/dL). No other
blood Pb measurements were significantly associated with any ERG outcomes.
As part of the longitudinal Normative Aging Study, Schaumberg et al. (2004) analyzed
prevalence of cataracts in adult males (n = 642), mean age 69 years (range 60-93 years).
Subjects were stratified by blood Pb, patella bone Pb, or tibia bone Pb quintiles for a logistic
regression analysis of the odds ratios for cataracts (first quintile as reference). Covariate
adjusted odds ratio for cataracts in the fifth tibia bone Pb quintile was significant (3.19 [95% CI:
1.48, .90]). Odds ratios for cataracts were not significantly associated with patella bone Pb
(1.88 [95% CI: 0.88, 4.02]) or blood Pb (0.89 [95% CI: 0.46, 1.72]). The first and fifth quintile
Pb levels were 0-11 jig/g and 31-126 jig/g for tibia bone; 1-16 jig/g and 43-165 jig/g for patella
bone; and 1.0-3.0 jig/g and 8-35 |ig/dL for blood. Covariates retained in the regression model
were age, smoking, history of diabetes; and daily intake of vitamin C, vitamin E, and
carotenoids.
Cavalleri et al. (1982) measured visual fields of male workers exposed to Pb stearate in a
polyvinyl pipe manufacturing facility (n = 35). Workers in a reference group (n = 350) were
individually matched for age, smoking, and alcohol consumption. Visual sensitivity was
significantly lower in Pb workers compared to the reference group; however, visual sensitivity
index was not significantly associated with blood or urine Pb. Prevalence of mesopic scotoma
(retinal light insensitivity under low illumination conditions) was 28.5% in the Pb workers and
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0% in the reference group. Mean blood Pb levels were 46 |ig/dL (range 21-82) in the Pb

workers and 30 |ig/dL (range 21-42) in the reference group.

6.9.8 Summary of the Epidemiologic Evidence for the Effects of Lead
on Other Organ Systems

The following are a listing of key findings discussed above for Pb effects on other organ

systems.

• Biochemical Effects of Lead. Evidence for disruption of heme synthesis derives from
numerous studies in which Pb exposure has been associated with decreased activities of
enzymes in the heme synthesis pathway (i.e., ALAS, ferrochelatase, cytochrome P450)
and increased levels of substrates for heme synthesis (i.e., ALA, coproporphyrin,
erythrocyte protoporphyrin) in both children and adults. Quantitative relationships
between blood Pb concentration and the above biomarkers of impaired heme synthesis
are highly consistent across studies. Increases in blood Pb concentration of -20-
30 |ig/dL are sufficient to halve erythrocyte ALAD activity and sufficiently inhibit
ferrochelatase to double erythrocyte protoporphyrin levels.

• Blood Lipids. Associations between occupational exposure to Pb and changes in blood
lipid composition have been observed. These include increased levels of lipid peroxides
in blood and/or serum, and increased serum levels of total and HDL cholesterol. Effects
on serum cholesterol levels were evident in association with a blood Pb concentration in
the range of 5-62 |ig/dL (-mean 14 jig/dL). Oxidative changes in blood lipids (e.g.,
increased levels of lipid peroxides and malondialdehyde levels) as well as decreased
levels of erythrocyte superoxide dismutase, catalase, G6PD, and GSH peroxidase; and
increased lymphocyte reactive oxygen species and depleted GSH levels, indicative of
increased oxidative stress, have been observed in Pb workers to be associated with blood
Pb concentrations >30 |ig/dL.

• Disruption of Hemoglobin Synthesis and Declines in Erythrocyte Numbers. Exposures
that result in blood Pb concentrations <40 |ig/dL appear to be tolerated without a decline
in blood hemoglobin levels or hematocrit. However, perturbation of erythropoiesis,
indicated by changes in serum erythropoietin and progenitor cells, occurs in association
with blood Pb concentrations <40 |ig/dL and in the absence of detectable changes in
blood hemoglobin levels or hematocrit in children and adults. Risk of clinical anemia in
children becomes appreciable at much higher blood Pb levels; a 10% decrease in
hematocrit has been estimated to occur in association with blood Pb concentrations
>85 |ig/dL; a 10% probability of anemia (hematocrit <35%) was estimated to be
associated with a blood Pb concentration of-20 |ig/dL at age 1 year, 50 |ig/dL at age
3 years, and 75 |ig/dL at age 5 years. In adults, with blood Pb levels below 25 |ig/dL,
increasing patella bone Pb, but not blood Pb, was associated with a significant decrease in
hematocrit.
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• Effects on the Endocrine System. Several studies have examined possible associations
between Pb exposures in children and adults and various biomarkers of endocrine
function, including the thyroid, male reproductive, and calcitropic endocrine systems.
The strongest study designs have yielded no associations, or weak associations, between
Pb exposure and thyroid hormone status. Studies of occupational exposures which
included subjects having blood Pb concentrations >100 |ig/dL have found depression of
serum T3 and/or T4 levels, without a detectable increase in serum TSH; however, studies
in which the blood Pb distribution was dominated by levels well below 100 |ig/dL, have
found either no effects or subclinical increases in serum T3, T4, with no change in TSH
levels.

• Reproductive Endocrine Function. Studies of the male reproductive system that
attempted to control for confounding effects of age have yielded mixed outcomes. Blood
Pb ranges in these studies were similar (4-90 |ig/dL), yet outcomes were mixed, with no
change, or subclinical decrease in serum testosterone (TES) in association with Pb
exposure. There are also mixed effects on serum follicle stimulating hormone (FSH) and
luteinizing hormone (LH) with increases, decreases, and no change in hormone levels
observed. The inconsistency in the direction of effects on TES and the two androgen-
regulating pituitary hormones, FSH and LH, is particularly noteworthy, in the absence of
evidence for effects of Pb exposure on GnRH-induced FSH (Erfurth et al., 2001).

• Calcitropic Endocrine Function. Children exposed to relatively a high level of Pb
(>30 |ig/dL) may exhibit depressed levels of circulating 1,25-(OH)2D3. However,
associations between serum vitamin D status and blood Pb may not be present in
calcium-replete children who have average lifetime blood Pb concentrations below
25 |ig/dL. In adults, exposures to Pb that result in blood Pb concentrations >40-60 |ig/dL
may increase, rather than decrease, circulating levels of 1,25-(OH)2D3 and PTH.

• Effects on the Hepatic System. Few studies of hepatic effects of Pb on humans have
been reported since the 1986 Lead AQCD. Studies of hepatic enzyme levels in serum
suggest that liver injury may be present in Pb workers; however, associations specifically
with Pb exposures are not evident. Studies of urinary metabolites of cytochrome P450
phenotypes CYP2A6 and CYP3 A4 suggest possible associations between Pb exposure
and suppression of hepatic enzyme activity. The effect on CYP2A6 activity was
observed in children with high Pb burdens (i.e., blood Pb >40 |ig/dL, EDTA-provoked
urinary Pb >500 |ig/dL). The effect on CYP3A4 was observed in association with blood
Pb ranges of-30-112 |ig/dL (based on reported serum Pb concentrations).

• Effects on the Gastrointestinal System. Several studies of prevalence of symptoms of GI
colic in Pb workers provide evidence for symptoms in association with blood Pb levels
>50-80 |ig/dL. Similar types of studies of children have not been reported.

• Effect on Bone and Teeth. There is limited, but suggestive evidence of an association
between Pb exposure and bone toxicity. However, in most studies, it is difficult to assess
the direct contribution of Pb on bone diseases or reduced growth. Several studies that
have explored possible associations between Pb exposure and risk of dental caries.
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Increased caries risk has been detected in association with increasing blood Pb
concentrations in populations with mean blood Pb concentrations of-2-3 |ig/dL.
• Ocular Health. Various disturbances of the visual system have been observed in
association with overt clinical Pb poisoning, including retinal stippling and edema,
cataracts, ocular muscle paralysis, and impaired vision. Two longitudinal studies
completed since the 1986 Lead AQCD provide evidence for possible associations
(a) between Pb exposure and visual evoked retinal responses in children of mothers with
blood Pb concentrations in mid-pregnancy of 10.5-32.5 |ig/dL and (b) between Pb
exposure and risk of cataracts in middle-aged males whose tibia bone Pb levels were
31-126|ig/g.
6.10 EPIDEMIOLOGIC CONSIDERATIONS AND SUMMARY OF
EVIDENCE FOR LEAD HEALTH EFFECTS
6.10.1 Introduction
A remarkable expansion has occurred since the 1990 Lead Supplement in the extent of
the database available for drawing inferences about the various expressions of Pb toxicity.
Moreover, the nature of the evidence available has changed as well. Many of the studies
conducted prior to 1990 focused on the issue of whether an observed observation was likely
to be real or the result of chance, selection bias, residual confounding, or some other
methodological error. The validity of any association still needs to be ensured. The studies
since 1990 mainly focus on characteristics of the pertinent concentration-response relationships
(including the functional forms of the relationships), the slopes of the relationships, the natural
histories of adverse effects, and the confounding or effect modifying influences of various co-
exposures and host characteristics. Discussed below are pertinent issues that need to be
considered in the evaluation and interpretation of the epidemiologic evidence regarding Pb health
effects. Measurement error in the exposure and outcome variables are first discussed, followed
by sections on potential confounding of Pb health effects and inferences of causality. Additional
issues, including the concentration-response relationship and the persistence of Pb health effects
are discussed in Chapter 8.
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6.10.2 Exposure and Outcome Assessment in Lead Epidemiologic Studies
6.10.2.1 Assessment of Lead Exposure and Body Burdens Using Biomarkers
Blood and bone Pb levels serve as valuable indicators of Pb exposure in epidemiologic
studies. Having a biomarker as a measure of pollutant exposure provides an important advantage
beyond air lead measurements in estimating health effects. The expanded discussion below of
the limitations in using biomarkers as indicators of exposure allows further understanding of the
uncertainties potentially involved.
For any health endpoint of interest, the most useful biomarker of exposure is one that
provides information about the Pb dose at the critical target organ and, moreover, reflects the
exposure averaging time that is appropriate to the underlying pathogenetic processes
(e.g., cumulative over lifetime, cumulative over a circumscribed age range, concurrent).
In recent studies of Pb and health, the exposure biomarkers most frequently used are Pb in blood
and bone (see discussion in Chapter 4). For outcomes other than those relating to hematopoiesis
and bone health, these biomarkers provide information about Pb dose that is some distance from
the target organ. For example, given that the central nervous system is considered the critical
target organ for childhood Pb toxicity, it would be most helpful to be able to measure, in vivo,
the Pb concentrations at the cellular site(s) of action in the brain. However, because such
measurements are not currently feasible, investigators must rely on measurements of Pb in the
more readily accessible but peripheral tissues. The relationship between brain Pb and Pb in each
of these surrogate tissues is still poorly understood, although the pharmacokinetics clearly differs
among these compartments. In both rodents and nonhuman primates, brain Pb level falls much
more slowly than blood Pb level following chelation with succimer and, in the rodent, in
nonchelated animals after cessation of exposure. These observations suggest that using blood Pb
as an index of Pb in the brain will result in exposure misclassification, although the magnitude of
this bias in any specific setting will be difficult to characterize. The most likely direction,
however, would be underestimation of the amount of Pb in the brain, at least under scenarios
involving chronic exposure.
As an exposure biomarker, blood Pb level has other limitations. Only about 5% of an
individual's total body Pb burden resides in blood. Furthermore, blood consists of several
subcompartments. More than 90% of Pb in whole blood is bound to red cell proteins such as
hemoglobin, with the balance in plasma. From a toxicological perspective, the unbound fraction
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is likely to be the most important subcompartment of blood Pb because of the ease with which it
diffuses into soft tissues. The concentration of Pb in plasma is much lower than in whole blood,
however. For example, in a group of pregnant women with blood Pb levels below 10 |ig/dL,
plasma Pb levels were less than 0.3% of the whole blood Pb level. The greater relative
abundance of Pb in whole blood makes its measurement much easier (and more affordable) than
the measurement of Pb in plasma. The use of whole blood Pb as a surrogate for plasma Pb could
be justified if the ratio of whole blood Pb to plasma Pb were well characterized, but this is not so.
At least some studies suggest that it varies several-fold among individuals with the same blood
Pb level. Moreover, the ability of red cells to bind Pb is limited, so the ratio of blood Pb to
plasma Pb would be expected to be nonlinear. Thus, interpreting whole blood Pb level as a
proxy for plasma Pb level, which, itself, is a proxy for brain Pb level, will result in some
exposure misclassification.
Although the use of blood Pb may not best reflect the actual dose of Pb in the specific
target organs of interest, of greater concern are the epidemiologic implications of its use. In a
regression model, the variation in Pb about its mean is correlated with the variation in outcome
about its mean. Because only variations about the mean contribute to the association, mean
differences between true and estimated levels become irrelevant. The measurement error in
considering blood Pb as a surrogate for brain Pb will bias the blood Pb effect towards the null
and is an example of classical measurement error. The error will be nonlinear if red cell binding
is limited. However, when interest centers on the low blood Pb level, the error should be
approximately additive, multiplicative, or both. Another example of measurement error with
epidemiologic implications is the Berksonian error. Berkson error arises when averages of blood
Pb levels are used in a regression rather than individual data. For example, if a regression of IQ
is performed on averages of children's blood Pb concentrations grouped between intervals, the
Berkson error model will apply. The slope will be unbiased, but the standard errors will be
inflated.
There are additional issues to consider in the use of blood as a marker of Pb exposure.
The residence time of Pb in blood is closely linked to red cell lifetime, with a half-time on the
order of 30 days. Thus, a high blood Pb level does not necessarily indicate a high body Pb
burden. Similarly, individuals who have the same blood Pb level will not necessarily have
similar body burdens or exposure histories. The rate at which blood Pb level changes with
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time/age depends on exposure history due to re-equilibration of Pb stored in the various body
pools. In nonchelated children, the time for blood Pb to decline to a value less than 10 |ig/dL
was linearly related to baseline blood Pb level. A single blood Pb measurement might therefore
provide limited information about an individual's Pb exposure history, a difficulty frequently
cited with respect to the interpretation of cross-sectional studies of pediatric Pb toxicity, in which
children's blood Pb level is often measured only once, and sometimes only well after the period
when blood Pb levels typically peak (18-30 months of age). If it is exposures to Pb in the early
postnatal years that are most detrimental to children's development, categorizing a child's
exposure status based on a blood Pb level contemporaneous with the measurement of
neurodevelopment at school-age could result in exposure misclassification. This concern must
be qualified, however, by recent data from some longitudinal studies indicating that concurrent
blood Pb level, even at ages well beyond 18 to 30 months, is sometimes the strongest predictor
of late outcomes (Canfield et al., 2003a; Dietrich et al., 1993a,b; Tong et al., 1996; Wasserman
et al., 2000b). Changes in blood Pb concentration in children are found to closely parallel
changes in total Pb body burden. Empirical evidence in support of this comes from longitudinal
studies in which relatively high correlations (r = 0.85) were found between concurrent or lifetime
average blood Pb concentrations and tibia bone Pb concentrations (measured by XRF) in a
sample of children in which average blood Pb concentrations exceeded 20 |ig/dL; the
correlations were much weaker (r = <0.15) among children who had average blood Pb
concentration <10 |ig/dL (Wasserman et al., 1994).
Age-related changes in vulnerability, and the reasons why it might differ across studies,
remain uncertain. It may be that among children with chronically elevated exposure, but not in
those with relatively low lifetime exposure, blood Pb level measured at school-age is a
reasonably good marker of cumulative exposure. That concurrent blood Pb level is, under some
circumstances, a stronger predictor of school-age outcomes than is blood Pb level in the early
postnatal years does not necessarily imply greater vulnerability of the brain to ongoing than to
past exposure. Due to the high intercorrelation among blood Pb measures taken at different time
points, it is not feasible to examine exposures during any given age for evidence of a sensitive
neurodevelopmental period.
The development of X-ray-fluorescence (XRF) methods for measuring Pb in mineralized
tissues offers another approach for characterization and reconstruction of exposure history. Such
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tissues are long-term Pb storage sites, with a half-life measured in decades and contain -90% of
the total body Pb burden in adults and 70% in children. Thus, bone Pb is an index with a long
exposure averaging time. XRF methods have proven useful in studying individuals with
occupational Pb exposure, those living in highly polluted environments, and those for whom
community Pb exposures are or, in the past, were relatively high (e.g., Korrick et al., 1999;
Schwartz et al., 2000a,b,c,d). In a relatively highly exposed cohort of pregnant women in
Mexico City, higher bone Pb levels at one month postpartum were associated with reduced birth
weight, less infant weight gain, smaller head circumference and birth length, and slower infant
development (Gomaa et al., 2002; Gonzalez-Cossio et al., 1997; Hernandez-Avila et al., 2002;
Sanin et al., 2001). Among children living near a large Pb smelter in Yugoslavia, IQ at age
10-12 years was more strongly associated, inversely, with tibia Pb level than with blood Pb level
(Wasserman et al., 2003).
Current XRF methods for measuring bone Pb levels have limitations, however. Temporal
features of exposure history cannot readily be discerned. Some progress has been made toward
this goal by examining the spatial distribution of Pb in teeth in relation to the relative abundance
of stable Pb isotopes, but the specialized technologies needed to carry out these analyses are
unlikely ever to be widely available, and the unpredictability of tooth exfoliation makes this
tissue difficult to collect unless the study design involves contact with (and the cooperation of)
participants at the appropriate ages. Current XRF methods might not be sufficiently sensitive for
studies of the health effects of low-dose community exposures. The bone Pb levels of a large
percentage of subjects might be below the detection limit, e.g., 80% in a case-control study of
bone Pb levels and juvenile delinquency in which the minimum detection limit was 21.5 |ig/g
bone mineral (Needleman et al., 2002). Even among individuals known to have histories of
substantial Pb exposures, such as adolescents and young adults who grew up near the Bunker
Hill smelter in Idaho (McNeill et al., 2000), bone Pb levels tend to be low. Lead appears to be
deposited at sites of most active calcification. In children, this is trabecular bone, in which the
rate of fractional resorption in early childhood is high. Depending on the amount of the child's
ongoing exposure, Pb deposited in bone might not remain there for decades, making bone Pb
level an imprecise index of lifetime Pb exposure. This concern also exists in the use of tooth Pb
to represent cumulative Pb exposure in children. Rabinowitz et al. (1993) observed that a child's
tooth Pb level was more strongly related to blood Pb level around the time of tooth exfoliation
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than to an integrated index of blood Pb level prior to exfoliation. Finally, it is difficult to
compare the performance of different laboratories using XRF methods to measure bone Pb
because of the absence of standard reference materials. Nevertheless, efforts continue to modify
the instrumentation or measurement protocols to reduce the detection limit.
A major research need is the development and validation of biomarkers of critical dose
that, compared to blood Pb or bone Pb, are fewer toxicokinetic steps removed from the sites of
Pb's actions in the brain. One promising front in the effort to deduce the contents of the "black
box" separating external dose and clinical disease is the measurement of processes and products
that potentially mediate the association between them. For example, magnetic resonance
spectroscopy (MRS) has been used in small case series to measure the ratio of N-acetylaspartate
to creatine, which are a marker of neuronal and axonal damage and thus, an early biological
effect rather than a biomarker of exposure. In children, higher Pb exposures are associated with
lower N-acetylaspartate to creatine ratios in the frontal gray matter and, to a lesser extent, in
frontal white matter (Trope et al., 1998, 2001). Similarly, an adult who had higher bone and
blood Pb levels than did his monozygotic twin had both greater neuropsychological deficits and
lower N-acetylaspartate to creatine ratios in the hippocampus, frontal lobe, and midbrain
(Weisskopf et al., 2004b). While much remains uncertain about the interpretation of MRS, the
use of this and other biochemical imaging methods, in combination with more conventional
structural and functional imaging methods, may improve the current understanding of the
mechanisms of Pb neurotoxicity.
Despite these limitations, blood and bone Pb levels both provide relevant and valuable
measures of Pb exposure in epidemiologic studies. Strong and consistent associations have been
observed in the epidemiologic literature between blood and/or bone Pb levels and various health
effects, most notably for neurotoxicity in children and cardiovascular effects in adults.
Evaluation of relative strength of associations of particular health endpoints with blood versus
bone Pb measures has been useful in a number of studies in attributing the relative likelihood of
effects being due to concurrent recent Pb exposures versus past peak or long-term cummulative
Pb exposures. Given the number of toxicokinetic steps separating Pb levels at the critical target
organs from the usual exposure biomarkers, the progress made in characterizing concentration-
response relationships for various Pb-related health outcomes is remarkable.
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6.10.2.2 Assessment of Health Outcomes
Outcome measurement and outcome classification have generally received less attention
from investigators than have exposure measurement and misclassification. The specific
problems are, to some extent, endpoint domain-specific. With regard to neurodevelopmental
toxicities, critical issues are whether the assessment instruments used are psychometrically sound
and appropriate for the study cohort, the data generated will support adequate tests of the study
hypotheses, and whether the instruments have been administered and scored consistently and
correctly. With regard to the cardiovascular toxicities of increased blood pressure/prevalence of
hypertension, the critical issue is whether the blood pressure value recorded for a participant is
an accurate estimate. Multiple measurements of blood pressure are frequently made in a study
but investigators usually have not taken advantage of the collected information to quantify the
amount of error in the measurements. This information can be used to improve the reliability of
the measurements, which would be expected to improve the precision of the associations
estimated. Similarly, aggregating scores to estimate latent variables representing, for instance,
"language skills" or "visual-spatial skills" is an approach that might take advantage of the
overlapping information provided by the multiple tests included in neurobehavioral test batteries,
producing more reliable endpoint variables. This approach, however, has not been widely
applied in Pb studies. Concerns regarding the presence of measurement error in the outcome
variable need to be considered in the context of the exposure of interest. If the measurement
error in outcome is uncorrelated with exposure, it will not induce bias in the estimate of the
effect of Pb. However, it should be noted that the measurement error will lead to reduced power
to detect a significant effect.

6.10.3 Confounding of Lead Health Effects
6.10.3.1 Methods Used to Adjust for Confounding in Epidemiologic Studies of Lead
The possibility that the adverse health effects associated with increased Pb exposure in
epidemiologic studies are, in fact, due to risk factors with which increased Pb exposure is
associated remains the most important impediment to drawing causal inferences. It is important
to note that confounding is not an inherent characteristic of an association between Pb exposure
and a health outcome. Rather it is a bias that arises from the particular setting in which the
association is being investigated, and its source is the patterns of covariance between Pb, the
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outcome, and other determinants of the outcome. Therefore, the extent to which it represents an
interpretational challenge is, to some extent, study-specific. Various approaches have been taken
to reduce the uncertainty this creates. Some investigators have specified the sampling frame or
the eligibility criteria so as to increase the homogeneity of the study participants on factors
known to be strong risk factors for the outcome of interest, thereby reducing both (a) the
correlation between them and Pb and (b) their potential to confound any association observed
between increased Pb exposure and poor outcome. An example is the recruitment of a birth
cohort from a maternity hospital that largely served a relatively affluent catchment area, resulting
in high umbilical cord blood Pb levels being associated with higher, rather than lower, social
class standing (Bellinger et al., 1984). Reducing confounding by means of such design decisions
has the disadvantage that an investigator cannot determine whether the impact of Pb on the
outcome varies depending on the factor whose range of potential values has been restricted.
More frequently, however, investigators have relied on statistical procedures, applied post data
collection, to identify and control for potential confounding. Unlike sample restriction, this
approach preserves the opportunity to explore possible modification of the Pb effect by
cofactors.
Adjustment for confounding has been performed primarily using multiple regression
analyses and data stratification. For multiple regression modeling, stepwise regression has been
frequently used for covariate selection. Stepwise regression has many faults and is often less
acceptable then the use of a few well-chosen covariates. However, the stepwise regression
methodology may be considered to have less bias, as it selects from a class of variables that
represent a wide scientific viewpoint rather than the narrower one of the investigator. One
problem with stepwise regression pointed out by Bellinger (2004) is that the usual adjustment
strategy assumes that all the variance in the response shared by the exposure and the confounder
belongs to the confounder. In some settings, this is likely to be excessively conservative,
because confounders can, to some extent, also be proxies for exposure. This is further discussed
in the next section.
Splitting the data set into smaller data sets (partitioning or stratification) and analyzing
those data sets separately was used in some of the studies examining the relationship between
blood pressure and Pb. This practice also has some advantages and disadvantages. Use of
an advanced statistical method could be helpful to determine how the partitioning should be done
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(Young and Hawkins, 1998), which could reveal relationships that would otherwise not be
possible to detect using usual regression techniques. A disadvantage of partitioning a small data
set is that the smaller sample size may lack sufficient power to detect otherwise detectable
associations and to yield reliable estimates.

6.10.3.2 Effects of Confounding Adjustment on Lead Health Effect Estimates
The ability of the investigator to determine how much of the apparent association between
a Pb biomarker and an outcome reflects residual confounding by a cofactor depends on the
characteristics of the joint distribution of Pb and the cofactor. For example, with respect to
neurodevelopment, important cofactors include maternal IQ, quality of the rearing environment,
maternal smoking, alcohol use, and birth weight, among others. Some of these cofactors are
truly independent predictors and can be adjusted for using multiple regression analyses. Under
some circumstances, however, Pb and the cofactor may be so highly related that one cannot be
confident that their associations with the outcome have been disentangled by the statistical
methods applied. Moreover, the true causal relationships among Pb, the cofactors, and the
outcome might not be sufficiently well understood that the outcome variance shared by Pb and
the cofactors can be characterized appropriately in the analyses.
In studies of Pb and neurodevelopment, the magnitude of the Pb coefficient, reflecting the
decline in test score per unit increase in the Pb biomarker, is substantially reduced, often by half
or more, by adjusting for markers of the social environment. However, as noted above, the
extent of confounding is study-specific, so the impact of adjustment for confounders on the Pb
coefficient is also study-specific. With respect to the Port Pirie study, long and Lu (2001)
observed that adjustment for four factors (i.e., quality of home environment, SES, maternal
intelligence, and parental smoking behavior) reduced the magnitude of the estimated association
between Pb and IQ by 40% and inclusion of additional factors resulted in another 10% reduction.
Similarly, in the pooled analysis by Lanphear et al. (2005) that included seven prospective
studies, the crude coefficient for concurrent Pb and childhood IQ score was -4.66 (95% CI:
-5.72, -3.60), while the coefficient adjusted for study site, quality of the home environment
(HOME score), birth weight, maternal IQ, and maternal education was -2.70 (95% CI: -3.74,
-1.66). When expressed as the percentage of variance accounted for in a health outcome, the
contributions of Pb have been characterized as modest in magnitude. For example, Roller et al.
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(2004) noted that blood Pb typically accounts for 1 to 4% of the variance in child IQ scores,
compared to 40% or more by social and parenting factors. During the 1980s, adjustment for
parental IQ and HOME scores became almost mandatory if the findings of a study of Pb and
children's cognitive outcomes were to be considered credible. Simulation analyses conducted by
Mink et al. (2004) suggested that relatively small differences in confounding variables between
"exposed" and "unexposed" groups could produce spurious differences in cognitive test scores if
unmeasured and unaccounted for in the analysis. As noted by Bellinger (2004), however, the
problem usually is not that such cofactors were unmeasured in a Pb study, but that they were not
measured well.
More important yet is the fact that the conceptual models that frame the interpretation of
the resulting models usually fail to reflect adequately the complexity of the associations among
Pb exposure, the outcome, and the cofactors. Although both HOME score and parental IQ surely
strongly influence child outcomes in ways that are independent of Pb, a case can also be made
that Pb might contribute to the associations. That is, a parent's IQ presumably reflects the
parent's early Pb exposure and, assuming that the physical environments in which a parent and
child grow up are not completely unrelated to one another are likely to provide similar Pb
exposure opportunities. Adjusting for parental IQ in evaluating the association between a child's
Pb exposure and his or her IQ, therefore, will result in an underestimate of the contribution of the
child's Pb exposure to his or her IQ. Similarly, if early Pb exposure alters child behavior, the
transactional model of child development would generate the prediction that the changes will
elicit different behaviors from parents, altering the characteristics of the child-rearing
environment. For instance, increased Pb exposure might result in an infant being more irritable,
less soothable, and the parent less nurturing. In so far as measurement of the quality of the
rearing environment in studies occurs after the children have experienced some Pb exposure, the
hypothesis that Pb is responsible for shaping some aspects of that environment cannot be entirely
dismissed, and control for HOME scores might be excessively conservative.
Other aspects of model building in assessing the association of Pb with health outcomes
also warrant comment. In many studies of Pb and cognitive outcomes in children, investigators
have adjusted for factors such as birth weight or length of gestation that might, themselves,
reflect adverse effects of Pb, i.e., mediating factors that lie between Pb and condition on the
causal pathway. The coefficient estimated for Pb in a model that contained such factors would
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be smaller in magnitude than it would be if terms for such mediating factors had not been
included.
Recognizing imperfections in the ability to measure such factors well, a concern is
expressed that the Pb coefficient could be reduced further, perhaps all the way to the null,
if better, more comprehensive methods of measurement were applied. On the other hand, the
methods used to adjust for such factors may be excessively conservative insofar as they attribute
to a factor all of the outcome variance that it shares with Pb, despite the likelihood that the true
relationships among Pb, social factors, and outcome are unlikely to be as simple as this model
assumes. Some factors might, in part, be markers of Pb exposure opportunities. For example,
both Pb biomarker levels and lower cognitive function in children are associated with lower
social class standing. Social class is a complex construct that conveys information about a
multitude of factors that might influence children's health, including the amount of Pb in
environmental media. Thus, some of the association between lower social class and poorer
health might reflect the effect of higher Pb exposure. If so, routine adjustment of health outcome
for social class in assessing the association between increased Pb exposure and poorer health in
children will fail to distinguish these Pb-related and non-Pb-related components of the
association between social class and health, and, in fact, will assume that all of it is non-Pb-
associated (Bellinger et al., 1989). It is nearly impossible to actually determine if the problem of
overadjustment exists in a particular data set. There are several statistical methods which
attempt to address this problem. These include using partial F tests, ridge regression, path
analysis, and structural equations. None of these methods are completely satisfactory.

6.10.4 Inferences of Causality
Even with more sophisticated and nuanced models, however, any conclusions about the
causal forces generating the results of any observational epidemiologic study are necessarily
uncertain. In the absence of random assignment to exposure group, residual confounding will
always be a possible explanation of an observed association. As in other areas of epidemiology,
a weight-of-evidence approach remains the best option available as a basis for drawing of causal
inferences. If the association between a Pb biomarker and a health outcome of interest is
observed in settings that vary widely in terms of the characteristics of the social environment
including sociodemographic and cultural characteristics, characteristics of the study participants,
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including nutritional status, genetic factors, and lifestyle factors, the likelihood that the
association is attributable, in its entirety, to residual confounding is reduced. For instance, the
pooled analyses of data contributed by many of the international prospective studies provide a
compelling demonstration that the association between blood Pb level and child IQ is remarkably
robust across disparate sociocultural settings (Lanphear et al., 2005). Even such consistency in
the effect estimate across diverse settings is only indirect and weak evidence of causality,
however. In general, epidemiologic studies rarely provide data that enhance understanding of
the "black box" between biomarkers of Pb burden and indicators of health status. Epidemiologic
data identify associations between exposure biomarkers and health indicators, but are not highly
informative regarding possible mechanisms of Pb toxicity that underlie the associations.
A critical stage in applying the overall weight-of-evidence approach is the examination of the
epidemiologic data in the context of data from experimental animal behavioral and mechanistic
studies. Although such data have their own limitations, they are not subject to many of the most
important potential biases that can becloud the interpretation of the epidemiologic data.

6.10.5 Summary of Key Findings and Conclusions Derived from Lead
Epidemiology Studies
The remarkable progress made since the mid-1980s in understanding the effects of Pb on
health can be gauged by noting the changes that have occurred over time in the questions that
investigators have addressed. In the 1980s, the question of interest was often, "Does low-level
Pb exposure affect health?" The questions asked in more recent studies have more often focused
on details of the associations, including the shapes of concentration-response relationships,
especially at levels well within the range of general population exposures, biological and
socioenvironmental factors that either increase or decrease an individual's risk, the prognoses
associated with Pb-associated effects, the efficacy of interventions to reduce adverse effects, and
so on. In fact, "low-level," a term long-used to describe exposures not sufficiently high to
produce clinical signs and symptoms, is increasingly being recognized as a descriptor that has
little biological meaning and is interpretable only in a specific historical context. What was
considered "low" in the 1980s is an order of magnitude higher than the current mean blood Pb
level in the U.S. population, and the current mean remains perhaps as much as two orders of
magnitude above "natural" background blood Pb levels in humans. The current CDC screening
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guideline for children of 10 |ig/dL is not a "bright line" separating toxicity from safety, but
merely a risk management tool. There is no level of Pb exposure that can yet be clearly
identified, with confidence, as clearly not being associated with potentially increased risk of
deleterious health effects. Recent studies of Pb neurotoxicity in children consistently indicate
that blood Pb levels <10 |ig/dL are associated with neurocognitive deficits. The data are also
suggestive that these effects may be seen at blood Pb levels ranging down to 5 |ig/dL, or perhaps
somewhat lower, but the evidence is less definitive. Public health interventions have resulted in
declines over the past 25 years of more than 90% in the mean blood Pb level within all age and
gender subgroups of the U.S. population, substantially decreasing the numbers of individuals at
risk for toxic effects of Pb. The following provides a listing of most salient key findings for
various classes of health outcomes discussed in this chapter:
• Neurotoxic effects of lead in children. Lead effects on neurobehavior in children have
been observed with remarkable consistency across numerous studies of various designs,
populations, and developmental assessment protocols. The negative impacts of Pb on
neurocognitive ability and other neurobehavioral outcomes persist in most recent studies
even after adjustment for numerous confounding factors, including social class, quality of
caregiving, and parental intelligence. These effects appear to persist into adolescence and
young adulthood. Collectively, the prospective cohort and cross-sectional studies offer
evidence that exposure to Pb affects the intellectual attainment of preschool and school age
children at blood Pb levels <10 |ig/dL (most clearly in the 5 to 10 |ig/dL range, but, less
definitively, possibly lower). Epidemiologic studies have demonstrated that Pb may also
be associated with increased risk for antisocial and delinquent behavior, which may be a
consequence of attention problems and academic underachievement among children who
may have suffered higher exposures to Pb during their formative years. Direct measures
of brain damage using Magnetic Resonance Imaging (MRI) and Magnetic Resonance
Spectroscopy (MRS) also provide evidence suggestive of neurologic harm due to Pb
exposure. Also, pharmacological or nutritional intervention strategies generally have not
been found to be effective in reducing or eliminating Pb-associated neurodevelopmental
morbidities in the absence markedly reduced environmental exposures.
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Neurotoxic effects of lead in adults. In the limited literature examining environmental
Pb exposure, mixed evidence exists regarding associations between Pb and impaired
cognitive performance in adults. Studies using concurrent blood Pb levels as the marker
for Pb exposure found no association between cognitive performance and Pb exposure.
However, significant associations were seen in relation to bone Pb concentrations,
suggesting that long-term cumulative exposure may be crucial in contributing to
neurocognitive deficits in adults. Numerous studies of occupational Pb exposure observed
associations of blood Pb with peripheral sensory nerve impairment, visuomotor and
memory impairment, and postural sway abnormalities. Past high-level occupational Pb
exposures have also been associated with increased risk of developing amyotrophic lateral
sclerosis (ALS), motor neuron disease, and essential tremor. The odds of developing ALS
and essential tremor were significantly increased in individuals with the ALAD2 allele.
These neurobehavioral impairments in occupationally-exposed individuals have typically
been associated with notably elevated blood Pb levels (-30 to 40 jig/dL); however,
essential tremor has been found to be associated with much lower blood Pb levels (mean
3 |ig/dL).

Renal effects of lead. In the general population, both cumulative and circulating Pb has
been found to be associated with longitudinal decline in renal functions. In the large
NHANES III study, alterations in urinary creatinine excretion rate (one indicator of
possible renal dysfunction) was observed in hypertensives at a mean blood Pb of only
4.2 |ig/dL. These results provide suggestive evidence that the kidney may well be a target
organ for effects from Pb in adults at current U.S. environmental exposure levels. The
magnitude of the effect of Pb on renal function ranged from 0.2 to -1.8 mL/min change in
creatinine clearance per 1.0 |ig/dL increase in blood Pb in general population studies.
However, the full significance of this effect is unclear, given that other evidence of more
marked signs of renal dysfunction have not been detected at blood Pb levels below
30-40 |ig/dL among thousands of occupationally-exposed Pb workers that have been
studied. The renal impact of environmental Pb exposure in children is difficult to assess,
because most studies have only measured early biological effect markers and their
prognostic value is uncertain. Studies involving the longitudinal assessment of renal
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function decline in susceptible patient populations have observed that low levels of blood
Pb (<5 |ig/dL) and chelatable Pb levels were associated with decline in glomerular
filtration rate over a 4-year follow-up period in patients with chronic renal insufficiency.

• Cardiovascular effects of lead. Epidemiologic studies support the relationship between
increased Pb exposure and increased deleterious cardiovascular outcomes, including
increased blood pressure and increased incidence of hypertension. A recent meta-analysis
reported that a doubling of blood Pb level was associated with a 1.0 mm Hg increase in
systolic blood pressure and a 0.6 mm Hg increase in diastolic pressure. Studies also have
found that cumulative past Pb exposure (as indexed by bone Pb) may be as important, if
not more, than present exposure in assessing cardiovascular effects. The evidence for an
association of Pb with cardiovascular morbidity and mortality is limited but supportive.

• Reproductive and developmental effects of lead. The epidemiologic evidence suggests
small associations between exposure to Pb and male reproductive outcomes, including
perturbed semen quality and increased time to pregnancy. These associations appear at
blood Pb levels >45 |ig/dL, as most studies have only considered exposure in the
occupational setting. There are no adequate data to evaluate associations between Pb
exposure and female fertility. For many other outcomes, the observed associations are
fairly small, especially at the levels of exposure that are currently of interest. However,
there may be populations that are highly susceptible to Pb-related reproductive effects,
especially if they have additional risk factors for these outcomes.

• Genotoxic and carcinogenic effects of lead. Studies of genotoxicity consistently find
associations of Pb exposure with DNA damage and micronuclei formation; however, the
associations with the more established indicator of cancer risk, chromosomal aberrations,
are inconsistent. Epidemiologic studies of highly-exposed occupational populations
suggest a relationship between Pb and cancers of the lung and the stomach; however the
evidence is limited by the presence of various potential confounders, including
coexposures (e.g., to arsenic and/or Cd), smoking, and dietary habits. The 2003 NTP and
2004 IARC reviews concluded that Pb and Pb compounds were probable carcinogens
based on limited evidence in humans and sufficient evidence in animals. Similarly,
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Pb compounds would likely be classified as probable human carcinogens according to
U.S. EPA (2005) Cancer Guidelines based on experimentally demonstrated carcinogenic
effects in animals, although the human evidence would be considered inadequate
according to those new 2005 guidelines.

Effects of lead on the immune system. Several studies have examined possible
associations between Pb exposures and biomarkers of immune function. Findings from
recent epidemiologic studies suggest that Pb exposure may be associated with effects on
cellular and humoral immunity. These effects include changes in serum immunoglobulin
levels; perturbation of peripheral lymphocyte phenotype profiles, including decreases in
peripheral blood T-cell abundance and changes in T-cell to B-cell abundance ratios;
suppression of lymphocyte activation; and suppression of neutrophil chemotaxis and
phagocytosis. Studies of biomarkers of humoral immunity in children have consistently
found significant associations between increasing blood Pb concentrations and serum
IgE levels at blood Pb levels <10 |ig/dL.

Effects of lead on the hematopoietic system. Lead exposure has been associated with
disruption of heme synthesis in both children and adults. Increases in blood Pb
concentration to -20 to 30 |ig/dL are sufficient to halve erythrocyte ALAD activity and
sufficiently inhibit ferrochelatase to double erythrocyte protoporphyrin levels.
Perturbation of erythropoiesis, indicated by changes in serum erythropoietin and
progenitor cells, occurs in the absence of detectable changes in blood hemoglobin levels or
hematocrit in children and adults at blood Pb levels <40 |ig/dL. A 10% probability of
anemia (hematocrit <35%) is estimated to be associated with a blood Pb level of
-20 |ig/dL at age 1 year.
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7. ENVIRONMENTAL EFFECTS OF LEAD
7.1 TERRESTRIAL ECOSYSTEMS
Surface soils across the United States are enriched in lead (Pb) relative to levels expected
from natural (geogenic) inputs (Erel and Patterson, 1994; Francek, 1992; Friedland et al., 1984;
Marsh and Siccama, 1997; Murray et al., 2004; Yanai et al., 2004). While some of this
contaminant Pb is attributed to paint, salvage yards, shooting ranges, and the use of Pb arsenate
as a pesticide in localized areas (Francek, 1997), Pb contamination of surface soils is essentially
ubiquitous because of atmospheric pollution associated with iron and steel foundaries, boilers
and process heaters, the combustion of fossil fuels in automobiles, trucks, airplanes, and ships,
the manufacturing of cement, and other industrial processes (Table 2-8, Newhook et al., 2003;
Polissar et al., 2001). However, Pb inputs to terrestrial ecosystems in the United States have
declined dramatically in the past 30 years. The primary reason for this decline has been the
almost complete elimination of alkyl-Pb additives in gasoline in North America. Also, emissions
from smelters have declined as older plants have been shut down or fitted with improved
emissions controls.
Most terrestrial ecosystems in North America remain sinks for Pb, despite reductions in
atmospheric Pb deposition of more than 95%. Lead released from forest floor soils in the past
has been largely immobilized in mineral soils (Miller and Friedland, 1994; Johnson et al., 1995,
2004; Kaste et al., 2003; Watmough et al., 2004). The amount of Pb that has leached into the
mineral soil to date ranges from 20 to 90% of the total anthropogenic Pb deposition, depending
on forest type, climate, and litter cycling. While inputs of Pb to ecosystems are currently low,
Pb export from watersheds via groundwater and streams is substantially lower than inputs.
Reported concentrations of Pb in waters draining natural terrestrial ecosystems have always been
low (Bacon and Bain, 1995; Johnson et al., 1995b; Wang et al., 1995; Vinogradoff et al., 2005),
generally less than 1 ng L l, even at moderately polluted sites (Laskowski et al., 1995). Thus,
even at current input levels, watersheds are accumulating industrial Pb (Wang et al., 1995;
Scudlark et al., 2005). However, burial/movement of Pb over time down into lower
soil/sediment layers also tends to sequester it away from more biologically active parts of the
watershed (unless later disturbed or redistributed, e.g., by flooding, dredging, etc.).
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The current chapter first summarizes the most relevant information from the 1986 Lead
Air Quality Criteria Document (Lead AQCD) (U.S. Environmental Protection Agency, 1986)
and then assesses new information that has become available on the potential effects of
atmospheric Pb inputs on the terrestrial ecosystem. It has been organized to address:
methodologies used in terrestrial ecosystem research (Section 7.1.1); the distribution of
atmospherically delivered Pb in terrestrial ecosystems (Section 7.1.2); Pb uptake and
mechanisms of action (Section 7.1.3); toxic effects of Pb on terrestrial organisms (Section 7.1.4);
and, Pb effects on natural terrestrial ecosystems (Section 7.1.5). The major findings and
conclusions from each corresponding Annex section for these subject areas are summarized here.

7.1.1 Methodologies Used in Terrestrial Ecosystem Research
Several methodologies used in terrestrial ecosystems research are described in Annex
Section AX7.1.1, with additional discussion in AX7.1.2 of the application of these methods to
the study of the distribution of atmospherically delivered Pb. One of the key factors necessary
for understanding ecological risks is related to bioavailability. The National Research Council
(NRC) 2002 review on bioavailability defined the "bioavailability processes" in terms of three
key processes. One of these processes, contaminant interactions between phases, is more
commonly referred to as "speciation." For a given metal or metalloid, the term speciation
describes the chemical's ability to interact with its biological or chemical surroundings by
characterizing its physicochemical properties that are relevant to bioavailability.
Methods to address bioavailability (speciation), and methods used to reduce Pb
bioavailability, are summarized in this section.

Analytical Tools and Models
A wide variety of analytical tools have been used to characterize a metal's speciation as it
is found in various media:
• XRD - X-ray diffraction;
• EPMA - electron probe microanalysis;
• PIXE and //PIXE - particle induced X-ray emission;
• XPS - X-ray photoelectron spectroscopy;
• XAS - X-ray absorption spectroscopy;
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• SIMS - secondary ion mass spectrometry;
• sequential extractions; and,
• single chemical extractions.
EPA techniques provide the greatest information on metal speciation. Other techniques,
such as EXAFS (extended X-ray absorption fine structure) and EXANES (extended X-ray
absorption near edge spectroscopy), show great promise and will be important in solving key
mechanistic questions. In the case of phytotoxicity, the speciation of metals by direct
measurement or chemical models of pore water chemistry is most valuable.
The tools that have been used most often to evaluate speciation for metal particles in
various media include the following computer-based models: SOILCHEM, MINTEQL,
REDEQL2, ECOSAT, MINTEQA2, HYDRAQL, PHREEQE, and WATEQ4F.

