EPA/600/R-10/075F | June 2013 | www.epa.gov
United States
Environmental Protection
Agency
Integrated Science Assessment
for Lead
ESBHSEHSEBESHB
Contains Errata Sheet created 5/12/2014
Office of Research and Development
National Center for Environmental Assessment, Research Triangle Park, NC
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xvEPA
United States June 2013
Environmental Protect,on EPA/600/R-10/075F
Errata Sheet created 5/12/2014
Integrated Science Assessment
for Lead
National Center for Environmental Assessment-RTF Division
Office of Research and Development
U.S. Environmental Protection Agency
Research Triangle Park, NC
-------
DISCLAIMER
This document has been reviewed in accordance with U.S. Environmental Protection Agency (EPA)
policy and approved for publication. Mention of trade names or commercial products does not
constitute endorsement or recommendation for use.
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Errata Sheet Created 5/12/2014
For the document titled: Integrated Science Assessment (ISA) for Lead(Pb), June 2013, Final
Table or Figure Page Correction or Comments
3-37 Replaced "at a concentration of 0.001 ug/m3" with
"at a concentration of 1 ug/m3"
Figure 4-2, Table 4-3, 4-62,
Figure 4-15, Table 4-16 4-70,
4-72,
4-76,
4-124,
4-254,
4-256,
4-284,
4-285
Two errors were identified in the dataset analyzed by
Lanphear et al. (2005) that resulted in slight
modifications to values reported in the ISA.
The page numbers of the ISA that are affected by these
errors are noted here; however, the final ISA has not
been corrected. Rather, recalculations using the
corrected dataset are presented in the Table on page A-2
of the memo to the docket (Kirrane and Patel, 2014),
EPA-HQ-ORD-2011-0051-0050. EPA has concluded that
the conclusions drawn in the 2013 Pb ISA are not
materially affected by these newly identified errors.
Table 6-5, column 5
6-307 Replaced "delayed in the 100 mg/L treatment" with
"delayed in the 100 ug Pb/L treatment"
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CONTENTS
LEAD PROJECT TEAM xvii
AUTHORS, CONTRIBUTORS, AND REVIEWERS xx
CLEAN AIR SCIENTIFIC ADVISORY COMMITTEE LEAD NAAQS REVIEW PANEL xxvi
ACRONYMS AND ABBREVIATIONS xxvii
PREAMBLE xliv
Process of ISA Development xliv
Figure I Illustration of the key steps in the process of the review of National
Ambient Air Quality Standards. xlv
Figure II Illustration of processes for literature search and study selection
used for development of ISAs. xlvi
Figure III Characterization of the general process of ISA development. /
EPA Framework for Causal Determination //
Evaluating Evidence for Inferring Causation li
Consideration of Evidence from Scientific Disciplines
Application of Framework for Causal Determination Iviii
Table I Aspects to aid in judging causality. lix
Determination of Causality Ix
Table II Weight of evidence for causal determination. Ixii
Quantitative Relationships: Effects on Human Populations Ixiii
Quantitative Relationships: Effects on Ecosystems or Public Welfare Ixiv
Concepts in Evaluating Adversity of Health Effects Ixv
Concepts in Evaluating Adversity of Ecological Effects Ixvi
References for Preamble Ixvii
LEGISLATIVE AND HISTORICAL BACKGROUND Ixix
Legislative Requirements for the NAAQS Review Ixix
History of the NAAQS for Pb Ixxi
References for Legislative and Historical Background Ixxv
EXECUTIVE SUMMARY Ixxvii
Introduction Ixxvii
Sources, Fate and Transport of Lead in the Environment, and the Resulting Human Exposure and
Dose Ixxviii
Figure ES-1 Conceptual model of multimedia Pb exposure. Ixxx
Integrative Overview of Health and Ecological Effects Ixxxi
Figure ES-2 Schematic representation of the relationships between the various
MOAs by which Pb exerts its effects. Ixxxii
Health Effects of Pb Ixxxii
Table ES-1 Summary of causal determinations for the relationship between
exposure to Pb and health effects. Ixxxiii
Ecological Effects of Pb Ixxxviii
Table ES-2 Summary of causal determinations for the relationship between Pb
exposure and effect on plants, invertebrates, and vertebrates. xc
Policy Relevant Considerations xciii
Summary xcv
References for Executive Summary xcvi
CHAPTER 1 INTEGRATIVE SUMMARY 1-1
1.1 ISA Development and Scope 1-1
111
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1.2 Ambient Pb: Source to Concentration 1-6
1.2.1 Sources, Fate and Transport of Ambient Pb 1-6
1.2.2 Monitoring and Concentrations of Ambient Air Pb 1-7
1.2.3 Ambient Pb Concentrations in Non-Air Media and Biota 1-8
Table 1-1 Pb concentrations in non-air media and biota considered for
ecological assessment. 1-10
1.3 Exposure to Ambient Pb 1-11
1.4 Toxicokinetics 1-12
1.5 Pb Biomarkers 1-13
1.6 Health Effects 1-14
Table 1-2 Summary of causal determinations for the relationship between
exposure to Pb and health effects. 1-15
1.6.1 Nervous System Effects 1-20
1.6.2 Cardiovascular Effects 1-27
1.6.3 Renal Effects 1-30
1.6.4 Immune System Effects 1-31
1.6.5 Hematological Effects 1-33
1.6.6 Reproductive and Developmental Effects 1-34
1.6.7 Cancer 1-37
1.7 Ecological Effects of Pb 1-38
1.7.1 Summary of Effects on Terrestrial Ecosystems 1-39
1.7.2 Summary of Effects on Aquatic Ecosystems 1-43
1.7.3 Determinations of Causality for Effects on Ecosystems 1-48
Table 1-3 Summary of Pb causal determinations for plants, invertebrates, and
vertebrates. 1-48
1.8 Integration of Health and Ecological Effects 1-61
Table 1-4 Summary of causal determinations for health and ecological effects. _ 1-61
1.8.1 Modes of Action Relevant to Downstream Health and Ecological Effects 1-63
Table 1-5 Modes of action, their related health effects, and information on
concentrations eliciting the MOAs. 1-64
1.9 Policy Relevant Considerations 1-68
1.9.1 Public Health Significance 1-68
Figure 1-1 Distributions of IQ scores. 1-69
1.9.2 Air-Pb-to-Blood-Pb Relationships 1-70
Table 1-6 Summary of estimated slopes for blood Pb to air Pb relationships
in children. 1-71
1.9.3 Concentration-Response Relationships for Human Health Effects 1-72
1.9.4 Pb Exposure and Neurodevelopmental Deficits in Children 1-75
1.9.5 Reversibility and Persistence of Neurotoxic Effects of Pb 1-76
1.9.6 Populations Potentially At-Risk for Health Effects 1-78
Table 1-7 Summary of evidence for factors that potentially increase the risk of
Pb-related health effects. 1-79
1.9.7 Ecological Effects and Corresponding Pb Concentrations 1-82
1.10 Summary 1-82
Table 1-8 Summary of evidence from epidemiologic, animal toxicological and
ecological studies on the effects associated with exposure to Pb. 1-84
References for Chapter 1 1-95
CHAPTER 2 AMBIENT LEAD: SOURCE TO CONCENTRATION 2-1
2.1
2.2
2.
2.
Introduction
Sources of Atmospheric Pb
2.1 National Emissions Inventory
Figure 2-1 Trends in Pb emissions (thousand tons) from stationary and mobile
sources in the U.S., 1970-2008.
Figure 2-2 Trends in Pb emissions (thousand tons) from stationary and mobile
sources in the U.S., 1990-2008.
Figure 2-3 Nationwide stationary and mobile source Pb emissions (tons) in the
U.S. by source sector in 2008.
Figure 2-4 County-level Pb emissions (tons) in the U. S. in 2008.
Figure 2-5 Pb facilities estimated to emit 0. 5 tons or more in 2008.
2.2 Anthropogenic Sources
2-1
2-1
2-2
2-3
2-4
2-5
2-6
2-7
2-7
IV
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Table 2-1 Pb compounds observed in the environment. 2-8
Figure 2-6 Five-year totals for Pb mining, primary and secondary production,
imports, and exports, 1991-2010. 2-9
Figure 2-7 Total U. S. Pb additives in on-road gasoline used in on-road vehicles,
1927-1995. 2-22
Figure 2-8 Estimated Pb aerosol inputs from on-road gasoline into 90 U. S.
urbanized areas (UAs), from 1950 through 1982. 2-24
2.3 Fate and Transport of Pb 2-24
Figure 2-9 Fate of atmospheric Pb. 2-25
2.3.1 Air 2-25
Figure 2-10 Scales of turbulence within an urban environment. 2-32
2.3.2 Water 2-35
Table 2-2 Surface sediment Pb concentrations for various continental shelves. _ 2-39
2.3.3 Soil 2-49
Figure 2-11 Schematic model summarizing the estimated flux of Pb within a
typical podzol profile from northern Sweden using data from
Klaminder et al. (2006a) and Klaminder et al. (2006b). 2-57
Figure 2-12 Eh-pH diagram for Pb in shooting range soils, Jefferson National
Forest, VA. 2-60
2.4 Monitoring of Ambient Pb 2-62
2.4.1 Measurement Techniques 2-62
Figure 2-13 Comparison of particle collection efficiency among different
TSP sampler types. 2-63
Table 2-3 Airborne PM sampling methods potentially applicable for
Pb sampling. 2-66
2.4.2 Network Design 2-79
Figure 2-14 Map of monitoring sites in current Pb NAAQS monitoring network. 2-82
Figure 2-15 Fifteen U.S. locations where a study is currently being performed on
airport Pb emissions. 2-83
Table 2-4 Ust of 15 airports included in the airport study. 2-84
Figure 2-16 Pb-PM2.s monitoring sites for CSN and IMPROVE networks. 2-86
Figure 2-17 Pb-PMio monitoring sites for NATTS network. 2-87
2.5 Ambient Air Pb Concentrations 2-87
2.5.1 Spatial Distribution of Air Pb 2-88
Table 2-5 Summary data for source-oriented Pb monitors across the U. S.,
2008-2010. 2-90
Figure 2-18 Highest county-level source-oriented Pb-TSP concentrations
(ug/m3), maximum 3-month average, 2008-2010. 2-90
Table 2-6 Summary data for non-source-oriented Pb monitors across the U.S.,
2008-2010. 2-97
Figure 2-19 Highest county-level non-source-oriented Pb-TSP concentrations
(ug/m3), maximum 3-month average, 2008-2010. 2-91
Figure 2-20 Time series of monthly average Pb-TSP concentration at five near-
road monitors. 2-93
Table 2-7 Sample of U.S. near-road Pb-TSP monitors. 2-94
2.5.2 Temporal Variability 2-96
Figure 2-21 National trends in Pb concentration (ug/m3), 74 trends sites,
1990-2010. 2-97
Figure 2-22 Boxplots of average monthly Pb-PM2 5 concentrations measured at
four IMPROVE sites, 2001-2010. 2-98
2.5.3 Size Distribution of Pb-Bearing PM 2-99
Table 2-8 Summary of comparison data for co-located ambient air Pb
monitors. 2-101
Table 2-9 Summary of studies reporting Pb size distribution in the peer-
reviewed literature. 2-703
Figure 2-23 Comparison of urban background and near-road size fractions of Pb-
bearing PM. 2-709
2.5.4 Pb Concentrations in a Multipollutant Context 2-111
Figure 2-24 Pearson correlations of monitored non-source daily average Pb-TSP
concentration with daily averages of copollutant concentrations,
2008-2010. 2-112
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Figure 2-25 Pearson correlations of monitored Pb-PM2.5 concentration with
copollutant concentrations, 2008-2010.
Background Pb Concentrations
2.5.5
2.6 Ambient Pb Concentrations in Non-Air Media and Biota
2.6.1 Soils
Table 2-10
Figure 2-26
Figure 2-27
2.6.2 Sediments
So/7 concentrations in various cities, 1992-2005.
Map of median Pb content in soil in New Orleans, 1998-2000.
Size distribution of Pb-containing dust collected near busy (HWY 1)
and low traffic (HWY 17) highways.
Figure 2-28
Figure 2-29
WACAP data for Pb concentration in sediment at eight National
Parks and/or Preserves.
Sediment core data (1992-1994) for the lakes and reservoirs along
the Apalachicola, Chattahoochee, and Flint River Basin (ACF)._
2-113
2-115
2-117
2-118
2-119
2-122
2-124
2-125
2-727
2-728
2.6.3
2.6.4
Figure 2-30
Rain
Figure 2-31
Snowpack
Sediment core data (1975-1995) for the lakes and reservoirs along
the Apalachicola, Chattahoochee, and Flint River Basin (ACF).
Trends in Pb concentration in precipitation from various sites in
Norway over the period 1980-2005.
2-129
2-130
2-737
2-132
Figure 2-32 Box plots illustrating Pb concentration in snow melt at nine National
Parks and Preserves.
2.6.5 Natural Waters
Figure 2-33 Boxplots of Pb concentration in surface waters measured at five
National Parks and Preserves.
Table 2-11 Pb concentrations from stream food-webs; in mining-disturbed areas
of Missouri and the western U.S.
2.6.6 Vegetation
Figure 2-34
Boxplots of Pb concentration in lichen measured at seven National
Parks and Preserves.
2.6.7
2.6.8
Figure 2-35 Trends in regional pollution near a copper (Cu) smelter in Canada
and Pb concentrations at the boundary ofheartwood trees within
roughly 75 km of the smelter.
Aquatic Bivalves_
Vertebrate Populations
Figure 2-36 Boxplots of Pb concentration in fish fillet and fish liver, measured at
eight National Parks and/or Preserves._
Figure 2-37 Boxplots of Pb concentration in moose meat and moose liver
measured at Denali National Park and Preserve (Alaska).
2.7 Summary and Conclusions
2.7.1 Sources of Atmospheric Pb
2.7.2
2.7.3
2.7.4
2.7.5
Fate and Transport of Pb_
Ambient Pb Monitoring
Ambient Air Pb Concentrations
Ambient Pb Concentrations in Non-Air Media and Biota
2.8 Chapter 2 Appendix (Supplemental Material)
2.8.1 Variability across the U.S.
Table 2-12 Distribution of 1-month average Pb-TSP concentrations (ug/m )
nationwide, source-oriented monitors, 2008-2010.
Distribution of 1-month average Pb-TSP concentrations (ug/m )
nationwide, non-source-oriented monitors, 2008-2010.
Table 2-13
Table 2-14 Distribution of 3-month moving average Pb-TSP concentrations
(pg/m3) nationwide, source-oriented monitors, 2008-2010.
Table 2-15 Distribution of 3-month moving average Pb-TSP concentrations
(ug/m3) nationwide, non-source-oriented monitors, 2008-2010.
Table 2-16 Distribution of annual 1-month site maxima TSP Pb concentrations
(ug/m3) nationwide, source-oriented monitors, 2008-2010.
Table 2-17 Distribution of annual 1-month site maxima TSP Pb concentrations
(ug/m3) nationwide, non-source-oriented monitors, 2008-2010.
Table 2-18 Distribution of annual 3-month site maxima Pb-TSP concentrations
(ug/m3) nationwide, source-oriented monitors, 2008-2010.
2-734
2-135
2-737
2-738
2-138
2-739
2-747
2-142
2-142
2-743
2-744
2-745
2-145
2-145
2-146
2-147
2-148
2-749
2-149
2-749
2-753
2-757
2-767
2-764
2-765
2-766
VI
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2.8.2
Table 2-19 Distribution of annual 3-month site maxima Pb-TSP concentrations
(ug/m3) nationwide, non-source-oriented monitors, 2008-2010.
Figure 2-38 Highest county-level Pb-PMio concentrations (ug/m3), maximum
3-month average, 2007-2009.
Figure 2-39 Highest county-level Pb-PM2.5 concentrations (ug/m3), maximum
3-month average, 2007-2009.
Intra-urban Variability
Figure 2-40 Pb TSP monitor and source locations within Los Angeles County,
CA (06-037), 2007-2009.
Figure 2-41 Wind roses for Los Angeles County, CA, from meteorological data at
the Los Angeles International Airport, 1961-1990.
Figure 2-42 Box plots of annual and seasonal 24-h Pb TSP concentrations
(ug/m3) from source-oriented and non-source-oriented monitors
within Los Angeles County, CA (06-037), 2007-2009.
Table 2-20
Comparisons between Pb TSP concentrations from source-oriented
and non-source-oriented monitors within Los Angeles County, CA
(06-037), 2007-2009.
Figure 2-43 Pb TSP monitor locations within Hillsborough and Pinellas Counties,
FL (12-057 and 12-103), 2007-2009.
Figure 2-44 Wind roses for Hillsborough/Pinellas Counties, FL, obtained from
meteorological data at Tampa International Airport, 1961-1990.
Figure 2-45 Box plots of annual and seasonal 24-h Pb TSP concentrations
(ug/m3) from source-oriented and non-source-oriented monitors
within Hillsborough and Pinellas Counties, FL (12-057 and 12-103),
2007-2009.
Table 2-21
Correlations between Pb TSP concentrations from source-oriented
and non-source-oriented monitors within Hillsborough and Pinellas
Counties, FL (12-057 and 12-103), 2007-2009.
Figure 2-46 Pb TSP Monitor locations within Cook County, IL (17-031),
2007-2009.
Figure 2-47 Wind roses for Cook County, IL, obtained from meteorological data
at O'Hare International Airport, 1961-1990.
Figure 2-48 Box plots of annual and seasonal 24-h Pb TSP concentrations
(ug/m3) from source-oriented and non-source-oriented monitors
within Cook County, IL (17-031), 2007-2009..
Table 2-22
Correlations between Pb TSP concentrations from source-oriented
and non-source-oriented monitors within Cook County, IL (17-031),
2007-2009.
Figure 2-49 Pb TSP Monitor locations within Jefferson County, MO (29-099),
2007-2009.
Figure 2-50 Wind roses for Jefferson County, MO, obtained from meteorological
data at St. Louis/Lambert International Airport, 1961-1990.
Figure 2-51
Table 2-23
Box plots of annual and seasonal 24-h Pb TSP concentrations
(ug/m3) from source-oriented and non-source-oriented monitors
within Jefferson County, MO (29-099), 2007-2009.
Correlations between Pb TSP concentrations from source-oriented
and non-source-oriented monitors within Jefferson County, MO
(29-099), 2007-2009.
Figure 2-52 Pb TSP Monitor locations within Cuyahoga County, OH (39-035),
2007-2009.
Figure 2-53 Wind roses for Cuyahoga County, OH, obtained from meteorological
data at Cleveland/Hopkins International Airport, 1961-1990.
Figure 2-54 Box plots of annual and seasonal 24-h Pb TSP concentrations
(ug/m3) from source-oriented and non-source-oriented monitors
within Cuyahoga County, OH (39-035), 2007-2009.
Table 2-24
Correlations between Pb TSP concentrations from source-oriented
and non-source-oriented monitors within Cuyahoga County, OH
(39-035), 2007-2009.
Figure 2-55 Pb TSP Monitor locations within Sullivan County, TN (47-163),
2007-2009.
Figure 2-56 Wind roses for Sullivan County, TN, obtained from meteorological
data at Bristol/Tri City Airport, 1961-1990.
2-767
2-768
2-769
2-170
2-773
2-774
2-775
2-776
2-779
2-780
2-787
2-782
2-785
2-786
2-787
2-788
2-790
2-797
2-792
2-793
2-796
2-797
2-798
2-799
2-207
2-202
VII
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Figure 2-57 Box plots of annual and seasonal 24-h Pb TSP concentrations
(ug/m3) from source-oriented monitors within Sullivan County, TN
(47-163), 2007-2009. 2-203
Table 2-25 Correlations between Pb TSP concentrations from source-oriented
monitors within Sullivan County, TN (47-163), 2007-2009. 2-204
2.8.3 Seasonal Variation in Pb Concentrations 2-204
Figure 2-58 Monthly source-oriented Pb-TSP average (ug/m ) over 12 months of
the year, 2008-2010. 2-205
Figure 2-59 Monthly non-source-oriented Pb-TSP average (ug/m ) over 12
months of the year, 2008-2010 2-206
Figure 2-60 Monthly Pb-PM 10 average (ug/m ) over 12 months of the year,
2007-2009. 2-206
Figure 2-61 Monthly Pb-PM 2 5 average (ug/m3) over 12 months of the year,
2007-2009. 2-207
2.8.4 Size Distribution of Pb-bearing PM 2-207
Table 2-26 Correlations and average of the concentration ratios for co-located
monitors, TSP versus PMW, TSP versus PM25, and PMW versus
PM2.5. 2-208
Table 2-27 Metadata for studies of Pb-PM size distribution. 2-213
Table 2-28 Size distribution data for various studies described in Table 2-27. 2-218
2.8.5 Pb Concentration in a Multipollutant Context 2-226
Figure 2-62 Spearman correlations of monitored non-source Pb-TSP
concentration with daily averages of copollutant concentrations,
2008-2010. 2-226
Figure 2-63 Seasonal correlations of monitored Pb-TSP concentration with
copollutant concentrations, 2007-2008. 2-227
Figure 2-64 Seasonal correlations of monitored Pb-TSP concentration with
copollutant concentrations, 2009. 2-228
Figure 2-65 Seasonal correlations of monitored Pb-TSP concentration with
copollutant concentrations, 2009. 2-229
Figure 2-66 Seasonal correlations of monitored Pb-PM2.5 concentration with
copollutant concentrations, 2007-2009. 2-230
Figure 2-67 Seasonal correlations of monitored Pb-PM2.5 concentration with
copollutant concentrations, 2007-2009. 2-231
Table 2-29 Copollutant exposures for various trace metal studies. 2-232
References for Chapter 2 2-234
CHAPTER 3 EXPOSURE, TOXICOKINETICS, AND BIOMARKERS 3-1
3.1 Exposure Assessment 3-1
3.1.1 Pathways for Pb Exposure 3-1
Figure 3-1 Conceptual model of multimedia Pb exposure. 3-3
Table 3-1 Estimates of Pb measurements for EPA Region 5 from the
NHEXAS study. 3-4
Table 3-2 Predicted concurrent blood Pb levels and source contributions for
children in their seventh year of life. 3-9
3.1.2 Environmental Exposure Assessment Methodologies 3-9
3.1.3 Exposure Studies 3-11
Table 3-3 Estimates of fixed effects multivariate modeling of Pb levels
measured during the NHEXAS-MD study. 3-12
Table 3-4 Comparison of personal, indoor, and outdoor Pb-PM measurements
from several studies. 3-15
Table 3-5 Measurements of indoor dust Pb concentration from 2006-2011
studies. 3-16
Figure 3-2 Market basket survey results for Pb concentration in foods. 3-27
Table 3-6 Pb bioaccumulation data for various plants. Bioaccumulation is
expressed as percent of Pb concentration in the plant to the Pb
concentration in the soil. 3-27
Table 3-7 Pb content in various consumer products. 3-37
3.2 Kinetics 3-33
3.2.1 Absorption 3-33
Vlll
-------
Figure 3-3 Estimated relative bioavailability (RBA, compared to Pb acetate) of
ingested Pb in mineral groups. 3-41
3.2.2 Distribution and Metabolism 3-43
Figure 3-4 Plot of blood and plasma Pb concentrations measured in adults and
children. 3-45
Figure 3-5 Relationship between Pb intake and blood Pb concentration in
infants (N = 105, age 13 weeks, formula-fed). 3-46
Figure 3-6 Simulation of quasi-steady state blood and plasma Pb
concentrations in a child (age 4 years) associated with varying Pb
ingestion rates. 3-47
3.2.3 Elimination 3-51
3.3 Pb Biomarkers 3-53
3.3.1 Bone-Pb Measurements 3-56
3.3.2 Blood-Pb Measurements 3-58
Figure 3-7 Simulation of temporal relationships between Pb exposure and blood
Pb concentration in children. 3-61
3.3.3 Urine-Pb Measurements 3-62
Figure 3-8 Simulation of relationship between urinary Pb excretion and body
burden in adults. 3-63
3.3.4 Pb in Other Biomarkers 3-64
3.3.5 Relationship between Pb in Blood and Pb in Bone 3-66
Figure 3-9 Simulation of relationship between blood Pb concentration and body
burden in children, with an elevated constant Pb intake from age
2 to 5 years. 3-71
Figure 3-10 Simulation of relationship between time-integrated blood Pb
concentration and cumulative Pb absorption in children. 3-72
Figure 3-11 Simulation of relationship between blood Pb concentration, bone Pb
and body burden in adults with relatively low Pb intake. 3-75
Figure 3-12 Simulation of relationship between blood Pb concentration, bone Pb
and body burden in adults with relatively high Pb intake. 3-76
3.3.6 Relationship between Pb in Blood and Pb in Soft Tissues 3-80
Figure 3-13 Simulation of blood and soft tissue (including brain) Pb in children
and adults who experience a period of increased Pb intake. 3-81
Figure 3-14 Simulation of blood and brain Pb in children and adults who
experience a period of increased Pb intake. 3-82
Figure 3-15 Relationship between Pb in urine, plasma, blood and bone. 3-84
3.4 Studies ofPb Biomarker Levels 3-85
3.4.1 Pb in Blood 3-85
Figure 3-16 Temporal trend in blood Pb concentration. 3-86
Table 3-8 Blood Pb concentrations in the U.S. population. 3-87
Figure 3-17 Box plots of blood Pb levels among U. S. children (1-5 years old)
from the NHANES survey, 1988-2010. 3-89
Figure 3-18 Blood Pb cohort means versus year of exam. 3-90
Figure 3-19 Percent distribution of blood Pb levels by race/ethnicity among U.S.
children (1-5 years) from the NHANES survey, 1988-1991 (top) and
1999-2004 (bottom). 3-97
Figure 3-20 Trends in 206Pbf°4Pb isotope ratio in blood Pb (a) and trends in
blood Pb levels (b) among Australian study populations of children
during the period 1990-2000. 3-95
3.4.2 Pb in Bone 3-96
Table 3-9 Epidemiologic studies that provide bone Pb measurements for non-
occupationally exposed populations. 3-96
Table 3-10 Epidemiologic studies that provide bone Pb measurements for
occupationally exposed populations. 3-709
3.4.3 Pb in Urine 3-115
Table 3-11 Urine Pb concentrations in the U.S. population. 3-775
3.4.4 Pbin Teeth 3-117
Figure 3-27 Comparison of relative temporal changes in tooth enamel Pb
concentration. 3-777
3.5 Empirical Models of Pb Exposure-Blood Pb Relationships 3-778
3.5.1 Air Pb-Blood Pb Relationships in Children 3-120
IX
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Table 3-12 Summary of estimated slopes for blood Pb to air Pb slope factors
3.5.2
3.5.3
in humans.
Figure 3-22 Predicted relationship between air Pb and blood Pb based on a
meta-analysis of 18 studies.
Figure 3-23 Blood Pb - air Pb slopes (ug/dL perug/mj) predicted from
epidemioloqic studies.
Table 3-13 Environmental Pb levels and blood Pb levels in children in Trail,
British Columbia.
Table 3-14 Predicted blood Pb levels and blood-air slopes for Mexico City
children (1987 and 1990 cohorts).
Table 3-15 U.S. gasoline Pb consumption and air Pb levels.
Table 3-16 Air Pb concentrations and blood Pb levels in children in Mumbai,
India.
Figure 3-24 Predicted relationship between air Pb and blood Pb based on data
from Chicago, IL in children age 0-5 years (1974-1988).
Figure 3-25 Effect of air-to-blood slope on estimated change in air-related blood
Pb with change in air Pb.
Air Pb-Blood Pb Relationships in Occupational Cohorts
Environmental Pb-Blood Pb Relationships
Table 3-1 7 Linear model relating environmental Pb exposure and blood Pb
3-121
3-123
3-124
3-125
3-128
3-130
3-131
3-132
3-133
3-134
3-135
concentration in children.
Figure 3-26 Predicted relationship between soil Pb concentration and blood Pb
concentration in children based on data collected in New Orleans,
Louisiana: 2000-2005.
Table 3-18
3.6
3.7
General linear model relating blood Pb concentration in children and
environmental Pb levels—Bunker Hill Superfund Site.
Biokinetic Models of Pb Exposure-Blood Pb Relationships
Summary and Conclusions
3.7.1 Exposure
3.7.2 Toxicokinetics
3.7.3 Pb Biomarkers
3.7.4
Air Lead-Blood Lead Relationships
References for Chapter 3
CHAPTER 4
INTEGRATED HEALTH EFFECTS OF LEAD EXPOSURE
3-737
3-739
3-747
3-747
3-743
3-143
3-144
3-145
3-147
3-748
4-1
4. 7 Introduction
4.2 Modes of Action
4.2.1
4.2.2
Introduction
Figure 4-1
Altered Ion Status
Schematic representation of the relationships between the various
MOAs by which Pb exposure exerts its health effects.
4-1
4-3
4-3
4-4
4-4
Table 4-1 Enzymes and proteins potentially affected by exposure to Pb and the
metal cation cofactors necessary for their proper physiological
4.2.3
4.2.4
4.2.5
4.2.6
4.2.7
4.2.8
4.3 Ner
4.3.1
4.3.2
activity.
Protein Binding
Oxidative Stress
Inflammation
Endocrine Disruption
Cell Death and Genotoxicity
Summary
Table 4-2 MOAs, their related health effects, and information on concentrations
eliciting the MOAs.
vous System Effects
Introduction
Cognitive Function
Figure 4-2 Associations of blood Pb levels with full-scale IQ (FSIQ) in children.
Table 4-3 Additional characteristics and quantitative results for studies
represented in Figure 4-2
4-16
4-20
4-25
4-32
4-36
4-39
4-51
4-52
4-55
4-55
4-59
4-67
4-62
Table 4-4 Associations of blood Pb level with Bayley MDI in children ages 1 to
3 years.
4-78
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Table 4-5 Associations between blood Pb levels and performance on tests of
learning and memory in children. 4-86
Figure 4-3 Summary of Pb exposure-nervous system concentration-response
information from animal toxicological studies. 4-93
Table 4-6 Summary of findings from neurotoxicological concentration-response
array presented in Figure 4-3. 4-94
Table 4-7 Summary of effects of maternal and lifetime Pb exposure on
Fl performance observed by Cory-Slechta laboratory. 4-102
Figure 4-4 Changes in Fixed Interval performance in (A) female and (B) male
offspring with gestational/lactational Pb exposure plus various
stressors given in adulthood. 4-103
Figure 4-5 Mean basal and final corticosterone levels of female and male
offspring exposed to lifetime Pb. 4-105
Table 4-8 Associations between blood or tooth Pb levels and performance on
tests of executive function in children and young adults. 4-109
Table 4-9 Associations between blood or tooth Pb levels and measures of
academic performance and achievement in children and young
adults. 4-114
Figure 4-6 Greater reduction in End-of-Grade (EOG) scores for an increase in
blood Pb level in lower percentiles of the test score distribution. 4-120
Table 4-10 Associations of blood and bone Pb levels with cognitive function in
adults. 4-131
Figure 4-7 Nonlinear association between patella Pb level and the relative
change over 3.5 years in response latency on the pattern
comparison test in men from the Normative Aging Study. 4-141
Figure 4-8 Nonlinear association of tibia Pb level with annual rate of cognitive
decline, by hemochromatosis genotype in men from the Normative
Aging Study. 4-142
4.3.3 Externalizing Behaviors in Children 4-150
Figure 4-9 Associations of blood Pb levels with attention, impulsivity, and
hyperactivity in children. 4-155
Table 4-11 Additional characteristics and quantitative results for studies
presented in Figure 4-9. 4-156
Table 4-12 Associations between Pb biomarker levels and behaviors related to
conduct disorders in children and young adults. 4-178
4.3.4 Internalizing Behaviors in Children 4-194
4.3.5 Psychopathological Effects in Adults 4-200
Figure 4-10 Hypothetical representation of the contribution of Pb exposure to the
development of a phenotype consistent with schizophrenia. 4-204
4.3.6 Sensory Organ Function 4-205
Table 4-13 Summary of Pb-related effects observed on the visual system. 4-213
Figure 4-11 Retinal a-wave and b-wave electroretinogram (ERG) amplitude in
adult rats after prenatal plus early postnatal Pb exposure. 4-214
Figure 4-12 Retinal dopamine metabolism in adult control and gestationally Pb-
exposed (GLE) rats. 4-216
4.3.7 Motor Function 4-217
4.3.8 Seizures in Animals 4-219
4.3.9 Neurodegenerative Diseases 4-220
4.3.10 Modes of Action for Pb Nervous System Effects 4-229
Figure 4-13 Neurogenesis (production of new cells) in the rat hippocampal
dentate gyrus after early postnatal Pb exposure. 4-244
4.3.11 Lifestages and Time Periods of Pb Exposure and Neurodevelopmental Effects 4-247
Table 4-14 Associations of cognitive function with blood Pb levels measured at
various lifestages and time periods in prospective studies. 4-250
Table 4-15 Comparisons of blood Pb-cognitive function associations in groups
of children with different temporal trends in blood Pb levels. 4-258
Figure 4-14 Estimated FSIQ for three different temporal patterns in blood Pb
level from ages 2 to 6 years in Rochester and Cincinnati cohorts. 4-260
4.3.12 Examination of the Pb Concentration-Response Relationship 4-263
Figure 4-15 Comparison of associations between blood Pb level and cognitive
function among various blood Pb strata. 4-264
XI
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Table 4-16
Additional characteristics and quantitative results for studies
presented in Figure 4-15.
4.3.13 Confounding in Epidemiologic Studies of Nervous System Effects_
4-265
4-274
4.3.14 Public Health Significance of Associations between Pb Biomarkers and Neurodevelopmental
Effects
Figure 4-16 Hypothetical effect of increasing blood Pb level on the proportion of
the population with FSIQ <80 and <70 points.
4.3.15 Summary and Causal Determination
Table 4-17 Summary of Evidence Supporting Nervous System Causal
Determinations.
4.4 Cardiovascular Effects
4.4.1 Introduction
4.4.2 Blood Pressure and Hypertension
Figure 4-17 Associations of blood and bone Pb levels with systolic BP, diastolic
BP, and pulse pressure in adults.
Table 4-18 Additional characteristics and quantitative data for associations of
blood and bone Pb with BP measures for studies presented in
Figure 4-17.
Figure 4-18 Odds ratios (95% Cl) for associations of blood Pb and bone Pb with
hypertension prevalence and incidence.
Table 4-19 Additional characteristics and quantitative data for results presented
in Figure 4-18 for associations of blood and bone Pb with
hypertension measures.
Figure 4-19 The relationship between tibia Pb and estimated systolic BP (SBP)
for those with high self-reported stress versus those with low self-
reported stress.
Figure 4-20 Changes in BP after Pb exposure (represented as blood Pb level) in
unanesthetized adult rats across studies.
Table 4-20 Characteristics of studies of blood Pb with BP measures in animals
presented in Figure 4-20.
Figure 4-21 Meta-analysis of change in systolic BP (SBP), in mmHg, with 95%
Cl, associated with a doubling in the blood Pb concentration.
4.4.3
Figure 4-22 Meta-analysis of an increase in systolic BP (SBP) and diastolic BP
(DBF) and relative risk of hypertension per 10ug/g increase in bone
Pb levels.
Vascular Effects and Cardiotoxicity
Table 4-21 Characteristics and quantitative data for associations of blood and
bone Pb with other CVD measures, i.e., HRV, PAD, and IHD in
recent epidemiologic studies.
4.4.4 Cardiovascular Function and Blood Pressure in Children
Table 4-22 Studies of child cardiovascular endpoints and Pb biomarkers._
Figure 4-23 Children's adjusted total peripheral resistance (dyn-s/cm ) responses
4.4.5 Mortality
to acute stress tasks, as a function of early childhood Pb levels. _
Figure 4-24 Multivariate adjusted relative hazards of all-cause and
cardiovascular mortality per 3.4 pg/dL increase in blood Pb.
Figure 4-25 Multivariate-adjusted relative hazard (ten axis) of mortality
associated with blood Pb levels between 1 ug/dL and 10 ug/dL.
Figure 4-26 Relative risk of all cause mortality for different blood Pb levels
compared with referent level of 1.5 ug/dL (12.5th percentile).
4-279
4-287
4-282
4-311
4-324
4-324
4-326
4-329
4-330
4-336
4-337
4-345
4-348
4-349
4-367
4-363
4-364
4-377
4-378
4-380
4-385
4-389
4-397
4-392
4-393
Figure 4-27 Associations between patella bone Pb level and the log of hazard
ratio (logHR) for all-cause, cardiovascular, and ischemic heart
disease.
Figure 4-28 Multivariate adjusted relative hazard (ten axis) of mortality as a
function of blood Pb levels between 1 ug/dL and 15 ug/dL.
Figure 4-29 Hazard ratios for associations of blood Pb or bone Pb with all-cause
mortality and cardiovascular mortality.
Table 4-23
4.4.6 Air Pb-PM Studies
4.4.7 Summary and Causal Determination,
Additional characteristics and quantitative data for associations of
blood and bone Pb with mortality for studies presented in
Figure 4-29.
4-395
4-397
4-399
4-400
4-405
4-406
XII
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Table 4-24 Summary of evidence supporting cardiovascular causal
determinations. 4-415
4.5 Renal Effects 4-427
4.5.1 Introduction 4-421
4.5.2 Nephrotoxicity and Renal Pathology 4-424
Figure 4-30 Concentration-response relationships for associations between
blood Pb level and kidney function outcomes in adults. 4-429
Figure 4-31 Percent change in kidney outcomes across quartiles of blood Pb
level in NHANES. 4-430
Table 4-25 Additional characteristics and quantitative data for associations of
blood and bone Pb with kidney outcomes for results presented in
Figure 4-30 and Figure 4-31. 4-432
Table 4-26 Prospective patient population studies: kidney function decline. 4-438
Table 4-27 Clinical randomized chelation trials in chronic kidney disease
patients. 4-442
Figure 4-32 Added variable plot of association between serum creatinine and
blood Pb in 267 Korean Pb workers in the oldest age fertile. 4-446
Table 4-28 Animal toxicological studies reporting the effects of Pb exposure
(as blood Pb level) on kidney function. 4-453
Table 4-29 Indicators of renal damage in male rats exposed to 50 ppm Pb for 40
and 65 days, starting at parturition. 4-455
Table 4-30 Effects of Pb on the kidney/renal system related to exposure
duration: Evidence from animal (rodent) toxicology studies. 4-460
4.5.3 Modes of Action for Pb-lnduced Nephrotoxicity 4-461
4.5.4 Effects of Exposure to Pb Mixtures 4-472
4.5.5 Summary and Causal Determination 4-477
Table 4-31 Summary of evidence supporting renal causal determinations. 4-482
4.6 Immune System Effects 4-485
4.6.1 Introduction 4-485
Figure 4-33 Immunological pathways by which Pb exposure potentially may
increase risk of immune-related diseases. 4-486
4.6.2 Cell-Mediated Immunity 4-488
4.6.3 Humoral Immunity 4-501
Table 4-32 Comparison of serum immunoglobulin levels and B cell abundance
among various blood Pb groups. 4-503
4.6.4 Inflammation 4-513
4.6.5 Immune-based Diseases 4-516
Figure 4-34 Associations of blood Pb levels with asthma and allergy in children. _ 4-519
Table 4-33 Additional characteristics and quantitative results for studies
presented in Figure 4-34. 4-520
4.6.6 Modes of Action for Pb Immune Effects 4-531
4.6.7 Immune Effects of Pb within Mixtures 4-541
4.6.8 Summary and Causal Determination 4-542
Table 4-34 Summary of evidence supporting immune causal determinations. 4-550
4. 7 Hematoloqical Effects
4.7.1 Introduction
4.7.2 Red Blood Cell Function
4.7.3 Red Blood Cell Heme Metabolism
4-555
4-555
4-557
4-573
Figure 4-35 Schematic representation of the enzymatic steps involved in the
heme synthetic pathway. 4-573
4.7.4 Summary and Causal Determination 4-577
Table 4-35 Summary of evidence supporting RBC survival and heme synthesis
causal determinations. 4-584
4.8 Reproductive and Developmental Effects 4-589
4.8.1 Effects on Development 4-589
Table 4-36 Summary of epidemiologic studies of associations between blood Pb
levels and puberty for females. 4-597
Figure 4-36 Toxicological concentration-response array for reproductive and
developmental effects of Pb. 4-607
Table 4-37 Toxicological concentration-response array summary for
reproductive and developmental effects of Pb presented in
Figure 4-36. 4-602
Xlll
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Table 4-38 Summary of recent epidemic/logic studies of associations between
blood Pb levels and puberty for males. 4-606
Table 4-39 Summary of recent epidemiologic studies of associations between
Pb biomarker levels and postnatal growth. 4-673
4.8.2 Effects on Birth Outcomes 4-627
Table 4-40 Summary of recent epidemiologic studies of associations between
Pb exposure indicators and neural tube defects. 4-629
Table 4-41 Summary of recent epidemiologic studies of associations between
Pb exposure indicators and preterm birth. 4-637
Table 4-42 Summary of recent epidemiologic studies of associations between
Pb exposure indicators and low birth weight and fetal growth. 4-640
4.8.3 Effects on Male Reproductive Function 4-655
Table 4-43 Summary of recent epidemiologic studies of associations between
Pb biomarker levels and effects on sperm and semen. 4-657
Table 4-44 Summary of recent epidemiologic studies of associations between
Pb exposure indicators and hormones for males. 4-670
4.8.4 Effects on Female Reproductive Function 4-680
Table 4-45 Summary of recent epidemiologic studies of associations between
blood Pb levels and hormones for females. 4-682
Table 4-46 Summary of recent epidemiologic studies of associations between
Pb biomarker levels and fertility for females. 4-690
Table 4-47 Summary of recent epidemiologic studies of associations between
Pb biomarker levels and spontaneous abortions. 4-696
4.8.5 Summary and Causal Determination 4-703
Table 4-48 Summary of evidence supporting reproductive and developmental
causal determinations.
4.9 Effects on Other Organ Systems
4.9.1 Effects on the Hepatic System
4.9.2 Effects on the Gastrointestinal System
4.9.3 Effects on the Endocrine System
4.9.4 Effects on Bone and Teeth
4.9.5 Effects on Ocular Health
4.9.6 Effects on the Respiratory System
4.70 Cancer
4-708
4-713
4-713
4-722
4-725
4-730
4-736
4-738
4-739
4.10.1 Cancer Incidence and Mortality 4-740
Table 4-49 Summary of recent epidemiologic studies of cancer incidence
and overall cancer mortality. 4-747
4.10.2 Cancer Biomarkers 4-765
4.10.3 Modes of Action for Pb-induced Carcinogenicity 4-765
4.10.4 Effects of Pb within Mixtures 4-777
4.10.5 Summary and Causal Determination 4-779
Table 4-50 Summary of evidence supporting cancer and genotoxicity
causal determinations. 4-782
References for Chapter 4 4-784
CHAPTER 5 POTENTIALLY AT-RISK POPULATIONS 5-1
Introduction
A
5.7
5.2
5.
5.
5.
pproach to Classifying Potential At-Risk Factors
Table 5-1 Classification of evidence for potential at-risk factors.
Physiological Factors that Influence the Internal Distribution of Pb
Population Characteristics Potentially Related to Differential Pb Exposure
2.1 Age
Table 5-2 Blood Pb levels (uQ/dL) by age and sex, 2009-201 0 NHANES.
Table 5-3 Percentage of children within six categories/brackets of blood Pb
levels, 1999-2004 NHANES.
2.2 Sex
2.3 Race and Ethnicity
5-1
5-2
5-4
5-4
5-6
5-6
5-8
5-9
5-11
5-11
Figure 5-1 Percent distribution of blood Pb levels by race/ethnicity among U. S.
children (1-5 years). 5-73
Figure 5-2 So/7 Pb concentration exposure among the population of three
parishes within greater metropolitan New Orleans. 5-74
xiv
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5.2.4 Socioeconomic Status (SES) 5-15
5.2.5 Proximity to Pb Sources 5-16
5.2.6 Residential Factors 5-17
Table 5-4 Regression of log-transformed blood Pb level of children 12-60
months old on various factors related to housing condition, from
1999-2004 NHANES dataset. 5-18
5.3 Factors Potentially Related to Increased Risk of Pb-lnduced Health Effects 5-20
5.3.1 Age 5-20
5.3.2 Sex 5-23
5.3.3 Genetics 5-26
5.3.4 Pre-existing Diseases/Conditions 5-31
5.3.5 Smoking Status 5-33
5.3.6 Socioeconomic Status 5-34
5.3.7 Race/Ethnicity 5-34
5.3.8 Body Mass Index 5-36
5.3.9 Alcohol Consumption 5-36
5.3.10 Nutritional Factors 5-36
5.3.11 Stress 5-38
5.3.12 Maternal Self-Esteem 5-39
5.3.13 Cognitive Reserve 5-39
5.3.14 Other Metal Exposure 5-39
5.4 Summary 5-40
Table 5-5 Summary of evidence for factors that potentially increase the risk of
Pb-related health effects. 5-41
References for Chapter 5 5-45
CHAPTER 6 ECOLOGICAL EFFECTS OF LEAD 6-57
6.1 Introduction to Ecological Concepts 6-57
6.1.1 Ecosystem Scale, Function, and Structure 6-58
6.1.2 Ecosystem Services 6-60
6.1.3 Critical Loads as an Organizing Principle for Ecological Effects of Atmospheric Deposition 6-61
6.2 Fate and Transport of Pb in Ecosystems 6-62
Figure 6-1 Fate of atmospheric Pb in ecosystems. 6-63
6.2.1 Fate and Transport 6-63
6.2.2 Ecosystem Exposure, Lag Time and Re-entrainment of Historically Deposited Pb 6-66
Table 6-1 Comparison among several metals: Time to achieve 95% of steady
state metal concentration in soil; example in a temperate system. 6-67
6.2.3 Concentrations in Non-Air Media 6-68
Table 6-2 Pb concentrations in non-air media and biota considered for
ecological assessment. 6-70
6.3 Terrestrial Ecosystem Effects 6-77
6.3.1 Introduction to Effects of Pb on Terrestrial Ecosystems 6-71
6.3.2 Soil Biogeochemistry and its Influence on Bioavailability 6-73
6.3.3 Bioavailability in Terrestrial Systems 6-77
Figure 6-2 Conceptual diagram for evaluating bioavailability processes and
bioaccessibility for metals in soil, sediment, or aquatic systems. 6-79
Figure 6-3 Schematic diagram of the biotic ligand model. 6-80
6.3.4 Biological Effects of Pb in Terrestrial Systems 6-98
6.3.5 Exposure and Response of Terrestrial Species 6-113
6.3.6 Terrestrial Community and Ecosystem Effects 6-116
6.3.7 Critical Loads in Terrestrial Systems 6-122
6.3.8 Soil Screening Levels 6-124
6.3.9 Characterization of Sensitivity and Vulnerability 6-125
6.3.10 Ecosystem Services Associated with Terrestrial Systems 6-127
6.3.11 Synthesis of New Evidence for Pb Effects in Terrestrial Systems 6-129
6.3.12 Causal Determinations for Pb in Terrestrial Systems 6-134
6.4 Aquatic Ecosystem Effects 6-745
6.4.1 Introduction to Effects of Pb on Aquatic Ecosystems 6-145
6.4.2 Biogeochemistry and Chemical Effects of Pb in Freshwater and Saltwater Systems 6-147
6.4.3 Introduction to Bioavailability and Biological Effects of Pb in Freshwater Ecosystems 6-153
6.4.4 Bioavailability in Freshwater Systems 6-156
xv
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6.4.5 Biological Effects of Pb in Freshwater Systems 6-177
6.4.6 Exposure and Response of Freshwater Species 6-198
6.4.7 Freshwater Community and Ecosystem Effects 6-203
6.4.8 Critical Loads in Freshwater Systems 6-207
6.4.9 Characterization of Sensitivity and Vulnerability in Freshwater Systems 6-208
6.4.10 Ecosystem Services Associated with Freshwater Systems 6-212
6.4.11 Synthesis of New Evidence for Pb Effects in Freshwater Ecosystems 6-213
6.4.12 Causal Determinations for Pb in Freshwater Systems 6-216
6.4.13 Introduction to Bioavailability and Biological Effects of Pb in Saltwater Ecosystems 6-231
6.4.14 Bioavailability of Pb in Saltwater Systems 6-233
6.4.15 Biological Effects of Pb in Saltwater Systems 6-240
6.4.16 Exposure and Response of Saltwater Species 6-245
6.4.17 Community and Ecosystem Effects in Saltwater Systems 6-247
6.4.18 Characterization of Sensitivity and Vulnerability in Saltwater Species 6-249
6.4.19 Ecosystem Services Associated with Saltwater Systems 6-251
6.4.20 Synthesis of New Evidence for Pb Effects in Saltwater Systems 6-252
6.4.21 Causal Determinations for Pb in Saltwater Systems 6-255
6.5 Causal Determinations for Ecological Effects of Pb 6-262
Table 6-3 Summary of Pb causal determinations for plants, invertebrates and
vertebrates. 6-265
6.6 Supplemental Material 6-266
Table 6-4 Recent evidence for Pb effects on terrestrial plants, invertebrates
and vertebrates; growth, reproduction and survival. 6-266
Figure 6-4 Subset of concentration-response data reported in Table 6-4 for
Pb effects on growth, reproduction, and survival in some terrestrial
invertebrates. 6-281
Table 6-5 Recent evidence for Pb effects on freshwater plants, invertebrates
and vertebrates; growth, reproduction, and survival. 6-282
Figure 6-5 Subset of concentration-response data reported in Table 6-5 for Pb
effects on growth, reproduction, and survival in some freshwater
invertebrates and fish. 6-308
Table 6-6 Recent evidence for Pb effects on saltwater plants, invertebrates,
and vertebrates: growth, reproduction, and survival. 6-309
References for Chapter 6 6-37 7
xvi
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Lead Project Team
Executive Direction
Dr. John Vandenberg (Director)—National Center for Environmental Assessment-RTF
Division, Office of Research and Development, U.S. Environmental Protection
Agency, Research Triangle Park, NC
Ms. Debra Walsh (Deputy Director)—National Center for Environmental Assessment-
RTF Division, Office of Research and Development, U.S. Environmental Protection
Agency, Research Triangle Park, NC
Dr. Mary Ross (Branch Chief)—National Center for Environmental Assessment, Office
of Research and Development, U.S. Environmental Protection Agency, Research
Triangle Park, NC
Dr. Reeder Sams III (Acting Deputy Director)—National Center for Environmental
Assessment-RTP Division, Office of Research and Development, U.S. Environmental
Protection Agency, Research Triangle Park, NC
Dr. Doug Johns (Acting Branch Chief)—National Center for Environmental Assessment,
Office of Research and Development, U.S. Environmental Protection Agency,
Research Triangle Park, NC
Scientific Staff
Dr. Ellen Kirrane (Team Leader, ISA for Lead)—National Center for Environmental
Assessment, Office of Research and Development, U.S. Environmental Protection
Agency, Research Triangle Park, NC
Dr. James Brown—National Center for Environmental Assessment, Office of Research
and Development, U.S. Environmental Protection Agency, Research Triangle Park,
NC
Mr. Allen Davis—National Center for Environmental Assessment, Office of Research
and Development, U.S. Environmental Protection Agency, Research Triangle Park,
NC
Dr. Jean-Jacques Dubois—National Center for Environmental Assessment, Office of
Research and Development, U.S. Environmental Protection Agency, Research
Triangle Park, NC
Dr. Tara Greaver—National Center for Environmental Assessment, Office of Research
and Development, U.S. Environmental Protection Agency, Research Triangle Park,
NC
Dr. Erin Hines—National Center for Environmental Assessment, Office of Research and
Development, U.S. Environmental Protection Agency, Research Triangle Park, NC
Dr. Dennis Kotchmar—National Center for Environmental Assessment, Office of
Research and Development, U.S. Environmental Protection Agency, Research
Triangle Park, NC
xvii
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Dr. Meredith Lassiter—National Center for Environmental Assessment, Office of
Research and Development, U.S. Environmental Protection Agency, Research
Triangle Park, NC
Dr. Stephen McDow—National Center for Environmental Assessment, Office of
Research and Development, U.S. Environmental Protection Agency, Research
Triangle Park, NC
Dr. Qingyu Meng—Oak Ridge Institute for Science and Education, National Center for
Environmental Assessment, Office of Research and Development,
U.S. Environmental Protection Agency, Research Triangle Park, NC
Dr. Elizabeth Oesterling Owens—National Center for Environmental Assessment, Office
of Research and Development, U.S. Environmental Protection Agency, Research
Triangle Park, NC
Dr. Molini Patel—National Center for Environmental Assessment, Office of Research
and Development, U.S. Environmental Protection Agency, Research Triangle Park,
NC
Dr Joseph P. Pinto—National Center for Environmental Assessment, Office of Research
and Development, U.S. Environmental Protection Agency, Research Triangle Park,
NC
Dr. Jennifer Richmond-Bryant—National Center for Environmental Assessment, Office
of Research and Development, U.S. Environmental Protection Agency, Research
Triangle Park, NC
Dr. Lindsay Stanek—National Center for Environmental Assessment, Office of Research
and Development, U.S. Environmental Protection Agency, Research Triangle Park,
NC
Dr. David Svendsgaard—National Center for Environmental Assessment, Office of
Research and Development, U.S. Environmental Protection Agency, Research
Triangle Park, NC
Dr. Lisa Vinikoor-Imler—National Center for Environmental Assessment, Office of
Research and Development, U.S. Environmental Protection Agency, Research
Triangle Park, NC
Technical Support Staff
Ms. Marieka Boyd—National Center for Environmental Assessment, Office of Research
and Development, U.S. Environmental Protection Agency, Research Triangle Park,
NC
Mr. Kenneth J. Breito-Senior Environmental Employment Program, National Center for
Environmental Assessment, Office of Research and Development,
U.S. Environmental Protection Agency, Research Triangle Park, NC
Mr. Gerald Gurevich—National Center for Environmental Assessment, Office of
Research and Development, U.S. Environmental Protection Agency, Research
Triangle Park, NC
Mr. Ryan Jones—National Center for Environmental Assessment, Office of Research and
Development, U.S. Environmental Protection Agency, Research Triangle Park, NC
xvin
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Ms. Ellen Lorang—National Center for Environmental Assessment, Office of Research
and Development, U.S. Environmental Protection Agency, Research Triangle Park,
NC
Mr. J. Sawyer Lucy-Student Services Authority, National Center for Environmental
Assessment, Office of Research and Development, U.S. Environmental Protection
Agency, Research Triangle Park, NC
Ms. Deborah Wales—National Center for Environmental Assessment, Office of Research
and Development, U.S. Environmental Protection Agency, Research Triangle Park,
NC
Mr. Richard N. Wilson-National Center for Environmental Assessment, Office of
Research and Development, U.S. Environmental Protection Agency, Research
Triangle Park, NC
Ms. Barbara Wright—Senior Environmental Employment Program, National Center for
Environmental Assessment, Office of Research and Development,
U.S. Environmental Protection Agency, Research Triangle Park, NC
xix
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Authors, Contributors, and Reviewers
Authors
Dr. Ellen Kirrane (Team Leader, ISA for Lead)—National Center for Environmental
Assessment, Office of Research and Development, U.S. Environmental Protection
Agency, Research Triangle Park, NC
Dr. Robyn Blain— Energy, Environment and Transportation, Environmental Science &
Policy, ICF International, Lexington, MA
Dr. James Brown—National Center for Environmental Assessment, Office of Research
and Development, U.S. Environmental Protection Agency, Research Triangle Park,
NC
Dr. Philip Bushnell, National Health and Environmental Effects Research Laboratory,
Office of Research and Development, U.S. Environmental Protection Agency,
Research Triangle Park, NC
Mr. Allen Davis—National Center for Environmental Assessment, Office of Research
and Development, U.S. Environmental Protection Agency, Research Triangle Park,
NC
Dr. Gary Diamond—Syracuse Research Corporation, Akron, NY
Dr. Rodney Dietert—Cornell University College of Veterinary Medicine, Veterinary
Medical Center, Ithaca, NY
Dr. Jean-Jacques Dubois—National Center for Environmental Assessment, Office of
Research and Development, U.S. Environmental Protection Agency, Research
Triangle Park, NC
Dr. Anne Fairbrother—Exponent, Inc., Bellevue, WA
Dr. Jay Gandy—Department of Environmental and Occupational Health, University of
Arkansas for Medical Sciences, Little Rock, AR
Dr. Harvey Gonick—David Geffen School of Medicine, University of California-Los
Angeles, Los Angeles, CA
Dr. Margaret Graham—School of Geosciences, University of Edinburgh, Edinburgh,
Scotland
Dr. Tara Greaver—National Center for Environmental Assessment, Office of Research
and Development, U.S. Environmental Protection Agency, Research Triangle Park,
NC
Dr. Erin Hines—National Center for Environmental Assessment, Office of Research and
Development, U.S. Environmental Protection Agency, Research Triangle Park, NC
Dr. Dennis Kotchmar—National Center for Environmental Assessment, Office of
Research and Development, U.S. Environmental Protection Agency, Research
Triangle Park, NC
xx
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Dr. Meredith Lassiter—National Center for Environmental Assessment, Office of
Research and Development, U.S. Environmental Protection Agency, Research
Triangle Park, NC
Dr. Stephen McDow—National Center for Environmental Assessment, Office of
Research and Development, U.S. Environmental Protection Agency, Research
Triangle Park, NC
Dr. Qingyu Meng—Oak Ridge Institute for Science and Education, National Center for
Environmental Assessment, Office of Research and Development,
U.S. Environmental Protection Agency, Research Triangle Park, NC
Dr. Bill Mendez—Energy, Environment and Transportation, Environmental Science &
Policy, ICF International, Fairfax, VA
Dr. Howard Mielke—Center for Bioenvironmental Research, Tulane/Xavier Universities,
New Orleans, LA
Ms. Chandrika Moudgal— Energy, Environment and Transportation, Environmental
Science & Policy, ICF International, Dublin, CA
Dr. Elizabeth Oesterling Owens—National Center for Environmental Assessment, Office
of Research and Development, U.S. Environmental Protection Agency, Research
Triangle Park, NC
Dr. Katherine Palmquist—Exponent, Inc., Bellevue, WA
Dr. Molini Patel—National Center for Environmental Assessment, Office of Research
and Development, U.S. Environmental Protection Agency, Research Triangle Park,
NC
Dr Joseph P. Pinto—National Center for Environmental Assessment, Office of Research
and Development, U.S. Environmental Protection Agency, Research Triangle Park,
NC
Dr. Jennifer Richmond-Bryant—National Center for Environmental Assessment, Office
of Research and Development, U.S. Environmental Protection Agency, Research
Triangle Park, NC
Dr. Stephen Rothenberg—National Institute of Public Health, Cuernavaca, Morelos,
Mexico
Dr. Mary Jane Selgrade—Energy, Environment and Transportation, Environmental
Science & Policy, ICF International, RTP, NC
Dr. Lindsay Stanek—National Center for Environmental Assessment, Office of Research
and Development, U.S. Environmental Protection Agency, Research Triangle Park,
NC
Dr. David Svendsgaard—National Center for Environmental Assessment, Office of
Research and Development, U.S. Environmental Protection Agency, Research
Triangle Park, NC
Dr. Lisa Vinikoor-Imler—National Center for Environmental Assessment, Office of
Research and Development, U.S. Environmental Protection Agency, Research
Triangle Park, NC
Dr. Virginia Weaver—Johns Hopkins Bloomberg School of Public Health, Baltimore,
MD
xxi
-------
Dr. Marc Weisskopf—Department of Environmental Health and Department of
Epidemiology, Harvard School of Public Health, Harvard University, Boston,
Massachusetts
Dr. John Pierce Wise, Sr.—Maine Center for Toxicology and Environmental Health,
Department of Applied Medical Sciences, Portland, ME
Dr. Rosalind Wright—Harvard Medical School and School of Public Health, Harvard
University, Boston, MA
Dr. Robert Wright—Harvard Medical School and School of Public Health, Harvard
University, Boston, MA
Contributors
Mr. Brian Adams—Oak Ridge Institute for Science and Education, National Center for
Environmental Assessment, Office of Research and Development,
U.S. Environmental Protection Agency, Research Triangle Park, NC
Dr. Halil Cakir— Office of Air Quality Planning and Standards, Office of Air and
Radiation, U.S. Environmental Protection Agency, Research Triangle Park, NC
Ms. Ye Cao—Oak Ridge Institute for Science and Education, National Center for
Environmental Assessment, Office of Research and Development,
U.S. Environmental Protection Agency, Research Triangle Park, NC
Ms. Laura Datko—Oak Ridge Institute for Science and Education, National Center for
Environmental Assessment, Office of Research and Development,
U.S. Environmental Protection Agency, Research Triangle Park, NC
Mr. Josh Drukenbrod—Office of Air Quality Planning and Standards, Office of Air and
Radiation, U.S. Environmental Protection Agency, RTP, NC
Ms. Meagan Madden—Oak Ridge Institute for Science and Education, National Center
for Environmental Assessment, Office of Research and Development,
U.S. Environmental Protection Agency, Research Triangle Park, NC
Mr. Mark Schmidt—Office of Air Quality Planning and Standards, Office of Air and
Radiation, U.S. Environmental Protection Agency, Research Triangle Park, NC
Ms. Katie Shumake—Oak Ridge Institute for Science and Education, National Center for
Environmental Assessment, Office of Research and Development,
U.S. Environmental Protection Agency, Research Triangle Park, NC
Ms. Kaylyn Siporin—Oak Ridge Institute for Science and Education, National Center for
Environmental Assessment, Office of Research and Development,
U.S. Environmental Protection Agency, Research Triangle Park, NC
Ms. Lauren Turtle—Oak Ridge Institute for Science and Education, National Center for
Environmental Assessment, Office of Research and Development,
U.S. Environmental Protection Agency, Research Triangle Park, NC
Ms. Adrien Wilkie—Oak Ridge Institute for Science and Education, National Center for
Environmental Assessment, Office of Research and Development,
U.S. Environmental Protection Agency, Research Triangle Park, NC
xxu
-------
Ms. Brianna Young—Oak Ridge Institute for Science and Education, National Center for
Environmental Assessment, Office of Research and Development,
U.S. Environmental Protection Agency, Research Triangle Park, NC
Reviewers
Dr. David Bellinger - Harvard Medical School and Department of Environmental Health,
Harvard School of Public Health, Boston, MA
Dr. Christal Bowman—National Center for Environmental Assessment, Office of
Research and Development, U.S. Environmental Protection Agency, Research
Triangle Park, NC
Dr. David Buchwalter—Department of Toxicology, North Carolina State University,
Raleigh, NC
Dr. Barbara Buckley—National Center for Environmental Assessment, Office of
Research and Development, U.S. Environmental Protection Agency, Research
Triangle Park, NC
Mr. Kevin Cavender—Office of Air Quality Planning and Standards, Office of Air and
Radiation, U.S. Environmental Protection Agency, Research Triangle Park, NC
Ms. Rebecca C. Dzubow—Office of Children's Health Protection, U.S. Environmental
Protection Agency, Washington, DC
Dr. David DeMarini—National Health and Environmental Effects Research Laboratory,
Office of Research and Development, U.S. Environmental Protection Agency,
Research Triangle Park, NC
Dr. Pam Factor-Litvak—Department of Epidemiology, Mailman School of Public Health,
New York, NY
Dr. Gabriel Filippelli—Department of Earth Sciences, Indiana University-Purdue
University, Indianapolis, IN
Dr. Andrew Friedland—Environmental Studies Program, Dartmouth College, Hanover,
NH
Dr. Barbara Glenn—National Center for Environmental Assessment, Office of Research
and Development, U.S. Environmental Protection Agency, Washington, DC
Dr. Jeff Herrick—National Center for Environmental Assessment, Office of Research
and Development, U.S. Environmental Protection Agency, Research Triangle Park,
NC
Dr. Marion Hoyer—Office of Transportation and Air Quality, Office of Air and
Radiation, U.S. Environmental Protection Agency, Ann Arbor, MI
Dr. Joseph Jacobson - Department of Psychiatry and Behavioral Neurosciences, Wayne
State University, Detroit, MI
Dr. Douglas Johns—National Center for Environmental Assessment, Office of Research
and Development, U.S. Environmental Protection Agency, Research Triangle Park,
NC
Dr. Jay S. Kaufman - Department of Epidemiology, Biostatistics and Occupational
Health, McGill University, Montreal, Canada
xxin
-------
Dr. Thomas Luben—National Center for Environmental Assessment, Office of Research
and Development, U.S. Environmental Protection Agency, Research Triangle Park,
NC
Dr. Karen Martin—Office of Air Quality Planning and Standards, Office of Air and
Radiation, U.S. Environmental Protection Agency, Research Triangle Park, NC
Ms. Connie Meacham—National Center for Environmental Assessment, Office of
Research and Development, U.S. Environmental Protection Agency, Research
Triangle Park, NC
Dr. Marie Lynn Miranda—Environmental Sciences and Policy, Nicholas School of the
Environment, Duke University, Durham, NC
Dr. Deirdre Murphy—Office of Air Quality Planning and Standards, Office of Air and
Radiation, U.S. Environmental Protection Agency, Research Triangle Park, NC
Dr. Paul Mushak—PB Associates, Durham NC
Dr. Kris Novak—National Center for Environmental Assessment, Office of Research and
Development, U.S. Environmental Protection Agency, Research Triangle Park, NC
Mr. David Orlin—Air and Radiation Law Office, Office of General Counsel, U.S.
Environmental Protection Agency, Washington, DC
Dr. Meredith Pedde—Office of Transportation and Air Quality, Office of Air and
Radiation, U.S. Environmental Protection Agency, Ann Arbor, MI
Dr. Pradeep Raj an—Office of Air Quality Planning and Standards, Office of Air and
Radiation, U.S. Environmental Protection Agency, Research Triangle Park, NC
Ms. Joanne Rice—Office of Air Quality Planning and Standards, Office of Air and
Radiation, U.S. Environmental Protection Agency, Research Triangle Park, NC
Dr. Mary Ross—National Center for Environmental Assessment, Office of Research and
Development, U.S. Environmental Protection Agency, Research Triangle Park, NC
Dr. Joel Schwartz—Department of Environmental Health, Harvard School of Public
Health, Boston, MA
Mr. Jason Sacks—National Center for Environmental Assessment, Office of Research
and Development, U.S. Environmental Protection Agency, Research Triangle Park,
NC
Ms. Ginger Tennant—Office of Air Quality Planning and Standards, Office of Air and
Radiation, U.S. Environmental Protection Agency, Research Triangle Park, NC
Dr. Jay Turner—Environmental and Chemical Engineering Department, Washington
University, St. Louis, MO
Dr. John Vandenberg—National Center for Environmental Assessment-RTP Division,
Office of Research and Development, U.S. Environmental Protection Agency,
Research Triangle Park, NC
Dr. Robert W. Vanderpool—National Exposure Research Laboratory, Office of Research
and Development, U.S. Environmental Protection Agency, Research Triangle Park,
NC
Dr. Nosratola Vaziri—Division of Nephrology and Hypertension, School of Medicine,
University of California, Irvine, Irvine, CA
xxiv
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Ms. Debra Walsh—National Center for Environmental Assessment-RTF Division, Office
of Research and Development, U.S. Environmental Protection Agency, Research
Triangle Park, NC
Dr. Nasser Zawia—Department of Biomedical and Pharmaceutical Sciences, University
of Rhode Island, Kingston, RI
xxv
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Clean Air Scientific Advisory Committee
Lead NAAQS Review Panel
Chair of the Lead Review Panel
Dr. Christopher H. Frey*, North Carolina State University, Raleigh, NC
Lead Review Panel Members
Mr. George A. Allen*, Northeast States for Coordinated Air Use Management
(NESCAUM), Boston, MA
Dr. Herbert Allen, University of Delaware, Newark, DE
Dr. Richard Canfield, Cornell University, Ithaca, NY
Dr. Deborah Cory-Slechta, University of Rochester, Rochester, NY
Dr. Cliff Davidson, Syracuse University, Syracuse, NY
Dr. Philip E. Goodrum, Cardno ENTRIX, Syracuse, NY
Dr. Sean Hays, Summit Toxicology, Allenspark, CO
Dr. Philip Hopke, Clarkson University, Potsdam, NY
Dr. Chris Johnson, Syracuse University, Syracuse, NY
Dr. Susan Korrick, Harvard Medical School, Boston, MA
Dr. Michael Kosnett, University of Colorado School of Medicine, Denver, CO
Dr. Roman Lanno, Ohio State University, Columbus, OH
Mr. Richard L. Poirot, Vermont Agency of Natural Resources, Waterbury, VT
Dr. Joel Pounds, Pacific Northwest National Laboratory, Richland, WA
Dr. Michael Rabinowitz, Marine Biological Laboratory, Newport, RI
Dr. William Stubblefield, Oregon State University, Corvallis, OR
Dr. Ian von Lindern, TerraGraphics Environmental Engineering, Inc., Moscow, ID
Dr. Gail Wasserman, Columbia University, New York, NY
Dr. Michael Weitzman, New York University School of Medicine, New York, NY
* Members of the statutory Clean Air Scientific Advisory Committee (CASAC)
appointed by the EPA Administrator
Science Advisory Board Staff
Mr. Aaron Yeow, Designated Federal Officer, U.S. Environmental Protection Agency,
Science Advisory Board (1400R), 1200 Pennsylvania Avenue, NW, Washington, D.C.
20460-0001, Phone: 202-564-2050, Fax: 202-565-2098, (yeow.aaron@epa.gov)
XXVI
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Acronyms and Abbreviations
Acronym/Abbreviation
a
aT
AA
AALM
AAS
Ab
ABL
ACE
ACF
ACh
ACP
ACR
Acyl-CoA
AD
ADHD
ADP
AE
AEC
AERMOD
AF
affd
A/G
Ag
AGL
A-horizon
AKI
Al
ALA
ALAD
Meaning
alpha
the extent of DNA denaturation
per cell
Angstrom (10"10 meter)
African American; arachidonic
acid, atomic absorption
All Ages Lead Model
atomic absorption
(spectrophotometry,
spectrometry, spectroscopy)
amyloid-beta peptide
atmospheric boundary layer
angiotensin converting enzyme
Apalachicola, Chattahoochee,
and Flint River Basin
acetylcholine
acid phosphatase
acute to chronic ratio
acyl-coenzyme A
axial diffusivity
attention deficit hyperactivity
disorder
adenosine diphosphate
anion exchanger
adenylate energy charge
atmospheric dispersion model
absorbed fraction; absorption
fraction
affirmed
albumin/globulin
silver
above-ground level
Topsoil horizon (surface soil)
acute kidney injury
aluminum
aminolevulinic acid
aminolevulinic acid
dehydratase;
Acronym/Abbreviation
ALAD 1-1
ALAD-2
ALD
ALM
ALP
ALS
ALT
AM
AMF
AMP
ANC
ANF
Angll
ANOVA
ANPR
AOP
AP-1
Apal
APC
APOE
APRT
AQCD
AQS
Ar
As
AST
ASV
ATLD
Meaning
aminolevulinate delta-
dehydratase 1-1
aminolevulinate delta-
dehydratase-2
aldehyde dehydrogenase
Adult Lead Methodology
alkaline phosphatase
Amyotrophic Lateral Sclerosis
(Lou Gehrig's disease)
alanine aminotransferase
Alveolar macrophages
arbuscular mycorrhizal fungi
adenosine monophosphate
acid neutralizing capacity;
absolute neutrophil counts
atrial natriuretic factor
renal angiotensin II
analysis of variance
advance notice of proposed
rulemaking
adverse outcome pathway
activator protein-1
polymorphism of the VDR in
humans
antigen-presenting cell
Apolipoprotein E
adenine
phosphoribosyltransferase
Air Quality Criteria Document
(U.S. EPA) Air Quality System
(database)
argon
arsenic
aspartate aminotransf erase
anode stripping voltammetry
ataxia-telangiectasia-like
disorder
XXVll
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Acronym/Abbreviation
ATOFMS
ATP
ATPase
ATS
ATSDR
Au
avg
AVS
a-wave
AWQC
P
3P-HSD
17P-HSD
Ba
BAF
BAL
BASC
BASC-PRS
BASC-TRS
BC
BCB
B-cell
BCF
Bcl-x
Bcl-xl
B-horizon
bio
Meaning
aerosol time-of-flight mass
spectrometry
adenosine-triphosphate
adenosine triphosphatase;
adenosine triphosphate
synthase
American Thoracic Society
Agency for Toxic Substances
and Disease Research
gold
average
acid-volatile sulfides
initial negative deflection in the
electroretinogram
Ambient Water Quality Criteria
Beta; Beta coefficient;
regression coefficient;
standardized coefficient
3 -beta-hy droxy steroid
dehydrogenase
17-beta-hydroxy steroid
dehydrogenase
barium
bioaccumulation factors
2,3-dimercaptopropanol
Behavior Assessment System
for Children
Behavior Assessment System
for Children-Parent Ratings
Scale
Behavior Assessment System
for Children-Teacher Rating
Scale
black carbon, soot
blood cerebrospinal fluid
barrier
Bone marrow-derived
lymphocytes, B lymphocyte
bioconcentration factors
member of the B-cell
lymphoma-2 protein family
B-cell lymphoma-extra large
subsoil horizon
biological
Acronym/Abbreviation
Bi2S3
BK
BLM
BMD
BMDL
BMI
BMP
BMS
BMW
BP
Br
BR
BrdU
8-Br-GMPc
Bs-horizon
BSI
BSID-II
BsmI
Bt20
BUN
bw
b-wave
C
Ca
Ca2+
CAA
CaBP
CaCl2
CaC03
CaEDTA
CaMKII
Meaning
bismuth (III) sulfide
biokinetics
biotic ligand model
benchmark dose; bone mineral
density
benchmark dose limit
body mass index
bone morphogenetic protein
Baltimore Memory Study
battery manufacturing workers
blood pressure
bromine
bronchial responsiveness
bromo-2'-deoxyuridine
8-bromo-cyclic guanosine
monophosphate
subsoil horizon with
accumulation of sesquioxides
Brief Symptom Inventory
Bay ley Scale for Infant
Development-II
polymorphism of the VDR in
humans
Birth-to-age Twenty (cohort)
blood urea nitrogen
body weight
initial positive deflection in the
electroretinogram
carbon; Celsius; soil or dry
sediment Pb concentration;
Caucasian; Cysteine
calcium
calcium ion
Clean Air Act
calcium binding protein
calcium chloride
calcium carbonate; calcite
calcium
ethylenediaminetetraacetic acid
calmodulin-dependent protein
kinase II
XXVlll
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Acronym/Abbreviation
cAMP
CASAC
CASM
CaSO4
CaSO4-2H2O
CAT
CBLI
CBSA
CCSEM
CD
Cd
Cd(II)
Cd2+
CD3+
CD4+
CDC
CEA
CEC
cent
cert.
cf
CFL
CFR
cGMP
C-H
CHAD
ChAT
CHD
CHL
CHO
C-horizon
Meaning
cyclic adenosine
monophosphate
Clean Air Scientific Advisory
Committee
Comprehensive Aquatic
Systems Model
calcium sulfate
gypsum
catalase
cumulative blood Pb index
core based statistical area
computer-controlled scanning
electron microscopy
cluster of differentiation
cadmium
cadmium (II)
cadmium ion
T lymphocyte
T helper cell
Centers for Disease Control
carcinoembryonic antigen
cation exchange capacity
central
certiorari
correction factor; latin
abbreviation for conferre (used
as "compared with)
constant flux layer
Code of Federal Regulations
cyclic guanosine
monophosphate
carbon-hydrogen (bond)
Consolidated Human Activity
Database
chlorine acetyltransferase
coronary heart disease
Chinese hamster lung
Chinese hamster ovary cell line
Soil horizon underneath A- and
B-horizons, may contain lumps
or shelves of rock and parent
material
Acronym/Abbreviation
CHV79
CI
Cir.