Metal Speciation for Plants
When considering the bioavailability of a metal to plants from soils and sediments, it is
generally assumed that both the kinetic rate of supply and the speciation of the metal to either the
root or shoot are highly important. In soils and sediments, generally only a small volume of
water is in contact with the chemical form, and although the proportion of a metal's
concentration in this pore water to the bulk soil/sediment concentration is small, it is this phase
that is directly available to plants. Therefore, pore water chemistry (i.e., metal concentration as
simple inorganic species, organic complexes, or colloid complexes) is most important.
Tools currently used for metal speciation for plants include (1) in situ measurements using
selective electrodes (Gundersen et al., 1992; Archer et al., 1989; Wehrli et al., 1994); (2) in situ
collection techniques using diffusive equilibrium thin films (DET) and diffusive gradient thin
films (DGT) followed by laboratory analyses (Davison et al., 1991, 1994; Davison and Zhang,
1994; Zhang et al., 1995); and (3) equilibrium models (SOILCHEM) (Sposito and Coves, 1988).

Influence of Soil Amendments on Bioavailability
The removal of contaminated soil to mitigate exposure of terrestrial ecosystem
components to Pb can often present both economic and logistical problems. Because of this,
recent studies have focused on in situ methodologies to lower soil-Pb relative bioavailability
(RBA) (Brown et al., 2003a,b). To date, the most common methods studied include the addition
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of soil amendments in an effort to either lower the solubility of the Pb form or to provide
sorption sites for fixation of pore-water Pb. These amendments typically fall within the
categories of phosphate, biosolid, and Al/Fe/Mn-oxide amendments.
Phosphate amendments have been studied extensively and, in some cases, offer the most
promising results (Brown et al., 1999; Ryan et al., 2001; Cotter-Howells and Caporn, 1996;
Hettiarachchi et al., 2001, 2003; Rabinowitz, 1993; Yang et al., 2001; Ma et al., 1995).
A number of potentially significant problems associated with phosphate amendments have been
recognized. The added phosphate poses the potential risk of eutrophication of nearby waterways
from soil runoff. There also may be both phyto- and earthworm toxicity (Ownby et al., 2005;
Cao et al., 2002; Rusek and Marshall, 2000), primarily associated with very high applications of
phosphorous and/or decreased soil pH. Indications of phytotoxicity are often balanced by studies
such as Zhu et al. (2004) that illustrate a 50 to 70% reduction in shoot-root uptake of Pb in
phosphate-amended soils. It also has been shown (Impellitteri, 2005; Smith et al., 2002; Chaney
and Ryan, 1994; Ruby et al., 1994) that the addition of phosphate would enhance arsenic
mobility (potentially moving arsenic down into the groundwater) through competitive anion
exchange. Some data (Lenoble et al., 2005) indicate that this problem can be mitigated if arsenic
and Pb contaminated soils could be amended with iron(III) phosphate, although there could still
be issues with drinking water quality.
Biosolids have been used historically in the restoration of coal mines (Haering et al.,
2000; Sopper, 1993). More recently, workers have demonstrated the feasibility of their use in
the restoration of mine tailings (Brown et al., 2000) and urban soils (Brown et al., 2003a; Farfel
et al., 2005). As with phosphate amendments, problems with biosolid application have also been
documented. Studies have shown that metal transport is significantly accelerated in soils
amended with biosolids (Al-Wabel et al., 2002; McBride et al., 1997, 1999; Lamy et al., 1993;
Richards et al., 1998, 2000).

7.1.2 Distribution of Atmospherically Delivered Lead in Terrestrial
Ecosystems
Advances in technology since the 1986 Lead AQCD have allowed for a more quantitative
determination of the mobility, distribution, uptake, speciation, and fluxes of atmospherically-
delivered Pb in terrestrial ecosystems. In most cases, Pb appears to be strongly bound to soils
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and sediments in terrestrial ecosystems, which prevents substantial mobility and uptake in the
terrestrial environment (e.g., Bacon et al., 2005). However, the controls on Pb speciation, and
thus on mobility and potential bioavailability are not completely understood, so there remains a
considerable need for more research on this topic.

Lead Speciation in Solid Phases
Lead can enter terrestrial ecosystems through natural rock weathering and by a variety of
anthropogenic pathways. During the hydrolysis and oxidation of Pb-containing minerals,
divalent Pb (Pb2+) is released to the soil solution where it is rapidly fixed by organic matter
and secondary mineral phases (Kabata-Pendias and Pendias, 1992; Erel et al., 1997). The
geochemical form of natural Pb in terrestrial ecosystems will be strongly controlled by soil type,
PH, and parent material (Emmanuel and Erel, 2002). In contrast, anthropogenically-introduced
Pb has a variety of different geochemical forms, depending on the specific source. While Pb in
soils from battery reclamation areas can be in the form of PbSC>4 or PbSiOs, Pb in soils from
shooting ranges and paint spills is commonly found as PbO and a variety of Pb carbonates
(Vantelon et al., 2005; Laperche et al., 1996; Manceau et al., 1996). Atmospherically-delivered
Pb resulting from fossil fuel combustion is typically introduced into terrestrial ecosystems as
Pb-sulfur compounds and Pb oxides (Olson and Skogerboe, 1975; Clevenger et al., 1991;
Batonneau et al., 2004; Utsunomiya et al., 2004). After deposition, Pb species are likely
transformed. Although the specific factors that control the speciation of anthropogenic Pb
speciation in soils are not well understood, there are many studies that have partitioned Pb into
its different geochemical phases. A thorough understanding of Pb speciation is very important in
order (a) to predict potential mobility and bioavailability and (b) to accurately apply a critical
loads methodology for determining air quality standards (Lawlor and Tipping, 2003; Paces,
1998). See Section 7.3 for more discussion of critical loads methodology.
Selective chemical extractions have been employed extensively for quantifying amounts
of a particular metal phase (e.g., PbS, Pb-humate, Pb-Fe, Mn oxide) present in soil rather than
total metal concentration. Selective extractions can be a relatively rapid, simple, and inexpensive
means for determining metal phases in soils, and the generated data can be linked to potential
mobility and bioavailability of the metal (Tessier and Campbell, 1987). However, some
problems persist with the selective extraction technique. First, extractions are rarely specific to a
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single phase. For example, while peroxide (H2O2) is often used to remove metals bound in
organic matter in soils, some researchers have demonstrated that this reagent destroys clay
minerals and sulfides (Ryan et al., 2002). Peroxide solutions may also be inefficient at removing
metals bound to humic acids and, in fact, could potentially result in the precipitation of metal-
humate substances. In addition to non-selectivity of reagents, significant metal redistribution has
been documented during sequential chemical extractions (Ho and Evans, 2000; Sulkowski and
Hirner, 2006), and many reagents may not extract targeted phases completely. Therefore, while
chemical extractions do provide some useful information on metal phases in soil and scenarios
for mobilization, the results should be treated as "operationally defined," e.g., "H2O2
liberated-Pb" rather than "organic Pb."
Synchrotron radiation (X-rays) allows researchers to probe the electron configuration of
metals in untreated soil samples. Because different elements have different electron binding
energies, X-rays can be focused in an energy window specific to a metal of interest. The precise
energy required to dislodge a core electron from a metal will be a function of the oxidation state
and covalency of the metal. Since the electron configuration of a Pb atom is directly governed
by its speciation (e.g., Pb bound to organics, Pb adsorbed to oxide surfaces, PbS, etc.), X-ray
absorption experiments are a powerful in situ technique for determining speciation that does not
suffer from some of the problems of chemical extractions (Bargar et al., 1997a,b; Bargar et al.,
1998).
Selective chemical extractions and synchrotron-based X-ray studies have shown that
industrial Pb can be strongly sequestered by organic matter and secondary minerals such as clays
and oxides of Al, Fe, and Mn (Miller and McFee, 1983; Jersak et al., 1997; Johnson and Petras,
1998; Kaste et al., 2005). More recent X-ray studies have demonstrated the importance of
biomineralization of Pb in soils by bacteria and nematodes (Jackson et al., 2005; Templeton
et al., 2003a,b; Xia et al., 1997).

Lead Solid-Solution Partitioning
The concentration of Pb species dissolved in soil solution is probably controlled by some
combination of a) Pb mineral solubility equilibria, b) adsorption reactions of dissolved Pb phases
on inorganic surfaces (e.g., oxides of Al, Fe, Si, Mn, etc., clay minerals), and c) adsorption
reactions of dissolved Pb phases on soil organic matter. Dissolved Pb phases in soil solution can
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be some combination of Pb2+ and its hydrolysis species, Pb bound to dissolved organic matter,
and Pb complexes with inorganic ligands such as Cl and SO42 . Alkaline soils typically have
solutions supersaturated with respect to PbCO3, Pb3(CO3)2(OH)2, Pb(OH)2, Pb3(PO4)2,
Pb5(PO4)3(OH), and Pb4O(PO4)2 (Badawy et al., 2002). Pb phosphate minerals in particular, are
very insoluble, and calculations based on thermodynamic data predict that these phases will
control dissolved Pb in soil solution under a variety of conditions (Nriagu, 1974; Ruby et al.,
1994). However, certain chelating agents, such as dissolved organic matter can prevent the
precipitation of Pb minerals (Lang and Kaupenjohann, 2003), and the natural formation of these
minerals has not yet been observed in terrestrial ecosystems (Kaste et al., 2006).
Increasing soil solution dissolved organic matter content and decreasing pH typically are
strongly correlated with increases in the concentration of dissolved Pb species (Badawy et al.,
2002; Sauve et al., 1998, 2000a,b, 2003; Tipping et al., 2003; Weng et al., 2002). In the case of
adsorption phenomena, the partitioning of Pb2+ to the solid phase is also controlled by total metal
loading: high Pb loadings will result in a lower fraction partitioned to the solid phase. Sauve
et al. (1998; 1997) demonstrated that only a fraction of the total Pb in solution was actually Pb2+
in soils treated with leaf compost. The fraction of Pb2+ to total dissolved Pb ranged from <1 to
60%, depending on pH and the availability of Pb-binding ligands. In acidic soils, Al species can
compete for sites on natural organic matter and inhibit Pb binding to surfaces (Gustafsson et al.,
2003).

Tracing the Fate of Atmospherically Delivered Lead
Radiogenic Pb isotopes offer a powerful tool for separating anthropogenic Pb from natural
Pb derived from mineral weathering (Erel and Patterson, 1994; Erel et al., 1997). This is
particularly useful for studying Pb in mineral soil, where geogenic Pb often dominates. The ore
bodies from which anthropogenic Pb are typically derived are usually enriched in 207Pb relative
to 206Pb and 208Pb when compared with Pb found in granitic rocks. Uranium-238 series 210Pb
also provides a tool for tracing atmospherically delivered Pb in soils. Fallout 210Pb is deposited
onto forests via wet and dry deposition, similar to anthropogenic Pb deposition in forests, and is
thus useful as a tracer for non-native Pb in soils. 210Pb is convenient to use for calculating the
residence time of Pb in soil layers because its atmospheric and soil fluxes can be assumed to be
in steady-state at undisturbed sites (Dorr, 1995; Dorr and Munnich, 1989; Kaste et al., 2003).
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Researchers assessing the fate of atmospheric Pb in soils have also relied on repeated
sampling of soils and vegetation for total Pb. This technique works best when anthropogenic Pb
accounts for the vast majority of total Pb in a particular reservoir. Johnson et al. (1995), Yanai
et al. (2004), and Friedland et al. (1992) used O horizon (forest floor) time series data to evaluate
the movement of gasoline-derived Pb in the soil profile. Surface soils sampled relatively
recently demonstrate that the upper soil horizons (O + A horizons) are retaining most of the
anthropogenic Pb burden introduced to the systems during the 20th century (Evans et al., 2005).
Miller and Friedland (1994) and Wang and Benoit (1997) suggested that the vertical movement
of organic particles dominated Pb transport in the soil profile.
By describing the movement of atmospherically-delivered Pb in terrestrial ecosystems, we
can begin to predict the Pb inventories of various ecosystem compartments that a particular
atmospheric deposition rate will support. This type of information is very pertinent to air quality
issues. For example, if the rate of Pb loss is known for a particular soil horizon and reasonable
assumptions can be made about biogeochemical cycling and chemical weathering inputs, then
steady-state Pb concentrations can be calculated for any constant deposition rate (in mass of Pb
deposited per square meter). First-order rate loss constants, &, have been calculated for organic
horizons using forest floor inventories, radiogenic 207Pb tracer techniques, and fallout 210Pb
(Miller and Friedland, 1994; Johnson et al., 1995; Kaste et al., 2003; Watmough et al., 2004;
Kaste et al., 2006). First order rate loss constants vary substantially, ranging between -0.003 to
-0.6 (1/y), depending on soil type and climate. Wang and Benoit (1997) used the first-order rate
loss technique to model forest floor Pb dynamics at the Hubbard Brook Experimental Forest in
New Hampshire. They concluded that with steady Pb deposition at 0.0065 kg/ha/y, the forest
floor would reach a steady-state Pb concentration of 1.4 ppm. Calculated steady-state Pb
contents of different ecosystem compartments can then be compared with experimentally-
derived toxicity thresholds (Liang and Tabatabai, 1977; 1978) to put deposition rates into context
with the terrestrial ecosystem.

7.1.3 Species Response/Mode of Action
The current document expands upon and updates knowledge since 1986 related to the
uptake, detoxification, physiological effects, and modifying factors of Pb toxicity to terrestrial
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organisms. Terrestrial organisms discussed in this chapter include soil organisms, plants, birds,
and mammals.

Uptake into Plants and Invertebrates
Recent work supports previous results and conclusions that surface deposition of Pb onto
above-ground vegetation from airborne sources may be significant (Dalenberg and Van Driel,
1990; Jones and Johnston, 1991; Angelova et al., 2004). In addition, most Pb is taken up by
plants via the symplastic route (through cell membranes) (Sieghardt, 1990) and remains in the
roots, with little translocation to shoots, leaves, or other plant parts. Different species of plants
and invertebrates accumulate different amounts of Pb (Pizl and Josens, 1995; Terhivuo et al.,
1994; Wierzbicka, 1999).
Recent work supports previous conclusions that the form of metal tested, and its
speciation in soil, influence uptake and toxicity to plants and invertebrates. The oxide form is
less toxic than the chloride or acetate forms, which are less toxic than the nitrate form of Pb
(Khan and Frankland, 1983; Lock and Janssen, 2002; Bongers et al., 2004). However, these
results must be interpreted with caution, as the counter ion (e.g., the nitrate ion) may be
contributing to the observed toxicity (Bongers et al., 2004).

Detoxification in Plants and Invertebrates
Lead may be deposited in root cell walls as a detoxification mechanism, and this may be
influenced by calcium (Antosiewicz, 2005). Yang et al. (2000) suggested that the oxalate
content in root and root exudates reduced the bioavailability of Pb in soil, and that this was an
important tolerance mechanism. Other hypotheses put forward recently include the presence of
sulfur ligands (Sharma et al., 2004) and the sequestration of Pb in old leaves (Szarek-
Lukaszewska et al., 2004) as detoxification mechanisms.
Lead detoxification has not been studied extensively in invertebrates. Glutathione
detoxification enzymes were measured in two species of spider (Wilczek et al., 2004). Lead may
be stored in waste nodules in earthworms (Hopkin, 1989) or as pyromorphite in the nematode
(Jackson et al., 2005).
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Physiological Effects
The effects on heme synthesis (as measured by 5-aminolaevulinic acid dehydratase
[ALAD] activity and protoporphyrin concentration, primarily) had been well-documented in the
1986 AQCD (U.S. Environmental Protection Agency, 1986) and continue to be studied (Schlick
et al., 1983; Scheuhammer, 1989; Redig et al., 1991; Henny et al., 1991; Beyer et al., 2000;
Hoffman et al., 2000a, b). However, Henny et al. (1991) caution that changes in ALAD and
other enzyme parameters are not always related to adverse effects, but simply indicate exposure.
Other effects on plasma enzymes, which may damage other organs, have been reported (Brar
et al., 1997a, b). Lead also may cause lipid peroxidation (Mateo and Hoffman, 2001) which may
be alleviated by Vitamin E, although lead poisoning may still result (Mateo et al., 2003b).
Changes in fatty acid production have been reported, which may influence immune response and
bone formation (Mateo et al., 2003a).

Response Modification
Genetics, biological factors, physical/environmental factors, nutritional factors and other
pollutants can modify terrestrial organism response to Pb. Fisher 344 rats were found to be more
sensitive to Pb than Sprague-Dawley rats (Dearth et al., 2004). Younger animals are more
sensitive than older animals (Eisler, 1988; Scheuhammer, 1991), and females generally are more
sensitive than males (Scheuhammer, 1987; Tejedor and Gonzalez, 1992; Snoeijs et al., 2005).
Monogastric animals are more sensitive than ruminants (Humphreys, 1991). Insectivorous
mammals may be more exposed to Pb than herbivores (Beyer et al., 1985; Sample et al., 1998),
and higher tropic-level consumers may be less exposed than lower trophic-level organisms
(Henny et al., 1991). Diets deficient in nutrients (including calcium) result in increased uptake
of Pb (Snoeijs et al., 2005) and greater toxicity (Douglas-Stroebel et al., 2005) in birds, relative
to diets containing adequate nutrient levels.
Mycorrhizal fungi may ameliorate Pb toxicity until a threshold is surpassed (Malcova and
Gryndler, 2003), which may explain why some studies show increased uptake into plants (Lin
et al., 2004) while others show no difference or less uptake (Dixon, 1988). Uptake of Pb into
plants and soil invertebrates increases with a decrease in soil pH. However, calcium content,
organic matter content, and cation exchange capacity of soils also had a significant influence on
uptake of Pb into plants and invertebrates (Beyer et al., 1987; Morgan and Morgan, 1988).
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Interactions of Pb with other metals are inconsistent, depending on the endpoint
measured, the tissue analyzed, the animal species, and the metal combination (Phillips et al.,
2003; An et al., 2004; Garcia and Corredor, 2004; He et al., 2004; Perottoni et al., 2005).

7.1.4 Exposure/Response of Terrestrial Species
The current document expands upon and updates knowledge related to the effects of Pb
on terrestrial primary producers, consumers and decomposers found in the 1986 Lead AQCD
(U.S. Environmental Protection Agency, 1986). Lead exposure may adversely affect organisms
at different levels of organization, i.e., individual organisms, populations, communities, or
ecosystems. Generally, however, there is insufficient information available for single materials
in controlled studies to permit evaluation of specific impacts on higher levels of organization
(beyond the individual organism). Potential effects at the population level or higher are, of
necessity, extrapolated from individual level studies. Available population, community, or
ecosystem level studies are typically conducted at sites that have been contaminated or adversely
affected by multiple stressors (several chemicals alone or combined with physical or biological
stressors). Therefore, the best documented links between Pb and effects on the environment are
with effects on individual organisms. Impacts on terrestrial ecosystems are discussed in
Section 7.1.5 and Annex AX7.1.5.

Primary Producers
Effects of Pb on terrestrial plants include decreased photosynthetic and transpiration rates,
and decreased growth and yield. The phytotoxicity of Pb is considered to be relatively low,
compared to other metals, and there are few reports of phytotoxicity from Pb exposure under
field conditions. Phytotoxicity data recently were reviewed for the development of the
ecological soil screening levels (Eco-SSL) (U.S. Environmental Protection Agency, 2005b).
Many of the toxicity data presented in U.S. Environmental Protection Agency (2005b) are lower
(i.e., they represent greater toxicity) than those discussed in the 1986 Lead AQCD, although both
documents acknowledge that toxicity is observed over a wide range of Pb concentrations in soil
(tens to thousands of mg/kg soil). This may be due to many factors, such as the soil conditions
(e.g., pH, organic matter) and differences in bioavailability of the Pb in spiked soils, perhaps due
to lack of equilibration of the Pb solution with the soil after spiking. Most phytotoxicity data
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continue to be developed for agricultural plant species (i.e., vegetable and grain crops). Few data
are available for trees or native herbaceous plants, although two of the five ecotoxicological
endpoints used to develop the Eco-SSL were for trees and two were for clover.

Consumers
Effects of Pb on avian and mammalian consumers include decreased survival,
reproduction, and growth, as well as effects on development and behavior. There remain few
field effects data for consumers, except from sites with multiple contaminants, for which it is
difficult to attribute toxicity specifically to Pb. Avian and mammalian toxicity data recently
were reviewed for the development of Eco-SSLs (U.S. Environmental Protection Agency,
2005b). Many of the toxicity data presented by EPA (U.S. Environmental Protection Agency,
2005b) are lower than those discussed in the 1986 Lead AQCD, (i.e., the Eco-SSL document
describes studies which report greater toxicity of Pb to various organisms), although EPA (U.S.
Environmental Protection Agency, 2005b) recognizes that toxicity is observed over a wide range
of doses (<1 to > 1,000 mg Pb/kg bw-day). Most toxicity data for birds are derived from chicken
and quail studies, and most data for mammals are derived from laboratory rat and mouse studies.
Data derived for other species would contribute to the understanding of Pb toxicity, particularly
for wildlife species with different gut physiologies. In addition, data derived using
environmentally-realistic exposures, such as from Pb-contaminated soil and food may be
recommended. Finally, data derived from inhalation exposures, which evaluate endpoints such
as survival, growth, and reproduction, would contribute to understanding the implications of
airborne releases of Pb.

Decomposers
Effects of Pb on soil invertebrates include decreased survival, growth and reproduction.
Effects on microorganisms include changes in nitrogen mineralization, and changes in enzyme
activities. Recent data on Pb toxicity to soil invertebrates and microorganisms are consistent
with those reported in the 1986 Lead AQCD, with toxicity generally being observed at
concentrations of hundreds to thousands of mg Pb/kg soil. Studies on microbial processes may
be influenced significantly by soil parameters, and the significance of the test results is not clear.
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Ecological Soil Screening Levels (Eco-SSLs)
Eco-SSLs are concentrations of contaminants in soils that would result in little or no
measurable effect on ecological receptors (U.S. Environmental Protection Agency, 2005a). They
were developed by U.S. EPA for use in the screening-level assessments at Superfund sites to
identify those contaminants needing further investigation, and also to identify those contaminants
that are not of potential ecological concern and do not need to be considered in the subsequent
analyses. They were developed following rigorous scientific protocols, and were subjected to
two rounds of peer review. However, several conservative factors were incorporated into their
development. For the plant and invertebrate Eco-SSLs, studies were scored to favor relatively
high bioavailability. For wildlife Eco-SSLs, only species with a clear exposure link to soil were
considered (generalist species, species with a link to the aquatic environment, or species which
consume aerial insects were excluded), simple diet classifications were used (100% plants, 100%
earthworms or 100% animal prey) when in reality wildlife consume a varied diet, species were
assumed to forage exclusively at the contaminated site, relative bioavailability or Pb in soil and
diet was assumed to be 1, and the TRY was selected as the geometric mean of NOAELs unless
this value was higher than the lowest bounded LOAEL for mortality, growth or reproduction
(U.S. Environmental Protection Agency, 2005a,b). The Eco-SSLs are intentionally conservative
in order to provide confidence that contaminants which could present an unacceptable risk are
not screened out early in the evaluation process. That is, at or below these levels, adverse effects
are considered unlikely. Due to conservative modeling assumptions (e.g., metal exists in most
toxic form or highly bioavailable form, high food ingestion rate, high soil ingestion rate), which
are common to screening processes, several Eco-SSLs are derived below the average background
soil concentration for a particular contaminant. For example, Scheuhammer et al. (2003) found
that ninety-one percent (64/70) of the soil samples analyzed in eastern Canada had <45 mg/kg
Pb, concentrations typical of noncontaminated rural soils in Canada and elsewhere in North
America (Breckenridge and Crockett, 1998; McKeague and Wolynetz, 1980). However, the
Eco-SSL for birds (based on the American woodcock) is recommended as 11 mg/kg (U.S.
Environmental Protection Agency, 2005a,b).
The Eco-SSLs for terrestrial plants, birds, mammals, and soil invertebrates are 120 mg/kg,
11 mg/kg, 56 mg/kg and 1700 mg/kg, respectively. See Annex Section AX7.1.4 for additional
information.
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7.1.5 Effects of Lead on Natural Terrestrial Ecosystems
Few significant effects of Pb pollution have been observed at sites that are not near point
sources of Pb. At present, industrial point sources such as smelter sites represent the greatest
Pb-related threat to the maintenance of sustainable, healthy, diverse, and high-functioning
terrestrial ecosystems in the United States. However, assessing the risks specifically associated
with Pb is difficult because these sites also experience elevated concentrations of other metals
and because of effects related to SC>2 emissions. Terrestrial ecosystems may respond to stress in
a variety of ways, including reductions in the vigor and/or growth of vegetation, reductions in
biodiversity, and effects on energy flow and biogeochemical cycling.

Influence of Acidification
Like most metals, the solubility of Pb increases as pH decreases (Stumm and Morgan,
1995), suggesting that enhanced mobility of Pb should be found in ecosystems under
acidification stress. However, Pb is also strongly bound to organic matter in soils and sediments.
Reductions in pH may cause a decrease in the solubility of dissolved organic matter (DOM), due
to the protonation of carboxylic functional groups (Tipping and Woof, 1990). Because of the
importance of Pb complexation with organic matter, lower DOM concentrations in soil solution
resulting from acidification may offset the increased solubility of Pb and hence decrease the
mobility of the organically bound metal. Increased mobility was only observed in very acidic
soils, those with pH <4.5 (Blake and Goulding, 2002). Acidification also may enhance Pb export
to drainage water in very sandy soils, with limited ability to retain organic matter (Swanson and
Johnson, 1980; Turner etal., 1985).

Influence of Land Use and Industry
Changes in land use represent potentially significant changes in the cycling of organic
matter in terrestrial ecosystems. Conversion of pasture and croplands to woodlands changes the
nature and quantity of organic matter inputs to the soil. The introduction of industrial activity
may have consequences for organic matter cycling, and subsequently, Pb mobilization. In a rare
long-term study of polluted soils, Egli et al. (1999) found that loss of soil carbon can induce the
mobilization and loss of Pb from terrestrial ecosystems. However, it is worth noting that the
decline in soil Pb was considerably smaller than the decline in organic carbon. This suggests
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that Pb mobilized during organic matter decomposition can resorb to remaining organic matter or
perhaps to alternate binding sites (e.g., Fe and Mn oxides).
Forest harvesting represents a severe disruption of the organic matter cycle in forest
ecosystems. However, observations from clear-cut sites in the United States and Europe indicate
that forest harvesting causes little or no mobilization or loss of Pb from forest soils (Berthelsen
and Steinnes, 1995; Fuller et al., 1988). The principal risk associated with forest harvesting is
the loss of Pb in particulate form to drainage waters through erosion.

Effects Observed Around Industrial Point Sources
The effects of Pb exposure on natural ecosystems are confounded by the fact that Pb
exposure cannot be decoupled from other factors that may also affect the ecosystem under
consideration. Principal among these factors are other trace metals and acidic deposition.
Emissions of Pb from smelting and other industrial activities are accompanied by other trace
metals (e.g., Zn, Cu, Cd) and sulfur dioxide (862) that may cause toxic effects independently or
in concert with Pb.
Natural terrestrial ecosystems near smelters, mines, and other industrial plants have
exhibited a variety of effects related to ecosystem structure and function. These effects include
decreases in species diversity, changes in floral and faunal community composition, and
decreasing vigor of terrestrial vegetation. All of these effects were observed in ecosystems
surrounding the Anaconda copper smelter in southwestern Montana, which operated between
1884 and 1980 (Galbraith et al., 1995; Kapustka et al., 1995). Similar observations were made in
the area surrounding Palmerton, Pennsylvania, where two zinc smelters operated between 1898
and 1980 (Jordan, 1975; Sopper, 1989; Storm et al., 1994). Subsequent to the effects on
vegetation, wind and erosion may remove litter and humus, leaving bare mineral soil, a nearly
sterile environment in which very little energy transfer takes place (Little and Martin, 1972;
Galbraith et al., 1995). Metal pollution around a Pb-Zn smelter near Bristol, England has not
resulted in the loss of oak woodlands within 3 km of the smelter, despite significant
accumulation of Pb, Cd, Cu, and Zn in soils and vegetation (Martin and Bullock, 1994).
However, the high metal concentrations have favored the growth of metal-tolerant species in the
woodland.
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The effects of Pb and other chemical emissions on terrestrial ecosystems near smelters
and other industrial sites decrease downwind from the source. Several studies using the soil
burden as an indicator have shown that much of the contamination occurs within a radius of 20
to 50 km around the emission source (e.g., Miller and McFee, 1983; Martin and Bullock, 1994;
Galbraith et al., 1995; Spurgeon and Hopkin, 1996). Elevated metal concentrations around
smelters have been found to persist despite significant reductions in emissions (Hrsak et al.,
2000). The confounding effect of other pollutants makes the assessment of Pb-specific
exposure-response relationships very difficult at the whole-ecosystem level.

Influence of Climate Change
Atmospheric Pb is not likely to contribute significantly to global climate change. The
potential linkages between climate-related stress and Pb cycling are poorly understood. Effects
related to alterations in organic matter cycling may influence Pb migration. For example, an
increase in temperature leading to increased rates of organic matter decomposition could lead to
temporary increases in DOM concentrations and smaller steady-state pools of soil organic
matter. There also is some evidence for recent increases in the frequency of soil freezing events
in the northeastern United States (Mitchell et al., 1996). Soil freezing occurs when soils have
little or no snow cover to insulate them from cold temperatures and results in an increased
release of nitrate and DOC from the O horizons of forest soils (Mitchell et al., 1996; Fitzhugh
et al., 2001). Increased fluctuations in precipitation may induce more frequent flooding,
potentially increasing inputs of Pb and other metals to floodplain soils (Kriiger and Grongroft,
2004). All of these factors could result in increased concentrations of Pb in waters draining
terrestrial ecosystems.

Influence on Energy Flow and Biogeochemical Cycling
Lead can have a significant effect on energy flow in terrestrial ecosystems. In terrestrial
ecosystems, energy flow is closely linked to the carbon cycle. The principal input of energy to
terrestrial ecosystems is through photosynthesis, in which CO2 is converted to biomass carbon.
Because of this link between photosynthesis and energy flow, any effect that Pb has on the
structure and function of terrestrial ecosystems influences the flow of energy into the ecosystem.
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At some sites severely affected by metal pollution, death of vegetation can occur, dramatically
reducing the input of carbon to the ecosystem (Jordan, 1975; Galbraith et al., 1995).
Lead influences energy transfer within terrestrial ecosystems, which begins with the
decomposition of litter and other detrital material by soil bacteria and fungi, and cascades
through the various components of the detrital food web. In acid- and metal-contaminated soils
or soils treated with Pb investigators have documented significant declines in litter
decomposition rates (Cotrufo et al., 1995; Johnson and Hale, 2004) and/or the rate of carbon
respiration (Laskowski et al., 1994; Cotrufo et al., 1995; Saviozzi et al., 1997; Niklinska et al.,
1998; Palmborg et al., 1998; Aka and Darici, 2004). The resulting accumulation of organic
matter on the soil surface can be dramatic.
Because Pb mobility in soils is closely tied to organic matter cycling, decomposition
processes are central to the biogeochemical cycle of Pb. Reduced decomposition rates in
polluted ecosystems are the result of the inhibition of soil bacteria and fungi and its effects on
microbial community structure (Baath, 1989). Lead and other metals also inhibit the
mineralization of nitrogen from soil organic matter and nitrification (Liang and Tabatabai, 1977,
1978; Senwo and Tabatabai, 1999; Acosta-Martinez and Tabatabai, 2000; Ekenler and
Tabatabai, 2002), resulting in lower nitrogen availability to plants. This suggests that the
inhibitory effect of Pb and other metals is broad-based, and not specific to any particular
metabolic pathway. It is important to note that terrestrial sites that have exhibited significant
disruption to energy flows and C processing are sites that have experienced severe metal
contamination from smelters or other metals-related activities.
7.2 AQUATIC ECOSYSTEMS
The overall intent of this Section 7.2 is to provide sufficient information to support
development of air quality criteria for Pb that is protective of aquatic ecosystems. To achieve
this objective, the logical starting points are to (1) gain a general understanding of the current
distribution and concentrations of Pb in the aquatic environment and (2) identify the threshold
levels for Pb effects on aquatic populations, communities, and ecosystems. Ambient water
quality criteria for Pb and other chemicals represent surface water concentrations intended to be
protective of aquatic communities, including recreationally and commercially important species.
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The EPA derives Ambient Water Quality Criteria (AWQC) to provide guidance to States and
Tribes that are authorized to establish water quality standards under the Clean Water Act
(CWA). Similarly, EPA has recommended sediment quality benchmarks for Pb and other
divalent metals that, although not truly being criteria, do represent concentrations in sediment
that are derived to be protective of benthic (sediment) organisms. As summarized further below
and in subsequent sections, the U.S. EPA has increasingly focused on developing AWQC and
sediment quality benchmarks for Pb and other metals that account for the bioavailability of the
metal to aquatic life. These criteria and benchmark concentrations in water and sediment
represent appropriate starting points to ensure that air quality criteria for Pb are adequately
protective of aquatic life.
Since publication of the 1986 Lead AQCD (U.S. Environmental Protection Agency,
1986), knowledge has expanded with regard to the fate and effects of Pb in aquatic ecosystems
and on the distribution and concentrations of Pb in surface waters throughout the United States.
In addition, chemical, physical, and biological properties of Pb are discussed. The following
provides a general overview of the key information found in corresponding Annex sections
(Sections AX7.2.1 through AX7.2.5).

7.2.1 Methodologies Used in Aquatic Ecosystem Research
Ambient Water Quality Criteria and Bioavailability
The U.S. EPA guidelines for developing AWQC (Stephan et al., 1985) were published
more than 20 years ago. Scientific advances in aquatic toxicology and risk assessment have been
made since the 1980s. For example, the toxicological importance of dietary metals has been
increasingly recognized and approaches for incorporating dietary metals into regulatory criteria
are being evaluated (Meyer et al., 2005). Other issues include consideration of certain sublethal
endpoints that are currently not directly incorporated into AWQC development (e.g., endocrine
toxicity, behavioral responses) and protection of threatened and endangered (T&E) species (U.S.
Environmental Protection Agency, 2003). In deriving appropriate and scientifically defensible
air quality criteria for Pb, it will be important that the state-of-the-science for metals toxicity in
aquatic systems be considered in the development process.
9-1-
The primary form of Pb in freshwater and marine environments is divalent Pb (Pb ).
In surface waters, the bioavailability of Pb to aquatic biota is driven by a variety of factors,
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including calcium, dissolved organic carbon (DOC), pH, alkalinity, and total suspended solids
(TSS). Accounting for the influence of calcium and magnesium ions on Pb bioavailability, the
current AWQC for Pb are normalized to the hardness of the receiving water (Table 7-1).
Table 7-1. Summary of Lead Ambient Water Quality Criteria for Freshwater
Organisms at Different Hardness Levels1 (criteria expressed as dissolved lead)
Hardness Acute Criterion Chronic Criterion
(mg/L as CaCO3) (jig/L) (ug/L)
50 30 1.2
100 65 2.5
200 136 5.3
1 The acute and chronic criteria values are based on empirical data on the relationship between toxicity
and hardness.
More recently, the biotic ligand model (BLM), which considers the binding of free metal
ion to the site of toxic action and competition between metal species and other ions, has been
developed to predict the toxicity of several metals under a variety of water quality conditions.
However, there are limitations to this tool in deriving AWQC because, currently, only limited
work has been conducted in developing chronic BLMs (for any metals, let alone Pb) and the
acute BLMs to-date do not account for dietary metal exposures.
The U.S. EPA is currently revising the aquatic life AWQC for Pb, which will include
toxicity data published after the 1985 AWQC were released and incorporation of the BLM is
being evaluated.

Sediment Quality Benchmarks and Bioavailability
As in surface waters, there are a number of factors in sediment that can influence Pb
bioavailability to benthic (sediment) organisms. Although sediment quality criteria have not
been formally adopted, the EPA has published an equilibrium partitioning procedure for
developing sediment criteria for metals (U.S. Environmental Protection Agency 2005c).
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Equilibrium partitioning (EqP) theory predicts that metals partition in sediment between acid

volatile sulfide, pore water, benthic organisms, and other sediment phases, such as organic

carbon. When the sum of the molar concentrations of simultaneously extracted metal (ZSEM)

minus the molar concentration of AVS is less than zero, it can accurately be predicted that

sediments are not toxic because of these metals. Further, if ZSEM-AVS is normalized to the

fraction of organic carbon (i.e., (ZSEM-AVS)/fOC), mortality can be more reliably predicted by

accounting for both the site-specific organic carbon and AVS concentrations (Table 7-2).
Table 7-2. Summary of Sediment Quality Benchmarks and
Guidelines for Lead
Benchmark/
Guideline Type
Source
Effect Level
Value
Equilibrium U.S. Environmental
partitioning Protection Agency (2005c)
Low risk of adverse
biological effects

May have adverse
biological effects

Adverse biological
effects expected
(SEM-AVS)//OC
< 130 umol/goc

130 umol/goc < (SEM-
AVS)//OC< 3,000 umol/goc

(SEM-AVS)//OC
> 3,000 umol/goc
Bulk sediment MacDonald et al. (2000)

Ingersolletal. (1996)

Long etal. (1995)

TEC
PEC
ERL
ERM
ERL
ERM
35.8 ug/gdry wt.
128 ug/g dry wt.
55 ug/g dry wt.
99 ug/g dry wt.
46.7 ug/g dry wt.
218 ug/gdry wt.
AVS = Acid volatile sulfide; ERL = Effects range - low (sediment concentration below which adverse effects are
rarely observed or predicted among sensitive species, Long et al. [1995]); ERM = Effects range - median
(sediment concentration above which effects are frequently or always observed or predicted among most species,
Long et al. [1995]); oc = Organic carbon (foc = fraction organic carbon, goc = grams organic carbon); PEC =
Probably effect concentration (sediment concentration above which harmful effects are likely to be observed,
MacDonald et al. [2000]); SEM = Simultaneously extracted metal; TEC = Threshold effect concentration
(sediment concentration below which harmful effects are unlikely to be observed, MacDonald et al. [2000]).
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An alternative approach for developing sediment quality guidelines is to use empirical
correlations between metal concentrations in bulk sediment to associated biological effects,
based on sediment toxicity tests (Table 7-2). These guidelines are based on total metal
concentrations in sediment and do not account for the bioavailability of metals between
sediments.
It should be noted that although EPA is favoring the AVS-SEM approach for regulating
metals in sediments, there is not scientific consensus on this issue. Various studies suggest that
ingestion of sediment particles by benthic organisms is an important exposure route not
accounted for by AVS-SEM (e.g., Lee et al., 2000; Griscom et al., 2002) or that the AVS-SEM
approach may not be the most accurate approach available for predicting non-toxic and toxic
results in laboratory studies.

7.2.2 Distribution of Lead in Aquatic Ecosystems
Speciation of Lead in Aquatic Ecosystems
The speciation of Pb in the aquatic environment is controlled by many factors, such as,
pH, salinity, sorption, and biotransformation processes. Lead is typically present in acidic
aquatic environments as PbSC>4, PbCU, ionic Pb, cationic forms of Pb hydroxide, and ordinary
hydroxide Pb(OH)2. In alkaline, waters common species of Pb include anionic forms of Pb
carbonate Pb(CO3) and hydroxide Pb(OH)2. In freshwaters, Pb typically forms strong complexes
with inorganic OH" and CO32 and weak complexes with Cl (Bodek et al., 1988; Long and
Angino, 1977). The primary form of Pb in freshwaters at low pH (<6.5) is predominantly Pb2+
and less abundant inorganic forms include Pb(HCO)3, Pb(SO4)22~, PbCl, PbCO3, and
Pb2(OH)2CO3. At higher pH (>7.5) Pb forms hydroxide complexes (PbOH+, Pb(OH)2,
Pb(OH)3 , Pb(OH)42 ). Lead speciation in seawater is a function of chloride concentration and
the primary species are PbCl3~ > PbCO3 > PbCl2 > PbCl+ > and Pb(OH)+ (Fernando, 1995).
Lead sorption to suspended or bed sediments or suspended organic matter typically
increases with increasing pH, increasing amounts of iron or manganese; and with the polarity of
particulate matter (e.g., clays). Adsorption decreases with water hardness (Syracuse Research
Corporation [SRC], 1999). At higher pH, Pb precipitates as Pb(OH)+ and PbHCO3+ into bed
sediments (Weber, 1993). Conversely, at low pH, Pb is negatively sorbed (repelled from the
adsorbent surface) (U.S. Environmental Protection Agency, 1979; Gao et al., 2003). In addition,
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Pb may be remobilized from sediment due to a decrease in metal concentration in the solution
phase, complexation with chelating agents (e.g., EDTA), and changing redox conditions
(Gao et al., 2003). Changes in water chemistry (e.g., reduced pH or ionic composition) can
cause sediment Pb to become re-mobilized and potentially bioavailable to aquatic organisms
(Weber, 1993). Methylation may result in Pb's remobilization and reintroduction into the
aqueous environmental compartment and its subsequent release into the atmosphere (SRC,
1999). However, methylation is not a significant environmental pathway controlling Pb fate in
the aquatic environment.

Lead Concentrations in United States Surface Waters
Nationwide, data for Pb in surface waters, from 1991 onward, were compiled using the
United States Geological Survey's (USGS) National Water-Quality Assessment (NAWQA)
database. Data were compiled from locations categorized as "ambient" or "natural." Ambient
refers to data collected from all sampling locations, while natural refers to data collected from
sampling locations categorized as forest, rangeland, or reference. Summary statistics for surface
water, sediment (bulk, <63 jim), and fish tissue (whole body and liver) are summarized in Table
7-3. Overall atmospheric sources of Pb are generally decreasing as regulations have removed Pb
from gasoline and other products (Eisenreich et al., 1986); however, elevated Pb concentrations
remain at sites near ongoing sources, such as near mining wastes or wastewater effluents.
Lead concentrations in lakes and oceans were generally found to be much lower than
those measured in the lotic waters assessed by NAWQA. Surface water concentrations of
dissolved Pb measured in Hall Lake, Washington in 1990 ranged from 2.1 to 1015.3 ng/L
(Balistrieri et al., 1994). Nriagu et al., 1996 found that the average surface water dissolved Pb
concentrations measured in the Great Lakes (Superior, Erie, and Ontario) between 1991 and
1993 were 3.2, 6.0, and 9.9 ng/L, respectively. Pb concentrations ranged from 3.2 to 11 ng/L
across all three lakes. Similarly, 101 surface water total Pb concentrations measured at the
Hawaii Ocean Time-series (HOT) station ALOHA between 1998 and 2002 ranged from 25 to
57 pmol/kg (5 to 11 ng/kg; (Boyle et al., 2005). Based on the fact that Pb is predominately found
in the dissolved form in the open ocean (<90%; Schaule and Patterson, 1981), dissolved Pb
concentrations measured at these locations would likely have been even lower than the total Pb
concentrations reported.
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Table 7-3. Summary of Lead Concentrations in United States Surface Water,
Sediment, and Fish Tissue
Surface Water -
Dissolved
Statistic
n
%ND
Min
Mean
Median
90th %ile
95th %ile
Max
Ambient
3,445
86
0.04
0.66
0.50
0.50
1.10
29.78
Natural
430
88
0.04
0.52
0.50
0.50
0.50
8.40
Sediment -
Bulk, <63 urn
(jig/g dry wt.)
Ambient
1,466
0.48
0.50
120
28
120
200
12,000
Natural
258
1.2
0.50
109
22
66
162
12,000
Fish Tissue (jig/g dry wt.)
Whole
Ambient
332
39
0.08
1.03
0.59
2.27
3.24
22.6
Organism
Natural
93
51
0.08
0.95
0.35
1.40
2.50
22.6
Liver

Ambient Natural
559
71
0.01
0.36
0.15
0.59
1.06
12.7
83
89
0.01
0.28
0.11
0.37
1.26
3.37
%ND = Percentage not detected.
In open waters of the North Atlantic the decline of Pb concentrations has been associated
with the phasing out of leaded gasoline in North America and western Europe (Veron et al.,
1998). Likewise, restrictions reducing Pb in gasoline appear to have been effective in reducing
atmospheric Pb loading to the Okefenokee Swamp in southern Georgia/northern Florida
(Jackson et al., 2004). Based on sediment cores from the Okefenokee Swamp, Pb concentrations
were -0.5 mg/kg prior to industrial development, reached a maximum of-31 mg/kg from about
1935 to 1965, and following passage of the Clean Air Act in 1970 concentrations declined to
about 18 mg/kg in 1990 (Jackson et al., 2004). However, in estuarine systems, it appears that
similar declines following the phase-out of leaded gasoline are not necessarily as rapid. Steding
et al. (2000) used isotopic evidence to demonstrate the continued cycling of Pb in the San
Francisco Bay estuary. In the southern arm of San Francisco Bay, which has an average depth of
<2 m, Steding et al. (2000) found that isotopic compositions were essentially invariant, with 90%
of the Pb derived from 1960s-1970s leaded gasoline. The authors attributed this to the limited
hydraulic flushing and remobilization of Pb from bottom sediments. In the northern arm of San
Francisco Bay, although seasonal and decadal variations in Pb isotope composition were
observed, mass balance calculations indicate that only a small fraction of leaded gasoline fallout
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from the late 1980s had been washed out of the San Joaquin and Sacramento rivers' drainage
basin by 1995 and, consequently, freshwater inputs remain a Pb source to the bay (Steding et al.,
2000). The authors suggest that the continuous source of Pb from the river systems draining into
the bay, coupled with benthic remobilization of Pb, indicates that historic gasoline deposits may
remain in the combined riparian/estuarine system for decades.
In addition to directly measuring Pb concentrations in various aquatic compartments, it is
useful to study the vertical distribution of Pb. Sediment profiling and core dating is a method
used to determine the extent of accumulation of atmospheric Pb and provides information on
potential anthropogenic sources. Sediment concentration profiles are typically coupled with Pb
isotopic analysis. The isotope fingerprinting method utilizes measurements of the abundance of
common Pb isotopes (204Pb, 206Pb, 207Pb, 208Pb) to distinguish between natural Pb over geologic
time and potential anthropogenic sources. Studies of sediment profiles have suggested that
observed increases in Pb concentrations in the upper sediment layer are concomitant with
increases in anthropogenic inputs (Bloom and Crecelius, 1987; Case et al., 1989; Ritson et al.,
1999; Chillrud et al., 2003). Isotopic ratios have been used to link increases in sediment
concentrations with specific anthropogenic sources and to estimate historic records of Pb fluxes
to surface waters and sediments (Flegal et al., 1987, 1989; Blais, 1996; Bindler et al., 1999). For
example, Gallon et al. (2006) collected sediment cores from Canadian Shield headwater lakes
along a 300 km transect extending from a nonferrous metal smelter and used 206Pb/207Pb ratios to
differentiate Pb contributions from smelter emissions relative to Pb contributions from other
anthropogenic inputs. The 206Pb/207Pb ratio for smelter emissions was 0.993, compared to ratios
on the order of 1.15 to 1.22 in aerosols collected at sites remote from point sources in Eastern
Canada and the United States. Based on these isotopic signatures, Gallon et al. (2006) were able
to estimate the amounts of smelter-derived Pb in sediment collected along the 300-km transect.

7.2.3 Species Response/Mode of Action
Lead Uptake
Lead can bioaccumulate in the tissues of aquatic organisms through ingestion of food and
water, and adsorption from water, and can subsequently lead to adverse effects if tissue levels are
sufficiently high (Vink, 2002; Rainbow, 1996). The accumulation of Pb is influenced by pH and
decreasing pH favors bioavailability and bioaccumulation. Bioconcentration factors (BCFs)
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have been reported in the scientific literature for various organisms and range from 840 to 20,000
(aquatic plants), 499 to 3,670 (aquatic invertebrates), and 42 to 45 (fish). Organisms that
bioaccumulate Pb with little excretion must partition the metal such that it has limited
bioavailability, otherwise toxicity will occur if a sufficiently high concentration is reached.

Resistance Mechanisms
Aquatic organisms have various methods to resist the toxic effects of metals such as Pb.
Resistance processes include detoxification and avoidance responses. Mechanisms of resistance
and detoxification vary among aquatic biota. These processes can include translocation,
excretion, chelation, adsorption, and vacuolar storage and deposition. For example, protists and
plants produce intracellular polypeptides that form complexes with Pb (Zenk, 1996; Morelli and
Scarano, 2001). Some macrophytes and wetland plants have developed translocation strategies
for tolerance and detoxification (Knowlton et al., 1983; Deng et al., 2004). Various aquatic
invertebrates may sequester Pb in the exoskeleton (Boisson et al., 2002; Knowlton et al., 1983)
or have developed specialized excretion processes (Vogt and Quinitio, 1994). Fish scales and
mucous may chelate Pb in the water column and potentially reduce Pb uptake (Coello and Khan,
1996).
Avoidance responses are actions performed to evade a perceived threat. Some aquatic
organisms have been shown to be quite adept at avoiding Pb in aquatic systems, while others
seem incapable of detecting its presence. Snails have been shown to be sensitive to Pb, and
avoid it at high concentrations (Lefcort et al., 2004). Conversely, anuran (frog and toad) species
lack an avoidance response up to 1000 jig Pb/L (Steele et al., 1991). Fish avoidance of chemical
toxicants has been well established, and is a dominant sublethal response in polluted waters
(Svecevicius, 2001). However, studies examining avoidance behavior of Pb in fish are lacking.
In addition to the presence of toxic metals, light and pH can also alter preference-avoidance
responses.

Physiological Effects of Lead
Physiological effects of Pb on aquatic biota can occur at the biochemical, cellular and
tissue levels of organization. Lead has been shown to affect brain receptors in fish (Rademacher
et al., 2005) and serum enzyme activity (e.g., EROD and ALAD) in fish and amphibians (Kutlu
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and Susuz, 2004; Blasco and Puppo, 1999; Gill et al., 1991; Vogiatzis and Loumbourdis, 1999).
Studies examining the effects of Pb on fish blood chemistry have indicated alterations from acute
and chronic exposures ranging from 100 to 10,000 |ig/L (Gill et al., 1991; Allen, 1993; Gopal
et al., 1997). Lead exposure has also been shown to negatively affect the growth of aquatic
invertebrates (Arai et al., 2002).

Factors that Modify Organism Response to Lead
There are several factors that may influence organism response to Pb exposure. These
may include the size or age of an organism, genetics, environmental factors (e.g., pH, salinity),
nutrition, and the presence of other contaminants. Lead accumulation in living organisms is
controlled, in part, by metabolic rates (Farkas et al., 2003) and by the physiological conditions of
an organism. Relationships between age, size and Pb body burden in aquatic invertebrates and
fish are variable and depend on many environmental variables (e.g., exposure) (Farkas et al.,
2003). For example, examination of Pb exposure (up to 100 |ig/L) in aquatic invertebrates
showed little relationship between body size and Pb accumulation (MacLean et al., 1996;
Canli and Furness, 1993), whereas Pb accumulation and fish size were found to be positively
correlated (Douben, 1989; Kock et al., 1996).
The genetics of an organism and/or population may alter the response to Pb exposure
through one of two processes: (1) a contaminant may influence selection, by selecting for certain
phenotypes that enable populations to better cope with the chemical, or (2) a contaminant can be
genotoxic, meaning it can produce alterations in nucleic acids at sublethal exposure levels,
resulting in changes in hereditary characteristics or DNA inactivation (Shugart, 1995). Genetic
selection has been observed in aquatic organisms due to Pb tolerance. Because tolerant
individuals have a selective advantage over vulnerable individuals in polluted environments, the
frequency of tolerance genes will increase in exposed populations over time (Beaty et al., 1998).
Several studies have shown that heavy metals can alter population gene pools resulting in
decreased genetic diversity (Duan et al., 2000; Kim et al., 2003). Laboratory studies have shown
that exposure to Pb at 10 mg Pb2+/mL of blood leads to chromosomal aberrations in some aquatic
organisms (Cestari et al., 2004). Lead exposure in water (50 |ig/L) over four weeks resulted in
DNA strand breakage in the freshwater mussel Anodonta grandis (Black et al., 1996). More
recently, Cestari et al. (2004) observed similar results (increase in the frequency of chromosomal
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aberrations and DNA damage in kidney cell cultures) in fish (Hoplias malabaricus) that were fed
Pb contaminated food over 18, 41 and 64 days.
Environmental factors can alter the availability, uptake and toxicity of Pb to aquatic
organisms. Van Hattum et al. (1996) studied the influence of abiotic variables, including
dissolved organic carbon (DOC) on Pb concentrations in freshwater isopods and found that as
DOC concentrations increased, BCFs decreased in P. meridianus and A aquaticus, indicating
that DOC acts to inhibit the availability of Pb to these isopods. Schwartz et al. (2004) collected
natural organic matter (NOM) from several aquatic sites across Canada and investigated the
effects of NOM on Pb toxicity in rainbow trout (Oncorhynchus mykiss). The results showed that
NOM in test water almost always increased LT50 (time to reach 50% mortality), and optically
dark NOM tended to decrease Pb toxicity more than did optically light NOM in rainbow trout.
Studies generally agree that the toxicity of Pb decreases as pH increases (MacDonald et al.,
2002; Home and Dunson, 1995a,b,c). As pH decreases, Pb becomes more soluble and more
readily bioavailable to aquatic organisms (Weber, 1993). Acute and chronic toxicity of Pb
increases with decreasing water hardness, as Pb becomes more soluble and bioavailable to
aquatic organisms (Home and Dunson, 1995c; Borgmann et al., 2005). There is some evidence
that water hardness and pH work together to increase or decrease the toxicity of Pb. High Ca2+
concentrations have been shown to protect against the toxic effects of Pb (Sayer et al., 1989;
Rogers and Wood, 2004; MacDonald et al., 2002; Hassler et al., 2004). Ca2+ affects the
permeability and integrity of cell membranes and intracellular contents (Sayer et al., 1989).
As Ca2+ concentrations decrease, the passive flux of ions (e.g., Pb) and water increases. Finally,
increasing salinity was found to decrease Pb toxicity (Verslycke et al., 2003). The reduction in
toxicity was attributed to increased complexation of Pb2+ with Cl~ ions.
Nutrients (e.g., nitrate, carbonate) have been shown to affect Pb toxicity in some aquatic
organisms. Jampani (1988) looked at the impact of various nutrients (i.e., sodium acetate, citric
acid, sodium carbonate, nitrogen, and phosphates) on reducing growth inhibition in blue-green
algae (Synechococcus aeruginosus) exposed to 200 mg Pb/L. Results indicated that additional
nitrogen, phosphates, and some carbon sources, including sodium acetate, citric acid and sodium
carbonate, all protected the algae from Pb toxicity at 200 mg Pb/L. One hypothesis was that
nutrients were able to reverse toxic effects. The second hypothesis was that nutrients directly
interacted with Pb, in some way sequestering the metal so as to inhibit its metabolic interaction
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with the organism (Rao and Reddy, 1985; Jampani, 1988). Rai and Raizada (1989) investigated
the effects of Pb on nitrate and ammonium uptake and results indicated that Pb exposure can
affect the uptake of some nutrients in N. muscorum. Thus, nutrients seem to be capable of
reducing toxicity, though the mechanisms have not been well established.

Interactions with Other Pollutants
Predicting the response of organisms to mixtures of chemicals is a daunting task
(Norwood et al., 2003). There are two major approaches to predict mixture toxicity including:
(1) examining the combined mode of action of the individual mixture substances; and
(2) determining whether an organism response to the mixture is additive, or some deviation from
additive (synergistic or antagonistic). In addition, researchers may report mixture toxicity in
terms of additive concentrations or additive effects, which can cause confusion in the
interpretation of study results. For the studies presented in this section, the authors primarily
report mixture toxicity in terms of additive concentrations (i.e., the sum of the concentrations of
each individual chemical in the mixture will result in a level of effect similar to the simple sum
of the effects observed if each chemical were applied separately).
When two or more metals compete for the same binding sites or interfere with transport
through cell walls or membranes, the interaction is termed less than strictly additive or
antagonistic. Antagonistic interactions can reduce metal bioavailability when metals are present
in combination, and may lead to reduced potential for toxicity (Hassler et al., 2004). There are a
number of elements (Ca2+, Cd2+, Mg2+, Na+ and Cl~) that act in an antagonistic fashion with Pb
(Niyogi and Wood, 2004; Rogers and Wood, 2003, 2004; Ahern and Morris, 1998; Li et al.,
2004). For example, Pb is a well-known antagonist to Ca2+ (Hassler et al., 2004; Niyogi and
Wood, 2004). Calcium is an essential element, required for a number of physiological processes
in most organisms.
Hassler et al. (2004) reported that in the presence of copper (Cu2+) there was a
significantly higher rate of internalization of Pb in the green algae Chlorella kesserii. It was
suggested that Cu2+ may have affected organism physiology through the disruption of cell
membrane integrity. This would allow increased cation (i.e., Pb2+) permeability and therefore
substantially increased internalization of Pb. Synergism is likely the result of increased
bioavailability of one or more of the metal ions due to the presence of other metals (Hassler
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et al., 2004).Synergistic interactions have also been observed with lead and other metals (Cd, Cu,
Ni, and Zn) (Hagopian-Schlekat et al., 2001).
Norwood et al. (2003) reported, in a review and re-interpretation of published data on the
interactions of metals in binary mixtures (n = 15 studies), that antagonistic (n = 6) and additive
interactions (n = 6) were the most common for Pb. The two most commonly reported
Pb-element interactions are between Pb and calcium and Pb and zinc. Both calcium and zinc are
essential elements in organisms, and the interaction of Pb with these ions can lead to adverse
effects both by increased Pb uptake and by a decrease in Ca and Zn required for normal
metabolic functions.

7.2.4 Exposure/Response of Aquatic Species
Lead exposure may adversely affect organisms at different levels of organization, i.e.,
individual organisms, populations, communities, or ecosystems. Generally, however, there is
insufficient information available for single materials in controlled studies to permit evaluation
of specific impacts on higher levels of organization (beyond the individual organism). Potential
effects at the population level or higher are, of necessity, extrapolated from individual level
studies. Available population, community, or ecosystem level studies are typically conducted at
sites that have been contaminated or adversely affected by multiple stressors (several chemicals
alone or combined with physical or biological stressors). Therefore, the best documented links
between Pb and effects on the environment are with effects on individual organisms.

Effects of Lead on Primary Producers
In the 1986 Lead AQCD (U.S. Environmental Protection Agency, 1986), several authors
reported that some algal species (e.g., Scenedesmus sp.) were found to exhibit physiological
changes when exposed to high Pb or organolead concentrations in situ. The observed changes
included increasing numbers of vacuoles, deformations in cell organelles, and increased autolytic
activity. Increased vacuolization was assumed to be a tolerance mechanism by which Pb was
immobilized within cell vacuoles.
Several studies have been conducted since the 1986 Lead AQCD on the toxicity of Pb to
primary producers (Rai and Raizada, 1989; Jampani, 1988; Adam and Abdel-Basset, 1990; Gaur
et al., 1994; Gupta and Chandra, 1994). Effects to algal growth (Chlorella vulgaris, Closterium
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acerosum, Pediastrum simplex., Scenedesmus quadricauda), ranging from minimal to complete
inhibition, have been reported at Pb concentrations between 100 and 200,000 |ig/L (Bilgrami and
Kumar, 1997; Jampani, 1988). The toxicity of Pb to aquatic plant growth has been studied using
Spirodelapolyrhiza, Azollapinnata, and Lemna gibba (Gaur et al., 1994; Gupta and Chandra,
1994; Miranda and Ilangovan, 1996). Test durations ranged from 4 to 25 days, and test
concentrations ranged between 49.7 and 500,000 |ig/L (Gaur et al., 1994; Miranda and
Ilangovan, 1996). Research on aquatic plants has been focused on the effects of Pb on aquatic
plant growth, chlorophyll and protein content.
Algae and other aquatic plants have a wide range in sensitivity to the effects of Pb in
water. Both groups of primary producers experience ECso values for growth inhibition between
-1,000 and >100,000 |ig/L (Bilgrami and Kumar, 1997; Jampani, 1988; Gaur et al., 1994). The
most sensitive primary producers reported in the literature for effects on growth were Closterium
acersoum and Azollapinnata (Bilgrami and Kumar, 1997; Gaur et al., 1994). The least sensitive
primary producers reported in the literature for effects to growth were Synechococcus
aeruginosus and L. gibba (Jampani, 1988; Miranda and Ilangovan, 1996). Exposure to Pb in
combination with other metals generally inhibits growth less than exposure to Pb alone. Studies
have shown that Pb adversely affects the metabolic processes of nitrate uptake, nitrogen fixation,
ammonium uptake, and carbon fixation (Rai and Raizada, 1989). Lead in combination with
nickel or chromium produced synergistic effects for nitrate uptake, nitrogenase activities,
ammonium uptake, and carbon fixation (Rai and Raizada, 1989).

Effects of Lead on Consumers
The 1986 Lead AQCD (U.S. Environmental Protection Agency, 1986) reported that
hematological and neurological responses are the most commonly reported effects to aquatic
vertebrates. These effects include red blood cell destruction and inhibition of the enzyme
ALAD, required for hemoglobin synthesis. The lowest reported exposure concentration causing
either hematological or neurological effects was 8 jig Pb/L (U.S. Environmental Protection
Agency, 1986).
Recent literature on the toxicity of Pb to fish and aquatic invertebrates has been
summarized by Eisler (2000). Exposure of invertebrates to Pb can lead to adverse effects on
reproduction, growth, survival, and metabolism (Eisler, 2000). Water-borne Pb is highly toxic to
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aquatic organisms, with toxicity varying, depending on the species and life stage tested, duration
of exposure, the form of Pb tested, and water quality characteristics. Among the species tested,
aquatic invertebrates (such as amphipods and water fleas) were the most sensitive to the effects
of Pb with adverse effects being reported at concentrations as low as 0.45 |ig/L (range: 0.45 to
8000 |ig/L). Freshwater fish demonstrated adverse effects at concentrations ranging from 10 to
>5400 |ig/L, generally depending upon water quality parameters (e.g., pH, hardness, salinity).
Amphibians tend to be relatively tolerant of Pb; however, they may exhibit decreased enzyme
activity (e.g., ALAD reduction) and changes in behavior (e.g., hypoxia response behavior)
(Eisler 2000). Lead tends to be more toxic in longer-term exposures, with chronic toxicity
thresholds for reproduction in water fleas ranging as low as 30 jig/L (e.g., Kraak et al., 1994).

7.2.5 Effects of Lead on Natural Aquatic Ecosystems
The effects of Pb on natural aquatic ecosystems were examined for this report following
the conceptual framework developed by the EPA Science Advisory Board (Young and Sanzone,
2002). The essential attributes used to describe ecological condition include landscape
condition, biotic condition, chemical and physical characteristics, ecological processes,
hydrology and geomorphology and natural disturbance regimes. For the biotic condition, the
Science Advisory Board (SAB) framework identifies community extent, community
composition, trophic structure, community dynamics, and physical structure as factors for
assessing ecosystem health. The majority of the published literature pertaining to Pb and natural
aquatic ecosystems focuses on the biotic condition and identifies effects on energy flow or
nutrient cycling, community structure, community level effects, and predator-prey interactions.
Other factors for assessing the biotic condition such as effects of Pb on species, populations, and
organism conditions (e.g., physiological status) were discussed earlier in Sections 7.2.3 and 7.2.4
(see also Annex Sections AX7.2.3 and AX7.2.4).
Recent studies have attributed the presence of Pb to reduced primary productivity,
respiration, and alterations of community structure. Specifically, Pb (6 to 80 mg/L) was found to
reduce primary productivity and increase respiration in an algal community (Jayaraj et al., 1992).
Laboratory microcosm studies have indicated reduced species abundance and diversity in
protozoan communities exposed to 0.02 to 1 mg Pb/L (Fernandez-Leborans and Novillo, 1992,
1994; Fernandez-Leborans and Antonio-Garcia, 1988). Numerous field studies have associated
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the presence or bioaccumulation of Pb with reductions in species abundance, richness, or
diversity, particularly in benthic macroinvertebrate communities (Deacon et al., 2001; Mize and
Deacon, 2002; Mucha et al., 2003; Poulton et al., 1995; Rhea et al., 2004; Maret et al., 2003).
However, in natural aquatic ecosystems, Pb is often found coexisting with other metals and other
stressors. Thus, understanding the effects of Pb in natural systems is challenging given that
observed effects may be due to cumulative toxicity from multiple stressors.
Exposure to Pb in laboratory studies and simulated ecosystems may alter competitive
behaviors of species, predator-prey interactions, and contaminant avoidance behaviors.
Alteration of these interactions may have negative effects on species abundance and community
structure. For example, Pb concentrations ranging from 0.3-1.0 mg/L altered feeding behavior
and affected predator avoidance in mummichogs (Weis and Weis 1998). Lead concentrations
ranging from 0.5-1.0 mg/L altered feeding behavior in fathead minnows (Weber 1996), but did
not elicit an avoidance response in American toads (Steele et al., 1991). The feeding behaviors
of competitive species in some aquatic organisms (e.g., snails and tadpoles) are also influenced
by the presence of Pb (Lefcort et al., 2000).
The effects of Pb have primarily been studied in instances of point source pollution rather
than area-wide atmospheric deposition. Thus, the effects of atmospheric Pb on aquatic
ecological condition remain to be defined. There is a paucity of data in the general literature
that explores the effects of Pb in conjunction with all or several of the various components of
ecological condition as defined by the EPA (Young and Sanzone, 2002). However, numerous
studies are available associating the presence of Pb with effects on biotic conditions.
7.3 CRITICAL LOADS FOR LEAD IN TERRESTRIAL AND
AQUATIC ECOSYSTEMS
This section defines critical loads, describes various concepts and methods that are related
to the estimation of critical loads, and provides a review of the relevant literature on critical
loads.
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7.3.1 Definitions
Critical loads are defined in a variety of ways depending on the chemicals and endpoints
of concern (Paces, 1998; Skeffmgton, 1999; U.S. Environmental Protection Agency, 2004).
For the purposes of this section, critical loads are defined as threshold deposition rates of air
pollutants that current knowledge indicates will not cause long-term adverse effects to ecosystem
structure and function. A critical load is related to an ecosystem's sensitivity to anthropogenic
inputs of a specific chemical. If future inputs of a chemical exceed the critical load for an
ecosystem, the chemical is expected to reach or persist at potentially toxic levels in the future.
A critical load indicates a potential for future impacts only; a current exceedance of a critical
load does not specify whether the current deposition rate of a chemical presents a hazard to the
ecosystem. A current exceedance may or may not indicate a potential hazard; for the
management application in which the critical load concept is applied, the calculation is for
evaluating future impacts.
In order to determine a critical load, the lowest concentration in the receiving medium that
poses a potential hazard to a defined ecosystem must first be determined. This concentration,
known in the critical loads literature as the critical limit (De Vries et al., 2004), is equal to the
effects-based criteria for the most sensitive endpoint in the ecosystem. The critical limit
indicates the current potential for adverse effects to an ecosystem.
In contrast to a critical load, a stand-still load is the highest deposition rate of a chemical
that will not result in future increases of its concentrations in the environmental media,
regardless of the potential for adverse effects at those concentrations. Stand-still loads are also
called "acceptable loads" or critical loads calculated using a "stand-still" approach (De Vries
et al., 2004) and should not be confused with effects-based critical loads.

7.3.2 Historical Perspective
In the 1960s, scientists demonstrated that sulfur emissions on the European continent
were contributing to the acidification of Scandinavian lakes. During the 1970s, evidence
mounted that air pollutants could travel thousands of miles before deposition occurred, implying
that international cooperation was necessary to control acidification. To this end, the European
Community (EC) and 34 governments signed the Convention on Long-range Transboundary of
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Air Pollution (CLRTAP) in 1979 under the auspices of the United Nations Economic
Commission for Europe (United Nations Economic Commission for Europe (UNECE),
2004a; 2004b).
CLRTAP has since been extended to include eight protocols that regulate air pollutants
such as sulfur, nitrogen oxides, heavy metals, persistent organic pollutants, volatile organic
compounds, and ozone. In 1988, CLRTAP adopted the critical-load concept, making it basic to
the future development of international agreements concerning limitation of the emissions of air
pollutants. In 1991, The Coordination Center for Effects (CCE) issued a Technical Report
entitled "Mapping Critical Loads for Europe" which presented the first maps of critical loads that
were produced as part of the work conducted under the UNECE. Each individual country
created maps detailing critical loads and levels of acidity within its boundaries. The maps were
then used by CCE to create a Europe-wide map of critical loads (Hettelingh et al., 1991) that is
used in combination with air emissions and deposition data to guide negotiations between
nations and reduce the gap between critical loads and deposition (Skeffington, 1999). The first
international agreement on pollution control based on critical loads was the second Sulfur
Protocol, which was established in Oslo (United Nations Economic Commission for Europe
(UNECE), 1994) within CLRTAP.
Since 1991, CCE has issued biennial technical status reports on critical loads and critical
thresholds of acidification, eutrophication, sulfur, nitrogen, and nitrogen oxide (Coordination
Center for Effects (CCE), 2005). Progress on data and methodologies is reviewed annually in
CCE Mapping workshops. Recent CCE reports focus on scientific and technical support for the
revision of protocols as well as time horizons for recovery from ecosystem damage.
Many of the signatory governments to CLRTAP have adopted the critical load concept for
determining national emission control polices. Canada has also committed to a critical load
approach for controlling acid deposition. In 1998, federal, provincial, and territorial Energy and
Environment Ministers signed The Canada-wide Acid Rain Strategy for Post-2000. According
to Environment Canada, the primary long-term goal of the Strategy is to achieve critical loads
(or the threshold level) for acidic deposition across Canada (Environment Canada, 2003).
The Ministry of Environment in the Netherlands took the initiative to develop analogous
methods for the calculation of critical loads for heavy metals, methods that would be valid in the
context of CLPTRP (De Vries et al., 2004). Beginning in the mid-1990s, these methods were
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developed through a series of manuals, international workshops, and expert meetings (De Vries
et al., 2004). Participating nations completed a voluntary preliminary critical load mapping
exercise for Pb and cadmium in Europe in 2002 (Hettelingh et al., 2002).
Germany has the National Focal Center (NFC) which coordinates and funds critical load
applications for Germany through a private company (www.okekodata.com). That site presents
preliminary data and applications of critical loads for several heavy metals in Sachsen and
Nordrhein-We stphal en.
The 1986 Pb AQCD (U.S. Environmental Protection Agency, 1986) largely predates the
development of the concept of critical loads and did not discuss this topic. The 2004 AQCD for
Particulate Matter (U.S. Environmental Protection Agency, 2004) includes a brief discussion of
the key elements of the critical loads framework as generally relevant to any air pollutant.
To date, the critical loads framework has not been used for regulatory purposes in the United
States for any chemical.

7.3.3 Application of Critical Loads to Terrestrial and Aquatic Ecosystems
A combinatorial application of critical limit and critical load allows one to assess current
risk while simultaneously estimating future risk from exposure to a chemical (De Vries et al.,
2004). Figure 7-1 shows that four combinations of critical load and limit exceedance or
non-exceedance are possible for a given ecosystem (Figure 1 of De Vries et al. [2004]).
For example, if a current risk is indicated by an exceedance of the critical limit for Pb due to
historical Pb deposition, but current inputs of Pb to the ecosystem are below the critical load
(upper right corner), the critical load model predicts that Pb concentrations will fall below the
critical limit at some point in the future if Pb deposition is maintained at the present level.
If current soil concentrations are below the critical limit (lower left corner), inputs greater than
the critical load will not result in exceedance of the critical limit for some period of time, but
continued exceedance of a critical load will eventually lead to an exceedance of the critical limit.
The time until a critical limit is exceeded (critical time) can also be predicted using the
critical load model (Paces, 1998). This requires knowledge of current concentrations, the critical
load, and predicted deposition rates. Critical times may be useful for setting priorities between
ecosystems with critical load exceedances or between different chemicals.
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No critical load
exceedance
Critical load
exceedance
No critical limit exceedance
No damage at present or foreseen:
Concentration
critical
present
• CL
• PL2
PL1
Tims
•^ Keep the Present .Load
(more stringent than Critical
Load)
Future damage foreseen:
Concentration
critical
present
PL4
Time

"^ Consider Critical Load
(emissions must decrease, even if
concentrations in the ecosystem
are allowed to increase further at
critical load)
Critical limit exceedance
Present damage but recovery in progress:

Concentration

present

critical
•CL
PL2
PL1
TT
Time
"^ Keep the Present Load
(more stringent than Criticnl
I ,oad)
or
•^ Consider Tjyjjci_Load to reach
the critical limit in a defined rime
period (more stringent than
Critical Load)
Present damage, no recovery foreseen:
Concentration

present

critical
PL4
TT
Time
"^ (Consider < jitictl Load
(decrease of concentrations in the
ecosystem down to critical limit
in the long term)
or
^ Consider Target I .oad to reach
the critical limit in a defined time
period (more stringent than
Critical Load)
CL - Critical load; PL - present load (2 cases); SI. - Stand-still load; TL - Target load; TT - Target time
Figure 7-1. The predicted development of metal concentrations in ecosystems for four
cases of exceedance or non-exceedance of critical limits and critical loads of
heavy metals, respectively.

Source: Taken from DeVries et al. (2004).
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7.3.4 Calculation of Critical Loads
This section summarizes the various methods used to calculate critical loads (De Vries
et al., 2001, 2002, 2004; Groenenberg et al., 2002), with an emphasis on the most recent
material.

7.3.4.1 Critical Limits
To determine the critical limit, effects-based criteria for the major ecological endpoints
should be developed for the ecosystem of concern. Criteria may be developed for any receptor
that is exposed to the chemical of concern deposited in the ecosystem. In terrestrial ecosystems,
possible ecological endpoints include effects from direct contact of invertebrates or plants
with soil and ingestion of plants by herbivores. Effects-based criteria for use in defining the
critical limit should be derived from ecotoxicological data appropriate to the most sensitive
endpoint (De Vries et al., 2004). Regardless of the selected endpoint, the critical limit should be
defined as a concentration in the medium that receives the depositional load, typically soil in
terrestrial ecosystems and surface water in aquatic ecosystems. To derive these values, uptake
and/or food- chain modeling may be necessary.
Many critical load calculations rely on ecological effects criteria developed by
government agencies in individual countries (Paces, 1998; De Vries et al., 1998; Van Den Hout
et al., 1999; Skjelkvale et al., 2001). Criteria for Pb vary widely and can be the largest source of
uncertainty in a critical load calculation (Van Den Hout et al., 1999). One reason for the wide
range in estimates of effects criteria is that Pb speciation is often not taken into account. This
can result in variation in estimates of concentration for total Pb that is associated with adverse
effects, since the fraction of Pb available to cause a toxic effect depends on chemical factors such
as the pH or organic matter content (Lofts et al., 2004). To develop effects-based criteria that are
applicable to media with a pH or organic matter content different from the test medium, it is
more appropriate to develop criteria based on the free concentration of Pb rather than the total
concentration of Pb. For example, Figure 7-2 shows the relationship between the critical limit of
Pb in soil as a function of organic matter content and pH.
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Figure 7-2. The relationship between the critical limit of Pb in soil as a function of organic
matter and pH.
Source: Adapted from Lofts et al. (2004)
(http://www.york.ac.uk/depts/eeem/research/projects/criticalloads(stage3)/critlimitsstage3.htm)
7.3.4.2 Models
Critical loads for heavy metals are typically calculated using a steady state model that
ignores internal metal cycling and keeps the calculations as simple as possible (De Vries et al.,
2004). The critical load is equal to the atmospheric input flux, which equals the sum of the
output fluxes from the system minus the other input fluxes (e.g., weathering) when the
concentration of Pb is at the critical limit. The input flux of heavy metals via weathering is
sometimes neglected, because quantitative estimates are highly uncertain, and weathering is
generally thought to be a relatively minor process (De Vries et al., 2004; Scudlark et al., 2005).
More complex methods may be used to calculate critical loads. For example, dynamic
models can be used to model the change of concentrations in soil or water over time (Paces,
1998). These models are most valuable when the time to steady state is very long compared to
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the time of interest. Using these models, the critical load is the deposition rate that leads to
concentrations equal to the critical limit as the model approaches steady state. Fate and transport
models that include explicit modeling of internal cycling and other refinements that may lead to
improved accuracy of the models can be used in place of simple mass-balance models (Doyle
et al., 2003).

Terrestrial Model
If internal cycling and weathering of Pb is neglected and atmospheric deposition is the
only important source of Pb to the system, the critical load in a terrestrial ecosystem is equal to
the sum of the most important fluxes out of the system, leaching, and uptake by harvested plants:
where:
CL(Pb)
Pbu
CL(Pb) = Pbu + Pble(cnt) (7-1)
critical load of Pb (mass per area-year)
metal net uptake in harvestable parts of plants at the critical limit
(mass per area-year)
leaching flux of Pb (dissolved and paniculate) from the soil layer at
the critical limit (mass per area-year)
When applying a mass balance model, it is important to define the boundaries of the
compartment such that all significant fluxes in and out of the compartment can be accounted for.
Uptake of Pb by harvested vegetation may be an important flux out of agricultural soil or
forested soil that is actively logged. In ecosystems that are not harvested, the steady state model
assumes that uptake by plants is balanced by deposition of Pb from decaying vegetation.
The flux out of the system due to uptake in harvested plants (Pbu) is calculated as follows:
Pbu = fPb,u,z * Yha * [Pb]ha
(7-2)
where:
fpb,u,z = fraction of net Pb uptake from soil within the considered layer
(dimensionless)
Yha = annual yield of harvestable biomass (mass per area-year)
[Pb]ha = metal concentration of harvestable parts of plants (Pb per unit mass)
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The net fraction of metal uptake from soil within the considered layer corrects for Pb
measured in harvested vegetation that is taken up via direct deposition onto the plant or from soil
outside of the considered soil layer.
The yield of harvestable biomass should only include the parts of plants that are removed
from the system. Tree leaves, stalks remaining after harvest of agricultural land, roots, and other
parts that remain in the considered terrestrial ecosystem should not be included in the yield.
De Vries et al. (2004) recommends that data for metal content in harvestable biomass
should be taken from unpolluted areas. If the selected endpoint for the critical limit is related to
the concentration in harvested plants rather than a concentration in soil, that critical
concentration should be used in place of actual metal content in harvestable biomass.
The critical leaching flux from the topsoil can be calculated as follows:

Pbcl(crit) = Qle * [Pb]tot,sdw(crit) (7-3)
where:
Qie = flux of drainage water leaching from the considered soil layer
(volume/year)
[Pb]tot,sdw(crit) = critical total concentration of Pb in soil drainage water
(mass per volume)

The total concentration of Pb in soil drainage water is the sum of all species of dissolved
and particulate Pb that leach out of the system in drainage water. De Vries et al. (2004) suggests
that Pb that is sorbed to suspended particulate matter should be neglected so that total Pb is equal
to dissolved Pb, as concentrations of suspended solids are difficult to estimate. Dissolved Pb
may exist as free ions, organic complexes, or inorganic complexes.
The drainage water flux leaching from the topsoil (Qie) can be calculated as follows:
Qle = P-E1-Es-fEt,z*Et (7-4)
where:

P = Precipitation (volume per area-time)
E; = Interception evaporation (volume per area-time)
Es = Soil evaporation within the topsoil (volume per area-time)
f£t,z = Plant transpiration (volume per area-time)
Et = Fraction of water uptake within the topsoil by roots (unitless)
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De Vries et al. (2004) recommends default values for some of these parameters and
provides an alternative calculation method for sites with detailed hydrologic data as part of the
guidance document.

Aquatic Model
If internal cycling and weathering of Pb is neglected and atmospheric deposition is the
only important source of Pb to the system, the critical load in an aquatic ecosystem is equal to
the sum of the most important fluxes out of the system, uptake by harvested plants in the
catchment, sedimentation, and lateral outflow from the catchment:
CL(Pb) = Pbu + Pbsed(cnt) * A! / Ac + PbloC;Cnt (7-5)
where:
CL(Pb) = critical load of Pb (mass per area-year)
Pbu = removal of Pb by harvesting of vegetation in the catchment
(mass per area-time)
PbSed(crit) = removal of Pb by sedimentation at the critical load
(mass per area-time)
Pbioc,crit = lateral Pb outflow from the catchment at the critical load
(mass per area-time)
AI = lake area
Ac = catchment area

It is important to carefully define the boundaries of the aquatic system, so that all inflows
and outflows may be fully accounted for. Current guidance recommends including the entire
watershed within the system, rather than confining the system to a single lake or stream
(De Vries et al., 2004). In stream water, removal of Pb due to sedimentation does not need to be
considered, simplifying the equation to the following:
CL(Pb) = Pbu + Pbloc,cnt (7-6)

De Vries et al. (2004) recommends that critical loads should be calculated for stream
waters only, due to a high level of uncertainty in the rate of removal via sedimentation or other
removal mechanisms within a lake. Critical loads for streams are protective of nearby lakes,
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because the critical loads calculated using this methodology will be lower for streams than
for lakes.
Calculation of removal of Pb by harvesting of vegetation in the catchment is similar to
that in terrestrial ecosystems, with fpt,,u equal to 1, since the entire catchment is now included.
The critical lateral Pb outflow from the catchment is the product of the lateral outflow flux
of water and the total concentration of Pb in the outflow water at the critical limit. The outflow
flux of water is calculated from the outflow divided by the catchment area.

7.3.5 Critical Loads in Terrestrial Ecosystems
Critical loads of Pb have been calculated using simple mass balance, dynamic, and
probabilistic models for forested and agricultural land in Europe and Canada in a handful of
preliminary studies. The methods and model assumptions used to calculate critical loads vary
widely between these studies and little attempt has been made to validate the models that were
used, so it is not known how much various simplifying assumptions affect the results.
Paces (1998) used data from a small agricultural catchment in the Czech Republic that is
typical of agricultural land in that country to calculate critical loads for Pb and other heavy
metals. The critical loads were calculated using a simple dynamic box model. The fluxes into
the system included atmospheric deposition, agricultural inputs, and weathering of bedrock and
the fluxes out of the system included biological uptake and runoff. The model assumed that
inputs of metals to the system are independent of their concentrations in soil but that outputs are
proportional to the concentration of biologically active metal. The author defined biologically
active metal as the concentration of metal in soil that can be extracted in a 2 M nitric acid
solution. This method was used to set a Czech state norm designed to be protective for soil
systems that is used as the critical limit in this study. Using the model, Paces determined that the
critical limit was not presently exceeded, but that the critical load is exceeded. However, the
critical time was almost 1,000 years. Therefore, the model predicts that Pb will continue to
accumulate in Czech agricultural soil and will eventually pose a potential risk if current inputs
continue. The author identified the simplifying assumptions used to calculate fluxes out of the
system as the major source of uncertainty.
Van den Hout et al. (1999) calculated critical loads for Pb and other pollutants in the
organic and mineral soil layers of forested ecosystems. Atmospheric deposition was assumed to
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be the only inflow, and outflows from soil were assumed to occur due to biological uptake and
leaching. Net heavy metal uptake by the forest was set equal to the rate of water uptake by
vegetation multiplied by the water concentration and a "preference factor" that indicates the
preference of the vegetation for the metal relative to water. Water flux was estimated from
precipitation, soil evaporation, and transpiration data. An equilibrium speciation model that
takes inorganic and organic ligands into account was used to estimate dissolved concentrations
of Pb in leachate. Results were strongly dependant on the critical limits that were chosen. Using
the most stringent levels, critical loads were exceeded over much of Europe. The time to steady
state was estimated to be hundreds of years. Speciation of Pb was identified as an important
source of uncertainty.
Reinds et al. (2002) used the guidance prepared by De Vries et al. (2002b) to calculate
critical loads in the mineral topsoil of forested and agricultural ecosystems across 80,000 acres of
the European continent. The median critical load for Pb in Europe was 25 g ha'1 year"1 using
this methodology. The drainage water flux leaching from the topsoil was the dominant term in
the model, so critical loads followed the spatial pattern of net runoff (excess precipitation)
across Europe.
Probst et al. (2003) calculated critical loads for Pb for forested sites in France.
Weathering rates were determined using a model for representative French soil samples. The
biomass uptake of Pb was derived using National Forestry Inventory data for the average annual
biomass growth and data for the Pb content in biomass. An uptake factor scaled down to the
considered depth was applied. Leaching of Pb was calculated using runoff data and dissolved Pb
concentrations in soil solution. Critical loads at the French site varied over a wide range (4.9 to
133 g ha- year"1). Critical loads were controlled mainly by net runoff. Weathering rates were
small compared to leaching and biomass uptake rates.
Doyle et al. (2003) used a probabilistic assessment to calculate critical loads in terrestrial
and aquatic (see following section) ecosystems on the Canadian Shield. The terrestrial model
used an analytical solution to the convection/dispersion equation. The model only considered
soluble metal in the flux to soil and assumed that the insoluble fraction was not available. Metals
were assumed to be sorbed onto immobile soil solids according to an equilibrium distribution
(Kd) relationship. The input parameters were selected to represent boreal forest and Canadian
Shield conditions. Best estimate inputs were used for deterministic evaluation and distributions
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of values were used in a probabilistic assessment. The model inputs included net water flux,
effective water velocity, moisture content of soil, pH, dispersion coefficient, and Kd. The
25th percentile critical loads (47 mg/m3 per year for Pb) were compared to current deposition
rates to evaluate risk.
In spite of the variation in methods and model assumptions used to calculate critical loads
for Pb in the studies discussed above, some general conclusions may be drawn. The critical limit
is the most important value for determining the value of the critical load. Wide variations in
available effects levels makes this parameter one of the most important sources of uncertainty
when calculating critical loads in terrestrial ecosystems. Spatial variations in critical loads for Pb
are largely controlled by net runoff. Weathering and uptake by harvestable vegetation were less
important. The time to reach steady state is several hundred years in the two studies that used
dynamic models to determine critical loads.

7.3.6 Critical Loads in Aquatic Ecosystems
Doyle et al. (2003) modeled critical loads in surface water bodies assuming complete
mixing with dilution water entering from the terrestrial catchment area. Loss of metal was also
assumed to occur through downstream flushing and burial in sediment. Transfer of metal to
sediment was modeled as a first-order process dependant on the dissolved concentration and pH.
The inputs to the model included the following: water body area, terrestrial catchment area,
water body depth, sediment accumulation rate, thickness of biologically active sediment, net
precipitation, and water pH. The fist-order rate constant for transfer to sediment was correlated
with pH. The model reached steady state within a few years. Transfer of Pb from the terrestrial
catchment to the water body was neglected, because the time to steady state could be on the
order of 10,000 years if the model included this source of Pb. However, the authors cited a
separate calculation that indicated that neglect of transfer of Pb from the catchment may lead to a
5-fold underestimation of Pb concentrations in the surface water.
These results indicate that Pb run-off from soil is more important than direct atmospheric
deposition to the surface water bodies considered in this study. Due to the long times required to
achieve steady state, the critical load methodology may not be appropriate for Pb in aquatic
systems.
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7.3.7 Limitations and Uncertainties
The largest sources of uncertainty identified in studies of critical loads for Pb include the
following:
• Steady-state assumption
• Derivation of the critical limit
• Lead speciation
• Soil runoff as an input to aquatic ecosystems
The critical load is calculated for steady state conditions, but the time for Pb to reach
steady-state concentrations can be as long as several centuries. Thus, dynamic models are often
used to predict Pb concentrations over shorter time frames. Dynamic modeling requires
additional knowledge about current concentrations in the considered ecosystem. For regulatory
purposes, use of dynamic modeling requires that a target time be set in order to calculate a
critical load.
Criteria for the protection of soil and for the protection of aquatic organisms vary over a
wide range from country to country. Use of the critical loads method for international
negotiations will require implementation of a consistent calculation methodology that takes into
account the effect of Pb speciation on toxicity over a range of soil types and chemical conditions.
Speciation strongly influences the toxicity of Pb in soil and water and partitioning
between dissolved and solid phases determines the concentration of Pb in soil drainage water,
but it has not been taken into account in most of the critical load calculations for Pb performed to
date. Recent guidance for heavy metals has begun to emphasize the importance of speciation to
critical load calculations and suggest methods to calculate speciation (De Vries et al., 2004).
To this end, Lofts et al. (2004) developed critical limit functions for several metals, including Pb,
that take into account the effects of pH, organic matter, and the protective effects of cations on
speciation.
Runoff of Pb from soil may be the major source of Pb into aquatic systems. However,
little attempt has been made to include this source into critical load calculations for aquatic
systems due to the complexity of including this source in the critical load models.
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7.3.8 Conclusions
Preliminary efforts to calculate critical loads for Pb in terrestrial and aquatic ecosystems
have so far relied on a variety of calculation methods and model assumptions. Efforts are
ongoing to refine and standardize methods for the calculation of critical loads for heavy metals
which are valid in the context of CLPTRP. At this time, the methods and models commonly
used for the calculation of critical loads have not been validated for Pb. Many of the methods
neglect the speciation of Pb when estimating critical limits, the uptake of Pb into plants, and the
outflux of Pb in drainage water, limiting the utility of current models.
Future efforts should focus on fully incorporating the role of Pb speciation into critical
load models, and validating the assumptions used by the models.
7-46
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8. INTEGRATIVE SYNTHESIS: MULTIMEDIA LEAD
EXPOSURE, HUMAN HEALTH EFFECTS, AND
ECOSYSTEM EFFECTS
8.1 INTRODUCTION
This integrative synthesis is structured to provide a coherent framework to support the
assessment of multimedia exposures, human health risks, and ecological effects associated with
ambient airborne lead (Pb) in the United States. The main goal of the chapter is to integrate
newly available scientific information with key findings and conclusions from the 1986 Air
Quality Criteria Document (Lead AQCD) and its associated Addendum (U.S. Environmental
Protection Agency, 1986a,b), and their 1990 Supplement (U.S. Environmental Protection
Agency, 1990), to address issues central to the EPA's assessment of evidence needed to support
the current ongoing periodic review of the Pb NAAQS. The integrated assessment of key
findings and conclusions provided here and elsewhere in this document provides key inputs to
further analyses of such findings and their policy implications as delineated in a Lead Staff Paper
prepared by EPA's Office of Air Quality Planning and Standards (OAQPS). The analyses
provided in that Staff Paper aim to "bridge the gap" between scientific assessments in this
criteria document and judgments required of the EPA administrator in evaluating whether to
retain or, possibly, to revise the current primary and/or secondary Pb NAAQS.