CKD
CKD-EPI
CL
Cl
cr
C12
CLACE 5
CLS
CO
C02
C032'
Co
CoA
COD
Coeff
COMPaT
Con
Cone.
Cong.
Con-
COX
COX-2
cPLA2
CPRI
CPRS-R
Cr
C-R
CrIII
CRAC
Meaning
Chinese hamster lung cell line
confidence interval
circuit
chronic kidney disease
Chronic Kidney Disease
Epidemiology Collaboration
confidence limit
chlorine
chlorine ion
molecular chlorine
Fifth Cloud and Aerosol
Characterization Experiment in
the Free Troposphere campaign
Cincinnati Lead Study
carbon monoxide
carbon dioxide
carbonate ion
cobalt
coenzyme A
coefficient of difference
coefficient
The percentage of sperm with
increased sensitivity to DNA
denaturation
control
concentration
congress
correlation
cyclooxygenase; cytochrome
oxidase subunits
cyclooxygenase-2
cytosolic phospholipidase A2
coarse particle rotary impactor
Conners' Parent Rating Scale-
Revised
chromium; creatine
concentration-response
(relationship)
chromium III
Ca2+ release activated calcium
XXIX
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Acronym/Abbreviation
CRACI
CREB
CRP
CSF
CSN
CT
Cu
Cu(II)
CV
CVD
CYP
CYPlAl,CyplAl
CYP 1A2, CyplA2
CYP P450
A
A5-3P-HSD
5-ALA
5-ALAD
D2,D3
D50
d
db,dB
DbH
DBF
DENA
dep
dev.
DEX
DG
2-dG
DHAA
diff
Meaning
calcium release activated
calcium influx
cyclic
adeno sinemonopho sphate
(cAMP) response element-
binding
C-reactive protein
colony-stimulating factor
Chemical Speciation Network
zinc-adequate control
copper
copper (II)
coefficient of variation
cardiovascular disease
cytochrome
cytochrome P450 family
1 member A1
cytochrome P450 family 1
member A2
cytochrome P450
delta, difference, change
delta-5-3-beta-hydroxy steroid
dehydrogenase
5-aminolevulinic acid; delta-
aminolevulinic acid
delta-aminolevulinic acid
dehydratase
dopamine receptors
size at 50% efficiency
day(s); depth
decibel
dopamine beta-hydroxylase
diastolic blood pressure
Denali National Park and
Preserve, Alaska
dependent
deviation
exogenous dexamethasone
degenerate gyrus
2-deoxyguanosine
dehydroascorbate
differentiation
Acronym/Abbreviation
BIT
DMPS
DMSA
DMSO
DNA
DoAD
DOC
DOM
DP-109
DP-460
DR
DRD4
DRD4.7
DRUM
D-serine
DSM-IV
DTK
DTPA
E
E2
e
EC
EC10
EC 20
EC 5o
ECG
ECOD
Eco-SSLs
ED10
Meaning
developmental immunotoxicity
2,3-dimercaptopropane-l-
sulfonic acid
dimercaptosuccinic acid
dimethyl sulfoxide
deoxyribonucleic acid
developmental origins of adult
disease
dissolved organic carbon
dissolved organic matter
metal chelator
metal chelator
diet-restricted
dopamine 4 receptor
dopamine 4 receptor repeat
alleles
Davis Rotating Unit for
Monitoring
neuronal signal
Diagnostic Statistical Manual-
IV
delay ed-type hypersensitivity
diethylene triamine pentaacetic
acid; technetium-
diethylenetriamine-pentaacetic
acid
east; expression for exposure
estradiol
exponential function
elemental carbon, endothelial
cell
effect concentration for 10% of
test population
effect concentration for 20% of
test population
effect concentration for 50% of
test population
electrocardiography;
electrocardiogram
7-ethoxycoumarin-o-deethylase
ecological soil screening levels
effect dose for 10% of
population
XXX
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Acronym/Abbreviation
EDTA
EPS
EOF
EGFR
eGFR
Eh
E-horizon
EI-MS
ELPI
eNOS
EOG
EPA
EPT
ER
Erg-1
ERG
ERK
ERK1/2
EROD
ESCA
ESI-MS
ESRD
ET
ET-1
ETA-type receptors
ETS
EU
EURO
eV
Meaning
ethylenediaminetetraacetic acid
electrical field stimulus
epidermal growth factor
epidermal growth factor
receptor
estimated glomerular filtration
rate
electrochemical potential
Eluviated horizon; soil horizon
which is eluviated or leached of
mineral and/or organic content
electron impact ionization mass
spectrometry
electrical low-pressure
impactor
endothelial nitric oxide
synthase
end-of-grade
U.S. Environmental Protection
Agency
ephemeroptera, plecoptera,
trichoptera
endoplasmic reticulum
ether-a-go-go related gene
electroretinogram
extracellular signal regulated
kinase
extracellular signal-regulated
kinases 1 and 2
7-ethoxyresorufin-o-deethylase
electron spectroscopy for
chemical analysis
electrospray ionization mass
spectrometry
end stage renal disease
endothelin
vasoconstrictor endothelin-1
endothelin type A receptors
environmental tobacco smoke
European Union
European emission standard
electronvolts
Acronym/Abbreviation
EXAFS
F2
FAA
FAI
FAS
Fas-L
Fe
Fe(III)
FEM
FEV1
FI
FI-Ext
Fl
Fokl
FR
FrA
FR-FI
FRM
FSH
FSIQ
FT3
FT4
G
G2
g, kg, mg, (ig, ng, pg
G93A
GAAR
GABA
Meaning
X-ray absorption fine structure
spectroscopy
filial "zero" generation
(parental stock)
first filial generation (offspring
ofF0)
second filial generation
(offspring ofF^
Federal Aviation Agency
free androgen index
apoptosis stimulating fragment
apoptosis stimulating fragment
ligand
iron
iron III
Federal equivalence method
forced expiratory volume in 1
second
fixed interval
fixed interval with extinction
fluoride
polymorphism of the VDR in
humans
Federal Register (Notice)
fractional anisotropy
fixed ratio-fixed interval
Federal reference method
follicle-stimulating hormone
full scale intelligence quotient
(IQ)
free triiodothyronine
free thyroxine
pregnancy; guanine
gap 2 Phase
gram(s), kilogram(s),
milligram(s), microgram(s),
nanogram(s), picogram(s)
mouse model
Gates of the Arctic National
Park and Preserve, Alaska
y-aminobutyric acid; gamma
aminobutyric acid
XXXI
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Acronym/Abbreviation
GABAergic
GAD
GC
G-CSF
GD
GEE
GFAAS
GFAP
GFR
GGT
GH
GHRH
GI
GIS
G+L
GLAC
GLE
GM
GMR
GnRH
G6PD
GPEI
GPT
GPx
GPX1
GR
GRP78
GRP94
Grp
GSD
Meaning
inhibitory neurons that release
the neurotransmitter GABA
generalized anxiety disorder
gas chromatography
granulocyte colony-stimulating
factor
gestational day
generalized estimating
equations
graphite furnace atomic
absorption spectrometry
glial fibrillary acidic protein
glomerular filtration rate
gamma-glutamyl
transpeptidase
growth hormone
growth-hormone releasing
hormone
gastrointestinal
Geographic Information
System
pregnancy plus lactation
Glacier National Park,
Montana
gestationally-lead exposed
geometric mean
geometric mean blood Pb ratio
gonadotropin-releasing
hormone
glucose-6-phosphate
dehydrogenase
glutathione transferase P (GST-
P) enhancer I
glutamate pyruvate
transaminase
glutathione peroxidase
gene encoding for glutathione
peroxidase 1
glutathione reductase
glucose-regulated protein 78
glucose-regulated protein 94
glucose-regulated protein
geometric standard deviation
Acronym/Abbreviation
GSH
GSSG
GST
GSTM1
GST-P
GTP
H
FT
h
ha
HAD
HAP
Hb
HC5
HC10
HC1
HC03-
Hct
HDL
HERO
HEW
HF
HFE
HFE C282Y
HFE H63D
Hg
HgCl2
5-fflAA
HIV
HLA-DRB
HMEC
Meaning
glutathione
glutathione disulfide
glutathione S-transferase
glutathione S-transferase Mu 1
glutathione transferase P
guano sine-5'-tripho sphate;
guanine tripho sphate
hydrogen
hydrogen ion
hour(s)
hectare
hydroxy-alkenals
hazardous air pollutant
hemoglobin
acute toxicity hazardous
concentration for 5% of species
acute toxicity hazardous
concentration for 10% of
species
hydrochloric acid
bicarbonate; hydrogen
carbonate
hematocrit
high-density lipoprotein
Health and Environmental
Research online (database)
U.S. Department of Health,
Education, and Welfare
hydrogen fluoride
hemochromatosis gene
hemochromatosis gene with
C282Y mutation
hemochromatosis gene with
H63D mutation
mercury
mercury(II) chloride
5-hydroxyindoleacetic acid
human immunodeficiency virus
human leukocyte antigen genes
human dermal micro vascular
endothelial cells
XXXll
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Acronym/Abbreviation
HMGR
Meaning
3-hydroxy-3-methylglutaryl-
CoA reductase
Acronym/Abbreviation
ID
IDA
HMOX-1
HNO3
HO-1
H20
H202
HOME
HPA
HPb, h-Pb
HPG
HPLC
HPRT
HPT
HR
HRV
hsp
5HT
5-HT
5-HT2B
hTERT
HVA
I
IARC
IC50
ICAP
ICP-AES
ICPMS, ICP-MS
ICR
ICRP
heme oxygenase-1
nitric acid
heme oxygenase; heme
oxidase-1
water
hydrogen peroxide
Home Observation for
Measurement of the
Environment
hypothalamic-pituitary-adrenal
high Pb
hypothalamic-pituitary-gonadal
high-performance liquid
chromatography
hypoxanthine-guanine
phosphoribosyltransf erase
hyperparathyroidism;
hypothalamic-pituitary -thyroid
heart rate; hazard ratio
heart rate variability
heat shock proteins
serotonin
5-hydroxytryptamine
5-hydroxytryptamine receptor
2B
telomerase reverse
transcriptase
homovanillic acid
interstate
International Agency for
Research on Cancer
half maximal inhibitory
concentration
inductively coupled argon
plasma
Inductively coupled plasma
atomic emission spectroscopy
Inductively coupled plasma
mass spectrometry
imprinting control region
International Commission on
IDE
IEPA
IEUBK
IFN-y
Ig
IgA
IgE
IGF-1
IgG
IgM
IHD
IL
IL-lp
IL-2
IL-4
IL-5
IL-6
IL-8
IL-10
IL-12
IMPROVE
IMT
INL
iNOS
IOM
i.p.
IQ
IQR
IRE1
IRP
ISA
Meaning
identification
iron-deficiency anemia
insulin-degrading enzyme
Illinois Environmental
Protection Agency
Integrated Exposure Uptake
Biokinetic
interferon-gamma
immunoglobulin
immunoglobulin A
immunoglobulin E
insulin-like growth factor 1
immunoglobulin G
immunoglobulin M
ischemic heart disease
interleukin
interleukin-1 Beta
interleukin-2
interleukin-4
interleukin-5
interleukin-6
interleukin-8
interleukin-10
interleukin-12
Interagency Monitoring of
Protected Visual Environment
intimal medial thickening
inner neuroblastic layers of the
retina
inducible nitric oxide synthase
Institute of Medicine (provides
health information to the NAS
[National Academy of
Sciences])
intraperitoneal (route)
intelligence quotient
interquartile range
inositol-requiring enzyme-1
integrated review plan
Integrated Science Assessment
Radiological Protection
XXXlll
-------
Acronym/Abbreviation
ISC-PRIME
Meaning
Industrial Source Complex-
Plume Rise Model
Enhancements
Acronym/Abbreviation
LA-ICP-MS
ISF
ISL
ISO
i.v.
IVBA
IVF
JNK
K
K+
Kfl.5
intake slope factor
inertial sublayer
International Standards
Organization
intravenous
in vitro bioaccessibility
in vitro fertilization
jun N-terminal kinase
Kelvin (temperature);
potassium; resuspension factor
potassium ion
concentration of free metal
LC50
LD50
LDH
LDL
LFH-horizons
LF/HF
LH
LHRH
KART
Kd
Kd
kDa,kD
KEDI-WISC
6-keto-PGFla
keV
Ki-67
Kim-1
Kinder-KITAP
K-ras
K-XRF
A
L
L, dL, mL
giving half maximal metal-
dependent release
Karters of American Racing
Triad
dissociation constant
partition coefficient; ratio of
the metal concentration in soil
to that in soil solution
kiloDalton
Korean Educational
Development Institute-
Wechsler Intelligence Scale for
Children
6-keto-prostaglandin F1 a
(vasodilatory prostaglandin)
kiloelectron volt
antigen, cell cycle and tumor
growth marker
kidney injury molecule-1
Kinder-Testbatterie zur
Aufmerksamkeitsprufung fur
Kinder
specific proto-oncogene
K-x-ray fluorescence method
of scanning for bone Pb
lambda; resuspension rate
length
Liter(s)[1000mL/L],
deciliters) [100 mL/dL],
milliliter(s) [1 mL/mL]
LINE
LINE-1
LLNA
In
L-NAME
L-NOARG
LOD
LOEC
log
LPb
LPS
LSO
LTP
M
M, mM, uM, nM, pM
m, km, cm, mm, um, nm
MAP
MAPK
Meaning
laser ablation inductively
coupled plasma mass
spectrometry
lethal concentration (at which
50% of exposed organisms die)
lethal dose (at which 50% of
exposed organisms die)
lactate dehydrogenase
low-density lipoproteins
organic soil horizons located
above well-drained surface soil
low frequency to high
frequency ratio
luteinizing hormone
luteinizing hormone releasing
hormone
long interspersed nuclear
element
long interspersed nucleotide
elements-1
local lymph node assay
natural logarithm
L-NG-nitroarginine methyl
ester
L-nitroarginine
limit of detection
lowest-observed-effect
concentration
logarithm
lowPb
lipopolysaccharide
lateral superior olive
long-term potentiation
metal
Molar, millimolar (10"3 M),
micromolar (10"6 M),
nanomolar (10" M), picomolar
(10'12M)
meter(s), kilometer(s),
centimeter(s), millimeter(s),
micrometer(s) [micron(s)],
nanometer(s)
mean arterial pressure
mitogen-activated protein
kinase(s), MAP kinase
XXXIV
-------
Acronym/Abbreviation
MATC
max
MBP
MCH
MCHC
MchDMSA
MCL
MCP-1
MCV
MD
MDA
MDD
MDI
MDL
MDRD
Med, med
MEK1
MEK2
MENTOR
Mg
Mg2+
MHC
MI
ml
min
MKK1/2
ML
MMAD
MMDD
MMF
mmHg
Meaning
maximum acceptable toxicant
concentration
maximum, maxima
myelin basic protein
mean corpuscular hemoglobin
mean corpuscular hemoglobin
concentration
mono-cyclohexyl
dimercaptosuccinic acid
maximum containment level
monocyte chemotactic
protein-1
mean corpuscular volume
mean diffusivity
malondialdehyde
major depressive disorder
Mental Development Index
method detection limit
Modification of Diet in Kidney
Disease
median
dual specificity mitogen-
activated protein kinase 1
dual specificity mitogen-
activated protein kinase 2
Modeling Environment for
Total Risk (framework)
magnesium
magnesium ion
major histocompatibility
complex
myocardial infarction, "heart
attack;" myocardial ischemia
myoinositol
minimum; minima; minute(s)
MAPK kinase 1 and 2
mixed layer
mass median aerodynamic
diameter
mental retardation or
developmental disabilities
mycophenolate mofetil
millimeters of mercury
Acronym/Abbreviation
mmol, (imol, nmol
MN
Mn
MNE
MnO2
Mo
mo
MOA(s)
MORA
MOUDI
MPb, m-Pb
MPO
MRI
mRNA
MRS
MS
MSC
MSWI
Mt
MTHFR
MTP
MW
MZ
N
n
Na
Na+
NAAQS
NAC
Na,CaEDTA
Meaning
millimole(s), micromole(s),
nanomole(s)
micronuclei formation;
mononuclear
manganese
micronucleated erythrocytes
per thousand
manganese dioxide
molybdenum
month(s)
mode(s) of action
Mount Rainier National Park,
Washington State
multi-orifice uniform deposit
impactor
moderate Pb
myeloperoxidase
magnetic resonance imaging
messenger ribonucleic acid
magnetic resonance
spectroscopy
maternal stress
mesenchymal cell
municipal solid waste
incineration
metallothionein
methylenetetrahydrofolate
reductase
mitochondria! transmembrane
pore
molecular weight
marginal zinc
nitrogen; normal; north;
number; population
number of observations
sodium
sodium ion
National Ambient Air Quality
Standards
N-acetyl cysteine; nucleus
accumbens
calcium disodium
ethylenediaminetetraacetic acid
XXXV
-------
Acronym/Abbreviation
NaCl
NAD
NADH
NADP
NADPH, NAD(P)H
NAEC
NAG
NaHC03
NANC
NAS
NASCAR
NATTS
NAWQA
NCAM
NCEA
NCore
N.D.
NDMAR
NE
NECAT
NEI
NFI
NF-KB
NGAL
NGF
NH
NHANES
Meaning
sodium chloride
nicotinamide adenine
dinucleotide
nicotinamide adenine
dinucleotide dehydrogenase
nicotinamide adenine
dinucleotide phosphate
reduced nicotinamide adenine
dinucleotide phosphate
no-adverse-effect concentration
N-acetyl-p-D-glucosaminidase;
N-acetylglucosamine
sodium bicarbonate; sodium
hydrogen carbonate
non-adrenergic non-cholinergic
U.S. Department of Veteran's
Affair's Normative Aging
Study; National Academy of
Sciences
National Association for Stock
Car Automobile Racing
National Air Toxics Trends
Station
National Water Quality
Assessment
neural cell adhesion molecule
National Center for
Environmental Assessment
National Core multi-pollutant
monitoring network
not detected
N-nitrosodimethylamine
receptor
norepinephrine
New England Children's
Amalgam Trial
National Emissions Inventory
non-fixed interval
nuclear factor kappa B
neutrophil gelatinase-
associated lipocalin
nerve growth factor
non-hispanic
National Health and Nutrition
Examination Survey
Acronym/Abbreviation
NH4C1
NHEJ
NHEXAS
NH4OAc
7-NI
Ni
NICA
NIOSH
NIST
NK
NKF-K/DOQI
NMDA
NMR
nNOS
NO
NO 2
No.
NOAA
NOAEL
NOAT
NOCA
NOEC
NOEL
NOS
NOX
NP
NPSH
NQ01
Meaning
ammonium chloride
non-homologous end joining
National Human Exposure
Assessment Survey
ammonium acetate
7-nitroinidazole
nickel
non-ideal competitive
absorption
National Institute for
Occupational Safety and Health
National Institute of Standards
and Technology
natural killer
National Kidney Foundation -
Kidney Disease Outcomes
Quality Initiative
N-methyl-D-aspartate
nuclear magnetic resonance
neuronal nitric oxide synthase
(NOS)
nitric oxide; nitric oxide
radical, nitrogen monoxide
nitrogen dioxide
number
National Oceanic and
Atmospheric Administration
no observed adverse effect
level
Noatak National Preserve,
Alaska
North Cascades National Park,
Washington State
no-observed-effect
concentration
no-observed-effect level
nitric oxide synthase; nitric
oxide systems
nitrogen oxides, oxides of
nitrogen (NO + NO 2)
nanoparticle
nonprotein sulfhydryl
NAD(P)H-quinone
oxidoreductase (genotype)
XXXVI
-------
Acronym/Abbreviation
NRC
NRCS
Meaning
National Research Council
Natural Resources
Acronym/Abbreviation
OSHA
Nrf2
NS
NTP
NTPDase
NW
NYC
NZ
02
o2-
03
9-O-Ac-GD3
OAQPS
OAR
OBS
OC
OEPA
OH"
1,25-(OH)2D3
O-horizon
OLC
OLYM
OM
ONL
ONOO"
OR
ORD
OS
Conservation Service
nuclear factor erythroid 2-
related factor 2
not specified
National Toxicology Program
nucleoside triphosphate
diphosphohydrolase
northwest
New York City
New Zealand
molecular oxygen
superoxide, superoxide free
radical
ozone
9-O-acetylated-GD3
U.S. EPA Office of Air Quality
Planning and Standards, in
OAR
U.S. EPA Office of Air and
Radiation
observations
organic carbon
Ohio Environmental Protection
Agency
hydroxide ion
1,25-dihydroxy vitamin D
horizon forest floor, organic
soil horizon (above surface
soil)
osteoblast-like cells
Olympic National Park,
Washington State
organic matter
outer neuroblastic layers of the
retina
peroxynitrate ion
odds ratio
U.S. EPA Office of Research
and Development
offspring stress
OVA
8-oxo-dG
P
Po
P450
P
PA
PAD
PAH(s)
Pb
203Pb
204Pb
206Pb
207p,
r D
208pb
210pb
Pb++
Pb°
Pb(II)
Pb2+
Pb(Ac)2
PbB
PbBrCl
Pb(C2H3O2)2
PbCf
PbCl2
PbCl3
PbCl4
PbC03
Pb(C03)2
Pb(C03)2(OH)2
PbCrO4
PbD
Meaning
Occupational Safety and Health
Administration
ovalbumin
8-hydroxy-2'-deoxyguanosine
percentile; phosphorus
parent generation
cytochrome P450
probability value; number of
paired hourly observations;
statistical significance
policy assessment
peripheral arterial disease
polycyclic aromatic
hydrocarbon(s)
lead
lead-203 radionuclide
stable isotope of lead-204
stable isotope of lead-206
stable isotope of lead-207
stable isotope of lead-208
stable isotope of lead-210
divalent Pb ion
elemental lead
lead (II)
lead ion
lead acetate
blood lead concentration
lead bromochloride
lead (II) acetate
lead chloride
lead chloride
lead (III) chloride; lead
trichloride
lead (IV) chloride; lead
tetrachloride
cerussite; lead carbonate
lead (IV) carbonate
hydrocerussite
lead (II) chromate
floor dust lead
XXXVll
-------
Acronym/Abbreviation
PbFe6(S04)4(OH)12
PEG
Pb(N03)2
Pb-NS
PbO
Pb02
Pb(IV)02
Pb304
Pb(OH)2
Pb5(P04)3Cl
Pb5(PO4)3OH
PbS
PbSe
PbS04
Pb4SO4(CO3)2(OH)3
PbxS
Pb5(V04)3Cl
PC12
PCA
PCE
PCR
Pet
PCV
PD
PDI
PEC
PEL
PER
PG
PGE2,PGE2
PGF2
PH
PHA
PHE
Meaning
plumbjarosite
porphobilinogen
lead(II) nitrate
lead-no stress
lead oxide; litharge; massicot
lead dioxide
lead dioxide
minimum or "red Pb"
lead hydroxide
pyromorphite
hy droxy pyromorphite
galena; lead sulfide; soil lead
concentration
lead selenide
anglesite; lead sulfate
macphersonite
lead by stress
vanadinite
pheochromocytoma 12 (adrenal
/ neuronal cell line)
principal component analysis
polychromatic erythrocyte
polymerase chain reaction
percent
packed cell volume
Parkinson's disease
Psychomotor Development
Index
probable effect concentration
permissible exposure limit
partial exfiltration reactor
prostaglandin
prostaglandin E2
prostaglandin F2
relative acidity; Log of the
reciprocal of the hydrogen ion
concentration
polyhy droxy alkanoates
phenylalanine
Acronym/Abbreviation
PIH
PIQ
PIR
PIXE
PKC
PLP
PM
PMX
Meaning
pregnancy induced
hypertension
performance intelligence
quotient (IQ)
poverty-income ratio
particle induced X-Ray
emission; proton-induced x-ray
emission
protein kinase C
proteolipid protein
participate matter
Particulate matter of a specific
size range not defined for
regulatory use. Usually X
refers to the 50% cut point, the
aerodynamic diameter at which
the sampler collects 50% of the
particles and rejects 50% of the
particles. The collection
efficiency, given by a
penetration curve, increases for
particles with smaller diameters
and decreases for particles with
larger diameters. The definition
of PMX is sometimes
abbreviated as "particles with a
nominal aerodynamic diameter
less than or equal to X |im"
although X is usually a 50%
cut point.
XXXVlll
-------
Acronym/Abbreviation
PM1n
PM2<
Meaning
In general terms, particulate
matter with an aerodynamic
diameter less than or equal to a
nominal 10 |im; a measurement
of thoracic particles (i.e., that
subset of inhalable particles
thought small enough to
penetrate beyond the larynx
into the thoracic region of the
respiratory tract). In regulatory
terms, particles with an upper
50% cut-point of 10 ± 0.5 (im
aerodynamic diameter (the
50% cut point diameter is the
diameter at which the sampler
collects 50% of the particles
and rejects 50% of the
particles) and a penetration
curve as measured by a
reference method based on
Appendix J of 40 CFR Part 50
and designated in accordance
with 40 CFR Part 53 or by an
equivalent method designated
in accordance with 40 CFR
Part 53.
In general terms, particulate
matter with an aerodynamic
diameter less than or equal to a
nominal 2.5 |im; a
measurement of fine particles.
In regulatory terms, particles
with an upper 50% cut-point of
2.5 |im aerodynamic diameter
(the 50% cut point diameter is
the diameter at which the
sampler collects 50% of the
particles and rejects 50% of the
particles) and a penetration
curve as measured by a
reference method based on
Appendix L of 40 CFR Part 50
and designated in accordance
with 40 CFR Part 53, by an
equivalent method designated
in accordance with 40 CFR
Part 53, or by an approved
regional method designated in
accordance with Appendix C of
40 CFR Part 58.
Acronym/Abbreviation
PM
PM,
p38MAPK
PMN
P5N
PND
POC
PP
ppb
ppm
PRP
PS
PSA
PSA-NCAM
PT
PTFE
PTH
Meaning
In general terms, particulate
matter with an aerodynamic
diameter less than or equal to a
nominal 10 |im and greater
than a nominal 2.5 |im; a
measurement of thoracic coarse
particulate matter or the coarse
fraction of PM10. In regulatory
terms, particles with an upper
50% cut-point of 10 |im
aerodynamic diameter and a
lower 50% cut-point of 2.5 |im
aerodynamic diameter (the
50% cut point diameter is the
diameter at which the sampler
collects 50% of the particles
and rejects 50% of the
particles) as measured by a
reference method based on
Appendix O of 40 CFR Part 50
and designated in accordance
with 40 CFR Part 53 or by an
equivalent method designated
in accordance with 40 CFR
Part 53.
The PM10-2.5 concentration of
PM10-2.5 measured by the 40
CFR Part 50 Appendix O
reference method which
consists of currently operated,
co-located low-volume
(16.7 Lpm) PM10 and PM2.5
reference method samplers.
p38 mitogen-activated protein
kinase(s)
polymorphonuclear leukocyte
pyrimidine 5'-nucleotidase
post natal day
particulate organic carbon
polypropylene; pulse pressure
parts per billion
parts per million
post-reinforcement pause
dam stress; prenatal stress;
phosphatidylserine
prostate specific antigen
polysialylated-neural cell
adhesion molecule
proximal tubule
polytetrafluoroethylene
parathyroid hormone
XXXIX
-------
Acronym/Abbreviation
PTHrP
PUFA
PVC
PVD
Q
QRS
QT
QTc
P
pS
R
r
R2
r2
RAAS
RAC2
RBA
RBC
RBP
RD
REA
Ref
RI-RI
RL
220Rn
222Rn
RNA
ROI
ROMO
ROS
RR
Meaning
parathyroid hormone-related
protein
polyunsaturated fatty acid
polyvinyl chloride
peripheral vascular disease
quantile; quartile; quintile
QRS complex in ECG
QT interval in ECG
corrected QT Interval
rho; bulk density; correlation
Pearson's r correlation
coefficient
net drainage loss out of soil
depth of concern; Spearman
correlation coefficient; upward
resuspension flux; correlation
Pearson correlation coefficient
multiple regression correlation
coefficient
correlation coefficient
renin-angiotensin-aldosterone
system
gene encoding for Rac2
relative bioavailability
red blood cell
retinol binding protein
radial diffusivity
Risk/Exposure Assessment
reference (group)
concurrent random interval
repeated learning
radon isotope
stable isotope of radon-222
ribonucleic acid
reactive oxygen
intermediate/superoxide anion;
regions of interest
Rocky Mountain National
Park, Colorado
reactive oxygen species
relative risk; risk ratio
Acronym/Abbreviation
RSL
rtPCR
a
S
SAB
SATs
SBP
SCE
Sena
SD
SDN
SE
Se
sec
SEKI
SEM
SES
Sess.
SFU
SGA
sGC
sGC-pl
SGOT
SGPT
SHBG
SHEDS
SHM
siRNA
SJW
SLAMS
SMC
Meaning
roughness sublayer (transition
layer, wake layer, interfacial
layer)
reverse transcription
polymerase chain reaction
sigma, standard deviation
south; sulfur; synthesis phase
U.S. EPA Science Advisory
Board
Standard Assessment Tests
systolic blood pressure
sister chromatid exchange
a-synuclein
standard deviation
sexually dimorphic nucleus
standard error
selenium
second(s)
Sequoia and Kings Canyon
National Park, California
scanning electron microscopy;
simultaneously extracted metal;
standard error of the mean
socioeconomic status
Session
stacked filter unit(s)
small for gestational age
soluble guanylate cyclase
soluble guanylate cyclase-beta
1
serum glutamic oxaloacetic
transaminase
serum glutamic pyruvic
transaminase
sex hormone binding globulin
Stochastic Human Exposure
and Dose (model)
Stockholm humic model
small interfering RNA
silver jewelry workers
State and Local Air Monitoring
Stations
smooth muscle cells
xl
-------
Acronym/Abbreviation
SNAP-25
SNARE
SNP
SNS
SO
SO2
So
soc
SOD
SOD1
SOF
SOM
SP
SPl,Spl
SPM
SPT
SREBP-2
S. Rep.
SRIXE
StAR
STAT
STATS
STATS
STD.
ST Interval
STN
Syb
Syn
Syt
SZn
Meaning
synaptosomal-associated
protein 25
soluble NSF attachment
receptor
single-nucleotide
polymorphism; sodium
nitroprusside
sympathetic nervous system
stratum oriens
sulfur dioxide
south
superior olivary complex
superoxide dismutase
superoxide dismutase-1
study of osteoporotic fractures
self-organizing map; soil
organic matter
spray painters
specificity protein 1
suspended particulate matter
skin prick test
sterol regulatory element
binding protein-2
Senate Report
synchrotron radiation induced
X-ray emission
steroidogenic acute regulatory
protein
signal transducer and activator
of transcription
signal transducer and activator
of transcription 3
signal transducer and activator
of transcription 5
Standard
measured from the J point to
the end of the T wave in an
ECG
Speciation Trends Network
synaptobrevin
synaptophysin
synaptotagmin
supplemental zinc
Acronym/Abbreviation
T,t
T3,T3
T4,T4
tl/2
TEARS
T cell, T-cell
TE
TEC
TEOM
TF
TFIIIA
Tg
TGF
TGF-P
TGFP1,TGF-P1
TH
THl,Thl
TH2, Th2
Th
TIMP-1
TIMS
TLC
T/LH
TNF
TNP-Ficoll
TNP-OVA
TPR
TS
Meaning
time
triiodothyronine
thyroxine
half-life (-lives); time required
to reduce the initial
concentration by 50%
thioBarbituric acid reactive
substances; thiobarbituric acid-
reactive species
T lymphocyte
trace elements
threshold effect concentrations
tapered element oscillating
microbalance, type of PM
sampler
ratio of the metal concentration
in plant to that in soil;
transferrin
transcription factor IIIA
transgenic
transforming growth factor
P transforming growth factor
pi transforming growth factor
tyrosine hydroxylase
T-derived lymphocyte helper 1
T-derived lymphocyte helper 2
T-helper lymphocyte
tissue inhibitor of
metalloproteinases-1
thermal ionization mass
spectrometry
Treatment of Lead-exposed
Children (study)
testosterone/luteinizing
hormone - measure of Ley dig
cell function
tumor necrosis factor
(e.g., TNF-a)
trinitrophenyl-Ficoll
trinitrophenyl-ovalbumin
total peripheral vascular
resistance
transferrin saturation
xli
-------
Acronym/Abbreviation
TSH
TSP
TSS
TXB2
U
UA
UBL
UCL
UDDS
UDPGT
UIUC
U.K.
U.S.
use
U.S. EPA
USF
USGS
USL
UUDS
UV
UWM
V
V79
VA
VAChAT
VAMP-2
VA-NAS
VDAC
VDR
VGAT
VGCC
Meaning
thyroid stimulating hormone;
total sulfhydryl
total suspended particles
total suspended solids
thromboxane
urbanized area
urban boundary layer
urban canopy layer
urban dynamometer driving
schedule
uridine diphosphate (UDP)-
glucuronosyltransferase(s)
University of Illinois atUrbana
Champaign
United Kingdom
United States of America
U.S. Code
U.S. Environmental Protection
Agency
uptake slope factor
U.S. Geological Survey
urban surface layer
urban dynamic driving
schedule
ultraviolet radiation
Unit World Model
vanadium
Chinese hamster lung cell line
Veterans Administration
vesicular acetylcholine
transporter
vesicle-associated membrane
protein-2
Veterans Administration
Normative Aging Study
voltage-dependent anion
channel
vitamin D receptor
vesicular gamma aminobutyric
acid (GABA) transporter
voltage gated calcium
channel(s)
Acronym/Abbreviation
VGLUT1
VIQ
VLPb
VMAT2
vo43-
VOC(s)
VS., V.
VSCC
VSMC
WACAP
WBC
WCST
WHAM
WHO
WIAT
WINS
wise
WISC-R
wk
WML
WPPSI-III
WPPSI-R
WRAT
W/S
WT
wt.
XAFS
XANES
XDH
Meaning
vesicular glutamate transporter
1
verbal intelligence quotient
(IQ)
very low Pb
vesicular monoamine
transporter-2
vanadate ion
volatile organic compound(s)
versus
very sharp cut cyclone
vascular smooth muscle cells
Western Airborne
Contaminants Assessment
Project
white blood cell
Wisconsin Card Sorting Test
Windermere humic aqueous
model
World Health Organization
Wechsler Individual
Achievement Test
well impactor ninety six
Wechsler Intelligence Scale for
Children
Wechsler Intelligence Scale for
Children-Revised
week(s)
white matter lesions
Wechsler Preschool and
Primary Scales of Intelligence-
Ill
Wechsler Preschool and
Primary Scale of Intelligence-
Revised
Wide Range Achievement Test
winter/summer
wild type
weight
X-ray absorption fine structure
X-ray absorption near edge
structure
xanthine dehydrogenase
xlii
-------
Acronym/Abbreviation Meaning Acronym/Abbreviation Meaning
Xy observed hourly concentrations Zn zinc
for time period i at site j 7+ . .
Zn zinc ion
Xjt observed hourly concentrations
for time penodi at site k ZPP zirconium-potassium
perchlorate; zinc
XPS X-ray photoelectron protoporphyrin
spectroscopy , ,
Z-score standard score
XRF X-ray fluorescence
yr year(s)
xliii
-------
Preamble
Process of ISA Development
This preamble outlines the general process for developing an Integrated Science
Assessment (ISA) including the framework for evaluating weight of evidence and
drawing scientific conclusions and causal judgments. The ISA provides a concise review,
synthesis, and evaluation of the most policy-relevant science to serve as a scientific
foundation for the review of the National Ambient Air Quality Standards (NAAQS). The
general process for NAAQS reviews is described at
http://www.epa.gov/ttn/naaqs/review.html. Figure I depicts the general NAAQS review
process and information for individual NAAQS reviews is available at
www.epa.gov/ttn/naaqs. This preamble is a general discussion of the basic steps and
criteria used in developing an ISA; for each ISA, specific details and considerations are
included in the introductory section for that assessment.
The fundamental process for developing an ISA includes:
• literature searches;
• study selection;
• evaluation and integration of the evidence; and
• development of scientific conclusions and causal judgments.
An initial step in this process is publication of a call for information in the Federal
Register that invites the public to provide information relevant to the assessment, such as
new or recent publications on health or welfare1 effects of the pollutant, or from
atmospheric and exposure sciences fields. The U.S. Environmental Protection Agency
(EPA) maintains an ongoing literature search process for identification of relevant
scientific studies published since the last review of the NAAQS. Search strategies are
designed for pollutants and scientific disciplines and iteratively modified to optimize
identification of pertinent publications. Papers are identified for inclusion in several
additional ways: specialized searches on specific topics; independent review of tables of
contents for journals in which relevant papers may be published; independent
identification of relevant literature by expert scientists; review of citations in previous
assessments and identification by the public and the Clean Air Scientific Advisory
Committee (CASAC) during the external review process. This literature search and study
1 Welfare effects as defined in Clean Air Act (CAA) Section 302(h) [42 U.S.C. 7602(h)] include, but are not limited
to, "effects on soils, water, crops, vegetation, man-made materials, animals, wildlife, weather, visibility and climate,
damage to and deterioration of property, and hazards to transportation, as well as effects on economic values and on
personal comfort and well-being."
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selection process is depicted in Figure II. Publications considered for inclusion in the ISA
are added to the Health and Environmental Research Online (HERO) database developed
by EPA (http://hero.epa.gov/); the references in the ISA include a hyperlink to the
database.
Studies that have undergone scientific peer review and have been published or accepted
for publication and reports that have undergone review are considered for inclusion in the
ISA. Analyses conducted by EPA using publicly available data are also considered for
inclusion in the ISA. All relevant epidemiologic, controlled human exposure,
toxicological, and ecological and welfare effects studies published since the last review
are considered, including those related to exposure-response relationships, mode(s) of
action (MOA), and potentially at-risk populations and lifestages. Studies on atmospheric
chemistry, environmental fate and transport, dosimetry, toxicokinetics and exposure are
also considered for inclusion in the document, as well as analyses of air quality and
emissions data. References that were considered for inclusion in a specific ISA can be
found using the HERO website (http://hero.epa.gov).
EPA .
proposed
decisions on
. standards •
Integrated Review Plan (IRP): timeline and key
policy-relevant issues and scientific questions
Integrated Science Assessment (ISA): evaluation and
synthesis of most policy-relevant studies
Risk/Exposure Assessment (REA):
quantitative assessment, as warranted, focused
on key results, observations, and uncertainties
Policy Assessment (PA): staff analysis of
policy options based on integration and
interpretation of information in Ihe ISA and REA
r
Public hearings
and comments
on proposal
Agency decision
making and draft
final notice
Interagency
review
Clean Air Scientific
Advisory Committee
(CASAC) review
Public comment
Figure I
Illustration of the key steps in the process of the review of
National Ambient Air Quality Standards.
xlv
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Literature
Search
Strategies
Recommendations
during Peer Review
Figure II
Citations from
Past Assessments
Criteriafor study evaluation
include:
•Are the study populations, subjects,
or animal models adequately
selected, and are they sufficiently
well defined to allow for meaningful
comparisons between study or
exposure groups?
•Are the statistical analyses
appropriate, properly performed, and
properly interpreted? Are likely
covariates adequately controlled or
taken into account in the study
design and statistical analysis?
•Are the air quality data, exposure,
or dose metrics of adequate quality
and sufficiently representative of
information regarding ambient
conditions?
•Are the health, ecological or welfare
effect measurements meaningful,
valid and reliable?
•Do the analytical methods provide
adequate sensitivity and precision to
support conclusions?
Illustration of processes for literature search and study selection
used for development of ISAs.
Each ISA builds upon the conclusions of previous assessments for the pollutant under
review. EPA focuses on peer reviewed literature published following the completion of
the previous review and on any new interpretations of previous literature, integrating the
results of recent scientific studies with previous findings. Important earlier studies may
be discussed in detail to reinforce key concepts and conclusions or for reinterpretation in
light of newer data. Earlier studies also are the primary focus in some areas of the
document where research efforts have subsided, or if these earlier studies remain the
definitive works available in the literature.
Selection of studies for inclusion in the ISA is based on the general scientific quality of
the study, and consideration of the extent to which the study is informative and policy-
relevant. Policy-relevant and informative studies include those that provide a basis for or
describe the relationship between the criteria pollutant and effects, including studies that
offer innovation in method or design and studies that reduce uncertainty on critical issues,
such as analyses of confounding or effect modification by copollutants or other variables,
analyses of concentration-response or dose-response relationships, or analyses related to
time between exposure and response. Emphasis is placed on studies that examine effects
associated with pollutant concentrations relevant to current population and ecosystem
exposures, and particularly those pertaining to concentrations currently found in ambient
xlvi
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air. Other studies are included if they contain unique data, such as a previously
unreported effect or MOA for an observed effect, or examine multiple concentrations to
elucidate exposure-response relationships. In general, in assessing the scientific quality
and relevance of health and welfare effects studies, the following considerations have
been taken into account when selecting studies for inclusion in the ISA.
• Are the study populations, subjects, or animal models adequately selected, and
are they sufficiently well defined to allow for meaningful comparisons between
study or exposure groups?
• Are the statistical analyses appropriate, properly performed, and properly
interpreted? Are likely covariates adequately controlled or taken into account in
the study design and statistical analysis?
• Are the air quality data, exposure, or dose metrics of adequate quality and
sufficiently representative of information regarding ambient conditions?
• Are the health, ecological or welfare effect measurements meaningful, valid and
reliable?
• Do the analytical methods provide adequate sensitivity and precision to support
conclusions?
Additional considerations are specific to particular disciplines. In selecting epidemiologic
studies, EPA considers whether a given study: (1) presents information on associations
with short- or long-term pollutant exposures at or near conditions relevant to ambient
exposures; (2) addresses potential confounding by other pollutants; (3) assesses potential
effect modifiers; (4) evaluates health endpoints and populations not previously
extensively researched; and (5) evaluates important methodological issues related to
interpretation of the health evidence (e.g., lag or time period between exposure and
effects, model specifications, thresholds, mortality displacement).
Considerations for the selection of research evaluating controlled human exposure or
animal toxicological studies include a focus on studies conducted using relevant pollutant
exposures. For both types of studies, relevant pollutant exposures are considered to be
those generally within one or two orders of magnitude of ambient concentrations. Studies
in which higher doses were used may also be considered if they provide information
relevant to understanding MOA or mechanisms, as noted below.
Evaluation of controlled human exposure studies focuses on those that approximated
expected human exposure conditions in terms of concentration and duration. Studies
should include control exposures to filtered air, as appropriate. In the selection of
controlled human exposure studies, emphasis is placed on studies that: (1) investigate
potentially at-risk populations and lifestages such as people with asthma or
cardiovascular diseases, children or older adults; (2) address issues such as concentration-
xlvii
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response or time-course of responses; and (3) have sufficient statistical power to assess
findings.
Review of the animal toxicological evidence focuses on studies that approximate
expected human dose conditions, which vary depending on the dosimetry, toxicokinetics
and biological sensitivity of the particular laboratory animal species or strains studied.
Emphasis is placed on studies that: (1) investigate animal models of disease that can
provide information on populations potentially at increased risk of effects; (2) address
issues such as concentration-response or time-course of responses; and (3) have sufficient
statistical power to assess findings. Due to resource constraints on exposure duration and
numbers of animals tested, animal studies typically utilize high-concentration exposures
to acquire data relating to mechanisms and assure a measurable response. Emphasis is
placed on studies using doses or concentrations generally within 1-2 orders of magnitude
of current levels. Studies with higher concentration exposures or doses are considered to
the extent that they provide useful information to inform understanding of interspecies
differences and differences between healthy and at-risk human populations. Results from
in vitro studies may also be included if they provide mechanistic insight or further
support for results demonstrated in vivo.
These criteria provide benchmarks for evaluating various studies and for focusing on the
policy-relevant studies in assessing the body of health, ecological and welfare effects
evidence. As stated initially, the intent of the ISA is to provide a concise review,
synthesis, and evaluation of the most policy-relevant science to serve as a scientific
foundation for the review of the NAAQS, not extensive summaries of all health,
ecological and welfare effects studies for a pollutant. Of most relevance for inclusion of
studies is whether they provide useful qualitative or quantitative information on
exposure-effect or exposure-response relationships for effects associated with pollutant
exposures at doses or concentrations relevant to ambient conditions that can inform
decisions on whether to retain or revise the standards.
The general process for ISA development is illustrated in Figure III. In developing an
ISA, EPA reviews and summarizes the evidence from: studies of atmospheric sciences;
human exposure, animal toxicological, controlled human exposure and epidemiologic
studies; and studies of ecological and welfare effects. In the process of developing the
first draft ISA, EPA may convene a peer input meeting in which the scientific content of
preliminary draft materials is reviewed to ensure that the ISA is up to date and is focused
on the most policy-relevant findings, and to assist EPA with integration of evidence
within and across disciplines. EPA integrates the evidence from across scientific
disciplines or study types and characterizes the weight of evidence for relationships
between the pollutant and various outcomes. The integration of evidence on health, and
xlviii
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ecological or welfare effects, involves collaboration between scientists from various
disciplines. As an example, an evaluation of health effects evidence would include the
integration of the results from epidemiologic, controlled human exposure, and
toxicological studies, and application of the causal framework (described below) to draw
conclusions. Integration of results on health or ecological effects that are logically or
mechanistically connected (e.g., a spectrum of effects on the respiratory system) informs
judgments of causality. Using the causal framework described in the following section,
EPA scientists consider aspects such as strength, consistency, coherence, and biological
plausibility of the evidence, and develop causality determinations on the nature of the
relationships. Causality determinations often entail an iterative process of review and
evaluation of the evidence. Two drafts of the ISA are typically released for review by the
CAS AC and the public, and comments received on the characterization of the science as
well as the implementation of the causal framework are carefully considered in revising
and completing the final ISA.
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Integrated Science Assessment Development Process
Literature Search and
Study Selection
(See Figure 2)
Characterization of Evidence
Develop initial sections or "building blocks" of scientific evidence for assessment:
review and summarize new study results, by outcome/effect category and
discipline, for example, toxicological studies of lung function. Summarize findings
and conclusions from previous assessment. As appropriate, develop initial
conclusions about the available evidence.
Peer Input Consultation
Review of initial draft materials for scientific quality of "building blocks" from
scientists from both outside and within EPA; also includes discussion of results to
facilitate integration of findings. Structure varies; may be public meeting or public
teleconference organized by contractor.
Evaluation, Synthesis, and Integration of Evidence
Integrate evidence from scientific disciplines or study types - for example,
toxicological, controlled human exposure and epidemiologic study findings for
particular health outcome. Evaluate evidence for related groups of endpoints or
outcomes to draw conclusions regarding health or welfare effect categories.
Development of Conclusions and Causal Determinations
Evaluate weight of evidence and develop judgments regarding causality for health
or welfare effect categories, integrating health or welfare effects evidence with
information on mode of action and exposure assessment. Develop conclusions
regarding concentration- or dose-response relationships, potentially at-risk
populations or ecosystems.
Draft Integrated Science Assessment
Final Integrated Science Assessment
Clean Air Scientific Advisory Committee
Review in public meeting; anticipated review of two
drafts of ISA
Public Comments
Comments on draft ISA solicited by EPA
Figure
Characterization of the general process of ISA development.
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EPA Framework for Causal Determination
EPA has developed a consistent and transparent basis for integration of scientific
evidence and evaluation of the causal nature of air pollution-related health or welfare
effects for use in developing ISAs. The framework described below establishes uniform
language concerning causality and brings more specificity to the findings. This
standardized language was drawn from sources across the federal government and wider
scientific community, especially the National Academy of Sciences (NAS) Institute of
Medicine (IOM) document, Improving the Presumptive Disability Decision-Making
Process for Veterans (2008). a comprehensive report on evaluating causality. This
framework:
• describes the kinds of scientific evidence used in establishing a general causal
relationship between exposure and health effects;
• characterizes the process for integration and evaluation of evidence necessary to
reach a conclusion about the existence of a causal relationship;
• identifies issues and approaches related to uncertainty; and
• provides a framework for classifying and characterizing the weight of evidence
in support of a general causal relationship.
Approaches to assessing the separate and combined lines of evidence
(e.g., epidemiologic, controlled human exposure, and animal toxicological studies) have
been formulated by a number of regulatory and science agencies, including the IOM of
the NAS (2008). International Agency for Research on Cancer (IARC) (2006b),
U.S. EPA (2005c). and Centers for Disease Control and Prevention (CDC) (2004). Causal
inference criteria have also been described for ecological effects evidence (U.S. EPA.
1998; Fox. 1991). These formalized approaches offer guidance for assessing causality.
The frameworks are similar in nature, although adapted to different purposes, and have
proven effective in providing a uniform structure and language for causal determinations.
Evaluating Evidence for Inferring Causation
The 1964 Surgeon General's report defined "cause" as a "significant, effectual
relationship between an agent and an associated disorder or disease in the host" (HEW.
1964). More generally, a cause is defined as an agent that brings about an effect or a
result. An association is the statistical relationship among variables; alone, however, it is
insufficient proof of a causal relationship between an exposure and a health outcome.
Unlike an association, a causal claim supports the creation of counterfactual claims, that
is, a claim about what the world would have been like under different or changed
circumstances [Samet and Bodurow, eds. (IOM. 2008)].
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Many of the health and environmental outcomes reported in these studies have complex
etiologies. Diseases such as asthma, coronary heart disease (CHD) or cancer are typically
initiated by multiple agents. Outcomes depend on a variety of factors, such as age,
genetic susceptibility, nutritional status, immune competence, and social factors (IOM.
2008; Gee and Payne-Sturges. 2004). Effects on ecosystems are often also multifactorial
with a complex web of causation. Further, exposure to a combination of agents could
cause synergistic or antagonistic effects. Thus, the observed risk may represent the net
effect of many actions and counteractions.
Scientific findings incorporate uncertainty. "Uncertainty" can be defined as having
limited knowledge to exactly describe an existing state or future outcome, e.g., the lack of
knowledge about the correct value for a specific measure or estimate. Uncertainty
analysis may be qualitative or quantitative in nature. In many cases, the analysis is
qualitative, and can include professional judgment or inferences based on analogy with
similar situations. Quantitative uncertainty analysis may include use of simple measures
(e.g., ranges) and analytical techniques. Quantitative uncertainty analysis might progress
to more complex measures and techniques, if needed for decision support. Various
approaches to evaluating uncertainty include classical statistical methods, sensitivity
analysis, or probabilistic uncertainty analysis, in order of increasing complexity and data
requirements. However, data may not be available for all aspects of an assessment and
those data that are available may be of questionable or unknown quality. Ultimately, the
assessment is based on a number of assumptions with varying degrees of uncertainty. The
ISA generally evaluates uncertainties qualitatively in assessing the evidence from across
studies; in some situations quantitative analysis approaches, such as meta-regression, may
be used.
Publication bias is a source of uncertainty regarding the magnitude of health risk
estimates. It is well understood that studies reporting non-null findings are more likely to
be published than reports of null findings. Publication bias can result in overestimation of
effect estimate sizes (loannidis, 2008). For example, effect estimates from single-city
epidemiologic studies have been found to be generally larger than those from multicity
studies which is an indication of publication bias in that null or negative single-city
results may be reported in a multicity analyses but might not be published independently
(Bell et al.. 2005).
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Consideration of Evidence from Scientific Disciplines
Moving from association to causation involves the elimination of alternative explanations
for the association. The ISA focuses on evaluation of the findings from the body of
evidence, drawing upon the results of all studies determined to meet the criteria described
previously. Causality determinations are based on the evaluation, integration, and
synthesis of evidence from across scientific disciplines. The relative importance of
different types of evidence varies by pollutant or assessment, as does the availability of
different types of evidence for causality determination. Three general types of studies
inform consideration of human health effects: controlled human exposure, epidemiologic
and toxicological studies. Evidence on ecological or welfare effects may be drawn from a
variety of experimental approaches (e.g., greenhouse, laboratory, and field) and
numerous disciplines (e.g., community ecology, biogeochemistry and
paleontological/historical reconstructions).
Direct evidence of a relationship between pollutant exposures and human health effects
comes from controlled human exposure studies. Such studies experimentally evaluate the
health effects of administered exposures in human volunteers under highly controlled
laboratory conditions. Also referred to as human clinical studies, these experiments allow
investigators to expose subjects to known concentrations of air pollutants under carefully
regulated environmental conditions and activity levels. These studies provide important
information on the biological plausibility of associations observed in epidemiologic
studies. In some instances, controlled human exposure studies can also be used to
characterize concentration-response relationships at pollutant concentrations relevant to
ambient conditions. Controlled human exposures are typically conducted using a
randomized crossover design, with subjects exposed both to the pollutant and a clean air
control. In this way, subjects serve as their own experimental controls, effectively
limiting the variance associated with many potential confounders. Considerations for
evaluating controlled human study findings include the generally small sample size and
short exposure time used in experimental studies, and that severe health outcomes are not
assessed. By experimental design, controlled human exposure studies are structured to
evaluate physiological or biomolecular outcomes in response to exposure to a specific air
pollutant and/or combination of pollutants. In addition, the study design generally
precludes inclusion of subjects with serious health conditions, and therefore the results
often cannot be generalized to an entire population. Although some controlled human
exposure studies have included health-compromised individuals such as those with
respiratory or cardiovascular disease, these individuals may also be relatively healthy and
may not represent the most sensitive individuals in the population. Thus, observed effects
in these studies may underestimate the response in certain populations. In addition, the
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study design is limited to exposures and endpoints that are not expected to result in
severe health outcomes.
Epidemiologic studies provide important information on the associations between health
effects and exposure of human populations to ambient air pollution. In epidemiologic or
observational studies of humans, the investigator generally does not control exposures or
intervene with the study population. Broadly, observational studies can describe
associations between exposures and effects. These studies fall into several categories:
e.g., cross-sectional, prospective cohort, panel, and time-series studies. Cross-sectional
studies generally use health outcome, exposure and covariate data available at the
community level (e.g., annual mortality rates and pollutant concentrations), but do not
have individual-level data. Prospective cohort studies have some data collected at the
individual level, generally health outcome data, and in some cases individual-level data
on exposure and covariates are collected. Time-series studies evaluate the relationship for
changes in a health outcome with changes in exposure indicators, such as an association
between daily changes in mortality with air pollution. Panel studies include repeated
measurements of health outcomes, such as respiratory symptoms or heart rhythm
variability, at the individual level. "Natural experiments" offer the opportunity to
investigate changes in health related to a change in exposure, such as closure of a
pollution source.
In evaluating epidemiologic studies, consideration of many study design factors and
issues must be taken into account to properly inform their interpretation. One key
consideration is evaluation of the potential contribution of the pollutant to a health
outcome when it is a component of a complex air pollutant mixture. Reported effect
estimates in epidemiologic studies may reflect (1) independent effects on health
outcomes; (2) effects of the pollutant acting as an indicator of a copollutant or a complex
ambient air pollution mixture; and (3) effects resulting from interactions between that
pollutant and copollutants.
In the evaluation of epidemiologic evidence, one important consideration is potential
confounding. Confounding is "... a confusion of effects. Specifically, the apparent effect
of the exposure of interest is distorted because the effect of an extraneous factor is
mistaken for or mixed with the actual exposure effect (which may be null)" (Rothman
and Greenland. 1998). One approach to remove spurious associations due to possible
confounders is to control for characteristics that may differ between exposed and
unexposed persons; this is frequently termed "adjustment." Scientific judgment is needed
to evaluate likely sources and extent of confounding, together with consideration of how
well the existing constellation of study designs, results, and analyses address the potential
for erroneous inferences. A confounder is associated with both the exposure and the
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effect; for example, confounding can occur between correlated pollutants that are
associated with the same effect.
Several statistical methods are available to detect and control for potential confounders;
however, none of these methods is completely satisfactory. Multivariable regression
models constitute one tool for estimating the association between exposure and outcome
after adjusting for characteristics of participants that might confound the results. The use
of multipollutant regression models has been the prevailing approach for controlling
potential confounding by copollutants in air pollution health effects studies. Finding the
likely causal pollutant from multipollutant regression models is made difficult by the
possibility that one or more air pollutants may be acting as a surrogate for an unmeasured
or poorly measured pollutant or for a particular mixture of pollutants. In addition,
pollutants may independently exert effects on the same system; for example, several
pollutants may be associated with respiratory effect through either the same or different
modes of action. The number and degree of diversity of covariates, as well as their
relevance to the potential confounders, remain matters of scientific judgment. Despite
these limitations, the use of multipollutant models is still the prevailing approach
employed in most air pollution epidemiologic studies and provides some insight into the
potential for confounding or interaction among pollutants.
Confidence that unmeasured confounders are not producing the findings is increased
when multiple studies are conducted in various settings using different subjects or
exposures, each of which might eliminate another source of confounding from
consideration. For example, multicity studies can provide insight on potential
confounding through the use of a consistent method to analyze data from across locations
with different levels of copollutants and other covariates. Intervention studies, because of
their quasi-experimental nature, can be particularly useful in characterizing causation.
Another important consideration in the evaluation of epidemiologic evidence is effect
modification, which occurs when the effect differs between subgroups or strata; for
example, effect estimates that vary by age group or potential risk factor. As stated by
Rothman and Greenland (1998):
"Effect-measure modification differs from confounding in several ways. The
main difference is that, whereas confounding is a bias that the investigator hopes
to prevent or remove from the effect estimate, effect-measure modification is a
property of the effect under study ... In epidemiologic analysis one tries to
eliminate confounding but one tries to detect and estimate effect-measure
modification."
When a risk factor is a confounder, it is the true cause of the association observed
between the exposure and the outcome; when a risk factor is an effect modifier, it
changes the magnitude of the association between the exposure and the outcome in
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stratified analyses. For example, the presence of a preexisting disease or indicator of low
socioeconomic status may act as effect modifiers if they are associated with increased
risk of effects related to air pollution exposure. It is often possible to stratify the
relationship between health outcome and exposure by one or more of these potential
effect modifiers. For variables that modify the association, effect estimates in each
stratum will be different from one another and different from the overall estimate,
indicating a different exposure-response relationship may exist in populations represented
by these variables.
Exposure measurement error, which refers to the uncertainty associated with the exposure
metrics used to represent exposure of an individual or population, can be an important
contributor to uncertainty in air pollution epidemiologic study results. Exposure error can
influence observed epidemiologic associations between ambient pollutant concentrations
and health outcomes by biasing effect estimates toward or away from the null and
widening confidence intervals around those estimates (Zeger et al.. 2000). There are
several components that contribute to exposure measurement error in air pollution
epidemiologic studies, including the difference between true and measured ambient
concentrations, the difference between average personal exposure to ambient pollutants
and ambient concentrations at central monitoring sites, and the use of average population
exposure rather than individual exposure estimates. Factors that could influence exposure
estimates include nonambient sources of exposure, topography of the natural and built
environment, meteorology, measurement errors, time-location-activity patterns, and the
extent to which ambient pollutants penetrate indoor environments. The importance of
exposure error varies with study design and is dependent on the spatial and temporal
aspects of the design.
The third main type of health effects evidence, animal toxicological studies, provides
information on the pollutant's biological action under controlled and monitored exposure
circumstances. Taking into account physiological differences of the experimental species
from humans, these studies inform characterization of health effects of concern,
exposure-response relationships and MOAs. Further, animal models can inform
determinations of at-risk populations. These studies evaluate the effects of exposures to a
variety of pollutants in a highly controlled laboratory setting and allow exploration of
toxicological pathways or mechanisms by which a pollutant may cause effects.
Understanding the biological mechanisms underlying various health outcomes can prove
crucial in establishing or negating causality. In the absence of human studies data,
extensive, well-conducted animal toxicological studies can support determinations of
causality, if the evidence base indicates that similar responses are expected in humans
under ambient exposure conditions.
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Interpretations of animal toxicological studies are affected by limitations associated with
extrapolation between animal and human responses. The differences between humans
and other species have to be taken into consideration, including metabolism, hormonal
regulation, breathing pattern, and differences in lung structure and anatomy. Also, in spite
of a high degree of homology and the existence of a high percentage of orthologous
genes across humans and rodents (particularly mice), extrapolation of molecular
alterations at the gene level is complicated by species-specific differences in
transcriptional regulation. Given these differences, there are uncertainties associated with
quantitative extrapolations of observed pollutant-induced pathophysiological alterations
between laboratory animals and humans, as those alterations are under the control of
widely varying biochemical, endocrine, and neuronal factors.
For ecological effects assessment, both laboratory and field studies (including field
experiments and observational studies) can provide useful data for causality
determination. Because conditions can be controlled in laboratory studies, responses may
be less variable and smaller differences may be easier to detect. However, the control
conditions may limit the range of responses (e.g., animals may not be able to seek
alternative food sources) or incompletely reflect pollutant bioavailability, so they may not
reflect responses that would occur in the natural environment. In addition, larger-scale
processes are difficult to reproduce in the laboratory.
Field observational studies measure biological changes in uncontrolled situations, and
describe an association between a disturbance and an ecological effect. Field data can
provide important information for assessments of multiple stressors or where site-specific
factors significantly influence exposure. They are also often useful for analyses of larger
geographic scales and higher levels of biological organization. However, because
conditions are not controlled, variability is expected to be higher and differences harder
to detect. Field surveys are most useful for linking stressors with effects when stressor
and effect levels are measured concurrently. The presence of confounding factors can
make it difficult to attribute observed effects to specific stressors.
Some studies are considered "intermediate" and are categorized as being between
laboratory and field are studies. Some use environmental media collected from the field
to examine the responses in the laboratory. Others are experiments that are performed in
the natural environment while controlling for some, but not all, of the environmental
conditions (i.e., mesocosm studies). This type of study in manipulated natural
environments can be considered a hybrid between a field experiment and laboratory study
since some aspects are performed under controlled conditions but others are not. They
make it possible to observe community and/or ecosystem dynamics, and provide strong
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evidence for causality when combined with findings of studies that have been made
under more controlled conditions.
Application of Framework for Causal Determination
In its evaluation and integration of the scientific evidence on health or welfare effects of
criteria pollutants, EPA determines the weight of evidence in support of causation and
characterizes the strength of any resulting causal classification. EPA also evaluates the
quantitative evidence and draws scientific conclusions, to the extent possible, regarding
the concentration-response relationships and the loads to ecosystems, exposures, doses or
concentrations, exposure duration, and pattern of exposures at which effects are observed.
To aid judgment, various "aspects"1 of causality have been discussed by many
philosophers and scientists. The 1964 Surgeon General's report on tobacco smoking
discussed criteria for the evaluation of epidemiologic studies, focusing on consistency,
strength, specificity, temporal relationship, and coherence (HEW. 1964). Sir Austin
Bradford Hill (Hill. 1965) articulated aspects of causality in epidemiology and public
health that have been widely used (TOM. 2008; IARC. 2006b: U.S. EPA. 2005c: CDC.
2004). These aspects (Hill. 1965) have been modified (Table I) for use in causal
determinations specific to health and welfare effects for pollutant exposures (U.S. EPA.
2009a).2 Although these aspects provide a framework for assessing the evidence, they do
not lend themselves to being considered in terms of simple formulas or fixed rules of
evidence leading to conclusions about causality (Hill. 1965). For example, one cannot
simply count the number of studies reporting statistically significant results or
statistically nonsignificant results and reach credible conclusions about the relative
weight of the evidence and the likelihood of causality. Rather, these aspects provide a
framework for systematic appraisal of the body of evidence, informed by peer and public
comment and advice, which includes weighing alternative views on controversial issues.
In addition, it is important to note that the aspects in Table I cannot be used as a strict
checklist, but rather to determine the weight of the evidence for inferring causality. In
particular, not meeting one or more of the principles does not automatically preclude a
determination of causality [see discussion in (CDC. 2004)].
1 The "aspects" described by Sir Austin Bradford Hill (Hill. 1965) have become, in the subsequent literature, more
commonly described as "criteria." The original term "aspects" is used here to avoid confusion with "criteria" as it is
used, with different meaning, in the Clean Air Act.
2 The Hill aspects were developed for interpretation of epidemiologic results. They have been modified here for use
with a broader array of data, i.e., epidemiologic, controlled human exposure, ecological, and animal toxicological
studies, as well as in vitro data, and to be more consistent with EPA's Guidelines for Carcinogen Risk Assessment.
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Table I
Aspects to aid in judging causality.
Aspect
Description
An inference of causality is strengthened when a pattern of elevated risks is observed
across several independent studies. The reproducibility of findings constitutes one of the
strongest arguments for causality. If there are discordant results among investigations,
possible reasons such as differences in exposure, confounding factors, and the power of
the study are considered.
Consistency of the
observed association
An inference of causality from one line of evidence (e.g., epidemiologic, clinical, or
animal studies) may be strengthened by other lines of evidence that support a
cause-and-effect interpretation of the association. Evidence on ecological or welfare
effects may be drawn from a variety of experimental approaches (e.g., greenhouse,
Coherence laboratory, and field) and subdisciplines of ecology (e.g., community ecology,
biogeochemistry, and paleontological/historical reconstructions). The coherence of
evidence from various fields greatly adds to the strength of an inference of causality. In
addition, there may be coherence in demonstrating effects across multiple study designs
or related health endpoints within one scientific line of evidence.
An inference of causality tends to be strengthened by consistency with data from
experimental studies or other sources demonstrating plausible biological mechanisms. A
proposed mechanistic linking between an effect and exposure to the agent is an
important source of support for causality, especially when data establishing the existence
and functioning of those mechanistic links are available.
Biological plausibility.
Biological gradient
(exposure-response
relationship)
A well-characterized exposure-response relationship (e.g., increasing effects associated
with greater exposure) strongly suggests cause and effect, especially when such
relationships are also observed for duration of exposure (e.g., increasing effects
observed following longer exposure times).
Strength of the observed
association
Experimental evidence
Temporal relationship of
the observed association
The finding of large, precise risks increases confidence that the association is not likely
due to chance, bias, or other factors. However, it is noted that a small magnitude in an
effect estimate may represent a substantial effect in a population.
Strong evidence for causality can be provided through "natural experiments" when a
change in exposure is found to result in a change in occurrence or frequency of health or
welfare effects.
Evidence of a temporal sequence between the introduction of an agent, and appearance
of the effect, constitutes another argument in favor of causality.
Evidence linking a specific outcome to an exposure can provide a strong argument for
causation. However, it must be recognized that rarely, if ever, does exposure to a
pollutant invariably predict the occurrence of an outcome, and that a given outcome may
have multiple causes.
Specificity of the
observed association
Structure activity relationships and information on the agent's structural analogs can
. . provide insight into whether an association is causal. Similarly, information on mode of
^" action for a chemical, as one of many structural analogs, can inform decisions regarding
likely causality.
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Determination of Causality
In the ISA, EPA assesses the body of relevant literature, building upon evidence available
during previous NAAQS reviews, to draw conclusions on the causal relationships
between relevant pollutant exposures and health or environmental effects. ISAs use a
five-level hierarchy that classifies the weight of evidence for causation1. In developing
this hierarchy, EPA has drawn on the work of previous evaluations, most prominently the
lOM's Improving the Presumptive Disability Decision-Making Process for Veterans
[Samet and Bodurow, eds. (TOM. 2008). EPA's Guidelines for Carcinogen Risk
Assessment (U.S. EPA. 2005c). and the U.S. Surgeon General's smoking report (CDC.
2004). This weight of evidence evaluation is based on integration of findings from
various lines of evidence from across the health and environmental effects disciplines.
These separate judgments are integrated into a qualitative statement about the overall
weight of the evidence and causality. The five descriptors for causal determination are
described in Table II.
Determination of causality involves the evaluation and integration of evidence for
different types of health, ecological or welfare effects associated with short- and long-
term exposure periods. In making determinations of causality, evidence is evaluated for
major outcome categories or groups of related endpoints (e.g., respiratory effects,
vegetation growth), integrating evidence from across disciplines, and evaluating the
coherence of evidence across a spectrum of related endpoints to draw conclusions
regarding causality. In discussing the causal determination, EPA characterizes the
evidence on which the judgment is based, including strength of evidence for individual
endpoints within the outcome category or group of related endpoints.
In drawing judgments regarding causality for the criteria air pollutants, the ISA focuses
on evidence of effects in the range of relevant pollutant exposures or doses, and not on
determination of causality at any dose. Emphasis is placed on evidence of effects at doses
(e.g., blood Pb concentration) or exposures (e.g., air concentrations) that are relevant to,
or somewhat above, those currently experienced by the population. The extent to which
studies of higher concentrations are considered varies by pollutant and major outcome
category, but generally includes those with doses or exposures in the range of one to two
orders of magnitude above current or ambient conditions. Studies that use higher doses or
exposures may also be considered to the extent that they provide useful information to
inform understanding of mode of action, interspecies differences, or factors that may
increase risk of effects for a population. Thus, a causality determination is based on
1 The Center for Disease Control (CDC) and IOM frameworks use a four-category hierarchy for the strength of the
evidence. A five-level hierarchy is used here to be consistent with the EPA Guidelines for Carcinogen Risk
Assessment and to provide a more nuanced set of categories.
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weight of evidence evaluation for health, ecological or welfare effects, focusing on the
evidence from exposures or doses generally ranging from current levels to one or two
orders of magnitude above current levels.
In addition, EPA evaluates evidence relevant to understand the quantitative relationships
between pollutant exposures and health, ecological or welfare effects. This includes
evaluation of the form of concentration-response or dose-response relationships and, to
the extent possible, drawing conclusions on the levels at which effects are observed. The
ISA also draws scientific conclusions regarding important exposure conditions for effects
and populations that may be at greater risk for effects, as described in the following
section.
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Table II Weight of evidence for causal determination.
Causal
Determination
Health Effects
Ecological and Welfare Effects
Evidence is sufficient to conclude that there is a
causal relationship with relevant pollutant exposures
(i.e., doses or exposures generally within one to two
orders of magnitude of current levels). That is, the
pollutant has been shown to result in health effects in
studies in which chance, bias, and confounding could
Causal be ruled out with reasonable confidence. For
relationship example: a) controlled human exposure studies that
demonstrate consistent effects; or b) observational
studies that cannot be explained by plausible
alternatives or are supported by other lines of
evidence (e.g., animal studies or mode of action
information). Evidence includes multiple high-quality
studies
Evidence is sufficient to conclude that there is a
causal relationship with relevant pollutant exposures
i.e., doses or exposures generally within one to two
orders of magnitude of current levels). That is, the
pollutant has been shown to result in effects in
studies in which chance, bias, and confounding could
be ruled out with reasonable confidence. Controlled
exposure studies (laboratory or small- to
medium-scale field studies) provide the strongest
evidence for causality, but the scope of inference
may be limited. Generally, determination is based on
multiple studies conducted by multiple research
groups, and evidence that is considered sufficient to
infer a causal relationship is usually obtained from the
joint consideration of many lines of evidence that
reinforce each other.
Likely to be a
causal
relationship
Evidence is sufficient to conclude that a causal
relationship is likely to exist with relevant pollutant
exposures, but important uncertainties remain. That
is, the pollutant has been shown to result in health
effects in studies in which chance and bias can be
ruled out with reasonable confidence but potential
issues remain. For example: a) observational studies
show an association, but copollutant exposures are
difficult to address and/or other lines of evidence
(controlled human exposure, animal, or mode of
action information) are limited or inconsistent; or b)
animal toxicological evidence from multiple studies
from different laboratories that demonstrate effects,
but limited or no human data are available. Evidence
generally includes multiple high-quality studies.
Evidence is sufficient to conclude that there is a likely
causal association with relevant pollutant exposures.
That is, an association has been observed between
the pollutant and the outcome in studies in which
chance, bias, and confounding are minimized, but
uncertainties remain. For example, field studies show
a relationship, but suspected interacting factors
cannot be controlled, and other lines of evidence are
limited or inconsistent. Generally, determination is
based on multiple studies in multiple research
groups.
Evidence is suggestive of a causal relationship with
relevant pollutant exposures, but is limited. For
example, (a) at least one high-quality epidemiologic
Suggestive of a study shows an association with a given health
causal outcome but the results of other studies are
relationship inconsistent; or (b) a well-conducted toxicological
study, such as those conducted in the National
Toxicology Program (NTP), shows effects in animal
species,
Evidence is suggestive of a causal relationship with
relevant pollutant exposures, but chance, bias and
confounding cannot be ruled out. For example, at
least one high-quality study shows an effect, but the
results of other studies are inconsistent.
Inadequate to
infer a causal
relationship
Evidence is inadequate to determine that a causal
relationship exists with relevant pollutant exposures.
The available studies are of insufficient quantity,
quality, consistency, or statistical power to permit a
conclusion regarding the presence or absence of an
effect.
The available studies are of insufficient quality,
consistency, or statistical power to permit a
conclusion regarding the presence or absence of an
effect.
Not likely to be a
causal
relationship
Evidence is suggestive of no causal relationship with
relevant pollutant exposures. Several adequate
studies, covering the full range of levels of exposure
that human beings are known to encounter and
considering at-risk populations, are mutually
consistent in not showing an effect at any level of
exposure.
Several adequate studies, examining relationships
with relevant exposures, are consistent in failing to
show an effect at any level of exposure.
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Quantitative Relationships: Effects on Human Populations
Once a determination is made regarding the causal relationship between the pollutant and
outcome category, important questions regarding quantitative relationships include:
• What is the concentration-response, exposure-response, or dose-response
relationship in the human population?
• What is the interrelationship between incidence and severity of effect?
• What exposure conditions (dose or exposure, duration and pattern) are
important?
• What populations and lifestages appear to be differentially affected (i.e., more at
risk of experiencing effects)?
In order to address these questions, the entirety of quantitative evidence is evaluated to
characterize pollutant concentrations and exposure durations at which effects were
observed for exposed populations, including populations and lifestages potentially at
increased risk. To accomplish this, evidence is considered from multiple and diverse
types of studies, and a study or set of studies that best approximates the concentration-
response relationships between health outcomes and the pollutant may be identified.
Controlled human exposure studies provide the most direct and quantifiable exposure-
response data on the human health effects of pollutant exposures. To the extent available,
the ISA evaluates results from epidemiologic studies that characterize the form of
relationships between the pollutant and health outcomes and draws conclusions on the
shape of these relationships. Animal data may also inform evaluation of
concentration-response relationships, particularly relative to MOAs and characteristics of
at-risk populations.
An important consideration in characterizing the public health impacts associated with
exposure to a pollutant is whether the concentration-response relationship is linear across
the range of concentrations or if nonlinear relationships exist along any part of this range.
The shape of the concentration-response curve at and below the level of the current
standards is of particular interest. Various sources of variability and uncertainty, such as
low data density in the lower concentration range, possible influence of exposure
measurement error, and variability between individuals in susceptibility to air pollution
health effects, tend to smooth and "linearize" the concentration-response function, and
thus can obscure the existence of a threshold or nonlinear relationship. Since individual
thresholds vary from person to person due to individual differences such as genetic level
susceptibility or preexisting disease conditions (and even can vary from one time to
another for a given person), it can be difficult to demonstrate that a threshold exists in a
population study. These sources of variability and uncertainty may explain why the
available human data at ambient concentrations for some environmental pollutants
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(e.g., participate matter [PM], O3, lead [Pb], environmental tobacco smoke [ETS],
radiation) do not exhibit thresholds for cancer or noncancer health effects, even though
likely mechanisms include nonlinear processes for some key events.