8.1.1 Historical Background
In 1971, U.S. EPA promulgated national ambient air quality standards for several major
"criteria" pollutants (see Federal Register, 1971), but did not include Pb among them at that time.
Later, on October 5, 1978, the EPA promulgated primary and secondary Pb NAAQS under
Section 109 of the CAA (43 FR 46258), as announced in the Federal Register (1979). Identical
primary and secondary Pb standards were then established as: 1.5 |ig/m3 as a calendar quarterly
average (maximum arithmetic mean averaged over 90 days). Those standards were based on
scientific assessments in EPA's original Air Quality Criteria for Lead (U.S. Environmental
Protection Agency, 1977) or "1977 Lead AQCD."
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In 1986, the EPA published a revised Lead AQCD (U.S. Environmental Protection
Agency, 1986a), which assessed newly available scientific information on health and welfare
effects associated with exposure to various concentrations of Pb in ambient air, based on
literature published through 1985. That 1986 document mainly assessed the health and welfare
effects of Pb, but other scientific data were also discussed in order to provide a better
understanding of the pollutant in the environment. Thus, the 1986 Lead AQCD included
chapters that discussed the atmospheric chemistry and physics of the pollutant; analytical
approaches; environmental concentrations; human exposure and dosimetry; physiological,
lexicological, clinical, and epidemiological aspects of Pb health effects; and Pb effects
on ecosystems. An Addendum to the 1986 Lead AQCD was also published along with it
(U.S. Environmental Protection Agency, 1986b). Then, a Supplement to the 1986 Lead
AQCD/Addendum was published by EPA in 1990 (U.S. Environmental Protection Agency,
1990a). That 1990 Supplement evaluated still newer information emerging in the published
literature concerning (a) Pb effects on blood pressure and other cardiovascular endpoints and
(b) the effects of Pb exposure during pregnancy or during the early postnatal period on birth
outcomes and/or on the neonatal physical and neuropsychological development of infants and
children.
The 1986 Lead AQCD/Addendum and the 1990 Supplement provided scientific inputs to
support decision-making regarding CAA-mandated periodic review and, as appropriate, revision
of the Pb NAAQS; and they were drawn upon in preparation of an associated OAQPS Lead Staff
Paper (U.S. Environmental Protection Agency, 1990b). Based on scientific assessments in the
1986 Lead AQCD/Addendum and the 1990 Supplement, as well as associated exposure/risk
analyses, the 1990 Staff Paper recommended that the EPA Administrator consider a range of
standards for the primary Pb NAAQS of 0.5 to 1.5 |ig/m3 (30-day arithmetic mean). After
considering those evaluations, EPA chose not to propose revision of the Pb NAAQS. At the
time, as part of implementing a broad 1991 U.S. EPA Strategy for Reducing Lead Exposures
(U.S. Environmental Protection Agency, 1991), the Agency focused primarily on regulatory and
remedial clean-up efforts aimed at reducing Pb exposures from a variety of non-air sources
judged to pose more extensive public health risks to U.S. populations, as well as on other actions
to reduce Pb emissions to air. By 1990, average ambient air Pb levels had dropped to 0.15 to
0.25 |ig/m3 across U.S. urban areas due to the phasedown of Pb in gasoline.
8-2
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8.1.2 Chapter Organization
The ensuing chapter sections collectively address the following topics: (1) ambient
airborne lead compounds, sources, emissions, and air quality; (2) ambient Pb exposures
pathways and dosimetric considerations; (3) epidemiologic and toxicologic evidence for
associations between Pb exposure of human populations and various health effects,
demonstrating a broad array of pathophysiologic responses of humans and animals to acute and
chronic Pb exposures; (4) characterization of applicable dose-response relationships for various
types of Pb-exposure effects; (5) persistence of key Pb exposure effects; (6) factors that enhance
or lessen susceptibility or vulnerability to Pb health effects; (7) identification of susceptible and
vulnerable human population groups likely at increased risk for Pb-related health effects;
(8) potential public health implications of low-level Pb exposures; and (9) delineation of
ecological effects of Pb.
8.2 OVERVIEW OF MULTIMEDIA LEAD, SOURCES, EMISSIONS,
AND CONCENTRATIONS IN THE UNITED STATES
Lead has been observed in measurable quantities in nearly every environmental medium
all over the world. Human exposure to Pb occurs through several routes, as shown in Figure 8-1,
which provides a simplified diagram of various routes of exposure through different
environmental media, with a main focus on the ambient air. The multimedia aspects of Pb
exposure can be seen in that Pb emissions to the air contribute to Pb concentrations in water, soil,
and dusts; Pb in soil and dust also can make important contributions to Pb concentrations in
ambient air. The relative contributions of Pb from different media and different sources on
human exposure depend on factors such as the proximity of major sources to the residence and
workplace of the individual, the condition of the residence (especially the presence and condition
of lead-based paint) and whether the residence is in an urban, suburban, or rural location. This
section briefly summarizes available evidence concerning multimedia Pb sources and exposure
pathways, with main emphasis on pathways involving airborne Pb components.
-------
SURFACE AND
GROUND WATER
DRINKING
WATER
Figure 8-1. Principal pathways of lead from the environment to human consumption.
Heavy arrows are those pathways discussed in greatest detail in this chapter.
8.2.1 Sources of Lead Emissions into Ambient Air
In ambient air, Pb occurs mainly as a component of organometallic compounds and
various salts or other compounds (as summarized in Chapter 2, Section 2.1) rather than as
elemental Pb because, at ambient atmospheric temperatures, elemental Pb deposits to surfaces or
forms a component of atmospheric aerosol. Those salts and covalently bound Pb compounds
that are of significance in the environment include: sulfates (PbSO4); chlorides (PbCb);
carbonates (PbCO3, Pb(HCO3) 2); hydroxides (Pb(OH) 2); nitrates (Pb(NO3) 2); phosphates
(PbPO4, Pb(HPO4) 2); oxides (PbO, Pb3O4), silicates, and PbS. With the exception of the
covalently-bound sulfide and oxide, these compounds are derived from acids (or the related
anions) that are common in the environment, such as sulfuric acid (H2SO4), nitric acid (HNO3),
carbonic acid (H2CO3, an acid that forms when CC>2 dissolves in water), and phosphoric acid
8-4
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(H3PO4). Lead salts, once formed, tend to be only slightly soluble in neutral solutions, but are
quite soluble in the presence of acid. Another form of Pb-containing compounds is the
tetravalent Pb (IV) organometallic compounds, such as the well-known fuel additives,
tetramethyllead (TML) and tetraethyllead (TEL).
Natural sources of Pb emissions to the air include volcanoes, sea-salt spray, biogenic
sources, forest fires, and wind-blown soil. There is significant variability in Pb emissions from
these sources, but it has been estimated that they contribute 10 to 20 thousand tons per year in
annual emissions of Pb, worldwide (see Chapter 2, Section 2.2.1). In addition to the typically
relatively limited inputs of natural sources to ambient air, Pb emitted into the air from a wide
variety of anthropogenic sources can contribute to human exposure via a number of often
interlinking multimedia exposure pathways, as illustrated in Figure 8-1 and discussed below.
Historically, mobile sources constituted a major source of Pb emissions into the ambient
air, due to the use of leaded gasoline (Section 2.2.4). Although its phase down began in 1974,
some Pb was still added to gasoline in the United States as an anti-knock additive at the time of
1986 Lead AQCD/Addendum. Accordingly, airborne Pb concentrations nationwide have fallen
dramatically over the past 20 years; and this represents one of the most important public and
environmental health successes in history. Remaining mobile source-related emissions of Pb
include brake wear, resuspended road dust, and emissions from vehicles that continue to use
leaded gasoline (e.g., some types of race cars and aircraft).
The dramatic decreases in Pb emissions to U.S. ambient air during recent decades,
including the notable decreases in Pb emissions from mobile sources, are shown in Figure 8-2.
Nationwide, ambient air Pb emissions fell 98% between 1970 and 2002 (U.S. Environmental
Protection Agency, 2003), primarily due to elimination of alkyl lead additives to automotive
gasoline. The decreasing contributions of mobile sources to ambient airborne Pb have been
documented by National Emissions Inventory data for Pb emissions from various sources for the
United States in 1990 and 2002 (see Table 2-8). In 1990, mobile sources still constituted the
largest single source of U.S. Pb emissions, even though substantial reductions in airborne Pb had
already occurred due to the phasedown of Pb in gasoline. However, the emissions inventory data
from 2002 show that, while mobile sources continue to make some contributions to Pb
emissions, industrial sources now play a much more significant proportional role (as can be seen
in Table 2-8).
-------
80,000
60,000
V)
c
o
t: 40,000
o
(0
20,000
in 1985, EPA refined its methods for estimating emissions
Transportation
| Industrial Processes
\ Fuel Combustion
1982-02: 93% decrease
1993-02: 5% decrease
85
92 93 94 95 96 97 98 99 00 01 02
Year
Figure 8-2. Trends in U.S. air lead emissions during the 1982 to 2002 period.
Source: U.S. Environmental Protection Agency (2003).

As discussed in Section 2.2.2, the largest Pb emitters into the ambient air are now in the
manufacturing sector, which includes combustion sources (such as industrial or utility boilers
and municipal or hazardous waste incinerators), iron and steel foundries, primary and secondary
smelters, and other, mainly, stationary sources of Pb emissions to the air. Other stationary
sources of airborne Pb emissions include smelters for other metals, such as copper or nickel,
Pb-acid battery manufacturing, cement manufacturing and mining or processing of Pb.
One observation that can be drawn from the data on trends in Pb emissions is that
currently the occurrence of airborne Pb concentrations in the United States is influenced heavily
by localized industrial or other stationary sources of Pb, in contrast to the situation a few decades
ago when elevated U.S. ambient air Pb concentrations were widespread mainly as a result of
leaded fuel use.
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8.2.2 Transport and Secondary Dispersal of Atmospheric Lead
Lead can be transported in the atmosphere and undergo secondary dispersal via the
deposition and resuspension of particles containing Pb, as discussed in Section 2.3.2 of
Chapter 2. As discussed in Section 2.3.1, numerous studies have analyzed Pb concentrations in
media such as soil, sediments, ocean water, peat bogs, plants, snowpacks, or ice cores to evaluate
the historical record of deposition of Pb. Sediments can provide records dating back several
million years, peat bogs can reach back to the late glacial period (-15000 years ago), corals and
trees can record up to several hundred years, and lichens and mosses can provide recent
deposition data (Weiss et al., 1999). Many studies have shown a pattern of sediment Pb
concentrations increasing to reach peak concentrations in layers representing deposition during
the 1970's followed by marked declines in more recent years. For example, Figure 8-3 presents
data on Pb concentrations in sediment samples from 12 lakes in the Great Lakes area (Yohn
et al., 2004).

Deposition of Airborne Lead
Dry deposition is the process by which pollutants are removed from the atmosphere in the
absence of precipitation. The size of depositing particles is arguably the most important factor
affecting dry deposition rates. For very small particles, Brownian motion is the dominant
mechanism that transports particles through the viscous sublayer that borders surfaces. For large
particles, sedimentation is the most important process governing particle deposition.
For intermediate particles, impaction and interception largely determine deposition rates.
The highest extent of uncertainty applies to deposition velocities for the intermediate sized
particles. As an example, in one study, although most of the airborne Pb mass was associated
with submicron particles, only about 0.5% of the Pb particle mass undergoing dry deposition in
Chicago was <2.5 |im in diameter. Also, more than 90% of Pb particle mass that undergoes dry
deposition is in an insoluble chemical form. Overall, dry deposition velocities for Pb are in the
range of 0.05 to 1.3 cm/s and dry deposition flux rates have been estimated to be in the range of
-1-2 mg/m2-year.
Wet deposition is the process by which airborne pollutants are scavenged by precipitation
and removed from the atmosphere. The size of particles can also influence wet deposition rates.
Large particles are scavenged more efficiently. Lead, which is found in particles primarily in
5-7
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Elk
Crystal
Gratiot
Cadillac
Cass
Gu
B
1815
0,00 0,20 0,40 0,60 0,80 1,00 1,20
concentrations normalized to

Figure 8-3. Lead concentrations in sediment samples in 12 Michigan lakes. The
concentrations are normalized by the peak Pb concentration in each lake;
peak Pb concentrations ranged from approximately 50 to 300 mg/kg.

Source: Yohn et al. (2004).
the submicron size range, does not undergo wet deposition as easily as many of the crustal

elements. Wet deposition flux has been estimated to range from about 300-1000 |ig/m2-year

in U.S. locations, as discussed in Section 2.3.2.
Resuspension of Lead in Soil

The resuspension of soil-bound Pb particles and contaminated road dust is a significant

source of airborne Pb. The main sources of resuspension are typically wind and vehicular traffic,

although resuspension through other mechanical processes, e.g., construction, pedestrian traffic,
-------
agricultural operations, and even raindrop impaction, is possible. In general, mechanical stresses
are more effective than the wind in resuspending particles.
Understanding the physics of resuspension from natural winds requires analyzing the
wind stresses on individual particles, including frictional drag, form drag, gravitation, and the
Bernoulli effect. Although this analysis can be accurate on a small scale, predicting resuspension
on a large scale generally focuses on empirical data for continual soil movement due to three
processes: saltation, surface creep, and suspension. Saltation is the process by which particles in
the 100 to 500 jim size range bounce or jump close to the surface. The low angle at which these
particles strike the surface transfers momentum to smaller particles, allowing them to be
suspended into the atmosphere. Depending on soil conditions, saltation can be responsible for
moving 50 to 75% of surface particles. Surface creep is the rolling or sliding motion of particles
induced by wind stress or momentum exchanged from other moving particles. This generally
applies to large particles 500 to 1000 jim in diameter and moves 5 to 25% of soil by weight.
Suspension is the process that actually ejects particles into the air. This affects particles
< 100 jim in diameter and moves 3 to 40% of soil by weight. Resuspension may occur as a series
of events. Short episodes of high windspeeds, dry conditions, and other factors conducive to
resuspension may dominate annual averages of upward flux.
Soil-Pb concentrations vary significantly throughout urban areas, depending on proximity
to roadways and stationary sources and on wind speed and direction, as noted in Section 3.2.1.
Some of the highest soil-Pb concentrations are observed near major roadways. For example,
surface soil-Pb concentrations measured near a major freeway in Cincinnati, OH, were between
59 ppm and 1980 ppm, levels well above background. These concentrations dropped off
dramatically with soil depth. An estimated 40% of Pb from automobile exhaust was retained in
the nearby soil. Lead-contaminated soils and dusts can be significant sources of Pb exposure for
human populations.
Lead in soil is also highly elevated near stationary sources of Pb emissions. In particular,
areas around smelters and battery disposal sites can have very high levels of soil Pb
(Section 3.2.2). Concentrations of soil Pb are highly elevated near mines as well. Lead and zinc
mines, in particular, typically have large deposits of Pb in nearby soil, but mines used for
extracting other metals can also have Pb-contaminated soil. Blood-Pb levels are typically
elevated in people living near Pb mines.
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The resuspension of soil particles historically contaminated by past deposition of airborne
Pb emitted from smelters and other stationary sources, as well as resulting from past combustion
of leaded gasoline, represents a continuing source of current air Pb.

8.2.3 Ambient Air Lead Concentrations
There are four ambient monitoring networks that measure Pb concentrations in the United
States, as discussed in Section 3.2.1 of Chapter 3. Determination of compliance with the current
Pb NAAQS is based on measurements taken at Federal Reference Method (FRM) monitors,
which measure Pb in total suspended particulate matter (TSP), i.e., particles up to about 30 jim in
diameter. In 2005, there were about 250 FRM sampling sites in operation across the
United States.
Data on airborne Pb concentrations are also available from two other U.S. networks that
measure Pb in fine particulate matter (<2.5 jim in diameter). There are -200 sites, primarily in
U.S. urban locations, in the PM2.5 speciation network; and there are over 100 sites in the
Interagency Monitoring of Protected Visual Environments (IMPROVE) network that are located
in U.S. national parks or wilderness areas. In addition, Pb concentrations are measured in PMio
samples collected at the National Air Toxics Trends Stations (NATTS) network of 24 U.S. sites.
As was seen for emissions of Pb, ambient air Pb concentrations have also markedly
declined over the past several decades. Between 1983 and 2002, ambient air Pb concentrations
measured at FRM monitors decreased by -94%, as shown in Figure 8-4. Data from the FRM
monitors and from the PM2.5 speciation, IMPROVE and NATTS networks all show a consistent
pattern of ambient air Pb concentrations, i.e., a long period of measured ambient levels
substantially lower than the current Pb NAAQS, except in a few local areas. For example, Pb
concentrations measured at the FRM monitors in 2000 to 2004 are quite low, on average, with
the mean level ranging from 0.03 to 0.05 |ig/m3 (excluding point source-related monitors) and
0.10 to 0.22 (including point source-related monitors). However, when data from point source-
oriented monitors are included, one to five U.S. locations (from among -200 sites) had measured
calendar quarterly maximum Pb levels that exceeded the NAAQS level (1.5 |ig/m3, quarterly
max average) in any given year during 2000 to 2004. As for data from PMio monitors in the
NATTS network, the highest quarterly max Pb concentration observed was 0.039 |ig/m3 during
8-10
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c
o
0)
o
c
o
O
1.6
1.4
1.2
1.0
0.8
0.6
0,4
0.2
0.0
NAAQS

- 90% of sites have concentrations below this line
1983-02: 94% decrease
1993-02: 57% decrease
10% of sites have concentrations below this line
83 84 85 86 87 88 89 90 91 92 93 94 95 96 97 98 99 00 01 02
Year
Figure 8-4. Airborne Pb concentrations measured at FRM sites, averaged across the
United States for the years 1983 through 2002. The data are plotted in terms
of maximum arithmetic mean averaged over a calendar quarter and are
shown in relation to the Pb NAAQS of 1.5 ug/m3
Source: U.S. Environmental Protection Agency (2003).
2002 to 2005. Using data from the PM2.5 speciation network for 2002 to 2005, the highest
quarterly max Pb concentration reported was 0.168 |ig/m3.
Descriptive statistics for Pb concentrations determined from several particle size fractions
are presented in Table 8-1. Focusing on the Pb concentrations reported from TSP, PMio and
PM2.5 samples in urban areas (i.e., not from the IMPROVE network), it can be seen that the
mean and median values are not markedly different, though in general PM2.5 mass is about 50%
of the mass of PMi0, which is then about 50% of the mass of TSP depending on the given area.
As summarized in section 3.1.1, recent studies suggest that Pb is somewhat more likely to be
found in fine fraction particles than in larger particle sizes. Overall, ambient air Pb
concentrations in the United States are generally well below the current NAAQS level, except
for a few scattered locations influenced by local sources.
8-11
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Table 8-1. Descriptive Statistics for Lead Measurements (in ug/m3) from Monitors*
Using Different Size Fractions of PM for Recent Years**
Particle size (network)
TSP*** (FRM; n~200)
XSP*** (FRM, n~200) excluding
point source sites
PM10 (NATTS, n=26)
PM2 5 (Speciation, n=272)
PM25 (IMPROVE, n=167)
Minimum
0.00
0.00
0.0027
0.002
0.0005
Mean
0.01-0.22
0.03-0.05
0.0116
0.008
0.0016
Median
0.02-0.04
0.01-0.02
0.0101
0.005
0.0013
Maximum
1.92-9.13
0.26-1.75
0.039
0.168
0.0065
* Excluding monitors representative of point source emissions.
* * 2000-2004 for data from IMPROVE and TSP; 2002-2005 for data from the PM2 5 speciation network
and NATTS.
*** Data for TSP presented as range of values for each year.
8.2.4 Non-air Environmental Lead Exposure Routes
In addition to ambient air, major non-air environmental routes for exposure to Pb include:
Pb in house dust; Pb-based paint in older homes; drinking water; and Pb-contaminated food.
Lead exposure can also occur at times due to other idiosyncratic sources such as calcium
supplements, Pb-based glazes, certain kinds of miniblinds, hair dye, and other consumer products
that can widely vary in their prevalence and the potential risk posed by them.
Given the large amount of time people spend indoors, exposure to Pb in dusts and indoor
air can be significant (see Section 3.2.3). For children, dust ingested via hand-to-mouth activity
may be a more important source of Pb exposure than inhalation. However, dust can be
resuspended through household activities, thereby posing an inhalation risk as well. A number
of different sources can contribute to Pb in housedust, both from sources outside the home and
from Pb-based paint.
Throughout early childhood, floor dust Pb contamination is a source of exposure. Lead-
contaminated windowsill dust becomes an additional source of Pb intake during the second year
of life when children stand upright. Because of normal mouthing behaviors and increased
mobility, the highest blood Pb levels are seen in children between 18 and 36 months of age. This
typically is observed after a rapid rise in blood Pb levels between 6 and 12 months. Even at low
concentrations, Pb in housedust can have a notable effect on children's blood Pb levels.
8-12
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For example, studies discussed in Section 3.2.3 show that, at a median floor dust Pb level of
5 |ig/ft2 (54 |ig/m2), -5% of children had blood Pb levels > 10 ng/dL. At a floor dust Pb loading
of 50 |ig/ft2 (540 jig/m2), the percentage of children with blood Pb levels > 10 |ig/dL rose to 20%.
In another study, children exposed to floor dust Pb loadings in excess of 25 |ig/ft2 (270 |ig/m2)
were at eight times greater risk of having blood Pb levels > 10 |ig/dL compared to children
exposed to levels below 2.5 |ig/ft2 (27 jig/m2).
Soil Pb is a significant contributor to elevated blood-Pb levels, especially among children,
in populations residing near certain Superfund sites, as discussed in Section 3.2.2. For example,
Pb levels in soil collected at residences near the Tar Creek Superfund Site (a Pb mining area in
northeastern Oklahoma) reflected contamination by wind-dispersed mine wastes. More than
20% of residential soil samples exceeded the EPA action level of 500 ppm, and children's blood
Pb levels tended to be higher in comparison to those of children living outside the Superfund
towns. In this same area, blood-Pb levels were found to be highest among African-American,
Mexican-American, and poor children. Blood-Pb levels were most commonly correlated with
mean floor dust Pb loading and with soil Pb, especially front yard soil. Another study found that
homes at the Jasper County Superfund Site in southwestern Missouri had significantly higher
soil and dust Pb levels and significantly higher blood-Pb levels than areas outside of the
Superfund site. There was a strong statistical relationship observed between blood-Pb levels and
soil, dust, and paint Pb concentrations.
Lead-based paint was the most widely used, dominant form of house paint for many
decades, and a significant percentage of homes (especially those built before 1978) still contain
Pb-based paint on some surfaces, as discussed in Section 3.5.1. As Pb-based paint degrades, it
becomes incorporated into house dust, as noted earlier in this chapter. Lead-based paint poses a
potential exposure risk due to ingestion of Pb-contaminated dusts via normal hand-to-mouth
activities and/or pica (which are common in children) or due to inhalation during renovation or
demolition projects. Lead-based paint can pose a particularly serious inhalation risk for both
adults and children during renovation activities that form easily inhaled Pb particles. The
ingestion and/or inhalation of Pb derived from Pb-based paint has long been one of the most
common causes of clinical Pb toxicity in the United States.
As discussed in Section 3.3, most U.S. drinking water distribution systems serving more
than 3,000 people typically supply drinking water that meets the EPA tap water limit of
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0.015mg/L(15ppb)setin 1991. Of 18 major U.S. cities illustrated in Table 3-11 of Chapter 3
as exceeding the EPA water Pb Action Level in the early 1990s, 14 had decreased their 90th
percentile tap water concentrations to below the 15-ppb Action Level during recent monitoring
periods (since the year 2000). On the other hand, with the introduction of chloramine as an
alternative water treatment used in some cities across the United States, some increases in tap
water-Pb concentrations have been detected in some municipal water supplies, raising concern
about possible resultant increases in blood-Pb levels among affected water-use populations.
Very little Pb in drinking water comes from public water utility supplies per se, versus
from leaded solder in plumbing or Pb in plumbing fixtures in buildings attached to utility
distribution lines. Thus, Pb in drinking water occurs primarily as a result of corrosion from Pb
pipes, Pb-based solder, or brass or bronze fixtures within a residence, as noted in Chapter 3.
However, Pb in drinking water, although generally found at low concentrations in the United
States, has been linked to elevated blood Pb concentrations in general population groups. In one
U.S. prospective study, for example, children exposed to water with Pb concentrations >5 ppb
had blood Pb levels -1.0 |ig/dL higher than children with water Pb levels <5 ppb (Lanphear
et al., 2002). In another study of mothers and infants in Glasgow, Scotland, tap water was the
main correlate of elevated maternal blood Pb levels (Watt et al., 1996). Thus, under certain
conditions, water may not be a trivial source of Pb exposure in some locations.
Although marked reductions of Pb in U.S. market basket food supplies have occurred
during the past several decades, Pb-contaminated food still can be an important route of Pb
exposure (see Section 3.4). As shown in Table 8-2, dietary Pb intake levels for various U.S.
general population groups have notably declined over recent decades, due largely both to
reduced Pb air emissions from automotive gasoline as well as reduced use of solder in cans in the
United States. Several recent studies in the U.S. and Australia indicate that daily dietary Pb
intake is in the range of 2 to 10 |ig/day (see Section 3.4). Lastly, some U.S. population groups
that frequently consume canned foods imported from non-U.S. countries that still allow use of
lead-soldered cans may be at distinctly greater risk for exposure to Pb via dietary intake and
consequent higher blood-Pb concentration.
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Table 8-2. Estimated Dietary Lead Intake in U.S. Population Groups
in 1982-1984 versus 1994-1996.
Gender and Age Groups
Infants, 6-11 months
Children, 2 years
Children, 6 years
Females, 14-16 years
Males, 14-16 years
Females, 20-25 years
Males, 20-25 years
Females, 60-65 years
Males, 60-65 years
1982-1984 Total Diet Study
Dietary Pb intake (jig/day)
16.7
23.0
—
28.7
41.3
29.6
40.9
30.4
37.6
1994-1996 Total Diet Study
Dietary Pb intake (jig/day) *
0.8-5.7
2.4-10.1
3.5-13.2
3.6-14.9
4.0-17.1
3.5-15.6
4.2-18.8
3.6-16.0
4.5-19.5
*Note - dietary intakes presented as ranges based on use of different methods to account for measurements below
limit of detection.
8.3 TOXICOKINETICS, BIOLOGICAL MARKERS, AND MODELS OF
LEAD BURDEN IN HUMANS
Understanding the relationships between human exposure to Pb in external media (air,
food, water, soil/dust) and internal Pb burden in blood and other body tissues is a key issue of
much importance in carrying out risk assessments that evaluate the potential risk for adverse
health effects to occur in response to various Pb exposure scenarios. Use of biomarkers to index
Pb exposures is predicated on knowledge concerning Pb toxickinetics. Blood-Pb concentrations
have long been the most widely used biomarker by which to index Pb exposures in children and
adults (as discussed extensively in the 1977 Lead AQCD and the 1986 Lead AQCD/Addendum).
At the time of the 1986 Lead AQCD, it was recognized that Pb distributed to and accumulated in
several bone compartments which exhibited differing mobility profiles. It was also recognized
that a larger fraction of total body burden of Pb is found in the bones of adults relative to
children. The possibility of bone-Pb serving as a source of long-term internal Pb exposure was
considered. New studies that have since been published on the kinetics of Pb movement into and
out of bone demonstrate the importance of bone-Pb stores as a source of Pb to the blood in
retired lead workers and during pregnancy, as discussed in Chapter 4 of this document.
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Additional information regarding Pb absorption, distribution, and elimination in humans is also
discussed in Chapter 4, and some of the most important points regarding these and other aspects
related to Pb toxickinetics are summarized below.

8.3.1 Biokinetics of Lead Uptake and Internal Distribution
Humans are exposed to Pb mainly by ingestion and inhalation. The absorption of Pb is
affected by factors such as an individual's age and diet, as well as chemical and physical
properties of the ingested or inhaled Pb, as discussed in Section 4.2.1. Lead absorption appears
to be increased by both iron and calcium deficiency. Fasting also increases the absorption of Pb
from ingested soil. Lead absorption in humans may be a capacity limited process, such that the
fraction of ingested Pb that is absorbed may decrease with increasing rate of Pb intake.
The available studies to date, however, do not provide a firm basis for discerning whether the
gastrointestinal absorption of Pb is limited by dose. The size of ingested Pb particles also affects
absorption, with absorption decreasing as particle size increases.
In general, the Pb burden in the body may be viewed as being divided between a dominant
slow compartment (bone) and a smaller fast compartment (soft tissues). This distribution of Pb
in the body and factors affecting the exchange of Pb between bone and blood are discussed in
detail in Sections 4.2.2, 4.3.1, and 4.3.2. In human adults, more than 90% of the total Pb body
burden is found in the bones, whereas bone Pb accounts for -70% of the body burden in
children. The highest soft tissue concentrations in adults also occur in liver and kidney cortex.
Lead in blood is exchanged between both of these compartments. The contribution of bone Pb to
blood Pb changes with the duration and intensity of the Pb exposure, age, and various
physiological variables (e.g., nutritional status, pregnancy, menopause).
As also discussed in Chapter 4, Pb accumulates in bone regions having the most active
calcification at the time of exposure. Lead accumulation is thought to occur predominantly in
trabecular bone during childhood and in both cortical and trabecular bone in adulthood. Lead
concentrations in bone increase with age throughout life, indicative of a relatively slow turnover
of Pb in adult bone. Lead content in some bones (i.e., mid femur and pelvic bone) increases into
adulthood, plateaus at middle age, and then decreases at older ages. This decrease is most
pronounced in postmenopausal females and may be due to osteoporosis and the release of Pb
from resorbed bone to blood. Lead in adult bone can serve to maintain blood-Pb levels long after
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external exposure has ceased. During pregnancy, bone Pb can also serve as a Pb source with the
resorption of maternal bone for production of the fetal skeleton, and maternal bone Pb can
continue postnatally to serve as a source of Pb exposure to the offspring via maternal
lactation/breastfeeding (see Section 4.3.2.5).
In contrast to Pb in bone, which accumulates with continued exposure in adulthood, Pb
concentrations in soft tissues (e.g., liver and kidney) are relatively constant in adults, reflecting a
faster turnover of lead in soft tissue relative to bone (as discussed in Chapter 4). It is also noted
that Pb in soft tissues exists predominantly bound to protein (see Section 5.11). High affinity
cytosolic Pb-binding proteins (PbBPs) have been identified in rat kidney and brain. Other high-
affinity Pb-binding proteins have been isolated in human kidney, two of which have been
identified as a 5 kD peptide, thymosin 4, and a 9 kD peptide, acyl-CoA binding protein.
Lead in blood is found primarily (-99%) in the red blood cells. As discussed in
Sections 4.2.2 and 4.3.1, 5-aminolevulinic acid dehydratase (ALAD) is the primary Pb-binding
ligand in erythrocytes. Lead binding to ALAD is saturable; the binding capacity has been
estimated to be -850 |ig/dL red blood cells (or -340 |ig/dL whole blood), with an apparent
dissociation constant of-1.5 |ig/L. It has been suggested that the small fraction of Pb in plasma
(<0.3%) may be the more biologically labile and lexicologically active fraction of circulating Pb.
Several authors have proposed that Pb released from the skeleton was preferentially partitioned
into serum compared with red cells. About 40 to 75% of Pb in the plasma is bound to proteins,
of which albumin appears to be the dominant ligand. Lead in serum not bound to protein exists
largely as complexes with low molecular weight sulfhydryl compounds (e.g., cysteine,
homocysteine) and other ligands.

8.3.2 Selection of Blood-Lead Concentration as Key Index of Lead Exposure
Blood-Pb concentration is extensively used in epidemiologic studies as an index of
exposure and body burden mainly due to the feasibility of incorporating its measurement into
human studies relative to other potential dose indicators, e.g., lead in kidney, plasma, urine, or
bone. Section 4.3.1 considers the use of blood Pb as a marker of Pb exposure and body burden,
and the contribution of bone Pb to the blood is specifically discussed in Section 4.3.2.4. A single
blood-Pb measurement may not distinguish between a history of long-term lower level Pb
exposure from a history that includes higher acute exposures, as discussed by Mushak (1998).
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An additional complication is that the relationship between Pb intake and blood-Pb concentration
is curvilinear; that is, the increment in blood-Pb concentration per unit of Pb intake decreases
with increasing blood-Pb concentration, both in children and in adults. In general, higher blood
Pb concentrations can be interpreted as indicating higher exposures (or lead uptakes); however,
they do not necessarily predict higher overall body burdens. Similar blood-Pb concentrations in
two individuals (or populations) do not necessarily translate to similar body burdens or similar
exposure histories. The disparity in the kinetics of blood Pb and cumulative body burden may
have important implications for the interpretation of blood-Pb concentration measurements in
some epidemiology studies, depending on the health outcome being evaluated.
Bone Pb, as also indicated in Chapter 4, has begun to be accorded increasing attention as
another potentially useful marker for Pb exposure. It is thought that bone-Pb measurements
likely constitute a better indication of overall past cumulative Pb exposure history than do blood
Pb concentrations, which are more strongly influenced by recent Pb exposures. Approaches to
measurement of bone Pb in living human or animal subjects are discussed in Section 4.3.2.2 and
mainly involve different x-ray techniques that have undergone extensive intercomparison testing
and refinements during the past decade or so. Still, in contrast to blood-Pb concentrations,
bone-Pb measurements have not yet gained widespread use in epidemiologic studies as a key
biomarker for Pb exposure.
In addition to blood Pb and/or bone Pb, concentrations of Pb in hair and urine have at
times also been used as biomarkers of Pb exposure (see Sections 4.3.4 and 4.3.5). However, an
empirical basis for interpreting hair Pb measures in terms of body burden or exposure has not
been firmly established. As discussed in Chapter 4, hair Pb measurements are subject to error
due to contamination of the hair surface with environmental Pb and contaminants in artificial
hair treatments (e.g., dyeing, bleaching, permanents) and, as such, are a relatively poor predictor
of blood-Pb concentration, particularly at low blood-Pb levels (< 10 to 12 |ig/dL). Spontaneous
urine-Pb excretion also provides little reliable information, unless adjusted to account for
unmeasured variability in urine flow rate. Analogous to blood-Pb concentration measurements,
spontaneous urinary Pb excretion measured in an individual at a single point in time mainly
reflects the recent exposure history. As a result, spontaneous urinary-Pb measurement may serve
as a feasible surrogate for plasma-Pb concentration, and may be useful for exploring dose-
response relationships for effect outcomes that may be more strongly associated with plasma-Pb
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level than overall Pb body burden. On the other hand, measurement of notably increased urinary
Pb excretion in response to Succimer or other approved chelant challenge has proven to be
reliable in both pediatric clinical and occupational settings as reflecting a history of excessive Pb
exposure.

8.3.3 Trends in U.S. Blood Lead Levels
As discussed in Section 4.3.1.3, blood-Pb concentrations in the U.S. general population
have been monitored over the past three decades via the National Health and Nutrition
Examination Survey (NHANES) conducted by the Centers for Disease Control and Prevention.
Data from the most recent survey (NHANES IV, Centers for Disease Control, 2005) are shown
in Tables 8-3 and 8-4. For survey years 2001-2002, the geometric mean blood-Pb level for ages
>1 year (n = 8,945) was 1.45 |ig/dL (95% CI: 1.39, 1.52); with the geometric mean in males
(n = 4,339) being 1.78 |ig/dL (95% CI: 1.71, 1.86) and in females (n = 4,606) being 1.19 |ig/dL
(95% CI: 1.14, 1.25). Blood-Pb concentrations in the U.S. general population have decreased
over the past three decades as regulations regarding leaded fuels, leaded paint, and lead-
containing plumbing materials have decreased Pb exposure among the general population.
Changes in average blood-Pb concentrations among U.S. children over time are shown in
Figure 8-5.
Blood Pb concentrations can vary considerably as a function of age, physiological state
(e.g., pregnancy, lactation, menopause), and numerous other factors that affect exposure to Pb.
The NHANES data provide estimates for average blood lead concentrations in various
demographic strata of the U.S. population. NHANES III Phase 2 samples were collected during
1991 to 1994. Geometric mean blood-Pb concentrations of U.S. adults, ages 20 to 49 years,
estimated from the NHANES III Phase 2, were 2.1 |ig/dL (95% CI, 2.0, 2.2). Among adults,
blood-Pb concentrations were highest in the strata that included ages 70 years and older
(3.4 |ig/dL; 95% CI, 3.3, 3.6). The geometric mean blood-Pb concentration of children, ages 1 to
5 years, was 2.7 (95% CI, 2.5, 3.0) for the 1991 to 1994 survey period; however, the mean varied
with socioeconomic (SES) status and other demographic characteristics that have been linked to
Pb exposure (e.g., age of housing). Central estimates from the NHANES III Phase 2 (1991 to
1994), when compared to those from NHANES III Phase 1 (1988 to 1991) and the NHANES II
(1976 to 1980), indicate a clear downward temporal trend in U.S. blood-Pb concentrations over
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Table 8-3. Blood Lead Concentrations in United States by Age, NHANES IV (1999-2002)
Age
Survey Period
N
Blood Lead
(Mg/dL)a
1-5 years
1999-2000 2001-2002
723 898
2.23 1.70
(1.96,2.53) (1.55,1.87)
6-11 years
1999-2000 2001-2002
909 1,044
1.51 1.25
(1.36, 1.66) (1.14, 1.36)
12-19 years
1999-2000 2001-2002
2,135 2,231
1.10 0.94
(1.04,1.17) (0.90,0.99)
^20 years
1999-2000 2001-2002
4,207 4,772
1.75 1.56
(1.68, 1.81) (1.49, 1.62)
aBlood lead concentrations presented are geometric means (95% CI).
oo

to
o
Table 8-4. Blood Lead Concentrations in United States by Gender, NHANES IV (1999-2002)
Gender
Survey
n
Blood
(Mg/dL
Period

Lead
Males
1999-2000
3,913
2.01
(1.93,2.09)
2001-2002
4,339
1.78
(1.71, 1.86)
Females
1999-2000
4,057
1.37
(1.32, 1.43)
2001-2002
4,606
1.19
(1.14, 1.25)
aBlood lead concentrations presented are geometric means (95% CI).
-------
0)
T3
O
_O
CO
18
16
14^
12-
10-
8
6

2
0
1976-1980 1988-1991 1991-1994 1999-2000
Survey Period
2001-2002
Figure 8-5. Blood lead concentrations in U.S. children, 1-5 years of age. Shown are
geometric means and 95% confidence intervals as reported from the
NHANES II (1976-1980) and NHANES III Phase 1 (1988-1991; Pirkle et al.,
1994); NHANES III Phase 2 (1991-1994; Pirkle et al., 1998); and NHANES IV
(1999-2000, 2001-2002; Centers for Disease Control, 2005).
the past 20 years or so. It should be noted, however, that blood-Pb levels have been declining at
differential rates for various general subpopulations, as a function of income, race, and certain
other demographic indicators such as age of housing. Also, substantial caution should be
exercised with regard to use of NHANES data for risk assessment purposes, in that the
nationally-representative NHANES quantitative results (e.g., national mean blood-Pb levels or
strata-classified national mean blood-Pb values) vary with regard to how reflective they may be
of specific regions or communities.

8.3.4 Approaches to Predictive Estimation of Pb-Exposure Impacts on
Distribution to Internal Tissues
As indicated in Chapter 4, a key issue of much importance in carrying out lead risk
assessments that evaluate the potential likelihood of Pb-induced health effects is the estimation
of external Pb-exposure impacts on internal Pb tissue concentrations. This includes estimation of
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typical Pb exposure impacts on internal distribution of Pb to blood and bone (as key biomarkers
of Pb exposure), as well as to other "soft tissue" target organs (e.g., brain, kidney, etc.). Earlier
criteria assessments in the 1977 and 1986 Lead AQCDs extensively discussed then available
slope factor and/or other regression models of external Pb exposure impacts on blood-Pb
concentrations in human adults and children. The older slope factor analyses discussed in the
1977 and 1986 Lead AQCDs noted that at relatively low air-Pb concentrations (<2 |ig/m3),
pediatric blood-Pb levels generally increase by ~2 |ig/dL per each 1 |ig/m3 increment in air-Pb
concentration. Further refinements in regression modeling of Pb impacts on blood or bone Pb
are discussed in Chapter 4.
Several new studies discussed in Chapter 4 have investigated relationships between Pb
exposure and blood Pb in children (see Section 4.4.2). These studies support the concept that
contact with Pb in surface dust (interior and exterior) is a major contributor to Pb intake in
children. In one meta-analysis, the most common exposure pathway to emerge as notably
influencing blood-Pb concentration was exterior soil, operating through its effect on interior dust
Pb and hand Pb. Using a structural equation model, other analyses also found that the exposure
pathway component that was most influential on blood Pb was interior dust Pb loading, directly
or through its influence on hand Pb. Both soil and paint Pb influenced interior dust Pb.
However, interior and exterior paints were more significant contributors to childrens' blood-Pb
levels in urban (heavily paint-impacted) areas than at western U.S. extractive (mining/smelting)
industry sites, and dust Pb was more significantly linked to soil Pb than paint Pb at such western
sites. Still, these and other studies of populations near active sources of air emissions (e.g.,
smelters, etc.), substantiate the effect of airborne Pb and resuspended soil Pb on interior dust and
blood Pb.
Both exterior soil and paint Pb contribute to interior dust Pb levels. It has been estimated
that for every 1000 ppm increase in soil-Pb concentration, pediatric blood-Pb levels generally
increase by ~1 to 5 |ig/dL in exposed infants and children <6 years old. All ingested lead is not
absorbed to the same extent, in that intake of soil-Pb with low bioaccessibility or bioavailability
characteristics can yield distinctly lower-than-typical blood-Pb increments. Factors such as an
individual's age and diet, as well as chemical and physical properties of Pb, affect absorption,
e.g., absorption is increased by fasting and dietary iron or calcium deficiencies.
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Additional information on Pb biokinetics, bone mineral metabolism, and Pb exposures has
led to refinements and expansions of earlier modeling efforts. In particular, there are three
pharmacokinetic models that are currently being used or considered for broad application in Pb
risk assessment: (1) the Integrated Exposure Uptake BioKinetic (IEUBK) model for Pb in
children developed by EPA (U.S. Environmental Protection Agency, 1994a,b; White et al.,
1998); (2) the Leggett model, which also simulates Pb kinetics from birth through adulthood
(Leggett, 1993); and (3) the O'Flaherty model, which simulates Pb kinetics from birth through
adulthood (O'Flaherty, 1993, 1995). The above three models have been individually evaluated
to varying degrees, against empirical physiological data on animals and humans and data on
blood-Pb concentrations in individuals and/or populations (U.S. Environmental Protection
Agency, 1994a,b; Leggett, 1993; O'Flaherty, 1993). In evaluating models for use in risk
assessment, exposure data collected at hazardous waste sites have mainly been used as inputs to
model simulations (Bowers and Mattuck, 2001; Hogan et al., 1998). The exposure module in the
IEUBK model makes this type of evaluation feasible. Exposure-biokinetics models both
illustrate exposure-blood-body burden relationships and provide a means for making predictions
about these relationships that can be experimentally or epidemiologically tested.
The EPA IEUBK model for Pb has gained widespread use for risk assessment purposes in
the United States, and it is currently clearly the model of choice in evaluating multimedia Pb
exposure impacts on blood-Pb levels and distribution of Pb to bone and other tissues in young
children <7 years old. The EPA All Ages Lead Model (AALM), now under development, aims
to extend beyond IEUBK capabilities to model external Pb exposure impacts (including over
many years) on internal Pb distribution not only in young children, but also in older children,
adolescents, young adults, and other adults well into older years (up to 90 years of age).
The AALM essentially uses adaptations of IEUBK exposure module features, coupled with
adaptations of IEUBK biokinetics components (for young children) and of Leggett model
biokinetics components (for older children and adults). However, the AALM has not yet
undergone sufficient development and validation for it to be recommended for general risk
assessment use.
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8.4 LEAD-INDUCED TOXICITY: INTEGRATION OF TOXICOLOGIC
AND EPIDEMIOLOGIC EVIDENCE
8.4.1 Introduction
As discussed in the previous two chapters (Chapters 5 and 6) dealing with the toxicology
and epidemiology of Pb-induced health effects, Pb has been shown to exert a broad array of
deleterious effects on multiple organ systems via widely diverse mechanisms of action. Truly
remarkable progress has been made during the past several decades with regard to (a) more fully
delineating over time the wide variety of pathophysiologic effects associated with Pb exposure of
human population groups and laboratory animals and (b) the characterization of applicable
exposure durations and dose-response relationships for the induction of the multifaceted Pb
effects. This progress has been well documented by the previous Pb NAAQS criteria reviews
carried out by EPA in the late 1970s and during the 1980s, as well as being well reflected by
previous chapters of this document.
The 1977 Lead AQCD (U.S. Environmental Protection Agency, 1977) that provided key
scientific bases for the setting in 1978 of the current Pb NAAQS included discussion of both:
(a) historical literature accumulated during several preceding decades that established Pb
encephalopathy and other signs and symptoms of persisting severe central and/or peripheral
nervous system damage, as well as renal and hepatic damage, and anemia as typifying the classic
syndrome of acute and/or chronic high-level Pb poisoning among human pediatric and /or adult
population groups, and (b) evaluation of then newly-emerging evidence for more subtle and
difficult-to-detect "subclinical" Pb effects on IQ, other neurological endpoints, and moderate
blood hemoglobin deficits or other erythropoietic indicators of heme synthesis impairment,
which collectively were judged to constitute an array of adverse Pb health effects associated with
Pb exposures indexed by blood Pb concentrations ranging down to -30 |ig/dL. The next Pb
NAAQS criteria review during the 1980's, as contained in the 1986 Lead AQCD/Addendum and
its 1990 Supplement (U.S. Environmental Protection Agency, 1986a, b, 1990) documented
further rapid advances in Pb health effects research that provided (a) increasingly stronger
evidence that substantiated still lower fetal and/or postnatal Pb-exposure levels (indexed by
blood-Pb levels extending to as low as 10 to 15 |ig/dL or, possibly, below) as being associated
with slowed physical and neurobehavioral development, lower IQ, impaired learning, and/or
other indicators of adverse neurological impacts and (b) other pathophysiological effects of Pb
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on cardiovascular function, immune system components, calcium and vitamin D metabolism,
and other selected health endpoints.
Newly available scientific information published since the 1986 Lead AQCD/Addendum
and the 1990 Supplement, as assessed in previous chapters of this document, further expands our
understanding of a wide array of Pb-induced health effects, underlying mechanisms, and factors
that enhance or lessen susceptibility to Pb effects. Very importantly, the newly available
toxicologic and epidemiologic information, as integrated below, includes assessment of new
evidence substantiating risks of deleterious effects on certain health endpoints being induced by
distinctly lower than previously demonstrated Pb exposures indexed by blood-Pb levels
extending well below 10 |ig/dL in children and/or adults.
The ensuing subsections provide concise summarization and integrative synthesis of the
most salient health-related findings and conclusions derived from the current criteria assessment.
This includes discussion of new toxicologic and/or epidemiologic evidence concerning Pb-
induced (a) effects on neurobehavioral development and other indicators of nervous system
effects; (b) cardiovascular effects; (c) heme synthesis effects; (d) renal effects; (e) immune
system functions; (f) effects on calcium and vitamin D metabolism; (g) inter-relationships to
bone and teeth formation and demineralization; (h) effects on reproduction and other
neuroendocrine effects; and (i) genotoxicity and carcinogenic effects.