Finally, identification of the population groups or lifestages that may be at greater risk of
health effects from air pollutant exposures contributes to an understanding of the public
health impact of pollutant exposures. In the ISA, the term "at-risk population" is used to
encompass populations or lifestages that have a greater likelihood of experiencing health
effects related to exposure to an air pollutant due to a variety of factors; other terms used
in the literature include susceptible, vulnerable, and sensitive. These factors may be
intrinsic, such as genetic or developmental factors, race, gender, lifestage, or the presence
of preexisting diseases, or they may be extrinsic, such as socioeconomic status (SES),
activity pattern and exercise level, reduced access to health care, low educational
attainment, or increased pollutant exposures (e.g., near roadways). Epidemiologic studies
can help identify populations potentially at increased risk of effects by evaluating health
responses in the study population. Examples include testing for interactions or effect
modification by factors such as gender, age group, or health status. Experimental studies
using animal models of susceptibility or disease can also inform the extent to which
health risks are likely greater in specific population groups.
Quantitative Relationships: Effects on Ecosystems or Public
Welfare
Key questions for understanding the quantitative relationships between exposure (or
concentration or deposition) to a pollutant and risk to ecosystems or the public welfare
include:
• What elements of the ecosystem (e.g., types, regions, taxonomic groups,
populations, functions, etc.) appear to be affected, or are more sensitive to
effects? Are there differences between locations or materials in welfare effects
responses, such as impaired visibility or materials damage?
• Under what exposure conditions (amount deposited or concentration, duration
and pattern) are effects seen?
• What is the shape of the concentration-response or exposure-response
relationship?
Evaluations of causality generally consider the probability of quantitative changes in
ecological and welfare effects in response to exposure. A challenge to the quantification
of exposure-response relationships for ecological effects is the great regional and local
spatial variability, as well as temporal variability, in ecosystems. Thus, exposure-
response relationships are often determined for a specific ecological system and scale,
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rather than at the national or even regional scale. Quantitative relationships therefore are
estimated site by site and may differ greatly between ecosystems.
Concepts in Evaluating Adversity of Health Effects
In evaluating health evidence, a number of factors can be considered in delineating
between adverse and nonadverse health effects resulting from exposure to air pollution.
Some health outcomes, such as hospitalization for respiratory or cardiovascular diseases,
are clearly considered adverse. It is more difficult to determine the extent of change that
constitutes adversity in more subtle health measures. These include a wide variety of
responses, such as alterations in markers of inflammation or oxidative stress, changes in
pulmonary function or heart rate variability, or alterations in neurocognitive function
measures. The challenge is determining the magnitude of change in these measures when
there is no clear point at which a change becomes adverse. The extent to which a change
in health measure constitutes an adverse health effect may vary between populations.
Some changes that may not be considered adverse in healthy individuals would be
potentially adverse in more at-risk individuals.
The extent to which changes in lung function are adverse has been discussed by the
American Thoracic Society (ATS) in an official statement titled What Constitutes an
Adverse Health Effect of Air Pollution? (ATS. 2000). An air pollution-induced shift in
the population distribution of a given risk factor for a health outcome was viewed as
adverse, even though it may not increase the risk of any one individual to an unacceptable
level. For example, a population of asthmatics could have a distribution of lung function
such that no identifiable individual has a level associated with significant impairment.
Exposure to air pollution could shift the distribution such that no identifiable individual
experiences clinically relevant effects. This shift toward decreased lung function,
however, would be considered adverse because individuals within the population would
have diminished reserve function and therefore would be at increased risk to further
environmental insult. The committee also observed that elevations of biomarkers, such as
cell number and types, cytokines and reactive oxygen species, may signal risk for ongoing
injury and clinical effects or may simply indicate transient responses that can provide
insights into mechanisms of injury, thus illustrating the lack of clear boundaries that
separate adverse from nonadverse effects.
The more subtle health outcomes may be connected mechanistically to health events that
are clearly adverse. For example, air pollution may affect markers of transient myocardial
ischemia such as ST-segment abnormalities or onset of exertional angina. These effects
may not be apparent to the individual, yet may still increase the risk of a number of
cardiac events, including myocardial infarction and sudden death. Thus, small changes in
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physiological measures may not appear to be clearly adverse when considered alone, but
may be a part of a coherent and biologically plausible chain of related health outcomes
that range up to responses that are very clearly adverse, such as hospitalization or
mortality.
Concepts in Evaluating Adversity of Ecological Effects
Adversity of ecological effects can be understood in terms ranging in biological level of
organization; from the cellular level to the individual organism and to the population,
community, and ecosystem levels. In the context of ecology, a population is a group of
individuals of the same species, and a community is an assemblage of populations of
different species interacting with one another that inhabit an area. An ecosystem is the
interactive system formed from all living organisms and their abiotic (physical and
chemical) environment within a given area (IPCC, 2007). The boundaries of what could
be called an ecosystem are somewhat arbitrary, depending on the focus of interest or
study. Thus, the extent of an ecosystem may range from very small spatial scales to,
ultimately, the entire Earth (TPCC. 2007).
Effects on an individual organism are generally not considered to be adverse to public
welfare. However if effects occur to enough individuals within a population, then
communities and ecosystems may be disrupted. Changes to populations, communities,
and ecosystems can in turn result in an alteration of ecosystem processes. Ecosystem
processes are defined as the metabolic functions of ecosystems including energy flow,
elemental cycling, and the production, consumption and decomposition of organic matter
(U.S. EPA. 2002a). Growth, reproduction, and mortality are species-level endpoints that
can be clearly linked to community and ecosystem effects and are considered to be
adverse when negatively affected. Other endpoints such as changes in behavior and
physiological stress can decrease ecological fitness of an organism, but are harder to link
unequivocally to effects at the population, community, and ecosystem level. The degree
to which pollutant exposure is considered adverse may also depend on the location and its
intended use (i.e., city park, commercial, and cropland). Support for consideration of
adversity beyond the species level by making explicit the linkages between stress-related
effects at the species and effects at the ecosystem level is found in A Framework for
Assessing and Reporting on Ecological Condition: an SAB report (U.S. EPA. 2002a).
Additionally, the National Acid Precipitation Assessment Program (NAPAP. 1991) uses
the following working definition of "adverse ecological effects" in the preparation of
reports to Congress mandated by the Clean Air Act: "any injury (i.e., loss of chemical or
physical quality or viability) to any ecological or ecosystem component, up to and
including at the regional level, over both long and short terms."
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On a broader scale, ecosystem services may provide indicators for ecological impacts.
Ecosystem services are the benefits that people obtain from ecosystems (1JNEP. 2003).
According to the Millennium Ecosystem Assessment, ecosystem services include:
"provisioning services such as food and water; regulating services such as regulation of
floods, drought, land degradation, and disease; supporting services such as soil formation
and nutrient cycling; and cultural services such as recreational, spiritual, religious, and
other nonmaterial benefits." For example, a more subtle ecological effect of pollution
exposure may result in a clearly adverse impact on ecosystem services if it results in a
population decline in a species that is recreationally or culturally important.
References for Preamble
ATS (American Thoracic Society). (2000). What constitutes an adverse health effect of air pollution?
This official statement of the American Thoracic Society was adopted by the ATS Board of
Directors, July 1999. Am J Respir Crit Care Med 161: 665-673.
Bell. ML; Dominici. F; Samet. JM. (2005). A meta-analysis of time-series studies of ozone and
mortality with comparison to the national morbidity, mortality, and air pollution study.
Epidemiology 16: 436-445. http://dx.doi.org/10.1097/01.ede.0000165817.40152.85
CDC (Centers for Disease Control and Prevention). (2004). The health consequences of smoking: A
report of the Surgeon General. Washington, DC: U.S. Department of Health and Human Services.
http ://www. surgeongeneral.gov/library/smokingconsequences/
Fox. GA. (1991). Practical causal inference for ecoepidemiologists. JToxicol Environ Health A 33:
359-373. http://dx.doi.org/10.1080/15287399109531535
Gee. GC: Pavne-Sturges. DC. (2004). Environmental health disparities: A framework integrating
psychosocial and environmental concepts [Review]. Environ Health Perspect 112: 1645-1653.
http://dx.doi.org/10.1289/ehp.7074
HEW (U.S. Department of Health, Education and Welfare). (1964). Smoking and health: Report of the
advisory committee to the surgeon general of the public health service. Washington, DC: U.S.
Department of Health, Education, and Welfare.
http://profiles.nlm.nih.gov/ps/retrieve/ResourceMetadata/NNBBMQ
Hill. AB. (1965). The environment and disease: Association or causation? Proc R Soc Med 58: 295-
300.
IARC (International Agency for Research on Cancer). (2006b). Preamble to the IARC monographs.
Lyon, France. http://monographs.iarc.fr/ENG/Preamble/
loannidis. JPA. (2008). Why most discovered true associations are inflated [Review]. Epidemiology
19: 640-648. http://dx.doi.org/10.1097/EDE.Ob013e31818131e7
IOM (Institute of Medicine). (2008). Improving the presumptive disability decision-making process
for veterans. In JM Samet; CC Bodurow (Eds.). Washington, DC: National Academies Press.
http://www.nap.edu/openbook.php7record id=l 1908
IPCC (Intergovernmental Panel on Climate Change). (2007). Climate change 2007: Impacts,
adaptation and vulnerability. Cambridge, UK: Cambridge University Press.
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NAPAP (National Acid Precipitation Assessment Program). (1991). The experience and legacy of
NAPAP: Report of the Oversight Review Board of the National Acid Precipitation Assessment
Program. Washington, DC.
Rothman. KJ; Greenland. S. (1998). Modern epidemiology (2nd ed.). Philadelphia, PA: Lippincott,
Williams, & Wilkins.
U.S. EPA (U.S. Environmental Protection Agency). (1998). Guidelines for ecological risk assessment
[EPA Report]. (EPA/630/R-95/002F). Washington, DC.
http://www.epa.gov/raf/publications/guidelines-ecological-risk-assessment.htm
U.S. EPA (U.S. Environmental Protection Agency). (2002a). A framework for assessing and reporting
on ecological condition: An SAB report [EPA Report]. (EPA-SAB-EPEC-02-009). Washington,
DC. http://www.ntis.gov/search/product.aspx?ABBR=PB2004100741
U.S. EPA (U.S. Environmental Protection Agency). (2005c). Guidelines for carcinogen risk
assessment [EPA Report]. (EPA/630/P-03/001F). Washington, DC.
http://www.epa.gov/cancerguidelines/
U.S. EPA (U.S. Environmental Protection Agency). (2009a). Integrated science assessment for
paniculate matter [EPA Report]. (EPA/600/R-08/139F). Research Triangle Park, NC.
http://cfpub.epa.gov/ncea/cfm/recordisplav.cfm?deid=216546
UNEP (United Nations Environment Programme). (2003). Ecosystems and human well-being: A
framework for assessment. Washington, DC: Island Press.
Zeger. SL; Thomas. D; Dominici. F; Samet. JM; Schwartz. J; Dockery. D; Cohen. A. (2000). Exposure
measurement error in time-series studies of air pollution: Concepts and consequences. Environ
Health Perspect 108: 419-426.
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Legislative and Historical Background
Legislative Requirements for the NAAQS Review
Two sections of the Clean Air Act (CAA) govern the establishment and revision of the
NAAQS. Section 108 (42:U.S.C.:7408) directs the Administrator to identify and list
certain air pollutants and then to issue air quality criteria for those pollutants. The
Administrator is to list those air pollutants that in her "... judgment, cause or contribute
to air pollution which may reasonably be anticipated to endanger public health or
welfare; ..." and, "... the presence of which in the ambient air results from numerous or
diverse mobile or stationary sources;" and, "... for which ... [the Administrator] plans to
issue air quality criteria...." Air quality criteria are intended to "accurately reflect the
latest scientific knowledge useful in indicating the kind and extent of all identifiable
effects on public health or welfare which may be expected from the presence of [a]
pollutant in the ambient air ..." (42:U.S.C.:7408([b]). Section 109 (42:U.S.C.:7409)
directs the Administrator to propose and promulgate "primary" and "secondary" NAAQS
for pollutants for which air quality criteria are issued. Section 109(b)(l) defines a primary
standard as one ".. .the attainment and maintenance of which in the judgment of the
Administrator, based on such criteria and allowing an adequate margin of safety, are
requisite to protect the public health." The legislative history of Section 109 indicates that
a primary standard is to be set at "... the maximum permissible ambient air level ...
which will protect the health of any [sensitive] group of the population," and that for this
purpose "... reference should be made to a representative sample of persons comprising
the sensitive group rather than to a single person in such a group..." (S. Rep. No.
91:1196, 91st Cong., 2d Sess. 10 [1970]). A secondary standard, as defined in Section
109(b)(2), must "... specify a level of air quality the attainment and maintenance of
which, in the judgment of the Administrator, based on such criteria, is requisite to protect
the public welfare from any known or anticipated adverse effects associated with the
presence of [the] pollutant in the ambient air." Welfare effects (as defined in Section
302(h); 42:U.S.C.:7602[h]) include, but are not limited to, "... effects on soils, water,
crops, vegetation, man-made materials, animals, wildlife, weather, visibility and climate,
damage to and deterioration of property, and hazards to transportation, as well as effects
on economic values and on personal comfort and well-being."
The requirement that primary standards provide an adequate margin of safety was
intended to address uncertainties associated with inconclusive scientific and technical
information available at the time of standard setting. It was also intended to provide a
reasonable degree of protection against hazards that research has not yet identified (Lead
Industries Association v. EPA, 647:F.2d: 1130-1154 [D.C.Cir 1980]; American Petroleum
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Institute v. Costle, 665:F.2d: 1176-1186 [D.C.Cir. 1981]; American Farm Bureau
Federation v. EPA, 559:F.3d:512-533 [D.C. Cir. 2009]; Association of Battery Recyclers
v. EPA, 604:F.3d:613, 617-618 [D.C. Cir. 2010]). Both kinds of uncertainties are
components of the risk associated with pollution at levels below those at which human
health effects can be said to occur with reasonable scientific certainty. Thus, in selecting
primary standards that provide an adequate margin of safety, the Administrator is seeking
not only to prevent pollution levels that have been demonstrated to be harmful but also to
prevent lower pollutant levels that may pose an unacceptable risk of harm, even if the risk
is not precisely identified as to nature or degree. The CAA does not require the
Administrator to establish a primary NAAQS at a zero-risk level or at background
concentration levels (LeadIndustries v. EPA, [647:F.2d:at 1156 n.51]), but rather at a
level that reduces risk sufficiently so as to protect public health with an adequate margin
of safety.
In addressing the requirement for an adequate margin of safety, the EPA considers such
factors as the nature and severity of the health effects involved, the size of sensitive
population(s) at risk, and the kind and degree of the uncertainties that must be addressed.
The selection of any particular approach to providing an adequate margin of safety is a
policy choice left specifically to the Administrator's judgment (Lead Industries
Association v. EPA, [647:F.2d: 1161-1162]; Whitman v. American Trucking Associations,
[531:U.S.:457-495(2001)]).
In setting standards that are "requisite" to protect public health and welfare as provided in
Section 109(b), EPA's task is to establish standards that are neither more nor less
stringent than necessary for these purposes. In so doing, EPA may not consider the costs
of implementing the standards (see generally, Whitman v. American Trucking
Associations, [531:U.S.:457, 465-472, 475-476 (2001)]). Likewise, "... Attainability and
technological feasibility are not relevant considerations in the promulgation of national
ambient air quality standards." (American Petroleum Institute v. Costle,
[665:F.2d:1185]).
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Section 109(d)(l) requires that "not later than December 31, 1980, and at 5-year intervals
thereafter, the Administrator shall complete a thorough review of the criteria published
under Section 108 and the national ambient air quality standards ... and shall make such
revisions in such criteria and standards and promulgate such new standards as may be
appropriate ... ." Section 109(d)(2) requires that an independent scientific review
committee "shall complete a review of the criteria ... and the national primary and
secondary ambient air quality standards ... and shall recommend to the Administrator any
new ... standards and revisions of existing criteria and standards as may be appropriate
... ." Since the early 1980's, this independent review function has been performed by the
Clean Air Scientific Advisory Committee (CASAC).
History of the NAAQS for Pb
Unlike pollutants such as PM and carbon monoxide (CO), air quality criteria had not
been issued for Pb as of the enactment of the Clean Air Act of 1970, which first set forth
the requirement to set national ambient air quality standards for criteria pollutants. EPA
did not intend to issue air quality criteria for lead, and accordingly had not listed lead
under Section 108. EPA had determined to control lead air pollution through regulations
to phase-out use of lead additives in gasoline and EPA viewed those controls, and
possibly additional federal controls, as the best approach to controlling lead emissions
(See 41 FR 14921 (April 8, 1976). However, the decision not to list lead under Section
108 was challenged by environmental and public health groups and the U.S. District
Court for the Southern District of New York concluded that EPA was required to list lead
under Section 108. (Natural Resources Defense Council v. EPA, 411 F. Supp. 864
[S.D. N.Y. 1976], aff d, 545 F.2d 320 [2d Cir. 1978]).
Accordingly, on April 8, 1976, EPA published a notice that Pb had been listed under
Section 108 as a criteria pollutant (41 FR 14921) and on October 5, 1978, EPA
promulgated primary and secondary NAAQS for Pb under Section 109 of the Act
(43 FR 46246). Both primary and secondary standards were set at a level of
1.5 micrograms per cubic meter (ug/m3), measured as Pb in total suspended particles (Pb-
TSP), not to be exceeded by the maximum arithmetic mean concentration averaged over
a calendar quarter. These standards were based on the 1977 Pb Air Quality Criteria for
Lead Document (AQCD) (U.S. EPA. 1977).
The first review of the Pb standards was initiated in the mid-1980s. The scientific
assessment for that review is described in the 1986 Pb AQCD (U.S. EPA. 1986a). the
associated Addendum (U.S. EPA. 1986c) and the 1990 Supplement (U.S. EPA. 1990a).
As part of the review, the Agency designed and performed human exposure and health
risk analyses (U.S. EPA. 1989). the results of which were presented in a 1990 Staff Paper
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(U.S. EPA. 1990c). Based on the scientific assessment and the human exposure and
health risk analyses, the 1990 Staff Paper presented recommendations for consideration
by the Administrator (U.S. EPA. 1990c). After consideration of the documents developed
during the review and the significantly changed circumstances since Pb was listed in
1976, the Agency did not propose any revisions to the 1978 Pb NAAQS. In a parallel
effort, the Agency developed the broad, multi-program, multimedia, integrated U.S.
Strategy for Reducing Lead Exposure (U.S. EPA. 1991). As part of implementing this
strategy, the Agency focused efforts primarily on regulatory and 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 actions to reduce Pb
emissions to air, such as bringing more areas into compliance with the existing Pb
NAAQS (U.S. EPA. 1991).
The most recent review of the Pb air quality criteria and standards was initiated in
November, 2004 (69 FR 64926) and the Agency's plans for preparation of the Air
Quality Criteria Document (AQCD) and conduct of the NAAQS review were contained
in two documents: Project Work Plan for Revised Air Quality Criteria for Lead (U.S.
EPA. 2005e); and Plan for Review of the 'National Ambient Air Quality Standards for
Lead (U.S. EPA. 2006f). The schedule for completion of this review was governed by a
judicial order in Missouri Coalition for the Environment v. EPA (No. 4:04CV00660
ERW, Sept. 14, 2005; and amended on April 29, 2008 and July 1, 2008), which specified
a schedule for the review of duration substantially shorter than five years.
The scientific assessment for the review is described in the 2006 Air Quality Criteria for
Lead [2006 Pb AQCD; (U.S. EPA. 2006b)1. multiple drafts of which received review by
CASAC and the public. EPA also conducted human exposure and health risk assessments
and a pilot ecological risk assessment for the review, after consultation with CASAC and
receiving public comment on a draft analysis plan (U.S. EPA. 2006d). Drafts of these
quantitative assessments were reviewed by CASAC and the public. The pilot ecological
risk assessment was released in December 2006 (ICF. 2006) and the final health risk
assessment report was released in November 2007 (U.S. EPA. 2007g). The policy
assessment based on both of these assessments, air quality analyses and key evidence
from the AQCD was presented in the Staff Paper (U.S. EPA. 2006g). a draft of which
also received CASAC and public review. The final Staff Paper presented OAQPS staffs
evaluation of the public health and welfare policy implications of the key studies and
scientific information contained in the 2006 Pb AQCD and presented and interpreted
results from the quantitative risk/exposure analyses conducted for this review. Based on
this evaluation, the Staff Paper presented OAQPS staff recommendations that the
Administrator give consideration to substantially revising the primary and secondary
standards to a range of levels at or below 0.2 (ig/m3.
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Immediately subsequent to completion of the Staff Paper, EPA issued an advance notice
of proposed rulemaking (ANPR) that was signed by the Administrator on December 5,
2007 (72 FR 71488).: CASAC provided advice and recommendations to the
Administrator with regard to the Pb NAAQS based on its review of the ANPR and the
previously released final Staff Paper and risk assessment reports. The proposed decision
on revisions to the Pb NAAQS was signed on May 1, 2008 and published in the Federal
Register on May 20, 2008 (73 FR 29184). Members of the public provided both written
and, at two public hearings, oral comments and the CASAC Pb Panel also provided
advice and recommendations to the Administrator based on its review of the proposal
notice. The final decision on revisions to the Pb NAAQS was signed on October 15, 2008
and published in the Federal Register on November 12, 2008 (73 FR 66964).
The November 2008 notice described EPA's decision to revise the primary and
secondary NAAQS for Pb from a level of 1.5 (ig/m3 to a level of 0.15 (ig/m3. EPA's
decision on the level for the primary standard was based on the much-expanded health
effects evidence on neurocognitive effects of Pb in children. The level of 0.15 ug/m3 was
established to protect against air Pb-related health effects, including intelligence quotient
(IQ) decrements in the most highly exposed children, those exposed at the level of the
standard. Results of the quantitative risk assessment were judged supportive of the
evidence-based framework estimates. The averaging time was revised to a rolling
three-month period with a maximum (not-to-be-exceeded) form, evaluated over a
three-year period. As compared to the previous averaging time of calendar quarter, this
revision was considered to be more scientifically appropriate and more health protective.
The rolling average gives equal weight to all three-month periods, and the new
calculation method gives equal weight to each month within each three-month period.
Further, the rolling average yields 12 three-month averages each year to be compared to
the NAAQS versus four averages in each year for the block calendar quarters pertaining
to the previous standard. The indicator of Pb-TSP was retained, reflecting the evidence
that Pb particles of all sizes pose health risks. The secondary standard was revised to be
identical in all respects to the revised primary standards.2
Revisions to the NAAQS were accompanied by revisions to the data handling
procedures, the treatment of exceptional events, and the ambient air monitoring and
1 The ANPR was one of the features of the revised NAAQS review process that EPA instituted in 2006. In 2009, this
component of the process was replaced by reinstatement of the OAQPS policy assessment (previously termed the
Staff Paper).
2 The 2008 NAAQS for Pb are specified at 40 CFR 50.16.
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reporting requirements, as well as emissions inventory reporting requirements.1 One
aspect of the new data handling requirements is the allowance for the use of Pb-PMi0
monitoring for Pb NAAQS attainment purposes in certain limited circumstances at
non-source-oriented sites. The monitoring network requirements resulted in a substantial
number of new monitors being required as of January 2010, Subsequent to the 2008
rulemaking, additional revisions were made to the monitoring network requirements,
which required additional monitors as of December 2011; the complete current
requirements are described in Section 2.4.
On February 26, 2010 (75 FR 8934), EPA formally initiated its current review of the air
quality criteria for Pb, requesting the submission of recent scientific information on
specified topics. Soon after, a science policy workshop was held to identify key policy
issues and questions to frame the review of the Pb NAAQS (75 FR 20843). Drawing
from the workshop discussions, a draft IRP [Integrated Review Plan for the National
Ambient Air Quality Standards for Lead (U.S. EPA. 201 Id)], was developed and made
available in late March, 2011 for public comment and consultation with CASAC and was
discussed by the CASAC via a publicly accessible teleconference consultation on May 5,
2011 (76 FR 20347, 76 FR21346). The final IRP (U.S. EPA. 201 Ic) was released in
November, 2011 (76 FR 76972).
As part of the science assessment phase of the current review, EPA held a workshop in
December 2010 (75 FR 69078) to discuss, with invited scientific experts, preliminary
draft materials prepared during the ongoing development of the Pb ISA. The first external
review draft ISA for Lead was released on May 6, 2011 (U.S. EPA. 201 le). The CASAC
Pb Review Panel met at a public meeting on July 20, 2011 to review the draft ISA
(76 FR 36120). Subsequently, on December 9, 2011, the CASAC panel provided a
consensus letter for their review to the Administrator of the EPA (Frey and Samet. 2011).
The second external review draft ISA for Lead, (U.S. EPA. 2012a) was discussed at a
public meeting of the CASAC Pb Review Panel on April 10, 2012 (77 FR 14783).
Subsequently, the CASAC panel provided a consensus letter for their review to the
Administrator of the EPA(Frev. 2012). The third external review draft ISA for Lead was
released on November 27, 2012(U.S. EPA. 2012b). The CASAC panel met at a public
meeting on February 5, 2013, to review the draft ISA (78 FR 938). Subsequently, on June
4, 2013, the CASAC provided a consensus letter for their review to the Administrator of
the EPA (Frev. 2013).
1 The 2008 federal regulatory measurement methods for Pb are specified in 40 CFR 50, Appendix G and 40 CFR
part 53. Consideration of ambient air measurements with regard to judging attainment of the standards is specified in
40 CFR 50, Appendix R. The Pb monitoring network requirements are specified in 40 CFR 58, Appendix D,
Section 4.5. Guidance on the approach for implementation of the new standards was described in the Federal
Register notices for the proposed and final rules (73 FR 29184; 73 FR 66964).
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References for Legislative and Historical Background
Frev. C: Samet JM. (2011). CASAC review of the EPA's integrated science assessment for lead (first
external review draft- May 2011). (EPA-CASAC-12-002). Washington, DC: U.S. Environmental
Protection Agency, Clean Air Scientific Advisory Committee.
Frev. C. (2012). CASAC review of the EPA's integrated science assessment for lead (second external
review draft- February 2012). (EPA-CASAC-12-005). Washington, D.C.: U.S. Environmental
Protection Agency, Clean Air Scientific Advisory Committee.
Frev. C. (2013). CASAC review of the EPA's integrated science assessment for lead (third external review
draft- November 2012). (EPA-CASAC-13-004). Washington, D.C.: U.S. Environmental Protection
Agency, Clean Air Scientific Advisory Committee.
ICF (ICF International). (2006). Lead human exposure and health risk assessments and ecological risk
assessment for selected areas: Pilot phase: External review draft technical report. Research Triangle
Park, NC: U.S. Environmental Protection Agency, Office of Air Quality Planning and Standards.
U.S. EPA (U.S. Environmental Protection Agency). (1977). Air quality criteria for lead [EPA Report].
(EPA-600/8-77-017). Washington, DC.
http://nepis.epa.gov/Exe/ZvPURL.cgi?Dockev=20013GWR.txt
U.S. EPA (U.S. Environmental Protection Agency). (1986a). Air quality criteria for lead [EPA Report].
(EPA/600/8-83/028aF-dF). Research Triangle Park, NC.
http://cfpub.epa.gov/ncea/cfm/recordisplav.cfm?deid=32647
U.S. EPA (U.S. Environmental Protection Agency). (1986c). Lead effects on cardiovascular function,
early development, and stature: An addendum to U.S. EPA Air Quality Criteria for Lead (1986) [EPA
Report]. (EPA-600/8-83/028aF). Washington, DC.
U.S. EPA (U.S. Environmental Protection Agency). (1989). Review of the national ambient air quality
standards for lead: Exposure analysis methodology and validation: OAQPS staff report [EPA Report].
(EPA-450/2-89-011). Research Triangle Park, NC.
U.S. EPA (U.S. Environmental Protection Agency). (1990a). Air quality criteria for lead: Supplement to
the 1986 addendum [EPA Report]. (EPA/600/8-89/049F). Washington, DC.
U.S. EPA (U.S. Environmental Protection Agency). (1990c). Review of the national ambient air quality
standards for lead: Assessment of scientific and technical information: OAQPS staff paper [EPA
Report]. (EPA-450/2-89-022). Research Triangle Park, NC.
U.S. EPA (U.S. Environmental Protection Agency). (1991). Strategy for reducing lead exposures [EPA
Report]. Washington, DC. http://www.epa.gov/ttn/naaqs/standards/pb/data/leadstrategyl991.pdf
U.S. EPA (U.S. Environmental Protection Agency). (2005e). Project work plan for revised air quality
criteria for lead [EPA Report]. (NCEA-R-1465). Research Triangle Park, NC.
http://cfpub.epa.gov/ncea/cfm/recordisplav.cfm?deid=l 13963
U.S. EPA (U.S. Environmental Protection Agency). (2006b). Air quality criteria for lead: Volume I of II
[EPA Report]. (EPA/600/R-05/144aF). Research Triangle Park, NC.
http ://cfpub. epa. gov/ncea/CFM/recordisplav. cfm?deid= 158823
U.S. EPA (U.S. Environmental Protection Agency). (2006d). Analysis plan for human health and
ecological risk assessment for the review of the lead national ambient air quality standards (draft)
[EPA Report]. Research Triangle Park, NC.
http://www.epa.gOv/ttn/naaqs/standards/pb/s pb cr pd.html
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U.S. EPA (U.S. Environmental Protection Agency). (2006f). Plan for review of the national ambient air
quality standards for lead [EPA Report]. Research Triangle Park, NC.
http://www.epa.gov/tm/naaqs/standards/pb/s_pb cr_pd.html
U.S. EPA (U.S. Environmental Protection Agency). (2006g). Review of the national ambient air quality
standards for lead: Policy assessment of scientific and technical information: OAQPS staff paper - first
draft [EPA Report]. (EPA-452/P-06-002). Research Triangle Park, NC.
U.S. EPA (U.S. Environmental Protection Agency). (2007g). Lead: Human exposure and health risk
assessments for selected case studies: Volume 1: Human exposure and health risk assessments - full-
scale [EPA Report]. (EPA-452/R-07-014a). Research Triangle Park, NC.
http://www.ntis. gov/search/product.aspx?ABBR=PB2008102573
U.S. EPA (U.S. Environmental Protection Agency). (20lie). Integrated review plan for the national
ambient air quality standards for lead [EPA Report]. (EPA-452/R-11-008). Research Triangle Park,
NC.
U.S. EPA (U.S. Environmental Protection Agency). (2011d). Integrated review plan for the national
ambient air quality standards for lead: External review draft [EPA Report]. (EPA-452/D-11-001).
Research Triangle Park, NC.
U.S. EPA (U.S. Environmental Protection Agency). (20lie). Integrated science assessment for lead (first
external review draft) [EPA Report]. http://cfpub.epa.gov/ncea/isa/recordisplav.cfm?deid=226323
U.S. EPA (U.S. Environmental Protection Agency). (2012a). Integrated science assessment for lead (2nd
external review draft) [EPA Report]. (EPA/600/R-10/075B). RTF, NC.
U.S. EPA (U.S. Environmental Protection Agency). (2012b). Integrated science assessment for lead (3rd
external review draft) [EPA Report]. (EPA/600/R-10/075C). Research Triangle Park, NC.
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Executive Summary
Introduction
This Integrated Science Assessment (ISA) is a synthesis and evaluation of the most
policy-relevant science that forms the scientific foundation for the review of the primary
(health-based) and secondary (welfare-based)1 national ambient air quality standard
(NAAQS) for Lead (Pb). In 2008, the levels of the primary and secondary NAAQS for
Pb were lowered ten-fold, from the 1978 level of 1.5 ug/m3, to a level of 0.15 ug/m3. The
averaging time was revised to a rolling three-month period with a maximum (not-to-be-
exceeded) form, evaluated over a three-year period. EPA's decision on the level for the
revised primary standard in 2008 was based on the substantially expanded body of health
effects evidence available at that time, including that on cognitive effects of Pb in
children. The revised standard was established to protect against air Pb-related human
health effects, including intelligence quotient (IQ) loss, in the most highly exposed
children.
The Preamble discusses the general framework for conducting the science assessment
and developing an ISA, including criteria for selecting studies for inclusion in the ISA,
evaluating and integrating the scientific evidence and developing scientific conclusions
regarding the causal association of air pollution with health and environmental effects.
As described in Table II of the Preamble, the ISA uses a five-level hierarchy that
classifies the weight of evidence for causation:
• Causal relationship
• Likely to be a causal relationship
• Suggestive of a causal relationship
• Inadequate to infer a causal relationship
• Not likely to be a causal relationship
Studies most relevant to the review of the NAAQS are highlighted in the ISA. In drawing
judgments regarding causality for the criteria air pollutants, evidence of health and
environmental effects in the range of relevant pollutant exposures or biomarker
concentrations is considered. Considerations that are specific to the causal determinations
drawn for the health and ecological effects of Pb are described in more detail in
1 Welfare effects as defined in Clean Air Act (CAA) Section 302(h) [42 U.S.C. 7602(h)] include, but are not limited
to, "effects on soils, water, crops, vegetation, man-made materials, animals, wildlife, weather, visibility and climate,
damage to and deterioration of property, and hazards to transportation, as well as effects on economic values and on
personal comfort and well-being."
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Section 1.1. Briefly, both long and short-term Pb exposures were considered. Evidence
from experimental animal studies observing effects of exposures resulting in blood Pb
levels within an order of magnitude of those currently experienced by the U.S. general
population was emphasized. With regard to the epidemiologic evidence,
population-based studies using Pb biomarkers (i.e., blood or bone Pb concentrations)
were emphasized over occupational studies. Relevant concentrations for drawing
causality judgments for the welfare effects of Pb were determined considering Pb
concentrations generally within one or two orders of magnitude above those which have
been observed in the environment and the available evidence for concentrations at which
effects were observed in plants, invertebrates, and vertebrates. Once conclusions are
drawn, evidence relevant to quantifying the health or environmental risks is assessed.
Sources, Fate and Transport of Lead in the Environment, and the
Resulting Human Exposure and Dose
Emissions of Pb to ambient air declined by more than two orders of magnitude over the
period 1970 to 2008 following the ban on alkyl-Pb additives for on-road gasoline and
tightened industrial emission standards. Air emissions in the U.S. were estimated to be
950 tons in 2008, a small fraction of the Pb processed in U.S. Pb-related industries. More
than half of these emissions were from piston-engine aircraft. Other important sources of
ambient air Pb, beginning with the next largest, include metals processing, fossil fuel
combustion, other industrial sources, and roadway-related sources.
During the same period that saw the dramatic decrease in Pb emissions, ambient air Pb
concentrations1 also declined. The median value (across monitoring sites) for the
maximum 3-month average concentration in 2010, 0.03 micrograms per cubic meter
(ug/m3), was approximately thirty times lower than it was in 1980. The sharpest drop in
Pb concentration occurred from 1980-1990; concentrations continued to decline up to
2010. Specific levels near Pb sources as well as away from Pb sources have also shown a
sharp decrease (Section 1.2.2).
The history of atmospheric deposition has led to measurable Pb concentrations in rain,
snowpack, soil, surface waters, sediments, agricultural plants, livestock, and wildlife
across the world, with the highest concentrations near Pb sources, such as smelters. After
1 The original indicator for the Pb NAAQS is the mass of the Pb portion of total suspended particles (Pb-TSP). The
Pb-TSP indicator was retained in 2008 in recognition of the role of all paniculate matter (PM) sizes in ambient air
Pb exposures (Section 1.2.2). The Federal Reference Method (FRM) Pb-TSP sampler's size-selective performance
is known to be affected by wind speed and direction, and collection efficiency has been demonstrated to decline with
particle size. Under certain conditions regulatory Pb monitoring can also be performed for ambient Pb sampled
using the FRM for Pb sampled in particles with an upper 50% cut-point of 10 ± 0.5 micrometer (^m) aerodynamic
diameter (Pb-PM10). Pb-PM10 is allowed in certain instances where the expected Pb concentration does not approach
the NAAQS and no sources of ultracoarse Pb particles are nearby.
Ixxviii
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the phase-out of Pb from on-road gasoline and declining industrial emissions in the U.S.,
Pb concentrations have decreased considerably in rain, snowpack, and surface waters. Pb
is retained in soils and sediments, where it provides a historical record of deposition and
associated concentrations. The national average Pb concentration in soil in non-urban
locations was 18.9 milligrams of Pb per kilogram (mg Pb/kg), measured in over 1,300
non-urban, generally vegetated sampling locations from 1961 through 1997. The national
median fresh surface water dissolved Pb concentration (1991-2003) was 0.5 micrograms
per liter ((ig/L) (Section 1.2.3). In remote lakes, sediment profiles indicate higher Pb
concentrations in near surface sediment as compared to pre-industrial era sediment from
greater depth; sediment profiles indicate peak Pb concentrations between 1960 and 1980
(when industrial and mobile source Pb emissions in the U.S. were at their peak).
The size distribution of Pb-bearing particulate matter (PM), (i.e., PM having Pb as one of
its components) depends on whether there are contributions from industrial sources or
near-road environments (Section 1.2.2). Airborne particles containing Pb (Pb bearing
PM1) tend to be small (much of the distribution <10 (im) compared with soil or dust
particles containing Pb (-50 (im to several hundred (im). Coarse Pb-bearing PM (i.e.,
approximately 2.5 - 10 (im) deposits to a great extent near its source, contributing to
local soil Pb contamination, while fine Pb-bearing PM (i.e., smaller than approximately
2.5 (im) can be transported long distances and possibly deposit in remote areas.
Depending on local conditions, deposited particles may be resuspended and redeposited
multiple times before further transport becomes unlikely.
There are multiple sources of Pb in the environment, and human exposure to Pb involves
multiple pathways. Figure ES-1 shows how Pb can move through multiple environmental
media. The "air/soil/water" arrows of the figure depict Pb exposures to plants, animals,
and/or humans through contact with Pb-containing media. Air-related pathways of
ambient Pb exposure are the focus of this assessment. Pathways of exposure to Pb from
ambient air include both inhalation of Pb and ingestion of Pb in dust or soil that
originated in the ambient air. For example, dietary Pb exposure may be air-related if
ambient air Pb deposits on plant materials or in water that becomes available for human
consumption. Dust and soil particles containing Pb are typically in the size range that is
ingested rather than inhaled. However, soil can act as a reservoir for deposited Pb
emissions, and exposure to soil contaminated with deposited Pb can occur through
resuspended PM as well as hand-to-mouth contact, which is the main pathway of
childhood exposure to Pb. The primary contribution of ambient air Pb to young children's
blood Pb concentrations is generally due to ingestion of Pb following its deposition in
soils and dusts rather than inhalation of ambient air (Section 3.1.1.2). Non-ambient, non-
1 Pb-bearing PM larger than 10 ^m have a sharp concentration gradient with distance from the source, because
larger particles have greater settling velocities.
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air-related exposures include hand-to-mouth contact with Pb-containing consumer goods,
hand-to-mouth contact with dust or chips of peeling Pb-containing paint, or ingestion of
Pb in drinking water conveyed through Pb pipes. As a result of the multitude of possible
exposure pathways, the contribution from specific pathways (e.g., consumer products,
diet, soil, ambient air) to blood Pb concentrations is situation specific.
Newly Emitted Pb
Historically Emitted Pb
OUJDOORSOIL
NDDUST
NATURAL WATERS
ND SEDIMENTS
Non-air Pb
eleases
AIR
SOIL
WATER
AIR
SOIL
WATER
AIR
SOIL
WATER
PLANT
EXPOSURE
HUMAN
EXPOSURE
ANIMAL
EXPOSURE
OCCUPATION
( COSMETICS
OYS etc
Note: This Venn diagram illustrates the passage of Pb through multiple environmental media compartments through which plant,
animal, and human exposures can occur.
Figure ES-1 Conceptual model of multimedia Pb exposure.
The majority of Pb in the body is stored in bone (roughly 90% in adults, 70% in
children). Much of the remaining Pb is found in soft tissues; only about 1% of Pb is
found in the blood. Pb in blood is primarily (-99%) bound to red blood cells [RBCs]).
The small fraction of Pb in blood plasma (<1% of Pb in blood) may be the more
biologically labile and lexicologically active fraction of the circulating Pb. Both
Pb uptake to and elimination from soft tissues are much faster than they are in bone. Pb
accumulates in bone regions undergoing the most active calcification at the time of
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exposure. Pb in bone becomes distributed in trabecular (e.g., patella [knee cap]) and the
more dense cortical bones (e.g., tibia [shin bone]).
Blood Pb is the most common index of Pb exposure in epidemiologic studies of Pb health
effects. Overall, blood Pb levels have been decreasing among U.S. children and adults for
the past thirty-five years. The median blood Pb level for the U.S. population is 1.1
micrograms per deciliter (ug/dL), with a 95th percentile blood Pb level of 3.3 ug/dL
based on the 2009-2010 National Health and Nutrition Examination Survey (NHANES)
data. Among children aged 1-5 years, the median and 95th percentiles are slightly higher
at 1.2 ug/dL and 3.4 ug/dL, respectively. Other common Pb exposure metrics used in
epidemiologic studies are Pb in bone, which generally reflects cumulative exposure over
long periods (months to years), and Pb in cord blood, which is an indicator of prenatal
and neonatal blood Pb concentration.
Blood Pb is dependent on both the recent exposure history of the individual, as well as
the long-term exposure history that determines total body burden and the amount of Pb
stored in the bone. The contribution of bone Pb to blood Pb changes throughout an
individual's life time, and depends on the duration and intensity of the exposure, age, and
various other physiological stressors (e.g., nutritional status, pregnancy, menopause,
extended bed rest, hyperparathyroidism) that may affect bone remodeling, which
normally and continuously occurs. In children, largely due to faster exchange of Pb to
and from bone, blood Pb is both an index of recent exposure and potentially an index of
body burden. Generally, bone Pb is an index of cumulative exposure and body burden. Pb
is sequestered in two types of bone compartments; Pb in trabecular bone exchanges more
rapidly with the blood than Pb in denser cortical bone. Therefore, Pb in cortical bone is a
better marker of cumulative exposure, and Pb in trabecular bone is more likely to be
correlated with blood Pb concentration. During pregnancy, Pb is transferred from the
mother to the fetus. Transplacental transfer of Pb may be facilitated by an increase in the
plasma/blood Pb concentration ratio during pregnancy. Maternal-to-fetal transfer of Pb
appears to be related partly to the mobilization of Pb from the maternal skeleton.
Integrative Overview of Health and Ecological Effects
There is substantial overlap between the ecological and human health endpoints related to
Pb exposure, which can be mediated through multiple, interconnected modes of action
(MOAs). The cellular/subcellular effect constituting the principal MOA for human health
and ecological endpoints is altered ion status. Other related MOAs include protein
binding, oxidative stress, inflammation, endocrine disruption, and cell death and
genotoxicity (Figure ES-2). Since the mechanisms of Pb toxicity in some organ systems
are the same or similar across species, many of the downstream health and ecological
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effects are similar across species from invertebrates to vertebrates, including humans
(Section 1.8.1).
Altered Ion
Status
(4.2.2)
Protein
Binding
(4.2.3)
Ł
Oxi dative
Stress
(4.2.4)
^\
/
Cell Death
and
Genotoxicity
(4.2.7)
Inflammation
(4.2.5)
Endocrine
Disruption
(4.2.6)
Note: The subsections where these MOAs are discussed are indicated in parentheses.
(Section 4.2.2: Section 4.2.3: Section 4.2.4: Section 4.2.5: Section 4.2.6: and Section 4.2.7).
Figure ES-2
Schematic representation of the relationships between the
various MOAs by which Pb exerts its effects.
Health Effects of Pb
Evidence from epidemiologic and toxicological studies was considered in combination
with the evidence from other disciplines such as toxicokinetics in determining the causal
relationships for the health endpoints discussed in this assessment. Detailed discussions
of the evidence relating to conclusions regarding the health effects of Pb are in
Section 1.6 and Chapter 4. The major conclusions regarding health effects from Pb
exposure in children and adults are presented in Table ES-1 and summarized below.
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Table ES-1 Summary of causal determinations for the relationship between
exposure to Pb and health effects.
Causality Determination3
Health Outcome (Table with Key Evidence)
Nervous System Effects (Section 1.6.1)
Children - Nervous System Effects (Section 1.6.1.1)
Cognitive Function Decrements Causal Relationship (Table 4-17)
Clear evidence of cognitive function decrements (as measured by Full Scale IQ, academic performance, and
executive function) in young children (4 to 1 1 years old) with mean or group blood Pb levels measured at
various lifestages and time periods between 2 and 8 ug/dL. Clear support from animal toxicological studies that
demonstrate decrements in learning, memory, and executive function with dietary exposures resulting in
relevant blood Pb levels of 10-25 ug/dL. Plausible MOAs are demonstrated.
Re.a.ionshlp
Clear evidence of attention decrements, impulsivity and hyperactivity (assessed using objective
neuropsychological tests and parent and teacher ratings) in children 7-17 years and young adults ages 19-20
years. The strongest evidence for blood Pb-associated increases in these behaviors was found in prospective
studies examining prenatal (maternal or cord), age 3-60 months, age 6 years, or lifetime average (to age 11-13
years) mean blood Pb levels of 7 to 14 ug/dL and groups with early childhood (age 30 months) blood Pb levels
>10 ug/dL. Biological plausibility is provided by animal toxicological studies demonstrating impulsivity or
impaired response inhibition with relevant prenatal, lactational, post-lactational and lifetime Pb exposures.
Plausible MOAs are demonstrated.
Externalizing Behaviors:
Conduct Disorders in Children and Young Adults'
c, d Likely Causal Relationship (Table 4-17)
Prospective epidemiologic studies find that early childhood (age 30 months, 6 years) or lifetime average (to age
11-13 years) blood Pb levels or tooth Pb levels (from ages 6-8 years) are associated with criminal offenses in
young adults ages 19-24 years and with higher parent and teacher ratings of behaviors related to conduct
disorders in children ages 8-17 years. Pb-associated increases in conduct disorders were found in populations
with mean blood Pb levels 7 to14 ug/dL; associations with lower blood Pb levels as observed in cross-sectional
studies were likely to be influenced by higher earlier Pb exposures. There is coherence in epidemiologic findings
among related measures of conduct disorders. Evidence of Pb induced aggression in animals was mixed, with
increases in aggression found in some studies of adult animals with gestational plus lifetime Pb exposure but
not juvenile animals. The lack of clear biological plausibility produces some uncertainty.
Internalizing Behaviors Likely Causal Relationship (Table 4-17)
Prospective epidemiologic studies find associations of higher lifetime average blood (mean: -14 ug/dL) or
childhood tooth (from ages 6-8 years) Pb levels with higher parent and teacher ratings of internalizing behaviors
such as symptoms of depression or anxiety, and withdrawn behavior in children ages 8-13 years. Consideration
of potential confounding by parental caregiving was not consistent and findings from cross-sectional studies in
populations ages 5 and 7 years with mean blood Pb levels of 5 ug/dL were mixed. Animal toxicological studies
demonstrate depression-like behaviors and increases in emotionality with relevant lactational exposures.
Plausible MOAs are demonstrated.
Auditory Function Decrements Likely Causal Relationship (Table 4-17)
A prospective epidemiologic study and large cross-sectional studies indicate associations between blood Pb
levels and increased hearing thresholds at ages 4-19 years. Across studies, associations were found with blood
Pb levels measured at various time periods, including prenatal maternal, neonatal (10 day, mean 4.8 ug/dL),
lifetime average, and concurrent (ages 4-19 years) blood Pb levels (median 8 ug/dL). Plausible MOAs are
demonstrated. The lack of biological plausibility in animals with relevant exposures produces some uncertainty.
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Table ES-1 (Continued): Summary of causal determinations for the relationship between
exposure to Pb and health effects.
Causality Determination3
Health Outcome (Table with Key Evidence)
Visual Function Decrements Inadequate to Infer a Causal Relationship (Table 4-17)
The available epidemiologic and toxicological evidence is of insufficient, quantity, quality and consistency.
Motor Function Decrements Likely Causal Relationship (Table 4-17)
Prospective epidemiologic studies provide evidence of associations of fine and gross motor function decrements
in children ages 4-17 years with lifetime average blood Pb levels and with blood Pb levels measured at various
time periods with means generally ranging from 4.8 to 12 ug/dL. Results were inconsistent in cross sectional
studies with concurrent blood Pb level means 2-5 ug/dL. Limited evidence in animal toxicological studies with
relevant Pb exposures.
Adults - Nervous System Effects (Section 1.6.1.2)
Cognitive Function Decrements Likely Causal Relationship (Table 4-17)
Prospective studies indicate associations of higher baseline bone Pb levels with declines in cognitive function
(executive function, visuospatial skills, learning and memory) in adults (>age 50 years) over 2-to 4-year
periods. Cross-sectional studies provide additional support. Uncertainties remain regarding the timing,
frequency, duration and level of the Pb exposures contributing to the effects observed and residual confounding
by age. Biological plausibility is provided by findings that relevant lifetime Pb exposures from gestation, birth, or
after weaning induce learning impairments in adult animals and by evidence demonstrating plausible MOAs.
Psychopathological Effects Likely Causal Relationship (Table 4-17)
Cross-sectional studies in a few populations demonstrate associations of higher concurrent blood or tibia Pb
levels with self-reported symptoms of depression and anxiety in adults. Uncertainties remain regarding the
timing, frequency, duration and level of Pb exposures contributing to the observed associations and residual
confounding by age. Observations of depression-like behavior in animals with dietary lactational Pb exposure,
with some evidence at relevant blood Pb levels, and evidence demonstrating plausible MOAs in experimental
animals provides support.
Auditory Function Decrements Suggestive of a Causal Relationship (Table 4-17)
A high-quality prospective epidemiologic study finds associations of higher tibia Pb level with a greater rate of
elevations in hearing threshold over 20 years. Some evidence indicates effects on relevant MOAs but important
uncertainties remain related to effects on auditory function in animals with relevant Pb exposures.
Visual Function Decrements Inadequate to Infer a Causal Relationship (Table 4-17)
The available epidemiologic and toxicological evidence is of insufficient, quantity, quality and consistency.
Neurodegenerative Diseases Inadequate to Infer a Causal Relationship (Table 4-17)
The available epidemiologic and toxicological evidence is of insufficient, quantity, quality and consistency.
Cardiovascular Effects (Section 1.6.2)
Hypertension Causal Relationship (Table 4-24)
Prospective epidemiologic studies with adjustment for multiple potential confounders consistently find
associations of blood and bone Pb levels with hypertension incidence and increased blood pressure (BP) in
adults. Cross-sectional studies provide supporting evidence. Meta-analyses underscore the consistency and
reproducibility of the Pb associated increase in blood pressure and hypertension (a doubling of concurrent blood
Pb level (between 1 and 40 ug/dL) is associated with a 1 mmHg increase in systolic BP); however, uncertainties
remain regarding the timing, frequency, duration and level of Pb exposures contributing to the effects observed in
epidemiologic studies. Experimental animal studies demonstrate effects on BP after long-term Pb exposure
resulting in mean blood Pb levels of 10 ug/dL or greater. Plausible MOAs are demonstrated.
Subclinical Atherosclerosis Suggestive of a Causal Relationship (Table 4-24)
Cross-sectional analyses of NHANES data find associations of blood Pb level with peripheral artery disease
(PAD) in adults. Animal toxicological evidence is limited to studies of MOA (oxidative stress, inflammation,
endothelial cell dysfunction) that demonstrate biologically plausible mechanisms through which Pb exposure may
initiate atherosclerotic vessel disease.
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Table ES-1 (Continued): Summary of causal determinations for the relationship between
exposure to Pb and health effects.
Causality Determination3
Health Outcome (Table with Key Evidence)
Coronary Heart Disease Causal Relationship (Table 4-24)
Prospective epidemiologic studies consistently find associations of Pb biomarkers with cardiovascular mortality
and morbidity, specifically myocardial infarction (Ml), ischemic heart disease (IHD), or HRV; however,
uncertainties remain regarding the timing, frequency, duration and level of Pb exposures contributing to the
effects observed in epidemiologic studies. Thrombus formation was observed in animals after relevant long term
exposure and MOAs (hypertension, decreased HRV, increased corrected QT (QTc) interval, and corrected QRS
complex (QRSc) duration in electrocardiogram [ECG]) are demonstrated in humans and animals.
Cerebrovascular Disease Inadequate to Infer a Causal Relationship (Table 4-24)
The available epidemiologic and toxicological evidence is of insufficient, quantity, quality, and/or consistency.
Plausible MOAs, which are shared with hypertension and atherosclerosis, are demonstrated.
Renal Effects (Section 1.6.3)
Reduced Kidney Function Suggestive of a Causal Relationship (Table 4-31)
Multiple high quality epidemiologic studies provide evidence that Pb exposure is associated with reduced kidney
function; however, uncertainty remains regarding the potential for reverse causality to explain findings in humans.
Further, inconsistencies and limitations in occupational studies, epidemiologic studies of children and clinical trials
of chelation of CKD patient preclude strong inferences to be drawn based on their results. Although longitudinal
studies found Pb-associated decrements in renal function in populations with mean blood Pb levels of 7 and 9
ug/dL, the contributions of higher past Pb exposures cannot be excluded. Animal toxicological studies
demonstrate Pb-induced kidney dysfunction at blood Pb levels greater than 30 ug/dL; however, evidence in
animals with blood Pb levels < 20 ug/dL is generally not available. At blood Pb levels between 20 and 30 ug/dL
studies provide some evidence for dysfunction in kidney function measures (e.g., decreased creatinine clearance,
increased serum creatinine, increased BUN). Plausible MOAs (Pb induced hypertension, renal oxidative stress
and inflammation, morphological changes, and increased uric acid) are demonstrated.
Immune System Effects (Section 1.6.4)
Atopic and Inflammatory Responses Likely Causal Relationship (Table 4-34)
Prospective studies of children ages 1-5 years indicate associations of prenatal cord and childhood blood Pb
levels with asthma and allergy. This evidence is supported by cross-sectional associations between higher
concurrent blood Pb levels (>10 ug/dl_) in children and higher IgE. Uncertainties related to potential confounding
by SES, smoking or allergen exposure are reduced through consideration of the evidence from experimental
animal studies. The biological plausibility for the effects of Pb on IgE is provided by consistent findings in animals
with gestational or gestational-lactational Pb exposures, with some evidence at blood Pb levels relevant to
humans. Strong evidence of Pb-induced increases in Th2 cytokine production and inflammation in animals
demonstrates MOA.
Decreased Host Resistance Likely Causal Relationship (Table 4-34)
Animal toxicological studies provide the majority of the evidence for Pb-induced decreased host resistance.
Dietary Pb exposure producing relevant blood Pb levels (7-25 ug/dL) results in increased susceptibility to bacterial
infection and suppressed delayed type hypersensitivity. Further, evidence demonstrating plausible MOA,
including suppressed production of Th1 cytokines and decreased macrophage function in animals, provides
coherence.
Autoimmunity Inadequate to Infer a Causal Relationship (Table 4-34)
The available toxicological and epidemiologic studies do not sufficiently inform Pb-induced generation of auto-
antibodies with relevant Pb exposures.
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Table ES-1 (Continued): Summary of causal determinations for the relationship between
exposure to Pb and health effects.
Causality Determination3
Health Outcome (Table with Key Evidence)
Hematologic Effects (Section 1.6.5)
** ^^ ^ ^ ^^ *"* Causal Relationship (Table 4-35)
Animal toxicological studies demonstrate that exposures resulting in blood Pb levels relevant to humans (2-7
ug/dL) alter several hematological parameters (Hemoglobin [Hb], Hematocrit [Hct], and mean corpuscular volume
[MCV]), increase measures of oxidative stress and increase cytotoxicity in red blood cell (RBC) precursor cells.
Limited body of epidemiologic studies provides additional support for the association of Pb exposure with these
endpoints. Plausible MOAs are demonstrated in experimental animals.
Altered Heme Synthesis Causal Relationship (Table 4-35)
Consistent findings from studies in experimental adult animal studies report that relevant exposures (e.g. blood
Pb levels of 6.5 ug/dL) cause decreased ALAD and ferrochelatase activities. Additional support is garnered from a
larger body of ecotoxicological studies demonstrating decreased ALAD activity across a wide range of species
and a limited body of epidemiologic studies. Plausible MOAs are demonstrated in experimental animals.
Reproductive and Developmental Effects (Section 1.6.6)
Development Causal Relationship (Table 4-48)
Multiple cross-sectional epidemiologic studies report associations between concurrent blood Pb levels and
delayed pubertal onset for girls (6-18 years) and boys (8-15 years). These associations are consistently observed
in populations with concurrent blood Pb levels 1 .2-9.5 ug/dL. Few studies consider confounding by nutrition.
Uncertainties remain regarding the timing, frequency, duration and level of Pb exposures contributing to the
effects observed in epidemiologic studies of older children. Experimental animal studies demonstrate delayed
onset of puberty in female pups with blood Pb levels of 1.3-13 ug/dL and delayed male sexual maturity at blood
Pb levels of 34 ug/dL.
abortion?0™5 (6'9" '°W Suggestive of Causal Relationship (Table 4-48)
Some well-conducted epidemiologic studies report associations of maternal Pb biomarkers or cord blood Pb with
preterm birth and low birth weight/fetal growth; however, the epidemiologic evidence is inconsistent overall and
findings from experimental animal studies are mixed.
Male Reproductive Function Causal Relationship (Table 4-48)
Key evidence is provided by toxicological studies in rodents, non-human primates, and rabbits showing
detrimental effects on semen quality, sperm and fecundity/fertility with supporting evidence in epidemiologic
studies. Toxicological studies with relevant Pb exposure routes leading to blood Pb concentrations ranging from
5-43 ug/dL reported effects on sperm quality and sperm production rate, sperm DMA damage, and histological or
ultrastructural damage to the male reproductive organs. Consistent associations in studies of occupational
populations with concurrent blood Pb levels of 25 ug/dL and greater, report detrimental effects of Pb on sperm;
however, uncertainties remain regarding the timing, frequency, duration and level of Pb exposures contributing to
the effects observed in epidemiologic studies.
Female Reproductive Function Suggestive of Causal Relationship (Table 4-48)
Although findings are mixed overall, the body of evidence include some high-quality epidemiologic and
toxicological studies, suggesting that Pb may affect some aspects of female reproductive function (hormone level,
placental pathology).
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Table ES-1 (Continued): Summary of causal determinations for the relationship between
exposure to Pb and health effects.
Causality Determination3
Health Outcome (Table with Key Evidence)
Cancer (Section 1.6.7)
Cancer Likely Causal Relationship (Table 4-50)
The animal toxicological literature provides the strong evidence for long-term exposure (i.e., 18 months or 2
years) to high concentrations of Pb (> 2,600 ppm) inducing tumor development; findings from epidemiologic
studies inconsistent. Plausible MOAs are demonstrated.
a In drawing conclusions regarding the causal relationship between Pb exposure and human health effects, evidence in the range of
relevant pollutant exposures or biomarker levels was considered. Specifically, population-based epidemiology studies were
emphasized with the recognition that many of the U.S populations studied included individuals with higher past than recent Pb
exposures. Evidence from toxicological studies of effects observed in experimental animals at biomarker levels (e.g. blood Pb)
comparable to those currently experienced by the U.S. general population were emphasized. Generally, studies with dietary
exposures resulting in blood Pb levels within one order of magnitude above the upper end of the distribution of U.S. blood Pb levels
were considered in forming concusions, with the majority of studies reporting blood Pb levels below 30 ug/dL. Studies with higher
blood Pb levels were considered if they informed the evaluation of MOA, mechanisms, or kinetics. (Preamble. Section 1.1).
b Within the attention deficit hyperactivity disorder domain of externalizing behaviors, studies of Pb exposure have focused primarily
on attention, impulsivity, and hyperactivity. Because the studies of ADHD were limited in terms of their design and did not
adequately consider potential confounding by factors such as SES, parental education, or parental caregiving quality, they were not
a major consideration in drawing conclusions about the relationship between Pb exposure and attention, impulsivity, and
hyperactivity.
0 Two domains of conduct disorders,(i.e., undersocialized aggressive conduct disorder and socialized aggressive conduct disorder),
are combined for the purpose of this assessment because it is difficult to differentiate between these two domains in the available
epidemiologic studies, which examine multiple endpoints such as delinquent behavior, aggression, antisocial behavior. Criminal
offenses are included in the evaluation because they can be predicted by earlier conduct disorders (Section 4.3.3.2).
d There was limited evaluation of potential confounding by parental psychopathology, which is a strong risk factor for externalizing
behaviors, in the majority of the epidemiologic studies; however, evidence of an association of between psychopathology in parents
and Pb exposure in their children is not available (Section 4.3.3).
e Strong evidence from experimental animal studies reduces uncertainty related to confounding generally.
Effects of Pb Exposure in Children
Multiple epidemiologic studies conducted in diverse populations of children consistently
demonstrate the harmful effects of Pb exposure on cognitive function (as measured by IQ
decrements, decreased academic performance and poorer performance on tests of
executive function). Blood Pb-associated effects on cognitive function were found in
populations of children (ages 4-10) with mean or group blood Pb levels measured
concurrently or earlier in the range of 2-8 (ig/dL1. Evidence suggests that some Pb-related
cognitive effects may be irreversible and that the neurodevelopmental effects of Pb
exposure may persist into adulthood (Section 1.9.4). Epidemiologic studies also
demonstrate that Pb exposure is associated with decreased attention, and increased
impulsivity and hyperactivity in children (externalizing behaviors). This is supported by
findings in animal studies demonstrating both analogous effects and biological
plausibility at relevant exposure levels. Pb exposure can also exert harmful effects on
blood cells and blood producing organs, and is likely to cause an increased risk of
symptoms of depression and anxiety and withdrawn behavior (internalizing behaviors),
The age range and blood Pb levels are based on studies described in detail in Section 4.3.2.
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decreases in auditory and motor function, asthma and allergy, as well as conduct
disorders in children and young adults. There is some uncertainty about the Pb exposures
contributing to the effects and blood Pb levels observed in epidemiologic studies;
however, these uncertainties are greater in studies of older children and adults than in
studies of young children (Section 1.9.5). Despite these uncertainties, it is clear that Pb
exposure in childhood presents a risk; further, there is no evidence of a threshold below
which there are no harmful effects on cognition from Pb exposure.
Effects of Pb Exposure in Adults
A large body of evidence from both epidemiologic studies of adults and experimental
studies in animals demonstrates the effect of long-term Pb exposure on increased blood
pressure (BP) and hypertension (Section 1.6.2). In addition to its effect on BP, Pb
exposure can also lead to coronary heart disease and death from cardiovascular causes
and is associated with cognitive function decrements, symptoms of depression and
anxiety, and immune effects in adult humans. The extent to which the effects of Pb on the
cardiovascular system are reversible is not well-characterized. Additionally, the
frequency, timing, level, and duration of Pb exposure causing the effects observed in
adults has not been pinpointed, and higher past exposures may contribute to the
development of health effects measured later in life. It is clear however, that Pb exposure
can result in harm to the cardiovascular system that is evident in adulthood and may also
affect a broad array of organ systems.
Ecological Effects of Pb
Ecological effects of Pb are summarized for terrestrial, freshwater and saltwater
ecosystems, and the ISA discusses endpoints common to plants, invertebrates and
vertebrates along with considerations of uncertainties in relating atmospheric Pb
concentrations to ecosystem effects. Effects of Pb in ecosystems are primarily associated
with Pb deposition onto soil and water, subsequent transport, and exposure through
environmental media (soil, water, sediment, biota). The 2006 Pb Air Quality Criteria
Document (AQCD) (U.S. EPA. 2006b) and previous EPA assessments reported effects of
Pb exposure on both terrestrial and aquatic organisms that included reduced survival,
reproduction and growth as well as effects on behavior, development, and heme
production. Studies reviewed in this ISA generally support the ecological findings of
previous Pb assessments with some effects observed in additional species and at lower
concentrations. Reproduction, growth, and survival are endpoints commonly used in
ecological risk assessment because they can lead to effects at the population, community,
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and ecosystem levels of biological organization. Impacts on hematological,
neurobehavioral and physiological stress endpoints may increase susceptibility to other
stressors and affect the fitness of individual organisms. Increasing exposures generally
result in increasing responses in laboratory and field experiments but the relationship of
exposure and responses is difficult to characterize quantitatively in natural systems
because of the influence of multiple environmental variables on both Pb bioavailability
and toxicity, and substantial species and lifestage differences in Pb sensitivity.
A brief discussion of the conclusions from this assessment and earlier Pb AQCDs
regarding Pb effects on reproduction, growth, and survival is provided below and
summarized in Table ES-2 along with effects of Pb on neurobehavior, hematological, and
stress endpoints. Causal determinations for ecological effects were based on integration
of information on biogeochemistry, bioavailability, biological effects, and exposure-
response relationships of Pb in terrestrial, freshwater, and saltwater environments. In
general, the number of studies available for assessing causality is greater for freshwater
organisms than for marine environments. A detailed discussion for all relevant welfare
effects (i.e., ecological effects) is provided in Section 1.7 and Chapter 6.
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Table ES-2 Summary of causal determinations for the relationship between Pb
exposure and effect on plants, invertebrates, and vertebrates.
Level
Effect
Terrestrial
Freshwater
Saltwater3
Community-
and Ecosystem
ation-Level Endpoints
3
Q.
0
Q.
Sub-organismal Organism-Level Responses
Responses
Community and Ecosystem Effects
(Section 1.7.3.7)
Reproductive and Developmental Effects-
Plants (Section 1.7.3.1)
Reproductive and Developmental Effects-
Invertebrates (Section 1.7.3.1)
Reproductive and Developmental Effects-
Vertebrates (Section 1.7.3.1)
Growth-Plants (Section 1.7.3.2)
Growth-Invertebrates (Section 1.7.3.2)
Growth-Vertebrates (Section 1.7.3.2)
Survival-Plants (Section 1.7.3.3)
Survival- Invertebrates (Section 1.7.3.3)
Survival- Vertebrates (Section 1.7.3.3)
Neurobehavioral Effects-
Invertebrates (Section 1.7.3.4)
Neurobehavioral Effects-
Vertebrates (Section 1.7.3.4)
Hematological Effects-
Invertebrates (Section 1.7.3.5)
Hematoloaical Effects-Vertebrates (Section 1.7.3.5)
Physiological Stress-Plants (Section 1.7.3.6)
Physiological Stress-Invertebrates (Section 1.7.3.6)
Physiological Stress-Vertebrates (Section 1.7.3.6)
Likely Causal
Inadequate
Causal
Causal
Causal
Likely Causal
Inadequate
Inadequate
Causal
Likely Causal
Likely Causal
Likely Causal
Inadequate
Causal
Causal
Likely Causal
Likely Causal
Likely Causal
Inadequate
Causal
Causal
Likely Causal
Causal
Inadequate
Inadequate
Causal
Causal
Likely Causal
Likely Causal
Likely Causal
Causal
Likely Causal
Likely Causal
Likely Causal
Inadequate
Inadequate
Suggestive
Inadequate
Inadequate
Inadequate
Inadequate
Inadequate
Inadequate
Inadequate
Inadequate
Inadequate
Suggestive
Inadequate
Inadequate
Suggestive
Inadequate
Conclusions are based on the weight of evidence for causal determination in Table II of the ISA Preamble. Ecological effects
observed at or near ambient Pb concentrations measured in soil, sediment and water in the most recent available studies (Table
1-1). were emphasized and studies generally within one to two orders of magnitude above the reported range of these values were
considered in the body of evidence for terrestrial (Section 6.3.12). freshwater (Section 6.4.12) and saltwater (Section 6.4.21)
ecosystems.
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Effects on Development and Reproduction
Reduced reproduction at the level of individual organisms can result in lowered
population numbers or extermination, decreased species diversity, and a decline in
relative or absolute population numbers at the community level. Effects of Pb on various
development, fertility, and hormone maintenance endpoints have been documented in
multiple species of terrestrial and freshwater organisms. In plants, only a few studies
have addressed reproductive effects of Pb exposure. Among the animal species tested,
freshwater invertebrates were the most sensitive to Pb with respect to reproduction
(Section 1.7.3.1V
Effects on Growth
Effects on growth observed at the species level can translate into effects at the ecosystem
level. Exposure to Pb has been shown to have effects on growth in plants and in some
species of invertebrates and vertebrates. Evidence for effects of Pb on growth is strongest
in terrestrial plants. These effects are typically found in laboratory studies with high Pb
exposure concentrations or in field studies near stationary sources such as metal
industries or mines where concentrations of multiple metals are elevated relative to non-
polluted locations. Many of those laboratory and field studies evaluate the effects of
increasing levels of Pb exposure, and find that effects on plant growth increase with
increasing exposure ("biological gradients"). Evidence for Pb effects on growth in
invertebrates has been observed most extensively in freshwater species, with growth
inhibition in a few sensitive species occurring in the range of Pb concentration values
available for U.S. surface waters. In general, juvenile organisms are more sensitive than
adults. There are only limited data on growth effects in vertebrates (Section 1.7.3.2).
Effects on Survival
Decreased survival of individuals within a population can have ecosystem-level impacts.
Pb is generally not toxic to aquatic or terrestrial plants at concentrations found in the
environment away from stationary sources, probably due to the fact that plants often
sequester large amounts of Pb in roots, with little translocation to other parts of the plant.
Aquatic invertebrates are generally more sensitive to Pb exposure than other types of
organisms, with survival reduced in laboratory studies of a few species at concentrations
occurring near Pb sources, as well as at concentrations occasionally encountered in the
general environment (that is, far from major Pb sources). Many terrestrial invertebrates
tolerate higher concentrations of Pb. Limited studies with vertebrates showed adverse
effects of Pb on survival at concentrations higher than typical ambient Pb levels in the
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environment, although juvenile organisms are usually more sensitive than adults
(Section 1.7.3.3).
Neurobehavioral Effects
Historical and recent evidence from Pb-exposed animals indicates that Pb affects
behaviors, such as food consumption, avoidance and escape from predators, behavioral
regulation of body temperature, and prey capture. Alterations to these behaviors can
decrease the overall fitness of the organism. Evidence from laboratory studies has shown
effects of Pb exposure on nervous system endpoints in both terrestrial and freshwater
animal taxa (Section 1.7.3.4).
Hematological Effects
Changes in hematological characteristics including ALAD (delta-aminolevulinic acid
dehydratase, an important rate-limiting enzyme needed for heme production) activity,
blood cell counts, and serum profiles are associated with Pb exposure in both aquatic and
terrestrial animals. It is commonly recognized that ALAD is an indicator of Pb exposure
across a wide range of animals as shown in both field and laboratory studies. Studies
conducted over the last two decades have shown that hematological responses are
associated with Pb in the environment (Section 1.7.3.5).
Effects on Physiological Stress
Increased levels of antioxidant enzymes (in response to oxidative stress or altered cell
signaling) and increased lipid peroxidation (the process by which free radicals induce the
oxidation of fatty acids, leading to cell membrane damage) are considered to be reliable
biomarkers of stress. Alterations in these biomarkers are associated with Pb exposure in
plants, invertebrates and vertebrates, and they may be indicative of increased
susceptibility to other stressors, as well as reduction in individual fitness. Markers of
oxidative damage and antioxidant activity have been observed in field studies in a wide
range of species in terrestrial and aquatic environments when Pb is present (along with
other chemicals), and also following laboratory exposures (Section 1.7.3.6).
Community and Ecosystem Effects
The effects of Pb on growth, reproduction, and survival at the level of individual
organisms, especially when considered cumulatively, are likely to result in effects on
population, community and ecosystem structure and function. Effects at those higher
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levels of biological organization are confirmed by both laboratory and field experiments.
In these experiments decreases in abundance, reduced species diversity, shifts in soil
microbial and plant community composition (in terrestrial ecosystems), and sediment-
associated and aquatic plant community composition (in freshwater ecosystems) have
been observed following Pb exposure. However, such ecosystem-wide effects can only
be tested directly in a few of the cases where individual organism effects are found.
Quantitative characterization of exposure-response relationships is difficult at the
community and ecosystem levels because potential confounders such as the presence of
other metals, physico-chemical variables and other stressors cannot be controlled and
their effects are incompletely characterized (Section 1.7.3.7).
Policy Relevant Considerations
Public Health Significance
The 2006 Pb AQCD (U.S. EPA. 2006b) concluded that neurodevelopmental effects in
children and cardiovascular effects in adults were of the greatest public health concern
because the evidence indicated that these effects occurred at the lowest blood Pb levels,
compared to other health effects. The evidence reviewed in the current assessment
supports and builds upon this conclusion. Small shifts in the population mean IQ can be
highly significant from a public health perspective because such shifts could translate into
a larger proportion of the population functioning at the low end of the IQ distribution
(Section 1.9.1). as well as a smaller proportion of population functioning at the high end
of the distribution1. Additionally, small Pb-associated increases in the population mean
blood pressure could result in an increase in the proportion of the population with
hypertension that is significant from a public health perspective.
Air Lead(Pb)-to-Blood Lead(Pb) Relationships
A limited number of epidemiologic studies evaluated relationships between air Pb and
blood Pb (Section 1.9.2). Regression models are typically used to produce slopes that
estimate the change in blood Pb per change in air Pb concentration ((ig/dL per (ig Pb/m3).
The larger the slope, the larger is the estimated contribution of air Pb to the blood Pb
level in exposed populations.
The range of air-to-blood slope estimates is 4 to 9 (ig/dL per (ig/m3 in studies of children.
The differences in the estimates across studies, at least in part, reflect the choice of model
1 This statement follows from the conceptual model described by Weiss et al. (1988X which assumes that the
incremental concentration-response between Pb exposure and IQ is similar across the full range of IQ and is not
based on actual data.
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(e.g., some models predict an increase in the blood Pb-air Pb slope with decreasing air Pb
concentration while other models predict a constant blood Pb-air Pb slope across all air
Pb concentrations). In addition, differences in the estimates across studies may reflect the
different terms that are included in the model (e.g., soil Pb); these terms may account for
some of the variation in blood Pb that is attributable to air Pb. Other factors that may
explain the variation in the derived blood Pb-air Pb slope include differences in the
populations examined and Pb sources (e.g., leaded gasoline or smelter).
Concentration-Response Relationships for Health Effects
Previous assessments found that progressively lower blood Pb levels were associated
with cognitive deficits in children, and newly available evidence is generally consistent
with findings of the previous review (Section 1.9.3). Compelling evidence for a larger
incremental effect of Pb on children's IQ at lower blood Pb levels compared to higher
blood Pb levels was presented in the 2006 Pb AQCD based on the international pooled
analysis of seven prospective cohort studies, as well as several individual studies. This
was supported by a subsequent reanalysis of the pooled data focusing on the shape of the
concentration-response function. Several recent studies also support the findings of the
original pooled analysis. The majority of the epidemiologic evidence from stratified
analyses comparing the lower and the higher ends of the blood Pb distributions also
indicates larger effect of Pb on IQ at lower blood Pb levels. The shape of concentration-
response relationships is not well characterized for association of health effects with
blood or bone Pb concentrations in adults (Section 1.9.3).
Pb Exposure and Neurodevelopmental Deficits in Children
Information about the patterns of exposure that contribute to the blood Pb levels and
effects observed in epidemiologic studies is generally lacking. Although, blood Pb may
reflect both recent exposures as well as past exposures because Pb is both taken up by
and released from the bone, uncertainty regarding the role of recent exposure is greater in
adults and older children than in young children who do not have lengthy exposure
histories. Several lines of evidence inform the interpretation of epidemiologic studies of
young children with regard to the exposures that contribute to observed health effects
(Section 1.9.4). Epidemiologic studies find associations of cognitive function with
several different blood Pb metrics that represent blood Pb during lifestages or time
periods from the prenatal period through adolescence. This epidemiologic evidence is
supported by studies of rodents and monkeys indicating that Pb exposures during
multiple lifestages and time periods, including prenatal only, prenatal plus lactational,
postnatal only, and lifetime are observed to induce impairments in learning. These
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findings are consistent with the fact that the nervous system continues to develop
throughout childhood.