8.4.2 Neurotoxic Effects
The neurotoxic effects of Pb exposure are among those most studied and most extensively
documented among human population groups. Also, extensive experimental laboratory animal
evidence has been generated that (a) substantiates well the plausibility of the epidemiologic
findings observed in human children and adults and (b) expands our understanding of likely
mechanisms underlying the neurotoxic effects. Two major issues are important in considering
the concordance of human and animal results: (1) comparability of blood Pb levels (or other
internal dose markers) among species; and (2) comparability of neurobehavioral tests for animals
and humans.
Animal models are extremely important in the characterization of Pb neurotoxicity,
because exposures can be controlled to address questions about sensitive periods of exposure.
Unlike typical human exposures reported in epidemiology studies, Pb dosing to animals can be
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stopped at any time to address questions about the reversibility and persistence of neurotoxic
effects. Also, with animals, dosing can be varied to include very low doses to examine effects
seen with more current pediatric exposures. Animal models, especially inbred strains of rodents,
can lessen the effects of the critical confounder of parental cognitive ability, which parallels
human IQ. Also eliminated in controlled animal exposures are the confounders of SES and
nutrition.
In a review, Davis et al. (1990) state that little effort has been directed toward making
direct comparisons of human and animal dose-response relationships because of the abundance
of human exposure-effect data. The 1986 Lead AQCD also reported that there exists some
uncertainty in extrapolating from animals to humans because blood-Pb levels may not be directly
comparable. Both rats and monkeys may require higher Pb exposure levels than humans to
achieve a comparable blood-Pb level. It was further recognized by Davis et al. (1990) that, due
to inadequate numbers of subjects and the resulting lack of statistical power, it may not be
possible to detect subtle Pb-induced neurotoxic effects in both epidemiologic and experimental
studies.
As discussed in the 1986 Lead AQCD, questions have also been raised regarding the
comparability between neurobehavioral effects in animals and effects on human behavior and
cognitive function. One major difficulty is the lack of standardized methodologies or a
consistent operational definition by which to compare behavioral endpoints. In addition,
behavior is difficult to compare meaningfully across species, because behavioral analogies do
not necessarily demonstrate behavioral homologies. Davis et al. (1990) examined the
comparative neurotoxicity of Pb in humans and animals and noted that a problem in comparing
behavior and identifying behavioral similarities is that behavior is not a phenomenological given,
but an event or series of events that must be represented by abstracting of one or more of its
features. They further state that it is important of be mindful of "the degree to which the model
faithfully reflects the mechanisms underlying its referent."
In assessing the comparability of measures of cognitive function in humans and animals,
Sharbaugh et al. (2003) also state that of ultimate importance is finding sensitive homologous or
parallel neurobehavioral tests in humans and animals. Homologous tests are those for which the
same procedure is followed in humans and the animal species. Examples of homologous tests
include Bayley Scales of Infant Development II, which tests a number of behavioral and reflect
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tasks, and the visual recognition memory test. Both tests are performed in human infants and
nonhuman primates. Parallel tests are those that are conducted in a different manner in humans
and animals, but for which it is believed that the same cognitive function is being measured,
e.g., tests of learning, recognition memory, and long-term memory in humans and rodents.
Generally measures of cognitive function for humans and nonhuman primates are homologous,
while those with rodents are parallel (Sharbaugh et al., 2003).
The most widely used measure of cognitive function in epidemiologic studies is the
intelligence quotient or IQ score. An IQ score is a global measure reflecting the integration of
numerous behavioral processes. There is no direct parallel to IQ tests for nonhuman primates or
rodents. However, in animals a wide variety of tests that assess attention, learning, and memory
suggests that Pb exposure results in a global deficit in functioning, just as it is indicated by
decrements in IQ scores in children (Rice, 1996).
Examination of the effect of Pb on behavioral processes in human and experimental
animals needs to focus beyond IQ, as noted by Cory-Slechta (1996). One strategy would be to
use the same behavioral baselines in human studies that have revealed Pb-related deficits in
cognitive functions in experimental animal studies, particularly those such as discrimination
learning, reversal learning, repeated learning of response sequences, and concurrent schedule
transitions. Rice (1996) concurs with this view and states further that the use of IQ has proven to
be a sensitive indicator of Pb exposure, but that using more specific tests could provide even
greater sensitivity. In the following sections, the epidemiologic and toxicologic evidence of Pb-
induced effects on global as well as specific neurobehavioral outcomes are integrated and
discussed.

8.4.2.1 Neurocognitive Ability
Global Measures of Cognitive Function - Intelligence Testing and Academic Achievement
Lead effects on human neurocognitive ability have been assessed in epidemiologic studies
largely by use of age-appropriate, standardized IQ tests (as discussed in Section 6.2.3 of
Chapter 6). Assessment of intelligence in infants and young children has been performed using a
number of scales, including the various Bayley Scales of Infant Development and the McCarthy
Scales of Children's Abilities. Most studies used the Weschler Intelligence Scales for Children-
Revised (WISC-R) in older children. As discussed by Rice (1996), it is generally recognized
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that early tests of intelligence such as the Bayley scales do not measure the same functions as
tests used at school age such as the WISC-R and have little predictive validity for individual
children (though the Bayley scales may have better predictive power for low-functioning
children). Regardless, numerous well-conducted longitudinal cohort and cross-sectional studies
that evaluated various study populations in several different countries have consistently found
Pb-related IQ deficits from infancy through at least early school age.
For example, in the largest available new cross-sectional study, Lanphear et al. (2000)
examined the relationship between blood Pb concentrations and cognitive deficits in a nationally
representative sample of 4,853 U.S. children aged 6 to 16 years (geometric mean blood Pb of
1.9 |ig/dL) who participated in NHANES III, with 97.9% of the children having blood-Pb
concentrations <10 |ig/dL. Two subtests of the WISC-R, Block Design (a measure of visual-
spatial skills) and Digit Span (a measure of short-term and working memory) were administered;
and numerous potential confounders were assessed in the multivariable analyses. Although no
data on maternal IQ or direct observations of caretaking quality in the home were available, other
variables such as the poverty index ratio and education level of the primary caregiver may have
served as adequate surrogate measures of these important potential confounders. In multivariate
analyses, a significant covariate-adjusted relationship was found between blood-Pb level and
scores on both WISC-R subtest for all children as well as among those with blood-Pb levels
<10 |ig/dL. Blood-Pb concentration was also significantly associated with Block Design when
the multivariate analysis was restricted to children with blood Pb levels <7.5 |ig/dL.
Other recent studies of the association of Pb with IQ in children with low Pb exposures
have consistently observed effects at blood-Pb concentrations below 10 |ig/dL (as discussed in
Section 6.2.3 of Chapter 6). Most notably, a large international pooled analysis of 1,333 children
from seven different cohorts by Lanphear et al. (2005) estimated a decline of 6.2 points (95%
CI: 3.8, 8.6) in full scale IQ for an increase in concurrent blood-Pb level from 1 to 10 |ig/dL.
A common observation among some of these studies of low-level Pb exposure is a non-linear
dose-response relationship between blood Pb and neurodevelopmental outcomes. Although this
may seem at odds with certain fundamental toxicological concepts, it is possible that the initial
neurodevelopmental lesions seen at lower Pb levels may be disrupting different biological
mechanisms (e.g., early developmental processes in the central nervous system) than the more
severe effects of high Pb exposures that result in symptomatic poisoning and frank mental
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retardation. One ad hoc explanation may be that the predominant mechanism at very low
blood-Pb levels is rapidly saturated and that a different, less-rapidly-saturated process, becomes
predominant at blood-Pb levels greater than 10 |ig/dL.
Another global measure of cognitive function is academic achievement. Compared to the
vast number of studies assessing the blood Pb-IQ relationship in children, there are relatively
little data available on the relationship between Pb exposure and objective measures of academic
achievement. These studies focused on the effect of Pb on school performance, including
reading, math, spelling, and handwriting (see Section 6.2.4).
Lanphear et al. (2000) examined the relationship between blood-Pb levels and a
standardized measure of academic achievement among 4,853 NHANES III children, aged 6 to
16 years (geometric mean blood-Pb of 1.9 jig/dL). Subjects were administered the Arithmetic
and Reading subtests of the Wide Range Achievement Test-Revised (WRAT-R). Multiple linear
regression revealed significant Pb-related decrements in Arithmetic and Reading scores.
In analyses stratified by blood-Pb levels, statistically significant inverse relationships between
blood-Pb levels and performance for both Reading and Arithmetic subtests were found for
children with concurrent blood-Pb concentrations <5 |ig/dL. However, possible attribution of
the observed associations of decrements in WRAT-R scores to earlier (but unmeasured) likely
somewhat higher peak blood-Pb concentrations cannot be ruled out.
Several other epidemiologic studies observed inverse associations between exposure to Pb
and academic achievement, for the endpoints noted above as well as class rankings and high
school graduation rates. Two studies specifically examined the effects of blood-Pb levels
<10 |ig/dL on academic achievement. One study examined 533 girls aged 6 to 12 years (mean
blood Pb level of 8.1 |ig/dL) in Riyadh, Saudi Arabia and observed that, in a subset of students
with blood-Pb levels <10 |ig/dL, class rank percentile was statistically significantly associated
with blood-Pb levels (Al-Saleh et al., 2001). In another study in Torreon, Mexico, a significant
inverse relationship was found between blood-Pb concentrations and math and vocabulary scores
in 594 second graders (mean blood Pb of 11.4 |ig/dL). In segmented regression analyses, slopes
for blood Pb associations with vocabulary and math scores were significantly steeper below
10 |ig/dL than above (Tellez-Rojo et al., 2006). Associations between Pb exposure and
academic achievement observed in the above-noted studies were significant even after
adjusting for IQ, suggesting that Pb-sensitive neuropsychological processing and learning factors
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not reflected by global intelligence indices might contribute to reduced performance on academic
tasks.

Specific Cognitive Abilities - Learning, Memory, and Attention
In addition to IQ and academic achievement, epidemiologic studies have evaluated Pb
effects on specific cognitive abilities, e.g., attention, executive functions, language, memory,
learning, and visuospatial processing. Results from these studies are most comparable to those
experimental animal studies examining Pb effects on learning ability, memory, and attention.
Executive functions refer to an individual's ability to regulate attention and engage
several related higher order cognitive processes such as strategic planning, control of impulses,
organized search, flexibility of thought and action, and self-monitoring of one's own behavior.
In some earlier studies, assessed in the 1986 Lead AQCD/Addendum and/or 1990 Supplement,
Pb exposure was associated with higher frequency of negative ratings by teachers and/or parents
on behaviors such as inattentiveness, impulsivity, distractibility, and lack of persistence on
assigned tasks, as well as slowed psychomotor responses and more errors on simple, serial, and
choice reaction time tasks. More recent studies (see Section 6.2.5) have observed inverse
relationships between exposure to Pb and attentional behaviors and executive function, even in
cohorts where more than 80% of the children had blood-Pb levels <10 |ig/dL. These
associations were observed across a wide range of age groups, from children 4-5 years to
19-20 years of age. Higher blood-Pb levels were also associated with impaired memory and
visual-spatial skills.
Whether the domains of executive functions, attention, memory, or visual-motor
integration per se are specifically sensitive to Pb is unknown, as there is rarely a one-to-one
correspondence between performance on a focused neuropsychological test and an underlying
neuropsychological process. For example, a low score on the visual-motor integration test may
reflect singular or multiple neurobehavioral deficits, e.g., difficulties with graphomotor control,
visual perception, behavioral monitoring (impulsivity), and/or planning (executive functions).
Early Pb exposure may be associated with poorer performance on executive/regulatory functions
that are thought to depend on the frontal or prefrontal brain regions. The prefrontal cortex is
highly innervated by neuronal projections from the midbrain and has the highest concentration of
dopamine of all cortical areas. The dopamine system, which plays a key role in cognitive
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abilities mediated by the prefrontal cortex, is particularly sensitive to Pb, based on data from
studies of rodents and nonhuman primates (see Section 5.3). These animal toxicology findings
provide strong biological plausibility in support of the concept that Pb may impact one or more
of these specific cognitive functions in humans.
Results from fixed interval (FI) studies in 4 species of laboratory animal models at
environmentally relevant doses (as shown in Figure 5-6) clearly demonstrate that Pb induces
increased response rates. The increased response rates are mostly due to shortened time to
initiate responding in the interval and more rapid response once the responding begins. This
pattern of effects has been compared to young human males diagnosed with Attention
Deficit/Hyperactivity Disorder (ADHD), and it is thought that increases in response rates found
in animal models parallel increases in impulsivity in self-control paradigms (as noted by
Cory-Slecta, 2003a).
As noted in Section 5.3.5, NMDAR function and ontogeny are affected by Pb exposure.
Functional NMDARs are necessary for spatial learning and memory, as tested by the Morris
water maze. Several studies that evaluated Pb effects with this learning paradigm have shown
that chronic exposure to 250 ppm Pb affected long-term memory. The effect of Pb on memory is
not clearly understood. In some studies, memory impairment was found at blood Pb levels of
10 |ig/dL, whereas numerous other studies found no Pb-induced effects on short term memory.
This parallels findings from most cross-sectional and prospective epidemiological studies, which
generally did not detect low-level Pb exposure effects on memory.
Studies of early developmental cognitive ability in monkeys postnatally exposed to Pb
(see Section 5.3.5) have used the Early Infant Behavioral Scale, which is modeled after the
Brazelton Neonatal Behavioral Assessment. The monkeys displayed both decreased visual
attentiveness and increased agitation. Other epidemiologic studies using the Brazelton scale
have shown analogous results for human infants.

8.4.2.2 Behavior, Mood, and Social Conduct
Investigating associations between Pb exposure and behavior, mood, and social conduct
of children has been an emerging area of research (see Section 6.2.6). Early studies indicated
linkages between lower-level Pb toxicity and behavioral problems (e.g., aggression, attentional
problems, and hyperactivity) in children. Blood-Pb and tooth-Pb levels have been associated
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with behavioral features of ADHD, including distractibility, poor organization, lacking
persistence in completing tasks, and daydreaming, in various cohorts of children with a wide
range of Pb exposures. In the Port Pirie, Australia cohort study, the relationship between Pb
exposure and emotional and behavioral problems at ages 11 to 13 years were examined after
stratifying the data set by gender. Stronger associations with Pb were observed for externalizing
behavior problems in boys compared to girls. In contrast, greater internalizing behavior
problems were observed for girls than in boys.
The relationship between Pb exposure and delinquent and criminal behavior also has been
addressed in several investigations. Studies linking attention deficits, aggressive and disruptive
behaviors, and poor self-regulation with Pb have raised the prospect that early exposure may
result in an increased likelihood of engaging in antisocial behaviors in later life. In two
prospective cohort studies conducted in Pittsburgh (Needleman et al., 1996) and Cincinnati
(Dietrich et al., 2001), elevated Pb levels were associated with several measures of behavioral
disturbance and delinquent behavior. It was also observed that bone-Pb levels in adjudicated
delinquents were significantly higher than in non-delinquent community control subjects in
Pittsburgh and surrounding Allegheny County, PA environs. In a Philadelphia survey of
987 African-American youths, a history of Pb poisoning was among the most significant
predictors of delinquency and adult criminality in males (Denno, 1990).
These results indicate that Pb may play a role in the epigenesis of behavioral problems in
inner-city children independent of other social and biomedical cofactors. The particular
biological mechanisms that may underlie Pb effects on aggression, impulsivity, and poor self-
regulation are not yet well understood. However, Pb impacts many brain sites and processes
involved in impulse control (Lidsky and Schneider, 2003). Also, the increased risk of
delinquency may indirectly be a consequence of attentional problems and academic under-
achievement among children who suffered higher Pb exposures during their formative years
(as noted by Needleman et al., 2002).
Lead has been shown to affect reactivity to the environment and social behavior in both
rodents and nonhuman primates at blood-Pb levels of 15 to 40 |ig/dL, though the literature has
some conflicting studies (see Section 5.3.5). In general, most studies show a Pb-induced
enhancement of social investigation and exploratory behavior. Aggression was increased in
hamsters, but not in rats, though the latter did display increased behavioral reactivity to stimuli.
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Early postnatal testing of Pb-exposed rhesus monkeys has shown lowered muscle tonus, greater
agitation, and decreased visual attentiveness. Chronically exposed rhesus monkeys exhibited Pb-
induced disruption of social play and increased self-stimulation and fearful behavior that
persisted for months after exposure ended. Thus, no clear pattern is yet apparent in the
experimental literature examining aggression that parallels the epidemiologic findings of
Pb-induced increases in aggression and delinquent behavior among humans. However, the
findings of increased reactivity to stimuli, impulsivity, and attention dysfunction found in both
Pb-exposed animals and humans may underlie some of the behavioral and emotional problems
observed epidemiologically.

8.4.2.3 Neurophysiologic Outcomes
Epidemiologic studies of the effect of Pb on sensory acuity have focused on hearing
thresholds and features of auditory processing in Pb-exposed children (see Section 6.2.7).
Schwartz and Otto (1987) observed significant Pb-associated elevations in pure-tone hearing
thresholds at various frequencies within the range of human speech among over 4,500 subjects
(4 to 19 years old) in NHANES II. These findings were replicated in a sample of
-3,000 subjects (6 to 19 years old) in the Hispanic Health and Nutrition Examination Survey
(HHANES) (Schwartz and Otto, 1991), including atblood-Pb levels <10 (ig/dL.
Dietrich et al. (1992) assessed the relationship between scores on a test of central auditory
processing (SCAN) and blood-Pb concentrations in 215 children 5 years of age drawn from the
Cincinnati Lead Study. Higher prenatal, neonatal, and postnatal blood-Pb concentrations were
associated with more incorrect identification of common monosyllabic words presented under
conditions of filtering (muffling). In another study, conducted in Poland, a significant
association between concurrent blood-Pb levels and increased hearing thresholds was also
observed among 155 children 4 to 14 years of age (median blood Pb of 7.2 |ig/dL) (Osman et al.,
1999). This relationship remained statistically significant when restricted to children with
blood-Pb levels <10 |ig/dL. The supportive evidence of a relationship between Pb exposure and
auditory processing suggests that Pb-related deficits in hearing and auditory processing may be
one plausible mechanism by which an increased Pb burden might impede a child's learning
(Bellinger, 1995).
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Animal studies have shown Pb-induced deficits in both auditory and visual acuity, which
may contribute to the cognitive deficits associated with Pb exposure. Blood Pb levels as low as
33 |ig/dL in nonhuman primates impair auditory function by increasing latencies in brainstem
auditory evoked potentials and elevating hearing thresholds. Blood-Pb levels of 19 |ig/dL in rats
have been found to cause selective effects on rod and bipolar cells, resulting in decreased
maximal ERG amplitude, decreased ERG sensitivity, and increased mean ERG latency. In a
review of Pb-induced auditory and visual dysfunction, Otto and Fox (1993) point to the
structural, biophysical, and photochemical similarities of rods in rats, monkeys and humans and
suggest that undetected visual or auditory deficits may profoundly impact both sensory motor
and mental development in children.
Electrophysiological evaluations have been conducted on Pb-exposed children in attempts
to obtain a more direct measure of the toxicant's impact on the nervous system (as discussed in
Section 6.2.9). Much of this work was conducted by Otto and colleagues during the 1980s and
demonstrated effects of Pb on neurosensory functioning (auditory and visual evoked potentials)
across a broad range of exposures. A more recent study examining the associations between Pb
exposure and brainstem auditory evoked responses observed results that were less consistent
(Rothenberg et al., 1994).
The methods of Magnetic Resonance Imaging (MRI) and Magnetic Resonance
Spectroscopy (MRS) have also been applied to evaluate Pb-exposed children. Several studies
compared subjects with elevated blood-Pb levels (blood Pb >23 |ig/dL) to control subjects
(blood Pb <10 |ig/dL). Although all of the participants had normal MRI examinations, the
Pb-exposed subjects exhibited a significant reduction in the ratios of N-acetylaspartate to
creatine and phosphocreatine in frontal gray matter compared to controls (Trope et al., 2001).
Similarly, reduced peak values of N-acetylaspartate, choline, and creatine were found in all four
brain regions in Pb-exposed children relative to control subjects (Meng et al., 2005).
The observed reductions in brain N-acetylaspartate levels may be related to decreased neuronal
density or neuronal loss. Also, reduced choline signal may indicate decreased cell membrane
turnover or myelin alterations that could lead to central nervous system hypertrophy, while lower
creatine may indicate reduced neuronal cell viability.
Using functional MRI (fMRI), a subsample of 48 young adults (aged 20-23 years) from
the Cincinnati Lead Study performed an integrated verb generation/finger tapping paradigm
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(Cecil et al., 2005; Yuan et al., 2006). Higher childhood average blood-Pb levels were
significantly associated with reduced activation in Broca's area, a recognized region of speech
production in the left hemisphere, and increased activation in the right temporal lobe, the
homologue of Wernicke's area (an area associated with speech production) in the left
hemisphere. This suggests that elevated childhood Pb exposure may influence neural substrates
underlying semantic language function in normal language areas, with concomitant recruitment
of contra-lateral regions causing a dose-dependent atypical organization of language function.

8.4.2.4 Neuromotor Function and Vocalization
Only a few recent epidemiologic studies have evaluated neuromotor deficits as an
outcome of early Pb exposure (see Section 6.2.8). In the Cincinnati Lead Study cohort, blood-Pb
levels, both neonatal and postnatal, were significantly associated with poorer scores on measures
of bilateral coordination, visual-motor control, upper-limb speed and dexterity, fine motor
composite from the Bruininks-Oseretsky scales, and postural stability in children 6 years of age
(Dietrich et al., 1993b). In general, the strongest and most consistent relationships were
observed with concurrent blood-Pb levels (mean 10.1 |ig/dL). At 16 years of age, 78-month
postnatal blood-Pb levels were significantly associated with poorer fine-motor skills, as indexed
by covariate-adjusted factor scores derived from a factor analysis of a comprehensive
neuropsychological battery. Variables loading highly on the fine-motor component came from
grooved pegboard and finger tapping tasks. In the Yugoslavian Prospective Study, lifetime
average blood-Pb concentration through 54 months of age was associated with poorer fine motor
and visual motor function, but was unrelated to gross motor function (Wasserman et al., 2000a).
Another recent study examined the effect of multiple exposures (including Pb, mercury,
and PCBs) on neuromotor functions in 110 preschool Inuit children residing in Canada (Despres
et al., 2005). Significant associations that were found only for blood-Pb concentrations (mean of
5.0 |ig/dL) were those associated with increased reaction time, sway oscillations, alternating arm
movements, and action tremor. Even after eliminating children with blood-Pb levels >10 |ig/dL
(10% of cohort) from the analyses, results generally remained consistent, suggesting that
neuromotor effects of Pb occurred at blood Pb levels <10 |ig/dL.
Changes in vocalization are a potential biomarker for Pb exposure. That is, analyses of
acoustical cries in babies showed that percent nasalization decreased progressively over cord
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blood-Pb ranging from 4 to 40 |ig/dL and that the number of cries was inversely related to cord
blood-Pb. These data may parallel Pb-induced changes in vocalization seen in developing rats
(see Section 5.3.5).
Earlier studies showed developmental lags in gross activity in rats with blood-Pb levels as
low as 14 |ig/dl, but other studies have found often contradictory results. More recent nonhuman
primate studies showed either no effects or subtle motor impairments, increased durations of
activity, failures to habituate, increased agitation, and fear. Rodent studies showed either no
effects or increases in locomotor activity and changes in vocalization patterns. Thus, no clear
pattern of Pb-induced effects on motor activity has yet emerged, though many studies do point to
an increase in activity, as seen with epidemiologic findings. However, Cory-Slechta (1989), in
discussing behavioral endpoints in Pb neurotoxicity, suggests that motor activity has little
correspondence with more complex functions important for human populations.

8.4.2.5 Neurochemical Alterations
Examination of Pb-induced biochemical alterations of the nervous system has largely
been limited to laboratory animal toxicologic studies. Although the linkage of neurochemical
alterations in animal to human neurobehavioral function is somewhat speculative, these studies
do provide some insight into possible neurochemical mediators of Pb neurotoxicity.
As summarized in Section 5.3.2, it has long been well known that Pb2+ acts as a Ca2+
mimetic. This affects neurotransmitter release in a dose-dependent fashion at glutamatergic,
cholinergic, and dopaminergic synapses. Glutamate, acetylcholine and dopamine systems play
very important roles in both cognitive function and brain development in both laboratory animals
and in humans. Extensive research has focused on chronic Pb exposure effects on NMDA
receptors. Much of the data point to an inhibition of NMDAR and changes in the ontogeny of
receptor subunit expression, though full characterization of the effects on specific subunits is not
available.
Considerable research has also focused on interactions of Pb2+ and Ca-dependent kinases
and phosphodiesterases. Lead alters the activity of many of these enzymes, which results in
changes in CREB, the transcription factor that controls expression of genes involved in learning,
memory, and synaptic plasticity. Protein kinase C (PKC) is also a Pb target, though the Pb
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effects on PKC in the intact animal have not been fully characterized. Thus, possible
relationships of Pb effects on this pathway to human cognitive function effects are not yet clear.

8.4.2.6 Assessment of Dose-Response Relationships for Neurotoxic Effects of
Lead Exposure
An important consideration in assessing potential public health impacts associated with
Pb exposure is whether concentration-response relationships are linear across the full exposure
range or, rather, shows nonlinearity. Also of interest is whether any thresholds can be discerned
for various types of health effects associated with Pb exposure. The 1986 AQCD/Addendum
and 1990 Supplement concluded that neurotoxic effects were related to blood Pb levels of 10 to
15 |ig/dL and possibly lower. Since then, the U.S. Centers for Disease Control and Prevention
(CDC) and the World Health Organization (WHO) have also lowered their definition of an
elevated blood Pb concentration to 10 jig/dL (CDC, 1991; WHO, 1995). Average blood-Pb
levels in U.S. children ages 1 to 5 years decreased from 15 |ig/dL in 1976-1980 to ~3 |ig/dL in
1991-1994 (CDC, 2000; Pirkle et al., 1998), allowing more recent studies to examine the effects
of low level Pb exposure on the neurodevelopment of children (as discussed in Section 6.2.3).
Several recent epidemiologic studies have observed significant Pb-induced IQ decrements
in children with peak blood Pb levels <10 |ig/dL (e.g., Canfield et al., 2003a; Lanphear et al.,
2005) and, in some cases, possibly below 5 |ig/dL (Bellinger and Needleman, 2003; Tellez-Rojo
et al., 2006). The most compelling evidence for effects below 10 |ig/dL, as well as a nonlinear
relationship between blood Pb levels and IQ, comes from the international pooled analysis of
seven prospective cohort studies (n = 1,333) by Lanphear et al. (2005). The slope for Pb effects
on IQ was steeper at lower blood-Pb levels, as indicated by the cubic spline function, the log-
linear model, and the piece-wise linear model. The shape of the spline function indicated that the
steepest declines in IQ were at blood-Pb concentrations <10 |ig/dL. Based on stratified analyses
using two cut points, a maximal blood-Pb of 7.5 and 10 |ig/dL, the effect estimate for children
with maximal blood-Pb levels <7.5 |ig/dL was significantly greater than for those with a
maximal blood-Pb >7.5 |ig/dL. Thus, recent epidemiologic evidence is highly indicative of Pb-
induced neurocognitive deficits in children at blood Pb levels below 10 |ig/dL and, possibly, as
low as 5 |ig/dL.
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In addition to IQ, significant associations were observed at low blood-Pb levels for other
neurotoxicity endpoints. In the large NHANES III study, children aged 6 to 16 years with
concurrent blood Pb <5 |ig/dL exhibited significant Pb-related decrements in Arithmetic and
Reading scores (Lanphear et al., 2000), but the possibility of earlier somewhat higher peaks in
blood-Pb levels of the same children around 2.5 years of age cannot be ruled out. Inverse
relationships between exposure to Pb and attentional behaviors and executive function were also
observed in cohorts where >80% of the children had blood Pb levels <10 |ig/dL (Canfield et al.,
2003b; Stiles and Bellinger, 1993). Other studies have also found significant Pb-induced
impairments of neuromotor function (Despres et al., 2005) and hearing (Osman et al., 1999;
Schwartz and Otto, 1987, 1991) in children with blood-Pb levels <10 |ig/dL. Collectively, these
studies most clearly indicate that Pb is associated with various neurodevelopmental endpoints in
children at blood-Pb levels as low as 5 to 10 |ig/dL. However, the shape of the concentration-
response curve has not been as extensively examined in these studies; thus, there is still some
question as to whether, for endpoints other than IQ, larger effects per incremental dose occur at
blood Pb levels <10 |ig/dL.
As stated in Section 5.3.7, there is little if any evidence from experimental animal studies
that allow for any clear delineation at this time of a threshold for neurotoxic effects of Pb.
Neurobehavioral changes have been reported in rodent studies at blood-Pb levels of-10 |ig/dL,
whereas neurochemical and neurophysiological changes have been reported at blood Pb levels of
-15 |ig/dL. However, these levels do not necessarily indicate a threshold for such effects but,
rather, may only reflect the levels of exposure that have been studied to date. Also, other
information, discussed in Chapter 4, suggests that blood-Pb concentrations in some animal
models (e.g., the rat or other rodents) may be more comparable to somewhat lower blood-Pb
levels seen in humans with more or less similar exposure conditions. Thus, blood-Pb levels
associated with neurobehavioral effects appear to be reasonably parallel between humans and
animals at reasonably comparable blood-Pb concentrations; and such effects appear likely to
occur in humans ranging down at least to 5-10 |ig/dL, or possibly lower (although the possibility
of a threshold for such neurotoxic effects cannot be ruled out at lower blood-Pb concentrations).
Lead appears to exhibit a curvilinear, or U-shaped, dose-effect relationship for a number
of toxicological endpoints. This effect is not unique to Pb, but occurs with other toxicants
(e.g., mercury chloride, chlordane, toluene, chlorpyrifos) as well, as reviewed by Calabrese
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(2005). In the case of Pb, this nonlinear dose-effect relationship occurs in the pattern of
glutamate release (Section 5.3.2), in the capacity for long term potentiation (LTP; Section 5.3.3),
and in conditioned operant responses (Section 5.3.5). The 1986 Lead AQCD also reported
U-shaped dose-effect relationships for maze performance, discrimination learning, auditory
evoked potential, and locomotor activity. Davis and Svendsgaard (1990) reviewed U-shaped
dose-response curves and their implications for Pb risk assessment. An important implication is
the uncertainty created in identification of thresholds and "no-observed-effect-levels" (NOELS).
As a nonlinear relationship is observed between IQ and low blood Pb levels in humans, as well
as in new toxicologic studies wherein neurotransmitter release and LTP show this same
relationship, it is plausible that these nonlinear cognitive outcomes may be due, in part, to
nonlinear mechanisms underlying these observed Pb neurotoxic effects.

8.4.2.7 Susceptibility and Vulnerability to Neurotoxic Effects from Lead Exposure
Several factors have emerged as likely affecting the relative likelihood that humans or
laboratory animals may experience Pb-induced neurotoxic effects under particular Pb exposure
conditions. Among the more important factors identified thus far are: age; gene-environment
interactions; gender; and socioeconomic status.

Age
Identifying discrete periods of development when the fetus or child is particularly
susceptible to Pb's effects on neurodevelopment is difficult as (1) age strongly predicts the
period of peak exposure (around 18-27 months when there is maximum hand-to-mouth activity),
making it difficult to distinguish whether greater neurotoxic effects resulted from increased
exposure or enhanced susceptibility at a particular age; and (2) despite changes in actual blood
Pb levels, children tend to maintain their relative rank order with regard to neurodevelopment
indicators through time, limiting the ability to examine critical periods of development.
One notable epidemiologic study has observed the strongest associations between IQ at
school age and academic achievement and blood Pb concentrations at 2 years of age (Bellinger,
et al., 1992). An understanding of human neurodevelopmental biology supports the notion that
the first 3 years of life represent a particularly vulnerable period. Maximal ingestion of Pb often
coincides with this same period of time when major events are occurring in the development of
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the human central nervous system, including some neurogenesis, rapid dendritic and axonal
outgrowth, synaptogenesis, synaptic pruning, and programmed cell death (seeNolte, 1993).
However, the human central nervous system continues to mature and be vulnerable to
neurotoxicants throughout the lifespan (Selevan et al., 2000, Weiss, 2000, Rice and Barone,
2000). Several prospective studies of children with both high and low Pb exposures found
concurrent or lifetime average blood-Pb levels to be more strongly associated than other earlier
blood-Pb measures with school age IQ and other measures of neurodevelopment (Canfield et al.,
2003a; Dietrich et al., 1993a,b; long et al., 1996; Wasserman et al., 2000b). Using data from the
Treatment of Lead-Exposed Children (TLC) study, Chen et al. (2005) examined whether cross-
sectional associations observed in school age children 84-90 months of age represented residual
effects from 2 years of age or "new" effects emerging among these children. Concurrent blood-
Pb concentration always had the strongest association with IQ. The strength of the cross-
sectional associations increased over time, despite lower blood-Pb concentrations in older
children. Adjustment for prior IQ did not fundamentally change the strength of the association
with concurrent blood-Pb level. These results suggest that Pb exposure continues to be toxic to
children as they reach school age, but does not support an interpretation that all of the damage
occurred by the time the child reaches 2 to 3 years of age. Examination of the toxicologic
evidence may be especially enlightening on this topic, given the difficulties involved in assessing
any periods of particularly increased susceptibility to Pb neurodevelopmental health effects in
the epidemiologic setting.
Cory-Slechta (1989) has reviewed age considerations in the neurotoxi col ogy of Pb and
concluded that: (1) though the presumed critical exposure period is prenatal and neonatal,
vulnerability extends well beyond this period in both rodents and humans; (2) for some
neurobehavioral endpoints (such as schedule-controlled behavior), the developmental period of
exposure can be relatively unimportant, whereas the body burden of Pb is more critical;
(3) enhanced vulnerability to Pb may also occur in later life, as ageing processes induce
degenerative changes in various organ systems; and (4) age-related shifts occur in the
toxicokinetics of Pb, such that Pb can be redistributed to brain and liver from bone during later
phases of life beyond the time of earlier Pb exposures.
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Gene-Environment Interactions
A few recent epidemiologic studies have examined susceptibility to Pb health effects as
related to genetic polymorphisms associated with Pb biokinetics and/or neurotransmitter
metabolism and function (as discussed in Section 6.2.10). Genetic polymorphisms in certain
genes have been implicated as influencing the absorption, retention and toxicokinetics of Pb in
humans. Although the ALAD gene has been the most studied, as of yet, the consequences of
different alleles for susceptibility to the neurodevelopmental consequences of Pb exposure in
children are unclear. For example, ALAD2 polymorphism has been implicated in influencing
vulnerability by raising blood Pb levels or by decreasing them by maintaining Pb in a
sequestered state in the bloodstream. Suggestive but limited evidence appears to indicate that
adolescents with the ALAD2 polymorphism tended to have lower dentin-Pb levels and
performed better in areas of attention and executive functioning when compared to subjects with
the ALAD1 polymorphism.
Another gene of interest is the vitamin D receptor or VDR gene, which is involved in
calcium absorption through the gut. The variant VDR alleles may modify Pb concentrations in
bone and the rate of resorption and excretion of Pb over time. The relationship between the
VDR Fokl polymorphism and blood Pb concentrations was evaluated in 275 children enrolled in
the Rochester Longitudinal Study. A significant interaction was found between floor dust-Pb
loading and VDR-Fokl genotypes on blood Pb concentration, with the FF genotypes (a marker
for increased calcium absorption) having the highest adjusted mean blood Pb concentrations at
2 years of age compared to children with Ff or ff genotypes. High prevalence of FF genotypes in
African-American children, compared to non-African American children, may partially explain
higher blood Pb concentrations often observed in African-American children. There have been
no studies to indicate which, if any, of the VDR polymorphisms are associated with increased
vulnerability to the neurodevelopmental toxicity of Pb. Animal toxicology studies have yet to
identify any role of genetic polymorphism in ALAD or VDR in affecting Pb toxicity.
Tiffany-Castiglioni et al. (2005), in an overview of genetic polymorphisms relating to
mechanisms of neurotoxicity, state that an understanding of the relationship among ALAD
polymorphisms, blood Pb levels, and Pb neurotoxicity is difficult at this time. They further note
that, though urinary ALA is a good marker for Pb exposure, it may not correlate with neuronal
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damage. There is a similar lack of information using animal models to characterize genetic
polymorphisms of the VDR and hemochromatosis genes.

Gender
Most surveys find that boys have higher blood-Pb levels than girls; yet the data are less
clear with regard to gender-related differences in Pb-associated neurodevelopmental effects.
As discussed in Section 6.2.10, a greater male vulnerability was seen in the Cincinnati Lead
Study at various assessments from birth to adolescence. Also, data from a cross-sectional study
in England showed more pronounced Pb-IQ deficit associations for boys at 6 years of age.
However, in a study of 764 children in Taiwan, the relationship between Pb exposure and IQ
scores was much stronger in girls; and, in the Port Pirie, Australia cohort study, Pb effects on
cognition were significantly stronger in girls at ages 2, 4, 7, and 11-13 years.
In the Cincinnati Lead Study (see Section 6.2.3.1), an extensive neuropsychological
battery administered to 15-17 year old subjects examined executive functions, attention,
memory, achievement, verbal abilities, visuoconstructional skills, and fine-motor coordination as
key endpoints. About 30% of the subjects had blood-Pb concentrations >25 |ig/dL during the
first 5 years of life, and 80% of the cohort had at least one blood Pb > 15 |ig/dL. A strong
"executive functions" factor did not emerge from a factor analysis of scores. However, the
analysis, following covariate-adjustment, revealed strong associations between Pb exposure and
the attention factor for males. This gender interaction suggests that neuromechanisms sub-
serving attention were affected by Pb in this cohort for boys but not for girls. This is not
surprising, given the heightened vulnerability of males for a wide range of developmental
perturbations. A substantial gender difference (greater incidence among boys) in the incidence
of Attention Deficit/Hyperactivity Disorder (ADHD) is well established, and one could speculate
that early exposure to Pb exacerbates a latent potential for such problems.
The Port Pirie, Australia cohort study examined relationships between Pb exposure and
emotional and behavioral problems at age's 11-13 years after stratifying the data set by gender.
Stronger associations with Pb were observed for externalizing behavior problems in boys versus
girls; but greater internalizing behavior problems were found for girls.
Early laboratory animal Pb toxicology studies did not evaluate gender differences in
responses to chronic or acute Pb exposure, with the exception of several that showed differences
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in social investigatory behavior and nonsocial activity. Some studies pointed to greater social
investigatory behavior in males compared to females. More recent work by Cory-Slechta and
colleagues (Section 5.3.1.7) has shown greater synergistic effects of Pb and stress in female rats,
coupled with permanently elevated corticosterone levels. Also, maternal Pb exposure and
restraint stress caused greater changes in operant behavior and stress responses in female
offspring. These studies point to clear gender differences in response to Pb and suggest possible
hypothalamic-pituitary-adrenal axis-modulated effects of Pb on CNS function.

Socioeconomic Status
Epidemiologic studies have shown that Pb exposure is typically higher among low
socioeconomic status (SES) children compared to other U.S. children. Chronic stress and
consequent increased levels of glucocorticoids are also associated with low SES. Cory-Slechta
et al. (2004) have pointed out that both elevated glucocorticoids and Pb can cause similar
behavioral changes and that both impact the mesocorticolimbic systems of the brain.
As discussed in Section 5.3.7, their data indicate a potential mechanism whereby Pb exposure
enhances susceptibility to cognitive deficits and disease states.

8.4.2.8 Persistence/Reversibility of Neurotoxic Effects from Lead Exposure
Much of the classic Pb poisoning literature substantiates well the persistence of serious
neurological damage resulting from extremely high Pb exposures. The persistence of more
subtle, but important, neurotoxic effects of lower level Pb exposure has been accorded much
attention during the past decade or so. Much of the pertinent human and animal data seem to
suggest that the neurotoxic effects of Pb may not generally be reversible. As noted in Chapter 6,
excessive accumulation of Pb in childhood has latent and/or persistent adverse health effects on
both the peripheral and central nervous systems of human adults assessed 19-29 years later.
Also, chelation studies in humans and animals (summarized in Tables AX6-2.10 and AX5-3.6)
show that chelation decreases total body Pb burden, but does not necessarily exert evident effects
on Pb-induced cognitive deficits. For example, the extensive multi-center TLC study
summarized by Rogan et al. (2001) indicates that medical interventions involving chelation
therapy (e.g., Succimer use) do not seem to fully reverse cognitive deficits associated with early
Pb exposure. Also, nonhuman primate studies evaluated the persistence of effects by limiting the
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Pb exposures to the first year of life, as discussed in the 1986 AQCD and more recently (see
Section 5.3.5). In these monkeys, deficits in performance of both spatial discrimination tasks
and delayed spatial alternation were seen up to 8 years post exposure, when blood Pb had
dropped to control levels. In one study, however, the Pb-treated monkeys performed better than
control subjects at 4 years of age. In addition, a few studies discussed in the 1986 Lead AQCD
and some more recent studies have suggested possible reversibility of observed Pb-induced
learning deficits. Such studies suggest that reversibility depends on the age of the organism at
the time of exposure, the exposure duration, dosage, and other exposure parameters. Also,
several animal studies (Section 5.3.5) demonstrate that environmental enrichment during
development may, at times, help to engender some recovery from cognitive effects of earlier
short-term low-level Pb exposures.
Davis et al. (1990), however, sound a cautionary note with regard to the interpretation of
neurobehavioral data in light of compensatory capacities of the nervous system. They note that
compensatory capacities may become overwhelmed with aging, concurrent disease state, stress
due to socioeconomic status, or other stressors. It may be only then, possibly decades following
earlier Pb exposure, that some Pb-induced neurobehavioral effects are manifested.