Potentially At-Risk Populations
The NAAQS are intended to protect public health with an adequate margin of safety. In
so doing, protection is provided for both the population as a whole and those groups at
increased risk for health effects in response to the air pollutant for which each NAAQS is
set. Children are at increased risk for the effects of Pb exposure. Among children, the
youngest age groups were observed to be most at risk of elevated blood Pb levels, with
levels decreasing with increasing age of the children. Evidence related to childhood and
other at-risk factors is described in Section 1.9.6.
Pb Concentrations Corresponding to Ecological Effects
There is limited evidence to relate ambient air concentrations of Pb to levels of deposition
onto terrestrial and aquatic ecosystems and to subsequent movement of atmospherically-
deposited Pb through environmental compartments (e.g., soil, sediment, water, and biota)
(Section 1.9.7). The contribution of atmospheric Pb to specific sites is not clear, and the
connection between air concentration of Pb and ecosystem exposure continues to be
poorly characterized. Furthermore, the level at which Pb elicits a specific effect is
difficult to establish in terrestrial and aquatic systems, due to the influence of other
environmental variables (e.g., pH, organic matter) on both Pb bioavailability and toxicity,
and also to substantial species differences in Pb sensitivity. Current evidence indicates
that Pb is bioaccumulated in biota; however, the sources of Pb in biota have only been
identified in a few studies, and the relative contribution of Pb from all sources is usually
not known.
Summary
Overall, the evidence evaluated for the current review expands upon findings of the
2006 Pb AQCD and previous assessments, which concluded that there was a strong body
of evidence substantiating the health effects from Pb exposure as well as strong evidence
of the effects from Pb exposure on some ecological endpoints.
Pb exposure exerts harmful effects on a broad array of organ systems. Cognitive function
decrements in children are the effects that are best substantiated as occurring at the lowest
blood Pb concentrations (Section 1.6.1.1). There is also a strong body of evidence
demonstrating that Pb exposure can cause cardiovascular effects; this evidence strongly
suggests that long-term Pb exposure plays a role. Since Pb exposures were generally
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higher in the past than they are today, uncertainties exist regarding the relative
importance of recent versus past exposure in the development of the Pb-related health
effects in the adult populations studied.
With regard to the ecological effects of Pb, uptake of Pb into fauna and subsequent
effects on reproduction, growth and survival are established and are further supported by
more recent evidence. These may lead to effects at the population, community, and
ecosystem level of biological organization. In both terrestrial and aquatic organisms,
gradients in response are observed with increasing concentration of Pb and some studies
report effects within the range of Pb detected in environmental media over the past
several decades. Specifically, effects on reproduction, growth, and survival in sensitive
freshwater invertebrates are well-characterized from controlled studies at concentrations
at or near Pb concentrations occasionally encountered in U.S. fresh surface waters.
Hematological and stress related responses in some terrestrial and aquatic species were
also associated with elevated Pb levels in polluted areas. However, in natural
environments, modifying factors affect Pb bioavailability and toxicity and there are
considerable uncertainties associated with generalizing effects observed in controlled
studies to effects at higher levels of biological organization. Furthermore, available
studies on community and ecosystem-level effects are usually from contaminated areas
where Pb concentrations are much higher than typically encountered in the environment.
The contribution of atmospheric Pb to specific sites is not clear and the connection
between air concentration of Pb and ecosystem exposure continues to be poorly
characterized. Furthermore, the level at which Pb elicits a specific effect is difficult to
establish in terrestrial and aquatic systems, due to the influence of other environmental
variables (e.g., pH, organic matter) on both Pb bioavailability and toxicity, and also to
substantial species differences in Pb sensitivity.
References for Executive Summary
U.S. EPA (U.S. Environmental Protection Agency). (2006b). Air quality criteria for lead: Volume I of
II [EPA Report]. (EPA/600/R-05/144aF). Research Triangle Park, NC.
http ://cfpub. epa. gov/ncea/CFM/recordisplav. cfm?deid= 158823
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CHAPTER 1 INTEGRATIVE SUMMARY
1.1 ISA Development and Scope
This chapter summarizes and synthesizes the recently available scientific evidence and is
intended to provide a concise synopsis of the ISA conclusions and findings that best
inform the review of the current NAAQS for lead (Pb). The Integrated Review Plan (IRP)
for the National Ambient Air Quality Standards for Lead (U.S. EPA. 201 Ic) identifies a
series of policy-relevant questions (in Chapter 3 of the plan) that provide the framework
for this assessment. These questions also frame the entire review of the NAAQS for Pb,
and thus are informed by both science and policy considerations.
The ISA organizes, presents, and integrates the scientific evidence, which is considered,
along with findings from risk analyses and policy considerations, to help the U.S.
Environmental Protection Agency (EPA) address these questions during the NAAQS
review for Pb. The ISA includes:
• An integration of the evidence on the human health effects associated with Pb
exposure, a discussion of important uncertainties identified in the interpretation
of the scientific evidence, and an integration across different scientific
disciplines and across individual endpoints within major outcome categories.
• An integration of the evidence on the welfare effects1 of Pb in terrestrial,
freshwater and saltwater ecosystems, discussion of endpoints common to plants,
invertebrates and vertebrates and consideration of uncertainties in relating
atmospheric Pb concentrations to welfare effects.
• An integration of the effects associated with exposure to Pb across the scientific
disciplines for health and ecology, focusing on common modes of action.
• Discussion of policy relevant considerations, such as potentially at-risk
populations and concentration-response relationships.
EPA has a systematic process for evaluating the scientific evidence and for drawing
conclusions and judgments regarding the causal association of air pollution with health
and environmental effects. The ISA process includes literature search strategies, criteria
for selecting and evaluating studies, approaches for evaluating weight of the evidence,
and a framework for making causality determinations. As part of this process, the ISA is
reviewed by the public and peer reviewed by a formal panel of scientific experts (the
Clean Air Scientific Advisory Committee [CASAC]). The process and causality
1 Welfare effects as defined in Clean Air Act (CAA) Section 302(h) [42 U.S.C. 7602(h)] include, but are not limited
to, "effects on soils, water, crops, vegetation, man-made materials, animals, wildlife, weather, visibility and climate,
damage to and deterioration of property, and hazards to transportation, as well as effects on economic values and on
personal comfort and well-being."
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framework are described in more detail in the Preamble to the ISA. This section provides
a brief overview of the process for development of this ISA.
EPA initiated the current review of the Pb NAAQS in February 2010 with a call for
information from the public (75 FR 8934). In addition, literature searches were conducted
routinely to identify studies published since the last review, focusing on studies published
from 2006 (close of the previous scientific assessment) through September 2011.
References that were considered for inclusion or actually cited in this ISA can be found at
http://hero.epa.gov/lead.
This ISA evaluates relevant epidemiologic, animal toxicological, and welfare effects
studies, including those related to concentration-response relationships, mode(s) of action
(MOA), and susceptible populations. Additionally, air quality and emissions data, studies
on environmental fate and transport, and issues related to Pb toxicokinetics and exposure
were considered for inclusion in the document. Previous AQCDs (U.S. EPA. 2006b.
1986b. 1977) have included an extensive body of evidence on these topics. In this ISA,
the conclusions and key findings from previous reviews are summarized at the beginning
of each section, to provide the foundation for consideration of evidence from recent
studies. Results of key studies from previous reviews are included in discussions or tables
and figures, as appropriate, and conclusions are drawn based on the synthesis of evidence
from recent studies with the extensive literature summarized in previous reviews.
The Preamble discusses the general framework for conducting the science assessment
and developing an ISA, including criteria for selecting studies for inclusion in the ISA
evaluating and integrating the scientific evidence and developing scientific conclusions.
In selecting the studies for inclusion in the Pb ISA, particular emphasis is placed on those
studies most relevant to the review of the NAAQS.
In drawing judgments regarding causality for the criteria air pollutants, evidence of health
effects in the range of relevant pollutant exposures or doses is considered. Evidence from
experimental animal studies observing effects at biomarker levels comparable to, or
somewhat above, those currently experienced by the U.S. general population were
emphasized. Generally studies with dietary exposures resulting in blood Pb levels within
one order of magnitude above the upper end of the distribution of U.S. blood Pb levels
were considered in forming conclusions1 with the majority of studies reporting blood Pb
levels below 30 (ig/dL. Studies with higher blood Pb levels were considered if they
informed the evaluation of modes of action, mechanisms, or kinetics. For toxicological
studies where blood Pb levels were not measured, judgments regarding how to
1 For example, the 95th percentile of the NHANES (2009-2010) distribution of blood Pb level in children 1-5 years
old is 3.4 ng/m3 (CDC. 2013): however, the proportion of individuals with blood Pb levels that exceed this
concentration varies depending on factors including age and sex (Section 3.4).
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distinguish high from the more relevant low doses were made considering the range of
doses across the available body of evidence and emphasizing studies at the lower end of
the range.
With respect to the epidemiologic evidence, population-based studies using Pb
biomarkers (i.e. blood or bone Pb concentrations) were emphasized with the recognition
that many of the U.S populations studied included individuals with higher past than
recent Pb exposures. For example, in U.S. population studies during years past (1968-
1980) when air concentrations in the U.S. were much higher than they are today, the
population geometric mean blood Pb levels were roughly an order of magnitude above
current population geometric mean blood Pb levels (Sections 3.4. 5.1. and 4.4.1). Recent
occupational studies of populations with relatively high mean blood Pb levels were
considered insofar as they addressed a topic area that was of particular relevance to the
NAAQS review (e.g., longitudinal studies designed to examine recent versus historical
Pb exposure).
Relevant concentrations for drawing causality judgments for the welfare effects of Pb
were determined considering the range of Pb concentrations in the environment and the
available evidence for concentrations at which effects were observed in plants,
invertebrates, and vertebrates. Effects observed at or near ambient Pb concentrations
measured in soil, sediment and water in the most recent available studies (Table 1-1)
were emphasized and studies generally within one to two orders of magnitude above the
reported range of these values were considered in the body of evidence for terrestrial,
freshwater and saltwater ecosystems. Studies at higher concentrations were used to the
extent that they informed modes of action and illustrated the wide range of sensitivity to
Pb across taxa.
The causal determinations for terrestrial, freshwater, and saltwater effects are divided into
two categories. The first category includes endpoints that are commonly used in
ecological risk assessment (reproduction, growth, and survival). Impacts on these
endpoints have the potential to lead to population-level (e.g., abundance, production,
extirpation), community-level (e.g., taxa richness, relative abundance) and ecosystem-
level effects (Anklev et al.. 2010; Suter et al. 2005). The second category includes
organism- and sub-organism-level responses such as physiological stress, hematological
effects, and neurobehavioral effects. As recognized in EPA's Framework for Ecological
Risk Assessment (U.S. EPA. 1992). and in the adverse outcome pathway (AOP)
framework (Anklev etal.. 2010) endpoints that are measured at one level of biological
organization may be related to an endpoint at a higher level. The AOP conceptual
framework was proposed to link mechanistic data from initiating events at the molecular
level through a series of higher order biological responses to growth, survival and
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reproductive endpoints that can be used in ecological risk assessment, i.e., at the
population level and higher. In the case of Pb, sub-organismal responses
(i.e., physiological stress, hematological effects) and organism-level responses
(neurobehavioral alterations) may decrease the overall fitness of an organism, even
though their connection to effects at higher levels of biological organization may not
have been characterized. Furthermore, the effects of Pb on ecosystems necessarily begin
with some initial effects at the molecular level of specific organisms within the
ecosystem (U.S. EPA. 1986b). There are many different molecular and cellular level
effects, and toxicity of Pb in ecosystems may be attained through multiple modes of
action.
The ISA considers evidence of health effects for both short- and long-term pollutant
exposures. Since biomarkers are typically used as an index of exposure or dose in
epidemiologic studies, there is uncertainty regarding the timing, frequency, level, and
duration of the exposure(s) associated with the observed effects and blood Pb (or other
biomarker) levels measured in these studies. Some animal toxicological studies provide
evidence to inform the exposure patterns that can induce effects in animals and these
studies are drawn upon to interpret the human health effects evidence. Exposure regimens
used in toxicological studies typically include chronic exposure (i.e., over 10% of the
lifespan of the animal), long-term exposure (e.g., greater than 4 weeks in rodents) and
acute or short-term exposure (e.g., less than 4 weeks in rodents). For the purpose of this
assessment, short-term human exposures are generally defined to include exposures of
months (e.g.,
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remain. A conclusion of "likely causal" may be appropriate if evidence is available from
multiple studies or several lines of evidence including cases where the weight of the
health effects evidence is largely derived from multiple animal toxicological studies. A
conclusion that the evidence is suggestive of a causal relationship reflects generally
limited evidence, but may include "at least one high-quality epidemiology study" or "a
well conducted toxicological study". Evidence is inadequate to determine whether a
causal relationship exists when the available studies are of insufficient quantity, quality,
consistency, or statistical power. If several adequate studies, covering the full range of
exposure levels that human beings are known to encounter, considering at-risk
populations, are mutually consistent in not showing an effect, the relationship may be
judged not likely to be causal.
Beyond judgments regarding causality are questions re levant to quantifying health or
environmental risks based on the understanding of the quantitative relationships between
pollutant exposures and health or welfare effects. Once a determination is made regarding
the causal relationship between the pollutant and outcome category, important questions
regarding quantitative relationships include:
• What is the concentration-response, exposure-response, or dose-response
relationship in the human population?
• What exposure conditions (dose or exposure, exposure pathways, duration and
pattern) are important?
• What populations and lifestages appear to be differentially affected i.e., at
greater risk of Pb-related health effects?
• What elements of the ecosystem (e.g., types, regions, taxonomic groups,
populations, functions, etc.) appear to be affected or are more sensitive to
effects?
This ISA is composed of a Preamble, a Legislative and Historical Background, an
Executive Summary, and six chapters. Chapter 1 presents an Integrative Summary.
Chapter 2 highlights key concepts or issues relevant to understanding the sources,
ambient concentrations, and fate and transport of Pb in the environment. Chapter 3
summarizes key concepts and recent findings on Pb exposures, toxicokinetics, and
biomarkers reflecting Pb exposure and body burden. Chapter 4 presents a discussion of
the MOA of Pb and evaluates and integrates epidemiologic and animal toxicological
information on health effects related to Pb exposure. Chapter 5 summarizes the evidence
on potentially at-risk populations. Chapter 6 evaluates welfare effects evidence that is
relevant to the review of the secondary NAAQS for Pb.
This chapter summarizes and integrates the newly available scientific evidence that best
informs consideration of the policy-relevant questions that frame this assessment. The
organization of this chapter generally follows the organization of the document as a
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whole, with several additional sections including: a discussion of the assessment
development and scope (Section 1.1); an integration of the evidence across the disciplines
of health and ecology (Section 1.8); a discussion of policy-relevant considerations
(Section 1.9); and, an overall summary (Section 1.10).
1.2 Ambient Pb: Source to Concentration
1.2.1 Sources, Fate and Transport of Ambient Pb
The findings of this review build upon those from the 2006 Pb AQCD (U.S. EPA.
2006b), which documented the decline in ambient air Pb emissions following the phase
out of alkyl-Pb additives for on-road gasoline and reductions in industrial facility
emissions of Pb. Pb emissions declined by 98% from 1970 to 1995 and then by an
additional 77% from 1995 to 2008. The 2008 National Emissions Inventory (NEI)
reported ambient air Pb emissions of 950 tons. Air Pb emissions represent just a small
fraction (by weight) of the Pb processed in U.S. Pb-related industries.
As at the time of the last review, the 2008 NEI (U.S. EPA, 201 la), indicates that
piston-engine aircraft emissions comprise the largest share (58%) of total atmospheric Pb
emissions in the U.S. Other sources of ambient air Pb, beginning with the next largest,
include metal working and mining, fossil fuel combustion, other industrial sources, and
miscellaneous sources. On a site-specific basis, emissions are greatest at metal industry
sites. Over the period 1991-2010, the amount of Pb used in secondary Pb processing
increased by 37%. Exports of Pb increased by 103%, with 2010 exports sent to Mexico as
refined Pb; to Canada, China, and Japan in spent Pb-acid batteries; and, to the Republic
of Korea as Pb in concentrate (USGS. 2012).
Global atmospheric Pb deposition peaked in the 1970s, followed by a decline
(Section 2.2). Pb deposition is greater near Pb emission sources. Both wet and dry
deposition are important mechanisms for removing Pb from the atmosphere, and the
atmosphere is the main environmental transport media for Pb which is deposited onto
surface water and soil. Wet deposition is more important for the fine fraction while the
coarse fraction is usually removed by dry deposition. Pb associated with coarse PM
deposits to a great extent near local industrial sources, contributing to soil Pb
concentrations in those locations, while fine Pb-bearing PM can be transported long
distances, contributing Pb contamination in remote areas. Depending on local conditions,
once they are deposited, particles may be resuspended and redeposited before reaching a
site where further transport is unlikely, especially for dry deposition (Section 2.3).
Surface waters act as an important reservoir, with Pb lifetimes in the water column
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largely controlled by deposition and resuspension of Pb in sediments. Substantial
amounts of Pb may be input to surface waters and sediments by wastewater discharges
and through transport of Pb from vehicle wear and building materials in runoff waters
without having become airborne. Pb containing sediment particles can be remobilized
into the water column (Section 2.3).
1.2.2 Monitoring and Concentrations of Ambient Air Pb
The indicator for the Pb NAAQS is Pb in total suspended particles (Pb-TSP). The Federal
Reference Method (FRM) for Pb-TSP specifies that ambient air is drawn through a
high-volume TSP sampler onto a glass fiber filter. The Pb-TSP sampler's size selective
performance is known to be affected by wind speed and direction, and collection
efficiency has been demonstrated to decline with increasing particle size with an
uncertain upper size limit (Wedding etal.. 1977). There have been only a few studies
since the publication of the 2006 Pb AQCD with regard to sampling error in the Pb-TSP
FRM or alternatives to the existing Pb-TSP sampling technology. In addition to monitors
used historically for sampling Pb-PM, several single stage and multi-stage impactors and
inlets used for sampling PM concentrations are also potential options for Pb-PM
monitoring when the majority of particles are smaller than 15 pirn. Ambient air Pb
monitoring requirements have undergone several changes since publication of the
2006 Pb AQCD. The current Pb monitoring network design requirements include two
types of FRM monitoring sites: source-oriented and non-source-oriented (Section 2.4).
For the purpose of analyzing data for the ISA, monitors reporting to the U.S. EPA Air
Quality System (AQS) database were considered to be source-oriented if they were
designated in AQS as source-oriented, or if they were located within 1 mile of a 0.5 ton
per year or greater source, identified using emissions estimates in the 2005 or 2008 NEI
(U.S. EPA. 2008a) (U.S. EPA. 201 la). Source-oriented FRM Pb-TSP monitoring sites
are required near sources of air Pb emissions which are expected to or have been shown
to contribute to ambient air Pb concentrations in excess of the NAAQS.
Non-source-oriented FRM (Pb-TSP or Pb-PM 10) monitoring is also required at national
core multipollutant monitoring network (NCore) sites in Core Based Statistical Areas
(CBSA) with a population of at least 500,000. In addition to FRM monitoring, Pb is also
routinely measured in smaller particle fractions in the chemical speciation network
(CSN), interagency monitoring of protected visual environment (IMPROVE), and the
national air toxics trends station (NATTS) networks. While monitoring in multiple
networks provides extensive geographic coverage, measurements between networks are
not directly comparable in all cases because there are differences in the methods,
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including the different particle size ranges sampled in the different networks. Depending
on monitoring network, Pb is monitored in TSP, PMi0, or PM2 5.
Ambient air Pb concentrations have declined drastically over the period 1980-2010
(Section 2.5). The median value (across monitoring sties) for the maximum 3-month
average concentration within a year has dropped by 97% from 0.87 ug/m3 in 1980 to
0.03 ug/m3 in 2010. The mean of maximum 3-month average Pb concentrations at
source-oriented sites was skewed toward the 75th percentile of the data distribution and
exceeded the level of the NAAQS, suggesting that ambient air Pb concentrations are high
near a subset of industrial sources of airborne Pb. Studies in the peer-reviewed literature
have shown slightly elevated Pb concentrations downwind of industrial sources and
airports. Estimates for the natural background Pb concentrations from sources including
volcanoes, sea-salt spray, and biogenic sources are -0.00002 to 0.001 ug/m3.
The size distribution of Pb-bearing PM has changed over time and varies by site
(Section 2.5.3). Recent study results indicate that the size distribution has generally
shifted upward since the 1980s, with the mode of the size distribution of Pb-PM particles
now falling between 2.5 um and 10 um (Cho et al., 2011). The Pb-PM size distribution
depends on whether there are contributions from industrial sources or near-road
environments. In contrast to Cho et al. (2011). analysis of the distributional properties of
the Pb-PM measured by the AQS monitors, which are often sited near sources, suggests
that the largest proportion of particles is still below 2.5 um in diameter.
1.2.3 Ambient Pb Concentrations in Non-Air Media and Biota
Releases of Pb to the atmosphere have contributed to measurable increases in Pb in rain,
snowpack, soil, surface waters, sediments, agricultural plants, livestock, and wildlife
across the world, with highest concentrations near Pb sources, such as smelters. After the
phase-out of Pb from on-road gasoline and with reductions in industrial emissions, Pb
concentrations have decreased considerably in rain, snowpack, and surface waters.
Declining Pb concentrations in tree foliage, trunk sections, and grasses, as well as surface
sediments and soils, have also been observed (U.S. EPA. 2006b).
Often, Pb is retained in soils and sediments, where it provides a historical record of
deposition. In remote lakes, sediment profiles indicate higher Pb concentrations in near
surface sediment as compared to pre-industrial era sediment from greater depth and
indicate peak concentrations between 1960 and 1980 (when leaded on-road gasoline was
at peak use). Concentrations of Pb in moss, lichens, peat, and aquatic bivalves have been
used to understand spatial and temporal distribution patterns of air Pb concentrations.
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Ingestion and water intake are the major routes of Pb exposure for aquatic organisms, and
food, drinking water, and inhalation are major routes of exposure for livestock and
terrestrial wildlife.
Overall, Pb concentrations have decreased substantially in media through which Pb is
rapidly transported, such as air and water. Substantial Pb remains in soil and sediment
sinks. In areas less affected by major local sources, the highest concentrations are below
the surface layers and reflect the phase-out of Pb from on-road gasoline and emission
reductions from other sources.
Information on ambient Pb concentrations in non-air media and biota is reported in
Section 2.6, and concentrations considered in the interpretation of the ecological evidence
are tabulated in Table 1-1. As noted in the Preamble, the ecological causal determinations
focus on studies where effects of Pb exposure are observed at or near ambient levels of
Pb and studies generally within the range of one to two orders of magnitude above
current or ambient conditions were also considered in the body of evidence.
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Table 1-1 Pb concentrations in non-air media and biota considered for
ecological assessment.
Media
Soil (non-urban)
Freshwater
Sediment
Saltwater
Sediment
Fresh Surface
Water
(Dissolved Pb)b
Saltwater0
Vegetation
Pb Concentration
Contiguous U.S. Median: 15 mg Pb/kg (dry weight)
Contiguous U.S. 95th Percentile:
50 mg Pb/kg (dry weight)
National Average: 18.9 mg Pb/kg (dry weight)
Range of state averages:
5-38.6 mg Pb/kg (dry weight)
Median: 73 mg Pb/kg (dry weight)
Median: 28 mg Pb/kgb (dry weight)
Range: 0.6 to 1,050 mg Pb/kga
Median: 0.50 ug Pb/Lb;
Max: 30 ug Pb/L, 95th percentile 1.1 ug Pb/L
Range: 0.0003-0.075 ug Pb/L
(Set of National Parks in western U.S.)
Range: 0.01-27 ug Pb/L
Lichens: 0.3-5 mg Pb/kg (dry weight)
(Set of National Parks in western U.S.)
Grasses:
Geometric Mean: 0.31 kg Pb/kg (dry weight)
Years Data
Obtained
1961-1976
1961-1997
1996-2001
1991-2003
Dates not
available
1991-2003
2002-2007
Dates not
available
2002-2007
1980s-2000s
References
Shaklette (1984)
U.S. EPA
(2007d, 2006b,
2003b)
Mahler et al.
(2006)
U.S. EPA
(2006b)
Sadiq (1992)
U.S. EPA
(2006b)
Field and
Sherrell (2003),
U.S. National
Park Service
(2011)
Sadiq (1992)
U.S. National
Park Service
(2011)
Vandenhove et
al. (2009)
Vertebrates
Fish:
Geometric Mean: 0.59 mg Pb/kg (dry weight)
(whole fish)
Geometric Mean: 0.15 mg Pb/kg (dry weight) (liver)
Range: 0.08-22.6 mg Pb/kg (dry weight) (whole fish)
Range: 0.01-12.7 mg Pb/kg (dry weight) (liver)
1991-2003
Fish (from a set of national parks in western U.S.):
0.0033 (fillet) to 0.97 (liver) mg Pb/kg (dry weight)
Moosed'e: 0.008-0.029 mg Pb/kg (dry weight) (meat)
Moosed'e: 0.012-0.023 mg Pb/kg (dry weight) (liver)
2002-2007
U.S. EPA
(2006b)
U.S. National
Park Service
(2011)
aNo information available regarding wet or dry weight
"Based on synthesis of National Water-Quality Assessment (NAWQA) data reported in 2006 Pb AQCD (U.S. EPA, 2006b)
°Data from a combination of brackish and marine saltwater samples. In general, Pb in seawater is higher in coastal areas and
estuaries since these locations are closer to sources of Pb contamination and loading from terrestrial systems.
d The reference cited and its source citations show that observations date from studies published in 1977-1990, indicating that the
data were obtained no later than those years. Further, these measurements seem to be for non-U.S. locations, including the max,
which is well above other reported values in these refs.
eThree moose in one Alaskan park
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1.3 Exposure to Ambient Pb
Human Pb exposure is difficult to assess because Pb has multiple sources in the
environment and passes through various media (Section 3.1). Air-related pathways of Pb
exposure are the focus of this assessment. In addition to inhalation of Pb in ambient air,
air-related Pb exposure pathways include inhalation and ingestion of Pb in indoor dust
and/or outdoor soil that originated from recent or historic ambient air (e.g., air Pb that has
penetrated into the residence either via the air or tracking of soil), ingestion of Pb in
drinking water drawn from surface water contaminated from atmospheric deposition or
contaminated from surface runoff of deposited Pb, and ingestion of Pb in dietary sources
after uptake by plants or grazing animals. Soil can act as a reservoir for deposited Pb
emissions. Exposure to soil contaminated with deposited Pb can occur through
resuspended PM as well as hand-to-mouth contact, which is the main pathway of
childhood air-related exposure to Pb. The primary contribution of ambient air Pb to
young children's blood Pb concentrations is generally due to ingestion of Pb following its
deposition in soils and dusts rather than inhalation of ambient air (Section 3.1.1.2).
Non-ambient air-related exposures include hand-to-mouth contact with dust or chips of
peeling Pb-containing paint, or ingestion of Pb in drinking water conveyed through Pb
pipes. Several study results indicate that Pb-containing paint in the home and home age
(often a surrogate for the presence of Pb paint) are important residential factors that
increase risk of elevated blood Pb (Sections 1.9.6 and 5.2.6). Most Pb biomarker studies
do not indicate species or isotopic signature. As a consequence, non-air exposures are
reviewed in this section, because they can also contribute to Pb body burden.
A number of monitoring and modeling techniques have been employed for ambient Pb
exposure assessment. Environmental Pb concentration data can be collected from
ambient air Pb monitors, soil Pb samples, dust Pb samples, and dietary Pb samples to
estimate human exposure. Exposure estimation error depends in part on the collection
efficiency of these methods; collection efficiency for ambient air Pb FRM samplers is
described in Section 2.4. Models, such as the Integrated Exposure Uptake Biokinetic
(IEUBK) model, simulate human exposure to Pb from multiple sources and through
various routes including inhalation and ingestion. IEUBK model inputs include soil-Pb
concentration, air-Pb concentration, dietary-Pb intake including drinking water, Pb-dust
ingestion, human activity, and biokinetic factors. The relative contribution from specific
exposure pathways (e.g., water, diet, soil, ambient air) to blood Pb concentrations is
situation specific. Measurements and/or assumptions can be utilized when formulating
the model inputs; errors in measurements and assumptions thus have the potential to
propagate through exposure models.
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The size distribution of dust particles containing Pb differs from the size distribution of
inhalable ambient Pb-bearing PM (Sections 2.5 and 3.1). Airborne particles containing Pb
tend to be small (much of the distribution <10 (im) compared with soil or dust particles
containing Pb (-50 (im to several hundred (im). Ingestion through hand-to-mouth contact
is the predominant exposure pathway for the larger particles in soil and dust containing
Pb.
1.4 Toxicokinetics
The majority of Pb in the body is found in bone (roughly 90% in adults, 70% in children);
only about 1% of Pb is found in the blood. Pb in blood is primarily (-99%) bound to red
blood cells (RBCs). It has been suggested that the small fraction of Pb in plasma (<1%)
may be the more biologically labile and lexicologically active fraction of the circulating
Pb. The relationship between Pb in blood and plasma is pseudo-linear at relatively low
daily Pb intakes (i.e., <10 ug/kg per day) and at blood Pb concentrations <25 (ig/dL, and
becomes curvilinear at higher blood Pb concentrations due to saturable binding to RBC
proteins. As blood Pb level increases and the higher affinity binding sites for Pb in RBCs
become saturated, a larger fraction of the blood Pb is available in plasma to distribute to
brain and other Pb-responsive tissues. See Section 3.2 for additional details.
The burden of Pb in the body may be viewed as divided between a dominant slow
(i.e., uptake and elimination) compartment (bone) and smaller fast compartment(s) (soft
tissues). Pb uptake to and elimination from soft tissues is much faster than in bone. Pb
accumulates in bone regions undergoing the most active calcification at the time of
exposure. During infancy and childhood, bone calcification is most active in trabecular
bone (e.g., patella); whereas, in adulthood, calcification occurs at sites of remodeling in
cortical (e.g., tibia) and trabecular bone (Aufderheide and Wittmers. 1992). A high bone
formation rate in early childhood results in the rapid uptake of circulating Pb into
mineralizing bone; however, in early childhood bone Pb is also recycled to other tissue
compartments or excreted in accordance with a high bone resorption rate (O'Flahertv.
1995). Thus, much of the Pb acquired early in life is not permanently fixed in the bone.
The exchange of Pb from plasma to the bone surface is a relatively rapid process. Pb in
bone becomes distributed in trabecular and the more dense cortical bone. The proportion
of cortical to trabecular bone in the human body varies by age, but on average is about
80% cortical to 20% trabecular. Of the bone types, trabecular bone is more reflective of
recent exposures than is cortical bone due to the slow turnover rate and lower blood
perfusion of cortical bone. Some Pb diffuses to kinetically deeper bone regions where it
is relatively inert, particularly in adults. These bone compartments are much more labile
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in infants and children than in adults as reflected by half-times for movement of Pb from
bone into the plasma (e.g., cortical half-time = 0.23 years at birth, 3.7 years at 15 years of
age, and 23 years in adults; trabecular half-time = 0.23 years at birth, 2.0 years at
15 years of age, and 3.8 years in adults) (Leggett 1993). See Section 3.2 for additional
details.
Evidence for maternal-to-fetal transfer of Pb in humans is derived from cord blood to
maternal blood Pb ratios (i.e., cord blood Pb concentration divided by mother's blood
Pb). Group mean ratios range from about 0.7 to 1.0 at the time of delivery for mean
maternal blood Pb levels ranging from 1.7 to 8.6 (ig/dL. Transplacental transfer of Pb
may be facilitated by an increase in the plasma/blood Pb concentration ratio during
pregnancy. Maternal-to-fetal transfer of Pb appears to be related partly to the
mobilization of Pb from the maternal skeleton. See Section 3.2.2.4 for additional details.
The dominant elimination phase of Pb kinetics in the blood, exhibited shortly after a
change in exposure occurs, has a half-life of-20-30 days. An abrupt change in Pb uptake
gives rise to a relatively rapid change in blood Pb, to a new quasi-steady state, achieved
in -75-100 days (i.e., 3-4 times the blood elimination half-life). A slower phase of Pb
clearance from the blood may become evident with longer observation periods following
a decrease in exposure due to the gradual redistribution of Pb among bone and other
compartments. See Section 3.3 for additional details.
1.5 Pb Biomarkers
Overall, blood Pb levels have been decreasing among U.S. children and adults over the
past 35 years (Section 3.4). The median blood Pb level for the entire U.S. population is
1.1 ug/dL and the 95th percentile blood Pb level is 3.3 ug/dL, based on the 2009-2010
National Health and Nutrition Examination Survey (NHANES) data (CDC. 2013).
Among children aged 1-5 years, the median and 95th percentiles were slightly higher, at
1.2 ug/dL and 3.4 ug/dL, respectively.
Blood Pb is dependent on both the recent exposure history of the individual, as well as
the long-term exposure history that determines body burden and Pb in bone. The
contribution of bone Pb to blood Pb changes, depending on the duration and intensity of
the exposure, age, and various other physiological stressors (e.g., nutritional status,
pregnancy, menopause, extended bed rest, hyperparathyroidism) that may affect bone
remodeling, which normally and continuously occurs. In children, largely due to faster
exchange of Pb to and from bone, blood Pb is both an index of recent exposure and
potentially an index of body burden. In adults and children whose exposure to Pb has
effectively ceased or greatly decreased, there is a rapid decline in blood Pb over the first
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few months followed by a more gradual, slow decline in blood Pb concentrations over the
period of years due to the gradual release of Pb from bone. Bone Pb is an index of
cumulative exposure and body burden. Even bone compartments should be recognized as
reflective of differing exposure periods with Pb in trabecular bone exchanging more
rapidly than Pb in cortical bone with the blood. Consequently, Pb in cortical bone is a
better marker of cumulative exposure, while Pb in trabecular bone is more likely to be
correlated with blood Pb, even in adults. See Section 3.3 for additional details.
Sampling frequency is an important consideration when evaluating blood Pb and bone Pb
levels in epidemiologic studies, particularly when the exposure is not well characterized.
It is difficult to determine what blood Pb is reflecting in cross-sectional studies that
sample blood Pb once, whether recent exposure or movement of Pb from bone into blood
from historical exposures. In contrast, cross-sectional studies of bone Pb and longitudinal
samples of blood Pb concentrations overtime provide more of an index of cumulative
exposure and are more reflective of average Pb body burdens overtime. The degree to
which repeated sampling will reflect the actual long-term time-weighted average blood
Pb concentration depends on the sampling frequency in relation to variability in
exposure. High variability in Pb exposures can produce episodic (or periodic) oscillations
in blood Pb concentration that may not be captured with low sampling frequencies.
Furthermore, similar blood Pb concentrations in two individuals (or populations),
regardless of their age, do not necessarily translate to similar body burdens or similar
exposure histories.
The concentration of Pb in urine follows blood Pb concentration. There is added
complexity with Pb in urine because concentration is also dependent upon urine flow
rate, which requires timed urine samples that is often not feasible in epidemiologic
studies. Other biomarkers have been utilized to a lesser extent (e.g., Pb in teeth). See
Section 3.3.
1.6 Health Effects
This section summarizes and evaluates the evidence from toxicological and
epidemiologic studies of the health effects associated with Pb exposure and integrates
that evidence across these disciplines. The coherence of the findings from experimental
animal and epidemiologic studies, including evidence for potential MOA, is evaluated to
establish biological plausibility and address uncertainties in the epidemiologic evidence
due to biases from factors such as reverse causality and confounding. Both short- and
long-term Pb exposures are considered (Section 1.1): information on the frequency,
timing, level and duration of exposure in animal toxicological studies is used to inform
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the interpretation of epidemiologic studies regarding the relevant patterns of exposure
that are likely to be associated with the health effects.
The health evidence is organized into groups of related endpoints (e.g., cognitive
function, externalizing behaviors, neurodegenerative diseases). This evidence is
considered in combination with the evidence from other fields (e.g., toxicokinetics) and
weighed against the attributes described in the framework for causal determination
(Table II of the Preamble) to draw conclusions regarding the causal relationship between
Pb exposure and the health effects evaluated in this assessment. A more detailed
discussion of the underlying evidence used to formulate each causal determination can be
found in Chapter 5 of this document. Table 1-2 summarizes the conclusions formed
regarding the causal relationships between Pb exposure and health effects.
Table 1-2 Summary of causal determinations for the relationship between
exposure to Pb and health effects.
Causality Determination3
Health Outcome (Table with Key Evidence)
Nervous System Effects (Section 4.3.15)
Children - Nervous System Effects
Cognitive Function Decrements Causal Relationship (Table 4-17)
Clear evidence of cognitive function decrements (as measured by Full Scale IQ, academic performance, and
executive function) in young children (4 to 11 years old) with mean or group blood Pb levels measured at
various lifestages and time periods between 2 and 8 ug/dL. Clear support from animal toxicological studies that
demonstrate decrements in learning, memory, and executive function with dietary exposures resulting in
relevant blood Pb levels of 10-25 ug/dL. Plausible MOAs are demonstrated.
Hyperac.M.y" •
Clear evidence of attention decrements, impulsivity and hyperactivity (assessed using objective
neuropsychological tests and parent and teacher ratings) in children 7-17 years and young adults ages 19-20
years. The strongest evidence for blood Pb-associated increases in these behaviors was found in prospective
studies examining prenatal (maternal or cord), age 3-60 months, age 6 years, or lifetime average (to age 11-13
years) mean blood Pb levels of 7 to 14 ug/dL and groups with early childhood (age 30 months) blood Pb levels
>10 ug/dL. Biological plausibility is provided by animal toxicological studies demonstrating impulsivity or
impaired response inhibition with relevant prenatal, lactational, post-lactational and lifetime Pb exposures.
Plausible MOAs are demonstrated.
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Table 1-2 (Continued): Summary of causal determinations for the relationship between
exposure to Pb and health effects.
Causality Determination3
Health Outcome (Table with Key Evidence)
and Young Adults
^ '
Prospective epidemiologic studies find that early childhood (age 30 months, 6 years) or lifetime average (to age
11-13 years) blood Pb levels or tooth Pb levels (from ages 6-8 years) are associated with criminal offenses in
young adults ages 19-24 years and with higher parent and teacher ratings of behaviors related to conduct
disorders in children ages 8-17 years. Pb-associated increases in conduct disorders were found in populations
with mean blood Pb levels 7 to14 ug/dL; associations with lower blood Pb levels as observed in cross-sectional
studies were likely to be influenced by higher earlier Pb exposures. There is coherence in epidemiologic findings
among related measures of conduct disorders. Evidence of Pb induced aggression in animals was mixed, with
increases in aggression found in some studies of adult animals with gestational plus lifetime Pb exposure but
not juvenile animals. The lack of clear biological plausibility produces some uncertainty.
Internalizing Behaviors Likely Causal Relationship (Table 4-17)
Prospective epidemiologic studies find associations of higher lifetime average blood (mean: -14 ug/dL) or
childhood tooth (from ages 6-8 years) Pb levels with higher parent and teacher ratings of internalizing behaviors
such as symptoms of depression or anxiety, and withdrawn behavior in children ages 8-1 3 years. Consideration
of potential confounding by parental caregiving was not consistent and findings from cross-sectional studies in
populations ages 5 and 7 years with mean blood Pb levels of 5 ug/dL were mixed. Animal toxicological studies
demonstrate depression-like behaviors and increases in emotionality with relevant lactational exposures.
Plausible MOAs are demonstrated.
Auditory Function Decrements Likely Causal Relationship (Table 4-17)
A prospective epidemiologic study and large cross-sectional studies indicate associations between blood Pb
levels and increased hearing thresholds at ages 4-19 years. Across studies, associations were found with blood
Pb levels measured at various time periods, including prenatal maternal, neonatal (10 day, mean 4.8 ug/dL),
lifetime average, and concurrent (ages 4-19 years) blood Pb levels (median 8 ug/dL). Plausible MOAs are
demonstrated. The lack of biological plausibility in animals with relevant exposures produces some uncertainty.
Visual Function Decrements Inadequate to Infer a Causal Relationship (Table 4-17)
The available epidemiologic and toxicological evidence is of insufficient, quantity, quality and consistency.
Motor Function Decrements Likely Causal Relationship (Table 4-17)
Prospective epidemiologic studies provide evidence of associations of fine and gross motor function decrements
in children ages 4-17 years with lifetime average blood Pb levels and with blood Pb levels measured at various
time periods with means generally ranging from 4.8 to 12 ug/dL. Results were inconsistent in cross sectional
studies with concurrent blood Pb level means 2-5 ug/dL. Limited evidence in animal toxicological studies with
relevant Pb exposures.
Adults - Nervous System Effects
Cognitive Function Decrements Likely Causal Relationship (Table 4-17)
Prospective studies indicate associations of higher baseline bone Pb levels with declines in cognitive function
(executive function, visuospatial skills, learning and memory) in adults (>age 50 years) over 2- to 4-year
periods. Cross-sectional studies provide additional support. Uncertainties remain regarding the timing,
frequency, duration and level of the Pb exposures contributing to the effects observed and residual confounding
by age. Biological plausibility is provided by findings that relevant lifetime Pb exposures from gestation, birth, or
after weaning induce learning impairments in adult animals and by evidence demonstrating plausible MOAs.
Psychopathological Effects Likely Causal Relationship (Table 4-17)
Cross-sectional studies in a few populations demonstrate associations of higher concurrent blood or tibia Pb
levels with self-reported symptoms of depression and anxiety in adults. Uncertainties remain regarding the
timing, frequency, duration and level of Pb exposures contributing to the observed associations and residual
confounding by age. Observations of depression-like behavior in animals with dietary lactational Pb exposure,
with some evidence at relevant blood Pb levels, and evidence demonstrating plausible MOAs in experimental
animals provides support.
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Table 1-2 (Continued): Summary of causal determinations for the relationship between
exposure to Pb and health effects.
Causality Determination3
Health Outcome (Table with Key Evidence)
Auditory Function Decrements Suggestive of a Causal Relationship (Table 4-17)
A high-quality prospective epidemiologic study finds associations of higher tibia Pb level with a greater rate of
elevations in hearing threshold over 20 years. Some evidence indicates effects on relevant MOAs but important
uncertainties remain related to effects on auditory function in animals with relevant Pb exposures.
Visual Function Decrements Inadequate to Infer a Causal Relationship (Table 4-17)
The available epidemiologic and toxicological evidence is of insufficient, quantity, quality and consistency.
Neurodegenerative Diseases Inadequate to Infer a Causal Relationship (Table 4-17)
The available epidemiologic and toxicological evidence is of insufficient, quantity, quality and consistency.
Cardiovascular Effects (Section 4.4.7)
Hypertension Causal Relationship (Table 4-24)
Prospective epidemiologic studies with adjustment for multiple potential confounders consistently find
associations of blood and bone Pb levels with hypertension incidence and increased blood pressure (BP) in
adults. Cross-sectional studies provide supporting evidence. Meta-analyses underscore the consistency and
reproducibility of the Pb associated increase in blood pressure and hypertension (a doubling of concurrent blood
Pb level (between 1 and 40 ug/dL) is associated with a 1 mmHg increase in systolic BP); however, uncertainties
remain regarding the timing, frequency, duration and level of Pb exposures contributing to the effects observed in
epidemiologic studies. Experimental animal studies demonstrate effects on BP after long-term Pb exposure
resulting in mean blood Pb levels of 10 ug/dL or greater. Plausible MOAs are demonstrated.
Subclinical Atherosclerosis Suggestive of a Causal Relationship (Table 4-24)
Cross-sectional analyses of NHANES data find associations of blood Pb level with peripheral artery disease
(PAD) in adults. Animal toxicological evidence is limited to studies of MOA (oxidative stress, inflammation,
endothelial cell dysfunction) that demonstrate biologically plausible mechanisms through which Pb exposure may
initiate atherosclerotic vessel disease.
Coronary Heart Disease Causal Relationship (Table 4-24)
Prospective epidemiologic studies consistently find associations of Pb biomarkers with cardiovascular mortality
and morbidity, specifically myocardial infarction (Ml), ischemic heart disease (IHD), or HRV; however,
uncertainties remain regarding the timing, frequency, duration and level of Pb exposures contributing to the
effects observed in epidemiologic studies. Thrombus formation was observed in animals after relevant long term
exposure and MOAs (hypertension, decreased HRV, increased corrected QT (QTc) interval, and corrected QRS
complex (QRSc) duration in electrocardiogram [ECG]) are demonstrated in humans and animals.
Cerebrovascular Disease Inadequate to Infer a Causal Relationship (Table 4-24)
The available epidemiologic and toxicological evidence is of insufficient, quantity, quality, and/or consistency.
Plausible MOAs, which are shared with hypertension and atherosclerosis, are demonstrated.
Renal Effects (Section 4.5.5)
Reduced Kidney Function Suggestive of a Causal Relationship (Table 4-31)
Multiple high quality epidemiologic studies provide evidence that Pb exposure is associated with reduced kidney
function; however, uncertainty remains regarding the potential for reverse causality to explain findings in humans.
Further, inconsistencies and limitations in occupational studies, epidemiologic studies of children and clinical trials
of chelation of CKD patient preclude strong inferences to be drawn based on their results. Although longitudinal
studies found Pb-associated decrements in renal function in populations with mean blood Pb levels of 7 and 9
ug/dL, the contributions of higher past Pb exposures cannot be excluded. Animal toxicological studies
demonstrate Pb-induced kidney dysfunction at blood Pb levels greater than 30 ug/dL; however, evidence in
animals with blood Pb levels < 20 ug/dL is generally not available. At blood Pb levels between 20 and 30 ug/dL
studies provide some evidence for dysfunction in kidney function measures (e.g., decreased creatinine clearance,
increased serum creatinine, increased BUN). Plausible MOAs (Pb induced hypertension, renal oxidative stress
and inflammation, morphological changes, and increased uric acid) are demonstrated.
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Table 1-2 (Continued): Summary of causal determinations for the relationship between
exposure to Pb and health effects.
Causality Determination3
Health Outcome (Table with Key Evidence)
Immune System Effects (Section 4.6.8)
Atopic and Inflammatory Responses Likely Causal Relationship (Table 4-34)
Prospective studies of children ages 1-5 years indicate associations of prenatal cord and childhood blood Pb
levels with asthma and allergy. This evidence is supported by cross-sectional associations between higher
concurrent blood Pb levels (>10 ug/dL) in children and higher IgE. Uncertainties related to potential confounding
by SES, smoking or allergen exposure are reduced through consideration of the evidence from experimental
animal studies. The biological plausibility for the effects of Pb on IgE is provided by consistent findings in animals
with gestational or gestational-lactational Pb exposures, with some evidence at blood Pb levels relevant to
humans. Strong evidence of Pb-induced increases in Th2 cytokine production and inflammation in animals
demonstrates MOA.
Decreased Host Resistance Likely Causal Relationship (Table 4-34)
Animal toxicological studies provide the majority of the evidence for Pb-induced decreased host resistance.
Dietary Pb exposure producing relevant blood Pb levels (7-25 ug/dL) results in increased susceptibility to bacterial
infection and suppressed delayed type hypersensitivity. Further, evidence demonstrating plausible MOA,
including suppressed production of Th1 cytokines and decreased macrophage function in animals, provides
coherence.
Autoimmunity Inadequate to Infer a Causal Relationship (Table 4-34)
The available toxicological and epidemiologic studies do not sufficiently inform Pb-induced generation of auto-
antibodies with relevant Pb exposures.
Hematologic Effects (Section 4.7.4)
** ^^ ^ ^ ^^ *"* Causal Relationship (Table 4-35)
Animal toxicological studies demonstrate that exposures resulting in blood Pb levels relevant to humans (2-7
ug/dL) alter several hematological parameters (Hemoglobin [Hb], Hematocrit [Hct], and mean corpuscular volume
[MCV]), increase measures of oxidative stress and increase cytotoxicity in red blood cell (RBC) precursor cells.
Limited body of epidemiologic studies provides additional support for the association of Pb exposure with these
endpoints. Plausible MOAs are demonstrated in experimental animals.
Altered Heme Synthesis Causal Relationship (Table 4-35)
Consistent findings from studies in experimental adult animal studies report that relevant exposures (e.g. blood
Pb levels of 6.5 ug/dL) cause decreased ALAD and ferrochelatase activities. Additional support is garnered from a
larger body of ecotoxicological studies demonstrating decreased ALAD activity across a wide range of species
and a limited body of epidemiologic studies. Plausible MOAs are demonstrated in experimental animals.
Reproductive and Developmental Effects (Section 4.8.5)
Development Causal Relationship (Table 4-48)
Multiple cross-sectional epidemiologic studies report associations between concurrent blood Pb levels and
delayed pubertal onset for girls (6-18 years) and boys (8-15 years). These associations are consistently observed
in populations with concurrent blood Pb levels 1 .2-9.5 ug/dL. Few studies consider confounding by nutrition.
Uncertainties remain regarding the timing, frequency, duration and level of Pb exposures contributing to the
effects observed in epidemiologic studies of older children. Experimental animal studies demonstrate delayed
onset of puberty in female pups with blood Pb levels of 1.3-13 ug/dL and delayed male sexual maturity at blood
Pb levels of 34 ug/dL.
abortion?0™5 (6'9" '°W birth ^^ SP°ntaneOUS Suggestive of Causal Relationship (Table 4-48)
Some well-conducted epidemiologic studies report associations of maternal Pb biomarkers or cord blood Pb with
preterm birth and low birth weight/fetal growth; however, the epidemiologic evidence is inconsistent overall and
findings from experimental animal studies are mixed.
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Table 1-2 (Continued): Summary of causal determinations for the relationship between
exposure to Pb and health effects.
Causality Determination3
Health Outcome (Table with Key Evidence)
Male Reproductive Function Causal Relationship (Table 4-48)
Key evidence is provided by toxicological studies in rodents, non-human primates, and rabbits showing
detrimental effects on semen quality, sperm and fecundity/fertility with supporting evidence in epidemiologic
studies. Toxicological studies with relevant Pb exposure routes leading to blood Pb concentrations ranging from
5-43 ug/dL reported effects on sperm quality and sperm production rate, sperm DMA damage, and histological or
ultrastructural damage to the male reproductive organs. Consistent associations in studies of occupational
populations with concurrent blood Pb levels of 25 ug/dL and greater, report detrimental effects of Pb on sperm;
however, uncertainties remain regarding the timing, frequency, duration and level of Pb exposures contributing to
the effects observed in epidemiologic studies.
Female Reproductive Function Suggestive of Causal Relationship (Table 4-48)
Although findings are mixed overall, the body of evidence include some high-quality epidemiologic and
toxicological studies, suggesting that Pb may affect some aspects of female reproductive function (hormone level,
placental pathology).
Cancer (Section 4.10.5)
Cancer Likely Causal Relationship (Table 4-50)
The animal toxicological literature provides the strong evidence for long-term exposure (i.e., 18 months or 2
years) to high concentrations of Pb (> 2,600 ppm) inducing tumor development; findings from epidemiologic
studies inconsistent. Plausible MOAs are demonstrated.
a In drawing conclusions regarding the causal relationship between Pb exposure and human health effects, evidence in the range of
relevant pollutant exposures or biomarker levels was considered. Specifically, population-based epidemiology studies were
emphasized with the recognition that many of the U.S populations studied included individuals with higher past than recent Pb
exposures. Evidence from toxicological studies of effects observed in experimental animals at biomarker levels (e.g. blood Pb)
comparable to those currently experienced by the U.S. general population were emphasized. Generally, studies with dietary
exposures resulting in blood Pb levels within one order of magnitude above the upper end of the distribution of U.S. blood Pb levels
were considered in forming concusions, with the majority of studies reporting blood Pb levels below 30 ug/dL. Studies with higher
blood Pb levels were considered if they informed the evaluation of MOA, mechanisms, or kinetics. (Preamble. Section 1.1).
b Within the attention deficit hyperactivity disorder domain of externalizing behaviors, studies of Pb exposure have focused primarily
on attention, impulsivity, and hyperactivity. Because the studies of ADHD were limited in terms of their design and did not
adequately consider potential confounding by factors such as SES, parental education, or parental caregiving quality, they were not
a major consideration in drawing conclusions about the relationship between Pb exposure and attention, impulsivity, and
hyperactivity.
0 Two domains of conduct disorders,(i.e., undersocialized aggressive conduct disorder and socialized aggressive conduct disorder),
are combined for the purpose of this assessment because it is difficult to differentiate between these two domains in the available
epidemiologic studies, which examine multiple endpoints such as delinquent behavior, aggression, antisocial behavior. Criminal
offenses are included in the evaluation because they can be predicted by earlier conduct disorders (Section 4.3.3.2).
d There was limited evaluation of potential confounding by parental psychopathology, which is a strong risk factor for externalizing
behaviors, in the majority of the epidemiologic studies; however, evidence of an association of between psychopathology in parents
and Pb exposure in their children is not available (Section 4.3.3).
e Strong evidence from experimental animal studies reduces uncertainty related to confounding generally.
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1.6.1 Nervous System Effects
The collective body of epidemiologic and toxicological evidence integrated across that
reviewed in the 2006 Pb AQCD (U.S. EPA. 2006b) coupled with recently available data
demonstrates the effects of Pb exposure on a range of nervous system effects. In children,
these effects include cognitive function (Sections 4.3.2.1. 4.3.2.2. 4.3.2.3. 4.3.2.4.
4.3.2.5). externalizing behaviors (Section 4.3.3). internalizing behaviors (Section 4.3.4).
auditory function (Section 4.3.6.1). visual function (Section 4.3.6.2). and motor function
(Section 4.3.7). In adults, nervous system effects examined in relation to Pb exposure
include cognitive function (Section 4.3.2.7). psychopathological effects (Section 4.3.5).
auditory function (Section 4.3.6.1). visual function (Section 4.3.6.2). and
neurodegenerative diseases (Section 4.3.9).
1.6.1.1 Children
Cognitive Function Decrements
Multiple prospective studies conducted in diverse populations consistently demonstrate
associations of higher blood and tooth Pb levels with lower full scale IQ (FSIQ),
executive function, and academic performance and achievement. Most studies examined
representative populations and had moderate to high follow-up participation without
indication of selective participation among children with higher blood Pb levels and
lower cognitive function. Associations between blood Pb level and cognitive function
decrements were found with adjustment for several potential confounding factors, most
commonly, socioeconomic status (SES), parental IQ, parental education, and parental
caregiving quality. In children ages 4-11 years, associations were found with prenatal,
early childhood, childhood average, and concurrent blood Pb levels in populations with
mean or group blood Pb levels in the range 2-8 (ig/dL (Section 4.3.2). Neither
epidemiologic nor toxicological evidence has identified an individual critical lifestage or
duration of Pb exposure within childhood that is associated with cognitive function
decrements. Several epidemiologic studies found a supralinear concentration-response
relationship. A threshold for cognitive function decrements is not discernable from the
available evidence (i.e., examination of early childhood blood Pb or concurrent [with
peak <10 (ig/dL] blood Pb in the range of < 1 to 10 (ig/dL). Evidence in children was
clearly supported by observations of Pb-induced impairments in learning, memory, and
executive function in juvenile animals. Several studies in animals indicated learning
impairments with prenatal, lactational, post-lactational and lifetime (with or without
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prenatal) Pb exposures that resulted in blood Pb levels of 10-25 (ig/dL. The mode of
action for Pb-associated cognitive function decrements is supported by observations of
Pb-induced impairments in neurogenesis, synaptogenesis and synaptic pruning, long term
potentiation, and neurotransmitter function in the hippocampus, prefrontal cortex, and
nucleus accumbens. The associations consistently found for FSIQ and other measures of
cognitive function in prospective studies of children with adjustment for SES, parental
education, and parental caregiving quality and the biological plausibility provided by
evidence in animals for impairments in learning, memory, and executive function with
relevant Pb exposures and evidence describing modes of action is sufficient to conclude
that there is a causal relationship between Pb exposure and decrements in cognitive
function in children.
Externalizing Behaviors
There are three domains of externalizing behaviors (Section 4.3.3). These domains are
attention deficit hyperactivity disorder, undersocialized aggressive conduct disorder, and
socialized aggressive conduct disorder. Studies of the effect of Pb exposure on the
domain of attention deficit hyperactivity disorder have focused primarily on attention,
impulsivity, and hyperactivity not diagnosis of ADHD. For the purpose of this
assessment, the two domains of conduct disorders are combined because it is difficult to
differentiate between these two domains based on the available epidemiologic studies of
Pb exposure, which examine multiple endpoints such as delinquent behavior, aggression,
and antisocial behavior. Criminal offenses are included within domain because they can
be predicted by earlier conduct disorders.
Attention, Impulsivity and Hyperactivity
Although examined less extensively than cognitive function, several prospective studies
demonstrated associations of blood or tooth Pb levels measured years before outcomes
with attention decrements and hyperactivity in children 7-17 years and young adults ages
19-20 years as assessed using objective neuropsychological tests and rated by parents and
teachers. Most of these prospective studies examined representative populations without
indication of participation conditional on blood Pb levels and behavior. The results from
prospective studies were adjusted for potential confounding by SES as well as parental
education and caregiving quality, with some studies also considering parental cognitive
function, birth outcomes, substance abuse, and nutritional factors. In prospective studies,
blood Pb-associated attention decrements and hyperactivity were found in populations
with prenatal (maternal or cord), age 3-60 month average, age 6 year, or lifetime average
(to age 11-13 years) mean blood Pb levels of 7 to 14 (ig/dL and groups with age
30 month blood Pb levels >10 (ig/dL. Most well-conducted cross-sectional studies that
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examined several potential confounding factors found associations of attention
decrements, impulsivity, and hyperactivity in children ages 5-7.5 years with concurrent
blood Pb levels with means of 5-5.4 (ig/dL but cannot establish temporality or exclude
the influence of higher blood Pb levels earlier in childhood. Biological plausibility for
observations in children is provided by several findings in animals for increases in
impulsivity or impaired response inhibition with relevant post-weaning and lifetime Pb
exposures that resulted in blood Pb levels of 11 to 30 (ig/dL. The mode of action for
Pb-associated attention decrements, impulsivity, and hyperactivity is supported by
observations of Pb-induced impairments in neurogenesis, synaptic pruning, and
dopamine transmission in the prefrontal cerebral cortex, cerebellum, and hippocampus.
The consistency of epidemiologic evidence for attention decrements and hyperactivity
from prospective studies and the biological plausibility provided by evidence for
Pb-induced impulsivity in animals and for underlying modes of action is sufficient to
conclude that there is a causal relationship between Pb exposure and effects on attention,
impulsivity, and hyperactivity in children.
Conduct Disorders in Children and Young Adults
Prospective studies consistently indicate that earlier childhood (e.g., age 30 months, 6
years) or lifetime average (to age 11-13 years) blood Pb levels or tooth (from ages 6-8
years) Pb levels are associated with criminal offenses in young adults ages 19-24 years
and with higher parent and teacher ratings of behaviors related to conduct disorders in
children ages 7-17 years. Pb-associated increases in conduct disorders were found in
populations with mean blood Pb levels 7-14 (ig/dL. Associations with lower blood Pb
levels that are not influenced by higher earlier Pb exposures are not well characterized.
These associations were found without indication of strong selection bias and with
adjustment for SES, parental education and IQ, parental caregiving quality, family
functioning, smoking, and substance abuse. Supporting evidence is provided by cross-
sectional evidence of children participating in NHANES and a meta-analysis of
prospective and cross-sectional studies. Additionally, there is coherence in epidemiologic
findings among related measures of conduct disorders. Evidence for Pb-induced
aggression in animals is mixed with increases in aggression found in some studies of
adult animals with gestational plus lifetime Pb exposure but not juvenile animals. The
consistent epidemiologic evidence from prospective and cross-sectional studies for
criminal offenses and ratings of behaviors related to conduct disorders but uncertainty
due to lack of clear evidence for aggression in animals is sufficient to conclude that a
causal relationship is likely to exist between Pb exposure and conduct disorders in
children and young adults.
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Internalizing Behaviors
Prospective studies in a few populations demonstrate associations of higher lifetime
average blood (mean: ~14 (ig/dL) or childhood tooth (from ages 6-8 years) Pb levels with
higher parent and teacher ratings of internalizing behaviors such withdrawn behavior and
symptoms of depression and anxiety in children ages 8-13 years. The lack of selective
participation by blood Pb level and associations found with parental and teacher ratings
do not provide strong indication of biased reporting of behaviors for children with higher
blood Pb levels. The few cross-sectional associations in populations with mean
concurrent blood Pb levels of 5 (ig/dL were inconsistent. Pb-associated increases in
internalizing behaviors were found with adjustment for maternal education and SES-
related variables. Consideration for potential confounding by parental caregiving quality
was inconsistent. Despite some uncertainty in the epidemiologic evidence, the biological
plausibility for the effects of Pb on internalizing behaviors is provided by a few findings
in animals with dietary lactational Pb exposure, with some evidence at blood Pb levels
relevant to humans. Additional toxicological evidence supports modes of action,
including Pb-induced changes in the HPA axis and dopaminergic and GABAergic
systems. The evidence from prospective studies in a few populations of children and the
coherence with evidence from a few animal studies with relevant Pb exposures and mode
of action but some uncertainty related to potential confounding by parental caregiving
quality in studies of children is sufficient to conclude that a causal relationship is likely to
exist between Pb exposure and internalizing behaviors in children.
Auditory Function Decrements
Evidence from a prospective study and cross-sectional studies in a few U.S. populations
of children indicates associations of higher blood Pb level with increases in hearing
thresholds and decreases auditory processing or auditory evoked potentials with
adjustment for potential confounding by SES in most studies and by child health and
nutritional factors in some studies. The high participation rates, particularly in the
prospective study with follow-up from birth, reduce the likelihood of biased participation
by children with higher blood Pb levels. Across studies, associations were found with
blood Pb levels measured at various time periods, including prenatal maternal, neonatal
(10 day, mean 4.8 (ig/dL), lifetime average (to age 5 years), and concurrent (ages 4-19
years) blood Pb levels (median 8 (ig/dL). Evidence for Pb-associated increases in hearing
thresholds or latencies of auditory evoked potentials was found in adult monkeys with
lifetime dietary Pb exposure. However, these effects in adult animals were found with
higher peak or concurrent blood Pb levels (i.e., 33-150 (ig/dL) than those relevant to this
ISA; thus, the biological plausibility for epidemiologic observations is unclear. The
evidence in children, particularly that from a prospective study and observations of
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decreased auditory evoked potentials in animals indicating a possible mode of action, but
uncertainties related to effects on auditory function in juvenile animals with relevant Pb
exposures, is sufficient to conclude that a causal relationship is likely to exist between Pb
exposure and decrements in auditory function in children.
Visual Function Decrements
A study in children and a few studies in animals show Pb-associated increases in
supernormal electroretinograms, the biological relevance of which is unclear. Because the
available epidemiologic and toxicological evidence is of insufficient quantity, quality,
and consistency, the evidence is inadequate to determine that a causal relationship exists
between Pb exposure and visual function in children.
Motor Function Decrements
Evidence from prospective studies in a few populations indicates associations of
decrements in fine and gross motor function with higher neonatal, concurrent, and
lifetime average blood Pb levels in children ages 4.5-6 years and with higher earlier
childhood (ages 0-5 year average, age 78 months) blood Pb levels in children ages 15-17
years. The means for these blood Pb metrics ranged from 4.8 to 12 (ig/dL. These
associations were found with adjustment for several potential confounding factors,
including SES, parental caregiving quality, and child health and without indication of
substantial selection bias. Evidence from cross-sectional studies was less consistent. The
biological plausibility for associations observed in children is provided by a study that
found poorer balance in male mice with relevant gestational to early postnatal (PND10)
Pb exposures. The limited available evidence in a few prospective cohorts of children,
but uncertainty because of the limited available findings in mice with relevant exposures
is sufficient to conclude that a causal relationship is likely to exist between Pb exposure
and decrements in motor function in children.
1.6.1.2 Adults
Cognitive Function
In adults without occupational exposure, recent prospective studies in the NAS and BMS
cohorts indicate associations of higher baseline tibia (means 19, 20 (ig/g) or patella (mean
25 (ig/g) Pb levels with declines in cognitive function in adults (>age 50 years) over 2- to
4-year periods. While the specific covariates differed between studies, these bone
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Pb-associated cognitive function decrements were found with adjustment for potential
confounding factors such as age, education, SES, current alcohol use, and current
smoking. Supporting evidence is provided by cross-sectional analyses of the NAS, BMS,
and the Nurses' Health Study, which found stronger associations with bone Pb level than
concurrent blood Pb level. Cross-sectional studies also considered more potential
confounding factors, including dietary factors, physical activity, medication use, and
comorbid conditions. The multiple exposures and health outcomes examined in many
studies reduces the likelihood of biased participation specifically by adults with higher Pb
exposure and lower cognitive function. The collective evidence indicates associations in
cohorts of white men and women and a cohort of more ethnically diverse men and
women. The specific timing, frequency, duration, and magnitude of Pb exposures
contributing to the associations observed with bone Pb levels are not discernable from the
evidence. Also, there is potential for residual confounding by age. The effects of recent
Pb exposures on cognitive function decrements in adults were indicated in Pb-exposed
workers by associations found with blood Pb levels, although these studies did not
consider potential confounding by other workplace exposures. The biological plausibility
for the effects of Pb exposure on cognitive function decrements in adults is provided by
findings that relevant lifetime Pb exposures from gestation, birth, or after weaning induce
learning impairments in adult animals and by evidence for the effects of Pb altering
neurotransmitter function in hippocampus, prefrontal cortex, and nucleus accumbens. The
associations between bone Pb level and cognitive function decrements consistently found
in the few prospective and cross-sectional studies of adults without occupational Pb
exposure, the coherence with animal findings, and toxicological evidence supporting
modes of action but uncertainties related to potential residual confounding by age in
epidemiologic studies are sufficient to conclude that a causal relationship is likely to exist
between long-term cumulative Pb exposure and cognitive function decrements in adults.
Psychopathological Effects
Cross-sectional studies in a few populations demonstrate associations of higher
concurrent blood or tibia Pb levels with self-reported symptoms of depression and
anxiety in adults. The examination of multiple exposures and outcomes in the available
studies does not provide strong indication of biased reporting of psychopathological
effects specifically by adults with higher Pb exposures. In adults, Pb-associated increases
in depression and anxiety were found with adjustment for age, SES, and in the NAS,
daily alcohol intake. The biological plausibility for epidemiologic evidence is provided
by observations of depression-like behavior in animals with dietary lactational Pb
exposure, with some evidence at relevant blood Pb levels and by toxicological evidence
supporting modes of action, including Pb-induced changes in the HPA axis and
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dopaminergic and GABAergic systems. The associations of blood and bone Pb level with
self-reported psychopathological effects found in the few studies of adults without
occupational Pb exposure, the biological plausibility provided by the coherence with
findings from a few animal studies and evidence for underlying modes of action, but
uncertainties related to residual confounding of bone Pb results by age in epidemiologic
studies are sufficient to conclude that a causal relationship is likely to exist between Pb
exposure and psychopathological effects in adults.
Auditory Function Decrements
The the evidence provided by the analysis of NAS men for associations of higher tibia Pb
level with a higher rate of elevations in hearing threshold over 20 years and observations
of decreased auditory evoked potentials in animals indicating a possible mode of action
but uncertainties related to effects on auditory function in adult animals with relevant Pb
exposures, is suggestive of a causal relationship between Pb exposure and auditory
function decrements in adults.
Visual Function Decrements
Studies in adult animals show differential effects on ERGs, depending on the timing and
concentration of exposure. A case control study finds higher Pb in retinal tissue from
macular degeneration cases but lacks rigorous statistical analysis and examination of
potential confounding. Because the available epidemiologic and toxicological evidence is
of insufficient quantity, quality, and consistency, the evidence is inadequate to determine
that a causal relationship exists between Pb exposure and visual function in adults.
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Neurodegenerative Disease
While evidence is inconclusive for Amyotrophic Lateral Sclerosis (ALS)and Alzheimer's
disease, a few case-control studies each found higher blood Pb levels in adults with
essential tremor and higher bone Pb levels in adults with Parkinson's disease. Because of
the inconclusive evidence for some diseases and limitations such as reverse causation for
essential tremor and the lack of consideration for potential confounding by manganese
(Mn) exposure for both essential tremor and Parkinson's disease, the evidence is
inadequate to determine that a causal relationship exists between Pb exposure and
neurodegenerative diseases.
1.6.2 Cardiovascular Effects
The 2006 Pb AQCD (U.S. EPA. 2006b) concluded that there was a relationship between
higher blood Pb and bone Pb and cardiovascular effects in adults, in particular increased
BP and increased incidence of hypertension, and recent evidence strengthens this
conclusion. For the evaluation of causal relationships with Pb exposure, evidence was
grouped in categories using the U.S. Surgeon General's Report on Smoking as a
guideline (CDC. 2004). In addition to hypertension (Section 4.4.7.1) these categories
include subclinical atherosclerosis (Section 4.4.7.2). coronary heart disease
(Section 4.4.7.3). and cerebrovascular disease (Section 4.4.7.4). Examination of measures
of subclinical atherosclerosis provides the opportunity to assess the pathogenesis of
vascular disease at an earlier stage. Studies that examine markers of subclinical
atherosclerosis (such as PAD [i.e., ankle-brachial index]) and generalized atherosclerosis
(i.e., IMT) are included in this category. Coronary heart disease (CHD) results from
interruption of the blood supply to a part of the heart resulting from atherosclerosis of the
coronary arteries, with acute injury and scarring leading to permanent damage to the heart
muscles. A disrupted HRV has been associated with a higher mortality after MI and is
used as a predictor of the physiological processes underlying CHD (Buccelletti et al..
2009). Studies that examine incidence of MI, IHD, HRV, and mortality from CHD, MI,
or IHD are included in this category. Cerebrovascular disease describes a group of
conditions involving the cerebral blood vessels that result in transient or permanent
disruption of blood flow to the brain. These conditions include stroke, transient ischemic
attack, and subarachnoid hemorrhage. Both hypertension and atherosclerosis are risk
factors for cerebrovascular disease and the mechanisms for these outcomes also apply to
cerebrovascular disease.
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Hypertension
Prospective epidemiologic studies consistently find associations of blood and bone Pb
levels with hypertension incidence and increased blood pressure (BP). These findings
have been replicated across multiple high-quality studies comprising large and diverse
populations. Further support is provided by multiple cross-sectional analyses. While the
adjustment for specific factors varied by study, the collective body of epidemiologic
evidence included adjustment for multiple potential key confounding factors, including
age, diet, sex, body mass index (BMI), blood pressure lowering medication use, SES,
race/ethnicity, alcohol consumption, cholesterol, smoking, pre-existing disease
(i.e., diabetes), measures of renal function, and co-exposures (i.e., Cd) thus reducing the
uncertainties related to confounding bias. Meta-analyses underscore the consistency and
reproducibility of the Pb-associated increases in BP and hypertension (Navas-Acien et
al.. 2008; Nawrot et al.. 2002). Nawrot et al. (2002) found that a doubling of concurrent
blood Pb level (between 1 and >40 ug/dL) was associated with a 1 mmHg increase in
systolic BP and a 0.6 mmHg increase in diastolic BP. Navas-Acien et al. (2008) found an
association of higher bone Pb levels with increased BP in their pooled analysis. Although
recent epidemiologic studies in adults report associations in populations with relatively
low mean concurrent blood Pb levels, the majority of individuals in these adult
populations were likely to have had higher levels of Pb exposure earlier in life. Thus,
there is uncertainty concerning the specific Pb exposure level, timing, frequency, and
duration contributing to the associations observed in the epidemiologic studies. A causal
relationship of Pb exposure with hypertension is supported by evidence from
experimental animal studies that demonstrate effects on BP after long-term Pb exposure
resulting in mean blood Pb levels of greater than 10 (ig/dL or greater (Figure 4-21).
Further, there was no evidence of a threshold below which no significant association of
blood Pb level with BP was observed among the NHANES II population, 20-74 years old
between 1976 and 1980 with concurrent blood Pb levels ranging from 7-34 (ig/dL
(Schwartz. 1991). Relevant mode of action is demonstrated and coherence for the
evidence of the effect of Pb exposure on BP and hypertension is further provided by
epidemiologic evidence indicating associations with cardiovascular conditions related to
increased BP including mortality, CHD, stroke, and cardiac failure. Thus, the overall
evidence is sufficient to conclude that there is a causal relationship between Pb exposure
and hypertension.
Subclinical Atherosclerosis
A limited number of studies have evaluated markers of subclinical atherosclerosis
following Pb exposure in humans or animals. An NHANES analysis of adults (1999-
2000) found an association between concurrent blood Pb and PAD that was robust to
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adjustment for cadmium (Cd) and other potential confounding factors (Navas-Acien et
al.. 2004). A second more recent analysis with adjustment for numerous potential
confounders reported an increasing trend in the odds of PAD across concurrent blood Pb
level groups in adults within the NHANES population (Muntner et al.. 2005). Evidence
of plausible biological mechanisms (e.g., oxidative stress, inflammation, endothelial cell
dysfunction) is clearly described in the animal toxicological literature but animal studies
have provided only limited evidence to suggest Pb exposure may initiate atherosclerotic
vessel disease. Further, the specific Pb exposure level, timing, frequency, and duration
contributing to the observed association with PAD is not discernable from the
epidemiologic evidence. Thus, the combined limited epidemiologic and toxicological
evidence is suggestive of a causal relationship between Pb exposure and subclinical
atherosclerosis.
Coronary Heart Disease
Prospective epidemiologic studies of cohorts of adults during the period 1976-1994
consistently report that blood Pb level is associated with risk of CHD-related mortality
from cardiovascular disease, specifically MI, IHD, and HRV (Figure 4-29 and Table
4-23). Several recent studies also report associations between biomarkers of Pb and
incidence of CHD-related outcomes including a prospective analysis reporting an
increased incidence of IHD (physician confirmed MI, angina pectoris) with blood and
bone Pb levels. In addition, Weisskopf et al. (2009) found that patella bone Pb levels
were associated with increased mortality from IHD (similar magnitude non-statistically
significant associations were observed with tibia Pb levels) among men enrolled in the
NAS. The level, timing, frequency, and duration of Pb exposure contributing to CHD in
adult populations with higher past than recent exposure is not discernable from the
evidence. However, coherence for the associations in humans is supported by the
observation of thrombus formation in animals after long term Pb exposure
(Sections 4.4.7.3 and 4.4.3) and mode of action for Pb-induced CHD (i.e., hypertension,
HRV, increased corrected QT (QTc) interval, and corrected QRS complex (QRSc)
duration in electrocardiogram [ECG]) in humans and animals (Sections 4.4.2 and
4.4.3.4). The overall evidence is sufficient to conclude that there is a causal relationship
between Pb exposure and coronary heart disease.
Cerebrovascular Disease
Both hypertension and atherosclerosis are risk factors for cerebrovascular disease and the
mechanisms for these outcomes also apply to cerebrovascular disease. Despite strong
evidence for hypertension and CHD and Pb exposure, very few epidemiologic or
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toxicological studies have examined the effects of Pb exposure on cerebrovascular
disease (Section 4.4.7). These epidemiologic studies provide limited evidence for
increased risk of mortality from stroke and are insufficient evidence to inform the
presence or absence of a causal relationship between cerebrovascular disease and Pb
exposure. Thus, the current evidence is inadequate to determine that a causal relationship
exists between Pb exposure and cerebrovascular disease.
1.6.3 Renal Effects
Recent epidemiologic and toxicological studies evaluated in the current review support
and expand upon the strong body of evidence presented in the 2006 Pb AQCD (U.S.
EPA. 2006b) indicating that Pb exposure is associated with reduced kidney function
(Section 4.5.5). The causal determination for reduced kidney function is informed by
evidence for reduced GFR, reduced creatinine clearance, and increased serum creatinine.
Reduced Kidney Function
There are multiple high quality epidemiologic studies and clear biological plausibility but
uncertainty regarding the potential for reverse causality to explain findings in humans.
Among epidemiologic studies, key evidence is provided by the few available longitudinal
studies which belter characterized the temporal sequence between Pb exposure and
changes in renal function by showing associations between baseline bone or blood Pb
levels and reduced kidney function over time in men in the Boston, MA area and greater
progression of kidney disease in CKD patients. Evidence from longitudinal studies also
addressed the potential for reverse causality by showing the persistence of blood Pb-
associated decrements in kidney function in the range of normal kidney function and
demonstrated stronger associations than cross-sectional analyses of the same data. Cross-
sectional adult studies provide supportive evidence but are weighed less than the
prospective studies in conclusions because by design they do not inform directionality
(i.e., reverse causality). Inconsistencies were noted in occupational studies and studies of
children, and important study design limitations were noted in clinical trials of chelation
in CKD patients. These inconsistencies and limitations preclude strong inferences from
the results of these three study groups. Longitudinal studies found Pb-associated
decrements in renal function in populations with mean blood Pb levels of 7 and 9 (ig/dL.
However, the contributions of higher past Pb exposures cannot be excluded. Animal
toxicological studies provide clear biological plausibility with evidence for Pb-induced
kidney dysfunction at blood Pb levels greater than 30 (ig/dL; however, evidence in
animals with blood Pb levels < 20 (ig/dL is generally not available. Studies with mean
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blood Pb levels between 20 and 30 (ig/dL provide some evidence for dysfunction in
kidney function measures (e.g., decreased creatinine clearance, increased serum
creatinine, increased BUN). Animal studies also provide biological plausibility for the
associations observed between blood Pb levels and reduced kidney function with
evidence for Pb-induced hypertension, renal oxidative stress and inflammation,
morphological changes, and increased uric acid. Collectively, the evidence integrated
across epidemiologic and toxicological studies with uncertainties related to the potential
for reverse causation, is suggestive of a causal relationship between Pb exposures and
reduced kidney function among adults.
1.6.4 Immune System Effects
The cumulative body of epidemiologic and toxicological evidence from the
2006 Pb AQCD (U.S. EPA. 2006b) and the current assessment describes several effects
of Pb exposure on the immune system related to a shift from T-derived lymphocyte
helper (Th) 1 - to Th2 -type responses, including an increase in atopic and inflammatory
conditions and a decrease in host resistance (Section 4.6.8). Outcomes related to an
increase in atopic and inflammatory conditions (Section 4.6.8.1) include asthma, allergy,
increased IgE, and mode of action endpoints such as selective differentiation of Th2 cells,
increased production of Th2 cytokines, B cell activation, and inflammation. Outcomes
related to decreased host resistance (Section 4.6.8.2) include enhanced susceptibility to
bacterial and viral infection, suppressed delayed type hypersensitivity (DTH), and those
describing mode of action, i.e., decreased production of Thl cytokines, reduced
phagocyte function, and increased inflammation. A small body of studies indicates the
effects of Pb exposure on autoimmunity (Section 4.6.8.3).
Atopic and Inflammatory Conditions
Prospective studies in a few populations of children ages 1-5 years indicate associations
of asthma and allergy with prenatal cord blood Pb levels or blood Pb levels measured
sometime before the outcome, with a cross-sectional study providing supporting evidence
with associations with concurrent blood Pb level. Prospective design, lack of selective
participation of subjects, and objective assessment of outcomes reduce the likelihood of
undue selection bias and reverse causation. These few studies varied in their
consideration for potential confounding by SES and exposure to smoking or allergens.
Thus, uncertainty remains regarding residual confounding in associations observed
between blood Pb levels and asthma and allergy in children. The evidence for asthma and
allergy is supported by cross-sectional associations found between higher concurrent
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blood Pb levels in children and higher IgE, an important mediator of asthma and allergy.
The biological plausibility for the effects of Pb on IgE is provided by consistent findings
in animals with gestational or gestational-lactational Pb exposures, with some evidence at
blood Pb levels relevant to humans. In epidemiologic studies, higher IgE and higher
asthma prevalence were examined and found mostly in children with concurrent blood Pb
levels >10 (ig/dL. Coherence for the evidence of Pb-associated increases in asthma,
allergy, and IgE is found with evidence for most of the examined endpoints related to
mode of action, i.e., Pb-induced increases in Th2 cytokine production and inflammation
in animals. Neither toxicological nor epidemiologic evidence clearly identifies an
individual critical lifestage or duration of Pb exposure associated with atopic and
inflammatory conditions but points to an influence of gestational and cumulative Pb
exposures. The combined epidemiologic evidence in a few populations and toxicological
evidence supporting a relationship between Pb exposure and asthma, allergy, and shift to
a Th2 phenotype as an underlying mode action but some uncertainty regarding potential
confounding is sufficient to conclude that a causal relationship is likely to exist between
Pb exposures and an increase in atopic and inflammatory conditions.
Decreases in Host Resistance
Much of the evidence on decreased host resistance was available in the 2006 Pb AQCD
(U.S. EPA. 2006b) and is summarized in Section 4.6.5.1 (and Section 4.6.8.2). Animal
toxicological observations are the primary contributors to the evidence for Pb-induced
decreased host resistance. Several studies in rodents show that dietary Pb exposure
producing relevant blood Pb levels (7-25 (ig/dL) results in increased susceptibility to
bacterial infection and suppressed DTH. Further, coherence is found with evidence
describing modes of action, including suppressed production of Thl cytokines and
decreased macrophage function in animals. These effects were found with gestational,
lactational, adult-only, and lifetime Pb exposures of animals, without an individual
critical lifestage of exposure identified. A few cross-sectional epidemiologic studies
indicate Pb-associated increases in respiratory infections but limitations, including the
lack of rigorous methodology and consideration for potential confounding produce
uncertainty in the epidemiologic evidence for decreased host resistance in humans. The
consistent toxicological evidence in animals for increased susceptibility to bacterial
infection, suppressed DTH, and related modes of action but uncertainty in the
epidemiologic evidence in humans is sufficient to conclude that a causal relationship is
likely to exist between Pb exposure and decreased host resistance.
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Autoimmunity
Toxicological evidence describes the potential of Pb to increase autoimmunity, with a
few previous studies showing Pb-induced generation of auto-antibodies (Hudson et al..
2003; Bunn et al.. 2000; El-Fawal et al.. 1999; Waterman et al.. 1994) and recent studies
providing indirect evidence by showing formation of neoantigens that could result in the
formation of auto-antibodies (Table 4-34). Several observations were made in animals
injected with Pb, which is a route of exposure with less relevance to humans. Higher
levels of auto-antibodies also were found in Pb-exposed battery workers; however,
implications are limited because of the high blood Pb levels (range: 10-40 (ig/dL) of
some of the workers and lack of consideration for potential confounding by several
factors, including other occupational exposures (El-Fawal et al.. 1999). Because results
from both available toxicological and epidemiologic studies do not sufficiently inform
Pb-induced generation of auto-antibodies with relevant Pb exposures, the evidence is
inadequate to determine if there is a causal relationship between Pb exposure and
autoimmunity.
1.6.5 Hematological Effects
Recent toxicological and epidemiologic studies support evidence presented in in previous
assessments including the 2006 Pb AQCD (U.S. EPA. 2006b) describes the effect of
exposure to Pb on hematological outcomes such as RBC survival and function
(Section 4.7.4.1) and altered heme synthesis (Section 4.7.4.2) . Endpoints considered
within the category of RBC survival and function include alterations in multiple
hematological parameters (e.g., Hb, Hct, MCV), oxidative stress (altered antioxidant
enzyme activities, decreased cellular glutathione [GSH], and increased lipid
peroxidation), increased cytotoxicity in RBC precursor cells, and mode of action
endpoints such as decreased intracellular calcium concentrations [Ca2+]l5 decreased
adenosine-triphosphase (ATPase) activity, and increased phosphatidylserine expression.
Endpoints related to altered heme synthesis include decreased activities of ALAD and
ferrochelatase, and decreased levels of Hb.
Decreased Red Blood Cell Survival and Function
Experimental animal studies demonstrate that exposures via drinking water and gavage,
resulting in blood Pb levels relevant to humans, alter several hematological parameters,
increase measures of oxidative stress and increase cytotoxicity in red blood cell (RBC)
precursor cells. Some of these effects have been observed in animal toxicological studies
with exposures resulting in blood Pb levels 2-7 (ig/dL. Support for these findings is
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provided by biologically plausible modes of action including decreased intracellular
calcium concentrations [Ca2+]l5 decreased ATPase activity, and increased
phosphatidylserine expression. Epidemiologic studies find associations in both adults and
children of blood Pb levels with altered hematological endpoints, increased measures of
oxidative stress, and altered hematopoiesis. Although the majority of these epidemiologic
studies are limited by their lack of rigorous methodology, some studies in children did
adjust for some potential confounding factors including age, sex, mouthing behavior,
anemia, dairy product consumption, maternal age, education, employment, marital status,
family structure, SES-related variables, strengthening their support for findings in
experimental animals. Collectively, the strong evidence from toxicological studies that is
supported by findings from mode of action and epidemiologic studies is sufficient to
conclude that there is a causal relationship between Pb exposures and decreased RBC
survival and function.
Heme Synthesis
Altered heme synthesis is demonstrated by a small, but consistent, body of studies in
adult animals reporting that exposures via drinking water and gavage resulting in blood
Pb levels relevant to humans (e.g., 6.5 (ig/dL result in decreased ALAD and
ferrochelatase activities. Supporting this evidence is a larger body of ecotoxicological
studies that demonstrate decreased ALAD activity across a wide range of taxa exposed to
Pb. Epidemiologic studies find associations in both adults and children between higher
blood Pb levels and decreased ALAD and ferrochelatase activities. Although the majority
of these studies are limited by their lack of rigorous methodology and consideration for
potential confounding, some studies in children did adjust for or consider potential
confounding factors (i.e. age, sex, urban/rural residence, height, weight, BMI),
strengthening their support for the findings in the animal toxicological studies. Evidence
for altered heme synthesis is also provided by a large body of toxicological and
epidemiologic studies that report decreased Hb concentrations in association with Pb
exposure or blood Pb levels. Collectively, the strong evidence from toxicological and
ecotoxicological studies that is supported by findings from epidemiologic studies is
sufficient to conclude that there is a causal relationship between Pb exposures and altered
heme synthesis.
1.6.6 Reproductive and Developmental Effects
Many epidemiologic and toxicological studies of the effects of Pb on reproductive and
developmental outcomes have been conducted since the 2006 Pb AQCD. The evaluation
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of causal relationships with Pb exposure focuses on four areas: developmental effects
(Section 4.8.5.1). birth outcomes (Section 4.8.5.2). male reproductive function
(Section 4.8.5.3). and female reproductive function (Section 4.8.5.4).
Development
In cross-sectional epidemiologic studies of girls (ages 6-18 years) with mean and/or
median concurrent blood Pb levels from 1.2-9.5 (ig/dL consistent associations with
delayed pubertal onset (measured by age at menarche, pubic hair development, and breast
development) were observed. Although fewer studies were conducted in boys overall,
associations between blood Pb levels and delayed puberty onset in boys (ages 8-15 years)
were observed in cross-sectional and one longitudinal study (mean and/or median blood
Pb levels 3-9.5 (ig/dL). Potential confounders considered in the epidemiologic studies of
both boys and girls that performed regression analyses varied across studies, with few
studies considering confounding by nutritional factors. A limitation across most of the
epidemiologic studies of blood Pb levels and delayed puberty is their cross-sectional
design, which does not allow for an understanding of temporality. There is uncertainty
with regard to the exposure frequency, timing, duration, and level that contributed to the
associations observed in these studies. Experimental animal studies demonstrate that
puberty onset in both males and females is delayed with Pb exposure. A recent animal
study indicates that delayed pubertal onset may be one of the more sensitive
developmental effects of Pb exposure with effects observed after gestational exposures
leading to blood Pb levels in the female pup of 1.3-13 (ig/dL. Experimental animal
studies have also reported delayed male sexual maturity as measured with sex organ
weight, among other outcomes, seeing significant decrements at blood Pb levels of
34 (ig/dL. Findings from epidemiologic studies of the effect of Pb on postnatal growth
are inconsistent and findings from the toxicological literature of the effect of Pb exposure
are mixed with recent growth findings showing adult onset male obesity after gestational
and lactational Pb exposure. Toxicological studies demonstrated effects of Pb exposure
on other organ systems (effects on the eye, and alterations in the hematopoietic, hepatic
systems and teeth.) The collective body of evidence integrated across epidemiologic and
toxicological studies, based on the findings of delayed pubertal onset among males and
females, is sufficient to conclude that there is a causal relationship between Pb exposure
and developmental effects.
Birth Outcomes
Overall, epidemiologic studies of the association of various Pb exposure indicators with
preterm birth report inconsistent findings. A recent epidemiologic study reported no
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association between maternal blood Pb and neural tube defects. Some associations were
observed between Pb and low birth weight when epidemiologic studies used measures of
postpartum maternal bone Pb or air exposures. The associations were less consistent for
maternal blood Pb measured during pregnancy or at delivery or umbilical cord and
placenta Pb (maternal blood Pb or umbilical cord and placenta Pb were the biomarkers
most commonly used in studies of low birth weight) but some associations between
increased Pb biomarker levels and decreased low birth weight/fetal growth were
observed. The effects of Pb exposure during gestation in animal toxicological studies
included mixed findings with some studies showing reduction in litter size, implantation,
and birth weight, and some showing no effect. Because some associations were observed
in well-conducted epidemiologic studies of preterm birth and low birth weight/fetal
growth, the evidence is suggestive of a causal relationship between Pb exposure and birth
outcomes.
Male Reproductive Function
Key evidence is provided by toxicological studies in rodents, non-human primates, and
rabbits showing detrimental effects on semen quality, sperm and fecundity/fertility with
supporting evidence in epidemiologic studies of associations between blood Pb levels and
detrimental effects on sperm. Pb exposures resulting in blood Pb levels from 5-43 (ig/dL
induced lower sperm quality and sperm production rate, sperm DNA damage, and
histological or ultrastructural damage to the male reproductive organs. These effects were
found in animals exposed to Pb during peripuberty or as adults for 1 week to 3 months.
Pb exposure of male rats also was associated with subfecundity in female mates and
lower fertilization of eggs in vitro. Detrimental effects of Pb on sperm were observed in
epidemiologic studies with concurrent blood Pb levels of 25 (ig/dL and greater among
men occupationally exposed; however, these studies were limited because of their lack of
consideration of potential confounding factors, including occupational exposures other
than Pb. Additional epidemiologic studies among men with lower Pb biomarker levels
were limited to infertility clinic studies that may lack generalizability; however, a well-
conducted epidemiologic study that controlled for other metals as well as smoking
reported a positive association with various detrimental effects in sperm (Telisman et al.,
2007). The median concurrent blood Pb level in this study was 4.92 (ig/dL. The specific
timing, frequency, duration and level of Pb exposure associated with the blood Pb level
and effects observed is not discernable from the epidemiologic evidence, however. Mode
of action support is provided by several recent animal toxicological studies that showed
that Pb induced oxidative stress within the male sex organs, increase apoptosis of
spermatocytes and germ cells, and impaired germ cell structure and function. Based on
the consistency and coherence of findings for the detrimental effects of Pb exposure on
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sperm and semen in the toxicological literature, the support from epidemiologic studies,
and biological plausibility provided by mode of action evidence, the evidence is
sufficient to conclude that there is a causal relationship between Pb exposures and male
reproductive function.
Female Reproductive Function
Epidemiologic and toxicological studies of reproductive function among females
investigated whether Pb biomarker levels were associated with hormone levels, fertility,
estrus cycle changes, and morphology or histology of female reproductive organs
including the placenta (Section 4.8.5.4). Toxicological studies of experimental animals
reported in the 2006 Pb AQCD (U.S. EPA. 2006b) demonstrated associations between Pb
exposure and female reproductive function, although little evidence was provided by
epidemiologic studies. Some studies have shown associations with concurrent blood Pb
levels and altered hormone levels in adults, with inconsistency across studies may be due
to the different hormones examined and the different timing in the menstrual and life
cycles. There is some evidence of a potential relationship between Pb exposure and
female fertility, but findings are also mixed. The majority of the epidemiologic studies
are cross-sectional, and adjustment for potential confounders varies from study to study,
with some potentially important confounders, such as BMI, not included in all studies.
Also, most of the studies have small samples sizes and are generally of women attending
infertility clinics. Animal toxicological studies that employ relevant prenatal or early
postnatal Pb exposures observe that Pb contributes to placental pathology and
inflammation, decreased ovarian antioxidant capacity, altered ovarian steroidogenesis and
aberrant gestational hormone levels. Although epidemiologic and toxicological studies
provide information on different exposure periods, both types of studies support the
conclusion that Pb possibly affects at least some aspects of female reproductive function.
Overall, the relationship observed with female reproductive outcomes, such as fertility,
placental pathology, and hormone levels in some epidemiologic and toxicological studies,
is sufficient to conclude that evidence is suggestive of a causal relationship between Pb
exposure and female reproductive function.
1.6.7 Cancer
The toxicological literature provides the strong evidence for the effect of long-term
exposure (i.e., 18 months or 2 years) to high concentrations of Pb (> 2,600 ppm) on
cancer. The consistent evidence indicating Pb-induced carcinogenicity in animal models
is substantiated by the mode of action findings from multiple high-quality toxicological
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studies in animal and in vitro models from different laboratories. Based on such evidence,
IARC has classified inorganic Pb compounds as a probable human carcinogen and the
National Toxicology Program has listed Pb and Pb compounds as "reasonably anticipated
to be human carcinogens." Strong evidence from animal toxicological studies
demonstrates an association between Pb and cancer, genotoxicity or epigenetic
modification. Carcinogenicity in animal toxicology studies with relevant routes of Pb
exposure has been reported in the kidneys, testes, brain, adrenals, prostate, pituitary, and
mammary gland, albeit at high doses of Pb. Epidemiologic studies of cancer incidence
and mortality reported inconsistent results; one strong epidemiologic study demonstrated
an association between blood Pb and increased cancer mortality, but the other studies
reported weak or no associations. In the 2006 Pb AQCD, various indicators of Pb
exposure were found to be associated with stomach cancer, and a recent study on
stomach cancer and occupational Pb exposure reported mixed findings depending on the
type of Pb exposure (organic Pb, inorganic Pb, or Pb from gasoline emissions). Similarly,
some studies in the 2006 Pb AQCD reported associations between Pb exposure indicators
and lung cancer. Recent epidemiologic studies of lung cancer focused on occupational
exposures and reported inconsistent associations. The majority of epidemiologic studies
of brain cancer had null results overall, but positive associations between Pb exposure
indicators and brain cancer were observed among individuals with certain genotypes.
Overall, the consistent and strong body of evidence from toxicological studies on
carcinogenicity and potential modes of action but inconsistent epidemiologic evidence is
sufficient to conclude that a causal relationship is likely to exist between Pb exposure and
cancer.
1.7 Ecological Effects of Pb
Sections 1.7.1 and 1.7.2 are summaries of the evidence evaluated in Chapter 6 in which
the effects of Pb on terrestrial and aquatic ecosystems are presented separately. The
evidence supporting ecological causal determinations is synthesized across endpoints
(reproduction, growth, survival, neurobehavioral effects, hematological effects,
physiological stress) common to terrestrial, freshwater and saltwater biota in Section
1.7.3 (Table 1-3). An integration of the evidence across endpoints examined in both
human health and ecological studies follows (Section 1.8). Consideration of atmospheric
deposition of Pb as related to ecological effects is discussed under policy relevant
considerations (Section 1.9.7).
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1.7.1 Summary of Effects on Terrestrial Ecosystems
Historically, Pb poisoning is one of the earliest recognized toxicoses of terrestrial biota,
occurring primarily through ingestion of spent shot by birds (Section 6.3.4.3). At the time
of the 1977 Pb AQCD, few studies of Pb exposure and effects in wild animals other than
birds were available. A limited number of rodent trapping studies and observations from
grazing animals near smelters provided evidence for differences in Pb sensitivity among
species and these findings were further supported in the 1986 and 2006 Pb AQCDs (U.S.
EPA. 2006b. 1986b. 1977). Commonly observed effects of Pb on terrestrial organisms
include decreased survival, reproduction, and growth, as well as effects on development,
behavior, and ALAD activity (U.S. EPA. 2006b. 1986b. 1977).
In plants, Pb effects have been studied for several decades. At the time of the 1977 Pb
AQCD, it was understood that Pb uptake in plants varied with species and with the size
of the pool of Pb in the soil, and that most of the Pb taken up from the soil by plants other
than trees remains in the roots, with translocation to other portions of the plant varying
with species (U.S. EPA. 1977). Plant growth was recognized as an endpoint of Pb
toxicity in plants in the 1977 Pb AQCD and additional effects of Pb on growth processes
were reported in subsequent Pb AQCDs (U.S. EPA. 2006b. 1986b. 1977). In the 1977 Pb
AQCD evidence for effects of Pb on forest-nutrient cycling and shifts in arthropod
community composition was found in one study conducted in the vicinity of a smelting
complex. In subsequent AQCDs, other ecosystem-level effects, including decreased
species diversity, changes in floral and faunal community composition, and decreasing
vigor of terrestrial vegetation have been reported near stationary sources of Pb (U.S.
EPA. 2006b. 1986b. 1977; Watson etal. 1976).
Pb is either deposited directly onto plant surfaces, or onto soil where it can bind with
organic matter or dissolve in pore water. The amount of Pb dissolved in soil pore water
determines the impact of soil Pb on terrestrial ecosystems to a much greater extent than
the total amount present. It has long been established that the amount of Pb dissolved in
soil solution is controlled by at least six factors: (1) solubility equilibria;
(2) adsorption-desorption relationship of total Pb with inorganic compounds;
(3) adsorption-desorption reactions of dissolved Pb phases on soil organic matter; (4) pH;
(5) cation exchange capacity (CEC); and (6) aging. Since 2006, further studies have
contributed to the understanding of the role of pH, CEC, organic matter, and aging.
Smolders et al. (2009) demonstrated that the two most important determinants of both Pb
solubility and toxicity in soils are pH and CEC. However, they had previously shown that
experimental aging, primarily in the form of initial leaching following addition of Pb,
decreases soluble metal fraction by approximately one order of magnitude (Smolders et
al.. 2009). Since 2006, organic matter has been confirmed as an important influence on
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Pb sequestration, leading to longer-term retention in soils with higher organic matter
content, and also creating the potential for later release of deposited Pb. Aging, both
under natural conditions and simulated through leaching, was shown to substantially
decrease bioavailability to plants, microbes, and vertebrates.
Evidence over several decades of research, previously reviewed in Pb AQCDs and in
more recent studies, shows that Pb accumulates in terrestrial plants, invertebrates and
vertebrates. Studies with herbaceous plant species growing at various distances from
smelters added to the existing strong evidence that atmospherically transported Pb is
taken up by those plants. In most species tested, soil Pb taken up by the roots is not
translocated into the stem and leaves. These studies did not establish the relative
proportion that originated from atmospheric Pb deposited in the soil, as opposed to that
taken up directly from the atmosphere through the leaves. In trees, studies have found
that soil Pb generally is translocated to other parts, in contrast to herbaceous plants, and
recent studies have shown that the proportion of Pb that is taken up through the leaves
and trunk is likely substantial. One study attempted to quantify this proportion Pb that is
taken up directly from the atmosphere suggested it amounts to 50% of the Pb contained
in Scots pine (Pinus sylvestris) in Sweden (Klaminder et al., 2005).
Since the 2006 Pb AQCD, various species of terrestrial snails have been found to
accumulate Pb from both diet and soil. Recent studies with earthworms have found that
both internal concentration of Pb and mortality increase with decreasing soil pH and
CEC, and the importance of the interaction of those factors with soil Pb has been strongly
confirmed, but only very partially quantified. Tissue concentration differences have been
found between species of earthworms that burrow in different soil layers, but it is
unknown whether those tissue concentration differences are a direct result of species
differences, a result of differences in soil variables such as in pH and CEC, or
interactions among those factors. Because earthworms often sequester Pb in granules,
some authors have suggested that earthworm Pb is not bioavailable to their predators.
There is some evidence that earthworm activity increases Pb availability in soil, but it is
inconsistent. In various arthropods collected at contaminated sites, recent studies found
gradients in accumulated Pb that corresponded to gradients in soil with increasing
distance from stationary sources.
There are a few recent studies of Pb bioavailability and uptake in birds since the
2006 Pb AQCD. Several found tissue levels in birds that indicated exposure to Pb, but
none of the locations for these studies was in proximity to stationary anthropogenic
sources, and the origin of the Pb could not be identified. A study at the Anaconda Smelter
Superfund site found increasing Pb accumulation in gophers with increasing soil Pb
around the location of capture. A study of swine fed various Pb-contaminated soils
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showed that the form of Pb determined accumulation. Recent studies were able to
measure Pb in the components of various food chains that included soil, plants,
invertebrates, and vertebrates. They confirmed that trophic transfer of Pb is pervasive, but
no consistent evidence of trophic magnification was found.
Evidence in this review further supports the findings of the previous Pb AQCDs that
biological effects of Pb on terrestrial organisms vary with species and lifestage, duration
of exposure, form of Pb, and soil characteristics. In photosynthetic organisms,
experimental studies have added to the existing evidence of photosynthesis impairment in
plants exposed to Pb, and have found damage to photosystem II due to alteration of
chlorophyll structure, as well as decreases in chlorophyll content in diverse taxa,
including lichens and mosses. Evidence of oxidative stress in response to Pb exposure has
also been observed in plants. Reactive oxygen species were found to increase in broad
bean and tomato plants exposed to increasing concentrations of soil Pb, and a
concomitant increase in superoxide dismutase, glutathione, peroxidases, and lipid
peroxidation, as well as decreases in catalase were observed in the same plants. Monocot,
dicot, and bryophytic taxa grown in Pb-contaminated soil or in experimentally spiked soil
all responded to increasing exposure with increased antioxidant activity. In addition,
reduced growth was observed in some experiments, as well as genotoxicity, decreased
germination, and pollen sterility.
In terrestrial invertebrates, evidence for Pb effects has included neurological and
reproductive endpoints. Recently published studies have shown neuronal damage in
nematodes exposed to concentrations of Pb [2.5 uM (0.5 mg Pb/L)] in laboratory settings,
accompanied by behavioral abnormalities. Reproductive adverse effects were found at
lower exposure in younger nematodes, and effects on longevity and fecundity were
shown to persist for several generations. Increased mortality was found in earthworms,
but was strongly dependent on soil characteristics including pH, CEC, and aging. Snails
exposed to Pb through either topical application or through consumption of Pb-exposed
plants had increased antioxidant activity, and decreased food consumption, growth, and
shell thickness. Effects on arthropods exposed through soil or diet varied with species and
exposure conditions, and included diminished growth and fecundity, endocrine and
reproductive anomalies, and body malformations. Within each study, increasing
concentration of Pb in the exposure medium generally resulted in increased effects, but
the relationship between concentration and effects varied between studies, even when the
same medium, e.g., soil, was used. Evidence suggested that aging and pH are important
modifiers.
ALAD was identified in the 1977 Pb AQCD as a sensitive indicator of exposure to Pb in
rats and waterfowl, and became regarded as a biomarker of exposure in many terrestrial
1-41
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vertebrates. Other effects of Pb on vertebrates reviewed in Pb AQCDs and the current
document include decreased white blood cell counts and behavioral anomalies observed
in amphibians and reptiles. However, large differences in effects were observed at the
same concentration of Pb in soil, depending on whether the soil was freshly amended or
field-collected from contaminated areas. As in most studies where the comparison was
made, effects were smaller when field-collected soils were used. In some birds, maternal
elevated blood Pb level was associated in recent studies with decreased hatching success,
smaller clutch size, high corticosteroid level, and abnormal behavior. Some species
evidenced little or no effect of elevated blood Pb level. Effects of dietary exposure were
studied in several mammalian species, and cognitive, endocrine, immunological, and
growth effects were observed.
Recent evidence reviewed in Sections 6.3.6 and 6.3.12.7 demonstrates that exposure to
Pb is generally associated with negative effects in terrestrial ecosystems. It also
demonstrates that many factors, including species and various soil physiochemical
properties, interact strongly with Pb concentration to modify those effects. In these
ecosystems, where soil is generally the main component of the exposure route, Pb aging
is a particularly important factor, and one that may be difficult to reproduce
experimentally. Without quantitative characterization of those interactions,
characterizations of exposure-response relationships would likely not be transferable
outside of experimental settings. Since the 2006 Pb AQCD, few studies of
exposure-response have been conducted, and results have been inconsistent. Table 6-4
summarizes studies of reproduction, growth, and survival in terrestrial organisms that
have been published since 2006, and in which concentration-response data were reported.
Recent evidence of effects of Pb at the community and ecosystem levels of biological
organization include several studies of the ameliorative effects of mycorrhizal fungi on
plant growth in the presence of Pb, attributed to decreased uptake of Pb by plants,
although both mycorrhizal fungus and plant were negatively affected at the exposures
assessed. Most recently published research on community and ecosystem-level effects of
Pb has focused on soil microbial communities, which have been shown to be impacted in
both composition and activity. Many of the recent studies of effects on soil microbial
communities have taken place in environments contaminated with multiple metals, and
some have attempted to separate the effects of individual metals when possible. Soil
microbial activity was generally diminished, but in some cases recovered overtime.
Species and genotype composition were consistently altered, and those changes were
long-lasting or permanent. Recent studies have addressed differences in sensitivity
between species explicitly, and have clearly demonstrated high variability between
related species, as well as within larger taxonomic groupings. Mammalian no observed
effect concentration (NOEC) values expressed as blood Pb levels were shown to vary by
1-42
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a factor of 8, while avian blood NOECs varied by a factor of 50 (Buekers et al. 2009).
Protective effects of dietary Ca2+ have been found in plants, birds, and invertebrates.
1.7.2 Summary of Effects on Aquatic Ecosystems
Effects of Pb on plants, invertebrates, and vertebrates are reported for both freshwater
and saltwater ecosystems. Although effects of Pb exposure are likely mediated through
common mode(s) of action across freshwater and marine/estuarine taxa, these ecosystems
are considered separately because of different environmental and physiological factors
that influence Pb toxicity such as bioavailability of the metal, form of Pb, water quality
parameters and adaptations in freshwater and saltwater organisms. Toxicity of Pb also
varies by organism, lifestage and duration of exposure. (U.S. EPA. 2006b. 1986a).
Closely related organisms can vary greatly in bioaccumulation of Pb and other
non-essential metals as well as in their susceptibility to Pb. Pb effects on aquatic biota
were previously assessed in the 1977 Pb AQCD, the 1986 Pb AQCD and the
2006 Pb AQCD (U.S. EPA. 2006b. 1986a. 1977V
Exposure of freshwater and estuarine organisms to Pb, and associated effects, are tied to
terrestrial systems via watershed processes (Section 6.2). Atmospherically-derived Pb can
enter aquatic systems through runoff from terrestrial systems or via direct deposition over
a water surface. In aquatic ecosystems affected by Pb, exposures are most likely
characterized as low dose, chronic exposures. Once Pb enters surface waters, its
solubility and subsequent bioavailability are influenced by Ca2+ concentration, pH,
alkalinity, total suspended solids, and dissolved organic carbon (DOC), including humic
acids. In saltwater, higher levels of ions additionally affect Pb bioavailability. In
sediments, Pb bioavailability may be influenced by the presence of other metals, sulfides,
iron (Fe-) and manganese (Mn-)oxides, and physical disturbance. Recent studies provide
further evidence for the role of modifying factors such as pH, DOC, and hardness.
Toxicity of the same concentration of Pb can vary greatly under different experimental
conditions.
As recognized in the 2006 Pb AQCD and further supported in this review, uptake of Pb
by aquatic invertebrates and vertebrates may preferentially occur via exposure routes
other than direct absorption from the water column such as ingestion of contaminated
food and water, uptake from sediment pore waters, or incidental ingestion of sediment
(U.S. EPA. 2006b). Currently available models for predicting bioavailability focus on
acute toxicity and do not consider all possible routes of uptake. They are therefore of
limited applicability, especially when considering species-dependent differences in
uptake and bioaccumulation of Pb. Recent evidence supports the 2006 Pb AQCD
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conclusion that processes such as Pb adsorption, complexation, and chelation alter
bioavailability to aquatic organisms.
Biological Effects of Pb on Freshwater Plants, Invertebrates and
Vertebrates
Recent evidence further supports the findings of the previous Pb AQCDs that waterborne
Pb is highly toxic to freshwater plants, invertebrates and vertebrates, with toxicity
varying with species and lifestage, duration of exposure, form of Pb, and water quality
characteristics. Concentration-response data from freshwater organisms indicate that
there is a gradient of response to increasing Pb concentration and that some effects in
sensitive species are observed at concentrations of Pb quantified in U.S. surface waters
(Table 1-2).
The toxicity of Pb to aquatic algae and plants has been recognized in earlier EPA reviews
of this metal. In the 1977 Pb AQCD, differences in sensitivity to Pb among different
species of algae were reported and concentrations of Pb varied within and between
genera. This observation was subsequently generalized across aquatic taxa (U.S. EPA.
1977). At the time of the 1977 Pb AQCD, the information available on effects of Pb on
freshwater plants was limited. For plants in general, Pb was recognized to affect
photosynthesis, mitosis, and growth, but at concentrations much higher than typically
found in the environment. Effects of Pb on plants reported in subsequent Pb AQCDs
included decreased growth, deformation of cells, and blocking of the pathways that lead
to pigment synthesis, thus affecting photosynthesis.
Effects of Pb on aquatic plants supported by additional evidence in this review include
oxidative damage, decreased photosynthesis, and reduced growth. Most recent studies
report effects on growth at concentrations much higher than Pb typically encountered in
the environment, however, some sublethal endpoints such as effects on chlorophyll were
reported at concentrations in the 100 to 200 (ig Pb/L range, albeit still much higher than
those typically encountered in U.S. surface waters (Table 1-1). Elevated levels of
antioxidant enzymes are commonly observed in aquatic plant, algae, and moss species
exposed to Pb (U.S. EPA. 1977) and recent evidence continues to support this
observation. Recent studies on uptake of Pb by aquatic plants support the findings of
previous Pb AQCDs that all such plants with roots tend to sequester larger amounts of Pb
in their roots than in their shoots, and provide additional evidence for species differences
in compartmentalization of sequestered Pb and in responses to Pb in water and sediments.
Exposure-response relationships in which increasing concentrations of Pb leads to
increasing effects have consistently been reported in freshwater algae and macrophytes,
suggesting that effects on growth and antioxidant activity are also occurring at lower
1-44
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concentrations, however, most current observations of Pb effects in freshwater plants are
at concentrations that exceed Pb concentration values available for U.S. surface waters
(Table 1-1).
The largest body of evidence for effects of Pb at or near concentrations encountered in
U.S. surface waters is from invertebrates. In the 1986 Pb AQCD (U.S. EPA. 1986a) and
2006 Pb AQCD (U.S. EPA. 2006b). reduced reproduction, growth, and survival were
reported in various species of freshwater invertebrates. In the 2006 Pb AQCD,
concentrations at which effects were observed in aquatic invertebrates ranged from 5 to
8,000 (ig Pb/L. Recent evidence for effects of Pb on reproduction, growth, and survival
supports findings in previous Pb AQCDs (Table 6-5). In a series of 48-hour acute toxicity
tests using a variety of natural waters across North America, LC50 values ranged from 29
to 180 (ig Pb/L tests with the cladoceran Ceriodaphnia dubia (Esbaugh et al.. 2011). In
this same species, increased DOC leads to an increased mean EC50 for reproduction as
low as 25 (ig Pb/L. Reproductive and growth effects have also been reported in rotifer,
midge and mayfly species near the range of Pb concentrations encountered in freshwater
habitats. Several studies in this review have provided evidence of growth effects at lower
concentrations. Among the most sensitive species, growth of juvenile freshwater snails
(Lymnaea stagnalis) was inhibited at an EC2o of <4 (ig Pb/L (Grosell and Brix. 2009;
Grosell et al.. 2006b). A chronic value of 10 (ig Pb/L, obtained in 28-day exposures of
2-month-old freshwater mussel (Lampsilis siliquoided) juveniles, was the lowest
genus-mean chronic value ever reported for Pb (Wang et al.. 2010f).
Since the 2006 Pb AQCD, there is additional evidence for Pb effects on antioxidant
enzymes, lipid peroxidation, stress response and osmoregulation in aquatic invertebrates,
as well as additional information on Pb bioaccumulation. Recent studies using stable
isotopes have enabled simultaneous measurement of uptake and elimination in several
aquatic organisms to assess the relative importance of water versus dietary uptake. In
uptake studies of various invertebrates, Pb was mainly found in the gills and digestive
gland/hepatopancreas.
Pb effects on freshwater vertebrates were previously assessed in the 1977 Pb AQCD, the
1986 Pb AQCD and the 2006 Pb AQCD (U.S. EPA. 2006b. 1986a. 1977). Evidence of
toxicity of Pb and other metals to freshwater organisms goes back to early observations
of contamination of natural areas by Pb mining leading to extirpation offish from streams
(U.S. EPA. 1977). Recent evidence supports the findings of effects on survival,
reproduction, and behavior reported in previous Pb AQCDs for freshwater vertebrates. In
a series of 96-hour acute toxicity tests with fathead minnow conducted in a variety of
natural waters across North America, LC50 values ranged from 41 to 3,598 (ig Pb/L
(Esbaugh et al.. 2011). Reproductive effects associated with water quality parameters
1-45
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were also noted with this species (Mager et al. 2010). In fish, several recent studies on
behavioral effects of Pb indicate decreased prey capture rate, slower swim speed and
decline in startle response and visual contrast with Pb exposure. These reported effects
provide additional evidence for toxicity of Pb to fish. Chronic NOEC and ECio values
reported for trout, a sensitive species, are within the range of Pb occasionally encountered
in U.S. surface waters (Table 6-2).
Observed responses offish to Pb reported in the 1986 Pb AQCD and the 2006 Pb AQCD
included inhibition of heme formation, alterations in brain receptors, effects on blood
chemistry and hormonal systems, and decreases in some enzyme activities (U.S. EPA.
2006b. 1986a). Since the 2006 Pb AQCD, possible molecular targets for Pb neurotoxicity
have been identified in fish and additional mechanisms of Pb toxicity have been
elucidated in the fish gill and the fish renal system. In the 2006 Pb AQCD, amphibians
were considered to be relatively tolerant to Pb. Observed responses to Pb exposure
included decreased enzyme activity (e.g., ALAD reduction) and changes in behavior.
Since the 2006 Pb AQCD, studies conducted at concentrations approaching
environmental levels of Pb have indicated sublethal effects on tadpoles including
deformities and decrements in growth and swimming ability.
In the 2006 Pb AQCD, adverse effects were found in freshwater fish at concentrations
ranging from 10 to >5,400 (ig Pb/L, generally depending on water quality variables
(e.g., pH, hardness, salinity). Additional testing of Pb toxicity under conditions of varied
alkalinity, DOC, and pH has been conducted since the last review. Effects in fish
observed in recent studies fall within the range of concentrations observed in the previous
Pb AQCD. Recent evidence also supports the 2006 conclusions that the gill is a major
site of Pb uptake in fish, and that there are species differences in the rate of Pb
accumulation and distribution of Pb within the organism. The anterior intestine has been
newly identified as a site of uptake of Pb through dietary exposure studies. At the time of
the publication of the 2006 Pb AQCD, trophic transfer of Pb through aquatic food chains
was considered to be negligible. Measured concentrations of Pb in the tissues of aquatic
organisms were generally higher in algae and benthic organisms than in consumers at
higher trophic levels, indicating that Pb was bioconcentrated but not biomagnified. Some
studies published since the 2006 Pb AQCD support the potential for transfer of Pb in
aquatic food webs, while other studies indicate that Pb concentration decreases with
increasing trophic level.
Ecosystem-level effects associated with Pb reported in previous Pb AQCDs include
alteration of predator-prey dynamics, species richness, species composition, and
biodiversity. Since the 2006 Pb AQCD, additional evidence for community and
ecosystem level effects of Pb reviewed in Sections 6.4.7 and 6.4.12.1 have been observed
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primarily in microcosm studies or field studies near contaminated areas (mining,
effluent). Findings from field studies of aquatic communities in the vicinity of heavily
contaminated sites include changes in species composition and species richness,
predator/prey interactions, nutrient cycling and energy flow; however, Pb is often found
coexisting with other metals and other stressors, which risk confounding the observed
effects. Recent studies provide evidence in additional habitats for these community and
ecological-level effects, specifically in aquatic macrophyte communities and
sediment-associated communities. Different species may exhibit different responses to
Pb-impacted ecosystems dependent not only upon other environmental factors
(e.g., temperature, pH), but also on species sensitivity, lifestage, or seasonally-affected
physiological state. Aquatic ecosystems with low pH and low dissolved organic matter
are likely to be the most sensitive to the effects of atmospherically-deposited Pb.
Biological Effects of Pb on Saltwater Plants, Invertebrates and Vertebrates
In general, Pb toxicity to marine/estuarine plants, invertebrates and vertebrates is less
well characterized than toxicity to Pb in freshwater systems due to an insufficient
quantity of studies on saltwater organisms. In marine algae, effects on growth are
observed in the most sensitive species at Pb concentrations that exceed amounts
measured in the open sea or estuaries (Table 1-1). The majority of available studies of Pb
effects on saltwater organisms are for invertebrate species. Evidence for Pb effects on
reproduction, growth and survival as well as neurobehavioral, hematological and
physiological stress endpoints are coherent with findings in freshwater invertebrates
although most effects are observed at concentrations above 100 (ig Pb/L which exceeds
Pb typically encountered in seawater (Table 1-1). Fewer studies are available for Pb in
marine sediments. In the amphipod, Elasmopus laevis, onset to reproduction was
significantly delayed at 118 mg/Pb kg sediment; a concentration that the authors indicate
is below the current marine sediment regulatory guideline for Pb (218 mg Pb/kg
sediment) (Ringenary et al.. 2007; NOAA. 1999). In the same study, no effects of Pb on
adult survival in 28-day or 60-day sediment exposures were observed. Additional studies
on reproduction, growth, and survival in marine invertebrates report effects above the
range considered for causal determinations (Table II, Preamble). Several field monitoring
studies with marine bivalves have used ALAD as a biomarker for Pb exposure and
correlated ALAD inhibition to increased Pb tissue content. Field and laboratory studies
provide evidence for antioxidant response to Pb exposure, however, most effects are
observed at concentrations of Pb that are higher than concentrations detected in marine
environments. No recent evidence for effects of Pb on marine vertebrates in controlled
exposures was available for review.
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1.7.3
Determinations of Causality for Effects on Ecosystems
Table 1-3 Summary of Pb causal determinations for plants, invertebrates, and
vertebrates.
Level
Effect
Terrestrial3 Freshwater3 Saltwater3
Community-
and
Ecosystem-
Level
+-
'o
Q.
•c
LU
.0
?
Q.
Sub-organismal Organism-Level Responses
rx6SpOnS6S
Community and Ecosystem Effects
(Sections 6.3.12.7, 6.4.12.7, and 6.4.21.7)
Reproductive and Developmental Effects - Plants
(Sections 6.3.12.1, 6.4.12.1, and 6.4.21.1)
Reproductive and Developmental Effects -
Invertebrates
(Sections 6.3.12.1, 6.4.12.1, and 6.4.21.1)
Reproductive and Developmental Effects -
Vertebrates
(Sections 6.3.12.1, 6.4.12.1, and 6.4.21.1)
Growth - Plants
(Sections 6.3.12.2, 6.4.12.2, and 6.4.21.2)
Likely Causal
Inadequate
Causal
Causal
Causal
Likely Causal
Inadequate
Causal
Causal
Likely Causal
Inadequate
Inadequate
Suggestive
Inadequate
Inadequate
Sons 6V3ei2b2ra64 12 2 and 6 4 21 2) LikelV CaUSal CaUSal lnade"Uate
Growth - Vertebrates
(Sections 6.3.12.2, 6.4.12.2, and 6.4.21.2)
Survival - Plants
(Sections 6.3.12.3, 6.4.12.3, and 6.4.21.3)
Survival - Invertebrates
(Sections 6.3.12.3, 6.4.12.3, and 6.4.21.3)
(Sections 6.3.12.3, 6.4.12.3, and 6.4.21.3)
Neurobehavioral Effects - Invertebrates
(Sections 6.3.12.4, 6.4.12.4, and 6.4.21.4)
Neurobehavioral Effects - Vertebrates
(Sections 6.3.12.4, 6.4.12.4, and 6.4.21.4)
Hematological Effects - Invertebrates
(Sections 6.3.12.5, 6.4.12.5, and 6.4.21.5)
Hematological Effects - Vertebrates
(Sections 6.3.12.5, 6.4.12.5, and 6.4.21.5)
(Sions'eS ?2r6SS6 Tw and 64216)
Inadequate
inadequate
Causal
Likely Causal
Likely Causal
Likely Causal
Inadequate
Causal
Causal
Inadequate
inadequate
Causal
Causal
Likely Causal
Likely Causal
Likely Causal
Causal
Likely Causal
Inadequate
inadequate
Inadequate
Inadequate
Inadequate
Inadequate
Suggestive
Inadequate
Inadequate
Ph vsioloo IC3J Strsss ~ Invsrtsbrstss • • .•
• LlKSlV C3US3I LlKSlV C3US3I oUQQSStlVS
( o 6 Ct lOnSDol^D O 4 1 ^ O 3nClD4^lD)
rSe^±?3.??6SV^2T±6.4.21^ ^ CaUSal ^ CaUSal lnad^Uate
aConclusions are based on the weight of evidence for causal determination in Table II of the ISA Preamble. Ecological effects
observed at or near ambient Pb concentrations measured in soil, sediment and water in the most recent available studies (Table
1-1). were emphasized and studies generally within one to two orders of magnitude above the reported range of these values were
considered in the body of evidence for terrestrial (Section 6.3.12). freshwater (Section 6.4.12) and saltwater (Section 6.4.21)
ecosystems.
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1.7.3.1 Effects on Development and Reproduction
Evidence from invertebrate and vertebrate studies from Pb AQCDs and in this review
indicates that Pb is affecting reproductive performance in multiple species. Various
endpoints have been measured in multiple taxa of terrestrial and aquatic organisms to
assess the effect of Pb on development, fecundity, and hormone homeostasis, and they
have demonstrated the presence of adverse effects. Reproductive effects are important
when considering effects of Pb because impaired fecundity at the organism level of
biological organization can result in a decline in abundance and/or extirpation of
populations, decreased taxa richness, and decreased relative or absolute abundance at the
community level (Suter et al.. 2005; U.S. EPA. 2003a). The evidence is sufficient to
conclude that there is a causal relationship between Pb exposures and developmental and
reproductive effects in terrestrial (Section 6.3.12.1) and freshwater (Section 6.4.12.1)
invertebrates and vertebrates. Although there is less evidence for reproductive and
developmental effects of Pb in marine systems, available evidence is suggestive of a
causal relationship between Pb exposure and reproductive and developmental effects in
saltwater invertebrates (Section 6.4.21.1). The evidence is inadequate to conclude that
there is a causal relationship between Pb exposures and developmental and reproductive
effects in saltwater vertebrates and in either terrestrial or aquatic plants.
Recent evidence for developmental and reproductive endpoints in terrestrial invertebrates
shown to be affected by Pb include hatching success in collembolans, increased
development time in fruit flies and aphids, and disrupted hormone homeostasis in moths.
These studies have generally used Pb concentrations that exceed Pb soil concentrations
found at most U.S. locations (Table 1-1). but many of them included multiple
concentrations, and responses increased with increasing concentration. In terrestrial
vertebrates, recent evidence for decreased sperm count and quality in deer at a location
contaminated by mining, and for decreased testis weight in lizards, support previous
associations between Pb exposure and reproductive and developmental effects. Few
studies are available that specifically address reproductive effects of Pb exposure in either
terrestrial or aquatic plants.
In terrestrial invertebrates, Pb can alter developmental timing, hatching success, sperm
morphology, and hormone homeostasis. In fruit flies, Pb exposure increased time to
pupation and decreased pre-adult development. Sperm morphology was altered in
earthworms exposed to Pb-contaminated soils. Pb may also disrupt hormonal
homeostasis in invertebrates as studies with moths have suggested. Evidence of
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multi-generational toxicity of Pb is also present in terrestrial invertebrates, specifically
springtails, mosquitoes, carabid beetles, and nematodes where decreased fecundity in
progeny of Pb-exposed individuals was observed. However, effects have only been
studied in a small number of species.
For freshwater invertebrates, exposure to Pb under controlled conditions has provided
evidence for reproductive effects on sensitive taxa (gastropods, amphipods, cladocerans)
at or near the range of Pb concentration values available for U.S. surface waters (Table
1-1). Reproductive effects were reported to begin at 19 (ig Pb/L for the snail Lymnaea
palustris and 27 jig Pb/L for Daphnia sp. as reported in the 1986 Pb AQCD (U.S. EPA.
1986b). In a 42-day chronic study reviewed in the 2006 Pb AQCD, the LOEC for
reproduction was 3.5 (ig Pb/L in amphipods receiving both waterborne and dietary Pb
(Besser et al. 2005). Several recent studies in snails, rotifers and other freshwater
invertebrates support previous findings of adverse impacts to reproduction (Table 6-5) .
Reproductive effects have also been observed in multi-generational studies with aquatic
invertebrates. Larval settlement rate and rate of population increase was decreased in
rotifers and marine amphipods. Rotifers have a reduced fertilization rate associated with
Pb exposure that appears to be due to decreased viability of sperm and eggs.
In freshwater vertebrates there is evidence for reproductive and developmental effects of
Pb. Recent evidence in frogs and freshwater fish continues to support developmental and
reproductive effects of Pb in aquatic vertebrates reported in earlier Pb AQCDs.
Pb-exposure in tadpoles has been demonstrated to delay metamorphosis, decrease larval
size, produce subtle skeletal malformations, and to result in slower swimming speed.
Previous Pb AQCDs have reported developmental effects in fish, specifically spinal
deformities in larvae at a concentration of 120 (ig Pb/L. In the 2006 Pb AQCD, decreased
spermatocyte development in rainbow trout was observed at 10 (ig Pb/L and in fathead
minnow testicular damage occurred at 500 (ig Pb/L. In more recent studies, reproduction
in fathead minnows was affected in breeding exposures following 300-day chronic
toxicity testing. However, reproductive performance was unaffected in zebrafish Danio
rerio exposed to Pb-contaminated prey. In fish, there is recent evidence of Pb effects on
steroid profiles from nominal exposure studies.
In terrestrial vertebrates, evidence from Pb AQCDs and more recent evidence support an
association between Pb exposure and observed adverse reproductive effects. In mammals,
few studies in the field have addressed Pb specifically: most available data in wild or
grazing animals are from near smelters, where animals are co-exposed to other metals.
Evidence obtained using mammals in the context of human health research demonstrates
adverse effects of Pb on sperm, and on onset of puberty in males and females (Chapter 4).
which is coherent with the partial evidence from mammals in the wild. Other
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reproductive endpoints including spontaneous abortions, pre-term birth, embryo
development, placental development, low birth weight, subfecundity, hormonal changes,
and teratology were also affected, but less consistently. Recent toxicological data from
animal studies support trans-generational effects, a finding that is also an area of
emerging interest in ecology.
Many studies of effects on reproductive and developmental endpoints in terrestrial
invertebrates and vertebrates have been conducted with soil Pb concentrations exceeding
those found in most U.S. locations. Recent and past studies that include multiple,
increasing concentrations of Pb, from background level to levels greater than those
associated with heavily contaminated sites, showed exposure-dependent responses. For
some sensitive aquatic species, recent evidence supports previous findings of
reproductive and developmental effects of Pb and differential lifestage response at or near
concentrations of Pb reported from natural environments.
1.7.3.2 Effects on Growth
Alterations in the growth of an organism can impact population, community and
ecosystem level variables. Evidence is sufficient to conclude that there is a causal
relationship between Pb exposures and effects on growth in terrestrial plants (Section
6.3.12.2) and freshwater invertebrates (Section 6.4.12.2). Evidence is sufficient to
conclude that a causal relationship is likely to exist between Pb exposure and effects on
growth in terrestrial invertebrates and freshwater plants. Evidence is inadequate to
establish a causal relationship between Pb exposures and effects on growth in terrestrial
and aquatic vertebrates and saltwater biota (Section 6.4.21.2).
Evidence for effects of Pb on growth is strongest in terrestrial plants. In invertebrates,
evidence for effects of Pb on growth has been observed most extensively in freshwater
taxa, with inhibition in sensitive species occurring in or near the range of Pb
concentration values found in surveys of U.S. surface waters (Table 1-1). Vertebrates,
particularly terrestrial, have been the object of a comparatively much smaller number of
studies of the effects of Pb on growth. Growth effects observed in both invertebrates and
vertebrates, however, underscore the importance of lifestage to overall Pb susceptibility.
In general, juvenile organisms are more sensitive than adults. Evidence for growth effects
of Pb in freshwater and terrestrial plant species is primarily supported by earlier Pb
AQCDs. In aquatic invertebrates, the weight of the evidence continues to support growth
effects of Pb with several recent studies reporting effects at < 10 (ig Pb/L, specifically in
snails and mussels (Table 6-5). Also, growth effects in frogs are reported at lower
concentrations in the current document than in earlier reviews.
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There is evidence over several decades of research that Pb inhibits photosynthesis and
respiration in plants, both of which reduce growth (U.S. EPA. 2006b. 1977). Many
toxicity studies conducted in laboratory and greenhouse settings have reported effects on
plants. These effects are typically observed in laboratory studies with high exposure
concentrations or in field studies near stationary sources and heavily contaminated sites,
but studies that include multiple concentrations of Pb show increased response with
increasing concentration. Pb has been shown to affect photosystem II, altering the
pigment structure, and decreasing the efficiency of visible light absorption by affected
plants. Decreases in chlorophyll a and b content have been observed in various algal and
plant species. Most primary producers experience EC50 values for growth in the range of
1,000 to 100,000 (ig Pb/L with minimal inhibition of growth observed as low as
100 jig Pb/L (U.S. EPA. 2006b).
Growth effects of Pb on aquatic invertebrates are reviewed in the draft Ambient Aquatic
Life Water Quality Criteria for Lead (U.S. EPA. 2008b) and the 2006 Pb AQCD (U.S.
EPA. 2006b). In the 2006 Pb AQCD, the LOEC for growth of freshwater amphipods
Hyalella azteca in 42-day chronic exposure to Pb was 16 (ig Pb/L (Besser et al.. 2005).
Recent studies summarized in Table 6-5 provide additional evidence for effects on
growth of aquatic invertebrates, with some effects observed at < 10 (ig Pb/L. Growth of
juvenile freshwater snails L. stagnalis was inhibited below the lowest concentration
tested resulting in an EC20 <4 (ig Pb/L (Grosell and Brix. 2009: Grosell et al.. 2006b). In
the same study, the NOEC was 12 (ig Pb/L and the LOEC was 16 (ig Pb/L. The authors
indicated that the observed effect level for Pb was very close to the current U.S. EPA
water quality criteria for Pb (3.3 (ig Pb/L normalized to test water hardness) (Grosell and
Brix. 2009). In the freshwater mussel, fatmucket (L. siliquoided) juveniles were the most
sensitive lifestage (Wang etal. 2010f). A chronic value of 10 (ig Pb/L in a 28-day
exposure of 2-month-old fatmucket juveniles was the lowest genus mean chronic value
ever reported for Pb. Growth effects are also reported in marine invertebrates at higher
concentrations of Pb than sensitive freshwater invertebrates.
In Pb AQCDs, a few studies have reported growth effects of Pb on vertebrates including
fish (growth inhibition), birds (changes in juvenile weight gain), and frogs (delayed
metamorphosis, smaller larvae). A recent study reviewed in this ISA supports findings of
growth effects in frogs and suggests that these effects may be occurring at lower
concentrations than previously reported: the growth rate of tadpoles of the northern
leopard frog exposed nominally to 100 (ig Pb/L from embryo to metamorphosis was
slower than the growth rate of the controls (Chen et al.. 2006b). In this study, Pb
concentrations in the tissues of tadpoles were quantified and the authors reported that
they were within the range of reported tissue concentrations reported in wild-caught
populations. Reports of Pb-associated growth effects in fish are inconsistent.
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1.7.3.3 Effects on Survival
Survival is a biologically important response that may have a direct impact on population
size and can lead to effects at the community and ecosystem level of biological
organization. The evidence is sufficient to conclude that there is a causal relationship
between Pb exposures and survival in terrestrial invertebrates (Section 6.3.12.3) and
freshwater invertebrates and vertebrates (Section 6.4.12.3). Evidence is sufficient to
conclude that a causal relationship is likely to exist between Pb exposure and survival in
terrestrial vertebrates (Section 6.3.12.3). The evidence is inadequate to conclude that
there is a causal relationship between Pb exposure and survival in terrestrial and
freshwater plants (Section 6.3.12.3. and Section 6.4.12.3). as well as in all saltwater biota
(Section 6.4.21.3). There is evidence for mortality in saltwater organisms at
concentrations that greatly exceed Pb concentrations typically encountered in seawater.
In general, marine organisms are less sensitive to Pb than freshwater species.
In terrestrial vertebrates, evidence for Pb effects on survival is primarily supported by Pb
AQCDs with no recent studies reporting effects on survival at lower concentrations. For
aquatic invertebrates recent studies support previous associations between Pb exposure
and mortality at concentrations near the range of Pb concentration values available for
U.S. surface waters (Table 1-1) in cladocerans, amphipods, and rotifers (Table 6-5). In
aquatic vertebrates, there is recent evidence for effects in fish at <100 (ig Pb/L. Pb is
generally not phytotoxic to freshwater or terrestrial plants at concentrations found in the
environment away from stationary sources and heavily contaminated sites, probably due
to the fact that plants often sequester large amounts of Pb in roots, and that translocation
to other parts of the plant is limited.
The relationship between Pb exposure and decreased survival rate has been well
demonstrated in terrestrial and aquatic species, as presented in Sections 6.3.12.3.
6.4.12.3. 6.4.21.3. and in previous Pb AQCDs. Toxicological studies have established
LC50 values for some species of plants, invertebrates, and vertebrates. From the LC50
data on Pb in this review and previous Pb AQCDs, a wide range of sensitivity to Pb is
evident across taxa. LC50 values are usually much higher than Pb concentrations near
contaminated sites, although physiological dysfunction that adversely impacts the fitness
of an organism often occurs well below concentrations that result in mortality.
Freshwater aquatic invertebrates are generally more sensitive to Pb exposure than other
taxa, with survival adversely impacted in a few species in laboratory studies at
concentrations near typical ambient levels. Freshwater biota that exhibit sensitivity to Pb
in the upper range of Pb concentrations measured in U.S. waters [median 0.50 (ig Pb/L,
range 0.04 to 30 (ig Pb/L (U.S. EPA. 2006b)1. include some species of gastropods,
amphipods, cladocerans, and rotifers although the toxicity of Pb is highly dependent upon
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water quality variables such as DOC, hardness, and pH. For example, amphipods tested
under various water conditions exhibited sensitivity to Pb at <10 (ig Pb/L (U.S. EPA.
2006c) and the present document). These impacted species may include endangered
species or candidates for the endangered species list, such as the freshwater mussel
Lampsilis rafinesqueana (Neosho mucket). The EC50 for foot movement (a measure of
viability) for newly transformed juveniles of this species was 188 (ig Pb/L. Other aquatic
invertebrates such as odonates may be tolerant of Pb concentrations that greatly exceed
Pb detected in aquatic ecosystems.
Terrestrial invertebrates typically tolerate high concentrations of Pb relative to
concentrations found in most uncontaminated soils. In the 1986 Pb AQCD it was reported
that Pb at environmental concentrations occasionally found near roadsides and smelters
(10,000 to 40,000 (ig Pb/g dry weight [mg Pb/kg]) can eliminate populations of bacteria
and fungi on leaf surfaces and in soil. LC50 values for soil nematodes vary from
10-1,550 mg Pb/kg dry weight dependent upon soil organic matter content and soil pH
(U.S. EPA. 2006b). In earthworms, 14 and 28 day LC50 values typically fall in the range
of 2,400-5,800 mg Pb/kg depending upon the species tested.
Data on mortality of saltwater species associated with exposure to Pb is limited; however,
in general, marine organisms are less sensitive to this metal than freshwater organisms
and the highest toxicity is observed in juveniles. In one study, effects of Pb on survival at
environmentally relevant concentrations of Pb in diet have been demonstrated through a
simulated marine food chain in which the primary producer, the microalgae Tetraselmis
suecica, was exposed nominally to 20 (ig Pb/L and subsequently fed to brine shrimp
Artemiafranciscana, (mean Pb content 12 to 15 (ig Pb/g) which were consumed by
white-leg shrimp Litopenaeus vannamei, itself consumed by grunt fish Haemulon
scudderi representing the top of the marine food chain (Soto-Jimenez et al.. 201 Ib).
Survival of brine shrimp was 25 to 35% lower than the control and both white shrimp and
grunt fish had significantly higher mortalities than controls.
In vertebrates, toxicity is observed in terrestrial avian and mammalian species in
laboratory studies over a wide range of doses (<1 to > 1,000 mg Pb/kg body weight
per day) as reviewed for the development of ecological soil screening levels (Eco-SSLs)
(U.S. EPA. 2005b). The NOAELs for survival ranged from 3.5 to 3,200 mg
Pb/kg per day. In freshwater vertebrates there is considerable historic information on Pb
toxicity to fish. Recent studies support earlier AQCD findings of Pb effects on fish
survival and indicate effects at lower concentrations when testing with juvenile lifestages.
In a series of 96-hour acute toxicity tests with fathead minnow conducted in a variety of
natural waters across North America, LC50 values ranged from 41 to 3,598 (ig Pb/L and
no Pb toxicity occurred in three highly alkaline waters (Esbaugh et al., 2011). Thirty day
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LC50 values for larval fathead minnows ranged from 39 to 1,903 (ig Pb/L in varying
concentrations of DOC and Ca2+ (as CaSO4) (Grosell et al.. 2006a). In a recent study of
rainbow trout fry at 2-4 weeks post-swim up, the 96-hour LC50 was 120 (ig Pb/L at a
hardness of 29 mg/L as calcium carbonate (CaCO3), a value much lower than in testing
with older fish (Mebane et al., 2008). In the same study, two chronic (>60 day) tests were
conducted with rainbow trout and the LC2o values for survival were 34 (ig Pb/L in the 69
day test and 113 (ig Pb/L in the 62 day test.
1.7.3.4 Neurobehavioral Effects
Overall, the evidence from terrestrial and freshwater systems is sufficient to conclude that
a causal relationship is likely to exist between Pb exposures and neurobehavioral effects
in invertebrates and vertebrates (Sections 6.3.12.4 and 6.4.12.4). Evidence is inadequate
to conclude that there is a causal relationship between Pb exposure and neurobehavioral
endpoints in saltwater species (Section 6.4.21.4).
Observations from laboratory studies reported in Chapter 6 and previous Pb AQCDs have
shown adverse effects of Pb on neurological endpoints in both terrestrial and freshwater
animal taxa. Studies that consider mode-of-action and molecular targets of Pb toxicity in
biota are now available for a few species. Recent studies have continued adding to the
evidence from both invertebrate and vertebrate studies that Pb adversely affects behaviors
such as food consumption, avoidance, and escape from predators, behavioral
thermoregulation, and prey capture. These changes are likely to decrease the overall
fitness of the organism. Recent evidence includes reports of behavioral responses across a
larger variety of organisms including fish larvae born from Pb-exposed adults and
reptiles, while some impairments in feeding and escaping behaviors were reported for the
first time.
Central nervous system effects in fish recognized in previous Pb AQCDs include effects
on spinal neurons and brain receptors. Recent evidence from this review identifies
possible molecular targets for Pb neurotoxicity in fish. Additionally, there is recent
evidence for neurotoxic action of Pb in invertebrates with exposure to Pb observed to
cause changes in the morphology of gamma aminobutyric acid (GAB A)-motor neurons
in nematodes (Caenorhabditis elegans) (Du and Wang. 2009).
Decreased food consumption of Pb-contaminated diet has been demonstrated in some
invertebrates (snails) and vertebrates (lizards, pigs, fish). Behavioral effects in grunt fish
H. scudderi, occupying the top level of a simulated marine food chain included lethargy
and decreased food intake in a 42-day feeding study (Soto-Jimenez et al.. 201 Ib). These
fish were fed white shrimp exposed to Pb via brine shrimp that were initially fed
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microalgae cultured nominally at 20 (ig Pb/L. In the same study, surfacing, reduction of
motility, and erratic swimming were observed in the white shrimp after 30 days of
exposure to Pb via diet. Pb may also decrease the ability of an organism to capture prey
or escape predation. For example, Pb exposure has been demonstrated to adversely affect
prey capture ability of certain fungal and fish species, and the motility of nematodes was
adversely affected in Pb-contaminated soils (Wang and Xing. 2008). Prey capture ability
was decreased in 10-day-old fathead minnows born from adult fish exposed to
120 (ig Pb/L for 300 days, then subsequently tested in a 21-day breeding assay (Mager_et
al.. 2010). Altered pattern of escape swimming in larval zebrafish exposed to Pb as
embryos was reported at low nominal concentrations of Pb (2 and 6 (ig Pb/L). Other
behavioral effects of Pb observed in fish include increased hyperactivity and decreased
ability to detect visual contrast. In a laboratory study, Pb-exposed gull chicks exhibited
abnormal behaviors such as decreased walking, erratic behavioral thermoregulation and
food begging that could make them more vulnerable in the wild (Burger and Gochfeld.
2005). The chicks were exposed to Pb via injection to produce feather Pb concentration
approximately equivalent to those observed in wild gulls. Lizards exposed to Pb through
diet in the laboratory exhibited abnormal coloration and posturing behaviors. These
findings show strong coherence with findings from studies in laboratory animals that
show that Pb induces changes in learning and memory (Section 4.3.2.3). as well as
attention (Section 4.3.3) and motor function (Section 4.3.7).
1.7.3.5 Hematological Effects
Based on observations in both terrestrial and freshwater organisms and additionally
supported by toxicological and epidemiological findings in laboratory animals and
humans, evidence is sufficient to conclude that there is a causal relationship between Pb
exposures and hematological effects in terrestrial and aquatic vertebrates (Sections
6.3.12.5 and 6.4.12.5). The evidence is sufficient to conclude that a causal relationship is
likely to exist between Pb exposures and hematological effects in freshwater
invertebrates and inadequate to conclude that there is a causal relationship between Pb
exposures and hematological effects in terrestrial invertebrates. Limited evidence from
marine invertebrates is suggestive of a causal relationship between Pb exposures and
hematological effects (Section 6.4.21.5). The mode of action of Pb on ALAD activity is
likely mediated through a common pathway in terrestrial, freshwater and saltwater
organisms.
Recent studies add support to the strong body of evidence presented in Pb AQCDs that
Pb exposure is associated with hematological responses in terrestrial and aquatic
vertebrates. Lower ALAD activity has been significantly correlated with elevated blood
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Pb levels in fish and mammals. In the 1986 Pb AQCD, decreases in RBC ALAD activity
following Pb exposure were well documented in birds and mammals (U.S. EPA. 1986a).
The draft Ambient Aquatic Life Water Quality Criteria for Lead summarized several
studies of ALAD activity in fish (U.S. EPA. 2008b). Further evidence from the
2006 Pb AQCD and this review suggests that this enzyme is an indicator for Pb exposure
across a wide range of taxa. Since the 2006 Pb AQCD, evidence of Pb effects on ALAD
activity has been found in additional species of invertebrates and fish, and has been
identified in bacteria. ALAD activity has been shown to vary greatly between species. In
addition to consideration of ALAD activity, there is recent evidence for deceased white
blood cell counts in amphibians affected by Pb exposure. The consistency and coherence
of these findings of effects on ALAD activity in vertebrates are also supported by some
evidence of Pb-induced alterations of blood chemistry in fish reported in the
2006 Pb AQCD (U.S. EPA. 2006b). This evidence is strongly coherent with observations
from human epidemiologic and animal toxicology studies where a causal relationship
was identified between Pb exposure and decreased RBC survival and function, and
altered heme synthesis in humans and laboratory animals (Sections 1.6.5 and 4.7).
In environmental assessments of metal-impacted habitats, ALAD is a recognized
biomarker of Pb exposure in invertebrates as well as vertebrates (U.S. EPA. 2006b).
Recent field studies of ALAD activity include observations in songbirds and owls near
historical mining areas and in bivalves collected from freshwater and estuarine
environments. Evidence for hematological effects of Pb in saltwater invertebrates is
limited primarily to field monitoring studies with bivalves.
1.7.3.6 Effects on Physiological Stress
Evidence is sufficient to conclude that there is a causal relationship between Pb
exposures and physiological stress in terrestrial plants (Section 6.3.12.6). Evidence is
sufficient to conclude that a causal relationship is likely to exist between Pb exposures
and physiological stress in terrestrial invertebrates and vertebrates (Section 6.3.12.6) as
well as freshwater plants, invertebrates and vertebrates (Section 6.4.12.6). Further
evidence in saltwater invertebrates is suggestive of a causal relationship between Pb
exposures and physiological stress (Section 6.4.21.6). Evidence is inadequate to conclude
that there is a causal relationship between Pb exposure and physiological stress responses
in saltwater plants and vertebrates.
Endpoints associated with physiological stress received no consideration prior to the
2006 Pb AQCD. Studies reviewed in the 2006 Pb AQCD reported stress-related effects
including upregulation of antioxidant enzymes and increased lipid peroxidation (U.S.
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EPA. 2006b). Recent evidence in additional species of terrestrial and freshwater plants,
invertebrates and vertebrates support, and expand upon findings in the previous Pb
AQCD. Some of these studies report findings within one to two orders of magnitude of
the range of Pb concentrations measured in terrestrial and freshwater environments
(Table 1-1). Recent studies include evidence for production of reactive oxygen species in
terrestrial plant species and in freshwater algae and fish in response to Pb exposure.
In the current document and the 2006 Pb AQCD (U.S. EPA. 2006b). there is strong
evidence of upregulation of antioxidant enzymes and increased lipid peroxidation
associated with Pb exposure in many species of plants, invertebrates and vertebrates. In
plants, increases of antioxidant enzymes with Pb exposure occur in algae, aquatic mosses,
floating and rooted aquatic macrophytes, and terrestrial species. Most observations of
antioxidant responses in plants typically occur at concentrations of Pb higher than found
in the environment. However, in a few terrestrial plant species, increases of antioxidant
enzymes occur at concentrations approaching the average Pb concentrations in U.S. soils
(Table 1-1) and limited transplantation studies with aquatic plants indicate elevated
antioxidant enzyme activity associated with Pb levels measured in sediments at polluted
sites. There is evidence for antioxidant activity in invertebrates, including gastropods,
mussels, and crustaceans, in response to Pb exposure. Some recent evidence for
invertebrate antioxidant responses in freshwater bivalves, and marine bivalve and
crustacean species indicates effects at Pb concentrations associated with polluted sites.
Markers of oxidative damage are also observed in fish, amphibians, and mammals, both
in the laboratory and in exposed natural environments. Across all biota, there are
differences in the induction of antioxidant enzymes that appear to be species-dependent.
Additional stress responses observed in terrestrial and freshwater invertebrates include
elevated heat shock proteins, osmotic stress and decreased glycogen levels. Heat shock
protein induction by Pb exposure has been observed in zebra mussels and mites. Tissue
volume regulation is adversely affected in freshwater crabs and glycogen levels
decreased in freshwater snails following Pb exposure. Although correlated with Pb
exposure, these responses are non-specific and may be altered with exposure to any
number of environmental stressors.
Upregulation of antioxidant enzymes and increased lipid peroxidation are considered to
be reliable biomarkers of stress, and suggest that Pb exposure induces a stress response in
those organisms, which may increase susceptibility to other stressors and reduce
individual fitness. The oxidative stress responses associated with Pb exposure are
consistent in terrestrial biota and in freshwater organisms. Furthermore, these responses
are also observed in experimental animal studies.
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1.7.3.7 Community and Ecosystem Effects
More evidence for Pb toxicity to terrestrial and aquatic biota has been reported from
single-species assays in laboratory studies than from whole ecosystem studies. The
evidence is strong for effects of Pb on growth, reproduction, and survival in very diverse
species, but considerable uncertainties exist in generalizing effects observed under
particular, small-scale conditions, up to the ecosystem level of biological organization. At
the ecosystem level, the presence of multiple stressors, variability in field conditions, and
differences in bioavailability of Pb make it difficult to measure the magnitude of effects,
and to quantify relationships between ambient concentrations of Pb and ecosystem
response. However, the cumulative evidence that has been reported for Pb effects at
higher levels of biological organization and for endpoints in single species with direct
relevance to population and ecosystem level effects (i.e., development and reproduction,
growth, survival) is sufficient to conclude that a causal relationship is likely to exist
between Pb exposures and the alteration of species richness, species composition and
biodiversity in terrestrial and freshwater ecosystems (Sections 6.3.12.7 and 6.4.12.7).
Evidence is inadequate to conclude that there is a causal relationship between Pb
exposure and effects at higher levels of biological organization in saltwater ecosystems
(Section 6.4.21.7).
Ecosystem-level studies in situ are complicated by the frequent confounding of Pb
exposure in Pb-polluted sites with other factors such as other trace metals and acidic
deposition. In those natural systems, Pb is often found co-existing with other stressors,
and observed effects may be due to cumulative toxicity. In laboratory studies and
simulated ecosystems, where it is possible to isolate the effect of Pb, this metal has been
shown to alter competitive behavior of species, predator-prey interactions, and
contaminant avoidance. At higher levels of biological organization, these effects may
change species abundance and community structure. Uptake of Pb into aquatic and
terrestrial organisms and its effects on survival, growth, and reproductive endpoints at the
organism level are expected to have ecosystem-level consequences. Where evidence of
effects is observed at the ecosystem level of organization, evidence from lower levels
brings consistency and plausibility for causality.
Most direct evidence of community and ecosystem level effects is from near stationary
sources and contaminated sites where Pb concentrations are higher than typically
observed in the environment. For terrestrial systems, evidence of impacts on natural
ecosystems near smelters, mines, and other industrial sources of Pb has been assembled
in previous decades. Those impacts include decreases in species diversity and changes in
floral and faunal community composition. For freshwater systems, the literature focuses
on evaluating ecological stress from Pb originating from urban and mining effluents
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rather than atmospheric deposition. Some organisms exhibit contaminant avoidance
behaviors when exposed to Pb-contaminated areas. For example, snails and fish avoid
higher concentrations of Pb while frogs and toads lack avoidance response. Recent
evidence, published since the 2006 Pb AQCD indicates that some species of worms will
avoid Pb-contaminated soils (Langdon et al., 2005).These dynamics are likely to change
species abundance and community structure at higher levels of biological organization.
Recent studies continue to demonstrate associations between Pb exposures and effects at
higher levels of biological organization that were shown in field and microcosm studies
in previous Pb AQCDs. Recent studies on plant and soil microbial communities and
sediment-associated and aquatic plant communities increase the total number of types of
ecological associations impacted by Pb. In terrestrial ecosystems, most studies show
decreases in microorganism abundance, diversity, and function with increasing soil Pb
concentration. Specifically, shifts in nematode communities, bacterial species, and fungal
diversity have been observed. Furthermore, presence of arbuscular mycorrhizal fungi
may protect plants growing in Pb-contaminated soils. Increased plant diversity
ameliorated effects of Pb contamination on a microbial community.