8.4.2.9 Summary of Toxicologic and Epidemiologic Evidence of Lead-
Induced Neurotoxicity
Findings from numerous experimental studies of rats and of nonhuman primates, as
discussed in Chapter 5, parallel the observed human neurocognitive deficits and the processes
responsible for them. Learning and other higher order cognitive processes show the greatest
similarities in Pb-induced deficits between humans and experimental animals. Deficits in
cognition are due to the combined and overlapping effects of Pb-induced perseveration, inability
to inhibit responding, inability to adapt to changing behavioral requirements, aversion to delays,
and distractibility. Higher level neurocognitive functions are affected in both animals and
humans at very low exposure levels (< 10 |ig/dL), more so than simple cognitive functions.
For example, the discrimination reversal paradigm is a more sensitive indicator of Pb-induced
learning impairment than simple discrimination. Many studies suggest that most Pb-induced
cognitive deficits are very persistent and that animals remain vulnerable to the effect of Pb
throughout development. Some studies, however, suggest that environmental enrichment during
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early development may confer some offsetting protection against Pb-induced cognitive effects or
that other factors (e.g., short-lived exposure duration/low concentration) may, at times, induce
detectable but transient cognitive deficits. Also, more evidence is emerging that substantiates
Pb-induced attentional deficits, which may contribute to persisting cognitive dysfunction, poorer
academic performance, and/or maladaptive anti-social behavior patterns (e.g., delinquency).
Other behavioral endpoints (e.g., social behavior, aggression, and locomotor activity)
evaluated in animal studies in relation to Pb exposure did not clearly indicate Pb-induced
impairments. This may be due to the lack of effect with low-level Pb exposure or to variables
(e.g., nutrition, age, gender, and strain) possibly not well controlled for experimentally.

8.4.3 Cardiovascular Effects
Epidemiologic studies that have examined the effects of blood-Pb levels on blood
pressure have generally found positive associations, even after controlling for confounding
factors such as tobacco smoking, exercise, body weight, alcohol consumption, and
socioeconomic status (discussed in Section 6.5.2). Recent meta-analyses of these studies have
reported robust, statistically-significant, though small effect-size, associations between blood-Pb
concentrations and blood pressure. For example, the meta-analysis of Nawrot et al. (2002)
indicated that a doubling of blood Pb corresponded to a 1 mm Hg increase in systolic blood
pressure. Although this magnitude of increase is not clinically meaningful for an individual, a
population shift of 1 mm Hg is important. The majority of the more recent studies employing
bone-Pb level have also found a strong association between long-term Pb exposure and arterial
pressure. Since the residence time of Pb in blood is relatively short but very long in bone, the
latter observations have provided compelling evidence for the positive relationship between Pb
exposure and a subsequent rise in arterial pressure in human adults.
Numerous experimental animal studies have shown that exposure to low levels of Pb for
extended periods results in an eventual onset of arterial hypertension (HTN) that persists long
after the cessation of Pb exposure in genetically normal animals. Many studies have been
conducted to explore the mechanisms by which chronic Pb exposure may cause HTN. Most of
these studies have examined various blood-pressure regulatory and vasoactive systems in animal
models of Pb-induced HTN. A number of studies have also utilized in vitro cell culture systems
such as endothelial and vascular smooth muscle cells to gain insight into molecular mechanisms
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implicated in this process. Key findings have emerged from the newly available in vivo and in
vitro studies of mechanisms including the following several important points.
During the past decade, several studies have shown that Pb exposure causes oxidative
stress, particularly in the kidney and cardiovascular tissues, as well as in cultured endothelial and
vascular smooth muscle cells (VSMC), as noted in Section 5.5.2.1. The in vivo studies have
further shown that Pb-induced oxidative stress is, at least in part, responsible for associated
hypertension (HTN) in experimental animals. Khalil-Manesh et al. (1994) were among the first
to suggest that oxidative stress may be involved in the pathogenesis of lead-induced HTN.
Gonick et al. (1997) later provided evidence for the occurrence of oxidative stress and
compensatory up regulation of NOS isotypes in the kidney of animals with lead-induced HTN.
Studies carried out with antioxidants, e.g., lazaroid compound, resulted in a significant
alleviation of oxidative stress, improved NO availability, and a marked attenuation of HTN
without affecting blood Pb concentration, further demonstrating that Pb-induced HTN is
associated with diminished NO availability and that the latter was mediated by oxidative stress
(Vaziri et al., 1997).
Numerous in vivo and in vitro studies on Pb-induced HTN, using endothelial and VSMC
with or without intervention by antioxidant therapeutics, suggest a role of oxidative stress and
NO in the pathogenesis of lead-induced HTN in the rat.
These observations provided compelling evidence that Pb-induced HTN causes oxidative
stress, which, in turn, promotes functional NO deficiency via ROS-mediated NO inactivation.
The latter, in turn, participates in the development and maintenance of HTN and cardiovascular
abnormalities. Also, the formation of the highly cytotoxic reactive nitrogen species peroxynitrite
(ONOO) from the NO-ROS interaction and the associated nitrosative stress could potentially
contribute to long-term cardiovascular, renal, and neurological consequences of Pb exposure.
Higher plasma levels of lipid peroxides in uncontrolled essential hypertension as
compared to normal controls suggested that free radicals are involved in the pathobiology of
human essential hypertension. Angiotensin II, a potent vasoconstrictor, was found to stimulate
free radical generation in normal leukocytes. This increase in free radical generation was
thought to inactivate NO, and possibly prostacyclin, which can lead to an increase in peripheral
vascular resistance and hypertension.
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In spite of such a wide range of experimental investigations into the cardiovascular effects
of Pb in animal studies, it is still not clear as to why low, but not high, levels of Pb exposure
cause HTN in experimental animals.

8.4.4 Heme Synthesis and Blood Effects
Lead exposure has long been recognized to be associated with disruption of heme
synthesis in both human children and adults. With extreme Pb exposure leading to blood-Pb
levels >30 |ig/dL, Pb-induced heme synthesis interference leads to notable reductions in
hemoglobin synthesis and, at blood-Pb >40 |ig/dL, to frank anemia (a classic clinical sign of
severe Pb poisoning).
Other indications of disruption of heme synthesis are readily detectable at distinctly lower
blood-Pb concentrations, but mainly tend to serve as highly useful biomarkers of Pb exposure.
Elevated blood-Pb concentrations of-20-30 |ig/dL, for example, are sufficient to halve
erythrocyte ALAD activity and sufficiently inhibit ferrochelatase so as to double erythrocyte
protoporphyrin (EP) levels. Erythrocyte ALAD activity ratio (the ratio of activated/non-
activated enzyme activity) has been shown to be a sensitive, dose-responsive measure of Pb
exposure, regardless of the mode of Pb administration. Competitive enzyme kinetic analyses in
RBCs from both humans and cynomolgus monkeys indicated similar inhibition profiles by Pb.
Decreased ALAD activity in rat RBCs have been reported at blood Pb levels of 10 |ig/dL.
The effects of various metals, including Pb, on RBC porphobilinogen synthase (PBG-S)
have been studied using human RBC hemolysate (see Section 5.2.3). Effects on the enzyme
were found to depend on the affinity of the metal for thiol groups at its active sites. Additional
studies utilizing rabbit erythrocyte PBG-S indicate that Pb acts as a potent effector of this
enzyme both in vitro and in vivo. Increased erythrocyte protoporphyrin levels seen at blood-Pb
concentrations > 15 |ig/dL represent another widely used biomarker of Pb exposure.
Comparison of pyrimidine-5'-nucleotidase (P5N) and deoxypyramidine-5-nucleotidase
levels in the RBC of Pb-exposed workers and matched controls also showed significantly lower
levels of P5N in Pb-exposed workers. Similar observations were reported for neonatal rat RBCs,
with the low levels of nucleotides being hypothesized to be due to inhibition of P5N activity by
Pb, as the depression in enzyme activity was correlated with blood-Pb levels (see Section 5.2.5).
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8.4.5 Renal System Effects
The nephrotoxic effects of Pb are mediated by alterations in the glomerular filtration rate
(GFR). A battery of tests used to screen both environmentally- and occupationally-exposed
individuals often include: (1) measures of glomerular integrity, (2) tubular absorption and
secretion, (3) measure of tubular integrity, (4) measure of glomerular and distal tubular function,
(5) glomerular structural proteins, and (6) measure of distal tubular function. Numerous new
epidemiologic studies discussed in Chapter 6 provide important new findings on associations
between Pb exposure and impacts on renal function (see Section 6.4).
Of particular importance are new analyses of associations between blood-Pb and renal
outcomes in 15,211 adult subjects in the NHANES III study (conducted during 1988-1994).
Dichotomous renal outcome measures analyzed included elevated serum creatinine and chronic
kidney disease. Mean blood-Pb level was 4.2 |ig/dL among the 4,813 hypertensives and
3.3 |ig/dL in normotensives, with prevalence of elevated serum creatinine being higher among
hypertensives than nonhypertensives but prevalence of chronic kidney disease being similar.
The authors noted that (1) the associations were strong, dose-dependent and consistent before
and after adjustment (e.g., for age, race, and gender) and (2) higher blood Pb was associated in
nonhypertensives with higher prevalence of chronic kidney disease in diabetics. This study is
notable for sample size, comprehensive adjustment for other renal risk factors, and the fact that
the study population is representative of the general U.S. population. In another study of
820 women (ages 53-64 years) in Sweden, significant negative associations were seen between
blood Pb (mean blood Pb of 2.2 |ig/dL) and both glomerular filtration rate (GFR) and creatinine
clearance, an association that was apparent over the entire dose range. This study also had the
additional advantage of blood and urinary cadmium assessment.
The above studies and other general population studies constitute some of the most
important types of research on Pb renal effects during the past 20 years, as discussed in Section
6.4.4.1. Overall, a number of strengths are present in this body of literature. These include study
design with longitudinal data in some studies; large populations in both the United States and
Europe; comprehensive assessment of Pb dose (including use of bone Pb as a measure of
cumulative Pb body burden in some studies); and statistical approaches using a range of
exposure and outcome measures, while adjusting for numerous renal risk factors. Associations
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between Pb exposure and worse renal function were observed in most of the general population
studies.
Residual confounding and reverse causality have both been proposed as alternative
explanations for the reported associations between Pb and renal dysfunction. As discussed in
Section 6.5, increased blood pressure has been associated with Pb exposure in general
populations. Adjustment for hypertension or blood pressure, although typical in Pb-renal
studies, carries the risk of underestimating the actual slope of the association between Pb dose
and renal dysfunction. Given the careful adjustment for confounding in the Pb-renal general
population literature, it is thought that residual confounding is not a likely explanation for the
observed Pb-renal dysfunction associations. Reverse causality, i.e. attributing increased Pb dose
to reduced Pb excretion as a consequence of renal insufficiency, is another possible explanation
posed to explain such associations. However, by examining temporal relationships between Pb
dose and renal function in longitudinal studies, it has been convincingly shown that Pb dose
predicts decline in renal function. Other evidence against possible reverse causality is the
positive impact of Pb chelation on renal function (see Section 6.4.4.3), although the possibility of
a direct beneficial effect of chelating agents on renal function cannot be ruled out.
Increased risk for nephrotoxicity has been observed at the lowest Pb exposure levels
studied epidemiologically to date. More specifically, the newly available general population
studies have shown associations between blood Pb and indicators of renal function impairment at
blood-Pb levels extending below 10 |ig/dL, with nephrotic effects having been reported among
some adults with mean concurrent blood-Pb levels as low as ~2 to 4 |ig/dL. However, the data
available to date are not sufficient to determine whether nephrotoxicity is related more to such
current blood-Pb levels, higher levels from past exposures, or both. An association between
cumulative Pb dose (indexed by mean tibia Pb of 21.5 jig/g bone mineral) and longitudinal
decline in renal function has been observed as well. Blood Pb levels <10 |ig/dL have also been
associated with altered creatinine clearance, as noted in Chapter 6 (Figure 6-8). Slopes ranged
from 0.2 to -1.8 mL/min change in creatinine clearance per each |ig/dL increase in blood Pb.
Animal toxicology studies reported that both low and high dose Pb-treated animals
showed a "hyperfiltration" phenomenon during the first 3 months of Pb exposure. This finding
could be invoked as a partial explanation for late changes of glomerulosclerosis seen in high-Pb
dose animals but cannot explain the lack of glomerular changes in the low-dose animals. These
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results support observations by several investigators in humans, leading some to argue that Pb
nephropathy should be added to diabetic nephropathy as diseases that lead to early
hyperfiltration. Also, animal toxicology studies that evaluated biochemical alterations in Pb-
induced renal toxicity suggest a role for oxidative stress and involvement of NO, with a
significant increase in nitrotyrosine and substantial fall in urinary excretion of NOX.
A few animal toxicology studies that evaluated the effect of coexposure to other metals
indicated that cadmium increases Pb in blood when both are given, but diminishes Pb in liver
and kidney. Selenium, an antioxidant, improves both parameters, as does thiamine or L-lysine
plus zinc. Iron deficiency increases intestinal absorption of Pb and the Pb content of soft tissues
and bone. Aluminum decreases kidney Pb content and serum creatinine in Pb-intoxicated
animals. Age also has an effect on Pb retention. There is higher Pb retention at a very young
age but lower bone and kidney Pb at old age, attributed in part to increased bone resorption and
decreased bone accretion and kidney Pb.
The above findings appear to indicate likely associations between some indicators of
altered kidney function (e.g., increased creatinine clearance) at relatively low blood-Pb levels
among the general population. However, the potential public health significance of such
findings is difficult to discern just yet. This is especially true in light of difficulties in resolving
discrepancies between these newly reported findings in the general population studies versus
observation among the more than 10,000 occupationally-exposed workers studied of notable Pb
effects on renal tubular function only when blood-Pb exceeds distinctly higher levels (e.g.,
>30-40 |ig/dL).

8.4.6 Lead-Associated Immune Outcomes
The effects of Pb exposure on the immune system of animals are described in Section 5.9
and are summarized in Figure 5-18. These include the targeting of T cells and macrophages
by Pb. Lead-induced alterations center on an increased inflammatory profile for macrophages
(i.e. elevated tumor necrosis factor-alpha, oxygen radical, and prostaglandin production) and a
skewing of the T cell response away from T helper 1 (Thl)-dependent functions toward
T helper 2 (Th2)-dependent functions. The resulting immune changes include an increased
production of Th2 cytokines (e.g., IL-4, IL-10) and certain immunoglobulins [e.g.,
immunoglobulin E (IgE)]. Concomitantly, there is a decrease in Thl-associated cytokines
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(interferon gamma and IL-12) and in Thl-functions, e.g., the delayed type hypersensitivity
(DTK) response. Significant age-related differences in immunotoxic sensitivity to Pb (based on
blood-Pb concentrations) approximate an order of magnitude difference in sensitivity between
the perinatal period and adulthood (see Table 5-10). Importantly, immune changes are
associated with blood-Pb levels well below 10 |ig/dL following gestational or perinatal
exposures (see Table 5-9). However, major immune cellular alterations are not a hallmark of
low-level Pb exposure, despite significant Pb-induced shifts in immune function. This lack of
major immune cell population changes becomes important for interpretation of human
epidemiologic results.
Human epidemiologic immune evaluations are hampered by the reality that the most
informative sources of functionally-reactive immune cells (e.g., those responding to antigens in
the lymphoid organs and local lymph nodes) are not available for routine human sampling. This
can be important in considerations of early-life associated immunotoxicity where functional
assessment of immune changes appear to be particularly important, as noted by Dietert and
Piepenbrink (2006). Instead, circulating lymphocytes and serum or plasma immunoglobulin
levels in humans must serve as easily accessible surrogates for a more comprehensive
determination of immune status. Despite this inherent limitation, the animal and human data for
Pb-induced immune alterations are in general agreement, including the association of blood-Pb
levels below 10 |ig/dL with significant neonatal/juvenile immune alterations.
The sentinel result suggesting that low-level Pb exposure produces similar immune
changes among animals and humans is the positive association of blood-Pb levels with IgE level.
This association has been observed at blood-Pb levels <10 |ig/dL following early life exposure in
both humans (Section 5.9.3.2) and animals (Section 5.9.3). Other animal studies also support
this by showing low-level Pb-induced increases in neonatal/juvenile IL-4, the hallmark cytokine
modulating IgE production. Similarly, in the adult human, a positive association between blood-
Pb level and IgE level has been reported for occupationally-exposed workers. It is not surprising
that human epidemiologic results showed less consistent changes in other immunoglobulins,
since Th biasing would be expected to produce shifts among immunoglobulin G (IgG) subclasses
without necessarily changing overall IgG concentrations, consistent with results in the rat.
It should be noted that, prior to 1992, no human epidemiology study involving Pb
reported comparisons of IgE levels. Also, since IgE is a minor immunglobulin component of
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human serum, several studies since 1992 did not include IgE quantitation in the evaluation.
The animal data suggest that IgE (as well as the supporting cytokine, IL-4) are among the most
sensitive parameters for modulation following low-level Pb exposure. Thus, in retrospect, those
human studies that did not evaluate IgE levels may have focused on Pb-insensitive immune
parameters.
Cell-mediated immunity in animals evaluated by the Th-dependent DTH reaction in most
animal studies (see Section 5.9.4) showed that this measure was particularly sensitive to Pb-
induced immunosuppression. In humans, the primary surrogate for cell-mediated immunity was
a non-functional measure of circulating leukocyte populations. Despite this difficulty in
evaluation, a majority of studies (see Table 6-6) that quantitated T, Th, Tc, and B cells reported
decreases in either T or Th cells relative to an increase in circulating B cells. This is consistent
with the profile described in the animal studies, where Th promotion of cell mediated immune
function is impaired by Pb exposure while humoral immunity remained either unchanged or
displayed increased IgE production.
Numerous animal studies reported that Pb produced elevated levels of TNF-alpha,
superoxide anion and prostaglandins while depressing production of nitric oxide by macrophages
(see Section 5.9.6). To the extent the same endpoints have been examined, the results are similar
between animals and humans. One study has reported that in vitro-activated monocytes from
Pb-exposed children were depressed in nitric oxide production, contrasting with a positive
association between blood-Pb level and production of superoxide anion. Based on these results,
the pattern of Pb-induced changes in major macrophage metabolites appears to be similar
between the animal experimental and human epidemiological data.
Comparison of the human and animal studies is quite feasible and is limited only by the
number of studies that incorporated comparable immune endpoints. In retrospect, several prior
human epidemiological studies measured endpoints that appear to be Pb-insensitive based on the
most recent animal data. Of the studies that evaluated similar parameters, the results are
strikingly in agreement.
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8.4.7 Reproduction and Development Effects
The majority of the experimental animal studies of Pb effects on reproduction and
development examined effects due to inorganic forms of Pb, with very little being known about
reproductive and developmental effects of organic Pb compounds.
Timing of exposure has been found to be critical to Pb-induced male reproductive toxicity
in rats. Studies conducted in nonhuman primates support the importance of exposure timing and
indicate that the adverse effects of Pb on male reproduction are dependent upon age (i.e.,
developmental stage at time of exposure) and duration of exposure. Numerous more recent
studies conducted in experimental animals support the earlier findings that Pb exposure during
early development can delay the onset of male puberty and alter reproductive function later in
life (see Section 5.4.2.1).
Other recent research supports the conclusion that mechanisms for endocrine disruption in
males involve Pb acting at multiple sites along the hypothalamic-pituitary-gonadal (HPG) axis.
However, variable findings regarding specific types of Pb effects have been attributed to
complex mechanisms involved in hormone regulation and the multiple sites of action for Pb.
It has been suggested that differences in results among studies may, in part, be attributed to an
adaptive mechanism in the hypothalamic-pituitary-gonadal axis that may render the expression
of some toxic effects dependent on dose and exposure duration (see Section 5.4.2). Thus,
adaptive or multiple effects on the HPG axis having different dose-duration-response
relationships may explain apparent inconsistencies among reported Pb effects on circulating
testosterone levels, sperm count, and sperm production.
A possible mode of action for Pb-induced testicular injury is oxidative stress (as discussed
in Section 5.4.2.4). Pb-induced oxygen free radical generation has been suggested as a plausible
mechanism of testicular injury in primates. This oxygen radical hypothesis is supported by
studies conducted in rodents; and the oxidative stress hypothesis is supported by observations of
increases in the percentage of apoptotic cells in the testes of rodents in response to Pb exposure.
Several modes of action for Pb-induced, endocrine disruption-mediated, alterations in
female reproduction have also been proposed, as discussed in Section 5.4.3. These include
changes in hormone synthesis or metabolism at the enzyme level and changes in hormone
receptor levels. In addition, Pb may alter sex hormone release and imprinting during early
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development. The latter effects would be consistent with observations of persistent changes in
estrogen receptor levels in the uterus and altered ovarian LH function in Pb-exposed animals.
A persistent effect of maternal Pb exposure (blood Pb 30 to 40 |ig/dL) has been seen on
corticosteroid levels in adult offspring (as discussed in Section 5.4.6). Both male and female
offspring born to dams exposed to Pb exhibited elevated corticosteroid levels as adults.
In female offspring, the Pb effect was potentiated when maternal Pb exposure occurred in
combination with environmental stress (administered as restraint). The interplay between Pb and
stress hormones is consistent with other animal toxicologic findings wherein neonatal exposure
to Pb (blood Pb 70 |ig/dL) decreased cold-water swimming endurance (a standard test for stress
endurance).
The literature provides convincing support for Pb-induced impairment of postnatal
growth, as discussed in Section 5.4.5. Although some early studies ascribed the reduction in
postnatal growth to reduced food consumption (suggesting an effect of Pb on satiety
mechanisms), more recent studies report impaired growth unrelated to changes in food
consumption. These and other findings suggest that Pb exposure may impair growth through a
mechanism that involves a suppressed pituitary response to hypothalamic stimulation. The
mechanism may involve Pb-induced reduction in plasma concentrations of IGFi.

8.4.8 Bone and Teeth Effects
Lead is readily taken up and stored in the bone of experimental animals, where it can
potentially manifest toxic effects that result in stunted skeletal growth. In experiments reported
since the 1986 Lead AQCD (see Section 5.8.3), uptake and retention of Pb were determined in
bone from rats exposed to plain water or water containing Pb-acetate (41.7 to 166.6 mg/L) for
12 to 16 weeks. After 4 weeks, the skeletal Pb in animals receiving the lowest dose was almost
5 times higher than in control animals (5.9 versus 1.2 jig Pb/g bone, respectively). Lead levels in
bones from animals receiving 83.3 mg/L and 166.6 mg/L were dose-dependently higher (at 11.7
and 17.0 jig Pb/g bone, respectively) after 4 weeks of exposure.
Results from several new animal studies have also yielded evidence for Pb exposure
adversely affecting bone growth and density, which may be potentially manifested through Pb
interference with growth and hormonal factors as well as toxic effects directly on bone. One of
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the studies suggested Pb was mediating its effect through 1, 25-(OH)2D3, rather than via a direct
action on the Calbindin-D protein.
The fact that Pb exposure has been associated with altered bone metabolism and
decreased growth and skeletal development is suggestive of potential Pb perturbation of one or
more endocrine factors, e.g., growth hormone. However, overall, available rat studies suggest
that differences in growth seen with Pb exposure may not necessarily be due to alterations in
secretion of growth hormone. Rather, effects on calcium uptake and/or metabolism may be more
crucial, as suggested by the results of several in vitro studies. The results suggest that the
calcium-ATPases of intracellular stores are potentially poisoned by Pb entering the cells.
An invaluable method to explore the kinetics of Pb transfer from bone to blood has been
developed and evaluated within the last decade (see Section 5.8.6). The method uses recent
administration of sequential doses of Pb mixes enriched in stable isotopes (204Pb, 206Pb, and
207Pb) to female cynomolgus monkeys that were earlier chronically administered a common Pb
isotope mix (1,300 to 1,500 jig Pb/kg body weight/day for > 10 years). The stable isotope mixes
serve as a marker of recent exogenous Pb exposure, whereas the chronically administered
common Pb serves as a marker of endogenous (principally bone) Pb. It was found that
administration of the first isotope label allows measurement of the contribution of historic bone
stores to blood Pb. Exposure to subsequent isotopic labels allows measurement of contributions
from historic bone Pb stores and the recently administered enriched isotopes that incorporated
into bone. In general, the contribution from historic bone Pb (common Pb) to blood Pb level was
relatively constant (-20%), but was augmented by spikes in total blood Pb due to current
administration of the stable isotopes (blood Pb ranged from 31.2 to 62.3 jig/100 g).
Using the above sequential stable isotope administration method, another study examined
flux of Pb from maternal bone during pregnancy of 5 female cynomolgus monkeys previously
exposed to common Pb (-1,100 tol,300 jig Pb/kg body weight) for about 14 years. In general,
Pb levels in maternal blood (as high as 65 jig/100 g) attributable to Pb from mobilized bone
dropped 29 to 56% below prepregnancy baseline levels during the first trimester of pregnancy.
This was ascribed to the known increase in maternal fluid volume, specific organ enlargement
(e.g., mammary glands, uterus, placenta), and increased metabolic activity that occurs during
pregnancy. During the two later trimesters, when there is a rapid growth in the fetal skeleton and
compensatory demand for calcium from the maternal blood, the Pb levels increased up to 44%
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over pre-pregnancy levels. Blood-Pb levels in the fetus generally corresponded to those found in
the mothers, both in total Pb concentration and in the proportion of Pb attributable to each
isotopic signature dose. From 7 to 25% of the Pb found in fetal bone originated from maternal
bone, with the balance derived from oral dosing of the mothers with isotope during pregnancy.
Of interest, in offspring from a low Pb exposure control monkey (blood Pb <5 jig/100 g) -39%
of Pb found in fetal bone was of maternal origin, suggesting enhanced transfer and retention of
Pb under low Pb conditions.
These studies show that Pb stored in bone is mobilized during pregnancy and lactation,
exposing both mother and fetus/nursing infant to potentially toxic blood/milk Pb levels. Also of
much concern, a significant proportion of Pb transferred from the mother is incorporated into the
offspring's developing skeletal system, where it can serve as a continuing source of toxic
exposure. The latter study illustrates the utility of sequentially administered stable isotopes in
pregnancy; however, its use may also be applicable in studies of lactation, menopause,
osteoporosis, and other disease states where mobilization of bone and release of Pb stores occurs.
Further, given that isotopic ratios of common Pbs vary by location and source of exposure, when
humans migrate from one area and source of exposure to another, it is possible to document
changes in mobilized Pb, especially during times of metabolic stress.
During pregnancy, transfer of Pb from mother to offspring has been documented. Still,
other available evidence also suggests that a more significant transfer from mother to offspring
occurs during lactation, when the Pb concentration in mother's milk can be several times higher
than corresponding blood-Pb levels.

8.4.9 Hepatic and Gastrointestinal System Effects
A large body of experimental animal toxicology database reviewed in this document
indicated hepatotoxic effects, including liver hyperplasia, at very high dose Pb exposures. Based
on the limited data available in these toxicology studies on blood-Pb levels, inhibition of liver
heme synthesis and inhibition of liver ALAD was reported at 15-20 |ig/dL, whereas alterations
in liver cholesterol metabolism and induction of hepatic oxidative stress was observed at
20-30 |ig/dL.
Studies of hepatic enzyme levels in serum suggest that liver injury may be present in lead
workers; however, associations specifically with Pb exposures were not evident. Also, children
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exposed to relatively high Pb levels (blood-Pb >30 |ig/dL) exhibit depressed levels of circulating
1,25-dihydroxy vitamin D (1,25-(OH)2D3). However, associations between serum vitamin D
status and blood Pb were not evident in a study of calcium-replete children who had average
lifetime blood-Pb concentrations <25 |ig/dL.
Investigations into the potential molecular mechanisms involved in these alterations
suggest induction of gene expression for CYP51 (Lanosterol 14a-demethylase), an essential
enzyme for cholesterol biosynthesis, in Pb-nitrate-induced liver hyperplasia, although other
cytochrome P450 enzymes involved in drug metabolism have been reported as being suppressed.
The induction of the cytokines interl eukin-la and TNF-a in rat liver prior to the induction of the
genes for these synthesis enzymes suggested that Pb-nitrate-induced cholesterol synthesis is
independent of sterol homeostasis regulation.
The effect of low-concentration Pb-acetate (0.1%) on the jejunal ultrastructure has been
studied in young male rats. The villi of jejunum of rats exposed to Pb for 30 days had a rough
appearance on the surface, which could be associated with a distortion of glycocalyx layer.
Areas of extensive degenerative lesions were also observed on the surface of most villi on the
60th day of exposure. All intestinal epithelial cells exhibited various degrees of glycocalyx
disturbance, indicating that pronounced toxic effects of Pb were related to modifications of the
biochemical properties of the surface coat of the cells.

8.4.10 Genotoxicity and Carcinogenicity
One study has investigated the carcinogenicity of a series of chromate compounds, i.e.,
Pb-chromate and several Pb-chromate-based compounds. The authors indicated that in this
design, Pb-chromate was not carcinogenic, but that 4 of the Pb chromate compounds did induce
a very rare tumor in the mice. The remaining five studies focused on Pb-acetate. In most
studies, this compound was administered in drinking water at concentrations from 0.5 to
4000 ppm, but one study considered effects from a subcutaneous (SC) injection both in mice and
in rats. Consistent with the findings in the 1986 Lead AQCD, Pb not only induced renal tumors,
but also induced other tumors (e.g., pituitary, thyroid, testicular), although the possible effect on
mammary tumors is difficult to interpret.
Overall, the above studies confirm that Pb is an animal carcinogen and extend our
understanding of mechanisms involved to include a role for metallothionein. Specifically, the
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recent data show that metallothionein may participate in Pb inclusion bodies and, thus, serves to
prevent or reduce Pb-induced tumorigenesis. Much more work is needed to determine the
potential exacerbating or ameliorating roles of calcium and selenium and to determine what role
Pb-induced immunomodulation may play in the promotion of tumors.
The data currently seem to indicate that Pb can induce anchorage independence in human
cells, but its ability to induce neoplastic transformation of human cells is uncertain. Further
study of different Pb compounds and the full assessment of their neoplastic potential (i.e.,
including studies of the ability of treated cells to form tumors in experimental animal models) are
needed before definitive conclusions can be drawn.
All together, animal cell culture studies suggest that Pb ions alone cannot transform
rodent cells; however, they may be co-carcinogenic or promote the carcinogenicity of other
compounds such as chromate.
The proliferative effects of various Pb salts (i.e., Pb-acetate, Pb-chloride, Pb-monoxide,
Pb-sulfate), have been evaluated, using liver-derived REL cells. All the Pb compounds tested
showed dose- and time-dependent effects on the proliferation of REL cells. Unlike other tumor
promoters, Pb compounds did not exhibit effects on cell junctional coupling. Liver hyperplasia
induced by Pb-nitrate has been shown to demonstrate sexual dimorphism in all phases of the
proliferation as well as in apoptosis.
Investigations of cell cycle-dependent expression of proto-oncogenes in Pb-nitrate
(10 |iM/100 g body wt)-induced liver cell proliferation showed that peak DNA synthesis
occurred at 36 h after a single injection of Pb-nitrate. In addition to DNA synthesis, Pb-induced
expression of c-fos, c-myc, and c-Ha-ras oncogenes was also observed in rat liver tissue.
Additional studies by the same group reported that Pb-nitrate-induced liver hyperplasia involved
an increased expression of c-jun in the absence of c-fos expression.
Differential activation of various PKC isoforms, down regulation of PKC-a, and marked
activation of PKC-e in Pb-nitrate-mediated liver hyperplasia suggested the involvement of these
PKC enzymes in DNA synthesis and related signal transduction pathways.
The majority of studies on genotoxicity of Pb compounds in animal models focused on
mice. Lead was administered by intraperitoneal (IP) or intravenous (IV) injection. Several
endpoints were considered, including chromosome aberrations, SCE, micronucleus formation,
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and DNA strand breaks. Overall, the results are ambiguous, due in part to study design and the
various endpoints considered. The results for SCE are consistently positive.
The potential mutagenicity of Pb compounds in rodent cells was evaluated by using three
mutagenesis systems: mutagenesis at the HPRT locus, the gpt locus, and mutations in sodium-
potassium ATPase. The results are highly variable and may be specific to the Pb compound
considered in each case. In particular, Pb-chromate and Pb-acetate appear to be nonmutagenic.
Lead acetate was positive but only at highly cytotoxic concentrations. By contrast, Pb-chloride
and Pb-sulfate appeared to be mutagenic at relatively nontoxic concentrations. Insufficient data
exist at this point to conclude whether or not Pb is mutagenic in animal cells.
Both Pb-chromate and Pb-nitrate induced DNA-protein crosslinks in cultured mammalian
cells. These data suggest that Pb is genotoxic in this manner; however, it is thought that the
Pb -chromate-induced DNA-protein crosslinks result from the chromate.
It is plausible that through this mechanism, Pb may act as a co-carcinogen by affecting the
metabolism of other chemicals or possibly as a direct carcinogen by enhancing endogenously-
induced damage. However, no studies have directly shown that such Pb effects are linked to
cancer or alter the potency of another chemical; and, thus, it remains only a plausible hypothesis.
Lead has been classified by IARC as a probable human carcinogen, based mainly on a
judgment that there is sufficient animal evidence. This classification is consistent with the
National Toxicology Program's Carcinogens Review Committee Report, which recommended
that Pb and Pb compounds be considered "reasonably anticipated to be human carcinogens."
Ten rat bioassays and one mouse assay have shown statistically significant increases in renal
tumors with dietary and subcutaneous exposure to several soluble Pb salts. Animal assays
provide reproducible results in several laboratories and in multiple rat strains, with some
evidence of multiple tumor sites. Also, short-term studies show that Pb affects gene expression.
Similarly, Pb and Pb compounds would likely be classified as likely to be carcinogenic to
humans according to the 2005 EPA Guidelines for Carcinogen Risk Assessment.
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8.5 KEY LOW-LEVEL LEAD EXPOSURE HEALTH EFFECTS
AND IDENTIFICATION OF FACTORS THAT AFFECT
SUSCEPTIBILITY TO LEAD TOXICITY
The numerous new studies that have become available since the 1986 Lead
AQCD/Addendum and the 1990 Supplement, as discussed above, provide extensive new data on
health effects of Pb across a wide exposure range, including information on concentration-
response relationships for key Pb-related health effects. Of particular interest for present
purposes is the delineation of lowest observed effect levels for those Pb-induced effects that are
most clearly associated with blood Pb <10|ig/dL in children and/or adults and are, therefore, of
greatest public health concern. Tables 8-5 and 8-6 highlight the most important such effects
observed in children and adults, respectively, as discussed in the preceding subsections.
As evident from those discussions, neurotoxic effects in children and cardiovascular effects in
adults are among those best substantiated as occurring at blood-Pb concentrations as low as 5 to
10 |ig/dL (or possibly lower); and these categories of effects are currently clearly of greatest
public health concern. Other newly demonstrated immune and renal system effects among
general population groups are also emerging as low-level Pb-exposure effects of potential public
health concern.
The remarkable progress made since the mid-1980s in understanding the effects of Pb on
health can be gauged by noting changes that have occurred over time in the questions
investigators have addressed. In the 1980s, the question of interest was often: "Does low-level
lead exposure affect health?" The questions asked in recent studies have more often focused on
details of the associations, including shapes of concentration-response relationships (especially at
levels well within the range of general population exposures); biological and socioenvironmental
factors that either increase or decrease an individual's risk; prognoses pertinent to Pb-associated
effects, efficacy of interventions to reduce adverse effects, and so on. In fact, "low-level," a term
long-used to describe exposures that are not sufficiently high to produce clinical signs and
symptoms, is increasingly being recognized as a descriptor that has little biological meaning and
is interpretable only in a specific historical context. What was considered "low" in the 1980s is
an order of magnitude higher than the current mean blood Pb level in the U.S. population, and
the current mean remains perhaps as much as two orders of magnitude above preindustrial
"natural" background levels in humans. The current CDC screening guideline for children of
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Table 8-5. Summary of Lowest Observed Effect Levels for Key Lead-Induced Health Effects in Children
Lowest Observed Effect
Blood Lead Level
Neurological Effects
Hematological Effects
Immune Effects
30 (ig/dL
Increased urinary 5-
aminolevulinic acid
15 (ig/dL
oo
10 (ig/dL
Behavioral disturbances
(e.g., inattention, delinquency)

Altered electrophysiological
responses

Effects on neuromotor function

CNS cognitive effects
(e.g., IQ deficits)
Erythrocyte protoporphyrin
(EP) elevation
Inhibition of 5-aminolevulinic
acid dehydratase (ALAD)

Pyrimidine-5 '-nuclotidase
Effects on humoral (| serum IgE)
and cell-mediated (J, T-cell
abundance) immunity

5 (ig/dL \
(Py5N) activ
( \
ty inhibition
f
(???)* (???)*
0 (ig/dL
*Note: Arrows depict cases where weight of overall evidence strongly substantiates likely occurrence of type of effect in association with blood-Pb
concentrations in range of 5-10 ug/dL, or possibly lower, as implied by (???). Although no evident threshold has yet been clearly established for those
effects, the existence of such effects at still lower blood-Pb levels cannot be ruled out based on available data.
Source: Adapted/updated from Table 1-17 of U.S. Environmental Protection Agency (1986a).
-------
Table 8-6. Summary of Lowest Observed Effect Levels for Key Lead-Induced Health Effects in Adults
Lowest Observed Effect
Blood Lead Level
Neurological Effects
Hematological Effects Cardiovascular Effects
Renal Effects
30 (ig/dL
Peripheral sensory nerve
impairment
Erythrocyte
protoporphyrin (EP)
elevation in males
Impaired Renal Tubular
Function
oo
I
to
20 (ig/dL

15 (ig/dL
10 (ig/dL

5 p.g/dL

0 (ig/dL
Cognitive impairment
Postural sway
Erythrocyte
protoporphyrin (EP)
elevation in females

Increased urinary
5-aminolevulinic acid

Inhibition of
5-aminolevulinic acid
dehydratase (ALAD)
Elevated blood pressure
(???)*
Elevated serum creatine
(I creatine clearance)
*Note: Arrows depict cases where weight of overall evidence strongly substantiates likely occurrence of type of effect in association with blood-Pb
concentrations in range of 5-10 ng/dL, or possibly lower, as implied by (???). Although no evident threshold has yet been clearly established for those
effects, the existence of such effects at still lower blood-Pb levels cannot be ruled out based on available data.
Source: Adapted/updated from Table 1-16 of U.S. Environmental Protection Agency (1986a).
-------
10 |ig/dL is not a "bright line" separating toxicity from safety, but merely a risk management
tool. There is no level of Pb exposure that can yet be identified, with confidence, as clearly not
being associated with some risk of deleterious health effects. Recent studies of Pb neurotoxicity
in infants have observed evidence of effects at population mean blood-Pb levels of only 1 or
2 |ig/dL and some cardiovascular, renal, and immune outcomes have been seen at blood-Pb
levels below 5 |ig/dL. Public health interventions have resulted in declines, over the last
25 years, of more than 90% in the mean blood-Pb level within all age and gender subgroups of
the U.S. population, substantially decreasing numbers of individuals at risk for Pb toxicity.
Recent studies have strengthened the consensus that the developing nervous system is the
organ system that is probably most sensitive to Pb toxicity in children. Based on new findings,
notable neurobehavioral deficits appear to occur at distinctly lower levels of exposure than had
been previously documented. The discussion in Section 8.5.1 below focuses on the functional
form of these observed relationships and their potential public health implications, starting with
blood-Pb/IQ relationships. Probably the most clearly established other Pb effects of concern are
cardiovascular effects, with several well-conducted new studies providing strong evidence that
elevations in blood Pb levels (even at <10 |ig/dL) are significantly associated with increased
systolic and diastolic blood pressure in adults (as discussed in Sections 6.5 and 6.10.8.2).

8.5.1 Concentration-Response Relationships for Neurotoxicity Effects
Newly accumulating data validate well the statement made in the 1996 AQCD/Addendum
and the 1990 Supplement that adverse effects occur at blood Pb levels of 10 to 15 |ig/dL or
"possibly lower." In a recent study of 6 to 16 year old children in the NHANES III survey,
concentration-related deficits in reading and arithmetic scores were found even when analyses
were restricted to children with concurrent blood Pb levels below 5 |ig/dL (Lanphear
et al., 2000), although these analyses were limited by the fact that direct adjustments could not be
made for certain important potential confounding factors (i.e., maternal IQ or caretaking quality
in the home) whose inclusion in regression models often notably reduces the size of the Pb
coefficient. Canfield et al. (2003a) applied semi-parametric models with penalized splines to
their data, essentially allowing the data to reveal the functional form that best described them.
These analyses showed that the IQ decline per |ig/dL increase in blood Pb was greater below
10 |ig/dL than it was above 10 |ig/dL. The estimated slope of the IQ decline per |ig/dL was
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greatest among children for whom the maximum blood Pb level measured over the course of
the study never exceeded 10 |ig/dL. Also, a similarly steeper slope was seen at lower than at
higher blood Pb levels in a re-analysis of the Boston prospective study (Bellinger and
Needleman, 2003).
Identifying the functional form that best fits a particular set of data and that presumably
represents the best description of the pertinent underlying concentration-response relationship is
clearly important. The linear model (Figure 8-6), as the name implies, is linear over the entire
range of the exposure data. For certain tests, the assumption is made that the residuals
(observed - predicted response) are normally distributed with constant variance, but violations of
this assumption in the presence of heteroscedasticity have no real effect on the estimation and
minimal effect on the tests of significance. If heteroscedasticity is present but all other
conditions are met, the regression model still yields unbiased estimators, but the standard errors
can be larger than when remedial efforts such as using weighted regression are employed. The
use of regression requires no assumption concerning the distribution of the independent variable
(i.e., Pb exposure marker).
105
100
(A
c
o
Q.

-------
However, when the form of the heteroscedasticity is an increase in variance with blood Pb
level and when the data are lognormally distributed or otherwise skewed, there are possibly a
large number of influential data points at high blood Pb where the data are least reliable. In this
case, a log transformation of blood-Pb values may result in more precise estimation of the slope
parameter. The log-linear model is concave upwards (assuming that the estimated coefficient is
negative). It approaches a linear function for very high exposure values, but approaches infinity
at very low exposure values. In other words, it implies that the adverse effect of Pb is greater at
lower than at higher blood-Pb levels. Blood Pb levels have been shown repeatedly to follow a
lognormal distribution (Azar et al., 1975; Billick et al., 1979; Hasselblad and Nelson, 1975;
Hasselblad et al., 1980; U.S. Environmental Protection Agency, 1986a; Yankel et al., 1977), but
this is not an argument for choosing the log-linear model. The choice of either log-linear or
linear may be based on the Akaike's Information Criteria (Akaike, 1973), J-test (Davidson and
MacKinnon, 1981), or other statistical tests if the choice is to be based on the best fitting model.
Rothenberg and Rothenberg (2005) compared the linear Pb model with the log-linear Pb model
for the pooled data from Lanphear et al. (2005) using the J-test. The J-test showed that the log
Pb specification was still significant (p = 0.009) in a model that also included the linear Pb
specification, indicating that the log Pb specification described the data significantly better than
did the linear Pb specification. Other models have been used, such as nonparametric models,
spline functions, and polynomial models, but the vast majority of the analyses have used either a
linear model or a log-linear model.
In a recent publication, Bowers and Beck (2006, p. 520) concluded that "a supralinear
slope is a required outcome of correlations between a data distribution where one is lognormally
distributed and the other is normally distributed." The authors' analyses were based on three
assumptions: that blood lead concentrations are lognormally distributed; that IQ is normally
distributed; and that the two have an inverse relationship. However, the authors' conclusions are
true only if those assumptions are met. In fact, IQ scores have not been forced into a normal
distribution in the epidemiologic analyses. Four of the seven studies included in the pooled
analysis by Lanphear et al. (2005) used IQ scores based on the WISC test, and these scores were
not normalized. Canfield et al. (2003), in one of the major studies cited by Bowers and Beck
(2006), used data from the Stanford-Binet test, but have also stated that the IQ data were not
normalized in their analyses(Canfield, pers. comm., September 8, 2006). In addition, the usual
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assumption of regression analysis is that the outcome distribution is normal conditional on the
predictors, unlike the assumption made by Bowers and Beck (2006) that the outcome is normally
distributed. Blood Pb, socioeconomic status and other variable have skewed distributions; when
the outcome is linearly related to these predictors, the outcome distribution will be skewed.
Therefore, while the conclusions drawn by Bowers and Beck (2006) may be true under certain
conditions, their assumptions are not generally the case in the epidemiologic analyses.
In a response to the report by Bowers and Beck (2006), Hornung et al. (2006) provided
evidence that the IQ data used in the pooled analysis of seven studies by Lanphear et al. (2005)
were not normalized. They state that for the individual studies, a linear relationship between IQ
and blood Pb provided an adequate fit over the narrower range of Pb values (<10 |ig/dL)
associated with each study. In the pooled analysis, the authors tested several different models,
and concluded that a loglinear model (a linear relationship between IQ and the log of blood Pb)
provided the best fit (Hornung et al., 2006).
The segmented line model consists of joined straight line segments, where the joined
points are chosen to best fit the data. The log-linear and the quadratic models have been shown
in several cases to better fit the biomarker-response relationship than the linear model. However,
these models are not considered practicable for extrapolation outside the range of the biomarker
variable. The segmented line model is suggested as a more reasonable model for extrapolation
into the low-concentration sparse-data region.
A biological mechanism for a steeper slope at lower than at higher blood-Pb levels has not
been identified. It is conceivable that the initial neurodevelopmental lesions at lower Pb levels
may be disrupting different biological mechanisms than the more severe effects of high
exposures that result in symptomatic poisoning or frank mental retardation (Dietrich et al., 2001).
Perhaps the predominant mechanism at very low blood-Pb levels is rapidly saturated, but a
different, less rapidly saturated process becomes predominant at blood-Pb levels >10 |ig/dL.
As Kordas et al. (2006) states, this might help explain why, within the range of exposures not
producing overt clinical effects, an increase in blood Pb beyond a certain concentration might
cause less additional impairment in children's cognitive functions. However, one must take care
not to interpret this as meaning that higher blood-Pb levels do not induce further toxic harm.
For example, blood-Pb levels >70-80 |ig/dL are still associated with encephalopathy and notable
risk for fatal outcome.
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The nonlinear concentration-response relationship observed between blood-Pb levels and
IQ in recent epidemiologic studies does not preclude the presence of a threshold. Patterson et al.
(1991) determined Pb concentrations in the tooth enamel, femur, and rib from buried skeletons of
Pre-Columbian Southwest American Indians. They found that the mean natural body burden of
adults Homo sapiens uncontaminated by technological Pb was 40 jig Pb/70 kg, which is about
one-thousandth of the mean body burden of present day American adults with no occupational
exposures and no record of childhood lead poisoning. This suggests that much reduced blood-Pb
levels in the 1-10 |ig/dL range are still orders of magnitude above pre-industrial natural levels.
Thus, a threshold for Pb neurotoxic effects may exist at levels distinctly lower than the lowest
exposures examined in these epidemiologic studies.