In aquatic ecosystems, Pb effects reviewed in the 2006 Pb AQCD (U.S. EPA. 2006b)
included reduced species abundance, richness and diversity, decreased primary
productivity, and altered predator-prey interactions. Since the 2006 Pb AQCD, there is
further evidence for effects of Pb in sediment-associated communities in both saltwater
and freshwater systems. Community structure and nematode diversity were altered in a
microcosm study with marine sediments (Mahmoudi et al., 2007). Sediment-bound Pb
contamination appears to differentially affect members of the benthic invertebrate
community, potentially altering ecosystem dynamics in small urban streams (Kominkova
and Nabelkova. 2005). Although surface water Pb concentrations in monitored streams
were determined to be very low, concentrations of the metal in sediment were high
enough to pose a risk to the benthic community (e.g., 34-101 mg Pb/kg). These risks
were observed to be linked to benthic invertebrate functional feeding group, with
collector-gatherer species exhibiting larger body burdens of heavy metals than benthic
predators and collector-filterers.
Changes to aquatic plant community composition have been observed in the presence of
elevated surface water Pb concentrations. A shift toward more Pb-tolerant species is also
observed in terrestrial plant communities near smelter sites (U.S. EPA, 2006b). Certain
types of plants such as rooted and submerged aquatic plants may be more susceptible to
aerially-deposited Pb resulting in shifts in Pb community composition. High Pb sediment
concentrations are linked to shifts in amphipod communities inhabiting plant structures.
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1.8 Integration of Health and Ecological Effects
The health and ecological effects considered for causal determination are summarized in
Table 1-4. The health outcomes were those related to the nervous, cardiovascular, renal,
and immune system, effects on heme synthesis and RBC function, reproductive and
developmental effects, and cancer. The ecological endpoints considered for causal
determination were: community and ecosystem level effects, reproductive and
developmental effects, growth, survival, neurobehavioral effects, hematological effects,
and physiological stress. The evidence relating to specific ecological endpoints is also
integrated across aquatic and terrestrial habitats. Further, the substantial overlap between
the ecological and health endpoints considered in the causal determinations allowed the
integration of the evidence across these disciplines.
Table 1-4 Summary of causal determinations for health and ecological effects.
Outcome/Effect
Human Health
Causal Determination3
Ecological Receptors
Causal Determination3
Nervous System
Effects'3
Cardiovascular
Effects
Renal Effects
Immune Effects
Hematological
Effects0
Causal Relationship: Cognition, Attention,
Impulsivity and Hyperactivity in Children
Causal Relationship: Hypertension and
Coronary Heart Disease
Likely Causal Relationship: Reduced
Kidney Function
Likely Causal Relationship: Atopic and
Inflammatory Conditions, Decreased Host
Resistance
Causal Relationship: RBC Function and
Survival, Altered Heme Synthesis
Likely Causal Relationship: Neurobehavioral
Effects in Terrestrial and Freshwater
Invertebrates and Vertebrates
N/Ae
N/Ae
N/Ae
Causal Relationship: ALAD Activity in Terrestrial
and Freshwater Vertebrates
Likely Causal Relationship: ALAD activity in
Freshwater Invertebrates
Reproductive and
Developmental
Effectsd
Causal Relationship: Development and
Male Reproductive Function
Causal Relationship:
Invertebrates and Vertebrates
Cancer
Likely to be a causal relationship
N/Ae
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Table 1-4 (Continued): Summary of causal determinations for health and ecological effects.
Outcome/Effect
Human Health
Causal Determination3
Ecological Receptors
Causal Determination3
Mortality
N/Ae (The strongest evidence of
Pb-induced mortality in humans was
observed for cardiovascular disease
related mortality and this evidence was
considered in determining the causal
relationship between Pb exposure and
coronary heart disease.)
Causal Relationship:
Survival Terrestrial Invertebrates and
Freshwater Invertebrates and Vertebrates
Likely Causal Relationship:
Terrestrial Vertebrates
Growth
N/Ae (There is mixed evidence from
toxicological and epidemiologic studies of
Pb effects on postnatal growth, which was
considered in determining the causal
association between Pb exposure and
developmental effects.)
Causal Relationship: Terrestrial Plants and
Freshwater Invertebrates
Likely Causal Relationship: Freshwater Plants
and Terrestrial Invertebrates
Physiological Stress
N/Ae (In Human Health, oxidative stress
was considered as an upstream event in
the modes of action of Pb, leading
downstream to various effects. Ecological
literature commonly uses oxidative stress
as a proxy indicator of overall fitness, and
thus treats it as an effect.)
Causal Relationship: Terrestrial Plants
Likely Causal Relationship:
Terrestrial and Freshwater Invertebrates and
Vertebrates and Freshwater Plants
Community and
Ecosystem Level
Effects
N/Ae
Likely Causal Relationship:
Terrestrial and Freshwater Ecosystems
aln drawing conclusions regarding the causal relationship between Pb exposure and human health effects, evidence in the range of
relevant pollutant exposures or biomarker levels was considered. Evidence from toxicological studies of effects observed in
experimental animals at biomarker levels (e.g. blood Pb) comparable to those currently experienced by the U.S. general population
were emphasized. Generally, studies with dietary exposures resulting in blood Pb levels within one order of magnitude above the
upper end of the distribution of U.S. blood Pb levels were considered in forming concusions, with the majority of studies reporting
blood Pb levels below 30 ug/dLIn forming conclusions, Ecological effects observed at or near ambient Pb concentrations measured
in soil, sediment and water in the most recent available studies (Table 1-1) were emphasized and studies generally within one to two
orders of magnitude above the reported range of these values were considered in the body of evidence for terrestrial, freshwater
and saltwater ecosystems.
bln ecological receptors, the causal determination was developed considering neurobehavioral effects that can be observed in
toxicological studies of animal models and studies of ecological effects in vertebrates and invertebrates.
°The ecological evidence considered for the causal determination included ALAD activity, blood cell count, and altered serum
profiles.
dFor health effects the strongest evidence was for delayed onset of puberty and detrimental effects on sperm. In the ecological
literature, a wide range of endpoints, including embryonic development, multigenerational studies, delayed metamorphosis, and
altered steroid profiles, was considered.
eN/A, not applicable, i.e., Endpoints were not directly comparable for the health and ecological evidence.
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1.8.1 Modes of Action Relevant to Downstream Health and Ecological
Effects
The diverse health and ecological effects of Pb are mediated through multiple,
interconnected modes of action. This section summarizes the principal
cellular/subcellular effects contributing to modes of action for human health endpoints
associated with Pb exposure and the concentrations at which those effects are observed.
Then, effects of Pb observed in aquatic and terrestrial species (Section 1.7) are evaluated
along with evidence from epidemiological and laboratory animal studies to determine the
extent to which common modes of action can be inferred from the observed effects. The
rationale for this approach is that the mechanism of Pb toxicity is likely conserved from
invertebrates to vertebrates to humans in some organ systems.
Each of the modes of action discussed in Section 4.2 has the potential to contribute to the
development of a number of Pb-induced health effects (Table 1-4). Evidence for the
majority of these modes of action is observed at low blood Pb levels in humans and
laboratory animals, between 2 and 5 (ig/dL, and at doses as low as the picomolar range in
animals and cells. The concentrations eliciting the modes of action (reported in Table
1-5) are drawn from the available data and do not imply that these modes of action are
not acting at lower exposure levels or that these doses represent the threshold of the
effect. Also, the data in presented this table does not inform regarding the exposure
frequency and duration required to elicit a particular MOA.
Ecosystem studies have presented evidence for the occurrence of many of these modes of
action in animals, and to some degree in plants, however the connection to ecological
outcomes must usually be inferred because ecological studies are typically not designed
to address mode of action directly. The level at which Pb elicits a specific effect is more
difficult to establish in terrestrial and aquatic systems due to the influence of
environmental variables on Pb bioavailability and toxicity and substantial species
differences in Pb susceptibility.
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Table 1-5 Modes of action, their related health effects, and information on
concentrations eliciting the MOAs.
Mode of Action
[Related Health Effects
(ISA Section)]
Concentrations or Doses (Conditions)3
Blood Pb
Dose
Altered Ion Status
[All Health Effects of Pb;
Chapter 41
3.5 |jg/dL
(Mean in cord blood; association with
cord blood Ca2+ATPase pump activity)
Huel et al. (2008)
0.00005 uM free Pb2+
(In vitro; 30 minutes; calmodulin
activation assay)
Kern et al. (2000)
Protein Binding
[Renal (45),
Heme Synthesis and
RBC Function (4.7)1
17.0ug/dL
(Concurrent mean in adult workers with
wildtype metallothionein expression;
increased BP susceptibility)
Chen et al. (201 Oa)
50 uM Pb glutamate
(In vitro; 24 hours; increased nuclear
protein in neurological cell)
Klann and Shelton (1989)
Oxidative Stress
[All Heath Effects of Pb
(Chapter 4)1
5.4 ug/dl_
(Concurrent mean in adult male
workers; decreased CAT activity in
blood)
Conterato et al. (2013)
0.1 uM Pb acetate
(In vitro; 48 hours; decreased cellular
GSH in neuroblastoma cells)
Chetty et al. (2005)
Inflammation
[Nervous System (4.3).
Cardiovascular (4.4).
Renal (45),
Immune (4.6),
Respiratory (4.6.5 and
4.9.6),
Hepatic (4.9.1)1
Among males with concurrent blood Pb
>2.5ug/dL
(Increased serum TNF-a and blood
WBC count)
Kim et al. (2007)
0.01 uM Pb acetate
(In vitro; 48 hours; increased cellular
PGE2 in neuroblastoma cells)
Chetty et al. (2005)
Endocrine Disruption
[Reproductive and
Developmental Effects (4.8),
Endocrine System (4.9.3),
Bone and Teeth (4.9.4)1
1.7ug/dL
(Lowest blood Pb level at which a
relationship could be detected in adult
women with both ovaries removed;
increased serum follicle stimulating
hormone [FSH])
Krieg (2007)
10 uM Pb nitrate
(In vitro; 30 minutes; displaced GHRH
binding to rat pituitary receptors)
Lau et al. (1991)
Cell Death/Genotoxicity
[Cancer (4.10),
Reproductive and
Developmental Effects (4.8),
Bone and Teeth (4.9.4)1
3.3 ug/dl_
(Concurrent median in adult women;
increased rate of hypoxanthine guanine
phospho ribosyltransferase reporter
gene [HPRT] mutation frequency)
Van et al. (2004)
0.03 uM Pb acetate
(In vitro; 18 hours; increased formation
of micronuclei)
Bonacker et al. (2005)
"This table provides examples of studies that report effects with low doses or concentration; they are not the full body of evidence
used to characterize the weight of the evidence. In addition, the levels cited are reflective of the data and methods available and
do not imply that these modes of action are not acting at lower Pb exposure or blood Pb levels or that these doses represent the
threshold of the effect. Additionally, the blood concentrations and doses (indicating Pb exposure concentrations from in vitro
systems) refer to the concentrations and doses at which these modes of action were observed. While the individual modes of
action are related back to specific health effects sections (e.g., Nervous System, Cardiovascular), the concentrations and doses
given should not be interpreted as levels at which those specific health effects occur.
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The alteration of cellular ion status (including disruption of Ca2+ homeostasis, altered ion
transport mechanisms, and perturbed protein function through displacement of metal
cofactors) appears to be the major unifying mode of action underlying all subsequent
modes of action in plants, animals, and humans (Figure 4-1). Pb can interfere with
endogenous cation homeostasis, necessary as a cell signal carrier mediating normal
cellular functions. Pb is able to displace metal ions, such as Zn, Mg, and Ca2+, from
proteins due to the flexible coordination numbers and multiple ligand binding ability of
Pb, leading to abnormal conformational changes to proteins and altered protein function.
Disruption of ion transport leading to increased intracellular Ca2+ levels is due in part to
the alteration of the activity of transport channels and proteins, such as Na+/K+ATPase
and voltage-sensitive Ca2+ channels. Pb can interfere with these proteins through direct
competition between Pb and the native metals present in the protein metal binding
domain or through disruption of proteins important in Ca2+-dependent cell signaling, such
as protein kinase C (PKC) or calmodulin.
This competition between metals has been reported not only in human systems, but also
in fish, snails, and plants. Altered Ca2+ channel activity and binding of Pb with
Na+/K+ATPase in the gills offish disrupts the Na+ and Cl" homeostasis, which may lead
to ionoregulatory failure and death. Ca2+ influx and ionoregulation has also been shown
to be inhibited by Pb exposure in a sensitive species of snail, leading to a reduction in
snail growth. In plants, substitution of the central atom of chlorophyll, Mg, by Pb
prevents light-harvesting, resulting in a breakdown of photosynthesis. Pb-exposed
animals also have decreased cellular energy production due to perturbation of
mitochondrial function.
Disruption of ion transport not only leads to altered Ca2+ homeostasis, but can also result
in perturbed neurotransmitter function. Evidence for these effects in Pb-exposed
experimental animals and cell cultures has been linked to altered neurobehavioral
endpoints and other neurotoxicity. Neurobehavioral changes that may decrease the
overall fitness of the organism have also been observed in aquatic and terrestrial
invertebrate and vertebrate studies. There is evidence in tadpoles and fish to suggest Pb
may alter neurotransmitter concentrations, possibly resulting in some of these
neurobehavioral changes.
Altered cellular ion status following Pb exposure can result in the inhibition of heme
synthesis. Pb exposure is commonly associated with altered hematological responses in
aquatic and terrestrial invertebrates, experimental animals, and human subjects. The
proteins involved in heme synthesis that are affected by Pb are highly conserved across
species, which may account for the common response seen in human health and
ecological studies. This evolutionarily conserved response to Pb is likely the result of the
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competition of Pb with the necessary metal cofactors in the proteins involved in heme
synthesis.
Although Pb will bind to proteins within cells through interactions with side group
moieties, thus potentially disrupting cellular function, protein binding of Pb may
represent a mechanism by which cells protect themselves against the toxic effects of Pb.
Intranuclear and intracytosolic inclusion body formation has been observed in the kidney,
liver, lung, and brain following Pb exposure in experimental animals. A number of
unique Pb binding proteins have been detected, constituting the observed inclusion
bodies. The major Pb binding protein in blood is ALAD with carriers of the ALAD-2
allele potentially exhibiting higher Pb binding affinity. Inhibition of ALAD activity is a
widely recognized response to Pb in environments where Pb is present and is considered
to be biomarker of Pb exposure in both terrestrial and aquatic biota. Additionally,
metallothionein is an important protein in the formation of inclusion bodies and
mitigation of the toxic effects of Pb. Protein binding of Pb is a recognized mechanism of
Pb detoxification in some terrestrial and aquatic biota. For example, plants can sequester
Pb through binding with phytochelatin and some fish have the ability to store
accumulated Pb in heat-stable proteins.
A second major mode of action of Pb is the development of oxidative stress, due in many
instances to the antagonism of normal metal ion functions. Disturbances of the normal
redox state of tissues can cause toxic effects and is involved in the majority of health and
ecological outcomes observed after Pb exposure. The origin of oxidative stress produced
after Pb exposure is likely a multi-pathway process. Studies in humans and experimental
animals provide evidence to conclude that oxidative stress results from oxidation of
5-ALA, NAD(P)H oxidase activation, membrane and lipid peroxidation, and antioxidant
enzyme depletion. Evidence of increased lipid peroxidation associated with Pb exposure
exists for many species of plants, invertebrates, and vertebrates. Enhanced lipid
peroxidation can also result from Pb potentiation of Fe2+ initiated lipid peroxidation and
alteration of membrane composition after Pb exposure. Increased Pb-induced ROS will
also sequester and inactivate biologically active nitric oxide or nitric oxide radical (NO),
leading to the increased production of the toxic product nitrotyrosine, increased
compensatory NOS, and decreased sGC protein. Pb-induced oxidative stress not only
results from increased ROS production but also through the alteration and reduction in
activity of the antioxidant defense enzymes. The biological actions of a number of these
enzymes are antagonized due to the displacement of the protein functional metal ions by
Pb. Increased ROS are often followed by a compensatory and protective upregulation in
antioxidant enzymes, such that this observation is indicative of oxidative stress
conditions. A number of studies in plants, invertebrates, and vertebrates present evidence
of increased antioxidant enzymes with Pb exposure. Additionally, continuous ROS
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production may overwhelm this defensive process leading to decreased antioxidant
activity and further oxidative stress and injury.
In a number of organ systems Pb-induced oxidative stress is accompanied by
misregulated inflammation. Pb exposure will modulate inflammatory cell function,
production of proinflammatory cytokines and metabolites, inflammatory chemical
messengers, and proinflammatory signaling cascades. Cytokine production is skewed
toward the production of proinflammatory cytokines like TNF-a and IL-6 as well as
leading to the promotion of Th2 response and suppression of Thl cytokines and
Thl-related responses.
Pb is a potent endocrine disrupting chemical. Steroid receptors and some endocrine
signaling pathways are known to be highly conserved over a broad expanse of animal
phylogeny. Pb can disrupt the HPG axis evidenced in humans, other mammals, and fish,
by a decrease in serum hormone levels, such as FSH, LH, testosterone, and estradiol. Pb
interacts with the hypothalamic-pituitary level hormone control causing a decrease in
pituitary hormones, altered growth dynamics, inhibition of LH secretion, and reduction in
StAR protein. Pb has also been shown to alter hormone receptor binding likely due to
interference of metal cations in secondary messenger systems and receptor ligand binding
and through generation of ROS. Pb disrupts hormonal homeostasis in invertebrates
necessary for reproduction and development. Pb also may disrupt the HPT axis by
alteration of a number of thyroid hormones, possibly due to oxidative stress. These
studies have been conducted in humans and other animals, including cattle; however the
results of these studies are mixed.
Genotoxicity and cell death has been investigated after Pb exposure in humans, animals,
plants, and cell models. High level Pb exposure to humans leads to increased DNA
damage, however lower blood Pb levels have been associated with these effects in
experimental animals and cells. Reports vary on the effect of Pb on DNA repair activity;
however, a number of studies report decreased repair processes following Pb exposure.
There is some evidence in plants, earthworms, freshwater mussels and fish for DNA
damage associated with Pb exposure. There is evidence of mutagenesis and clastogenicity
in highly exposed humans, however weak evidence has been shown in animals and cells
based systems. Human occupational studies provide limited evidence for micronucleus
formation (blood Pb levels >10 (ig/dL), supported by Pb-induced effects in both animal
and cell studies. Micronucleus formation has also been reported in amphibians. Animal
and plant studies have also provided evidence for Pb-induced chromosomal aberrations.
The observed increases in clastogenicity may be the result of increased oxidative damage
to DNA due to Pb exposure, as co-exposures with antioxidants ameliorate the observed
toxicities. Limited evidence of epigenetic effects is available, including DNA
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methylation, mitogenesis, and gene expression. Altered gene expression may come about
through Pb displacing Zn from multiple transcriptional factors, and thus perturbing their
normal cellular activities. Consistently positive results have provided evidence of
increased apoptosis of various cell types following Pb exposure.
Overall, Pb-induced health and ecological effects can occur through a number of
interconnected and evolutionarily well conserved modes of action that generally originate
with the alteration of ion status.
1.9 Policy Relevant Considerations
1.9.1 Public Health Significance
The rationale for establishing the public health significance of the various health
endpoints associated with Pb exposure is multifaceted. The 2006 Pb AQCD (U.S. EPA.
2006b) concluded that neurodevelopmental effects in children and cardiovascular effects
in adults were among the effects best substantiated as occurring at the lowest blood Pb
levels, and that these categories of effects were clearly of the greatest public health
concern. The evidence reviewed in the current assessment supports and builds upon this
conclusion. Evidence in a few cohorts of children that indicated the supralinear
concentration-response blood Pb-FSIQ relationships, did not identify a threshold for
Pb-associated neurodevelopmental effects in the range of blood Pb levels examined
(Sections 1.9.3 and 4.3.13V
The World Health Organization (WHO) definition of "health" is "the state of complete
physical, mental, and social well-being and not merely the absence of disease or
infirmity" (WHO. 1948). By this definition, decrements in health status that are not
severe enough to result in the assignment of a clinical diagnosis might reflect a decrement
in the well-being of an individual. Further, deficits in subtle indices of health or
well-being may not be observable except in aggregate, at the population level, so the
critical distinction between population and individual risk is essential for interpreting the
public health significance of study findings. This concept of population risk is relevant to
the interpretation of findings regarding both IQ and blood pressure in the assessment of
their public health significance.
Weiss et al. (1988) discusses the hypothetical impact of a small shift in a population
distribution of IQ Score. As shown in Figure 1-1. these authors anticipate that even a
small shift in the population mean IQ may be significant from a public health perspective
because such a shift could yield a larger proportion of individuals functioning in the low
1-68
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range of the IQ distribution, which is associated with increased risk of educational,
vocational, and social failure (Section 4.3.13). as well as reduce the proportion of
individuals with high IQ scores.
0.03-1
0.02
0.01
0.00
• 0.02-
0.01-
0.00
so
70
90 no
IQ Score
130
ISO
so
70
90 no
IQ Score
130
ISO
Note: Two distributions of intelligence test scores. (Left): Based on a mean of 100 (the standardized average, with a standard
deviation of 15. (Right):Demonstrating a 5% reduction model, based on a mean score of 95. This is a conceptual model that
assumes that the incremental concentration-response between Pb exposure and IQ is similar across the full range of IQ, and is not
based on actual data. The figure shows that the effect of a small shift in population mean IQ score may result in a larger proportion
of individuals with IQ scores below 70 and a smaller proportion with IQ scores above 130.
Source: Reprinted with permission of Elsevier; from Weiss et al. (1988)
Figure 1-1
Distributions of IQ scores.
It is also important to note that the change in a population mean observed in an
epidemiologic study may be small compared to the standard error of measurement for the
outcome. Measurement error in the outcome can affect the likelihood of detecting an
association but if a study is large enough it will have adequate statistical power to detect
small changes at the population level. Bias may be introduced if the measurement error of
the outcome is highly correlated with the exposure. There is no evidence to suggest that
individuals with higher blood Pb levels test systematically lower than their true IQ.
Pb-associated changes in blood pressure also increase an individual's risk for health
effects that are of greater clinical consequence than is suggested by a small individual
change in blood pressure. Nawrot et al. (2002) found that a doubling of blood Pb was
associated with an approximate 1 mmHg increase in systolic blood pressure. Results from
the Framingham Heart Study show that higher levels of blood pressure, even within the
nonhypertensive range, increase the risk 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
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cases, but mainly in those with high normal blood pressure (i.e., mild hypertension).
Kannel (2000a) emphasized that systolic blood pressure exerts a strong influence on more
serious cardiovascular events, as it is the primary cause of hypertension and its adverse
cardiovascular sequelae. Pb-associated effects on cardiovascular morbidity outcomes
such as ischemic heart disease (Section 4.4.3.6) and mortality (Section 4.4.5) have also
been observed. In addition, some groups within the population can be at greater risks for
cardiovascular effects; as summarized in Chapter 5. there is evidence for increased
cardiovascular effects based on race/ethnicity and several genetic markers. Overall, while
some of the specific health endpoints that have been associated with Pb exposure are
small physiological changes in an individual, these changes can represent substantial risk
at the population level.
1.9.2 Air-Pb-to-Blood-Pb Relationships
The 1986 Pb AQCD described epidemiologic studies of relationships between air Pb and
blood Pb. Much of the pertinent earlier literature for children described in the 1986 Pb
AQCD was included in a meta-analysis by Brunekreef (1984). Based on the studies
available at that time, the 1986 Pb AQCD concluded that "the blood Pb versus air Pb
slope (3 is much smaller at high blood and air levels." This is to say that the slope (3 was
much smaller for occupational exposures where high blood Pb levels (>40 ug/dL) and
high air Pb levels (much greater than 10 ug/m3) prevailed relative to lower environmental
exposures which showed lower blood Pb and air Pb concentrations (<30 ug/dL and
<3 ug/m3). For those environmental exposures, it was concluded that the relationship
between blood Pb and air Pb "... for direct inhalation appears to be approximately linear
in the range of normal ambient exposures (0.1-2.0 ug/m3)" (pp 1-98 of the 1986 Pb
AQCD). In addition to the meta-analysis of Brunekreef (1984). more recent studies have
provided data from which estimates of the blood Pb-air Pb slope can be derived for
children (Table 1-6. Table 3-12). The range of estimates from these studies is 4-9 ug/dL
per ug/m3, which encompasses the estimate from the Brunekreef (1984) meta-analysis.
Most studies have described the blood Pb-air Pb relationship as either log-log (Schnaas et
al.. 2004; Hayes et al.. 1994; Brunekreef. 1984). which predicts an increase in the blood
Pb-air Pb slope with decreasing air Pb concentration or linear (Hilts. 2003; Tripathi et al..
2001; Schwartz and Pitcher. 1989). which predicts a constant blood Pb-air Pb slope
regardless of air Pb concentrations. These differences may simply reflect model selection
by the investigators; alternative models are not reported in these studies.
The blood Pb-air Pb slope may also be affected in some studies by the inclusion of
parameters (e.g., soil Pb) that may account for some of the variance in blood Pb
attributable to air Pb. Other factors that likely contribute to the derived blood Pb-air Pb
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slope include differences in the populations examined and Pb sources, which varied
among individual studies. See Section 3.5 for a detailed discussion of studies that inform
air Pb-to blood-Pb relationships.
Table 1-6 Summary of estimated slopes for blood Pb to air Pb relationships
in children.
Reference
Study Methods
Model Description
Blood Pb-
AirPb
Slope3
Brunekreef (1984)
Location: Various countries
Years: 1974-1983
Subjects: Children (varying age ranges,
N>190,000)
Analysis: Meta analysis of 96 child populations
from 18 study locations
Model: Log-Log
Blood Pb: 5-76 ug/dL
(mean range for studies)
AirPb:0.1-24ug/m3
(mean range for studies)
All children:
4.6(1.5)b
Children
<20 ug/dL:
4.8 (0.54)c
Hayes et al. (1994)
Location: Chicago, IL
Years: 1974-1988
Subjects: 0.5-5 yr (N = 9,604)
Analysis: Regression of quarterly median blood
Pb and quarterly mean air Pb
Model: Log-Log
Blood Pb: 10-28ug/dL
(quarterly median range)
Air Pb: 0.05-1.2 ug/m3
(quarterly mean range)
8.2 (0.62)d
Hilts (2003)
Location: Trail, BC
Years: 1996-2001
Subjects: 0.5-5 yr, 1996-2000; 0.5-3 yr, 2001
(Estimated N = 220-460, based on 292-536
blood Pb measurements/yr with 75-85%
participation).
Analysis: Regression of blood Pb screening and
community air Pb following upgrading of a local
smelter
Model: Linear
Blood Pb: 4.7-11.5 ug/dL
(annual geometric mean
range)
Air Pb: 0.03-1.1 ug/m3
(annual geometric mean
range)
7.0 (0.48)e
Schwartz and
Pitcher (1989).
U.S. EPA(1986a)
Location: Chicago, IL
Years: 1976-1980
Subjects: Black children, 0-5 yr (N = 5,476)
Analysis: Multivariate regression of blood Pb
with mass of Pb in gasoline (derived from
gasoline consumption data and Pb
concentrations in gasoline for the U.S.)
Model: Linear
Blood Pb: 18-27ug/dL
(mean range)'
Air Pb: 0.36-1.22 ug/m3
(annual maximum
quarterly mean)h
8.6 (0.75)g
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Table 1-6 (Continued): Summary of estimated slopes for blood Pb to air Pb relationships
in children.
Blood Pb-
AirPb
Reference Study Methods Model Description Slope3
Location: Mumbai, India (multiple residential Model: Linear
locations) Blood Pb: 8.6-14.4 pg/dL
Years- 1984-1996 (GM ran9e for residential
Tripathi et al. (2001) locations) 3.6 (0.45)'
K v ' Subjects: 6-10 yr N = 544 ' 3 v '
1 y v ' Air Pb: 0.11-1.18 ug/m3
Analysis: Regression of residential location- (GM e for residential
specific average blood Pb and air Pb data locations)
"Slope is predicted change in blood Pb (|jg/dL per |jg/m3) evaluated at ± 0.01 ug/m3 from central estimate of air Pb for the study
(shown in parentheses). The central estimate for the Brunekreef (1984) study, is the median of air Pb concentrations, since it was a
meta-analysis; for all other studies the mean is presented. For multiple regression models, this is derived based only on air Pb
coefficient and intercept. Depending on the extent to which other variables modeled also represent air Pb, this method may
underestimate the slope attributable to air pathways. In single regression models, the extent to which non-modeled factors,
unrelated to air Pb exposures, exert an impact on blood Pb that co-varies with air Pb may lead to the slope presented here to over
represent the role of air Pb.
bln(PbB) = In(PbA) x 0.3485 + 2.853
°ln(PbB) = In(PbA) x 0.2159 + 2.620
dln(PbB) = In(PbA) x 0.24 + 3.17
ePbB = PbA x 7.0
'Observed blood Pb values not provided; data are for regressed adjusted blood Pb.
9PbB = PbA x 8.6
"Based on data for the U.S. [1986 Pb AQCD, (U.S. EPA. 1986a)1.
!PbB =PbA x3.6
GM, geometric mean; GSD, geometric standard deviation; PbB, blood Pb concentration (ug/dL); PbA, air-Pb concentration (ug/m3)
1.9.3 Concentration-Response Relationships for Human Health Effects
Concentration response (C-R) relationships have been examined most extensively in
studies of neurodevelopmental effects in children. Although relatively few studies
examined the shape of the concentration-response relationship between Pb in blood or
bone and effects in adults, several recent studies of adult endpoints (i.e., cognitive
function, cardiovascular and mortality effects) add to the evidence. Some of the
populations examined (e.g., NHANES, NAS) are likely to have had higher past than
recent Pb exposure. Other populations (e.g., worker populations) studied have ongoing
exposure to Pb. As described elsewhere in the document (Sections 3.3. 4.3. 4.4. and 4.5).
the interpretation of the study findings depends on the exposure history and the choice of
the biomarker in the context of what is known about that exposure history. There is
uncertainty regarding the frequency, duration, timing and level of exposure contributing
to the blood Pb or bone Pb levels in the adult populations studied.
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Cognitive and Behavioral Effects in Children
With each successive Pb AQCD and supplement, the epidemiologic and toxicological
study findings show that progressively lower blood Pb levels or Pb exposures are
associated with cognitive deficits in children (Section 4.3.13). For example, effects were
observed in association with blood Pb levels in the range of 10-15 (ig/dL in the 1986
Addendum (U.S. EPA. 1986c) and 1990 Supplement (U.S. EPA. 1990a). and 10 (ig/dL
and lower in the 2006 Pb AQCD (U.S. EPA. 2006b). No evidence of a threshold for the
effects of Pb on neurodevelopmental effects has been reported across the range of blood
Pb levels examined in epidemiologic studies.
Compelling evidence for a larger decrement in cognitive function per unit increase in
blood Pb among children with lower mean blood Pb concentrations compared to children
with higher mean blood Pb concentrations was presented in the 2006 Pb AQCD. Key
evidence was provided by studies that examined prenatal or early childhood blood Pb
levels or considered peak blood Pb levels in schoolaged children or concurrent blood Pb
levels in young children age 2 years (Section 4.3.12. Figure 4-16. and Table 4-16)
(Tellez-Rojo et al.. 2006) as well as the international pooled analysis of seven prospective
cohort studies by Lanphear et al. (2005). a subsequent reanalysis of these data focusing
on the shape of the concentration response function (Rothenberg and Rothenberg. 2005).
Attenuation of C-R relationships at higher exposure or dose levels has been reported in
the occupational literature for a range of exposures. Reasons proposed to explain the
attenuation include greater exposure measurement error and saturation of biological
mechanisms at higher levels as well depletion of the pool of susceptible individuals at
higher exposure levels (Stavner et al.. 2003). Possible explanations specific to nonlinear
relationships observed in studies of Pb exposure in children include a lower incremental
effect of Pb due to covarying risk factors such as low SES, poor caregiving environment,
and higher exposure to other environmental factors (Schwartz. 1994). differential activity
of mechanisms at different exposure levels, and confounding by omitted or misspecified
variables. Review of the evidence did not reveal a consistent set of covarying risk factors
to explain the differences in blood Pb IQ C-R relationship across high and low Pb
exposure groups observed in epidemiologic studies. Nonlinear concentration-response
relationships including U- or inverted U-shaped curves for various endpoints, including
those related to cognitive impairment were demonstrated in the toxicological literature.
However, these toxicological findings are distinct from epidemiologic findings of
supralinear relationships in that some U- or inverted U-shaped relationships do not
indicate Pb-induced impairments at higher exposure concentrations.
The supralinear relationship reported in multiple prospective studies does not provide
evidence of a threshold for Pb-associated cognitive function decrements. For example, as
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detailed in Section 4.3.12. higher blood Pb levels at age 2 years were associated with
FSIQ decrements in children aged 10 years whose blood Pb levels were in the range of
1.0-9.3 ug/dL, e.g.(Bellinger. 2008). Supporting evidence was provided by Pb-associated
decrements in academic performance observed in fourth grade children with earlier
childhood blood Pb levels categorized as 2 (ig/dL versus 1 (ig/dL (Miranda et al.. 2009;
2007a). The lack of a reference population with blood Pb levels reflecting still lower Pb
exposures limits the ability to identify a threshold in the current population.
Toxicological studies showed that lower Pb exposures (e.g., 50 ppm in drinking water)
induced learning and memory impairments in animals compared to control exposures or
higher Pb exposures (e.g., 150 ppm). Additional toxicological evidence suggests that
mechanisms may be differentially activated at lower and higher Pb exposures, and
reduced long-term potentiation (LTP) and hippocampal glutamate release with lower Pb
exposures may provide explanation for the impaired learning and memory observed with
lower Pb exposures in some animal studies.
Studies of Pb Effects in Adults
The shape of the C-R function (e.g., linear versus non-linear) was not examined in most
studies of the association of Pb biomarkers with cognitive function in adults
(Sections 4.3.2.7 and 4.3.13). Log-linear models were used to fit the data in NHANES
analyses. Nonlinearity in the relationship between bone Pb and cognitive function among
participants in the BMS and NAS cohorts was examined with the use of quadratic terms,
penalized splines, or visual inspection of bivariate plots. There was some evidence for
nonlinearity in prospective analyses of the NAS cohort (Figure 4-7 and Figure 4-8). but
not all results indicated a larger decrement in cognitive function per unit increase in bone
Pb level in the lower end of the bone Pb distribution. In the BMS cohort, observation of a
statistically nonsignificant quadratic bone Pb term indicated that a linear model fit the
relationship between tibia Pb level and various tests of cognitive function.
A meta-analysis of human studies found that each doubling of blood Pb level (between 1
and >40 ug/dL measured concurrently in most studies of adults for which past exposures
were likely higher than current exposures) was associated with a 1 mmHg increase in
systolic BP and a 0.6 mmHg increase in diastolic BP (Nawrot et al.. 2002). In this
analysis, effect sizes were adjusted for the purpose of pooling them depending on
whether a linear or log (common or natural) linear model was used. The functional form
of the C-R relationship was examined in few individual studies of cardiovascular effects
(Section 4.4.2.1). Weaver (2010). reported that a logarithmic function of blood Pb level
better described data from a cohort of Korean workers than the linear form. Only a small
number of studies that focused on Pb-induced hypertension in experimental animals have
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included more than two exposure concentrations, and these studies appear to show a
nonlinear concentration-response (Figure 4-21).
Studies investigating both all-cause and cardiovascular mortality report both linear and
non-linear relationships (Section 4.4.5). Although associations are consistently reported,
findings regarding the shape of the C-R relationship between blood Pb level and
mortality in NHANES analyses were mixed. For example, in the NAS cohort, C-R
relationships between bone Pb and mortality were approximately linear for patella Pb on
the log(heart rate [HR]) scale for all cardiovascular disease (CVD), but appear nonlinear
for IHD (Weisskopf et al.. 2009). It is important to note the wide confidence limits,
which increase uncertainty at the lower and upper bounds of patella Pb levels. The
strongest associations were observed between mortality and baseline patella Pb
concentration while tibia Pb levels were more weakly associated with CVD mortality.
1.9.4 Pb Exposure and Neurodevelopmental Deficits in Children
As discussed in Section 3.3.5. blood Pb may reflect both recent exposures as well as past
exposures because Pb is both taken up by and released from the bone. The relative
proportion of blood Pb from recent versus past exposure is uncertain in the absence of
specific information about the pattern of exposure contributing to observed blood Pb
levels. This uncertainty is greater in adults and older children, than in young children who
do not have lengthy exposure histories. Several lines of evidence, which are summarized
below, inform the interpretation of epidemiologic studies of young children with regard
to the patterns of exposure that contribute to observed health effects.
Pb can cross the placenta to affect the developing fetal nervous system and fetal Pb
exposure can occur from recent maternal exposure or from mobilization of bone Pb stores
from past exposures (Section 3.2.2.4). In very young children, ages <2 years, decrements
in mental development, as assessed with MDI, was associated with higher prenatal
(maternal and cord) and concurrent blood Pb levels (Section 4.3.2.2). Thus, both
postnatal child and maternal Pb exposures may contribute to neurodevelopmental effects
in children from infancy to age 2 years. There is some evidence that the relative influence
of maternal Pb levels on postnatal blood Pb level is substantially reduced soon afterbirth
(Section 3.4). There was also a good correlation between child blood Pb level and child
hand Pb loading (R2 = 0.70) in a study following children living in a contaminated area,
indicating the influence of concurrent Pb exposures on blood Pb during the early
childhood years (Simon et al.. 2007). In another study (Carbone etal.. 1998) blood Pb
levels of infants aged 6-12 months were significantly lower than their neonatal cord
blood Pb levels (2.24 ug/dL versus 4.87 ug/dL). Among infants born with blood Pb levels
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of greater than 7 ug/dL, who were followed for a week, there was a dramatic drop in the
blood Pb (Carbone et al.. 1998).
Epidemiologic studies consistently show that blood Pb levels measured during various
lifestages or time periods throughout childhood, as well as averaged over multiple years
during childhood, are associated with cognitive function decrements (Section 4.3.11). An
international pooled analysis of seven prospective studies found that increments in
concurrent and peak blood Pb levels were associated with a decrease in FSIQ measured
between ages 5 and 10 years (Lanphear et al., 2005). In individual studies, postnatal
(early childhood and concurrent) blood Pb levels are also consistently associated with
cognitive function decrements in children and adolescents (Figure 4-2. Table 4-3. Table
4-14).
Exposure metrics based on blood Pb measurements at different ages in childhood are
typically highly correlated. For example, analyses of serial blood Pb concentrations
measured in longitudinal epidemiologic studies find relatively strong correlations (e.g.,
r = 0.5-0.8) among each child's individual blood Pb concentrations measured after 6-12
months of age (Section 3.3.2). Consequently, the relative importance of various exposure
metrics, which are measured during different lifestages and time periods, is difficult to
discern in epidemiologic studies. Evidence in rodents and monkeys, however, indicates
that Pb exposures during multiple lifestages and time periods, including prenatal only,
prenatal plus lactational, postnatal only, lifetime are observed to induce impairments in
learning (Rice. 1992b. 1990; Rice and Gilbert. 1990b). These findings are consistent with
the understanding that the nervous system continues to develop (i.e., synaptogenesis and
synaptic pruning remains active) throughout childhood and into adolescence.
1.9.5 Reversibility and Persistence of Neurotoxic Effects of Pb
The 2006 Pb AQCD concluded that the human and animal evidence suggest that the
neurotoxic effects of Pb are not generally reversible (U.S. EPA. 2006b). Chelation studies
in humans and animals show that chelation decreases total body Pb burden, but does not
necessarily exert evident effects on Pb-induced cognitive deficits. For example, analysis
of multi-center study data indicates that medical interventions involving chelation therapy
(e.g., Succimer use) do not fully reverse cognitive deficits associated with early Pb
exposure (Liu et al.. 2002).
The persistence of neurodevelopmental effects from comparatively low-level Pb
exposure was considered in the 2006 Pb AQCD (U.S. EPA. 2006b). with some evidence
suggesting that the effects of Pb on neurodevelopmental outcomes persisted into
adolescence and young adulthood. The toxicological evidence continues to support a
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range of effects with prenatal and early postnatal Pb exposures that persist to adulthood
(Sections 4.3.2.3 and 4.3.3.1). In rats, persistent neurobehavioral deficits were observed
with prenatal, preweaning, and postweaning Pb exposure. In monkeys, impairments were
found in learning and short-term memory at ages 7 to 8 years (Rice and Karpinski. 1988)
and in attention at ages 9 to 10 years (Gilbert and Rice. 1987) with lifetime Pb exposure
that did not begin until postnatal day 400 and that produced peak blood-Pb levels <15 or
25 ug/dL and steady-state levels 11 and 13 ug/dL, indicating that postnatal juvenile Pb
exposures were sufficient to produce neurodevelopmental deficits.
A number of mechanisms, including changes in neurogenesis, synaptogenesis and
synaptic pruning, long term potentiation, and neurotransmitter function have been
identified that provide biological plausibility for epidemiologic and toxicological findings
of persistent cognitive and behavioral effects that result from Pb exposures during
prenatal and early childhood periods. Furthermore, the normal dynamic and rapid rate of
development that occurs early in life in the CNS makes insults early in life especially
problematic in that they can permanently change the trajectory of brain development such
that there are little or no compensatory pathways to replace the lost potential for proper
brain development (Bayer. 1989).
The persistence of effects appears to depend on the duration and window of exposure as
well as other factors that may affect an individual's ability to recover from an insult.
Several studies have observed improved cognition in children with declining blood Pb
levels (Hornung et al.. 2009: Chen et al.. 2005a: Liu et al.. 2002: Ruffetal.. 1993). There
is evidence that some cognitive effects of prenatal Pb exposure may be transient and that
recovery is greater among children reared in households with more optimal caregiving
characteristics and in children whose concurrent blood Pb levels were low (Bellinger et
al.. 1990): the animal toxicology literature supports these findings using studies of
Pb-exposed animals that live in enriched environments. However, the extent to which
such improvement represents biological reversibility of Pb-related effects, the influence
of enrichment related intervention, or the development of compensatory mechanisms
remains uncertain.
Toxicological studies in the 2006 Pb AQCD highlighted the importance of Pb exposure
during early life in promoting Alzheimer's-like pathologies in the adult rodent brain, with
Pb-induced neurodegeneration and formation of neurofibrillary tangles in aged animals in
which blood Pb levels had returned to control levels after an earlier life Pb exposure
(U.S. EPA. 2006b). Sensitive windows of early life Pb exposure or a Pb biomarker and
have been associated with changes in adulthood as demonstrated with animal models of
neurodegeneration, i.e., neurofibrillary tangle formation. Behavioral or cognitive testing
in these animal models has not been performed to assess these changes. These effects are
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not reflective of concurrent blood Pb levels at the age of manifestation of the pathology
but instead are associated with an earlier life Pb exposure.
1.9.6 Populations Potentially At-Risk for Health Effects
The NAAQS are intended to protect public health with an adequate margin of safety. In
so doing, protection is provided for both the population as a whole and those groups
potentially at increased risk for health effects from exposure to the air pollutant for which
each NAAQS is set (Preface to this ISA). To facilitate the identification of populations at
increased risk for Pb-related health effects, studies have evaluated various factors that
may contribute to susceptibility and/or vulnerability to Pb. These factors include genetic
background, race and ethnicity, sex, age, diet, pre-existing disease, SES, and
characteristics that may modify exposure or the response to Pb. Table 1-7 and Table 5-5
provide an overview of the factors examined as potentially increasing the risk of
Pb-related health effects based on the recent evidence integrated across disciplines. They
are classified according to the criteria discussed in the introduction to Chapters.
In consideration of the evidence base as a whole (e.g., stratified and longitudinal
analyses) and integrating across disciplines of toxicokinetics, exposure, and health, there
is adequate evidence to conclude that children are an at-risk population. It is recognized
that Pb can cross the placenta and affect the developing nervous system of the fetus
(Section 3.2.2.4). Children may have increased exposure to Pb compared with adults
because children's behaviors and activities (including increased hand-to-mouth contact,
crawling, and poor hand-washing), differences in diets, and biokinetic factors. There is
evidence of increased risk to the cognitive effects of Pb exposure during several
lifestages and time periods throughout gestation, childhood, and into adolescence
(Section 4.3.12). Findings from magnetic resonance imaging (MRI) studies indicate that
normal brain development remains dynamic throughout adolescence, and epidemiologic
studies have linked concurrent blood Pb level (as well as other blood Pb metrics) in
adolescents to cognitive function decrements and delinquent or criminal behavior
(Section 4.3.4). Delays in puberty onset (Section 4.8.1). and renal effects
(Section 4.5.2.2). are also observed in association with concurrent blood Pb level in
cross-sectional studies of adolescents. Since the populations of older children in these
studies generally had higher past exposures, the current evidence does not clearly
establish the link between a time and duration of Pb exposure and the observed health
effects in the adolescent populations studied. Elevated biomarkers levels, which may be
related to remobilization of stored Pb during bone loss and/or higher historical Pb
exposures, are observed in older adults. Studies of older adults report inconsistent
findings for effect measure modification of Pb-related mortality by age and no
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modification of other health effects was studied. However, toxicological studies support
the possibility of age-related differences in Pb-related health effects. The overall
evidence, based on limited epidemiologic evidence but support from toxicological studies
and differential exposure studies, is suggestive that older adults are potentially at risk of
Pb effects. However, there are uncertainties related to the exposure profile associated
with the effects in older populations.
Table 1-7 Summary of evidence for factors that potentially increase the risk of
Pb-related health effects.
Factor Evaluated
Childhood (Sections 5.2.1, 5.3.1)
Older Adulthood (Sections 5.2.1 and 5.3.1)
Sex (Sections 5.2.2, 5.3.2)
Genetics (Sections 5.3.3)
Pre-existina Disease3 (Section 5.3.4)
Smokinq Status (Section 5.3.5)
Socioeconomic Status (SES) (Sections 5.2.4, 5.3.6)
Race/Ethnicitv (Sections 5.2.3, 5.3.7)
Proximity to Pb Sources (Section 5.2.5)
Residential Factors (Section 5.2.6)
Body Mass Index (BMI) (Section 5.3.8)
Alcohol Consumption (Section 5.3.9)
Nutrition (Section 5.3.10)
Stress (Section 5.3.11)
Maternal Self-Esteem (Section 5.3.12)
Coqnitive Reserve3 (Section 5.3.13)
Other Metals (Section 5.3.14)
Classification
Adequate
Suggestive
Suggestive
Suggestive
Suggestive
Inadequate
Suggestive
Adequate
Adequate
Adequate
Inadequate
Inadequate
Adequate
Suggestive
Inadequate
Inadequate
Suggestive
"Possible mediator
The evidence regarding the other at-risk factors listed in the table above is summarized in
detail in Section 5.4. Some studies suggest that males at some ages have higher blood Pb
levels than comparably aged females; this was supported by stratifying the total sample
of NHANES subjects. Sex-based differences appeared to be prominent among the
adolescent and adult age groups but were not observed among the youngest age groups
(1-5 years and 6-11 years). Studies of effect measure modification of Pb and various
health endpoints by sex were inconsistent; although it appears that there are some
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differences in associations for males and females. This is also observed in toxicological
studies. Overall, there is suggestive evidence to conclude that sex is a potential at-risk
factor, with adolescent and adult males typically demonstrating higher blood Pb levels,
although evidence regarding health outcomes is limited due to inconsistencies in whether
males or females are at greater risk of certain outcomes in relation to Pb exposure.
Regarding race and ethnicity, recent data suggest that the difference in blood Pb levels
between black and white subjects is decreasing over time, but black subjects still tend to
have higher Pb body burden and Pb exposures than white subjects. Compared to whites,
non-white populations were observed to be more at risk of Pb-related health effects.
Studies of race/ethnicity provide adequate evidence that race/ethnicity is an at-risk factor
based on the higher exposure observed among non-white populations and some
modification observed in studies of associations between Pb levels and health effects. For
example two investigators report (Table 5-5). reported modification by race/ethnicity in
an analysis of hypertension among NHANES III participants. Muntner et al. (2005) sons
of the highest quartile of blood Pb to the lowest, the odds ratio for hypertension was 1.54
(95% CI: 0.99, 2.39) among Mexican Americans, 1.44 (95% CI: 0.89, 2.32) among Non-
Hispanic Blacks and 1.10 (95% 0.87, 1.41) among Non-Hispanic Whites (Scinicariello et
al..201Q).
The gap between SES groups with respect to Pb body burden appears to be diminishing.
However, biomarkers of Pb exposure have been shown to be higher among lower SES
groups even in recent studies in which differences among SES groups have lessened.
Studies of SES and its relationship with Pb-related health effects are few and report
inconsistent finding regarding low SES as a potential at-risk factor. Overall, the evidence
is suggestive that low SES is a potential at-risk factor for Pb-related health effects.
There is adequate evidence that proximity to areas with Pb sources, including areas with
large industrial sources, is associated with increased Pb exposure. Relatively high
concentrations of ambient air Pb have been measured near sources, compared with large
urban areas without sources and high Pb exposures have been documented near
Superfund sites. NHANES analyses report increased Pb biomarker levels related to
increase house dust Pb levels, homes built after 1950, and renovation of pre-1978 homes.
Thus, there is adequate evidence that residing in a residence with sources of Pb exposures
will increase the risk of Pb exposure and associated health effects.
There is suggestive evidence to conclude that various genetic variants potentially modify
the associations between Pb and health effects. Epidemiologic and toxicological studies
reported that ALAD variants may increase the risk of Pb-related health effects. Other
genes examined whose variants may also affect risk of Pb-related health effects were
VDR, DRD4, GSTM1, TNF-a, eNOS, and HFE. Overall the interaction between genetic
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variants and Pb exposure were examined in a small number of studies and these types of
analysis are potentially vulnerable to type II error if multiple statistical tests are
conducted. However, there may be a large potential impact of Pb exposure in specific at-
risk populations carrying specific gene variants. For example, Scinicariello et al. (2010)
found that Non-Hispanic white carriers of the ALAD2 genetic variant in the highest
blood Pb quartile had a 2-fold higher risk of hypertension compared with ALAD1
homozygous individuals (OR:2.00 [95%CI: 1.12, 3.55]). No evidence of effect
modification of the association of Pb with blood pressure by ALAD genotype was
observed in an occupational study of Korean Pb workers, however (Weaver et al.. 2008).
NAS subjects with the H63D polymorphism of the HFE gene had a larger Pb-associated
increase in pulse pressure compared to those with the C282Y variant [i.e., 3.3 mmHg
increase (95%CI: 0.16, 6.46) versus an 0.89 mmHg increase (95%CI 0-5.24) per 13 (ig/g
increase in tibia Pb (Zhang et al.. 2010a)].
Diets sufficient in minerals such as calcium (Ca2+), iron (Fe), and zinc (Zn) offer some
protection from Pb exposure by preventing or competing with Pb for absorption in the GI
tract. Additionally, those with Fe deficiencies were observed to be an at-risk population
for Pb-related health effects in both epidemiologic and toxicological studies. Thus, there
is adequate evidence across disciplines that some nutritional factors contribute to a
population being at increased risk. Other nutritional factors, such as Ca2+, Zn, and protein
intake, demonstrated the potential to modify associations between Pb and health effects
in toxicological studies.
There was suggestive evidence for several other factors as potentially increasing the risk
of Pb-related health effects: pre-existing diseases/conditions, stress, and co-exposure with
other metals. Pre-existing diseases/conditions have the potential to affect the risk of
Pb-related health effects. Recent epidemiologic studies did not support modification of
associations between Pb and health endpoints (i.e., mortality, HRV) by the prevalence of
diabetes; however, past studies have found individuals with diabetes to be an at-risk
population with regard to renal function. Studies of Pb biomarker levels and both renal
effects and HRV demonstrated greater odds of the associations among hypertensive
individuals compared to those who are normotensive. Stress was evaluated as a factor
that potentially increases the risk of Pb-related effects on cognitive function in adults and
hypertension and although limited by the small number of epidemiologic studies,
increased stress was observed to exacerbate the effects of Pb. Toxicological studies
supported this finding. High levels of other metals, such as Cd and Mn, were observed to
result in greater effects for the associations between Pb and various health endpoints
(e.g., renal function, cognitive function in children) but overall the evidence was limited.
Finally, there was inadequate evidence to conclude that smoking, BMI, alcohol
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consumption, maternal self-esteem, and cognitive reserve are potential at-risk factors due
to limited quantities of studies regarding their effect on Pb-related health outcomes.
1.9.7 Ecological Effects and Corresponding Pb Concentrations
Information on the effects of atmospherically-deposited Pb on some ecosystem receptors
in the vicinity of Pb sources is available from studies of terrestrial systems near mining
and smelting operations where Pb, as well as other metals, is present in high
concentrations. In these studies a decreasing gradient of exposure is observed, and effects
in a species or population of interest typically decrease with increasing distance from the
source. Thus, concentrations of Pb in moss, lichens and peat have been used to
understand spatial and temporal distribution patterns of air Pb concentrations. In other
environmental compartments such as sediment and aquatic biota, or in locations not close
to Pb sources, evidence that would permit clear attribution of Pb effects to atmospheric
deposition is insufficient. Pb that is released into air, soil, or water is then cycled through
any or all of these media before reaching an ecological receptor. When a plant,
invertebrate, or vertebrate is exposed to Pb, the proportion of observed effects attributable
to Pb from atmospheric sources is difficult to assess due to a lack of information not only
on deposition, but also on bioavailability, as affected by specific characteristics of the
receiving ecosystem and on kinetics of Pb distribution in long-term exposure scenarios.
Therefore, the connection between air concentration and ecosystem exposure continues to
be poorly characterized for Pb, and the contribution of atmospheric Pb to specific sites is
not clear.
Furthermore, the level at which Pb elicits a specific effect is difficult to establish in
terrestrial and aquatic systems, due to the influence of other environmental variables on
both Pb bioavailability and toxicity, and also to substantial species differences in Pb
susceptibility. Current evidence indicates that Pb is bioaccumulated in biota; however,
the sources of Pb in biota have only been identified in a few studies, and the relative
contribution of Pb from all sources is usually not known. There are large differences in
species sensitivity to Pb, and many environmental variables (e.g., pH, organic matter)
determine the bioavailability and toxicity of Pb.
1.10 Summary
Table 1-8 characterizes the evidence in the 2006 Pb AQCD (U.S. EPA. 2006b) and
previous assessments and compares it to the evidence evaluated in the current
assessment. Evidence regarding both the health and ecological effects of Pb are
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summarized. The purpose of the table is to highlight the extent to which recent evidence
may contribute to current conclusions. The critical assessment of body of evidence as a
whole, however, is discussed in Chapter 4 and Chapter 6 of this document, and
summarized in Sections 1.6 and 1.7. With regard to ecological effects, evidence pointing
to responses in species at ambient or near ambient concentrations is highlighted in the
table.
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Table 1-8 Summary of evidence from epidemiologic, animal toxicological and
ecological studies on the effects associated with exposure to Pb.
Endpoint
Evidence in the 2006 Pb AQCD
Evidence in the 2013 Pb ISA
Health Outcomes:
Nervous System Effects
Children
Cognitive Function
in Children
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 ug/dl_,
and possibly lower." Prenatal, early
childhood, lifetime average, and concurrent
blood Pb levels were associated with
decrements in IQ, learning and executive
function. In some cases, concurrent blood
Pb level was the strongest predictor.
Recent epidemiologic studies in children
continue to demonstrate associations of
concurrent blood Pb level with IQ
decrements; most recent evidence describes
associations of concurrent blood Pb levels
with decrements in cognitive abilities related
to executive function, and academic
performance.
Externalizing
Behaviors:
Attention,
Impulsivity and
Hyperactivity in
Children
Several epidemiologic studies reported
associations between blood and tooth Pb
levels and attention decrements in children
ages 7-17 years and young adults 19-20
years. Most studies examined prenatal or
lifetime average blood Pb levels (means 7,
14 ug/dL) or tooth Pb. The few studies of
concurrent blood Pb levels did not find
associations with attention in children ages
4-5 years. There were no studies
specifically examining ADHD diagnosis.
Uncertainty remained regarding whether Pb
exposure was an independent predictor of
neurobehavioral effects. Prenatal and
postnatal Pb exposure was found to reduce
ability to inhibit inappropriate responding
and increase distractibility in animals.
Recent studies in children continue to
support associations of blood Pb levels with
attention decrements, impulsivity, and
hyperactivity in children ages 7-17 years. A
few case-control studies found higher
concurrent blood Pb levels in children with
diagnosed ADHD; however, because ADHD
studies had various limitations,they were not
a major consideration in drawing
conclusions.
Externalizing
Behaviors:
Conduct Disorders
in Children and
Young Adults
Several epidemiologic studies reported
associations between Pb exposure and
behaviors related to conduct disorders as
rated by parents and teachers and criminal
offenses in children, adolescents, and
young adults. Most studies examined blood
Pb levels measured earlier in childhood
(means 10-14 ug/dl_), tooth Pb, or bone Pb.
There was little examination of concurrent
blood Pb levels.
Recent studies in children continue to
support associations of parent and teacher
ratings of behaviors related to conduct
disorders with early childhood blood Pb
levels and provide new evidence for
concurrent blood Pb levels. Additional follow-
up of previous cohorts to older ages, support
associations of early childhood blood Pb
levels or tooth Pb levels with criminal
offenses in adults ages 19-24 years.
Internalizing Several prospective studies reported
Behaviors in associations of concurrent or childhood
Children average blood, tooth, and bone Pb levels
with parent or teacher ratings of withdrawn
behavior and symptoms of depression,
tearfulness, and anxiety in children ages
3-13 years.
The few recent available studies found
associations between concurrent blood Pb
level and higher ratings of internalizing
behaviors in children ages 8-13 years but
these studies had limited consideration for
potential confounding.
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Table 1-8 (Continued): Summary of evidence from epidemiologic, animal toxicological and
ecological studies on the effects associated with exposure to Pb.
Endpoint
Evidence in the 2006 Pb AQCD
Evidence in the 2013 Pb ISA
Visual Function The selective action of Pb on retinal rod
Decrements in cells and bipolar cells (e.g., ERG effects) of
Children animals is well documented in earlier
AQCDs in animals and found in one study
in children.
Additional evidence for retinal effects was
found in female rats. High-level
developmental Pb exposure did not affect
visual acuity in infant monkeys.
Auditory Function
Decrements in
Childiren
The few available studies reported
associations between concurrent blood Pb
levels (population means 7-12 ug/dL) and
increased hearing thresholds in children.
There was coherence with findings in
animals but with high Pb exposures (e.g.,
blood Pb levels 89-150 ug/dl_).
The few available recent epidemiologic
studies on auditory function in children
examined children with high blood Pb levels
(means >30 ug/dL) and did not consider
potential confounding. Early postnatal Pb
exposure of monkeys increased hearing
thresholds.
Motor Function A small number of studies indicated
Decrements in associations of neonatal, earlier childhood
Children average, lifetime average, and concurrent
blood Pb levels (means: 5-12 ug/dL) with
poorer fine and motor function in children
ages 4.5-17 years. The few toxicological
studies did not consistently find Pb-induced
impairments in balance and coordination in
animals with blood Pb levels >60 ug/dL.
The few recent epidemiologic studies did not
consistently find associations between
concurrent blood Pb level and decrements in
fine motor function. A toxicological study
found poorer balance in male mice with
gestational plus early postnatal Pb exposure
(peak blood Pb level: <10 ug/dl_).
Adults
Cognitive Function
Decrements in
Adults
Among environmentally-exposed adults,
bone Pb levels but not blood Pb levels were
associated with poorer cognitive
performance. These findings point to an
effect of long-term cumulative Pb exposure.
Impaired learning, memory, and executive
function found in adult animals after lifetime
Pb exposures beginning in gestation or
infancy.
Recent studies support previous findings and
recent prospective studies provide new
evidence of associations of bone Pb levels
with subsequent declines in cognitive
function in environmentally-exposed adults
over 2-4 year periods.
Psychopathological
Effects in Adults
Environmentally-exposed adults were not
widely examined; however a study found
associations of concurrent blood and tibia
Pb level with self-reported symptoms of
depression and anxiety in men. Several
studies found higher prevalence of
symptoms related to mood disorders and
anxiety in Pb-exposed workers with mean
blood Pb levels 15-38 ug/dL
Concurrent blood Pb levels were associated
with self-reported symptoms of major
depressive disorder and general anxiety
disorder among men and women
participating in NHANES.
Auditory Function
Decrements in
Adults
A few studies found blood Pb level or
cumulative Pb exposure duration to be
associated with increased hearing
thresholds and hearing loss in Pb-exposed
workers.
A prospective study found higher tibia Pb
level to be associated with a faster rate of
increase in hearing threshold in
environmentally-exposed men over a median
of 23 years.
Visual Function Decreased visual acuity found in adult
Decrements in monkeys with high blood Pb levels
Adults (50-115 ug/dL) after lifetime Pb exposure.
Additional evidence for retinal effects was
found in female rats. A case-control study
found higher retinal Pb levels in adults with
macular degeneration but potential
confounding was not considered. Pb-induced
effects on ERGs in adult animals vary
depending on timing and dose of exposure.
1-85
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Table 1-8 (Continued): Summary of evidence from epidemiologic, animal toxicological and
ecological studies on the effects associated with exposure to Pb.
Endpoint
Evidence in the 2006 Pb AQCD
Evidence in the 2013 Pb ISA
Neurodegenerative
Diseases in Adults
In the limited body of epidemiologic studies,
occupational Pb exposure and brain Pb
levels were not associated with Alzheimer's
disease. Blood Pb levels were not
consistently associated with Amyotrophic
Lateral Sclerosis among environmentally-
exposed adults. A few studies found
associations between occupational Pb
exposure and Parkinson's disease and
blood Pb levels and essential tremor. Each
study had sufficient limitations.
Toxicological studies found Pb-induced
neuronal cell death loss and amyloid
plaques in aged animals with lactational Pb
exposures.
The few case-control studies reported
associations of bone Pb levels with
Parkinson's disease in environmentally-
exposed adults and blood Pb levels with
Amyotrophic Lateral Sclerosis and essential
tremor. Limitations of previous studies apply
to the recent evidence. Recent toxicological
evidence suggests that early-life, not adult-
only Pb exposure may be associated with
neurodegeneration in adult animals.
Cardiovascular Effects
Hypertension
A meta-analysis of numerous epidemiologic
studies estimated that a doubling of blood
Pb level (e.g., from 5 to 10 ug/dL) was
associated with a 1 mmHg increase in
systolic BP and a 0.6 mmHg increase in
diastolic BP."
Epidemiologic studies consistently
demonstrated associations between Pb and
incidence of hypertension with suggestive
evidence that bone Pb may be associated
with hypertension. Animal studies
demonstrated that long-term exposure to
Pb resulted in hypertension that persisted
after cessation of exposure.
Recent epidemiologic and toxicological
studies continue to support associations
between long-term Pb exposure and
increased BP.
Recent studies, including those using bone
Pb as a metric of cumulative exposure,
continue to demonstrate associations of
hypertension with Pb levels in adults. Recent
studies have emphasized the interaction of
cumulative exposure to Pb with other factors
including stress.
Subclinical One NHANES analysis reported an
Atherosclerosis association of blood Pb with PAD
Limited evidence for Pb-induced subclinical
atherosclerosis, including one high-quality
epidemiologic study that reports an
increasing trend in the odds of PAD and
concurrent blood Pb level in adults. Recent
toxicological studies describe a plausible
biological mechanism.
Coronary Heart The evidence for an association of Pb with
Disease cardiovascular mortality was limited but
supportive. A few cross-sectional studies
indicated associations between Pb
biomarker levels and increased risk of CHD
outcomes (i.e., Ml and left ventricular
hypertrophy).
Recent studies address limitations of
previous studies and provide additional
evidence for an association of Pb with
cardiovascular mortality in adults. Specific
causes of mortality that were associated with
Pb could be related to increased BP and
hypertension.
Cerebrovascular
Disease
Evidence available on the risk of
cerebrovascular disease from Pb exposure
was limited to one study of stroke.
Evidence for increased risk of mortality from
stroke is inconsistent.
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Table 1-8 (Continued): Summary of evidence from epidemiologic, animal toxicological and
ecological studies on the effects associated with exposure to Pb.
Endpoint
Evidence in the 2006 Pb AQCD
Evidence in the 2013 Pb ISA
Renal Effects
Reduced Kidney
Function
Circulating and cumulative Pb was
associated with longitudinal decline in renal
function in adults. Toxicological studies
demonstrated that initial accumulation of
absorbed Pb occurred primarily in the
kidneys and noted a hyperfiltration
phenomenon during the first 3 months of
exposure, followed by decrements in kidney
function.
Recent epidemiologic and toxicological
studies add to the body of evidence on
Pb exposure and kidney dysfunction
(e.g., lower creatinine clearance, higher
serum creatinine, and lower GFR) in
nonoccupationally-exposed adults.
Immune System Effects
Increases in Atopic
and Inflammatory
Conditions
Children:
Several epidemiologic studies suggested
that Pb exposure may be associated with
effects on cellular and humoral immunity in
children. The principal effect demonstrated
increases in serum immunoglobulin E (IgE)
levels with concurrent blood Pb levels
>10 ug/dL Toxicological evidence
supported these findings with extensive
evidence for prenatal and early postnatal
Pb exposures skewing toward Th2 cytokine
production and affecting downstream
events such as increases in IgE and
inflammation Several toxicological studies
found a Pb-induced shift to Th2 cytokine
production and a hyperinflammatory
phenotype of macrophages in animals with
long-term (>4 weeks) prenatal or postnatal
Pb exposure.
Recent studies in children added to the
evidence for associations of blood Pb levels
with asthma, allergy, and IgE. The
consistency and coherence of findings
among related immune effects that support a
shift from a Th1 to a Th2 phenotype supports
the biological plausibility for epidemiologic
observations of associations with asthma,
allergy and inflammation-related effects in
other organ systems.
Adults:
Pb exposure-associated immune effects
were not widely examined in
environmentally-exposed adults.
A small body of recently available studies
provide new evidence for increases in
cytokines and other indicators of
inflammation in association with higher
concurrent blood Pb level. A few available
toxicological studies find Pb-associated
increases in cytokines and effects on
dendritic cells in adult mice.
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Table 1-8 (Continued): Summary of evidence from epidemiologic, animal toxicological and
ecological studies on the effects associated with exposure to Pb.
Endpoint
Evidence in the 2006 Pb AQCD
Evidence in the 2013 Pb ISA
Decreases in Host
Resistance
Toxicological evidence demonstrated
Pb-induced increases in bacterial and viral
infection and suppressed DTH in animals.
These effects were supported by extensive
evidence for prenatal and early postnatal
Pb exposures decreasing Th1 cytokine
production, for short-term prenatal Pb
exposure decreasing nitric oxide production
by macrophages), and for long-term (>4
weeks) exposure Pb exposure inducing a
hyperinflammatory phenotype of
macrophages in adult animals.
A few epidemiologic studies found higher
prevalence of respiratory infections in
association with higher blood Pb levels in
children and occupational Pb exposure in
adults; however, studies did not consider
potential confounding.
In the large body of studies in
occupationally-exposed adults, the most
consistent findings were reduced neutrophil
functionality in workers with blood Pb levels
>14.8-91 ug/dL. Environmentally-exposed
adults were not widely examined.
A small body of recent studies supports
previous findings of decreased bacterial
resistance and decreased IFN-yTh1 cytokine
production in animals. Epidemiologic
evidence is limited to an ecological study of
soil and lichen Pb that lacked consideration
for potential confounding.
Autoimmunity
A small number of toxicological studies
found that prenatal and postnatal Pb
treatment, several by i.p. injection,
increased generation of auto-antibodies.
A study found higher auto-antibodies to
neural proteins in Pb-exposed workers with
blood Pb levels 10-40 ug/dL
A recent toxicological study provided indirect
evidence by showing Pb-induced increases
in the activation of neo-antigen specific T
cells, which have the potential to induce
formation of auto-antibodies.
Hematologic System
Red Blood Cell
Function and Heme
Synthesis
Children:
Pb exposure was associated with disruption
in heme synthesis with increases in blood
Pb levels of approximately 20 ug/dL
sufficient to halve ALAD activity and inhibit
ferrochelatase. Risk of clinical anemia in
children becomes apparent at high blood
Pb levels: 10% probability of anemia was
estimated to be associated with -20 ug/dL
Pb at 1 year of age, 50 ug/dL at 3 years of
age, and 75 ug/dL at 5 years of age.
Recent epidemiologic studies provide
evidence that exposure to Pb is associated
with numerous deleterious effects on the
hematological system in children, including
altered hematological parameters (Hb, MCV,
MCH, RBC count), perturbed heme
synthesis mediated through decreased
ALAD and ferrochelatase activities, and
oxidative stress.
Adults:
Pb exposure was associated with disruption
in heme synthesis with increases in blood
Pb levels of approximately 20 ug/dL
sufficient to halve ALAD activity and inhibit
ferrochelatase. Exposures to Pb resulting in
blood concentrations <40 ug/dL appear to
be tolerated without decreases in blood
hemoglobin or hematocrit, however
changes in erythropoiesis do occur at these
blood levels.
Recent epidemiologic studies provide
evidence exposure to Pb is associated with
numerous deleterious effects on the
hematological system in adults, including
altered hematological parameters (Hb, MCV,
MCH, RBC count), perturbed heme
synthesis mediated through decreased
ALAD and ferrochelatase activities,
decreased erythropoiesis, and oxidative
stress.
1-8
-------
Table 1-8 (Continued): Summary of evidence from epidemiologic, animal toxicological and
ecological studies on the effects associated with exposure to Pb.
Endpoint
Evidence in the 2006 Pb AQCD
Evidence in the 2013 Pb ISA
Developmental and Reproductive Effects
Development
Epidemiologic studies reported effects
including delayed puberty in girls. Animal
toxicological studies reported
Pb-associated developmental effects on
teeth, sensory organs, the Gl system, the
liver, and postnatal growth. Delayed
puberty was also observed in both male
and female populations in animal toxicology
studies showing associations with dam
blood Pb levels of -40 ug/dL and pup blood
Pb levels of 26 ug/dL.
Recent toxicological and epidemiologic
studies provide evidence for delayed onset
of puberty in males and females. Most
studies found delayed onset of puberty was
among children ages 6-18 years. These
findings were supported by studies in the
animal toxicological literature showing
effects on puberty onset at blood Pb levels of
3.5-13ug/dL.
Birth Outcomes Toxicological studies reviewed concluded
that Pb exposure can increase fetal
mortality and produce sublethal effects,
smaller litters, reduced birth weight, and
fewer implantation sites. Epidemiologic
studies on preterm birth and low birth
weight/fetal growth reported inconsistent
findings. Epidemiologic studies reported the
possibility of small associations between
increased Pb exposure and birth defects,
and toxicological studies demonstrated
associations between exposure to high
doses of Pb and increased incidences of
teratogenic effects.
Recent toxicological and epidemiologic
studies have reported inconsistent findings
for studies for birth defects, preterm birth,
and low birth weight/fetal growth. A few well-
conducted epidemiologic studies of preterm
birth and low birth weight/fetal growth
reported associations between increased Pb
levels and decreased gestational age and
birth weight/fetal growth.
Male Reproductive
Function
Epidemiologic evidence suggested small
associations between Pb exposure and
male reproductive outcomes including
perturbed semen quality and increased
time to pregnancy. Associations between
Pb exposure and male reproductive
endocrine status were not observed in the
occupational populations studied.
Toxicological studies provided evidence
that Pb produced effects on male and
female reproductive junction and
development and disrupts endocrine
function.
Recent toxicological studies provide strong
evidence for effects on sperm (blood Pb
levels 5-43 ug/dL). Epidemiologic studies
support the association observed in
toxicological studies of Pb exposure and
detrimental effects on sperm.
Female Reproductive
Function
Toxicological studies reported that Pb
exposure was associated with effects on
female reproductive function that can be
classified as alterations in female sexual
maturation, effects on fertility and menstrual
cycle, endocrine disruption, and changes in
morphology or histology of female
reproductive organs including the placenta.
Epidemiologic studies on Pb and female
reproductive function provided little
evidence for an association between Pb
biomarkers and effects on female
reproduction and fertility.
Epidemiologic studies of Pb levels and
hormones demonstrate associations but are
inconsistent overall and there is a lack of
large, well-conducted epidemiologic studies
examining associations between Pb levels
and fertility. Toxicological studies of Pb and
effects on female reproduction demonstrate
effects in some studies.
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Table 1-8 (Continued): Summary of evidence from epidemiologic, animal toxicological and
ecological studies on the effects associated with exposure to Pb.
Endpoint
Evidence in the 2006 Pb AQCD
Evidence in the 2013 Pb ISA
Cancer
Cancer
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 co-exposures (e.g., to As, 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.
Based on animal data and inadequate
human data Pb and Pb compounds would
be classified as likely carcinogens
according to the EPA Cancer Assessment
Guidelines for Carcinogen Risk
Assessment.
The toxicological literature continues to
provide the strongest evidence for Pb
exposure and cancer with supporting
evidence provided by the epidemiologic
literature. Epidemiologic studies of cancer
incidence and mortality reported inconsistent
results.
Ecological/Welfare Effects:
Endpoint
Evidence in the 2006 Pb AQCD
Evidence in the 2013 Pb ISA
Developmental and
Reproductive Effects
Terrestrial Organisms:
No information on reproduction in plants.
There is an insufficient number of studies
that consider Pb effects on plant
reproduction.
Limited evidence in invertebrates and
vertebrates.
Recent studies in a fewtaxa expand the
evidence for Pb effects on developmental
and reproductive endpoints for invertebrates
and vertebrates at concentrations that
generally exceed Pb levels in U.S. soils. In
some organisms, exposure-dependent
responses in development and reproductive
outcomes are observed in experiments
where exposure increases from background
concentrations to concentrations found in
heavily exposed sites near stationary
sources. Data on terrestrial species is
coherent with toxicological data from
mammals in the context of human health
research.
Aquatic Organisms:
No reviewed studies on reproductive effects
in aquatic plants.
Reproductive and developmental effects
reported in a few species of invertebrates at
<50 ug Pb/L and in fish at <150 ug Pb/L
Recent evidence supports previous findings
of reproductive and developmental effects of
Pb in freshwater invertebrates and
vertebrates and differential lifestage
response at near ambient concentrations of
Pb in some organisms. For saltwater
invertebrates there is limited evidence for
effects on early development at Pb
concentrations higher than typically detected
in marine environments.
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Table 1-8 (Continued): Summary of evidence from epidemiologic, animal toxicological and
ecological studies on the effects associated with exposure to Pb.
Endpoint
Evidence in the 2006 Pb AQCD
Evidence in the 2013 Pb ISA
Growth
Terrestrial Organisms:
Pb inhibits photosynthesis and respiration
in plants.
Limited evidence for growth effects in soil
invertebrates, avian and mammalian
consumers.
Recent studies support previous findings of
Pb effects on plant growth, with some
evidence for exposure-dependent decreases
in the biomass of some plant species grown
in Pb-amended or Pb-contaminated soil.
Recent data for soil invertebrates supports
previous evidence of increasing effects on
growth with increasing exposure.
Limited studies considered effects on growth
on vertebrates.
Aquatic Organisms:
Evidence for growth effects in algae,
aquatic plants and aquatic invertebrates
Most primary producers experience EC5o
values for growth in the range of 1,000 to
100,000 ug Pb/L
The weight of the evidence continues to
support growth effects of Pb in freshwater
plants and invertebrates. Recent studies on
growth in freshwater invertebrates find
effects of Pb at lower concentrations than
previously reported.