8.5.2 Persistence/Reversibility of Lead Neurotoxic Effects
Persistence or apparent "irreversibility" of effects can result from two different scenarios:
(1) organic damage has occurred without adequate repair or compensatory offsets, or
(2) exposure somehow persists. As Pb exposure can also derive from endogenous sources (e.g.,
bone), a performance deficit that remains detectable after external exposure has ended, rather
than indicating irreversibility, could reflect ongoing toxicity due to Pb remaining at the critical
target organ or Pb deposited at the organ post-exposure as the result of redistribution of Pb
among body pools.
The persistence of effect appears to depend on the duration of exposure as well as other
factors that may affect an individual's ability to recover from an insult. The likelihood of
reversibility also seems to be related, at least for the adverse effects observed in certain organ
systems, to both the age-at-exposure and the age-at-assessment. In occupationally-exposed
adults, the central and peripheral nervous system correlates of higher Pb burdens appear to
attenuate if exposure is reduced.
Data from the Treatment of Lead Exposed Children (TLC) study, a randomized controlled
trial of late outcomes of children treated for Pb poisoning (baseline blood Pb of 20 to 44 jig/dL),
support the hypothesis that the deficits associated with exposures of such magnitude are
persistent (Dietrich et al., 2004; Rogan et al., 2001). At 36-months post-treatment and at age
7 years, no significant differences in cognition or behavior were noted between the succimer and
placebo groups. Current blood Pb levels were significantly associated with cognitive
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performance at baseline, 36-months post-treatment, and at 7 years of age, and the regression
coefficients were similar in magnitude to those estimated in observational studies (i.e., ~3 point
IQ decline per 10 |ig/dL increase in blood Pb), providing a linkage between the results of the
observational studies and those of this experimental study. However, within-child analyses
indicated that changes in developmental test scores over time were not consistently associated
with changes over time in blood Pb level.
The prospective studies of childhood Pb exposure, using serial measurements of Pb
biomarkers and health outcomes, provide the best opportunities available to assess the natural
history of adversities associated with low-level Pb exposures. In some prospective studies,
associations observed in infancy between biomarkers of prenatal Pb exposure and slowed
neurodevelopment appeared to be attenuated by the time children reached preschool age. It can
be difficult to determine, however, whether this reflects actual disappearance of the effect or an
increased difficulty in detecting it due to the emergence of associations between Pb biomarkers
measured postnatally and neurodevelopment. It is notable, however, that in some prospective
studies of children, associations between biomarkers of prenatal Pb exposure and various
outcomes in middle adolescence have been reported, suggesting that the persistence of the
associations might be endpoint-specific. For example, among children in Kosovo, Yugoslavia,
IQ scores at the age of 8 years were inversely associated with a composite index of prenatal Pb
exposure (average of mothers' blood Pb levels at midpregnancy and at delivery) (Wasserman
et al., 2000b). This association was independent of changes in postnatal blood Pb levels.
Or, among 15 to 17 year old inner-city children in Cincinnati, OH, maternal blood-Pb levels
(ranging from 1 to -30 |ig/dL) in the first trimester were inversely related to attention and
vasoconstriction (Ris et al., 2004) and positively related to the frequency of self-reported
delinquent behaviors (Dietrich et al., 2001).
In most prospective cohort studies, the potential for true longitudinal analysis of the data
has not been fully exploited, with the data evaluated in what is effectively a series of cross-
sectional analyses. Nevertheless, the results of the prospective studies are consistent in showing
that higher postnatal Pb biomarkers are associated with neurocognitive deficits that persist, in
some studies, into early adulthood when the concurrent Pb exposures are generally much lower.
Ongoing external exposure does not appear to be necessary to maintain the deficits, although, as
noted previously, it is not possible to exclude entirely a role for ongoing endogenous exposures
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of the target organs resulting from the redistribution, over time, of Pb stores among different
compartments. These data are consistent with those from experimental nonhuman primate
studies, in which the temporal characteristics of exposure are manipulated as opposed to merely
observed, as in the human studies.
One study examined the persistence of lead-related cognitive impairment using an
intervention that resulted in a marked reduction in external Pb exposure (Soong et al., 1999).
The cognitive abilities of exposed children (n = 32, median blood-Pb level of 15.1 |ig/dL [range
7.7-31.7]) from a kindergarten located near a Pb-recycling plant in Taiwan were compared to a
referent group of children (n = 35, median blood-Pb of 8.4 |ig/dL [range 4.8-12.8]) from another
kindergarten 5 km away from the plant. Both groups of children were comparable with respect
to age, sex, birth order, sibling number, and parental education level. The exposed children were
found to have significantly lower IQ levels compared to the referent children, with a median
score of 94.5 points (range 60-121) compared to 101 points (range 76-129). The next year, the
school located near the Pb-recycling plant was moved an additional 2 km away. A follow-up
study was conducted with 28 in each group 21/2 years later. The median blood-Pb levels of the
previously exposed and referent children decreased to 8.5 |ig/dL (range 5.0-15.0, average decline
of 6.9 |ig/dL) and 7.0 |ig/dL (range 4.0-11.0, average decline of 1.7 |ig/dL), respectively. The
average IQ scores in the previously exposed children increased by 11.7 points (SD 13.2), with a
median value of 107 points (range 75-135). This value was not significantly different from the
median score of the referent children, 109.5 points (range 79-132). These results indicate that IQ
impairment resulting from blood-Pb elevations for a period of 1 to 3 years in 3 to 5 year old
children was at least partially reversible when external Pb exposure was reduced.
Only limited data are available on factors that influence the likelihood that an association
observed between an early Pb biomarker and later outcome will persist among children. In one
study, the association between prenatal exposure and cognitive development in infancy and the
preschool period appeared to attenuate among children living in more privileged circumstances
or in whom postnatal Pb exposures were lower (Bellinger et al., 1988, 1990). These findings are
consistent with those from cross-sectional epidemiologic studies showing that the effects of a
given level of Pb exposure are more severe among disadvantaged children (Lansdown et al.,
1986; Winneke and Kraemer, 1984) and from experimental animal studies showing that being
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raised in an enriched environment can reduce the apparent detrimental impact of Pb exposure on
learning (Guilarte et al., 2003; Schneider et al., 2001).

8.5.3 Factors Affecting Susceptibility to Lead Toxicity
Although increased Pb exposure has been linked to adverse health effects in many
different organ systems, scatterplots reveal tremendous variability of observed points about the
best fit lines representing the concentration-response relationships. In other words, individuals
for whom the Pb biomarker measured has the same value can have markedly different values on
the health indicator measured. Even for neurobehavioral deficits in children, the correlation
between biomarker level and test score rarely exceeds 0.2, indicating that the explained variance
in the test score generally does not exceed 5%. A major challenge is therefore to decompose this
variability, to distinguish components of it that reflect error from components that reflect
biological processes that determine an individual's response to Pb.
Deviation of the observed points from the fitted point can have many sources. Exposure
misclassification is one source. The Pb biomarker measured might not adequately capture the Pb
dose delivered to the target organ that, at the time, is most appropriate biologically. In general,
the error would be expected to be non-differential, i.e., it would not introduce a systematic bias
in the estimation of the concentration-response relationship. On average, such misclassification
would be expected to result both in an attenuation of the slope of the concentration-response
relationship and an increase in the scatter of the observations. As focus shifts to the risks
associated with lower and lower levels of Pb exposure, the importance of errors introduced by
poor dosimetry will assume greater importance insofar as the effects at such levels will
presumably be more subtle and increasingly difficult to detect amid the noise contributed by
exposure misclassification. Outcome misclassification is another source of error that is likely to
contribute to apparent interindividual variability in response. This results if the indicator of the
critical health effect that is measured is fallible, i.e., an imperfect measure of the target function.
Such misclassification would generally be expected to be non-differential, introducing random
noise rather than a systematic bias.
Another likely source of scatter in observed points is true interindividual variability in
response to a given Pb dose. That is, the magnitude of individual response to Pb might depend
on other characteristics of that individual. Three major categories of such effect modifying
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factors that might influence susceptibility to Pb toxicity are genetic polymorphisms, nutritional
status, and social environmental factors. Adequate data are not available to provide a
quantitative estimate of the amount of interindividual variability in susceptibility to Pb.

Influence of Genetic Polymorphisms on Risk
Genetic polymorphisms that are presumed to influence Pb toxicokinetics and/or
toxicodynamics have been identified, mostly in studies of adults who were occupationally
exposed to Pb. The magnitude of Pb-associated renal dysfunction appears to vary, in complex
ways, with the delta-aminolevulinic acid dehydratase (ALAD) polymorphism (Chia et al., 2005,
2006). Lead workers with the ATP1 A2(3') polymorphism appear to be at increased risk of
Pb-associated effects on blood pressure (Glenn et al., 2001). The slope of the association
between floor dust-Pb and blood-Pb is steeper among children with the less common variant of
the vitamin D receptor (Fok 1 or B) than among children with the wild-type allele (Haynes et al.,
2003). In adults, these same alleles are associated with higher blood-Pb levels and increased
blood pressure (Schwartz et al., 2000a; Lee et al., 2001). Greater Pb-associated reductions in
renal function have been observed in adults with a variant allele of nitric acid synthetase,
although cardiovascular outcomes, such as blood pressure and hypertension do not appear to
depend on the eNOS (endogenous nitric oxide synthase) allele (Weaver et al., 2003b). Adults
with variants of the hemochromatosis gene (C282Y and/or H63D) have higher patella Pb levels
(Wright et al., 2004). With regard to polymorphisms that modify Pb neurotoxicity, workers with
the apolipoprotein E4 allele showed greater Pb-associated decreases in neurobehavioral function
than did workers with the El, E2, or E3 alleles (Stewart et al., 2002). Chia et al. (2004)
speculated that the ALAD2 confers protection against Pb neurotoxicity, although Kamel et al.
(2003) reported that this variant allele is associated with an increased risk of amyotrophic lateral
sclerosis. This work is in its early stages and, while it promises to shed light on bases of
susceptibility to Pb toxicity, firm conclusions cannot yet be drawn.

Influence of Nutritional Status on Risk
Only limited epidemiologic data are available on the role of nutritional status in
modifying an individual's risk of Pb toxicity. Adjusting for severity of environmental Pb
contamination, iron-deficient children appear to have higher blood-Pb levels than iron-replete
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children (Bradman et al., 2001). One interpretation of these data is that children experiencing the
same external Pb dose can experience different internal doses. In another study of iron status,
a decline in blood-Pb level was associated with improved cognitive performance in iron-
sufficient but not in iron-deficient children (Ruff et al., 1996). Among the possible explanations
for this finding is that iron deficiency contributes to pharmacodynamic variability, increasing the
toxicity of a given Pb dose. Some evidence suggests that the intellectual deficit associated with
an elevated blood-Pb level is greater among undernourished children than well-nourished
children (Gardner et al., 1998).
Several studies have suggested that dietary calcium may have a protective role by
decreasing absorption of Pb in the gastrointestinal tract and decreasing the mobilization of Pb
from bone stores to blood, especially during periods of high metabolic activity of the bone such
as pregnancy and lactation. Lower calcium intake during pregnancy, especially the second half,
appears to increase the mobilization of Pb from bone compartments (Hernandez-Avila et al.,
1996). However, in other studies, calcium supplementation had no effect on bone-Pb levels in
pregnant and lactating women (Rothenberg et al., 2000; Tellez-Rojo et al., 2002).

Influence of Health Status on Risk
The influence of an individual's health status on susceptibility to Pb toxicity has been
demonstrated most clearly for renal outcomes. Individuals with diabetes, hypertension, and
chronic renal insufficiency are at increased risk of Pb-associated declines in renal function, and
indications of altered kidney function have been reported at blood Pb levels ranging somewhat
below 5 |ig/dL (Lin et al., 2001, 2003; Muntner et al., 2003; Tsaih et al., 2004). As noted in the
previous section, children with nutritional deficiencies also appear to be more vulnerable to Pb-
associated neurobehavioral deficits.

Influence of Coexposures on Risk
Epidemiologic studies do not provide an adequate basis for determining whether cigarette
smoking and/or alcohol affect the nature or severity of Pb health effects. Both factors have often
been included in models of both child and adult health outcomes to adjust for potential
confounding. Both have also been evaluated as pertinent pathways of adult exposure. However,
their possible roles as effect modifiers have not been well studied.
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Although most individuals are not exposed to Pb in isolation but rather to Pb in
combination with other toxicants (e.g., cadmium, arsenic, mercury, and polychlorinated
biphenyls), epidemiologic studies have generally focused solely on Pb. Other toxicant exposures
have sometimes been measured but are usually treated as potential confounders in the statistical
analyses, with their potential as possible modifiers of Pb toxicity left unexplored (Bellinger,
2000). Thus, available epidemiologic studies do not provide an adequate basis for determining
the extent to which co-exposure to other toxicants may affect the nature or severity of Pb-related
health effects.

Influence of Timing of Exposure on Risk
Children
Available studies do not provide a definitive answer to the question of whether Pb-
associated neurodevelopmental deficits are the result of exposure during a circumscribed critical
period or of cumulative exposure. Although support can be cited for the conclusion that it is
exposure within the first few postnatal years that is most important in determining long-term
outcomes (Bellinger et al., 1992), other studies suggest that concurrent blood-Pb level is as
predictive, or perhaps more predictive, of long-term outcomes than are early blood-Pb levels
(Canfield et al., 2003a; Dietrich et al., 1993a,b; long et al., 1996; Wasserman et al., 2000b).
Because of the complex kinetics of Pb, an accumulative toxicant, it is extremely difficult to draw
strong conclusions from these observational studies about windows of heightened vulnerability
in children. The high degree of intra-individual "tracking" of blood Pb levels over time,
especially among children in environments providing substantial, chronic exposure opportunities
(e.g., residence near a smelter or in older urban dwellings in poor repair), poses formidable
obstacles to identifying the time interval during which exposure to Pb caused the health effects
measured in a study. It could be that damage occurred during a circumscribed period when the
critical substrate was undergoing rapid development, but that the high correlation between serial
blood Pb levels impeded identification of the special significance of exposure at that time.
Under such circumstances, an index of cumulative blood Pb level or concurrent blood Pb
level, which might be a good marker of overall body burden under conditions of relatively
steady-state exposure, might bear the strongest association with the effect. Under these
circumstances, however, it might be incorrect to conclude that it was the later exposures,
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incurred around the time that the effect was detected, that was responsible for producing it.
While some observations in children as old as adolescence indicate that exposure biomarkers
measured concurrently are the strongest predictors of late outcomes, the interpretation of these
observations with regard to critical windows of vulnerability remains uncertain. Additional
research will be needed to distinguish effects that reflect the influence of later Pb exposures from
effects that reflect the persistent of effects resulting from exposure during some prior critical
window. Resolving this issue solely on the basis of data from observational studies will be
difficult due to the high intercorrelation among blood Pb measures taken at different ages.
Increasing attention is being devoted to determining the extent to which early childhood
Pb exposures increases the risk of adverse effects that are only apparent at older ages (i.e.,
delayed or latent effects). Among young adults who lived as children in an area heavily polluted
by a smelter and whose current Pb exposure was low, higher bone Pb levels were associated with
higher systolic and diastolic blood pressure (Gerr et al., 2002). In adult rats, greater early
exposures to Pb are associated with increased levels of amyloid protein precursor, a marker of
risk for neurodegenerative disease (Basha et al., 2005).

Aging Population
Increases in blood Pb for postmenopausal women have been attributed to release of Pb
from the skeleton associated with increased bone remodeling during menopause in both
occupationally- and environmentally-exposed women (Garrido-Latorre et al., 2003; Popovic
et al., 2005). Also, in middle-aged to elderly males from the Normative Aging Study, patella Pb
accounted for the dominant portion of variance in blood Pb (Hu et al., 1996). These findings
suggest that the skeleton serves as an endogenous source of Pb in the aging population.
Considerable evidence also suggests that indicators of cumulative or long-term Pb
exposure are associated with adverse effects in several organ systems, including the central
nervous, renal, and cardiovascular systems. Among occupationally-exposed men, higher tibia
Pb levels have been associated with increased cognitive decline over repeated assessments
(Schwartz et al., 2005). With regard to the renal system, increased Pb exposure may accelerate
the effects of normal aging, producing a steeper age-related decline in function. Weaver et al.
(2003a) observed that higher Pb exposure and dose were associated with worse renal function in
older workers, but with lower blood urea nitrogen and serum creatinine in young workers.
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Pregnancy
Mobilization of Pb from the skeleton also occurs during pregnancy and lactation due to
increased bone remodeling to meet the calcium requirements of the developing fetus
(Hertz-Picciotto et al., 2000; Manton, 1985; Silbergeld, 1991). In women who have been
exposed to Pb in childhood and have accumulated large stores in their bones, there may be
significant mobilization of Pb from bone to blood during late pregnancy and lactation. Lead
isotope studies on immigrant women to Australia reported increases of 20% to 99% during
pregnancy (Gulson et al., 1997, 1998). Skeletal Pb contribution to blood Pb was significantly
greater during the postpregnancy period than during the second and third trimesters. The highest
probability of Pb toxicity for the mothers will be in postpartum while they are lactating; the
infants will be particularly vulnerable during the prenatal period, especially in the last weeks of
pregnancy (Manton et al., 2003). Calcium supplementation appears to provide a modest
reduction in blood-Pb levels in pregnant or lactating women (Gulson et al., 2004;
Hernandez-Avila et al., 2003).
A variety of adverse reproductive outcomes have been associated with higher paternal or
maternal Pb exposures, including reduced fertility, spontaneous abortion, gestational
hypertension, congenital malformations, fetal growth deficits, and neurobehavioral deficits in
offspring. The levels of exposure at which different adverse outcomes occur vary. Increased
risks of spontaneous abortion, neurobehavioral deficits in offspring and, in some studies,
gestational hypertension, have been reported at pregnancy blood Pb levels below 10 |ig/dL
(Bellinger, 2005).
8.6 POTENTIAL PUBLIC HEALTH IMPLICATIONS OF LOW-LEVEL
LEAD EXPOSURE
8.6.1 Introduction
In studies of Pb toxicity, health endpoints have more often been continuously-distributed
indices such as blood pressure or IQ. A view that the endpoints should be diagnoses rather than
measured values on the underlying indices is that a change in the value of a health index that
does not exceed the criterion value defining the diagnosis is therefore without consequence for
an individual's health. The World Health Organization (WHO) definition of "health,"
8-75
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is: "Health is a state of complete physical, mental and social well-being and not merely the
absence of disease or infirmity" (World Health Organization, 1948). By this definition, even
decrements in health status that are not severe enough to result in the assignment of a diagnosis
might be undesirable if they reflect a decrement in an individual's well-being but are not severe
enough to meet diagnostic criteria. Deficits in health indices or well-being may not be
observable except in aggregate, at the population level. The American Thoracic Society
discusses similar concepts of shift in population distribution and health effects (American
Thoracic Society, 2000).
Sometimes, the importance of a Pb-associated change on a health index is evaluated by
comparing it to the standard error of measurement of the index, i.e., the statistic that defines the
range within which an individual's "true" value on the index is likely to lie. For instance the
standard error of measurement for full scale IQ is 3 to 4 points, leading some to conclude that the
estimated IQ decrement of 3 points per 10 |ig/dL increase in blood Pb level is "in the noise" of
measurement and, therefore, meaningless. A similar claim has been made with regard to the
magnitude of the association between Pb and blood pressure. The error in this argument is that
the estimated decrement of 3 IQ points per 10 |ig/dL applies to grouped, not individual, data.
For measurement error to provide an explanation for the observation of an association that is
approximately the size of the standard error of measurement, it would be necessary to postulate
that the true association is null, but that, by chance or because of some bias, the measured IQ
scores of the individuals with higher Pb exposures were systematically underestimated (i.e., their
true IQ scores lie in the upper tails of the 95% CI for the children's observed scores) and that the
measured IQ scores of the individuals with lower exposures were systematically overestimated
(i.e., their true IQ scores lie in the lower tails of the 95% CI). Thus, this argument requires an
assumption that the direction of measurement error is highly correlated with exposure status.
The fundamental flaw is using a statistic that pertains to individual-level data to draw inferences
about group-level data.
Nosology (the classification and naming of diseases) is dynamic as knowledge accrues.
The total serum cholesterol level that is considered indicative of hyperlipidemia has dropped
steadily over the past 40 years. Second, even within the range of health index values that are
sub-diagnostic, variations on the index are significantly associated with health outcomes.
For instance, even among children with birth weights greater than the cut-off used to define
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"low birth weight," birth weight is significantly associated with IQ at age 7 years (Matte et al.,
2001). Third, exposure-related changes on a health index can be markers or indicators of other
changes that are likely to have occurred whose significance is more certain. For instance, slower
completion of a commonly-used neuropsychological test, the Grooved Pegboard, is associated
with poorer handwriting, and reduced ability to copy a drawing is associated with a greater risk
of a need for remedial school services (Bellinger, 2004).
The critical distinction between population and individual risk, an issue pertinent to many
questions in chronic disease epidemiology, has often been blurred in discussions of the public
health implications of Pb-associated decrements in health. In regard to neurodevelopment,
although a two- or three-point decline in IQ might not be consequential for an individual, it is
important to recognize that this figure represents the central tendency of the distribution of
declines among individuals. Thus, some individuals might manifest declines that are much
greater in magnitude, while others manifest no decline at all, reflecting interindividual
differences in vulnerability. Moreover, the import of a decline for an individual's well-being is
likely to vary depending on the portion of the IQ distribution. For an individual functioning in
the low range due to the influence of developmental risk factors other than Pb, a Pb-associated
decline of several points might be sufficient to drop that individual into the range associated with
increase risk of educational, vocational, and social failure.
The point estimate indicating a modest mean change on a health index at the individual
level can have substantial implications at the population level. For example, although an
increase of a few mmHg in blood pressure might not be of concern for an individual's well-
being, the same increase in the population mean might be associated with substantial increases in
the percentages of individuals with values that are sufficiently extreme that they exceed the
criteria used to diagnose hypertension (Rose and Day, 1990). In other words, the mean value
conveys substantial information about the percentage of individuals with clinically relevant,
extreme values of the indicator. Moreover, interventions that shift the population mean, in a
beneficial direction, by an amount that is without clinical consequence for an individual have
been shown to produce substantial decreases in the percentage of individuals with values that are
clinically significant (Bellinger, 2004). The following subsections discuss quantitatively
Pb-related effects of a population level change in IQ and blood pressure.
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8.6.2 Potential Implications of Lead Effects on Intelligence
The outcome most often examined to investigate the neurotoxic effects of Pb is IQ.
Although the definition of "intelligence" is quite abstract, IQ remains a useful outcome measure
as it is correlated with important measures of life success, such as academic achievement,
earnings, and social status (Bellinger, 2003; Weiss, 2000). Several studies reported quantitative
relationships between measures of IQ and current blood Pb levels for children aged 2 to 11 years
old. The estimated relationships as reported by the authors are summarized in Table 8-7,
organized by the type of model used in the analysis. The pooled analysis by Lanphear et al.
(2005) included the studies of Baghurst et al. (1992), Bellinger et al. (1992), Canfield et al.
(2003a), Dietrich et al. (1993a), Ernhart et al. (1989) and Wasserman et al. (1997). That pooled
analysis also included the Mexico City study of Schnaas et al. (2000). The results from Schnaas
et al. (2000) are not included in Table 8-7, because the authors did not provide regression
coefficients in their paper, thus the concentration-response relationship was not estimable.
The study by Earnhart et al. (1989) also is not included, as slopes cannot be estimated from the
covariate variances presented for the adjusted models.
The curves over a range of blood Pb levels from the 10th percentile to the 90th percentile
are shown in Figure 8-7. The curves are restricted to that range because log-linear curves
become very steep at the lower end of the blood Pb levels, and this may be an artifact of the
model chosen. The percentiles are estimated using various methods and are only approximate
values. Studies which estimated a linear relationship are shown as reported, and similarly for the
log-linear relationships. Note that these are not forest plots of slopes or hazard ratios—they are
the actual estimated relationships.
Several conclusions can be drawn from these graphs. First, note that the overall IQ levels
are quite different. This results from different populations and from different applications of the
IQ tests. Second, all studies showed a decreasing IQ score as the blood-Pb level increased. It is
the slope of the studies that is relevant, not the actual IQ scores. Third, for studies with lower
blood-Pb levels, the slopes appear to be steeper. This is the reason that many authors choose to
use the log-linear model. However, for those studies where the blood-Pb levels were generally
high, the log-linear and linear models are almost identical. Thus, it is not surprising that some
authors chose a linear model instead of a log-linear model. The curves in Figure 8-7 do not show
evidence of a no-effect threshold because the slopes increase as the blood-Pb levels become
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Table 8-7. Summary of Studies with Quantitative Relationships for IQ and Blood Lead
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Dietrich et al. (1993)
,Wasserman et al. (1997)
10 15 20
Blood Lead (M9/dL)
25
30
Figure 8-7. Concentration-response relationships of IQ to blood lead for the individual
studies and the pooled analysis by Lanphear et al. (2005).
smaller. The observed mean adjusted IQ levels (for blood Pb <5, 5-10, 10-15, 15-20, and
>20 |ig/dL) reported by Lanphear et al. (2005) also show no evidence of a threshold, as seen in
Figure 8-8.
Weiss (1990) predicted, on purely statistical grounds, that a downward shift of five points
in mean IQ, if the amount of dispersion in the distribution remained the same, should be
accompanied by a doubling of the numbers of individuals with scores two or more standard
deviations below the mean and a reduction by half of the number of individuals with scores two
or more standard deviations above the mean. With respect to Pb, the general accuracy of this
prediction has been empirically demonstrated in two different datasets by Needleman et al.
(1982) and Bellinger (2004). An illustrative example is provided below, and it shows further
evidence of the change in percentages of individuals with IQ <80 or <70 points and >120 or
>130 points after restricting the analysis to those with blood-Pb levels <10 |ig/dL.
The slope of -0.9 points/|ig/dL was used in these calculations. This slope is the median
value from the estimated slopes for blood-Pb levels <10 |ig/dL presented in Table 8-7.
A nonexposed population was assumed to have a standard mean IQ of 100 and standard
8-80
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105
100
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85
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35
40
Figure 8-8. Mean blood lead levels adjusted for HOME score, maternal education,
maternal IQ, and birth weight from the pooled analysis of seven studies by
Lanphear et al. (2005). Mean adjusted IQ levels at blood lead levels of <5,
5 to 10,10 to 15,15 to 20, and >20 ug/dL are shown.
deviation of 15 at a blood-Pb exposure of 0 |ig/dL. The fraction of the population that would
have an IQ <80 or <70 as a function of blood-Pb level was then calculated. The results are
shown in Figure 8-9 A. The fraction of the population with an IQ level less than 80 more than
doubles from 9% with no Pb exposure to 23% with a blood-Pb level of 10 |ig/dL. The fraction
with an IQ level below 70, a level often requiring community support to live (World Health
Organization, 1992) increases from a little over 2% with no Pb exposure to about 8% with a
blood-Pb level of 10 jig/dL.
The Pb-related decrements in IQ are manifested fairly uniformly across the range of IQ
scores (Needleman et al., 1982). Thus, a shift in the mean value of a health indicator has
substantial importance for both extremes of the distribution. In the case of Pb, a downward shift
in the mean IQ value is not associated only with a substantial increase in the percentage of
individuals achieving very low scores, but also with substantial decreases in percentages
achieving very high scores. Based on the study by Bellinger et al. (1987) examining intelligence
test scores of Pb-exposed children, Weiss (1988) discussed the shift of the population
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distribution of IQ from a mean of 100 and a standard deviation of 15 to a mean of 95, a 5%
reduction. When the mean IQ level is 100, 2.3% of the individuals in a given population would
score above 130. However, with the population distribution shift and the resulting mean decline
in IQ, only 0.99% of the individuals would score above 130. Weiss states that the implication of
such a loss transcends the current circumscribed definitions of risk. Similar results were
observed using the slope of-0.9 points/ |ig/dL to examine the effects on the percentage of
individuals with an IQ >120 or >130 points at blood Pb levels <10 jig/dL (Figure 8-9B). The
fraction of individuals with an IQ >120 decreased from about 9% with no Pb exposure to less
than 3% at a blood Pb level of 10 |ig/dL. The fraction of individuals with an IQ >130 points
decreased from 2.25% to 0.5% with a blood Pb level change from 0 to 10 |ig/dL.
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<70 points (A) and IQ levels >120 or >130 points (B).
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8.6.3 Potential Implications of Cardiovascular Effects of Lead
In human epidemiology studies investigating the cardiovascular effects of Pb, blood
pressure has been examined most frequently, as discussed in Section 6.5.2 of Chapter 6. Results
from the Framingham Heart Study show that higher levels of blood pressure, even within the
nonhypertensive range, impose increased rates of cardiovascular disease (Kannel, 2000a,b).
A continuous graded increase in cardiovascular risk is observed as blood pressure increases, with
no evidence of a threshold value. Most events arise not in the most severe cases, but mainly in
those with high normal blood pressure (i.e., mild hypertension). This view is further supported
by the Seventh Report of the Joint National Committee on Prevention, Detection, Evaluation,
and Treatment of High Blood Pressure (Chobanian et al., 2003). Kannel (2000b) states that
reducing even moderate elevation in blood pressure is likely to be beneficial.
Kannel (2000a) emphasized that systolic blood pressure exerts a strong influence on more
serious cardiovascular events, as it is the prime causal function of hypertension and its adverse
cardiovascular sequelae. Cardiovascular events include coronary disease, stroke, peripheral
artery disease, and cardiac failure. Risk ratios are larger for cardiac failure and stroke, but
coronary disease (i.e., myocardial infarction, angina pectoris, sudden death) is the most common
and most lethal sequela of hypertension (Kannel, 1996). Kannel (2000a) notes that the
Framingham Heart Study has recognized that elevated blood pressure tends to occur alongside
other major risk factors of cardiovascular disease such as glucose intolerance, dyslipidemia,
abdominal obesity, and left ventricular hypertrophy, among others. If a cluster of multiple risk
factors is present, the hazard is formidable for coronary disease and stroke.
No single critical level for blood pressure is evident. The risk appears to be simply
proportional from the lowest to the highest level recorded. In the Multiple Risk Factor
Intervention Trial (MRFIT), Neaton et al. (1995) confirmed a continuing and graded influence of
systolic blood pressure on cardiovascular disease mortality extending down into the range of
<140 mm Hg. The Prospective Studies Collaboration (2002) meta-analysis of 61 prospective
studies relates blood pressure to vascular mortality without indication of a threshold down to
115/75 mm Hg. The absence of a demonstrable safe or critical level of blood pressure suggests
using the range of blood pressure rather than discrete categories such as hypertension.
Many studies have provided evidence for a relationship between blood Pb and systolic
blood pressure. In particular, the meta-analysis of Nawrot et al. (2002) indicated that a doubling
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of the blood Pb (e.g., from 5 to 10 |ig/dL) corresponded to a 1 mm Hg increase in systolic blood
pressure. As noted earlier, although this magnitude of increase in systolic blood pressure is not
particularly meaningful clinically for any given individual, a population shift of 1 mm Hg is
important.
The Framingham Heart Study results (Kannel, 2000a) were used to estimate a typical
population distribution of systolic blood pressure values (Figure 8-10). The distribution of
systolic blood pressure values was approximated well by a lognormal distribution for both
women and men (p > 0.4). The relationship between systolic blood pressure and the risk of
cardiovascular events was also given by Kannel (2000a), as shown in Figure 8-11. To estimate
population risk, it was assumed that the effect of blood Pb on blood pressure was to shift the
entire distribution by the amount given by Nawrot et al. (2002). For each shift in the
distribution, the entire distribution was integrated out over the risk given in Figure 8-11.
The result estimated was the expected number of cardiovascular events per 1,000 person
years, and this was plotted for blood-Pb levels ranging from 5 to 15 |ig/dL for both women and
men. The results are shown in Figure 8-12. Although the effects are modest, they translate into
a large number of events for a moderate population size. For example, a decrease in blood Pb
from 10 to 5 |ig/dL results in an annual decrease of 27 events per 100,000 women and 39 events
per 100,000 men.
In order to relate the effects of blood Pb levels to air Pb concentrations, an estimate of the
relationship of air Pb to blood Pb in adults is necessary. One such estimate, as an example, can
be derived from the Azar et al. (1975) study, which used personal monitors to estimate air Pb
exposure in 149 adults (as discussed in Chapter 11 of the 1986 Lead AQCD). In that study, the
estimated slope at an air Pb concentration of 1.0 |ig/m3 was a 2.57 |ig/dL increase in blood Pb
per 1 |ig/m3 increase in air Pb. Based on this slope estimate, a 0.25 |ig/m3 decrease in air Pb
would lead to a 0.64 |ig/dL decrease in blood Pb levels. Using both the relationship between
blood-Pb levels and blood pressure (i.e., a doubling of the blood-Pb corresponds to a 1 mm Hg
increase in systolic blood pressure) and the relationship between blood pressure and
cardiovascular events, a decrease of 0.64 |ig/dL in blood Pb from 5 |ig/dL to 4.36 |ig/dL would
be projected to lead to an annual decrease of 5 cardiovascular events per 100,000 for women and
8 events per 100,000 for men. For a city of 3 million people (about the size of Chicago) this
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Systolic Blood Pressure (mm Hg) in Men
Figure 8-10. Distribution of systolic blood pressure in women and men aged 35 to 64 years

from the Framingham Heart Study (Kannel, 2000a).
would translate to about 150 fewer events (e.g., heart attacks, strokes) for women and 240 fewer

events for men, respectively. For a city of 10 million people (about the size of New York City)

the estimated fewer serious cardiovascular events annually would be 500 and 800, respectively,

for women and men.
8-85
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Figure 8-11. Relationship of serious cardiovascular events (coronary disease, stroke,
peripheral artery disease, cardiac failure) to systolic blood pressure in
women and men aged 35 to 64 years from the Framingham Heart Study
(Kannel, 2000a).
Recent analyses of NHANES II and III data have yielded evidence supporting the
likelihood that long-term Pb exposure can increase the risk of cardiovascular-related mortality in
the general U.S. population, consistent with projected likely increases in serious cardiovascular
events (stroke, heart attack) resulting from even small Pb-induced increases in blood pressure (as
discussed above).
A recent follow-up of the NHANES II cohort provided mortality data used to associate
past blood Pb concentration with increased circulatory mortality in the U.S. population (Lustberg
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Figure 8-12. Effect of blood lead on expected annual risk of cardiovascular events per
1,000 person-years.
and Silbergeld, 2002). Blood Pb concentration as measured during 1976 to 1980 was divided
into three categories (<10 |ig/dL, 10-19 |ig/dL, and 20-29 |ig/dL) after eliminating 109 subjects
with blood Pb >30 |ig/dL, leaving 4,190 subjects 30-74 years of age in the mortality sample
followed to the end of 1992. During the follow-up period, 929 subjects died of all causes.
ICD-9 codes 390-459 (circulatory) accounted for 424 deaths. Proportional hazards models using
a priori selected potential confounding variables (age, sex, race, education, income, smoking,
BMI, exercise, and location) were used to calculate risk ratios of cardiovascular mortality for the
two higher Pb categories compared to a <10 |ig/dL reference. The 20-29 |ig/dL category showed
significantly elevated relative risk of 1.39 (95% CI: 1.01,1.91) for cardiovascular mortality.
Although the NHANES II analysis using data from 1976 to 1980 suggested an increased
risk of mortality at blood Pb levels above 20 |ig/dL, blood Pb levels have dramatically decreased
since the late 1970s. More recent data from NHANES have found that the geometric mean
blood Pb levels decreased from 12.8 jig/dL in 1976-1980 to 2.8 jig/dL in 1988-1991 (Annest
et al., 1983) and 2.3 jig/dL in 1991-1994 (CDC, 1997). NHANES III data (1988-1994) were
8-87
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used to further analyze risk of mortality in adults (age 40 years) at lower blood Pb levels
(Schober et al., 2006). A total of 9,757 subjects were followed for a median of 8.55 years during
which there were 2,515 deaths. An increased risk of cardiovascular mortality was associated
with blood Pb levels of 5-9 |ig/dL and 10 |ig/dL compared to 5 |ig/dL. The relative risk was
1.20 [95% CI: 0.93, 1.55] for 5-9 |ig/dL and 1.55 [95% CI: 1.16, 2.07] for 10 |ig/dL, and the
test for trend was statistically significant. Increased risks of all cause and cancer mortality also
were observed at blood Pb levels of 5-9 |ig/dL compared to <5 |ig/dL (relative risk of 1.24 [95%
CI: 1.05, 1.48] for all cause mortality and 1.44 [95% CI: 1.12, 1.86] for cancer mortality).
The authors noted that an important limitation of this study was that exposure classification was
based on one blood Pb level measurement taken at baseline (i.e., at 40 years old). These
individuals were more likely to have notably higher past peak and cumulative Pb exposure, and
their blood Pb levels might have been disproportionately influenced by release of Pb from bone
stores compared to younger individuals.
The effects of Pb on serious cardiovascular events and mortality were projected to be
more modest in two studies that estimated relative risks of serious cardiovascular outcomes from
Pb-induced blood pressure changes using data from the NHANES II cohort. Pirkle et al. (1985)
calculated that a 6.2 |ig/dL increase in blood Pb levels (from 10.5 to 16.7 |ig/dL) predicted a
4.6% increase in the incidence of fatal and nonfatal myocardial infarctions and 6.7% increase in
the incidence of fatal and nonfatal strokes over 10 years in White men aged 40-54 years. The
incidence of death from all causes was estimated to increase by 5.5% over 11.5 years with the
same increase in blood Pb levels. Schwartz (1991) noted that doubling of blood Pb levels was
associated with a relative risk of-1.05 for cardiovascular disease in men and women aged
20-74 years.
Collectively, the above analyses of NHANES II and III data suggest a significant effect of
Pb on cardiovascular mortality in the general U.S. population. Consideration of this health
outcome may be qualitatively useful in helping to more fully understand potential public health
impacts of Pb. However, the reasons for the notable differences between the above-noted
analyses in estimated increased risk of cardiovascular-related fatal outcomes associated with
incremental changes in blood Pb concentrations are unclear at this time. One likely explanation
is the overestimation of risk by the Schober et al. (2006) analyses by relating ongoing mortality
to relatively more recent baseline blood Pb levels in older adults than to past higher peak or
8-88
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cumulative Pb exposures. Conversely, the Schober et al. estimates may, in part, reflect the
contributions of Pb exposures to fatal cardiovascular outcomes mediated via other possible
underlying mechanisms (e.g., Pb effects on heart rate variability, cholesterol metabolism,
arthrosclerosis) in addition to Pb effects on blood pressure. Thus, until the Schober et al.
findings are replicated and more fully understood, the Schober et al. (2006) estimates for Pb-
induced cardiovascular mortality should probably not be used for quantitative risk assessment
purposes.

8.6.4 Potential Implications of Renal Effects of Lead
The potential clinical relevance of Pb renal effects for chronic kidney disease has recently
been examined. Chronic kidney disease is an important risk factor for cardiac disease and other
causes of mortality and morbidity. Increasing blood lead from the 5th to the 95th percentile
(3.5 |ig/dL) has the same adverse impact on glomerular filtration as increases in age and body
mass index (both known renal risk factors) among the general population. Further, a 10-fold
increase in blood Pb (e.g., from 1 to 10 |ig/dL) causes a 22.5% decrease in creatinine clearance
in populations at high risk for Pb exposure. The biomedical significance of such altered
creatinine clearance remains to be more fully elucidated, however, given observations in
occupationally exposed groups of more notable renal dysfunction signs only at substantially
higher blood-Pb levels (<30-40 (ig/dL).
A Pb-induced small downward shift in renal function among a general population may not
alone result in chronic kidney disease in identifiable individuals; however, the segment of the
population with the lowest renal reserve may be put at increased risk for chronic kidney disease
when Pb exposure is combined with one or more other renal risk factors. Effect estimates in
susceptible populations, such as those with diabetes, hypertension, or chronic renal insufficiency
from non-Pb related causes, are likely to be higher. Lead exposure in populations that are also at
increased risk for obesity, diabetes, and hypertension represent groups likely to be the most
impacted by Pb. Frequently both risk factors are present in the same lower socioeconomic status
groups.
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8.6.5 Potential Implications of Lead-Induced Immune System Effects
Disease implications associated with Pb-induced immune changes seen in animals are
likely to include an increased risk of allergic diseases, atopic manifestations and possibly later-
life autoimmunity as well as a reduced capacity to combat certain viral infections and cancers.
Diseases associated with hyperinflammation would also be of concern. A recent mechanistic
study in the mouse produced two major findings (see Section 5.9.8): (1) it confirmed the
capacity of Pb to induce a Th2 bias, increasing allergic disease concerns; and (2) it showed that
Pb exposure elevates immune reaction against neoantigens, thereby increasing the risk of
autoimmune reactions.

8.7 KEY LEAD ECOSYSTEM EFFECTS AND POTENTIAL
IMPLICATIONS
8.7.1 Terrestrial Ecosystems
Surface soils across the United States are enriched in lead (Pb) relative to levels expected
from natural (geogenic) inputs. While some of this Pb contamination is attributable to paint,
salvage yards, shooting ranges, and the use of Pb arsenate as a pesticide in localized areas,
Pb contamination of surface soils is essentially ubiquitous because of atmospheric pollution
associated with past widespread use of leaded gasoline and more contemporary metal smelting
and production, combustion of fossil fuels, and waste incineration (see Table 2-8). However,
lead inputs to terrestrial ecosystems in the United States have declined dramatically in the past
30 years, due to the almost complete elimination of alkyl-lead additives in gasoline in North
America. Also, emissions from smelters have declined as older plants have been shut down or
fitted with improved emissions controls.
Most terrestrial ecosystems in North America remain sinks for Pb, despite reductions in
atmospheric Pb deposition of more than 95% during the past several decades. Lead released
from forest floor soils in the past has been largely immobilized in mineral soils (see Section 8.1).
The amount of Pb that has leached into the mineral soil to date has been estimated to range from
20 to 90% of past total anthropogenic Pb deposition, depending on forest type, climate, and litter
cycling. While inputs of Pb to ecosystems are currently low, Pb export from watersheds via
groundwater and streams appears to be substantially lower, Pb concentrations in waters draining
natural terrestrial ecosystems always having been reported as low (generally less than 1 ng/L),
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even at moderately polluted sites. Therefore, even at current low input levels, U.S. watersheds
are accumulating industrial Pb. However, burial/movement of lead over time down into lower
soil/sediment layers also tends to sequester it away from more biologically active parts of the
watershed (unless later disturbed or redistributed, e.g., by flooding, dredging, etc.).

Metal Speciationfor Plants
When considering the bioavailability of a metal to plants from soils and sediments, it is
generally assumed that both the kinetic rate of supply and the speciation of the metal to either the
root or shoot are highly important. In soils and sediments, generally only a small volume of
water is in contact with the chemical form; and, although the proportion of the concentration of a
metal in this pore water to the bulk soil/sediment concentration is small, it is this phase that is
directly available to plants. Therefore, pore water chemistry, i.e., metal concentration as simple
inorganic species, organic complexes, or colloid complexes, is most important. Tools currently
used for metal speciation for plants include (1) in situ measurements using selective electrodes;
(2) in situ collection techniques using diffusive equilibrium thin films (DET) and diffusive
gradient thin films (DOT) followed by laboratory analyses; and (3) equilibrium models (e.g.,
SOILCHEM) (see Section 7.1.1 and AX7.1.1.2).