Growth inhibition in one species of
freshwater snail was observed at <4 ug Pb/L
in juveniles.
Lowest genus mean chronic value for Pb
reported at 10 ug Pb/L in a freshwater
mussel.
Survival
Terrestrial Organisms:
No information on mortality in plants.
Effects of Pb on invertebrates and
vertebrates include decreased survival.
In terrestrial and avian species toxicity was
observed in laboratory studies over a wide
range of doses (<1 to >1,000 mg Pb/kg
body weight per day) (U.S. EPA. 2005b).
Recent studies in invertebrates and
vertebrates support previous associations
between Pb exposure and mortality.
Aquatic Organisms:
No studies reviewed on mortality in plants
at current concentrations of Pb in the
environment.
Pb impacted survival of some aquatic
invertebrates at <20 ug Pb/L dependent
upon water quality variables (i.e., DOC,
hardness, pH).
Range of 96-hour LC50 values in fathead
minnow: 810->5,400 ug Pb/L
The weight of evidence continues to support
Pb effects on survival of freshwater
invertebrates and vertebrates and indicates
that there are effects in a few species at
lower concentrations than previously
reported.
Recent evidence for effects in a few
freshwater invertebrates at <20 ug Pb/L
Recent evidence in freshwater fish for
impacts to survival at <100 ug Pb/L
dependent upon water quality parameters
and lifestage.
96- hour LC50 values as low as 41 ug Pb/L in
fathead minnows tested in natural waters
from across the U.S.
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Table 1-8 (Continued): Summary of evidence from epidemiologic, animal toxicological and
ecological studies on the effects associated with exposure to Pb.
Endpoint
Evidence in the 2006 Pb AQCD
Evidence in the 2013 Pb ISA
Neurobehavioral
Effects
Terrestrial Organisms:
Exposure to Pb in laboratory studies and
simulated ecosystems may alter species
competitive behaviors, predator-prey
interactions, and contaminant avoidance
behaviors.
Recent studies continue to support previous
evidence that Pb exposure is associated with
behavioral alterations. Recent studies
identify possible molecular targets for Pb
neurotoxicity in invertebrates and there is
new evidence in a few invertebrate and
vertebrate species for behavioral effects
associated with Pb exposure (i.e., feeding
and escape behaviors).
Aquatic Organisms:
Exposure to Pb has been shown to affect
brain receptors in fish and may alter
avoidance behaviors and predator-prey
interactions.
Recent studies continue to support previous
evidence that Pb exposure is associated with
behavioral alterations. Recent studies
identify possible molecular targets for Pb
neurotoxicity in fish and provide additional
evidence for Pb effects on behaviors in
freshwater organisms that may impact
predator avoidance (swimming).
Hematological Effects Terrestrial Organisms:
Pb effects on heme synthesis were
documented in the 1986 Pb AQCD and
continue to be studied in terrestrial biota.
Changes in ALAD are not always related to
adverse effects but may simply indicate
exposure. The linkage between effects of
Pb on blood parameters is well
documented; however, the linkage between
hematological indicators and ecologically
relevant effects is less well understood.
Consistent with previous studies, the weight
of the evidence in recent studies continues
to support findings of Pb effects on heme
synthesis and ALAD enzyme activity. Recent
studies in birds near historical mining areas
and altered serum profiles and blood cell
counts in vertebrates provide evidence for
additional species in which hematological
endpoints are potentially affected by Pb.
Aquatic Organisms:
In metal impacted habitats, ALAD is a
recognized biomarker of Pb exposure.
Changes in ALAD are not always related to
adverse effects but may simply indicate
exposure. In fish, Pb effects on blood
chemistry have been documented with Pb
concentrations ranging from 100 to
10,000 ugPb/L
Consistent with previous studies, the weight
of the evidence in recent studies continues
to support findings of Pb effects on ALAD
and expands this evidence to additional
species of bacteria, invertebrates, and
vertebrates as well as in recent studies on
altered blood cell counts in vertebrates.
Additional field studies in both fresh and
saltwater bivalves report a correlation
between Pb and ALAD activity.
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Table 1-8 (Continued): Summary of evidence from epidemiologic, animal toxicological and
ecological studies on the effects associated with exposure to Pb.
Endpoint
Evidence in the 2006 Pb AQCD
Evidence in the 2013 Pb ISA
Physiological Stress Terrestrial Organisms:
Pb exposure may cause lipid peroxidation
and changes in glutathione concentrations.
There are species differences in resistance
to oxidative stress.
Recent studies continue to support previous
associations of Pb exposure with
physiological stress. New evidence includes
upregulation of antioxidant enzymes,
production of reactive oxygen species and
increased lipid peroxidation associated with
Pb exposure in additional species of
terrestrial plants, invertebrates and
vertebrates. Increasing effects follow
increasing experimental exposures from
background concentrations to concentrations
found in heavily exposed sites near
stationary sources.
Aquatic Organisms:
Pb exposure associated with alterations in
enzymes involved in physiological stress
responses.
Recent studies continue to support previous
associations of Pb exposure with
physiological stress. New evidence in
freshwater organisms includes upregulation
of antioxidant enzymes, production of
reactive oxygen species and increased lipid
peroxidation associated with Pb exposure.
Changes in antioxidant activity are reported
in some saltwater invertebrates. Observed
effects generally occurred at concentrations
that typically exceed Pb levels in U.S. waters
with limited evidence for effects associated
with Pb at polluted sites.
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Table 1-8 (Continued): Summary of evidence from epidemiologic, animal toxicological and
ecological studies on the effects associated with exposure to Pb.
Endpoint
Evidence in the 2006 Pb AQCD
Evidence in the 2013 Pb ISA
Community and
Ecosystem Level
Effects
Terrestrial Ecosystems:
Effects of Pb difficult to interpret because of
the presence of other stressors including
metals. The 1986 Pb AQCD reported shifts
toward Pb-tolerant communities at 500 to
1,000 mg Pb/kg soil.
In the 2006 Pb AQCD, decreased species
diversity and changes in community
composition were observed in ecosystems
surrounding former smelters.
Recent evidence for effects of Pb in soil
microbial communities adds to the body of
evidence for effects at higher levels of
biological organization. In addition, effects of
Pb uptake on reproduction, growth, and
survival at the species level are likely to lead
to effects at the population, community, and
ecosystem level. However, most evidence
for Pb toxicity to terrestrial biota is from
single-species assays, and there are
important uncertainties in generalizing from
effects observed under small-scale,
controlled conditions, up to effects at the
ecosystem level of biological organization.
Aquatic Ecosystems:
Most evidence of community and
ecosystem level effects is from near Pb
sources, usually mining effluents. Effects of
Pb difficult to interpret because of the
presence of other stressors including
metals.
Generally, 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).
Recent evidence for Pb effects on sediment-
associated and freshwater aquatic plant
communities add to the body of evidence of
effects at higher levels of biological
organization. However, most evidence for Pb
toxicity to aquatic biota is from single-
species assays. Uncertainties exist in
generalizing effects observed under small-
scale, predicted conditions up to effects at
the ecosystem-level however, uptake of Pb
into aquatic organisms and subsequent
effects on reproduction, growth, and survival
at the species level are likely to lead to
effects at the population, community, and
ecosystem level.
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CHAPTER 2 AMBIENT LEAD: SOURCE TO
CONCENTRATION
2.1 Introduction
This chapter reviews concepts and findings in atmospheric sciences that provide a
foundation for the detailed presentation of evidence of Pb exposure and Pb-related health
and ecological effects in subsequent chapters. Information in this chapter builds on
previous Pb AQCDs using more recent data and studies. This includes new knowledge of
Pb fate and transport, the latest developments in monitoring and analysis methodologies,
and recent data describing Pb concentrations as a function of size range. The chapter
focuses on Pb concentrations in the U.S. but includes non-U.S. studies to the extent that
they are informative regarding current conditions in the U.S. Description of the chemical
forms of Pb is not provided here, however, because this information is well established.
The reader is referred to the 2006 Pb AQCD for a description of the chemical forms of Pb
(U.S. EPA. 2006b).
Section 2.2 provides an overview of the sources of ambient air Pb. Section 2.3 provides a
description of the fate and transport of Pb in air, soil, and aqueous media. Descriptions of
Pb measurement methods, monitor siting requirements, and monitor locations are
presented in Section 2.4. Ambient Pb concentrations, their spatial and temporal
variability, size distributions of Pb-bearing particulate matter (PM), associations with
copollutants and background Pb concentrations are characterized in Section 2.5.
Concentrations of Pb in non-air media and biota are described in Section 2.6.
2.2 Sources of Atmospheric Pb
The following section reviews emissions estimates from the 2008 National Emissions
Inventory (NEI), version 3 (U.S. EPA. 2013). updated with EPA's best estimates of
piston-engine aircraft emissions1. This section also reviews updated information from the
peer-reviewed literature regarding sources of ambient Pb. Detailed information about
processes for anthropogenic emissions and naturally-occurring emissions can be found in
the 2006 Pb AQCD (U.S. EPA. 2006b). The papers cited herein generally utilized PM
sampling data, because a majority of ambient airborne Pb readily condenses to PM. The
mobile source category included combustion products from organic Pb antiknock
1 The piston-engine aircraft emissions inventory can be obtained from the following site:
http://www.epa.gov/ttn/chief/net/2008neiv2/2008 neiv2 tsd draft.pdf.
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additives used in piston-engine aircraft (hereafter referred to piston-engine aircraft
emissions).
2.2.1 National Emissions Inventory
The 2006 Pb AQCD (U.S. EPA. 2006b) listed the largest sources to be (in order):
industrial-commercial-institutional boilers and process heaters (17%), coal utilities
boilers (12%), mobile sources (10%), iron and steel foundries (8%), and miscellaneous
sources from industrial processes, incineration, and utilities, each contributing less than
5% (53%). The sources listed in the 2006 Pb AQCD were based on the 2002 NEI (U.S.
EPA. 2006a). Subsequent correction of computational errors prior to completion of the
2008 NAAQS review provided corrected estimates for the 2002 inventory which
indicated the largest sources to be (in order): mobile sources from leaded aviation gas
usage in piston-engine aircraft (45%), metallurgical industries (23%), manufacturing
(14%), incineration (8%), boilers (6%), and miscellaneous sources contributing less than
5% (U.S. EPA. 2007h). The 2002 and prior year inventories discussed in this document
reflect the corrected information.
Emissions of Pb have dropped substantially over the past forty years, as shown in Figure
2-1 and Figure 2-2. The reduction before 1990 is largely due to the phase-out of Pb as an
anti-knock agent in gasoline for on-road automobiles, as discussed in the 2006 Pb AQCD
(U.S. EPA. 2006b). This action resulted in a 98% reduction in Pb emissions from
1970-1995. Total Pb emissions over the period 1995-2008 decreased an additional 77%,
from 4,100 tons in 1995 to 950 tons in 2008. Additional emissions reductions are related
to enhanced control of the metals processing industry. In 1995, metals processing
accounted for 42% (2,200 tons) of total Pb emissions. By 2008, metals processing
accounted for 19% (178 tons) of total emissions. This represented more than an order of
magnitude decrease in Pb emissions from metals processing. Emissions from piston-
engine aircraft decreased 34% over this time period. In 1990, nonroad Pb emissions were
990 tons, 830 tons of which were generated from piston-engine aircraft, and represented
19% of total Pb emissions. In 2008, nonroad Pb emissions from piston-engine aircraft
were slightly lower at 550 tons,1 which represented 58% of all Pb emissions. 2008 piston-
engine aircraft emissions were comprised of 254 tons of Pb from emissions at or near
airports and 296 tons of Pb emitted in flight (i.e., outside the landing and take-off cycles).
"Miscellaneous" emissions from other industrial processes, solvent utilization,
agriculture, and construction constituted 11% of emissions (103 tons, 80 of which were
1 This reflects EPA's best estimates of piston-engine aircraft emissions. The piston-engine aircraft emissions
inventory can be obtained from the following site: http://www.epa.gov/ttn/chief/net/2008neiv2/2008_neiv2_tsd_draft.pdf.
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from industrial sources other than metal working and mining) in 2008 (U.S. EPA. 201 la.
2008a).
250
200
150
E 100
.0
Q.
50
Highway Vehicles
I Metal Workingand Mining
Fuel Combustion
I Piston Engine Aircraft
I Miscellaneous
1970
1975
1980
1985
1990
1995
1999
2002
2005
2008
Source: U.S. EPA (2011 a. 2008a)
Figure 2-1 Trends in Pb emissions (thousand tons) from stationary and
mobile sources in the U.S., 1970-2008.
20
-3
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6 -i
Highway Vehicles
Metal Workingand Mining
Fuel Combustion
Piston Engine Aircraft
Miscellaneous
1990
1995
1999
2002
2005
2008
Source: U.S. EPA (2011 a. 2008a)
Figure 2-2 Trends in Pb emissions (thousand tons) from stationary and
mobile sources in the U.S., 1990-2008.
Direct emissions of Pb into the atmosphere primarily come from piston-engine aircraft,
fuel combustion, and industrial activities. Direct Pb emissions estimated by the 2008 NEI
are shown in Figure 2-3. Piston-engine aircraft produced 58% of all emissions (550 tons).
Metal working and mining contributed 178 tons (19%) of Pb emissions in 2008, followed
by fuel combustion (13%), other industry (8%), and miscellaneous contributions from
agriculture, solvent utilization, and operation of commercial marine vessels and
locomotives (2%) (U.S. EPA. 201 la). Pb emissions from the "metal working and
mining" category include the single primary Pb smelter in the U.S., the Doe Run facility
in Herculaneum, MO; secondary Pb smelters, mostly designed to reclaim Pb for use in
Pb-acid batteries: and smelters for other metals.
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550
100
200 300 400
2008 Emissions (tons)
500
600
Source: U.S. EPA (2011 a)
Figure 2-3
Nationwide stationary and mobile source Pb emissions (tons) in
the U.S. by source sector in 2008.
There is substantial variability in Pb emissions across U.S. counties, as shown in Figure
2-4 for the continental U.S. The emissions levels, shown in units of tons, vary over
several orders of magnitude. Ninety-five percent of U.S. counties, territories, and tribal
areas had 2008 emissions below 1 ton; 50% of counties, territories, and tribal areas had
2008 emissions below 0.042 tons. Jefferson County, MO was the highest emitting single
county, with over 20 tons of airborne Pb emissions in 2008. Jefferson County is home to
the Doe Run primary Pb smelting facility, which is the only remaining operational
primary Pb smelter in the U.S. and is planning to cease the existing smelter operations at
this site by April, 2014 (DRRC. 2010). Pb emissions from piston-engine aircraft
operating on leaded fuel are estimated to occur at approximately 20,000 airports across
the U.S. Many of the more active airports are more numerous in highly populated
metropolitan regions, which suggests that emissions from piston-engine aircraft may be
higher in these locations compared with rural areas. In twenty-five counties, piston-
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engine aircraft are estimated to emit cumulatively greater than one ton of Pb in 2008 U.S.
EPA (201 la).
2008 NEI Pb Emissions by County (tons)
^B 0-0.13
^H 0.13 -0.40
| 0.40 - 0.82
| 0.82-1.54
| 1.54-2.76
| 2.76-5.49
| 5.49-10.13
I 10.13-20.10
Source: U.S. EPA (2011 a)
Figure 2-4 County-level Pb emissions (tons) in the U.S. in 2008.
Figure 2-5 illustrates the locations and relative magnitude of Pb emissions for 124
facilities in the U.S. emitting 0.5 tons or more in 2008 (U.S. EPA. 201 la). One facility,
Doe Run in Herculaneum (Jefferson Co.), MO, emitted more than 10 tons in 2008.
Additionally, the map illustrates several locations where there is a confluence of point
sources (not to be confused with total sources including non-point), including Jefferson
Co., MO, Lake Co., IN, Iron Co., MO, and Gila, AZ, in each of which are facilities that
were estimated to cumulatively emit between 5 and 10 tons. Among the facilities shown,
124 are non-airport facilities; eight of these were estimated to emit 5 tons or more, 53 to
emit between 1 and 5 tons, and 63 to emit between 0.5 and 1 tons in 2008. Figure 2-5
2-6
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additionally includes 58 airports, the six largest of which were estimated to emit between
1 and 5 tons.
Airports with estimates <1.0 and =>0.50 tpy**
Airports with estimates <5.0 and =>1.Qtpy*
Study Airports
Facilities with estimates <1.0 and =>0.50 tpy"
0 Facilities with estimates <5.0 and => 1.0 tpy*
A Facilities with estimates => 5.0 tpy
Figure 2-5 Pb facilities estimated to emit 0.5 tons or more in 2008.
2.2.2 Anthropogenic Sources
Anthropogenic Pb source categories are organized below in order of magnitude with
regard to the sum of emissions nationally reported on the 2008 NEI (U.S. EPA. 201 la).
Pb sources were reviewed in the 2006 Pb AQCD (U.S. EPA. 2006b) by species. Forms of
Pb commonly observed in the environment are carried forward from the 2006 Pb AQCD
(U.S. EPA. 2006b) and are presented in Table 2-1 to serve as a reference for the
categories of Pb sources described in Sections 2.2.1 and 2.2.2.
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Table 2-1 Pb compounds observed in the environment.
Emission Source
Observed Pb Compounds
Minerals
PbS (Galena)
PbO (Litharge, Massicot)
Pb3O4 ("Red Pb")
PbSO4 (Anglesite)
Smelting aerosols
Pb°, PbS
PbSO4, PbO
PbCO3
Pb silicates
Coal combustion aerosols
PbS
PbSe
Coal combustion flue gases
Pb, PbO, PbO2 (Above 1,150 K)
PbCI2 (Low rank coals, above 1,150 K)
PbSC-4 (Below 1,150 K)
Wood combustion
PbCO3
Waste incineration aerosols
PbCI2, PbO
Soils near mining operations
PbCO3
PbSO4
[PbFe6(S04)4(OH)12]
[Pb5(P04)3CI]
[Pb4S04(C03)2(OH)3]
PbS-Bi2S3
Pb oxides, silicates
Piston-engine aircraft emissions, racing vehicle
exhaust (combustion of leaded fuel)
PbBr2
Alkyl Pb
PbBrCI-NH4CI, PbBrCI-2NH4CI
Roadside dust
PbSO4, Pb°, PbSO4(NH4)SO4, Pb3O4, PbO-PbSO4
and 2PbCO3-Pb(OH)2
Brake wear, wheel weights
Pbu
Aircraft engine wear
Pbu
Source: Biggins and Harrison (1980. 1979): U.S. EPA (2006b)
-------
Pb emissions in the U.S. derive from a combination of mined, processed, and imported
Pb. Figure 2-6 illustrates trends in the origin of Pb used in various sectors described
below. Over the period 1991-2010, the amount of Pb used in secondary Pb processing
increased by 37%. Exports of Pb increased by 103%, with 2010 exports sent to Mexico as
refined Pb; to Canada, China, and Japan in spent Pb-acid batteries; and, to the Republic
of Korea as Pb in concentrate (USGS. 2012). Primary Pb processing decreased by 67%,
while Pb mining and imports fluctuated over 1991-2010 without a clear increasing or
decreasing trend. In 2008, 1.28 million tons of Pb were introduced to the market by
primary and secondary processing combined. Nine hundred sixty-four (964) tons of Pb
were emitted to the ambient air. Hence, 99.9% of Pb produced in 2008 remained in
products or was emitted directly to soil or water following disposal.
I Mine
I Primary Production
Secondary Production
I Imports
I Exports
Note: Exports are shown by negative numbers to illustrate that the Pb was leaving the U.S. Data were aggregated into five-year
totals to stabilize the data shown.
Source: U.S. Geological Survey (2012. 2006. 2001. 1996)
Figure 2-6 Five-year totals for Pb mining, primary and secondary production,
imports, and exports, 1991-2010.
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2.2.2.1 Pb Emissions from Piston-engine Aircraft Operating on
Leaded-Aviation Gasoline and Other Non-road Sources
The largest source of Pb in the NEI, in terms of total emissions nationally, is emissions
from piston-engine aircraft operating on leaded aviation gasoline (U.S. EPA. 201 la). As
outlined in Table 2-1, there are several forms of Pb emitted from engines operating on
leaded fuel. Dynamometer testing has indicated that Pb emissions from piston engines
operating on leaded fuel can occur in the particulate and gaseous forms. For example,
Gidney et al. (2010) performed dynamometer testing on automobiles operating on
standard gasoline and on gasoline with low levels of organometallic additives. Tetraethyl
Pb was included since it is still used in piston-engine aircraft fuel. Gidney et al. (2010)
point out that, where tetraethyl Pb is used as an additive in piston-engine aircraft fuel, the
fuel also contains ethylene dibromide, which reacts with Pb to form Pb bromide and Pb
oxybromides. Pb bromides and Pb oxybromides are more volatile than elemental Pb at
combustion temperatures and are therefore exhausted from the engine. After being
exhausted, the brominated Pb compounds cool to ambient temperatures and condense to
form sub-micron solid particles. In contrast, emissions of organic Pb would remain
largely in the vapor phase at ambient temperatures. Studies of Pb emissions within
enclosed microenvironments where automobiles were the dominant Pb source cited
within the 1986 Pb AQCD (U.S. EPA. 1986a). reported that organic Pb vapors
contributed less than 20% of total vehicular Pb emissions. A more recent study supports
this (Shotyk et al.. 2002). The 20% estimate of organic Pb emissions from the previous
studies of on-road Pb emissions may potentially provide an upper bound for organic Pb
emissions from current piston-engine aircraft.
Pb emission rates from piston aircraft vary with fuel consumption rates, which depend on
the engine/airframe combination and the mode of operation of the aircraft. The ASTM
specification for the maximum Pb content in "100 Low Lead," the most commonly used
leaded piston-engine aircraft fuel, is 2.12 g of elemental Pb/gallon (ASTM. 2007). Fuel
consumption rates can be obtained for some engine/aircraft combinations by running
FAA's Emissions and Dispersion Modeling System (FAA. 2011). Fuel consumption for
piston-engine aircraft operating at one airport in the U.S. was estimated to range from
1.6 grams per second (g/second) of fuel during taxi-out, to 15.3 g/second of fuel during
run-up preflight check for single-engine aircraft; and 5.1 g/second during taxi and
50 g/second during preflight run-up check for twin-engine aircraft (Carr et al.. 2011).
Fuel consumption rates for aircraft listed in FAA's Emissions and Dispersion Modeling
System were used to develop the Pb emissions inventory for piston aircraft that are
discussed in Section 2.2.1. EPA estimates that on average, 7.34 g of Pb is emitted during
a landing and take-off cycle conducted by piston-engine aircraft (ERG. 2011).
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2.2.2.2 Emissions from Metals Processing and Mining
High Pb emissions were observed in the 2008 NEI (U.S. EPA. 201 la) in Herculaneum,
MO, where the Doe Run Pb smelter is operated. Although it is set to cease smelting
operations in 2014 (DRRC. 2010). it is of interest to consider studies of primary smelter
emissions in the context of the data analyzed in this ISA. Batonneau et al. (2004) and
Sobanska et al. (1999) found that the Pb content in PM emitted from a primary Pb
smelter was 56.6% by weight, and the Pb content in PM from a Pb/Zn smelter was 19.0%
by weight. Choel et al. (2006) confirmed that Pb was strongly associated with sulfur in
Pb-Zn smelter emission PM, and that Pb sulfates and Pb oxy-sulfates were the most
abundant species, with important contributions from Pb oxides. Pb concentrations
1,800 meters downwind of the smelter (0.625-0.880 (ig/m3) were roughly thirty-five
times higher than a monitor 1,800 meters upwind (0.017-0.026 (ig/m3).
Fugitive emissions (i.e., unaccounted ambient air Pb emissions) from secondary Pb
processing (e.g., Pb recovery from batteries) can be substantial over the course of a year,
but they are difficult to estimate. Thurston et al. (2011) performed source apportionment
of PM25 found that Pb-PM2 5 concentrations from the Chemical Speciation Network
(CSN) were associated with the metals industry along with Zn-PM25. Goyal et al. (2005)
estimated fugitive emissions using concentration data obtained from samplers sited in
close vicinity of secondary Pb processing facilities and meteorological data from nearby
weather monitoring stations. Regression modeling and Bayesian hierarchical modeling
were both used to estimate fugitive and stack emissions from secondary Pb processing
facilities in Florida, Texas, and New York. Depending on the model used, median
fugitive emissions were estimated to be 1.0 * 10"6 to 4.4 x 10"5 g Pb/m2 per second at the
Florida site, 9.4 x 10"7 to 2.0 x 10"6 g/m2 per second for the Texas site, and 8.8 x 10"7 to
1.1 x 10"6 g/m2 per second at the New York site. Median stack emissions estimates varied
widely among the models, with the Florida site median ranging from 1.4 x 10"6 to
1.4 x 10"1 g Pb/second, the Texas site median ranging from 8.4 x 10"2 to
8.6 x 10"2 g/second, and the New York site ranging from 8.4 x 10"3 to 1.0 x 10"2 g/second.
Additionally, the Bayesian hierarchical model was used to estimate fugitive Pb emissions
from secondary Pb processing facilities nationwide using concentration data as prior
information. Nationwide median fugitive emissions from secondary Pb processing
facilities were estimated to be 9.4 x 10"7 to 3.3 x 10"6 g/m2 per second. Recently,
speciation of emissions from a battery recycling facility indicated that PbS was most
abundant, followed by Pb sulfates (PbSO4 and PbSO4-PbO), PbO and Pb° (Uzu et al..
2009).
In addition to secondary Pb smelting, Pb emissions occur from processing of other
metals. For example, a recent study examined Pb emissions from a sintering plant, a
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major component of the steel making process in southern France (Sammut et al.. 2010).
Cerussite, a Pb carbonate (PbCO3-2H2O), was observed to be the most abundant species
and contributed 20 g Pb/kg measured PM. In another example, Reinard et al. (2007) used
a real-time single particle mass spectrometer to characterize the composition of PMi
collected in Wilmington, Delaware in 2005 and 2006. Strong Pb-Zn-K-Na associations
were observed within 13% of PM samples. Comparison with stack emissions revealed
that a nearby steel manufacturing facility was an important source of Pb. Ambient PM
classes containing only a subset of such elements, e.g., Zn only, Pb-K only were
non-specific and so could not be mapped to individual sources. Ogulei et al. (2006)
observed that 6% of Pb in PM2 5, along with some O3, Cu, and Fe, was attributed to steel
processing in Baltimore, MD. Murphy et al. (2007) conducted a detailed study of the
distribution of Pb in single atmospheric particles during the fifth Cloud and Aerosol
Characterization Experiment in the Free Troposphere campaign at the Jungfraujoch High
Altitude Research Station in Switzerland and found that the predominant type of urban
Pb-bearing aerosols contained Pb together with K and Zn. The mode of the size
distribution for this type was around 200 nm.
Waste from current or defunct mines has been shown to present an additional fugitive
source of Pb. For example, distribution of Pb along a haul road connecting an active mine
to a port has been documented in Alaska (see Section 2.6.6). Additionally, Zheng and Li
(2009) applied source apportionment in three northeastern Oklahoma towns to identify
the influence of "chat," or waste piles from formerly operational Pb-Zn mines, on
PMio-2.5 and PM25. They estimated that mine waste was responsible for 88% of Pb in
PM 10-2.5 samples and 40% of Pb in PM25 samples.
2.2.2.3 Fossil Fuel Combustion
Murphy et al. (2007) found that the volatility of Pb and its compounds such as PbO may
result in its presence at high concentration in the submicron fraction of PM emitted from
coal emissions. PbSO4, also derived from coal combustion, has low water solubility
(Barrett et al.. 2010). PbSO4 was estimated to comprise 37% of Pb in PMi0 from a
2002-2003 study of PM in Shanghai, China (Tan et al.. 2006) and 0.6% of total PM10
from a study of PM in Taiyuan City, China (Xie et al.. 2009). Murphy et al. (2007)
presented an estimated U.S. mass budget for Pb emitted from consumption of select fuels
and crude oil. Fuel consumption estimates for 2005 were employed (Freme. 2004). Based
on an annual consumption of 1.0 x 109 metric tons coal with an average Pb concentration
of 20 mg/kg (range: 5 to 35 mg/kg) and using an emission factor (airborne fraction) of
approximately 0.01, coal contributed approximately 200 metric tons Pb/year to the
atmosphere. At the time of the Murphy et al. (2007) study, there were no emission factors
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available to estimate airborne Pb emissions for crude oil or residual oil, but these
represent potentially large sources (with total Pb in these sources estimated by Murphy et
al. (2007) to be as much as 100-500 metric tons/year and 25-700 metric tons/year,
respectively). These calculations imply that there is substantial uncertainty in estimates of
Pb emissions resulting from fuel combustion. It is important to note that Murphy et al.
(2007) state that the crude oil estimates are based on a limited number of samples and
that there was uncertainty in the estimates of Pb content in residual oil. Furthermore,
Murphy et al. (2007) was based on data ranging back in time from 1972 to 2005.
Therefore, the Murphy et al. (2007) findings do not necessarily conflict with reported
ambient air Pb emissions from the NEI. As part of recent rulemaking, EPA has developed
a draft Pb emission factor of 1.3xlO"5 Ib/mmBtu for boilers larger than 25 MW that use
#2 or #6 fuel oil (U.S. EPA. 201 Ib).The amounts of Pb emitted from these U.S. sources,
however, are several orders magnitude smaller than those estimated to arise from coal
combustion in China.
Coal combustion is considered to be a major source of Pb in the atmosphere now that
leaded gasoline has been phased out for use in on-road vehicles (Diaz-Somoano et al..
2009). Global Pb estimates are considered here to inform understanding of U.S. Pb
emissions from coal combustion. McConnell and Edwards (2008) examined correlations
of Pb with BC, Cd, Ce, sea salt Na, non-sea salt S, and Tl in a Greenland ice core and
observed high correlations for BC, Cd, non-sea salt S and Tl during the period
1860-1940, when coal combustion was the predominant energy source. With the
exceptions of non-sea salt S and Tl, the high correlations were not maintained into the
years 1940-2003, when oil combustion was the most prevalent energy source. This
suggests common industrial sources of PbS or PbSO4. Rauch and Pacyna (2009)
constructed global metal cycles using anthropogenic data from 2000. They confirmed that
the largest anthropogenic airborne Pb emissions arise from fossil fuel combustion, and
they quantified Pb emissions at 85,000 tons/year worldwide. Globally, Pb emissions from
stationary sources have been increasing and the north-south gradient in aerosol Pb
concentrations over the Atlantic Ocean has disappeared as a result of industrialization of
the southern hemisphere (Witt et al.. 2006; Pacyna and Pacyna. 2001). The Pb isotope
ratio values (mainly 206Pb/207Pb) for coal from around the world have been compared with
those for atmospheric aerosols. In most parts of the world, there has been a difference
between the signature for aerosols and that for coal, where the atmospheric 206Pb/207Pb
ratio values are lower, indicative of additional contributions from other sources. Zhang et
al. (2009a) used single particle aerosol mass spectrometry (ATOFMS) to find that PM
containing Pb along with OC and/or EC was attributed to coal combustion processes in
Shanghai, China; this accounted for roughly 45% of Pb-bearing PM.
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Seasonal effects of the contributions of Pb emissions from coal combustion have been
observed. For example, in Tianjin, northern China, the winter heating period starts in
November, and the contribution from coal combustion to the Pb aerosol becomes high
during the winter. This leads to both a high Pb content and a high 206Pb/207Pb ratio. Coal
consumption and Pb-bearing PM concentrations declined during the summer months, and
Pb from other sources, mainly vehicle exhaust emissions, became relatively more
pronounced (Wang et al., 2006c). This seasonal relationship contrasts with observations
for the U.S. when power stations are more active in summer months (EIA. 2012). The
increased energy use in summer periods in the U.S. may be attributable to increased
requirements for air-conditioning.
2.2.2.4 Waste Incineration
Waste incineration studies suggest that the Pb content varies by industrial or municipal
waste stream. For example, Ogulei et al. (2006) performed positive matrix factorization
of PM25 and gaseous copollutants for Baltimore, MD and observed that 63% of Pb in
PM2 5 was attributed to waste incineration during the six day study duration. Other
prevalent compounds associated with incineration included NO3", EC, Cd, Cu, Fe, Mn,
Se, Zn, O3, and NO2 (note that Cl was not observed in this study). Likewise, Song et al.
(2001) used PMF to deduce sources of PM25 measured at Washington, B.C., Brigantine,
NJ, and Underbill, VT during the years 1988-1999. They observed a waste incineration
source loaded with OC, EC, Pb, and Zn at all three sites. A study by Moffet et al. (2008b)
found that Pb-Zn-Cl-containing particles in PM2 5 samples collected from an industrial
area in Mexico City represented as much as 73% of fine PM. These were mainly in the
submicron size range and were typically mixed with elemental carbon (EC), suggesting a
combustion source. Zhang et al. (2009a) also observed high correlation between Pb and
Cl associated with waste incineration in Shanghai, China. Several Pb isotope studies have
also been used to distinguish contributions to incineration from industrial sources. Isotope
analysis is discussed in more detail in Section 2.4.1.5. Novak et al. (2008) evaluated
changes in the amounts and sources of Pb emissions in the U.K. and Czech Republic
during the 19th and 20th centuries and found uncertainty in the amount and the isotope
composition of Pb emanating from incineration plants. The isotopic signature of Pb
recycled into the atmosphere by incineration of various industrial wastes could have
shifted from relatively high 206Pb/207Pb ratios consistent with local Variscan ores to lower
values reflecting imported Precambrian ores. However, other environmental studies
concerning incineration have given highly consistent values for the Pb isotope ratio for
European incineration sources. For example, Cloquet et al. (2006) showed that the Pb
isotopic composition of urban waste incineration flue gases in northeastern France was
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-1.16. De la Cruz et al. (2009) reported that waste incineration was an important source
of Pb and showed that the 206Pb/207Pb and 208pb/207 Pb ratios for waste incineration Pb
emitted in European countries were 1.14-1.16 and 2.43 respectively (de la Cruz et al..
2009).
2.2.2.5 Wood Burning and Wildfires
Another potentially uncontrollable source is Pb deposited historically in forests and
remobilized during forest fires. Section 3.1.3.1 describes residential Pb-PM
concentrations related to in-home burning of wood contaminated with Pb of ambient
origin, while this section describes ambient air measurements of Pb attributed to wood
burning. The 2006 Pb AQCD (U.S. EPA. 2006b) presented data by Nriagu (1989)
estimating that 1,900 metric tons of Pb were emitted globally each year from wildfires.
Wildfire Pb emissions were not included in the NEI. Murphy et al. (2007) observed that a
fraction of particles contained small quantities of Pb on biomass particles measured using
ATOFMS to sample directly from forest fire plumes in northwest Canada and eastern
Alaska in July, 2004; these particles also typically contained SO4"2. Several studies
illustrate moderate-to-long range transport of biomass burning plumes containing Pb.
Using positive matrix factorization, Ogulei et al. (2006) estimated that 20% of Pb in
PM2 5 measured in Baltimore, MD was attributed to a July, 2002 episode of wildfires in
Quebec, Canada in his 6-day study. Other components strongly associated with the
Quebec wildfires included NO3", OC, EC, Cd, Mn, Zn, O3, and CO. Qureshi et al. (2006)
also observed a spike up to 42 ng/m3 in Pb-PM2 5 concentration in Queens, NY
coinciding with the Quebec wildfires; for comparison, the authors provide the 3-month
average from July to September of 5.1 ng/m3 for Pb-PM2 5 in Queens. Similarly, Anttilla
et al. (2008) measured PMi0 in Virolahti, Finland during a wildfire in Russia and
observed average Pb-PMi0 concentrations during the forest fire episodes to be 1.7-3.0
times higher than the reference concentration of 3.5 ng/m3. Hsu et al. (2009c) observed
Pb concentrations in Taiwan attributed to biomass burning in Northeastern China; Pb was
highly correlated with K attributed to biomass burning during these episodes. Odigie and
Flegal (2011) studied remobilization of Pb during the 2009 wildfires in Santa Barbara,
CA. Pb concentrations in ash samples obtained after the wildfire ranged from 4.3 to
51 mg/kg. Isotopic analysis of the ash suggested that the remobilized Pb was initially
emitted by a mix of contemporary and previous industrial sources and historic
combustion of leaded gasoline. Grouped with "miscellaneous" Pb emissions, fires from
agricultural field burning and prescribed fires accounted for 2.4 tons of U.S. Pb emissions
in 2008 (U.S. EPA. 201 la). Polissar et al. (1998) used positive matrix factorization to
apportion PM2 5 and found small Pb signals attributed to the forest fire factor at two of
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six Alaskan sites where a forest fire factor was detected; the forest fire factor was
dominated by a combination of BC, ft", and K.
Wood combustion emissions from commercial boilers present an additional potential Pb
source. For example, Pb concentrations in 132 samples of wood pellets and 23 samples of
wood chips from the northeastern US were presented by Chandrasekaran et al. (2012).
Some of the commercially available pellets contained waste wood that was thought to
include Pb-painted material. PM2 5 emissions from the boilers were enriched with Pb as
well as Cd, Ti, Rb, and Zn (Chandrasekaran et al.. 2011). The concentration of Pb in
PM2.5 was 96% and 87% of the Pb concentration in pre-combustion wood pellet fuels of
0.014 kg/kg and 0.027 kg/kg, respectively.
Several studies have explored the chemical properties of biomass emissions. Obernberger
et al. (2006) simulated biomass combustion in a laboratory setting to assess emissions.
They reported pre-combustion mean Pb content in wood, bark, and logging residues to
range from 2-5 mg/kg dry basis. They reported volatilization and subsequent
condensation of Pb emissions from combustion. Van Lith et al. (2008; 2006) studied the
inorganic element content of wood chips and particle board and the release of inorganic
elements during combustion of those materials in laboratory experiments. They measured
a Pb content of 16 mg/kg dry basis in particle board and of 0.44 mg/kg dry basis in
spruce wood chips. Using three different types of combustion for different materials, they
found that up to 10% of Pb was released at a combustion temperature of 500 °C and up to
85% was released at a temperature of 850 °C. At temperatures greater than 650 °C, PbO
gas was released under oxidizing conditions; under reducing conditions, Pb gas, PbCl
gas, and PbS gases were released at temperatures above 500 °C. Jimenez et al. (2008)
performed laboratory experiments of olive tree combustion and concluded that Pb
vaporizes upon combustion and then condenses between 900 °C and 560 °C. Jimenez et
al. (2008) also observed that Pb concentration in PM changes with oxygen content and
temperature, with concentrations converging toward 2,000 mg/kg for increasing percent
available oxygen and increasing temperature.
Pb deposition on trees has been documented in Acadia National Park in Maine with mean
foliar concentrations ranging from <0.5 to 3.1 mg/kg (Wiersma et al.. 2007). Tree ring
core samples obtained in the Czech Republic illustrate that the amount of Pb deposited on
trees from coal and leaded gasoline combustion sources tended to increase over the depth
of the core, with maximum concentrations corresponding to time periods of 1969-1972,
1957-1960, and 1963-1966 in three samples (Zunaet al.. 2011).
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2.2.2.6 Roadway-Related Sources
Contemporary Emissions from Vehicle Parts
Contemporary Pb emissions from motor vehicles may occur because several vehicle parts
still contain Pb. Wheel weights, used to balance tires, are clipped to the rims of tire
wheels in order to balance the tires, and may become loose and fall off. Pb wheel weights
have been banned in several states including Washington, Maine, and Vermont with
legislation considered in Iowa, California, and Maryland. However, Pb wheel weights are
a source in most states for the period of time covered in this assessment. Ambient air Pb
concentrations near heavily trafficked areas may be related to use of Pb-based wheel
weights that are prone to dislodgement. Root (2000) and Aucott and Caldarelli (2012)
estimated that 7.5 kg/km per year are deposited and that, among deposited weights,
2.7-5% of the mass is lost from the roadway daily. Aucott and Caldarelli (2012)
extrapolated their results for Mercer County, NJ to the U.S. to estimate that 480 tons of
Pb are deposited to roadways each year. On pavement they may be ground into dust by
the pounding forces of traffic (Root. 2000). For example, Aucott and Caldarelli (2012)
estimated that 13.8 ± 5.0% of the deposited mass of wheel weights are dispersed each
year through abrasion and grinding by traffic. Schauer et al. (2006) measured Pb
emissions in two traffic tunnels and found that the Pb-PM2 5 concentration did not exceed
17% of the Pb-PMio concentration in any of the runs. Schauer et al. (2006) suggested that
enrichment in the coarse fraction may have been related to wheel weights. Additionally,
Schauer et al. (2006) measured PMi0 and PM25 composition from brake dust and found
concentrations that were low but statistically significantly greater than zero for Pb in
PM10 (0.02 ± 0.01 mg/g) and Pb in PM25 (0.01 ± 0.00 mg/g) for semi-metallic brake
pads and for Pb in PM10 (0.01 ± 0.00 mg/g) for low-metallic brake pads. Song and Gao
(2011) speciated coarse and fine PM samples obtained next to the New Jersey Turnpike
in winter and summer of 2007-2008. Using principal component analysis, they found that
Pb was prevalent in the factor including automobile exhaust and brake wear. Pb was
observed to have a similar size distribution as Zn in the winter and Zn and Cd in the
summer, with higher concentrations in the fine fraction at a mode of 0.18-0.32 (im.
Fauser (1999) observed that 92% of particles generated by tire abrasion have
aerodynamic diameter smaller than 1 nm. Additionally, Hjortenkrans et al. (2007) used
material metal concentrations, traffic volume, emissions factors, and sales data to
estimate the quantity of Pb emitted from brake wear and tires in Stockholm, Sweden in
2005. They observed that 24 kg (0.026 ton) of Pb were emitted from brake wear each
year, compared with 2.6 kg (0.0029 ton) of Pb from tire tread wear; an estimated 549 kg
(0.61 ton) was estimated to have been emitted from brake wear in 1998. McKenzie et al.
(2009) determined the composition of various vehicle components including tires and
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brakes and found that tires were a possible source of Pb in stormwater, but no
identification of Pb-containing PM in stormwater was carried out. However, PM from tire
abrasion is usually found in coarser size ranges (Chon et al., 2010). while those in the
submicron range are more typically associated with combustion and incineration sources.
Road Paint
Some paints, frequently used for road markings and near-road structures, have contained
large quantities of Pb. Ozaki et al. (2004) analyzed paints used in urban and forested
areas of Japan to delineate traffic lanes on roads and measured high concentrations of Pb
(12 mg/g) and Cr (820 mg/g) in yellow paint, suggestive of PbCrO4. High concentrations
of Pb (5.3 mg/g) and Cr (6.7 mg/g) were measured in red paint, also suggesting presence
of PbCrO4. Gray paint and anticorrosive used to coat guardrails was also measured to
have high Pb concentrations of 451 mg/g.
Street dust particles containing Pb, Cr, or both have been associated with PbCrO4 in
street paints. Adachi and Tainosho (2004) used scanning electron microscopy to analyze
road dust composition and found particles with high Pb and Cr content, indicative of
PbCrO4 in yellow paint. Murakami et al. (2007) estimated enrichment of road dust by
paint particles. They measured 46% of the road dust particles to contain more than 0.2%
Pb, Cr, or both. Pb and Cr loading (mass deposited per area of surface) was also observed
for all road dust size fractions measured (ranging from < 43 pirn to > 841 pirn) (Lau and
Stenstrom. 2005).
Unleaded Fuel
Unleaded fuel contains Pb as an impurity within crude oil (Pacyna et al.. 2007). Schauer
et al. (2006) measured Pb in PM2 5 from tailpipe emissions and observed quantities in
on-road gasoline emissions that were statistically significantly different from zero
(83.5 ± 12.80 mg/kg), whereas emissions of Pb from diesel engines were not statistically
significantly different from zero. Hu et al. (2009a) investigated the heavy metal content
of diesel fuel and lubricating oil. They found
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factorization to decompose PM2 5 samples obtained from an Underbill, VT site [Song et
al. (2001) used data from 1988-1999 and Polissar et al. (2001) used data from 1988-1995]
and observed a "Canadian Mn" factor loaded with Mn and Pb. Methylcyclopentadienyl
manganese tricarbonyl had been used in Canada as an antiknock agent to replace
tetraethyl and tetramethyl Pb additives, but it is not clear if the co-occurrence of Mn with
Pb denotes that the phase-out had not been completed during part of the study period
(Canadian phase-out completed in 1993).
2.2.2.7 Deposited Pb
Soil Pb can serve as a reservoir for deposited Pb. The following subsections describe
studies of previously deposited Pb that originated from industrial activities, historical use
of leaded on-road gasoline, and urban sources such as paint and building materials. The
2006 Pb AQCD (U.S. EPA. 2006b) cited an estimate by Harris and Davidson (2005) that
more than 90% of airborne Pb emissions in the South Coast Basin of California were
from soil resuspension. This value was obtained by constructing mass balances rather
than from direct measurements of Pb along roads, and hence it is an estimate. Currently,
measured data are not available with sufficient spatial resolution to discern the specific
contribution of soil Pb resuspension to air Pb concentration, but resuspended soil Pb
cannot be eliminated as a potential source of airborne Pb. Section 2.5.1.2 includes recent
information on ambient air concentrations of Pb-TSP, sampled at 6 meters above ground
level (AGL), at a distance of 500 meters from the heavily trafficked Interstate 1-405, and
10 meters from a busy arterial road in Los Angeles. From this monitor, average
concentrations were not substantially higher than the local urban background
concentration (Sabin et al.. 2006b). Insufficient data are available to ascertain if the near
road Pb-TSP concentrations would be higher at lower monitor heights. The
2006 Pb AQCD (U.S. EPA. 2006b) also noted a smaller estimate of 40% for the Southern
California Air Basin (Lankey et al., 1998).
In a recent paper, Laidlaw and Filipelli (2008) analyzed Interagency Monitoring of
Protected Visual Environments (IMPROVE) data to explore conditions under which
PM2 5 particles estimated to be of crustal origins that may contain Pb may become
airborne. They observed a seasonal pattern in the concentration of PM25 of crustal
origins in the atmosphere, and they also found that at one IMPROVE site in central
Illinois, 83% of the variability in concentrations of crustal PM25 was predicted by
variability in meteorology and soil moisture content. The authors concluded that
seasonality and climate parameters could not be eliminated in relation to ambient air Pb
concentrations. Such mechanisms are described in more detail in Section 2.3. As
described in Section 2.2.2.6 and Section 2.6.1. there are many contemporary
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contributions of Pb to soil in urban areas, and studies summarized here have not
quantitatively differentiated the contributions of these various sources to Pb
concentrations in urban areas.
Pb from Industrial Activities
Several studies have indicated elevated levels of Pb are found in soil exposed to industrial
emissions, including brownfield sites (Dermont et al.. 2010; Verstraete and Van
Meirvenne. 2008; Jennings and Ma, 2007; van Herwijnen et al., 2007; Deng and
Jennings. 2006). Pb in industrial soils is described in Section 2.6.1. Recent Pb speciation
results also indicate a contribution from resuspended soils in areas with previous major
emission sources, but without current major sources. Data from airborne PM in the
vicinity of an inactive smelter in El Paso, TX were described as consistent with
Pb-humate as the major form of Pb in airborne PM, which the authors suggest relates to
soil resuspension since the local near-surface soils appeared to have had high humic
content (Pingitore et al.. 2009).
Pb from Paint and Building Materials
Exterior structures painted with Pb-based paint have long been known to be a source of
Pb in outdoor dust or grit (U.S. EPA. 2006b). Recent studies support earlier findings.
Mielke and Gonzales (2008) sampled exterior paint chips from paint applied prior to
1992 on 25 homes in New Orleans, LA, and they found elevated Pb levels in 24 of the 25
tested exterior paints. Weiss et al. (2006) studied the distribution of Pb concentration in
roadway grit in the vicinity of steel structures in New York City and contrasted those data
with roadway grit concentration data where no steel structure was nearby. In each case,
the difference was significant (p <0.006 at one site and p <0.0001 at 4 other sites), with
median Pb concentrations in the grit under the steel structures (median: 1,480 mg/kg)
collectively being 4.4 times higher than median Pb concentrations in the roadway grit not
near a structure (median: 340 mg/kg).
The studies described above considered paint as a source of Pb in outdoor dust through
gradual abrasion of the painted surfaces. However, atmospheric conditions may break
down polymers in aging paint, causing previously bound Pb-based pigments to be
released from the surface more readily. Edwards et al. (2009) performed experiments to
simulate one week of exposure of Pb-based paints to highly elevated levels of O3
(11.3 ± 0.8 mg/kg or 150 times the level of the 8-hour NAAQS) and NO2
(11.6 ± 0.9 mg/kg, or 220 times the level of the annual NAAQS). Following NO2
exposure, the Pb in wipe samples increased by a median of 260% (p <0.001), and
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following O3 exposure, the Pb in wipe samples increased by a median of 32%
(p = 0.004).
Building demolition was listed as a source of Pb in urban dust in the 2006 Pb AQCD
(U.S. EPA. 2006b). In a follow-up study to previous work cited therein, Farfel et al.
(2005) observed that surface loadings of dust containing Pb increased by 200% in streets,
by 138% in alleys, and by 26% in sidewalks immediately following demolition of an old
building compared with surface loadings of dust containing Pb prior to demolition. One
month later, Pb dust loadings were still elevated in alleys (18% higher than pre-
demolition) and sidewalks (18% higher than pre-demolition), although they had
decreased in streets by 29% compared with loadings prior to demolition. However, Farfel
et al. (2005) did not provide detailed time series samples from before or after demolition
to judge whether the observations made one month following demolition were within the
normal conditions of the urban area. These results suggest that building demolition may
be a short-term source of Pb in the environment, but it is unclear if demolition is related
to long-term Pb persistence in the environment.
Pb from Historic Automobile Emissions
Historic Pb emissions, or Pb emitted from on-road vehicles prior to the ban on use of
leaded automobile gasoline, deposited onto soil and in some areas may serve as a
potential source of airborne Pb. The historical use of leaded on-road gasoline has been
estimated from documents submitted by Ethyl Corporation to the U.S. Senate (1984) and
a report by the U.S. Geological Survey (USGS. 2005): see Mielke et al. (20 lie). These
estimates are presented in Figure 2-7. The peak U.S. use of Pb additives occurred
between 1968 and 1972 with an annual amount of over 200,000 metric tons. According to
Ethyl Corporation, the 1970 use of Pb additives was 211,000 metric tons. By 1980, the
annual use of Pb additives to on-road gasoline decreased to about 91,000 metric tons or a
57% reduction from its 1970 peak. From 1970 to 1990 there was a 92% decline in Pb
additive use. In 1990, the annual U.S. use of Pb additives decreased to
16,000 metric tons, a further 82% decline in Pb additive use from 1980. The final U.S.
ban on the use of Pb additives for highway use in on-road gasoline occurred in 1996.
After that time, Pb additives were only allowed in nonroad applications, including piston-
engine aircraft fuel, racing fuels, farm tractors, snowmobiles, and boats.
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Lead Additives in U.S. Gasoline
250,000
200,000
150,000
100,000
50,000
Year
Note: Estimates were derived from the proceedings of the U.S. Senate hearings on the Airborne Pb Reduction Act of 1984, S. 2609
(1984) and the U.S. Geological Survey Pb end use statistics (USGS. 2005).
Source: Reprinted with permission of Pergamon Press, Mielke et al. (2011c).
Figure 2-7 Total U.S. Pb additives in on-road gasoline used in on-road
vehicles, 1927-1995.
Pb emissions from on-road sources were estimated by the U.S. EPA (1986a), which
indicated that 75% of Pb additives were emitted as exhaust, while the remainder were
retained within the engine. The tonnages of relatively large >10 (im mass median
aerodynamic diameter (MMAD) Pb-PM probably settled locally. EPA Q986a) indicated
that 35% of the Pb-PM at that time were <0.25 urn in MMAD. In high traffic urbanized
areas, soil Pb from historic emissions as well as contemporary sources, are elevated
adjacent to roadways and decrease with distance away from roadways (Laidlaw and
Filippelli. 2008).
The use of Pb additives resulted in a national scale of influence. For example, variously
sized urbanized areas of the U.S. have different amounts of vehicle traffic associated with
Pb (Mielke et al.. 2010). Figure 2-8 illustrates the national scale of the estimated vehicle-
derived Pb aerosol emissions. Note that the estimated 1950-1982 Pb aerosol emissions in
the 90 cities below vary from 606 metric tons for Laredo, Texas, to nearly
150,000 metric tons for the Los Angeles-Long Beach-Santa Anna urbanized area.
Although this figure might imply that the soil Pb concentration in these areas would be
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proportional to the magnitude of historic on-road emissions in each city, it is recognized
that the atmospheric dispersion of emissions, as well as the atmospheric deposition and
subsequent distribution associated with surface runoff, will have varied substantially
among the cities illustrated. The 2006 and prior CDs described the wealth of evidence
documenting the long-range dispersion of Pb emitted from tail pipes beyond the U.S. and
depositing in polar ice and seawater of both oceans (U.S. EPA. 2006b. 1986a. 1977).
Additionally, the amount of soil turnover since 1982 may have varied substantially
among the cities illustrated in Figure 2-8. depending on the amount of highway
construction in those cities. As noted in Section 2.2.2.6. there have historically been, and
are currently, many additional sources of Pb contributing to near-roadway soil Pb
concentrations. Data are lacking that quantify the range of airborne Pb concentrations
originating from historic Pb in resuspended soil particles, but data on airborne
concentrations near roadways indicate measured air Pb concentrations (from all
contributing sources) to be generally less than 0.02 (ig/m3 (Section 2.5.3.2).
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U.S. Urbanized Areas
Pb 1950 -1982 (metric tons)
• 608
• 652-4.570
• 4.662 - 8,108
0 8.403-16.614
Ł 16.623-91.878
A 149.938
.r *• *
l
%
*
Note: The numbers on the map are rankings of each urbanized area (UA). The size of each dot refers to the magnitude of motor
vehicle gasoline-related emissions for each group of UAs.
Source: Reprinted with permission of Pergamon Press, Mielke et al. (2011c)
Figure 2-8 Estimated Pb aerosol inputs from on-road gasoline into 90 U.S.
urbanized areas (UAs), from 1950 through 1982.
2.3 Fate and Transport of Pb
There are multiple routes of exposure to Pb, including direct exposure to atmospheric Pb,
exposure to Pb deposited in other media after atmospheric transport, and exposure to Pb
in other media that does not originate from atmospheric deposition. As a result, an
understanding of transport within and between media such as air, surface water, soil, and
sediment is necessary for understanding direct and indirect impacts of atmospheric Pb as
well as the contribution of atmospheric Pb to total Pb exposure through inhalation and
ingestion pathways. Figure 2-9 describes relevant Pb transport pathways through
environmental media discussed in this chapter and their relationship to key environmental
and human exposure pathways for which some or all of the Pb is processed through the
atmosphere. This discussion includes recent research on atmospheric transport of Pb,
atmospheric deposition and resuspension of Pb, Pb transport in surface waters and
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sediments, and Pb transport in soil. Facets of fate and transport relevant to the ecological
effects of Pb are also summarized in Section 6.2.
Biota
Source
Agriculture +
Livestock
Human
Exposure
Groundwater
DrinkingWater
Transport
to Sea
Note: Media through which Pb is transported and deposited are shown in bold.
Figure 2-9 Fate of atmospheric Pb.
2.3.1
Air
The 2006 Pb AQCD (U.S. EPA. 2006b) concluded that Pb was primarily present in
submicron aerosols, but that bimodal size distributions were frequently observed. Pb-PM
in the fine fraction is transported long distances, found in remote areas, and can be
modeled using Gaussian plume models and Lagrangian or Eulerian continental transport
models as reported by several studies. Good agreement between measurements and these
models have been reported. Historical records of atmospheric deposition to soil,
sediments, peat, plants, snowpacks, and ice cores have provided valuable information on
trends and characteristics of atmospheric Pb transport. Numerous studies using a variety
of environmental media indicated a consistent pattern of Pb deposition peaking in the
1970s, followed by a more recent decline. These findings indicated that the elimination of
leaded gasoline for motor vehicles and systematic reductions in emissions from other Pb
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sources has not only led to lower atmospheric concentrations in areas impacted by
vehicles (Section 2.5). but a pervasive pattern of decreasing atmospheric Pb deposition
and decreasing concentrations in other environmental media even at great distances from
sources.
2.3.1.1 Transport
Recent research on long range transport as well as transport of Pb in urban areas has
advanced the understanding of Pb transport in the atmosphere. While the 2006 Pb AQCD
described long range Pb transport as essentially a process of submicron PM transport
(U.S. EPA. 2006b), much of the recent research on Pb transport has focused on
interactions between anthropogenic and coarser geogenic PM that leads to incorporation
of Pb into coarse PM as well as subsequent transformation on exposure to mineral
components of coarse PM. Using scanning electron microscopy (SEM), Schleicher et al.
(2010) observed interactions of anthropogenic soot and fly ash particles on the surfaces
of coarse geogenic mineral particles in Beijing, China and concluded that toxic metals
were often associated with TSP. Murphy et al. (2007) found that PM released from wild
fires and transported over long distances scavenged and accumulated Pb and sulfate
through coagulation with small Pb rich PM during transport and that Pb was associated
with PM over a wide size range. Erel et al. (2006) also found that Pb enrichment factors
calculated for PMi0 from dust storms collected in Israel were much greater than those
sampled at their north African source, suggesting that the dust samples had picked up
pollutant Pb in transit between the Saharan desert and Israel. Marx et al. (2008)
characterized dust samples collected from the surface of glaciers and in dust traps on the
remote west coast of New Zealand's South Island and observed that most of the dust
samples were enriched in metals, including Pb, compared with their source area
sediments.
Pb accumulated on mineral dusts is also subject to atmospheric transformations. PbSO4 is
one of the main constituents of Pb-containing aerosols resulting from coal combustion
(Giere et al.. 2006) and it has been shown to react with calcite, CaCO3, a PM mineral
component, to form Pb3(CO3)2(OH)2, Pb(CO3) and Ca(SO4)2'H2O on the surface of the
calcite (Falgayrac et al.. 2006). In laboratory experiments, (Ishizaka et al.. 2009) also
showed that PbSO4 could be converted to PbCO3 in the presence of water.
Approximately 60-80% was converted after only 24 hours for test samples immersed in a
water droplet. This compared with only 4% conversion for particles that had not been
immersed. As a result of recent research, there is considerable evidence that appreciable
amounts of Pb can accumulate on coarse PM during transport, and that the physical and
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chemical characteristics of Pb can be altered by this process due to accompanying
transformations.
2.3.1.2 Deposition
The 2006 Pb AQCD (U.S. EPA. 2006b) documented that soluble Pb was mostly removed
by wet deposition, and most of the insoluble Pb was mostly removed by dry deposition.
As a result, dry deposition was the major removal mechanism for Pb in coarse PM (which
is mainly insoluble and settles faster than fine PM), and wet deposition was the most
important removal mechanism for fine PM and Pb halides (which were more soluble).
Numerous studies reported that Pb dry deposition velocities in the U.S. were mostly
within a range of 0.05 to 1.0 cm/second and dry deposition fluxes ranging from 0.04 to
4 mg/m2 per year. Precipitation concentrations ranged mostly from 0.5 to 60 ug/L, but
with considerably lower concentrations in remote areas, and wet deposition fluxes in the
U.S. ranged from 0.3 to 1.0 mg/m2 per year. Wet deposition was linked to precipitation
intensity, with slow even rainfalls usually depositing more Pb than intense rain showers.
Rain concentrations decreased dramatically between the early 1980s and the 1990s,
reflecting the overall decreasing trend in Pb emissions due to elimination of leaded motor
vehicle gasoline. A summary of studies investigating total deposition including both wet
and dry deposition indicated typical deposition fluxes of 2-3 mg/m2 per year and dry to
wet deposition ratios ranging from 0.25 to 2.5. Seasonal deposition patterns can be
affected by both variations in local source emissions and vegetation cover, and as a result
a consistent seasonal pattern across studies has not been observed, although there have
been only a few investigations. The 2006 Pb AQCD (U.S. EPA. 2006b) concluded that
resuspension by wind and traffic contributes to airborne Pb near sources.
Wet Deposition
The 2006 Pb AQCD (U.S. EPA. 2006b) documented that dry deposition was the major
removal mechanism for Pb in coarse PM and wet deposition as the most important
removal mechanism for fine PM. Which process is most important for atmospheric
removal of metals by deposition is largely controlled by solubility in rain water. Metal
solubility in natural waters is determined by a complex multicomponent equilibrium
between metals and their soluble complexes and insoluble ionic solids formed with
hydroxide, oxide, and carbonate ions. This equilibrium is strongly dependent on pH and
ionic composition of the rain water. As pH increases, Pb solubility is reduced. Theodosi
et al. (2010) found that solubility of Pb was near 100% when rain water pH was
measured to be less than 4.5. As a consequence, it is possible that efforts to reduce acidity
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of precipitation could also reduce wet deposition of Pb. Recent research confirms the
general trend described in the 2006 Pb AQCD (U.S. EPA. 2006b) that Pb associated with
fine PM is usually more soluble in rain water than Pb associated with coarse PM, leading
to a relatively greater importance of wet deposition for fine Pb and of dry deposition for
coarse Pb. Theodosi et al. (2010) concluded that larger particles were less soluble,
because Pb solubility decreased with increasing dust loading. Likewise, Preciado and Li
(2006) observed that solubility decreased with increasing particle size. Moreover,
Theodosi et al. (2010) observed that 53% of wet deposition samples were comprised of
particulate Pb, not soluble Pb. This finding suggests that wash-out can be equally
important to wet deposition as solubility.
Although recent observations are consistent with previous findings, they also indicate
considerable spatial and seasonal variability. Birmili et al. (2006) found that Pb solubility
varied between the two main Pb-containing size fractions, <0.5 (im (-40%) and
1.5-3.0 (im (-10%), indicative of a different chemical speciation. However, the
observation that the amount of soluble Pb was higher in their U.K. samples than in an
analytically identical study carried out in Seville, Spain (Fernandez Espinosa et al..
2004). led them to conclude that Pb solubility in fine PM may vary on a regional basis
(Birmili et al.. 2006). For PM10 from Antarctica, 90 to 100% of the Pb was insoluble at
the beginning of the summer season (November), but by the end of the summer
(January), approximately 50% was soluble. Most of the Pb was from long range transport
(Annibaldi et al.. 2007). These studies illustrate the variable nature of atmospheric Pb
solubility.
Dry Deposition
Recent research on dry deposition has focused on differences between urban or industrial
sites and rural or less industrial areas. For locations outside of industrial areas, new
measurements of Pb dry deposition fluxes are similar to those reported in the
2006 Pb AQCD (U.S. EPA. 2006b). but in industrialized urban areas, they are
considerably greater than in nonindustrialized areas. Deposition is typically documented
by measurements of Pb concentrations on surface material or by measurements of flux.
For example, Hasselbach et al. (2005) documented Pb concentration in moss, and the
spatial distribution of Pb concentration in moss, as evidence of Pb deposition from truck
traffic between an Alaskan Zn-Pb mine and a port. Additionally, Maher et al. (2008)
measured Pb deposition onto leaves near a road in Norwich, U.K. Their results are
described in detail in Section 2.6.6.
Several studies presented measurements of dry deposition flux obtained by capturing
deposited particles onto a sampling substrate. Hence, these measurements did not provide
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information on net deposition following resuspension of deposited material.
Resuspension processes and measurements thereof are described in Section 2.3.1.3. Yi et
al. (2006) calculated dry deposition fluxes for trace elements including Pb in New York-
New Jersey harbor and observed much greater dry deposition fluxes for this urban
industrial site in Jersey City (mean: 50 (ig/m2-day) than for suburban New Brunswick
(mean: 8 (ig/m2-day). Sabin and Schiff (2008) measured dry Pb deposition flux along a
transect from Santa Barbara to San Diego, CA in 2006 and observed a range of
0.52-14 (ig/m2-d for the median values across the eight sites. The highest median Pb flux
was observed at Los Angeles Harbor, which is downwind of a harbor with a mix of
industrial (harbor-related) and urban activities (14 (ig/m2-day). The second highest
median Pb flux was observed at San Diego Bay, a military port (3.3 (ig/m2-day). This is
consistent with similar observations of dry deposition fluxes that were more than ten
times greater in urban Chicago than in rural South Haven, Michigan (Paode et al.. 1998).
These results illustrate the strongly localized nature of atmospheric Pb deposition in
source rich areas. In a study of Tokyo Bay, Sakata and Asakura (2008) reported an
average dry deposition velocity of 1.06 cm/second, which is near the upper end of dry
deposition velocities reported in the 2006 Pb AQCD (U.S. EPA. 2006b). They also
reported that dry deposition fluxes were greater in industrially impacted urban areas,
ranging from 12-17 mg/m2 per year.
Recent results also confirmed the trend of decreasing overall deposition fluxes after
removal of Pb from on-road gasoline, as described in the 2006 Pb AQCD (U.S. EPA.
2006b). Watmough and Dillon (2007) found that the bulk annual deposition of Pb in a
central Ontario forested watershed during 2002-2003 was 0.49 mg/m2 per year; this was
lower than the value of 1.30-1.90 mg/m2 per year for 1989-91 and represented a 75%
decline in Pb deposition. It was consistent with the decline more generally observed for
the Northeastern U.S. as a consequence of the restrictions to alkyl-Pb additives in on-road
gasoline. From previously published work, and in agreement with the precipitation data
described above, most of the decline in Ontario Pb deposition took place before the start
of the Watmough and Dillon (2007) study.
Within-day variation in deposition fluxes was observed to be related to the urban
boundary layer. Lim et al. (2006) observed higher deposition fluxes of Pb and other
metals during the nighttime in Los Angeles, when inversions are frequent occurrences.
Deposition mass was also greater for particles larger than 10 pirn in the urban areas where
measurements occurred. Larger particle deposition flux was greater during the day.
Several important observations can be highlighted from the few studies of atmospheric
Pb deposition carried out in the past several years. Deposition fluxes have greatly
declined since the removal of Pb additives from on-road gasoline. However, more recent
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results in industrial areas indicate that local deposition fluxes there are much higher than
under more typical conditions. In general, wet deposition appears to be more important
for Pb in fine PM, which is relatively soluble; and dry deposition appears to be generally
more important for Pb in coarse PM, which is relatively insoluble. However, the relative
importance of wet and dry deposition is highly variable with respect to location and
season, probably reflecting both variations in Pb speciation and variations in external
factors such as pH and rain water composition. Although industrial Pb emissions are
mainly associated with fine PM, and wet deposition is likely to be more important for this
size range, a substantial amount of Pb is apparently removed near industrial sources.
2.3.1.3 Resuspension of Pb from Surface Soil to Air after
Deposition
The following information focuses on issues regarding the transport processes affecting
resuspended soil Pb and dust Pb in urban environments. As described in Section 2.2.1.
the greatest point source Pb emissions in the U.S. occur in locations near specific major
facilities, such as secondary smelters, and other industrial operations involving large
scale metal processing or fuel combustion. However, in the absence of such sources and
in the vicinity of previous major sources, the 2006 Pb AQCD (U.S. EPA. 2006b)
concluded that resuspension by mechanical stressors such as traffic, construction, and
wind can be a source of airborne Pb above background levels near sources, with Pb
accounting for between 0.002 to 0.3% of the mass of resuspended PMi0. Additionally,
solubility of Pb increases its soil mobility, while adsorption to inorganic soil complexes
decreases mobility (McLean and Bledsoe. 1992). These factors influence the site of Pb
content within soil and hence its ability to become resuspended. Factors affecting Pb
mobility in soil are discussed in more detail in Section 2.3.3.3. Reentrainment of
deposited Pb complicates air related pathways of human (Section 3.1.1) and ecosystem
exposure (Section 6.2.2).
Results from several studies have suggested minor contributions from resuspension to
airborne Pb concentration from various sources, including city centers (Laidlaw and
Filippelli. 2008). major freeways (Sabin et al.. 2006b). and steel structures with abrading
paint (Weiss et al.. 2006). Recently, Laidlaw et al. (2012) modeled concurrent
measurements of (log-transformed) air Pb-PM2s as a function of (log-transformed)
airborne soil measured in PM25. They observed a statistically significant increase in air
Pb of 0.84% with a 1% increase in airborne soil (p <0.01). As noted in the
2006 Pb AQCD (U.S. EPA. 2006b). the contribution of resuspended soil and dust to the
airborne burden may be significant from highly contaminated sites (e.g., active or
abandoned industrial facilities and Superfund sites). In contrast, as summarized in
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Section 2.5.3. Pb concentrations near roads in urban areas are one to two orders of
magnitude below the current Pb NAAQS.
The urban environment can be considered quite different from natural landscapes because
it has been highly modified by human activity, including above- and below-ground
infrastructure, buildings, and pavement, and a high density of motorized transportation.
These factors may influence the distribution and redistribution of Pb-bearing PM. As
shown in Figure 2-10. urban turbulence occurs on several scales. Transport and
dispersion of urban grit is subject to air movement within the urban canopy layer, where
air movement is driven by air velocity within the urban boundary layer and urban
topographical conditions such as building shape, building facade, and street canyon
aspect ratio (Fernando. 2010). Within a street canyon, air circulates and tends to form
counter-rotating eddies along the height of the canyon (Figure 2-10). which result in
lower mean components of air movement, higher turbulence components, and higher
shear stress within the canyon compared with open field conditions (Kastner-Klein and
Rotach. 2004; Britter and Hanna. 2003). Recirculation around intersection corners and
two-way traffic conditions can also enhance turbulence levels, while one-way traffic
conditions increase air velocity along the street (Soulhac et al.. 2009; Kastner-Klein et al..
2003; Kastner-Klein et al.. 2001). Sedefian et al. (1981) measured the length scales of
turbulent eddies resulting from passing 50 mph (22.2 meters/second) traffic on a test road
and observed scales of 0.6-2.7 meters when winds were perpendicular to the test road and
scales of 1.8-2.7 meters when winds were parallel to the road.
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ABL
ABL. - Atmospheric boundary layer
UBL - Urban boundary layer
RSL - Roughness sublayer
(transition layer, wake layer, inlerfacial layer)
UCL - Urban canopy layer
USL - Urban surface layer
ML • Mixed layer
CFL - Constant flux layer
(Lejnertial sublayer- ISL)
Street
Canyon
Recirculation
Note: Top: multiple scales within the atmospheric boundary layer. Bottom: illustration of airflow recirculation within a single street
canyon located in the urban canopy layer.
Source: Adapted with permission of Annual Reviews, Fernando (2010)
Figure 2-10 Scales of turbulence within an urban environment.
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Recent research on urban PM transport is highly relevant to Pb transport and dispersion
because Pb is most prevalently particle-bound. Relevant results for Pb exposure in these
areas include observations that PM concentration peaks dissipate more rapidly on wider
streets than in narrow street canyons (Buonanno et al.. 2011); concentrations are typically
low next to a building because either less source material is available or less material
penetrates the boundary layer of the building (Buonanno et al.. 2011); and there are
stronger inverse relationships between mean wind speed and PM concentration
fluctuation intensities at middle sections of urban street blocks compared with
intersections (Hahn et al.. 2009). Patra et al. (2008) conducted experiments in London,
U.K. in which a "tracer" grit (i.e., rock salt) was applied to a road and then the grit's
dispersion by traffic was measured overtime to simulate resuspension and transport of
road dust. During the experiments, 0.039% of the tracer grit was measured to move down
the road with each passing vehicle, 0.0050% was estimated to be swept across the road
with each passing vehicle, and 0.031% was estimated to become airborne when a vehicle
passed.
Harris and Davidson (2008) developed a model of resuspension of single particles
initially at rest on a solid surface based on the balance of lift, drag, gravity, torque, and
adhesion forces on the particle in addition to turbulent wind fluctuations within a
simulated urban boundary layer. Their model simulations showed 2.5 (im and 10 (im
particles to reach a maximum height of 0.04-0.06 meter above ground level (AGL), while
50 (im particles reached a maximum of 0.2 meter AGL and 75 (im particles reached at
least 0.4 meter AGL, depending on friction velocity. Empirical analysis has shown that
lift force is proportional to particle diameter to the power of approximately 1.5, so that
large particles actually have larger initial vertical displacement than smaller particles. At
the same time, lateral travel distance following resuspension tended to decrease linearly
with increasing particle size, reflecting the counteracting force of gravity. For all cases
simulated, the resuspension and deposition were estimated to occur over time frames on
the order of seconds.
Early work described resuspension as an important process for wind erosion for particles
up to 100 (im but indicated that particles larger than this rarely became suspended, and
that the tendency of particles to remain airborne long enough for appreciable transport
decreases sharply beyond a size of 10 to 20 (im (Nicholson. 1988; Gillette et al.. 1974).
As a result, long range transport of dust is usually limited to particles smaller than 10 (im
(Prospero. 1999).
In urban environments the transport distance that must be traversed to penetrate indoors
can be very short, and at the same time resuspension and dispersion of larger particles
may be caused by motor vehicles. Resuspension of road dust by traffic becomes more
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difficult with decreasing particle size because adhesive forces are stronger than shear
force that is imparted by traffic-induced turbulent air movement (Harris and Davidson.
2008). The critical diameter at which resuspension occurs when a particle's settling
velocity becomes lower than the friction velocity of air needed to move the particle from
rest. The work of Gillette et al. (1974). in which a critical diameter of roughly 20 (im was
estimated, is based on wind in an open landscape. It would be reasonable to expect that
friction velocity would be higher for urban environments with traffic-induced turbulence
(Britter and Hanna. 2003). Hence, it is possible that larger particles are resuspended in a
heavily-trafficked urban setting (Nicholson and Branson. 1990).
Particle size determines the distance particles can travel and the height which they can
achieve before they are removed by gravitational settling. Song and Gao (2011) observed
that coarse mode Pb concentration was negatively correlated with wind speed (Dp = 14:
p = -0.62; Dp = 7.8; p = -0.76), which suggests that coarse Pb may be dispersed by wind.
Observations in near road environments indicate that roughly 15% of Pb in airborne dust
in areas impacted by heavy traffic is greater than 10 (im (Cho etal.. 2011; Lough et al..
2005; Zereini et al.. 2005). Sabin et al. (2006b) also collected three size fractions greater
than 11 (im and found that approximately 25% of all Pb mass was associated with
particles larger than 29 (im at a site 10 meters from a freeway, but only a very small
percentage of Pb mass was in this size fraction at an urban background site. These results
suggest that both size distribution and concentrations in the immediate vicinity of
roadways might differ from estimates based on concentrations from monitoring sites at
some distance from roads or on elevated rooftops. In these studies, only one size fraction
slightly greater than 10 (im was collected, but another study of road dust (not specific to
Pb) reported size fractions extending up to 100 (im with a mass median diameter of
greater than 60 (im (Yang etal.. 1999). Although the Yang et al. (1999) study did not
include Pb, the results suggest that resuspended dust can be larger than PMi0.
Collectively, the size distribution of Pb-containing resuspended dust is uncertain.
Recent resuspension studies complement previous research indicating street dust half-
lives on the order of one-hundred days (Allott et al.. 1989). with resuspension and street
run-off as major sinks (Vermette et al.. 1991) as well as observations of a strong
influence of street surface pollution on resuspension (Bukowiecki et al.. 2010).
observations of greater resuspension of smaller PM than coarser PM, leading to
enrichment of metal concentrations in resuspended PM relative to street dust (Wong et
al.. 2006) and observations of wind speed, wind direction, vehicular traffic, pedestrian
traffic, agricultural activities, street sweeping and construction operations as important
factors determining resuspension.