Lead Solid-solution Partitioning
The concentration of Pb species dissolved in soil solution is probably controlled by some
combination of (a) Pb mineral solubility equilibria; (b) adsorption reactions of dissolved Pb
phases on inorganic surfaces (e.g., oxides of Al, Fe, Si, Mn, etc., clay minerals); and
(c) adsorption reactions of dissolved Pb phases on soil organic matter. Dissolved Pb phases in
soil solution can be some combination of Pb2+ and its hydrolysis species, Pb bound to dissolved
organic matter, and Pb complexes with inorganic ligands such as Cl and SO42 . Alkaline soils
typically have solutions supersaturated with respect to PbCOs, Pb3(CO3)2(OH)2, Pb(OH>2,
Pb3(PO4)2, Pb5(PO4)3(OH), and Pb4O(PO4)2. Pb phosphate minerals in particular, are very
insoluble, and calculations based on thermodynamic data predict that these phases will control
dissolved Pb in soil solution under a variety of conditions. However, certain chelating agents,
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such as dissolved organic matter can prevent the precipitation of Pb minerals and the natural
formation of these minerals has not yet been observed in terrestrial ecosystems (see AX7.1.2.1).
Soil solution dissolved organic matter content and pH typically have a very strong
positive and negative correlation, respectively, with the concentration of dissolved Pb species.
In the case of adsorption phenomena, the partitioning of Pb2+ to the solid phase is also controlled
by total metal loading: high Pb loadings will result in a lower fraction partitioned to the solid
phase. It has been found that only a fraction of the total Pb in solution was actually Pb2+ in soils
treated with leaf compost. The fraction of Pb2+ to total dissolved Pb ranged from <1 to 60%,
depending on pH and the availability of Pb-binding ligands. In acidic soils, Al species can
compete for sites on natural organic matter and inhibit Pb binding to surfaces.

Tracing the Fate of Atmospherically Delivered Lead
Radiogenic Pb isotopes offer a powerful tool for separating anthropogenic Pb from natural
Pb derived from mineral weathering (see AX7.1.2.2). This is particularly useful for studying Pb
in mineral soil, where geogenic Pb often dominates. The ore bodies from which anthropogenic
Pb are typically derived are usually enriched in 207Pb relative to 206Pb and 208Pb when compared
with Pb found in granite rocks. Uranium-238 series 210Pb also provides a tool for tracing
atmospherically delivered Pb in soils. Fallout 210Pb is deposited onto forests via wet and dry
deposition, similar to anthropogenic Pb deposition in forests and is thusly useful as a tracer for
non-native Pb in soils. 210Pb is convenient to use for calculating the residence time of Pb in soil
layers because its atmospheric and soil fluxes can be assumed to be in steady-state at undisturbed
sites.
Researchers assessing the fate of atmospheric Pb in soils have also relied on repeated
sampling of soils and vegetation for total Pb. This technique works best when anthropogenic Pb
accounts for the vast majority of total Pb in a particular reservoir. Evans et al. (2005), for
example, have noted that surface soils sampled relatively recently demonstrate that the upper soil
horizons (O + A horizons) have been retaining most of the anthropogenic Pb burden introduced
to the systems during the 20th century, and others have suggested that lateral and vertical
movement of organic particles dominated Pb transport in the soil profile (see AX7.1.2.2).
By describing the movement of atmospherically-delivered Pb in terrestrial ecosystems, we
can begin to predict the Pb inventories of various ecosystem compartments that a particular
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atmospheric deposition rate will support. This type of information is very pertinent to air quality
issues. For example, if the rate of Pb loss is known for a particular soil horizon and reasonable
assumptions can be made about biogeochemical cycling and chemical weathering inputs, then
steady-state Pb concentrations can be calculated for any constant deposition rate (in mass of Pb
deposited per square meter). First-order rate loss constants, k, have been calculated for organic
horizons using forest floor inventories, radiogenic 207Pb tracer techniques, and fallout 210Pb.
First order rate loss constants vary substantially, ranging between -0.003 to -0.6 (1/y),
depending on soil type and climate. First-order rate loss techniques used to model forest floor Pb
dynamics at the Hubbard Brook Experimental Forest in New Hampshire revealed that with
steady Pb deposition at 0.0065 kg/ha/y, the forest floor would reach a steady-state Pb
concentration of 1.4 ppm. Calculated steady-state Pb contents of different ecosystem
compartments can then be compared with experimentally-derived toxicity thresholds to put
deposition rates into context with the terrestrial ecosystem.

Uptake into Plants and Invertebrates
Recent work supports previous conclusions that the form of metal tested, and its
speciation in soil, influence uptake and toxicity to plants and invertebrates. The oxide Pb form is
less toxic than the chloride or acetate forms, which are less toxic than the nitrate form of Pb.
However, these results must be interpreted with caution, as the counted on (e.g., the nitrate ion)
may be contributing to the observed toxicity (see AX7.1.3.1). Most Pb is taken up by plants via
the symplastic route (through cell membranes) and remains in the roots, with little translocation
to shoots, leaves, or other plant parts. Different species of plants and invertebrates accumulate
different amounts of lead.

Detoxification in Plants and Invertebrates
Lead may be deposited in root cell walls as a detoxification mechanism, and this may be
regulated by calcium precipitates in the cell wall. The oxalate content in root and root exudates
may reduce the bioavailability of Pb in soil, and constitute an important tolerance mechanism.
Other hypotheses put forward recently include (a) the presence of sulfur ligands and (b) the
sequestration of Pb in old leaves as detoxification mechanisms. Lead detoxification has not been
studied extensively in invertebrates. Glutathione detoxification enzymes were measured in two
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species of spider; and Pb may be stored in waste nodules in earthworms or as pyromorphite in
nematodes.

Physiological Effects
The effects on heme synthesis, as measured by 5-aminolaevulinic acid dehydratase
(ALAD) activity and protoporphyrin concentration, primarily, were well-documented in the 1986
AQCD (U.S. Environmental Protection Agency, 1986) and continue to be studied in birds and
mammals. However, Henny et al. (1991) caution that changes in ALAD and other enzyme
parameters are not always related to adverse effects, but may simply indicate exposure. Other
effects on plasma enzymes that may damage other organs have been reported (see AX7.1.3.3).
Lead also may cause lipid peroxidation which may be alleviated by Vitamin E, although Pb
poisoning may still result. Also, changes in fatty acid production have been reported, which may
influence immune response and bone formation.

Response Modification
Genetics, biological factors, physical/environmental factors, nutritional factors, and
interactions with other pollutants can all modify terrestrial organism response to Pb (see
AX7.1.3.4). Some species are more sensitive to Pb than others. For example, Fisher 344 rats
were found to be more sensitive to Pb than Sprague-Dawley rats. Also, younger animals are
more sensitive than older animals, and females generally more so than males. Too, monogastric
animals are more sensitive than ruminants, insectivorous mammals may be more exposed to Pb
than herbivores, and higher tropic-level consumers may be less exposed than lower trophic-level
organisms. Diets deficient in nutrients (including low calcium) result in increased uptake of Pb
and greater toxicity in birds, relative to diets containing adequate nutrient levels. Data on effects
of Pb interactions with other metals vary, depending on the endpoint measured, the tissue
analyzed, the animal species, and the metal combination.
Mycorrhizal fungi may ameliorate Pb toxicity until a threshold is surpassed, which may
explain why some studies show increased uptake into plants while others show no difference or
less uptake. Uptake of lead into plants and soil invertebrates increases with a decrease in soil
pH. However, calcium content, organic matter content, and cation exchange capacity of soils
also can significantly influence Pb uptake into plants and invertebrates (see AX7.1.3.4).
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Primary Producers
Effects of lead on terrestrial plants include decreased photosynthetic and transpiration
rates, and decreased growth and yield. The phytotoxicity of lead is considered to be relatively
low compared to other metals, and there are few reports of phytotoxicity from Pb exposure under
field conditions. Data on phytotoxicity were recently reviewed for development of ecological
soil screening levels (Eco-SSL) (U.S. Environmental Protection Agency, 2005b). Many of the
toxicity data presented in U.S. Environmental Protection Agency (2005b) are lower (i.e., they
represent greater toxicity) than those discussed in the 1986 Lead AQCD (U.S. Environmental
Protection Agency, 1986), although both documents acknowledge that toxicity is observed over a
wide range of concentrations of Pb in soil (tens to thousands of mg/kg soil). This may be due to
many factors, such as the soil conditions (e.g., pH, organic matter) and differences in
bioavailability of the Pb in spiked soils, perhaps due to lack of equilibration of the Pb solution
with the soil after spiking. Most phytotoxicity data continue to be developed for agricultural
plant species (i.e., vegetable and grain crops). Few data are available for trees or native
herbaceous plants, although two of the five ecotoxicological endpoints used to develop the Eco-
SSL were for trees and two were for clover.

Consumers
Lead effects on avian and mammalian consumers include decreased reproduction, growth,
and survival, as well as effects on development and behavior. Only relatively few field effects
data exist for consumers, except from sites with multiple contaminants, for which it is difficult to
attribute toxicity specifically to Pb. Much of the avian and mammalian toxicity data recently
reviewed for the development of Eco-SSLs (U.S. Environmental Protection Agency, 2005b) are
lower than those discussed in the 1986 Lead AQCD (i.e., the Eco-SSL document describes
studies which report greater toxicity of Pb to various organisms) although EPA (U.S.
Environmental Protection Agency, 2005b) recognizes that toxicity is observed over a wide range
of doses (<1 to > 1,000 mg Pb/kg bw-day). Most toxicity data for birds are derived from chicken
and quail studies, and most data for mammals are derived from laboratory rat and mouse studies.
Data derived for other species would contribute to increased understanding of Pb toxicity,
particularly for wildlife species with different gut physiologies. In addition, data derived using
environmentally-realistic exposures, such as from Pb-contaminated soil and food may be
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recommended. Finally, data derived from inhalation exposures that evaluate endpoints such as
survival, growth, and reproduction would enhance understanding the implications of airborne
releases of Pb.

Decomposers
Lead effects on soil invertebrates include decreased survival, growth and reproduction.
Effects on microorganisms include changes in nitrogen mineralization and in enzyme activities.
Recent data on Pb toxicity to soil invertebrates and microorganisms are consistent with those
reported in the 1986 Lead AQCD, with toxicity generally observed at concentrations of 100's to
1,000's of mg Pb/kg soil. Studies on microbial processes may be influenced significantly by soil
parameters, and the significance of the test results is not clear.

Ecological Soil Screening Levels (Eco-SSLs)
Eco-SSLs are concentrations of contaminants in soils that would result in little or no
measurable effect on ecological receptors. They were developed by U.S. EPA for use in the
screening level assessments at Superfund sites to identify those contaminants needing further
investigation, and also to identify those contaminants that are not of potential ecological concern
and do not need to be considered in the subsequent analyses. However, several conservative
factors were incorporated into their development. For the plant and invertebrate Eco-SSLs,
studies were scored to favor relatively high bioavailability. For wildlife Eco-SSLs, only species
with a clear exposure link to soil were considered (generalist species, species with a link to the
aquatic environment, or species which consume aerial insects were excluded), simple diet
classifications were used (100% plants, 100% earthworms or 100% animal prey) when in reality
wildlife consume a varied diet, species were assumed to forage exclusively at the contaminated
site, relative bioavailability or Pb in soil and diet was assumed to be 1, and the TRV was selected
as the geometric mean of NOAELs unless this value was higher than the lowest bounded
LOAEL for mortality, growth or reproduction. The Eco-SSLs are intentionally conservative in
order to provide confidence that contaminants which could present an unacceptable risk are not
screened out early in the evaluation process. That is, at or below these levels, adverse effects are
considered unlikely. Due to conservative modeling assumptions (e.g., metal exists in most toxic
form or highly bioavailable form, high food ingestion rate, high soil ingestion rate) which are
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common to screening processes, several Eco-SSLs are derived below the average background
soil concentration for a particular contaminant. The Pb Eco-SSLs for terrestrial plants, birds,
mammals, and soil invertebrates are 120 mg/kg, 11 mg/kg, 56 mg/kg, and 1700 mg/kg,
respectively. (For additional information see Annex Section AX7.1.4).

Effects of Lead on Natural Terrestrial Ecosystems
Few significant effects of Pb pollution have been observed at sites that are not near point
sources of Pb. At present, industrial point sources such as smelter sites represent the greatest Pb-
related threat to the maintenance of sustainable, healthy, diverse, and high-functioning terrestrial
ecosystems in the United States. However, assessing the risks specifically associated with Pb is
difficult because these sites also experience elevated concentrations of other metals and because
of effects related to SC>2 emissions. Terrestrial ecosystems may respond to stress in a variety of
ways, including reductions in the vigor and/or growth of vegetation, reductions in biodiversity,
and effects on energy flow and biogeochemical cycling.

Influence of Acidification
Like most metals, the solubility of Pb increases as pH decreases, suggesting that enhanced
mobility of Pb should be found in ecosystems under acidification stress. However, Pb is also
strongly bound to organic matter in soils and sediments. Reductions in pH may cause a decrease
in the solubility of dissolved organic matter (DOM), due to the protonation of carboxylic
functional groups. Because of the importance of Pb complexation with organic matter, lower
DOM concentrations in soil solution resulting from acidification may offset the increased
solubility of Pb and hence decrease the mobility of the organically bound metal. Increased
mobility was only observed in very acidic soils, those with pH <4.5 (see AX7.1.5.1).
Acidification also may enhance Pb export to drainage water in very sandy soils that have limited
ability to retain organic matter.

Influence of Land Use and Industry
Changes in land use represent potentially significant changes in the cycling of organic
matter in terrestrial ecosystems. Conversion of pasture and croplands to woodlands changes the
nature and quantity of organic matter inputs to the soil. The introduction of industrial activity
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may have consequences for organic matter cycling, and subsequently, Pb mobilization. In one
rare long-term study of polluted soils, loss of soil carbon was found to induce the mobilization
and loss of Pb from terrestrial ecosystems. However, it is worth noting that the decline in soil Pb
was considerably smaller than the decline in organic carbon. This suggests that Pb mobilized
during organic matter decomposition can resorb to remaining organic matter or perhaps to
alternate binding sites (e.g., Fe and Mn oxides).
Forest harvesting represents a severe disruption of the organic matter cycle in forest
ecosystems. However, observations from clear-cut sites in the United States and Europe indicate
that forest harvesting causes little or no mobilization or loss of Pb from forest soils. The
principal risk associated with forest harvesting is the loss of Pb in particulate form to drainage
waters through erosion.

Effects Observed Around Industrial Point Sources
The effects of Pb exposure on natural ecosystems are confounded by the fact that Pb
exposure cannot be decoupled from other factors that may also affect the ecosystem under
consideration. Principal among these factors are other trace metals and acidic deposition.
Emissions of Pb from smelting and other industrial activities are accompanied by other trace
metals (e.g., Zn, Cu, Cd) and sulfur dioxide (802) that may cause toxic effects independently or
in concert with Pb.
Natural terrestrial ecosystems near smelters, mines, and other industrial plants have
exhibited a variety of effects related to ecosystem structure and function. These effects include
decreases in species diversity, changes in floral and faunal community composition, and
decreasing vigor of terrestrial vegetation (see AX7.1.5.2). Subsequent to the effects on
vegetation, wind and erosion may remove litter and humus, leaving bare mineral soil, a nearly
sterile environment in which very little energy transfer takes place. In a rare case, metal
pollution around a Pb-Zn smelter near Bristol, England has not resulted in the loss of oak
woodlands within 3 km of the smelter, despite significant accumulation of Pb, Cd, Cu, and Zn in
soils and vegetation. However, the high metal concentrations have favored the growth of metal-
tolerant species in the woodland (see AX7.1.5.2). The effects of Pb and other chemical
emissions on terrestrial ecosystems near smelters and other industrial sites decrease downwind
from the Pb source. Several studies using the soil Pb burden as an indicator have shown that
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much of the contamination occurs within a radius of 20 to 50 km around the emission source.
Elevated metal concentrations around smelters have been found to persist despite significant
reductions in emissions. The confounding effect of other pollutants makes the assessment of
Pb-specific exposure-response relationships impossible at the whole-ecosystem level.

Influence of Climate Change
Atmospheric Pb is not likely to contribute significantly to global climate change. The
potential linkages between climate-related stress and Pb cycling are poorly understood. Climate
effects related to alterations in organic matter cycling may influence Pb migration. For example,
an increase in temperature leading to increased rates of organic matter decomposition could lead
to temporary increases in DOM concentrations and smaller steady-state pools of soil organic
matter. There also is some evidence for recent increases in the frequency of soil freezing events
in the northeastern United States. Soil freezing occurs when soils have little or no snow cover to
insulate them from cold temperatures and results in an increased release of nitrate and DOC from
the O horizons of forest soils. Increased fluctuations in precipitation may induce more frequent
flooding, potentially increasing inputs of Pb and other metals to floodplain soils. All of these
factors could result in increased concentrations of Pb in waters draining terrestrial ecosystems.

8.7.2 Aquatic Ecosystems
Sediment Quality Benchmarks and Bioavailability
There are a number of factors in sediment that can influence lead bioavailability to
benthic (sediment) organisms. Although sediment quality criteria have not been formally
adopted, the EPA has published an equilibrium partitioning procedure for developing sediment
criteria for metals. Equilibrium partitioning (EqP) theory predicts that metals partition in
sediment between acid volatile sulfide, pore water, benthic organisms, and other sediment
phases, such as organic carbon. Using this theory, sediment toxicity and organism mortality can
be more reliably predicted by accounting for both the site-specific organic carbon and AVS
concentrations. It should be noted that although EPA is favoring the AVS-SEM approach for
regulating metals in sediments, there is not scientific consensus on this issue. Various studies
suggest that ingestion of sediment particles by benthic organisms is an important exposure route
not accounted for by AVS-SEM or that the AVS-SEM approach may not be the most accurate
approach available for predicting non-toxic and toxic results in laboratory studies.

Speciation of Lead in Aquatic Ecosystems
The speciation of Pb in the aquatic environment is controlled by many factors, such as pH,
salinity, sorption, and biotransformation processes. Lead is typically present in acidic aquatic
environments as PbSC>4, PbCU, ionic Pb, cationic forms of Pb hydroxide, and ordinary hydroxide
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Pb(OH)2. In alkaline, waters common Pb species include anionic forms of Pb carbonate Pb(CO3)
and hydroxide Pb(OH)2. In freshwaters, Pb typically forms strong complexes with inorganic
OH" and CO32 and weak complexes with Cl . The primary form of Pb in freshwaters at low pH
(<6.5) is predominantly Pb2+; and less abundant inorganic forms include Pb(HCO)3, Pb(SO/t)22 ,
PbCl, PbCO3, and Pb2(OH)2CO3. At higher pH (>7.5) Pb forms hydroxide complexes (PbOH+,
Pb(OH)2, Pb(OH)3 , Pb(OH)42 ). Lead speciation in seawater is a function of chloride
concentration and the primary species are PbCl3 > PbCO3 > PbCl2 > PbCl+ > and Pb(OH)+
(seeAX7.2.2.1).
Lead sorption to suspended or bed sediments or suspended organic matter typically
increases with increasing pH, increasing amounts of iron or manganese; and with the polarity of
component particulate matter (e.g., clays). Adsorption decreases with water hardness. At higher
pH, Pb precipitates as Pb(OH)+ and PbHCO3+ into bed sediments. Conversely, at low pH, Pb is
negatively sorbed, i.e., repelled from the adsorbent surface (see AX7.2.2.1). Also, Pb may be
remobilized from sediment due to a decrease in metal concentration in the solution phase,
complexation with chelating agents (e.g., EDTA), and changing redox conditions. Changes in
water chemistry (e.g., reduced pH or ionic composition) can cause sediment Pb to become re-
mobilized and potentially bioavailable to aquatic organisms. Methylation may result in Pb
remobilization, its reintroduction into the aqueous environmental compartment, and its
subsequent release into the atmosphere. However, methylation is not a significant environmental
pathway controlling Pb fate in the aquatic environment.

Lead Concentrations in United States Surface Waters
Nationwide U.S. data for Pb in surface waters, from 1991 onward, were compiled using
the United States Geological Survey's (USGS) National Water-Quality Assessment (NAWQA)
database. Data were compiled from locations categorized as "ambient" or "natural." Ambient
refers to data collected from all sampling locations, while natural refers to data collected from
sampling locations categorized as forest, rangeland, or reference. Summary statistics for surface
water, sediment (bulk, <63 um), and fish tissue (whole body and liver) are summarized in
Table 8-8. Overall, atmospheric sources of Pb have generally decreased as regulations have
removed Pb from gasoline and other products; however, elevated Pb concentrations remain at
sites near ongoing sources, such as near mining wastes or wastewater effluents.
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Table 8-8. Summary of Lead Concentrations in United States Surface Water,
Sediment, and Fish Tissue
Surface Water -
Dissolved
Statistic
n
%ND
Min
Mean
Median
90th %ile
95th %ile
Max
Ambient
3,445
86
0.04
0.66
0.50
0.50
1.10
29.78
Natural
430
88
0.04
0.52
0.50
0.50
0.50
8.40
Sediment -
Bulk, <63 fim
(jig/g dry wt.)
Ambient
1,466
0.48
0.50
120
28
120
200
12,000
Natural
258
1.2
0.50
109
22
66
162
12,000
Fish Tissue (jig/g dry wt.)
Whole Organism
Ambient
332
39
0.08
1.03
0.59
2.27
3.24
22.6
Natural
93
51
0.08
0.95
0.35
1.40
2.50
22.6
Liver
Ambient
559
71
0.01
0.36
0.15
0.59
1.06
12.7
Natural
83
89
0.01
0.28
0.11
0.37
1.26
3.37
%ND = Percentage not detected

Lead concentrations in lakes and oceans were generally found to be much lower than
those measured in the lotic waters assessed by NAWQA. Surface water concentrations of
dissolved Pb measured in Hall Lake, Washington in 1990 ranged from 2.1 to 1015.3 ng/L, and
the average surface water dissolved Pb concentrations measured in the Great Lakes (Superior,
Erie, and Ontario) between 1991 and 1993 were 3.2, 6.0, and 9.9 ng/L, respectively. Lead
concentrations ranged from 3.2 to 11 ng/L across all three lakes. Similarly, 101 surface water
total Pb concentrations measured at the Hawaii Ocean Time-series (HOT) station ALOHA
between 1998 and 2002 ranged from 5 to 11 ng/kg. Based on the fact that Pb is predominately
found in the dissolved form in the open ocean (<90%), dissolved Pb concentrations measured at
these locations would likely have been even lower than the total Pb concentrations reported.
In addition to directly measuring Pb concentrations in various aquatic compartments, it is
useful to study the vertical distribution of Pb. Sediment profiling and core dating is a method
used to determine the extent of accumulation of atmospheric Pb and provides information on
potential anthropogenic sources. Sediment concentration profiles are typically coupled with lead
isotopic analysis. The isotope fingerprinting method utilizes measurements of the abundance of
common lead isotopes (204Pb, 206Pb, 207Pb, 208Pb) to distinguish between natural Pb over geologic
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time and potential anthropogenic sources. Studies of sediment profiles have suggested that
observed increases in Pb concentrations in the upper sediment layer are concomitant with
increases in anthropogenic inputs. Isotopic ratios have been used to link increases in sediment
concentrations with specific anthropogenic sources and to estimate historic records of Pb fluxes
to surface waters and sediments (see AX7.2.2.3).

Lead Uptake
Lead can bioaccumulate in the tissues of aquatic organisms through ingestion of food and
water, and adsorption from water, and can subsequently lead to adverse effects if tissue
concentrations are sufficiently high. The accumulation of Pb is influenced by pH and decreasing
pH favors bioavailability and bioaccumulation. Organisms that bioaccumulate Pb with little
excretion must partition the metal such that it has limited bioavailability, otherwise toxicity will
occur if a sufficiently high concentration is reached (see AX7.2.3.1).

Resistance Mechanisms
Aquatic organisms have various methods to resist the toxic effects of metals such as Pb.
Mechanisms of resistance vary among aquatic biota and may include detoxification and
avoidance responses. Detoxification processes can include translocation, excretion, chelation,
adsorption, and vacuolar storage and deposition. For example, protists and plants produce
intracellular polypeptides that form complexes with Pb. Some macrophytes and wetland plants
have developed translocation strategies for tolerance and detoxification. Various aquatic
invertebrates may sequester Pb in the exoskeleton or have developed specialized excretion
processes. Fish scales and mucous may chelate Pb in the water column and potentially reduce
Pb uptake (see AX7.2.3.2).
Avoidance responses are actions performed to evade a perceived threat. Some aquatic
organisms have been shown to be quite adept at avoiding Pb in aquatic systems, while others
seem incapable of detecting its presence. Snails have been shown to be sensitive to Pb, and
avoid it at high concentrations. Conversely, anuran (frog and toad) species lack an avoidance
response up to 1000 jig Pb/L. Fish avoidance of chemical toxicants has been well established
and is a dominant sublethal response in polluted waters. However, studies examining avoidance
behavior of Pb in fish are lacking (see AX7.2.3.2).
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Physiological Effects of Lead
Physiological effects of Pb on aquatic biota can occur at the biochemical, cellular and
tissue levels of organization. Lead has been shown to affect brain receptors in fish and serum
enzyme activity (e.g., EROD and ALAD) in fish and amphibians. Studies examining Pb effects
on fish blood chemistry have indicated alterations from acute and chronic exposures ranging
from 100 to 10,000 |ig/L. Lead exposure has also been shown to negatively affect the growth of
aquatic invertebrates (see AX7.2.3.3).

Factors that Modify Organism Response to Lead
There are several factors that may modify responses of aquatic organisms to Pb exposure.
These may include the size or age of an organism, genetics, environmental factors (e.g., pH,
salinity), nutrition, and the presence of other contaminants. Lead accumulation in living
organisms is controlled, in part, by metabolic rates and by the physiological conditions of an
organism. Relationships between age, size and Pb body burden in aquatic invertebrates and fish
are variable and depend on many environmental variables (e.g., exposure). For example,
examination of Pb exposure (up to 100 |ig/L) in aquatic invertebrates showed little relationship
between body size and Pb accumulation (MacLean et al., 1996; Canli and Furness, 1993) while
Pb accumulation and fish size were positively correlated (Douben, 1989; Kock et al., 1996).
The genetics of an organism and/or population may alter the response to Pb exposure
through one of two processes: (1) a contaminant may influence selection, by selecting for certain
phenotypes that enable populations to better cope with the chemical, or (2) a contaminant can be
genotoxic, meaning it can produce alterations in nucleic acids at sublethal exposure
concentrations, resulting in changes in hereditary characteristics or DNA inactivation. Genetic
selection has been observed in aquatic organisms due to lead tolerance. Because tolerant
individuals have a selective advantage over vulnerable individuals in polluted environments, the
frequency of tolerance genes will increase in exposed populations over time. Several studies
have shown that heavy metals can alter population gene pools resulting in decreased genetic
diversity. Laboratory studies have shown that Pb exposure at 10 mg Pb2+/mL of blood Pb to
chromosomal aberrations in some aquatic organisms. Lead exposure in water (50 |ig/L) over
four weeks resulted in DNA strand breakage in the freshwater mussel Anodonta grandis. More
recently, similar results (increase in the frequency of chromosomal aberrations and DNA damage
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in kidney cell cultures) were observed in fish (Hoplias malabaricus) fed Pb-contaminated food
over 18, 41, and 64 days (see AX7.2.3.4).
Environmental factors can alter the availability, uptake and toxicity of Pb to aquatic
organisms. A study of the influence of abiotic variables, including dissolved organic carbon
(DOC) on Pb concentrations in freshwater isopods found that, as DOC concentrations increased,
BCFs decreased in P. meridianus and A aquaticus, indicating that DOC acts to inhibit the
availability of Pb to these isopods. Schwartz et al. (2004) collected natural organic matter
(NOM) from several aquatic sites across Canada and investigated the effects of NOM on Pb
toxicity in rainbow trout (Oncorhynchus mykiss). The results showed that NOM in test water
almost always increased LT50 (time to reach 50% mortality), and optically dark NOM tended to
decrease Pb toxicity more than did optically light NOM in rainbow trout. Studies generally
agree that the toxicity of Pb decreases as pH increases. As pH decreases, Pb becomes more
soluble and more readily bioavailable to aquatic organisms (Weber, 1993). Acute and chronic
toxicity of Pb increases with decreasing water hardness, as Pb becomes more soluble and
bioavailable to aquatic organisms (Home and Dunson, 1995c; Borgmann et al., 2005). There is
some evidence that water hardness and pH work together to increase or decrease Pb toxicity
(seeAX7.2.3.4).
High Ca2+ concentrations have been shown to protect against the toxic effects of Pb.
Ca2+ affects the permeability and integrity of cell membranes and intracellular contents. As Ca2+
concentrations decrease, the passive flux of ions (e.g., lead) and water increases. Finally,
increasing salinity was found to decrease Pb toxicity. The reduction in toxicity was attributed to
increased complexation of Pb2+ with Cl ions (see AX7.2.3.4).
Also, nutrients (e.g., nitrate, carbonate) have been shown to affect Pb toxicity in some
aquatic organisms. A study of blue-green algae (Synechococcus aeruginosus) exposed to
200 mg Pb/L indicated that additional nitrogen, phosphates, and some carbon sources (including
sodium acetate, citric acid and sodium carbonate) all protected the algae from Pb toxicity at
200 mg Pb/L. The protective mechanism is still not clear. One hypothesis was that nutrients
were able to reverse toxic effects. The second hypothesis was that nutrients directly interacted
with Pb, in some way sequestering the metal so as to inhibit its metabolic interaction with the
organism (see AX7.2.3.4).
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Interactions with Other Pollutants
Predicting the response of organisms to mixtures of chemicals is a daunting task. There
are two major approaches to predict mixture toxicity including: 1) examining the combined
mode of action of the individual mixture substances; and 2) determining whether an organism
response to the mixture is additive, or some deviation from additive (synergistic or antagonistic).
In addition, researchers may report mixture toxicity in terms of additive concentrations or
additive effects, which can cause confusion in the interpretation of study results. For the studies
presented in Section AX7.2.3.4, the authors primarily report mixture toxicity in terms of additive
concentrations (i.e., the sum of the concentrations of each individual chemical in the mixture will
result in a level of effect similar to the simple sum of the effects observed if each chemical were
applied separately).
When two or more metals compete for the same binding sites or interfere with transport
through cell walls or membranes, the interaction is termed less than strictly additive or
antagonistic. Antagonistic interactions can reduce metal bioavailability when metals are present
in combination, and may lead to reduced potential for toxicity. There are a number of elements
(Ca2+, Cd2+, Mg2+, Na+ and Cl") that act in an antagonistic fashion with Pb (see AX7.2.3.4).
For example, Pb is a well-known antagonist to Ca2+, an essential element required for many
physiological processes in most organisms.
Synergism occurs when the interaction of two or more metals causes an effect that is
greater than the effect observed from the individual metals themselves. Synergism is likely the
result of increased bioavailability of one or more of the metal ions due to the presence of other
metals. Synergistic interactions have been observed with Pb and other metals (Cd, Cu, Ni, Zn)
(see discussion in AX7.2.3.4).
The two most commonly reported Pb-element interactions are between Pb and calcium
and Pb and zinc. Both calcium and zinc are essential elements in organisms, and the interaction
of Pb with these ions can lead to adverse effects both by increased Pb uptake and by a decrease
in Ca and Zn required for normal metabolic functions.

Effects of Lead on Primary Producers
Several studies have been conducted since the 1986 Lead AQCD on the toxicity of Pb to
primary producers. Effects on algal growth (Chlorella vulgaris, Closterium acerosum,
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Pediastrum simplex., Scenedesmus quadricauda\ ranging from minimal to complete inhibition,
have been reported at Pb concentrations between 100 and 200,000 |ig/L. The toxicity of Pb to
aquatic plant growth has been studied using Spirodelapolyrhiza, Azollapinnata, and Lemna
gibba. Test durations ranged from 4 to 25 days and test concentrations ranged between 49.7 and
500,000 |ig/L. Research on aquatic plants has been focused on Pb effects on aquatic plant
growth, chlorophyll and protein content (see AX7.2.4.2).
Algae and other aquatic plants have a wide range in sensitivity to the effects of Pb in
water. Both groups of primary producers experience ECso values for growth inhibition between
-1,000 and >100,000 |ig/L (see AX7.2.4.2). Exposure to Pb in combination with other metals
generally inhibits growth less than exposure to Pb alone. Studies have shown that Pb adversely
affects the metabolic processes of nitrate uptake, nitrogen fixation, ammonium uptake, and
carbon fixation. Lead in combination with nickel or chromium produced synergistic effects for
nitrate uptake, nitrogenase activities, ammonium uptake, and carbon fixation.

Effects of Lead on Consumers
The 1986 Lead AQCD (U.S. Environmental Protection Agency, 1986a) reported that
hematological and neurological responses are the most commonly reported Pb effects on aquatic
vertebrates. These effects include red blood cell destruction and inhibition of the enzyme
ALAD, required for hemoglobin synthesis. The lowest reported exposure concentration causing
either hematological or neurological effects was 8 jig Pb/L (U.S. Environmental Protection
Agency, 1986a).
More recent literature on the toxicity of lead to fish and aquatic invertebrates has been
summarized by Eisler (2000). Exposure of invertebrates to Pb can lead to adverse effects on
reproduction, growth, survival, and metabolism. Water-borne Pb is highly toxic to aquatic
organisms, with toxicity varying, depending on the species and life stage tested, duration of
exposure, the form of Pb tested, and water quality characteristics (see AX7.2.4.1). Among the
species tested, aquatic invertebrates, such as amphipods and water fleas, were the most sensitive
to Pb effects, with adverse effects being reported at water Pb concentrations as low as 0.45 |ig/L
(range: 0.45 to 8000 |ig/L). Freshwater fish demonstrated adverse effects at concentrations
ranging from 10 to >5400 |ig/L, generally depending upon water quality parameters (e.g., pH,
hardness, salinity). Amphibians tend to be relatively tolerant of Pb, but may exhibit decreased
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enzyme activity (e.g., ALAD reduction) and changes in behavior (e.g., hypoxia response
behavior). Lead tends to be more toxic in longer-term exposures, with chronic toxicity
thresholds for reproduction in water fleas ranging as low as 30 |ig/L.

Effects of Lead on Natural Aquatic Ecosystems
Lead exposure may adversely affect organisms at different levels of organization, i.e.,
individual organisms, populations, communities, or ecosystems. Generally, however, there is
insufficient information available for single materials in controlled studies to permit evaluation
of specific impacts on higher levels of organization (beyond the individual organism). Potential
effects at the population level or higher are, of necessity, extrapolated from individual level
studies. Available population, community, or ecosystem level studies are typically conducted at
sites that have been contaminated or adversely affected by multiple stressors (several chemicals
alone or combined with physical or biological stressors). Therefore, the best documented links
between lead and effects on the environment are with effects on individual organisms.
Recent studies on exposure to Pb in laboratory studies and simulated ecosystems indicate
that Pb may alter species competitive behaviors, predator-prey interactions, and contaminant
avoidance behaviors. Alteration of these interactions may have negative effects on species
abundance and community structure (see AX7.2.5.2). For example, reduced avoidance
behaviors have been observed at Pb concentrations ranging from 0.3 to 1.0 mg/L. The feeding
behaviors of competitive species in some aquatic organisms are also influenced by the presence
of Pb. Lead (6 to 80 mg/L) has also been found to reduce primary productivity and increase
respiration in an algal community; and laboratory microcosm studies found reduced species
abundance and diversity in protozoan communities exposed to 0.02 to 1 mg Pb/L. Lastly,
numerous field studies have associated the presence or bioaccumulation of Pb with reductions in
species abundance, richness, or diversity, particularly in benthic macroinvertebrate communities.
In natural aquatic ecosystems, Pb is often found coexisting with other metals or other stressors.
Thus, understanding the effects of Pb in natural systems is challenging given that observed
effects may be due to cumulative toxicity from multiple stressors.
The effects of Pb have primarily been studied in relation to point source pollution rather
than area-wide atmospheric deposition. Thus, the effects of atmospheric Pb on aquatic
ecological condition remain to be defined. There is a paucity of data in the general literature that
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explores Pb effects in conjunction with all or several of the various components of ecological
condition as defined by the EPA (Young and Sanzone, 2002). However, numerous studies are
available that associate the presence of Pb with effects on biotic conditions.

8.7.3 Application of Critical Loads to Terrestrial and Aquatic Ecosystems
For the purpose of this section, critical loads are defined as threshold deposition rates of
air pollutant that current knowledge indicates will not cause long-term adverse effects to
ecosystem structure and function (see Section 8.3.1). A combinatorial application of critical
limit and critical load allows one to assess current risk while simultaneously estimating future
risk from exposure to a chemical. Figure 8-13 shows that four combinations of critical load and
limit exceedance or non-exceedance are possible for a given ecosystem (Figure 1 of De Vries
et al. [2004]). For example, if a current risk is indicated by an exceedance of the critical limit for
Pb due to historical Pb deposition, but current inputs of Pb to the ecosystem are below the critical
load (upper right corner), the critical load model predicts that Pb concentrations will fall below
the critical limit at some point in the future if Pb deposition is maintained at the present level.
If current soil concentrations are below the critical limit (lower left corner), inputs greater than
the critical load will not result in exceedance of the critical limit for some period of time, but
continued exceedance of a critical load will eventually lead to an exceedance of the critical limit.
The time until a critical limit is exceeded (critical time) can also be predicted using the critical
load model. This requires knowledge of current concentrations, the critical load, and predicted
deposition rates. Critical times may be useful for setting priorities between ecosystems with
critical load exceedances or between different chemicals.

Calculation of Critical Loads
This section summarizes various methods used to calculate critical loads (De Vries et al.,
2001, 2002, 2004; Groenenberg et al., 2002), with an emphasis on the most recent material.

Critical Limits
To determine the critical limit, effects-based criteria for the major ecological endpoints
should be developed for the ecosystem of concern. Criteria may be developed for any receptor
that is exposed to the chemical of concern deposited in the ecosystem. In terrestrial ecosystems,
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No critical load
exceedance
Critical load
exceedance
No critical limit exceedance
No damage at present or foreseen:

Concentration
critical
present
CL
PL2
PL1
Time
T^ Keep the Present Load
(more stringent than Critical
Load)
Future damage foreseen:
Concentration
critical
present
PL4
Time
•^ Consider Critical ] ,oad
(emissions must decrease, even if
concentrations in the ecosystem
are allowed to increase further at
critical load)
Critical limit exceedance
Present damage but recovery in progress:

Concentration

present

critical
- CL
PL2
TT
Time
T^ Keep the Present Load
(more stringent than Critical
Load)
or
•^ Consider Target Load to reach
the critical limit in a defined time
period (more stringent than
Present damage, no recovery foreseen:
Concentration
present
critical
PL4
- PL3
- CL
TT
Time
•^ Consider Critical 1 ,oad
(decrease of concentrations in the
ecosystem down to critical limit
in the long term)
or
•^ Consider Target Load to reach
the critical limit in a defined time
period (more stringent than
Critical Load)
CL - Critical load; PL - present load (2 cases); SL - Stand-still load; TL - Target load; TT - Target time
Figure 8-13. The predicted development of metal concentrations in ecosystems for four
cases of exceedance or non-exceedance of critical limits and critical loads of
heavy metals, respectively.

Source: Taken from DeVries et al. (2004).
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possible ecological endpoints include effects from direct contact of invertebrates or plants
with soil and ingestion of plants by herbivores. Effects-based criteria for use in defining the
critical limit should be derived from ecotoxicological data appropriate to the most sensitive
endpoint (De Vries et al., 2004). Regardless of the selected endpoint, the critical limit should be
defined as a concentration in the medium that receives the depositional load, typically soil in
terrestrial ecosystems and surface water in aquatic ecosystems. To derive these values, uptake
and/or food- chain modeling may be necessary.
Criteria for Pb vary widely and can be the largest source of uncertainty in a critical load
calculation. One reason for the wide range in estimates of effects criteria is that Pb speciation is
often not taken into account. This can result in variation in estimates of concentration for total
Pb that is associated with adverse effects, since the fraction of Pb available to cause a toxic effect
depends on chemical factors such as the pH or organic matter content. To develop effects-based
criteria applicable to media with a pH or organic matter content different from the test medium,
it is more appropriate to develop criteria based on the free concentration of Pb rather than the
total Pb concentration.

Models
Critical loads for heavy metals are typically calculated using a steady state model that
ignores internal metal cycling and keeps the calculations as simple as possible (De Vries et al.,
2004). The critical load is equal to the atmospheric input flux, which equals the sum of the
output fluxes from the system minus the other input fluxes (e.g., weathering) when the
concentration of Pb is at the critical limit. The input flux of heavy metals via weathering is
sometimes neglected, because quantitative estimates are highly uncertain, and weathering is
generally thought to be a relatively minor process (see Section 7.3.4.2).

Critical Loads in Terrestrial Ecosystems
Critical loads for Pb have been calculated using simple mass balance, dynamic, and
probabilistic models for forested and agricultural land in Europe and Canada in a handful of
preliminary studies. The methods and model assumptions used to calculate critical loads vary
widely between these studies and little attempt has been made to validate the models that were
used, so it is not known how much various simplifying assumptions affect the results.
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In spite of the variation in methods and model assumptions used to calculate critical loads
for Pb, some general conclusions may be drawn. The critical limit is the most important value
for determining the value of the critical load. Wide variations in available effects levels makes
this parameter one of the most important sources of uncertainty when calculating critical loads in
terrestrial ecosystems. Spatial variations in critical loads for Pb are largely controlled by net
runoff. Weathering and uptake by harvestable vegetation were less important. The time to reach
steady state is several hundred years in the two studies that used dynamic models to determine
critical loads.

Critical Loads in Aquatic Ecosystems
Doyle et al. (2003) modeled critical loads in surface water bodies assuming complete
mixing with dilution water entering from the terrestrial catchment area. Loss of metal was also
assumed to occur through downstream flushing and burial in sediment. Transfer of metal to
sediment was modeled as a first-order process dependant on the dissolved concentration and pH.
The inputs to the model included the following: water body area, terrestrial catchment area,
water body depth, sediment accumulation rate, thickness of biologically active sediment, net
precipitation, and water pH. The fist-order rate constant for transfer to sediment was correlated
with pH. The model reached steady state within a few years. Transfer of Pb from the terrestrial
catchment to the water body was neglected, because the time to steady state could be on the
order of 10,000 years if the model included this source of Pb. However, the authors cited a
separate calculation that indicated that neglect of transfer of Pb from the catchment may lead to a
5-fold underestimation of Pb concentrations in the surface water. These results indicate that Pb
run-off from soil is more important than direct atmospheric deposition to the surface water
bodies considered in this study. Due to the long times required to achieve steady state, the
critical load methodology may not be appropriate for Pb in aquatic systems.

Limitations and Uncertainties
The largest sources of uncertainty identified in studies of critical loads for Pb include the
following: (1) steady-state assumption; (2) derivation of the critical limit; (3) Pb speciation; and
(4) soil runoff as an input to aquatic ecosystems
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The critical load is calculated for steady state conditions, but the time for Pb to reach
steady-state concentrations can be as long as several centuries. Thus, dynamic models are often
used to predict Pb concentrations over shorter time frames. Dynamic modeling requires
additional knowledge about current concentrations in the considered ecosystem. For regulatory
purposes, use of dynamic modeling requires that a target time be set in order to calculate a
critical load.
Speciation strongly influences the toxicity of Pb in soil and water and partitioning
between dissolved and solid phases determines the concentration of Pb in soil drainage water,
but it has not been taken into account in most of the critical load calculations for Pb performed to
date. Recent guidance for heavy metals has begun to emphasize the importance of speciation to
critical load calculations and suggest methods to calculate speciation (De Vries et al., 2004).
To this end, Lofts et al. (2004) developed critical limit functions for several metals, including Pb,
that take into account the effects of pH, organic matter, and the protective effects of cations on
speciation.
Runoff of Pb from soil may be the major source of Pb into aquatic systems. However,
little attempt has been made to include this source into critical load calculations for aquatic
systems due to the complexity of including this source in the critical load models.
Preliminary efforts to calculate critical loads for Pb in terrestrial and aquatic ecosystems
have so far relied on a variety of calculation methods and model assumptions. Efforts are
ongoing to refine and standardize methods for the calculation of critical loads for heavy metals
which are valid in the context of CLPTRP. At this time, the methods and models commonly
used for the calculation of critical loads have not been validated for Pb. Many of the methods
neglect the speciation of Pb when estimating critical limits, the uptake of Pb into plants, and the
outflux of Pb in drainage water, limiting the utility of current models.
Future efforts should focus on fully incorporating the role of Pb speciation into critical
load models, and validating the assumptions used by the models.
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