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2.3.2 Water
As described in the 2006 Pb AQCD (U.S. EPA. 2006b). atmospheric deposition, urban
runoff, and industrial discharge have been identified as major sources of Pb in surface
waters. Water columns have been described as transient reservoirs with Pb residence
times in lakes typically several months long, and shorter residence times expected in
turbulent waterways. Because dispersal in waterways is a relatively rapid process,
concentrations in surface waters are highest near sources of pollution before substantial
Pb removal by flushing, evaporation, and sedimentation occurs. Transport in surface
water is largely controlled by exchange with sediments, and the cycling of Pb between
water and sediments is governed by chemical, biological, and mechanical processes that
are affected by many factors, including salinity, organic complexation, oxidation-
reduction potential, and pH; ecological impacts of these factors are described in Section
6.4.9. As described in the 2006 Pb AQCD (U.S. EPA. 2006b). metals in waterways are
transported primarily as soluble chelates and ions, or adsorbed on colloidal surfaces,
including secondary clay minerals, iron (Fe) and manganese (Mn) oxides or hydroxides,
and organic matter, and adsorption on organic or inorganic colloids is particularly
important for Pb. The extent of sorption strongly depends on particle size as smaller
particles have larger collective surface areas. Aqueous Pb concentrations also increase
with increasing salinity. Pb is found predominantly as PbO or PbCO3 in aqueous
ecosystems. Pb is relatively stable in sediments, with long residence times and limited
mobility. However, Pb-containing sediment particles can be remobilized into the water
column. As a result trends in sediment concentration tend to follow those in overlying
waters. Fe and Mn oxides are especially susceptible to recycling with the overlying water
column. Although resuspension of sediments into overlying waters is generally small
compared to sedimentation, resuspension of contaminated sediments is often a more
important source than atmospheric deposition. Organic matter (OM) in sediments has a
high capacity for accumulating trace elements. In an anoxic environment, removal by
sulfides is particularly important. The following section highlights recent literature
regarding the fate and transport of Pb in water systems. Section 6.2 synthesizes this
information with ecosystem exposure data.
Runoff from storms was identified as a potentially important source of Pb in aquatic
systems (U.S. EPA. 2006b). Runoff from atmospheric deposition, painted buildings,
gutters, roofing materials and other housing materials were identified as major
contributors to Pb in runoff waters. Investigations of building material contributions
indicated runoff concentrations ranging from 2 to 88 mg/L, with the highest
concentrations observed from more than 10-year-old paint and the lowest concentrations
from residential roofs. There was some indication that Pb from roofing materials, siding,
and piping could be due to dissolution of Pb carbonate (cerussite) or related compounds.
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In several studies Pb in runoff was consistently mostly PM, with a relatively small
dissolved fraction. Runoff release was dependent on storm intensity and length of dry
periods between rain events, with greater runoff of Pb associated with more intense
storms and with longer periods between rain events. Several studies indicated a "first
flush effect," with highest runoff concentrations observed at the beginning of a rain
event.
2.3.2.1 Pb Transport in Water and Sediment
Recent publications provide additional detail regarding Pb adsorption on Fe-rich and
organic-rich colloids. Correlation between Pb concentration in unfiltered water with total
Fe was observed (Hassellov and von der Kammer. 2008). which is consistent with
previous research using cross flow filtration (Pokrovsky and Schott 2002; Ross and
Sherrell. 1999) and SEM examination of single particles (Taillefert et al.. 2000).
Two distinct colloidal phases, one organic-rich (0.5-3 nm in diameter) and the other Fe-
rich (>3 nm in diameter), have been observed to coexist in both soil isolates and river
water (Stolpe and Hassellov. 2007). Pb was observed to be predominantly associated with
Fe-oxide PM in river water but also associated with the organic colloids in the soil
isolates (Hassellov and von der Kammer. 2008). Investigation of Pb binding onto
ferrihydrite showed Pb binding data were consistent with Pb being held at the surface by
sorption processes, rather than enclosed within the particle structure (Hassellov and von
der Kammer. 2008).
Observations in boreal rivers and soil pore waters in permafrost dominated areas of
Central Siberia indicated that Pb was transported with colloids in Fe-rich waters. Trace
elements that normally exhibited limited mobility (including Pb) had 40-80% of their
annual flux in the nominal dissolved phase, operationally defined as material that passes
through a 0.45 (im pore-size filter, and that these metals had a higher affinity for organo-
mineral Fe-Al colloids (Pokrovsky et al.. 2006). Pokrovsky et al. (2006) postulated that
during the summer, rainwater interacts with degrading plant litter in the top soil leading
to the formation of Fe-Al-organic colloids with incorporated trace elements. Migration of
trace element-Fe-Al-OM colloids may result in export of Pb and other elements to
riverine systems. Most of the transport occurred after thawing had commenced. This
contrasts with permafrost free areas where trace elements such as Pb are incorporated
into Fe colloids during OM-stabilized Fe-oxyhydroxide formation at the redox boundary
of Fe(II)-rich waters and surficial DOC-rich horizons. Similarly, during a spring flood
(May) that exported 30-60% of total annual dissolved and suspended flux of elements
including Pb, Pb was mainly in the nominal dissolved phase, operationally defined as
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material that passes through a 0.45 (im pore-size filter (Pokrovsky et al.. 2010). Pb
adsorbed on colloidal surfaces rather than incorporated into particle structure is likely to
be more readily dissolved because dissolution of the entire particle is not required.
Recent research on retention of Pb in water bodies and sediments has focused on the
estuarine and marine environment, where considerable retention of Pb was observed in
estuarine sediments. For a large riparian system, the Trinity River, Texas, Warnken and
Santschi (2009) found that 80% of riverine Pb was retained in Lake Livingston, an
estuarine region, while an additional 16% was removed to estuarine sediments, and only
about 4% eventually reached the ocean. Geochemical (sorption by Fe oxyhydroxides),
biological (seasonal uptake by sinking algae in Lake Livingston) and hydrological
(dilution effects by increasing flow rates) processes were mainly responsible for
controlling dissolved trace metal concentrations rather than pollution sources.
Overall, recent research on Pb transport in aquatic systems has provided a large body of
observations confirming that Pb transport is dominated by Fe-rich and organic-rich
colloids. In addition, new results indicated that although the 2006 Pb AQCD (U.S. EPA.
2006b) described rivers and lakes as temporary reservoirs with Pb lifetimes of months or
less, estuaries can present a substantial barrier to transport into the open ocean.
2.3.2.2 Deposition of Pb within Bodies of Water and in Sediment
As described in the 2006 Pb AQCD (U.S. EPA. 2006b). in general Pb is relatively stable
in sediments, with long residence times and limited mobility. As described in previous
sections, Pb enters and is distributed in bodies of water largely in PM form. In rivers,
particle-bound metals can often account for > 75% of the total load, e.g., (Horowitz and
Stephens. 2008). Areas near historically Pb emitting industries and urbanized areas tend
to have greater aquatic Pb loads than areas more remote from Pb sources, as several
studies have shown the strong positive correlation between population density and river
or lake sediment Pb concentrations (Horowitz et al. 2008; Chalmers et al.. 2007). Indeed,
Chalmers et al. (2007) revealed that in river and lake sediments in New England, there
was an order of magnitude difference between Pb sediment concentrations in rural versus
urbanized areas.
The fate of Pb in the water column is determined by the chemical and physical properties
of the water (pH, salinity, oxidation status, flow rate and the suspended sediment load
and its constituents, etc). Desorption, dissolution, precipitation, sorption and
complexation processes can all occur concurrently and continuously, leading to
transformations and redistribution of Pb. The pH of water is of primary importance in
determining the likely chemical fate of Pb in terms of solubility, precipitation or organic
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complexation. In peatland areas, such as those in upland areas of the U.K., organic acids
draining from the surrounding peatlands can lower stream water pH to below 4. Under
these conditions, Pb-PM can be desorbed and released into solution, leading to elevated
dissolved Pb concentrations (Rothwell et al.. 2008). At the other end of the pH scale, Pb
tends to remain or become complexed, precipitated or sorbed to suspended sediments in
water, as observed by Das et al. (2008) who studied trace metal geochemistry in a South
African lake with water pH of 9. They also found marked differences in Pb
concentrations associated with increasing depth in the water column [e.g., the surface
Pb-PM concentration of 2 (ig/L increased to 60 (ig/L at depth and the Pb concentration in
the <0.45 (im fraction increased from 2 (ig/L at the surface to 19 (ig/L at depth (Das et
al.. 2008)]. This is suggestive of a settlement process in action.
In estuarine and wider marine environments the processes may be more complex because
of the additional perturbation caused by tidal action and the strong effects of salinity.
Again, PM forms of Pb are important in determining Pb distribution and behavior. Li et
al. (2010a) reported that PM Pb accounted for 85 ± 15% and 50 ± 22% in Boston Harbor
and Massachusetts Bay, respectively, while Lai et al. (2008b) reported a solid (acid
soluble): dissolved Pb ratio of 2.6 for areas of the Australian sector of the Southern
Ocean.
The accurate modeling of Pb behavior in marine waters (including estuaries) requires
consideration of many parameters such as hydrodynamics, salinity, pH, suspended PM,
fluxes between PM and dissolved phases (Hartnett and Berry. 2010). Several new
advances in the study of Pb cycling in these complex environments have been described
in recent publications. Li et al. (2010a) used particle organic carbon (POC) as a surrogate
for the primary sorption phase in the water column to describe and model the partitioning
of Pb between PM and dissolved forms. Huang and Conte (2009) observed that
considerable change in the composition of PM occurs as they sink in the marine
environment of the Sargasso Sea, with mineralization of OM resulting in increased
Pb-PM concentration with increased depth. As a result of this depletion of OM in sinking
particles, geochemical behavior at depth was dominated by inorganic processes,
e.g., adsorption onto surfaces, which were largely independent of Pb source. Sinking
rates in marine environments can vary, but a rate approximating 1 meter/day has been
used in some models of Pb transport and distribution in aquatic-sediment systems (Li et
al.. 2010a). Surface sediment Pb concentrations for various continental shelves were
collated and compared by Fang et al. (2009) (Table 2-2).
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Table 2-2 Surface sediment
Location
East China Sea
Mediterranean, Israel coast
Aegean Sea
Bane d'Arguin, Mauritania
Campeche shelf, Gulf of Mexico
Laptev Sea, Siberia
Pechora Sea, Russia
Pb concentrations for various
Digestion solution
HCI/HNOs/HF
HNO3
HCI/HNOs/HF
HCI/HNOs/HF
HCI/HNOs
HCI/HNOs/HF
Not reported
continental shelves.
Pba (mg/kg)
10-49(27)a
9.9-20
21-44(34)
2.8-8.9
0.22-20 (4.3)
12-22
9.0-22(14)
aValues in parentheses are the average, where calculable
Source: Data from Fang et al. (2009) and references therein.
2.3.2.3 Flux of Pb from Sediments
Sediments can be either a source or a sink for metals in the aquatic environment. Release
can be via re-suspension of the sediment bed via wind, wave and tidal action or by
dissolution from sediment to the water column. When external Pb inputs to bodies of
water are decreased by environmental improvement actions or regulations, contributions
of Pb to the water column from the existing sediments can become an increasingly
important source. Roulier et al. (2010) determined that Pb flux from sediments originated
mostly from organic fractions, but also partially from Mn and Fe components undergoing
reductive dissolution. The rate of release was controlled by OM content, particle size,
clay type and content, and silt fraction (Roulier et al.. 2010). The importance of sediment
particle size, OM content and acid volatile sulfide concentration in relation to metal
release was similarly identified (Cantwell et al.. 2008): ecosystem effects related to
sulfide concentration are described in Section 6.4.2.5. The effect of pH change on Pb
release from lake sediments has also been examined, revealing that 1.8 protons (FT) were
exchanged per divalent metal cation released (Lee et al.. 2008a). Processes governing Pb
release from lake sediments, including microbial reductive dissolution of Fe, biogenic
sulfide production and metal sorption-desorption, have been investigated in a basin
heavily contaminated by historical precious metal mining activities, and results indicated
that release of Pb from suboxic and anoxic zones of the heavily contaminated sediment
act as a Pb source to the overlying water of the lake (Sengor et al.. 2007). Bacardit and
Camarero (2010a. b) performed a mass balance of Pb, Zn, and As for three lakes in the
Central Pyrenees in France to identify dominant metals distribution processes. They
estimated that flux from the catchment accounted for 91-99% of the lakes' Pb inputs,
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while sediment flux accounted for 98-99% of Pb outputs. In this paper, sediment was
only modeled as an output.
Disturbance of bed sediments also occurs by tidal action contributing to re-suspension of
sediments. Benthic fluxes of dissolved metals released from sediments measured in
Boston Bay were calculated as strong enough that in the absence of Pb inputs such
benthic flux would reduce sediment Pb concentrations in Boston Bay to background
levels in 30-60 years (Li et al.. 2010a). In a related way, a half-life for sediment Pb
(considering benthic flux alone as the loss mechanism) of 5.3 years was estimated for
marine sediments off the Belgian coast (Gao et al.. 2009). Atkinson et al. (2007)
conducted experiments in an area contaminated by metal smelters, Lake Macquarie,
Australia, to assess the factors that influence flux of metals from marine sediment. Low
pH (pH = 6 ± 1), bioturbation, and other mixing processes were found to have stronger
influence over flux than binding to sulfides, which were thought to be sequestered in
deeper sediments.
Radakovitch et al. (2008) investigated the riverine transport of PM including Pb to the
Gulf of Lion, France, and also concluded that a major part of annual fluxes could be
delivered over a short time period. From budget calculations, riverine inputs were more
important than atmospheric deposition and Pb concentrations in the prodelta sediments
showed a strong correlation with OM content. These sediments, however, were not
considered to be a permanent sink, as resuspension in these shallow areas was an
important process. OM, Pb and other metals were enriched in resuspended PM compared
with the sediment.
In a heavily contaminated, high salinity embayment upstream from Sydney Harbor in
Australia, Birch and O'Hea (2007) reported higher total suspended solids, turbidity and
total water metal concentration in surface compared with bottom water as well as a
difference in suspended PM metal concentrations between surface water and bottom
sediments, demonstrating that storm water discharge was the dominant process of metal
transfer during high rainfall events. Total suspended sediments (and total water metals) in
bottom water were higher than in the surface water plume, indicating that resuspension of
bottom sediment is a greater contributor of total suspended sediments than storm water
during such events, especially in shallower regions of the bay. Soto-Jimenez and Paez-
Osuna (2010) determined diffusive and advective fluxes, geochemical partitioning of Pb
and Pb-isotopic signatures in a study of mobility and behavior of Pb in hypersaline salt
marsh sediments. They determined that sulfides were the main scavengers for Pb that was
diagenetically released Pb.
Overall, recent research on Pb flux from sediments in natural waters provided greater
detail on resuspension processes than was available in the 2006 Pb AQCD (U.S. EPA.
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2006b), and confirms previous findings that resuspended Pb is largely associated with
OM or Fe and Mn particles, but that anoxic or depleted oxygen environments in
sediments play an important role in Pb cycling. This newer research confirms previous
findings that resuspension and release from sediments largely occurs during discrete
events related to storms. It has also confirmed that resuspension is an important process
that strongly influences the lifetime of Pb in bodies of water.
2.3.2.4 Pb in Runoff
Runoff is a major source of Pb in surface waters. This complicates any evaluation of the
contribution of atmospheric Pb to watersheds, which must take into account direct
atmospheric deposition, runoff of atmospherically deposited Pb, and runoff of Pb from
sources such as mine tailings or paint chips that are shed from outdoor structures. The
2006 Pb AQCD (U.S. EPA. 2006b) concluded that runoff was consistently mostly PM,
with a relatively small dissolved fraction, and that dissolution of carbonate and related
compounds were important contributors to Pb pollution in runoff waters. It also described
Pb release into runoff as dependent on storm intensity and length of dry periods between
rain events, and a "first flush effect," with highest runoff concentrations observed at the
beginning of a rain event. Subsequent research has provided considerable new
information about the flux of Pb from roadway and urban runoff and snow melt to
watersheds.
Severe contamination due to export of anthropogenic Pb to adjacent ecosystems via urban
runoff and domestic wastewater discharge and to a lesser extent by direct atmospheric
deposition has been documented (Soto-Jimenez and Flegal 2009). Recent investigations
also confirm roof runoff as an important contributor to Pb pollution. Huston et al. (2009)
measured Pb concentrations in water from urban rainwater tanks and found Pb
concentrations in bulk deposition were consistently lower than in water in the rainwater
tanks, but that sludge in the tanks had a high Pb content, indicating that not all major
sources of Pb are from atmospheric deposition. Pb levels frequently exceeded drinking
water standards. Pb flashing on the roofs was implicated as the source of Pb in the
rainwater tanks although other possible sources include old paint and Pb stabilized PVC
drain pipes (Lasheen et al.. 2008; Weiss et al.. 2006; Al-Malack. 2001).
New research has improved the understanding of suspended PM size ranges, speciation,
and impacts of Pb runoff from urban soil and road dust. Soil and road dust have been
identified as major sources of Pb pollution to near-coastal waters, leading to high Pb
concentrations in stormwater runoff that became associated with dissolved and suspended
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PM phases as well as bedload, material moved by rolling, sliding, and saltating along the
bottom of a stream (Birch and McCreadv. 2009).
Several recent studies reported that the size distribution of Pb in PM transported in runoff
is relatively uniform. Characterization of the roadside dust in Australia showed that Pb in
PM was approximately uniformly distributed among PM size fractions of up to 250 (im.
The Pb-containing particles had the potential to be dispersed to some distance into
sensitive ecosystems (Pratt and Lottermoser. 2007). Pb in roadside dusts in Thessaloniki,
Greece was characterized by Ewen et al. (2009) and no difference in Pb concentration
was found between <75 (im and 75-125 (im PM size ranges, although a difference in the
chemical form of Pb between slightly versus highly contaminated areas was observed.
Ewen et al. (2009) reported that Pb was mainly in a more exchangeable form (similar to
that in an old auto-catalyst reference material) in small particles, but in the residual, or
least mobile fraction in larger particles. In urban road dust from Manchester U.K.,
Pb-bearing Fe-oxides were observed to be dominant in most of the size fractions, and
PbCrO4 comprised 8-34% of total Pb with the highest concentrations being found in the
largest and smallest size fractions. Pb(CO3)2 and Pb(OH)2 were measured in the two
middle size fractions, while PbO and PbSO4 were present in the largest and smallest size
fractions (Barrett etal. 2010).
Murakami et al. (2007) also emphasized the importance of PbCrO4 as an important
species of Pb in runoff from road surfaces. That study identified individual particles
containing high levels of Pb and Cr (> 0.2%), most likely from the yellow road line
markings. The identified PM constituted 46% of Pb in heavy traffic dust and 7-28% in
dust from residential areas and soakaway sediments. The presence of such particles in
soakaway sediments is consistent with their low environmental solubility.
Recent research also continues to document the first flush effect described in the
2006 Pb AQCD. Flint and Davis (2007) reported that in 13% of runoff events, more than
50% of Pb was flushed in the first 25% of event water. A second flush occurred less
frequently (4% of runoff events for Pb). In agreement with the 2006 Pb AQCD (U.S.
EPA. 2006b). most recent studies have concluded that, during storm events, Pb is
transported together with large PM. Some studies, however, found that Pb was
concentrated in the fine PM fraction and, occasionally, Pb was found predominantly in
the dissolved fraction. Tuccillo (2006) found that Pb was almost entirely in the >5 (im
size range and, indeed, may be associated with PM larger than 20 (im. (Sansalone et al..
2010) compared Pb-containing PM size distributions from Baton Rouge, LA, New
Orleans, LA; Little Rock, AR; North Little Rock, AR; and Cincinnati, OH and found no
common distribution pattern. Pb was associated with Cincinnati PM mainly in the
<75 (im fractions, at Baton Rouge and Little Rock Pb mainly in the 75-425 (im PM
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fractions, and at North Little Rock Pb predominantly in the >425 (im PM fractions. New
Orleans Pb was almost uniformly distributed among the smaller size PM fractions.
McKenzie et al. (2008) found that Pb was enriched in the finest PM (0.1-0.3 (im) in
stormwater samples collected in California, particularly for storms that occurred during
and after an extended dry period.
Guo et al. (2006a) investigated the effect of engineered partial exfiltration reactor (PER)
systems on the partitioning and speciation of Pb in rainfall-runoff at the upstream end of
an urban source area catchment that is part of the much larger urbanized and industrial
Mill Creek watershed in Hamilton County, Ohio. The catchment is paved to a large
extent with asphalt and is used for transportation. Guo et al. (2006a) investigated a
catchment that drained toward a wide grassy area and found that Pb was mainly
associated with dissolved organic matter (DOM) in the influent. The study suggested that
interaction of the rainfall-runoff with the grassy area may have resulted in removal of
PM-bound Pb. PM amount and size can also be influenced by the runoff surface. Guo et
al. (2006a) found that Pb entering the engineered PER system was mainly in the
dissolved, rather than particulate, fraction of the rainfall runoff with -76%.
There were several recent observations of a relationship between road traffic volume and
runoff Pb concentration, although a clear relationship was not always observed. At a
relatively clean location, Desta et al. (2007) studied highway runoff characteristics in
Ireland and found that although as expected, total Pb concentration was strongly
correlated with total suspended solids, no relationship between total suspended solids and
rainfall, rain intensity, antecedent dry days or runoff event duration were observed. They
concluded that runoff composition from site to site could be highly variable. Most other
studies did find a relationship between traffic volume and Pb concentration. A California
study of highway runoff by Kayhanian et al. (2007) reported that 70-80% Pb was in
particulate form for both non-urban and urban highways, and that the concentration of Pb
in runoff from low traffic flow (30,000-100,000 vehicles/day) urban highways was 50%
higher than that from non-urban highways (total Pb mean =16.6 (ig/L). Additionally, the
concentrations in runoff from high traffic flow (>100,000 vehicles/day) urban areas were
five times higher than those from non-urban highways. Helmreich et al. (2010)
characterized road runoff in Munich, Germany, with an average daily traffic load of
57,000 vehicles. The mean total Pb concentration, 56 (ig/L (maximum value = 405 (ig/L),
lay in between the values for low traffic flow and high traffic flow runoff from urban
areas in California, i.e., there was good agreement with Kayhanian et al. (2007). There
was no detectable dissolved Pb, i.e., 100% in PM form. Seasonal effects of highway
runoff have also been observed recently. Hallberg et al. (2007) found that summer
particle-bound Pb concentrations in runoff water in Stockholm ranged from
1.37-47.5 (ig/L while, in winter, the range was 1.06—296 (ig/L. There was a strong
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correlation between Pb (and most other elements) and total suspended solids (R2 = 0.89).
Helmreich et al. (2010) also found higher metal concentrations during cold seasons in
Stockholm but Pb concentrations increased only slightly during the snowmelt season.
There was no change in the distribution of Pb between dissolved and PM forms for the
rain and snowmelt periods. Runoff from urban snowmelt has been intensively
investigated since the 2006 Pb AQCD was published. The relocation of snow means that
the area receiving the snowmelt is not necessarily the same area that received the
snowfall. Magill and Sansalone (2010) also noted that plowed snowbanks alongside
roadways form a temporary linear reservoir for traffic generated constituents such as
metals and PM. Snowmelt concentrations of metals such as Pb can therefore be several
orders of magnitude higher than those in rainfall runoff (Sansalone and Buchberger.
1996). The melt process usually occurs in a sequence: pavement melt, followed by
roadside (impervious) and finally pervious area melt. As part of this sequence, rain-on-
snow can transport high loads of PM-associated pollutants (Oberts. 2000). Westerlund
and Viklander (2006) investigated differences in PM and Pb concentrations between
rainfall events occurring during snowmelt and rain periods. Runoff events occurring
during the snowmelt period (i.e., rain-on-snow) had about five times higher numbers of
particles (in the size range 4 to 120 (im)/Liter of runoff. The first rain-on-snow event was
characterized by an increase in the number of particles in the 4 to 25 (im size range. The
rain-on-snow gave a "flush" through the snow but this was still not sufficient to transport
the larger sized particles. Only the highest energy rain-on-snow events increased
transport of PM across the entire size spectrum. There was no difference in particle size
distributions between snowmelt and rain on snow events, although more was transported
during snowmelt. Pb concentrations were most strongly associated with the smaller PM
size fractions.
Overall, there was a significant difference between the melt period and the rain period in
terms of concentrations, loads, transportation and association of heavy metals with
particles in different size fractions (Westerlund and Viklander. 2006). Over a 4-year
period, Magill and Sansalone (2010) analyzed the distribution of metal in snow plowed to
the edge of roads in the Lake Tahoe catchment in Nevada, and concluded that metals
including Pb were mainly associated with large particles (179-542 (im). The PM-
associated metal could be readily separated from runoff water (e.g., in urban drainage
systems), but there is potential for leaching of metals from the PM within storage basins
(Ying and Sansalone. 2008). For adsorbed species that form outer sphere complexes, a
decrease in adsorption and an increase in aqueous complexes for pollutant metals is a
likely consequence of higher deicing salt concentrations. If metals form inner-sphere
complexes directly coordinated to adsorbent surfaces, background deicing salt ions would
have less impact. It is thought that physical and outer-sphere complexes predominate for
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coarse PM, as was the case in Nevada, and so leaching would be likely to cause an
increase in dissolved phase Pb concentrations.
Rural runoff has also been extensively studied since publication of the 2006 Pb AQCD
(U.S. EPA. 2006b). including several recent publications on a forested watershed (Lake
Plastic) in central Ontario (Landre etal.. 2010. 2009; Watmough and Dillon. 2007) and
nearby Kawagama Lake, Canada (Shotyk and Krachler. 2010). Results indicated that
bulk deposition substantially decreased to 0.49 mg/m2 in 2002 from 1.30-1.90 mg/m2 in
1989-91. The upland soils retained >95% of the Pb in bulk deposition, i.e., leaching
losses to stream water were small. The wetland area was, however, a net source of Pb
with annual Pb concentrations in stream water ranging from 0.38 to 0.77 (ig/L. Lake
sediments were efficient sinks for atmospherically deposited Pb with 80-91% of the Pb
input being retained. Up to 68% of the Pb entering the lake was derived from the
terrestrial catchment. Overall, the watershed effectively retained atmospherically
deposited Pb, but some Pb was then redistributed from the catchment to the lake
sediments; and the Pb in the near-surface lake sediments reflected terrestrially transported
soil material, rather Pb being deposited from the atmosphere. The highest concentrations
of dissolved organic carbon (DOC), Fe and Pb in the wetland draining stream occurred in
summer when Pb concentration frequently exceeded 1 (ig/L (Landre et al.. 2009).
Graham et al. (2006) observed two temporally separated mechanisms occurring during
storm events in a rural organic rich upland catchment. At the beginning of an event, Pb
was transported together with large particles in the >25 (im size range, but after several
hours Pb was mainly transported with colloidal or DOM (<0.45 (im), and the remaining
30-40% of storm related Pb was transported in this form. This indicated that overland
flow rapidly transported Pb-PM into the receiving streams at the very beginning of the
event, and this was followed within a few hours by transport of organic-colloidal Pb via
near-surface throughflow. The authors used a conservative estimate of Pb removal, based
on their observations that the catchment was continuing to act as a sink for Pb. These
observations about the transport and fate of Pb agree well with those of Watmough and
Dillon (2007) and Shotyk et al. (2010).
Soil type was also found to have a strong influence on runoff contributions. Dawson et al.
(2010) found that for organic-rich soils Pb was mobilized from near-surface soils together
with DOC, but for more minerogenic soils percolation of water allowed Pb bound to
DOC to be retained in mineral horizons and combine with other groundwater sources.
The resulting Pb in stream water had a more geogenic signature (Dawson et al.. 2010).
The findings of both Graham et al. (2006) and Dawson et al. (2010) were important
because the provenance and transport mechanisms of Pb may greatly affect the net export
to receiving waters.
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In another study, Rothwell et al. (2007b) observed stormflow Pb concentrations in a peat
catchment in Southern Pennines, U.K. almost three times higher than those reported by
Graham et al. (2006) for northeastern Scotland. The generally high dissolved Pb were due
to high soil Pb pools and high stream water DOC concentrations (Rothwell et al.. 2007b).
In a separate study, Rothwell et al. (2007a) showed that OM was the main vector for Pb
transport in the fluvial system. Some seasonal variability was observed: declining Pb
concentrations in autumn stormflow may indicate the exhaustion of DOC from the
acrotelm (the hydrologically active upper layer of peat which is subject to a fluctuating
water table and is generally aerobic) or a dilution effect from an increasing importance of
overland flow.
Erosion of agricultural soils and the effects of different types of storm events on soil
particle and Pb losses from these soils was characterized by Quinton and Cart (2007). A
close link between metal concentration and the silt, or clay and organic content of stream
sediments was consistent with enrichment of metals as a consequence of small erosion
events. They also noted that short intense events could produce the same amount of
sediment as longer low-intensity events. More intense events, however, could mobilize a
wider range of particle sizes whereas low intensity events mobilized finer but more
metal-rich material. Smaller events accounted for 52% of Pb losses from the agricultural
soil.
The Tinto River in Spain drains one of the largest polymetallic massive sulfide regions in
the world: the Iberian Pyrite Belt. Evaporitic sulfate salts, formed as a result of acid mine
drainage processes, are considered to be a temporary sink for many heavy metals. Upon
the arrival of rainfall, however, they rapidly dissolve, releasing acidity and contaminant
metals into receiving waters. Thus rivers in semi-arid climate regions such as the Tinto
River which alternate between long periods of drought and short but intense rainfall
events, can experience quick acidification and increases in metal concentration. In a study
of such events, Canovas et al. (2010) found that while many element concentrations
decreased during events, the concentrations of Fe, Cr, Pb and As increased. This was
attributed to the redissolution and transformation of Fe oxyhydroxysulfates and/or
desorption processes.
Dunlap et al. (2008) studied a large (>160,000 km2) riparian system (the Sacramento
River, CA) and showed that the present day flux of Pb into the river was dominated by Pb
from historical anthropogenic sources, which included a mixture of high-ratio hydraulic
Au mining-derived Pb and persistent historically-derived Pb from leaded on-road
gasoline. Outside of the mining region, 57-67% of Pb in suspended colloidal material was
derived from past on-road gasoline emissions and 33-43% was from hydraulic Au mining
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sediment. Periods of high surface runoff mobilize additional fluxes of Pb from these two
sources and carry them into the river.
Rothwell et al. (2007b) commented that although there have been substantial reductions
in sulfur deposition to U.K. uplands over the last few decades (Fowler et al., 2005).
anthropogenic acidification of upland waters is still possible if there is nitrogen leaching
from the surrounding catchment and this may increase with nitrogen saturation (Curtis et
al.. 2005). Rothwell et al. (2007b) predicted that if an increase in surface water
acidification is coupled with further increases in DOC export from organic-rich
catchments, metal export from peatland systems will increase. The deterioration of peat
soils by erosion is considered to be exacerbated by climatic change. Rothwell et al.
(2010) used digital terrain analysis to model suspended Pb concentrations in
contaminated peatland catchments. The peat soils of the Peak District are characterized
by extensive eroding gullies and so they were combined in an empirical relationship
between sediment-associated Pb concentrations and mean upslope gully depth with fine-
resolution mapping of the gully areas. This model will enable prediction of metal
contamination in receiving waters.
Klaminder et al. (2010) investigated the environmental recovery of sub-arctic lakes in
response to reduced atmospheric deposition over the last few decades. They found that
there had been no improvement in surface sediments and indeed the reduction in Pb
contamination had been much less than the 90% reduction in emissions over the last four
decades. The weak improvement in the 206Pb/207Pb ratio together with the Pb contaminant
concentrations suggests that catchment export processes of previously-deposited
atmospheric contaminants have had a considerable impact on the recent contaminant
burden of sub-arctic lakes. In Arctic regions, soil export of contaminants to surface
waters may dramatically increase in response to climate change if it triggers thawing of
frozen soil layers. It is thought that thawing may generate accelerated soil erosion, altered
hydrological flow paths, increased runoff and exposure of soluble compounds that had
previously been in the frozen layers. At this stage, however, the links between catchment
export and climate change have not yet been clearly established.
Coynel et al. (2007) also considered the effects of climate change on heavy metal
transport. In this case, the scenario of flood-related transport of PM in the Garonne-
Gironde fluvial-estuarine system was investigated. Export of suspended PM during a
five-day flood in December 2003 was estimated at -440,000 tons, accounting for -75%
of the annual suspended PM fluxes. Sediment remobilization accounted for -42% of the
total suspended particulate matter (SPM) flux during the flood event (-185,000 tons
suspended PM) and accounted for 61% of the 51 tons Pb that was exported. Coynel et al.
(2007) postulate that flood hazards and transport of highly polluted sediment may
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increase as a result of climate change and/or other anthropogenic impacts (flood
management, reservoir removal).
In heavily contaminated catchments [e.g., that of the Litavka River, Czech Republic (Zak
et al., 2009)1. the flux of heavy metals to the river during storm events can be substantial.
Even during a minor 4-day event, 2,954 kg of Pb was transported, and the majority was
associated with suspended PM. For the Adour River in a mountainous area of France, Pb
pollution predominantly originated from mining activities, and Point et al. (2007) showed
that 75% of annual soil fluxes into the river were transported in 30-40 days.
The consequences of flood management (dam flushing) practices on suspended PM and
heavy metal fluxes in a fluvial-estuarine system (Garonne-Gironde, France) were
considered by Coynel et al. (2007). Dam flushing enhanced mobilization of up to
30-year-old polluted sediment from reservoir lakes. Sediment remobilization accounted
for -42% of the total suspended PM fluxes during the flood and strongly contributed to
PM-bound metal transport (61% for Pb). They concluded that flood management will
need to be taken into consideration in future models for erosion and pollutant transport.
Bur et al. (2009) investigated the associations of Pb in stream-bed sediments of the
French Gascony region. They found that Pb enrichment in stream sediments was
positively correlated with catchment cover and increasing organic content whereas Pb
concentration was strongly linked with Fe-oxide content in cultivated catchments. For the
low-OM, anthropogenic Pb was associated with carbonates and Fe-oxides (preferentially,
the amorphous fraction). Fe-oxides became the most efficient anthropogenic Pb trapping
component as soon as the carbonate content is reduced. They noted, however, that OM
was always weakly involved. N'Guessan et al. (2009) also studied trace elements in
stream-bed sediments of the French Gascony region. They used enrichment factors to
show that only -20-22% of Pb was from anthropogenic sources with the remainder
originating from natural weathering processes.
Overall, research results from the last several years have greatly expanded the extent of
the knowledge concerning Pb from runoff. Substantial Pb input to estuarine and marine
ecosystems has been well documented. More detail concerning the origin of Pb from roof
runoff has led to the conclusion that roof flashing could be especially important. Research
on road runoff has provided valuable insight into the size and composition of suspended
sediment particles, indicating that size distributions for Pb-containing PM in runoff water
vary from study to study and from location to location. Recent studies confirmed the
"first flush" effect, releasing more Pb at the beginning of rainfall than subsequently, and
documented size distributions of Pb-containing PM also vary considerably when water
from the first flush is isolated. Influence of road traffic volume on runoff has also been
more fully documented in recent years. The role of urban snowmelt and rain-on-snow
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events is also better understood, and it has been observed that greater runoff occurs from
snowmelt and in rain-on-snow events than when snow is not present, and that metals,
including Pb, are often associated with coarse PM under these circumstances. Runoff in
rural areas is strongly controlled by soil type and the presence of vegetation with less
runoff and greater retention in mineral soils. When grass is present, there is more runoff
for soils high in OM. Runoff also follows atwo-step process of transport of larger
particles at the beginning of an event, followed within hours by transport of finer
colloidal material. Some initial research on the effects of climate change on runoff has
focused on documenting the association between increased runoff and more intense rain
events and greater thawing. Overall, recent research has provided greater detail on
amounts, particle size distributions, composition, and important processes involving Pb
transport. The understanding of Pb runoff has become more complete since publication of
the 2006 Pb AQCD (U.S. EPA, 2006b) (U.S. EPA. 2006R
2.3.3 Soil
The 2006 Pb AQCD (U.S. EPA. 2006b) summarized that Pb has a relatively long
retention time in the organic soil horizon, although its movement through the soil column
also suggests potential for contamination of groundwater. Leaching was consistently
observed to be a slower process for Pb than for other contaminants because Pb was only
weakly soluble in pore water, but anthropogenic Pb is more available for leaching than
natural Pb in soil. Pb can bind to many different surfaces and Pb sorption capacity is
influenced by hydraulic conductivity, solid composition, OM content, clay mineral
content, microbial activity, plant root channels, animal holes, and geochemical reactions.
As a result of Pb binding to soil components, leaching is retarded by partitioning to soils,
which is not only influenced by sorption capacity, but leaching also increases with
proximity to source, increasing pH, and increasing metal concentrations. Leaching is also
strongly influenced by pore water flow rates, with more complete sorption contributing to
slower leaching at lighter flows. Leaching rates are especially high in soils with a high Cl
content, but typically the most labile Pb fraction is adsorbed to colloidal particles that
include OM, clay, and carbonates. Transport through soils is enhanced by increasing the
amount of colloidal suspensions, increasing colloidal surface charge, increasing organic
content of colloids, increasing colloidal macroporosity, and decreasing colloidal size.
Acidity and alkalinity have a more complex influence, with sorption maximized at
neutral pH between pH = 5 and pH = 8.2, and greater mobility at higher and lower pH.
High Pb levels have been observed in leachates from some contaminated soils, but this
effect appears to be pH dependent. In several studies of contaminated soils a substantial
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fraction of Pb was associated with Mn and Fe oxides or carbonate. Influence of soil
chemistry on Pb effects in ecosystems is described in Section 6.3.2.
2.3.3.1 Deposition of Pb onto Soil from Air
As described in the 2006 Pb AQCD (U.S. EPA. 2006b). a considerable amount of Pb has
been deposited from air onto soils in urban areas and near stationary sources and mines.
Removal and translocation of Pb in soil is an ongoing process.
High geometric mean Pb soil concentrations (for soil particles between 180 nm and 2 mm
in size: 51 ppm; for soil particles smallerthan 180 urn in size: 105 ppm) were observed
near stationary sources such as smelters and battery disposal operations, and soil Pb
concentrations decreased rapidly with distance from the source. Several recent studies
continue to document high concentrations of Pb in soil. A study of soil Pb concentrations
in Queensland, Australia described atmospheric transport and deposition of Pb in soils
(due to ongoing emissions from nearby mining [which began in 1924] and smelting
[which began in 1931] activities) are continuing to impact the urban environment,
resulting in elevated soil Pb concentrations at urban property sites within 2 km of the
mines (Taylor et al.. 2010). Similarly, sediment cores from four remote Canadian Shield
headwater lakes located along a transect extending 300 km from a non-ferrous metal
smelter generated useful information about distance of Pb transport from the smelter prior
to deposition (Gallon et al., 2006). Shotyk and Krachler (2010) postulated that long-range
transport of Pb from a smelter at Rouyn-Noranda may still contribute to deposition on
these lakes. Recent measurements of deposition fluxes to soil in rural and remote areas
have ranged from approximately 0.5 mg/m2 per year to about 3 mg/m2 per year with fair
agreement between locations in Canada, Scandinavia, and Scotland. These measurements
showed a substantial decrease compared to when leaded on-road gasoline was in
widespread use (Shotbolt et al.. 2008; Watmough and Dillon. 2007; Fowler etal.. 2006;
Graham et al.. 2006).
There has been considerable interest in the response of soils to the decreasing aerosol Pb
concentrations and Pb deposition rates that have been recorded in recent years. Kaste et
al. (2006) resampled soils at 26 locations in the Northeast U.S. (during a 2001-2002
survey of soil sites originally sampled in 1980), and found no significant change in the
amount of Pb in the O-horizon at high altitude sites, suggested to be related to reduced
microbial activity at altitude. However, the amount of Pb in the O-horizon had decreased
at some locations in the southern part of the survey region (Connecticut, New York,
Pennsylvania), where the forest soils have typically thinner O-horizons, the reasons for
which are discussed further in Section 2.3.3.2. Relatively higher Pb concentrations were
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also found among sampled areas of Japan with differing geology (Takamatsu et al.,
2010).
Further support for the use of mosses as bioindicators or monitors for atmospheric Pb
inputs to peat bogs have recently been published by Kempter et al. (2010) who found that
high moss productivity did not cause a dilution of Pb concentrations in peat bogs. They
also found that productive plants were able to accumulate particles from the air and that
rates of net Pb accumulation by the mosses were in excellent agreement with the
atmospheric fluxes obtained by direct atmospheric measurements at nearby monitoring
stations. In addition, Bindler et al. (2008) used Pb isotopes to compare the distribution of
Pb in the forest soils with that of lake sediments where no "plant pumping" processes
could be invoked, and used Pb isotope ratios to demonstrate that observations were
consistent with anthropogenic Pb deposition to the soils rather than intermixing of natural
Pb from underlying mineral soil horizons.
Overall, recent studies provided deposition data that were consistent with deposition
fluxes reported in the 2006 Pb AQCD (U.S. EPA. 2006b). and demonstrated consistently
that Pb deposition to soils has decreased since the phase-out of leaded on-road gasoline.
Follow-up studies in several locations at high altitude indicated little change in soil Pb
concentrations since the phase-out of leaded on-road gasoline, although reductions in
surface soil Pb concentrations have been documented in some areas.
2.3.3.2 Sequestration of Pb from Water to Soil
The 2006 Pb AQCD described Pb as being more strongly retained in soil than other
metals because of its weak solubility in pore water, but that anthropogenic Pb was more
available for leaching than natural Pb (U.S. EPA. 2006b). It also described a complex
variety of factors that influence Pb retention, including hydraulic conductivity, solid
composition, OM content, clay mineral content, microbial activity, plant root channels,
geochemical reactions, colloid amounts, colloidal surface charge, and pH.
Recent research in this area has provided more insight into the details of the Pb
sequestration process. Importance of leaf litter was further investigated, and it was
observed that the absolute Pb content can be substantial because rain events cause
splashing of the leaf litter with soil thus placing the litter in direct contact with soil
metals. The resulting increase in leaf litter metal concentrations suggests that the litter
can act as a temporary sink for metals from the soil around and below leaves on the
ground. The low solubility of Pb in the leaf litter indicates that the Pb is tightly bound to
the decomposing litter, making the decomposing leaves act as an efficient metal storage
pool (Scheid et al., 2009). Differences between throughfall (i.e., water depositing onto the
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soil following collection on leaves) and litterfall (i.e., deposition of leaves, bark, and
other vegetative debris onto soil) in forested areas have been investigated in forested
areas, and the combined input of Pb to the forest floor from throughfall and litterfall was
approximately twice that measured in bulk deposition (Landre et al.. 2010). The
difference was attributed to a substantial contribution from internal forest cycling and
indicates that bulk deposition collectors may underestimate the amount of Pb reaching the
forest floor by about 50% (Landre etaL 2010).
New research has also provided details about the complexity of Pb sequestration during
soil OM decomposition. Schroth et al. (2008) investigated Pb sequestration in the surface
layer of forest soils and the transformation of Pb speciation during soil OM
decomposition. The pH range for forest floor soils in the Northeast U.S. is typically 3.5-5
and, under these conditions, dissolved Pb would adsorb strongly to soluble OM and to
Fe/Al/Mn oxides and oxyhydroxides. It had been thought that the high affinity of Pb for
organic ligands meant that sequestered atmospheric Pb would be preferentially bound to
soluble OM. As a consequence, decomposition of OM would lead to Pb migration to the
underlying mineral layers where it would be precipitated with the dissolved OC or
adsorbed to pedogenic mineral phases. However, recent research has revealed a more
complicated picture of gasoline-derived Pb associations in the forest floor. More recent
research indicates that, as decomposition progresses, Pb and Fe become more
concentrated in "hotspots" and Pb likely becomes increasingly distributed on surfaces
associated with Fe and Mn (and to some extent Ca). It was postulated that Pb was
initially bound to labile organic sites but, following decomposition, the Pb was adsorbed
at reactive sites on pedogenic mineral phases (Schroth et al., 2008). Differences in litter
types were also reported, with more rapid decomposition of OM in deciduous litter
mobilizing more Pb initially bound to labile OM than coniferous litter, and producing
more pedogenic minerals that could readily sequester the released Pb (Schroth et al..
2008). In the next stage of the study, the speciation of Pb in the O-horizon soils of
Northern Hardwood, Norway spruce and red pine forest soils were compared. In general
there was good agreement between the Pb speciation results for the soils and those for the
laboratory decomposition experiments. Specifically, for the Northern Hardwood forest
soil, a little more than 60% of the Pb was bound to soil organic matter (SOM) and this
percentage increased to -70% and -80% for the Norway spruce and red pine soils,
respectively. In all three cases, however, most of the remainder of the Pb was bound to
ferrihydrite rather than to birnessite. This was not considered to be surprising because of
the well-known leaching and cycling behavior of Mn that would be expected in the
natural system. Thus the prevalence of Mn phases in the field based samples would be
lessened (Schroth et al.. 2008).
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More generally, other studies have observed Pb sorption to Mn and Fe phases in soils.
For example, Boonfueng et al. (2006) investigated Pb sequestration on Mn oxide-coated
montmorillonite. Pb formed bidentate corner-sharing complexes. It was found that Pb
sorption to MnO2 occurred even when MnO2 was present as a coating on other minerals,
e.g., montmorillonite. Although their importance in the near-surface phases has clearly
been demonstrated by Schroth et al. (2008). ferrihydrite surfaces may not be a long-term
sink for Pb since reductive dissolution of this Fe(III) phase may release the surface-bound
Pb into the soil solution. Sturm et al. (2008) explored the fate of Pb during dissimilatory
Fe reduction. Pb was indeed released but was then incorporated into less reactive phases.
These phases could not, however, be identified. Even so, it was asserted that Pb should be
largely immobile under Fe-reducing conditions due to its incorporation into refractory
secondary minerals.
Kaste et al. (2006) found that Pb species currently in the soil O-horizons of the Northeast
U.S. differed considerably from those that were originally deposited from fossil fuel
combustion (including on-road gasoline). PbSO4 was considered to be the main form of
Pb that had been delivered from the atmosphere to the surface of the Earth and it was
postulated that the presence of sulfate may have facilitated the adsorption of Pb to
colloidal Fe phases within the organic-rich horizons.
Altogether, these new results enhance the understanding of Pb sequestration in forest
soils. First, the role of leaf litter as a major Pb reservoir is better understood. Second, the
effect of decomposition on Pb distribution and sequestration on minerals during OM
decomposition has been further characterized, and finally, the relative importance of Mn
and Fe in sequestration is better understood.
Recent research also addressed roadside soils. Jensen et al. (2006) found that Pb was
retained by an organic-rich blackish deposit with a high OM content and elevated soil Pb
concentrations, originating from total suspended solids in road runoff and from aerial
deposition. Hossain et al. (2007) observed that after long dry periods, OM oxidation may
potentially result in the release of Pb. Microbial activity may also breakdown OM and
have similar consequences (i.e., Pb release). Bouvet et al. (2007) investigated the effect
of pH on retention of Pb by roadside soils where municipal solid waste incineration
(MSWI) bottom ash had been used for road construction. They found that the Pb that had
leached from the road construction materials was retained by the proximal soils under the
prevailing environmental conditions (at pH = 7, <2% was released, but at pH = 4, slightly
more Pb (4-47%) was released) and the authors speculated that the phase from which Pb
had been released may have been Pb(CO3)2(OH)2, indicating that sequestration of Pb via
formation of oxycarbonate minerals is only effective at near-neutral to alkaline pH values
(Figure 2-12 in Section 2.3.3.3).
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Other recent research on Pb sequestration focused on microbial impacts and soil
amendments. There have been few if any previous observations of microbial
sequestration of Pb in soil. Perdrial et al. (2008) observed bacterial Pb sequestration and
proposed a mechanism of Pb complexation by polyphosphate. They also postulated that
bacterial transport of Pb could be important in sub-surface soil environments. Wu et al.
(2006) also concluded that Pb adsorption to the bacterial cell walls may be important with
respect to Pb transport in soils. This new area of research suggests that bacteria can play
an important role in both sequestration and transport of Pb. Phosphate addition to
immobilize Pb-contaminated soils has often been used to immobilize Pb in situ through
the formation of Pb phosphate minerals such as chloropyromorphite. Recent research
investigated factors affecting the long-term stability of such products, which depends on
the equilibrium solubility and the dissolution rate of the mineral, trace impurities, such as
Pb(OH)2, the presence of complexing agents, and pH (Xie and Giammar. 2007). Overall,
in agreement with the 2006 Pb AQCD (U.S. EPA. 2006b). the addition of phosphate can
enhance immobilization of Pb under certain conditions in the field but may cause
desorption and mobilization of anionic species of As, Cr and Se.
2.3.3.3 Movement of Pb within the Soil Column
The 2006 Pb AQCD summarized studies that demonstrated that Pb has a long retention
time in the organic soil horizon; it also has some capacity to leach through the soil
column and contaminate groundwater more than other contaminants do, because Pb is
only weakly soluble in pore water (U.S. EPA. 2006b). The fate of any metal transport in
soil is in response to a complex set of parameters including soil texture, mineralogy, pH
and redox potential, hydraulic conductivity, abundance of OM and oxyhydroxides of Al,
Fe, and Mn, in addition to climate, situation and nature of the parent material. As a
consequence, it is impossible to make general conclusions about the final fate of
anthropogenic Pb in soils. Indeed, Shotyk and LeRoux (2005) contend that the fate of Pb
in soils may have to be evaluated on the basis of soil type. Some generalizations are,
however, possible: Pb migration is likely to be greater under acidic soil conditions
(Shotyk and Le Roux. 2005). In this respect, it would be expected that there should be
considerable mobility of Pb in the surface layers of certain types of forest soils. This
section reviews recent research on movement of Pb through soil types by first focusing
on forest soils, followed by a broader treatment of a more diverse range of soils.
2-54
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Forest Soils and Wetlands
Several studies confirmed the slow downward movement of Pb within the soil column.
Kaste et al. (2006) found that the amount of Pb in O-horizon soils had remained constant
at 15 of 26 sites in remote forested areas of the Northeast U.S. that had been re-sampled
after a 21-year time period had elapsed, but that measured soil Pb concentrations were
lower than predicted concentrations from total deposition, strongly suggesting that the
O-horizon had not retained all of the atmospheric Pb, and that a proportion of the
atmospheric deposition must have leached into the underlying mineral layers. At some
sites, mainly those at the southern latitudes and lower altitude sites, the proportion of Pb
that had been leached downward from the O-horizon was quite considerable. Relative
retention of Pb was influenced by the rate of OM decomposition, depth of soil O-horizon,
and pH. For soils where Pb was strongly retained by the O-horizon, a relationship
between Pb and Fe-rich phase was observed, but Pb was also significantly correlated with
other metals. XANES data suggested a possible interaction with an amorphous Fe oxide,
but spectra were not entirely explained by Fe and oxygen and an additional spectral
feature suggested the presence of a sulfur (S) or phosphorus (P) atom, which could result
if OM functional groups were binding to Pb. Kaste et al. (2006) concluded that a
substantial fraction of Pb was associated with amorphous Fe-hydroxides. The strong
binding of Pb coupled with the low solubility of Fe phases under oxic conditions, helped
to explain the relatively long residence time of gasoline-derived Pb in forest floors which
had thick O-horizons and were well-drained. In the situations where Pb was leached
downward to a large extent, mobility was likely governed by OM decomposition and
colloidal transport of Pb associated with colloidal Fe and OM.
Klaminder et al. (2006b) also considered the transfer of Pb from the O-horizon to the
underlying mineral horizons (including the C-horizon). They concluded that atmospheric
pollution-derived Pb migrated at a rate about 10-1,000 times slower than water. They
assumed that Pb was mainly transported by dissolved OM and so the mean-residence-
time of Pb in the O-horizon depended on OM transport and turnover. The retardation rate
was a reflection of the slow mineralization and slow downward transport rates of organic-
Pb complexes, due to sorption and desorption reactions involving mineral surfaces.
In a study involving stable Pb isotopes, Bindler et al. (2008) showed that Pb with a
different isotopic composition could be detected in the soil down to a depth of at least
30 cm and sometimes down to 80 cm in Swedish soils. In comparison, in North American
podzols, pollution Pb is typically only identified to a depth of 10-20 cm (even with the
aid of isotopes). This difference is attributed to the longer history of metal pollution in
Europe (as has been traced using lake sediments).
2-55
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Several research groups have attempted to determine the mean residence time of Pb in the
O-horizon of forest soils. Klaminder et al. (2006a) used three independent methods to
estimate a mean residence time of about 250 years for Pb in the O-horizon of boreal
forests in Sweden, indicating that deposited atmospheric Pb pollution is stored in the
near-surface layers for a considerable period and, consequently, will respond only slowly
to the reduction in atmospheric inputs. It should be noted, however, the OM in the upper
parts of the O-horizon is continually being replaced by fresh litter and the mean residence
time of Pb in these horizons is only 1-2 years. Thus, the uppermost layer will respond
more quickly than the rest of the O-horizon to the decreases in Pb inputs.
Klaminder et al. (2008a) considered the biogeochemical behavior of atmospherically
derived Pb in boreal forest soils in Sweden (Figure 2-11). The estimated annual losses via
percolating soil water were ~2 mg/m2 per year (Klaminder et al.. 2008a) and so the
annual loss, assumed to be from the mor layer, the well-defined layer of acidic humus
that forms within the O-horizon in cool and moist soils, was greater than the atmospheric
input of ~0.5 mg/m2 per year. The upward transport of Pb did not compensate for the
losses either. In contrast, the amount of Pb being stored in the mineral soil layers was
increasing. The mean residence time of Pb in the mor layer was estimated to be
-300 years, in reasonable agreement with their earlier work (Klaminder et al.. 2006a).
These values were greater than the values of 2-150 years determined for U.S. forest soils,
e.g., CWatmough et al.. 2004; Kaste et al.. 2003) but the difference was attributed to the
lower decomposition rates of OM within the northern boreal forests of Sweden. They
concluded that more research was needed to determine the processes occurring within the
mor layer that control the release of Pb from this horizon.
Klaminder et al. (2008b) investigated in more detail the distribution and isotopic
signature of Pb persisting within the O-horizon (mor layer) of boreal forest soils. They
found that the mor layer preserved a record of past Pb emissions from a nearby smelter.
Minimal animal burrowing activity and low leaching rates observed at the sampling
location were important factors contributing to the preservation of this record. They
concluded that temporal changes in atmospheric fallout in addition to adsorption
processes need to be considered when interpreting Pb concentration changes within the
mor layer.
2-56
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0,05
0.05
Bs
B/C
Loss of lead
Buildup of
lead
0.1
i—> Present atmospheric deposition (mg m-J yr1)
- - - * Plant uptake (mg rrr2 yr1)
~* Soil water flux (mgm^yr1)
Notes: The atmospheric deposition rate is from (Klaminder et al.. 2006a). the plant uptake rates from (Klaminderet al.. 2005) and
estimated soil-water fluxes from (Klaminder et al.. 2006b).
Figure 2-11 Schematic model summarizing the estimated flux of Pb within a
typical podzol profile from northern Sweden using data from
Klaminder et al. (2006a) and Klaminder et al. (2006b).
Significantly higher O-horizon Pb concentrations have been observed in coniferous than
deciduous forest soils (McGee et al.. 2007). Steinnes et al. (2005) noted evidence for
downward migration of Pb from the O-horizon to the E-horizon of most soils and in some
cases the upper B-horizon. They found that the downward transport of Pb differed
considerably between the sites, e.g., from almost no anthropogenic Pb in the B-horizon at
some sites to -70% at other sites. The greater downwards transport in some locations was
attributed to climatic variations, with more extensive leaching and possibly a greater
turnover of OM at sites where higher mean annual temperatures were experienced.
Higher atmospheric deposition of acidifying substances in these locations was considered
the most important factor in Pb transport, causing release of Pb from exchange sites in the
humus layer and promoting downward leaching.
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Seasonal variation in Pb mobility has also been observed in forest soil. Other research
indicated that Pb concentrations correlated with DOC concentrations in the soil solution
from the O-horizon, and were lower during late winter and spring compared with summer
months (Landre et al. 2009). The degradation of OM in the O-horizon produced high
DOC concentrations in the soil solution. It was also shown that Pb was associated with
the DOC, and concluded that DOC production is a primary factor enhancing metal
mobility in this horizon. In the underlying mineral horizons, DOC concentrations
declined due to adsorption and cation exchange processes. The B-horizon retained most
of the DOC leached from the O-horizon and it has also been observed that Pb is similarly
retained.
Non-forested Soils
In contrast with forest soils, most non-forested soils are less acidic and so most studies of
Pb behavior in non-forested soils have focused on Pb immobility. However, there are
acid soils in some locations that are not forested. For these soils, as for forest soils, Pb
mobility is weak but correlated with OM. For example, Schwab et al. (2008) observed
that low molecular weight organic acids added to soil enhanced Pb movement only
slightly. Citric acid and tartaric acid enhanced Pb transport to the greatest degree but the
extent of mobilization was only slightly higher than that attained using deionized water
even at high concentrations. While the formation of stable solution complexes and more
acidic conditions favored mobilization of Zn and Cd, Pb remained strongly sorbed to soil
particles and so the presence of complexing agents and low pH (2.8-3.8) did not
substantially enhance Pb mobility. Similarly, limited penetration and leaching was
observed in an extremely complex temperate soil profile, with highest concentrations of
Pb (-80 mg/kg) found in the top 0-5 cm section of soil. For this uppermost soil section,
there was a strong correlation between Pb concentration and OC content, both for the
total soil fraction and the acid-extractable fraction. The Pb migration rate was calculated
to be 0.01 cm/year. It was estimated that Pb would be retained in the soil column for
20,000 years, with no evidence of rapid movement of anthropogenic Pb from the top
0-5 cm soil section into the soil profile Kylander et al. (2008).
Other recent studies also reported strong retention in non-forest soils and enhanced
mobility of Fe and OM colloids. Pb was strongly retained on acidic Mediterranean soil
columns, and association of Pb with the exchangeable OM and crystalline Fe oxide
fractions appeared to favor mobility, while Pb association with Mn oxides and
amorphous Fe oxides was linked with semi-irreversible retention of Pb in the solid phase
(Garrido et al.. 2008). In another study of Pb mobility within Mediterranean soils, Pb
infiltration velocity was measured to be 0.005 meter/year (Erel. 1998). The authors
attributed Pb movement within the soil column to advection and concluded that the soil
2-58
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profile of Pb is similar to the anthropogenic air Pb emissions trend. Pedrot et al. (2008)
studied colloid-mediated trace element release at the soil/water interface and showed that
Pb was mobilized by Fe nanoparticles that were bound to humic acids.
Soil pH value is probably the single most important factor affecting solubility, mobility
and phytoavailability but reducing conditions also results in increased Pb mobility, with
the release of Pb into an anoxic soil solution due to the combined effect of Fe(III)
reductive dissolution and dissolved OM release. Dissolved OM is more important than Fe
oxyhydroxides in determining Pb mobility. Under oxic conditions, Fe-Mn-hydroxides
often play an important role in the sorption of Pb to the solid phase soil (Schulz-Zunkel
and Krueger. 2009). In an agricultural soil, fate of Pb in soils is related to agricultural
management. Although Pb was found to be strongly sorbed to the soil, downward
migration was observed and the movement of Pb to deeper soils was due to the soil
mixing activities of earthworms (Fernandez et al., 2007). Thus in relatively unpolluted
non-forested soils, as in forested soils, colloidal Fe and OM, pH, and biophysical
transport all enhance Pb mobility in soil. Pb transport in more highly contaminated soils
has also been the subject of recent research. In a vegetated roadside soil, Pb was leached
from the upper 50 cm of the soil even though the pH was 7.2. Pb was transported on
mobile particles and colloids in the soil solution. Some of the colloids may have formed
from OM produced by roots and decaying shoots. The transport process was enhanced by
preferential flow triggered by intense rainfall events. This study suggested that the value
of the effective sorption coefficient estimated under dynamic conditions was unrelated to
values measured in conventional batch studies. This indicates that the use of batch studies
to derive input values for sorption coefficients in transport models requires caution. It
was concluded that the primary control of Pb transport in the long term was the degree of
preferential flow in the system (Roulier et al., 2008b).
Other studies also noted similarly low Pb mobility, but with substantial variation between
soil types and locations. A decline in O-horizon Pb concentrations and Pb accumulation
in mineral horizons was also observed for forest soils by Watmough and Dillon (2007),
but did not hold for nearby wetland areas from which a large amount of DOC is exported,
with approximately 10 times more Pb being associated with a given amount of DOC in
the leachate from the LFH-horizon of the wetland soil than with the DOC in the stream
water draining the wetland. This may reflect greater retention of Pb by the wetland and/or
a change in structure of DOC leading to a change in complexing capacity possibly
because of changes in pH or competition with Al and Fe.
Williams et al. (2006) characterized Pb speciation in a mine waste-derived fertilizer,
ironite. It was thought that PbS would be the main form of Pb, but instead was the
predominant form was PbSO4, which may move more easily through soil and enter
2-59
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proximal waters. In contrast, Courtin-Nomade et al. (2008) showed that Pb was
incorporated into barite rather than goethite in waste rock pile materials. The high-
stability phase formed was an anglesite-barite solid solution.
In weathering flotation residues of a Zn-Pb sulfide mine, more Pb was mobile in
weathered topsoil than in the unweathered subsoil (Schuwirth et al.. 2007). The topsoil
had a very high OM content and the Pb enrichment was attributed to an interaction with
soil OM. Overall, the results contrast strongly with most other studies but the
interpretation was supported by the sequential extraction results which showed that there
was a very large exchangeable Pb component in these surface soils. Scheetz and Rimstidt
(2009) characterized shooting range soils in Jefferson National Forest, VA, in which the
metallic Pb shot rapidly became corroded and developed a coating of hydrocerussite,
which dissolved at the pH values of 8-9; see Figure 2-12. which shows an Eh-pH diagram
indicating the solubility, equilibrium, and stability of these corroded Pb molecules in
terms of the activity of hydrogen ions (pH) versus the activity of electrons (Eh [in volts]).
The solubilized Pb was largely re-adsorbed by the Fe and Mn oxides and carbonate soil
fractions. The minimum solubility of hydrocerussite lies in the pH range 8-9 but
solubility increases by several orders of magnitude at pH below 6 (Scheetz and Rimstidt
2009).
Pb3(C03)2(OH)2
Hydrocerussite
0 1 2 3 4 5 6 7 8 9 10 11 12 13 14
Source: Reprinted with permission of Elsevier Publishing, Scheetz and Rimstidt (2009)
Figure 2-12 Eh-pH diagram for Pb in shooting range soils, Jefferson National
Forest, VA.
2-60
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Rooney et al. (2007) also investigated the controls on Pb solubility in soils contaminated
with Pb shot. Again, corrosion crusts were found to develop on Pb pellets. The
concentrations of Pb in the soil solution were, however, much lower than if they were
controlled by the solubility of the dominant crustal Pb compounds (mainly
hydrocerussite). Instead it was suggested that the concentrations were being controlled by
sorption of Pb by the soil solid phase. The pH range in this study was 4.5-6.5 and so
again dissolution of hydrocerussite would be expected. Sorption to solid phases in the soil
is also consistent with the findings of Scheetz and Rimstidt (2009). Overall, in contrast to
less polluted forested and non-forested soils, considerable mobility was often, but not
always observed in soils near roadways and mines and on shooting ranges, with colloid
transport and soil pH playing an important role in Pb mobility. Although there have been
steep declines in Pb deposition, surface soils in have been slow to recover (Bindler et al..
2008: Kaste et al.. 2006). As was concluded in the 2006 Pb AQCD (U.S. EPA. 2006b).
soils continue to act as a predominant sink for Pb.
While in some studies the flux of Pb, from the soil through aquatic ecosystems to lakes
has peaked and declined. In other studies, no recovery of lake sediments in response to
emission reductions was observed (Norton. 2007). For example, Klaminder et al. (2010)
has shown that the Pb concentrations in sub-Arctic lake sediments remain unchanged in
recent years, with the lack of recovery linked to the effects of soil warming, which affect
Pb-OM transport from soil to the receiving lake systems. Shotyk and Krachler (2010)
also reported a disconnect between atmospheric deposition and recent changes in Pb
concentration and isotope ratios in the lake sediments. Simulations of future metal
behavior suggest that the more strongly sorbing metals such as Pb will respond to
changes in metal inputs or acidification status only over centuries to millennia (Tipping et
al.. 2006).
Overall, recent research confirms the generally low mobility of Pb in soil. This limited
mobility is strongly dependent on both colloid amount and composition, as well as pH,
and may be greater in some contaminated soils. Mobility is so low that soils continue to
act as a sink for atmospheric Pb even though atmospheric Pb concentrations peaked
several decades ago.
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2.4 Monitoring of Ambient Pb
2.4.1 Measurement Techniques
2.4.1.1 Sample Collection
Federal Reference Methods
The indicator for the Pb NAAQS is Pb in total suspended particles (Pb-TSP) (73 FR
66964). In order to be used in regulatory decisions judging attainment of the Pb NAAQS,
ambient Pb concentration data must be obtained for this indicator using either the Federal
Reference Method (FRM) or a Federal Equivalent Method (FEM) defined for this
purpose. Accordingly, for enforcement of the air quality standards set forth under the
Clean Air Act, EPA has established provisions in the Code of Federal Regulations under
which analytical methods can be designated as FRM or FEM. Measurements for
determinations of NAAQS compliance must be made with FRMs or FEMs. FRMs and
FEMs for the Pb NAAQS exist for both sample collection and sample analysis.
There are two FRMs for sample collection in the Pb NAAQS monitoring network
(described in Section 2.4.2): (1) Reference Method for the Determination of Lead in
Suspended Particulate Matter Collected From Ambient Air (40 CFR part 50 Appendix
G), and (2) Reference Method for the Determination of Lead in Particulate Matter as
PM10 Collected From Ambient Air (40 CFR part 50, Appendix G). The Pb-TSP FRM
sample collection method is required for all source-oriented NAAQS monitors, and the
FRM for Pb-PMio is accepted for Pb NAAQS monitoring at non-source-oriented
monitors in specified situations.
The Pb-TSP FRM sample collection method specifies use of a high-volume TSP sampler
that meets specified design criteria (40 CFR part 50 Appendix B). Ambient airborne PM
is collected on a glass fiber filter for 24 hours using a high volume air sampler. It has
long been recognized that there is notable variability in high-volume TSP sample
measurements associated with the effects of wind speed and wind direction on collection
efficiency. This variability is predominantly associated with the capture efficiency for
particles larger than 10 (im, but the sampler's size selective performance is known to be
affected by wind speed and direction. For example, at a simulated wind speed of
2-62
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4.6 meters/second, a directional difference of 45 degrees can result in a nearly two-fold
difference in 15 (im particle collection efficiency and a nearly five-fold difference in
50 (im particle collection efficiency (Wedding et al., 1977). Effective D50 (size at 50%
efficiency) was observed to decrease from 50 (im at a 2 km/hour wind speed to 22 (im at
24 km/hour (Rodes and Evans. 1985). Figure 2-13 illustrates the effect of sampler
orientation on collection efficiency as a function of particle size.
00
z
LJ
oc
LJ
LJ
LLJ
uJ
CL
'J-)
60
40
20
U<* = 15 fps , 8 % TURBULENCE
--O MODIFIED I CFM ANDERSEN
-•-T -45"-,
__fc QOJHI-VOLUME SAMPLER
PROTOTYPE DICHOTOMOUS SAMPLER
ORIGINAL 1 CFM ANDERSEN
•D—..
•—D
10
20 30
PARTICLE DIAMETER , (
50
Different TSP sampler types: (1) Modified Andersen Sampler [open circles]; (2) Hi-volume Sampler (for different incident wind
direction (45° [open squares], 0° [closed squares]); (3) Prototype 15 urn Outpoint Dichotomous Sampler [open triangles]; and
(4) Original Andersen Sampler [closed circles].
Source: Reprinted with permission of the American Chemical Society (Wedding et al., 1977)
Figure 2-13 Comparison of particle collection efficiency among different
TSP sampler types.
Some existing commercially available sampler inlets are designed to collect particles
larger than 10 (im with greater than 50% efficiency (Kenny et al.. 2005). and these inlets
can be tested as potential replacements for TSP sampling. Efficient collection of particles
much larger than 10 (im is considerably more challenging because their greater inertia
and higher settling velocities hinder their efficient intake by samplers. The sampling
difficulties and the long history of research to develop adequate sampling technology for
large particles have been thoroughly reviewed (Garland and Nicholson. 1991). High
intake velocities and large inlet openings are necessary to minimize sampling bias for
2-63
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sampling ultra-coarse particles. At this time, no alternative to the FRM TSP sampler has
been identified that has been adequately characterized. As such, there is a continued need
to assess the feasibility of a revised TSP sampler design with improved control on
collection efficiency over a wider range of particle sizes, including ultra-coarse particles
(which are not captured with PM10 samplers).
The spatial scale for which ambient air Pb samples are representative varies depending on
particle sizes present, as discussed further in Section 2.5.3. Concentrations of particles
larger than 10 (im are likely to be very spatially and temporally heterogeneous, with
higher concentrations in the vicinity of their emissions sources. Under typical conditions,
PM 10-2.5 particles travel much shorter distances before settling out than finer particles
(U.S. EPA. 2009a). As a result, spatial and temporal heterogeneity is greater for PMi0-2.5
than for PM2 5, because coarser particles have greater settling velocities (Hinds. 1999).
and settling velocities are even greater for particles larger than 10 (im. Thus, spatial
gradients are steepest near sources, such that measured concentrations of larger particle
sizes tend to be most representative of the ambient air in areas in close proximity to the
monitor, with higher concentrations likely to occur closer to the source and decreasing
concentrations with increasing distance from the source. This issue has been thoroughly
discussed in the 2006 Pb AQCD (U.S. EPA. 2006b). It has also been acknowledged in
previous Pb AQCDs, with a lengthy discussion appearing in the 1977 AQCD (U.S. EPA.
1986b. 1977).
The low-volume Pb-PMi0 FRM sample collection method specifies use of a low-volume
PMio sampler that meets specified design criteria (40 CFRpart 50, Appendix Q).
Ambient airborne PM is collected on a polytetrafluoroethylene (PTFE) filter for 24 hours
using active sampling at local conditions with a low-volume PMi0 sampler and analyzed
by X-ray fluorescence (XRF). In recognition of the steep spatial gradients associated with
sources of ultracoarse particles, ambient Pb sampled using the FRM for Pb-PM10 is
allowed in certain instances where the expected Pb concentration does not approach the
NAAQS and no sources of ultracoarse Pb are nearby.
Alternative Sample Collection Methods
In addition to the FRMs for ambient Pb sample collection, a range of other PM sampling
methods are available for collecting samples for Pb analysis. These include FRM
sampling methods for PM that have also been used for collection of samples for Pb
analysis, sampling methods in use in other sampling networks such as the CSN,
IMPROVE and National Air Toxics Trends Stations (NATTS) networks described in
Section 2.4.2. and other sampling methods that have been used to measure airborne Pb
concentrations in research studies unrelated to network applications; these methods are
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listed in Table 2-3. Monitors that are not FRM sampling methods but are stated by their
manufacturer to capture TSP are called "manufacturer-designated TSP inlets" (denoted in
the text by "TSP") for this discussion. Most of these methods have been described in
considerable detail in the 2004 PM AQCD (U.S. EPA. 2004V Table 2-3 also lists key
conditions of capture for each method, including particle size, inlet type, collection
medium, and flow rate. Not all methods included in Table 2-3 and in subsequent
paragraphs have been applied to Pb-PM collection, but these methods represent potential
alternative methods for this purpose. It should also be noted that not all of the samplers
listed in Table 2-3 have been wind tunnel tested for a variety of aerodynamic particle
sizes, wind speeds, and wind directions. In addition, some of these samplers are no longer
commercially available.
Size discrimination is usually accomplished with impactors or cyclones. With impactors,
PM is forced through a jet at high speed, and particle inertia carries particles above a
given size into a collection surface downstream of the jet, while smaller particles follow
the air stream around the collector. In multistage impactors, a series of successive stages
of jets are used to collect a range of particle sizes. The micro-orifice uniform deposit
impactor (MOUDI) is a widely used multistage impactor. The impaction process and
performance of various impactors, including the WINS and MOUDI, has been described
in detail in the 2004 PM AQCD (U.S. EPA. 2004). The biggest concern in collection by
impaction is particle bounce, which occurs when particles collide with the collection
surface but bounce off the collection stage into the air stream and are not actually
collected. Considerable effort has been devoted to minimizing errors due to bounce in
FRM samplers, and this has been thoroughly discussed in the 2004 PM AQCD (U.S.
EPA. 2004). An alternative to impaction that also eliminates particle bounce is the use of
an air sampling cyclone. In the CSN and IMPROVE networks, cyclones are used to
remove particles larger than 2.5 (im. An air sampling cyclone brings air into a tangential
jet and directs flow against a circular wall, where particles larger than a given size are
removed by centrifugal and gravitational forces.
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Table 2-3 Airborne PM sampling methods potentially applicable for Pb sampling.
Sampler
High Volume TSP
Low Volume PM-io
PM2.5
Met One SASS
IMPROVE
MOUDI
Noll Impactor
SEAS
PM10 SSI HiVol
Andersen
Inhalable PM
Sampler
Sierra
Dichotomous
Sampler
TEOM
Network
Pb-FRM
PM-FRM,
NATTS
PM-FRM
CSN
IMPROVE
None
None
None
FRM
FRM
FRM
FRM
Sampler Type
Single
Channel
Single
Channel
Single Channel
Multiple Channel
Multiple Channel
Multistage
Impactor
Multistage
Impactor
Slurry
Single Channel
Single Channel
Dichotomous
Single Channel
Particle Size
Characteristics
TSP
MMAD <10 urn
MMAD <2.5 urn
MMAD <2.5 urn
MMAD <2.5 urn
8 stages MMAD
0.056-18 urn
4 stages MMAD
<108 urn
MMAD <1.2 urn
MMAD <10 urn
MMAD <10 urn
MMAD <10 urn,
MMAD <2.5 urn
MMAD <10 urn
Inlet or
Fractionator
Type
Gabled,
rectangular
Louvered Inlet +
PM10
Impactor
WINS Impactor or
(VSCC)
Cyclone
Cyclone
Impactor
Impactor
Impactor
PM-io Size
Selective Inlet
RAAS10 Inlet
10 urn Inlet +
Virtual Impactor
R&P PM-io Inlet or
Louvered Inlet
Collection
Medium
Glass
Teflon
Teflon
Teflon
Teflon
Teflon
Coated Mylar
Slurry
8" x 10" filter paper
Teflon
Teflon
Teflon Coated
Glass Fiber
Typical
Flow Rate
1130L/min
16.67 L/min
16.67 L/min
6.7 L/min
22.8 L/min
30 L/min
Rotating arm
90 L/min
11 30 L/min
16.67 L/min
16.7 L/min
16.7 L/min
Reference
U.S. EPA
(201 1f)
U.S. EPA
(201 1f)
U.S. EPA
(2Q11f)
MetOne (2009)
CNL (2001)
Marple et al.
(1991)
Noll (1970)
Pancras et al.
(2006), Ondov
et al. (2006)
U.S. EPA
(2Q11f)
U.S. EPA
(201 1f)
U.S. EPA
(201 1f)
U.S. EPA
(2Q11f)
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Table 2-3 (Continued): Airborne PM sampling methods potentially applicable for Pb sampling.
Sampler
Louvered
Inlet "TSP"a
Texas A&M
Lo-Vol "TSP"a
Andersen
Multistage
Impactor
UIUC Isokinetic
"TSP"a Sampler
Airmetric MiniVol
Stacked Filter
Units
ELPI
Wagner & Leith
Passive Sampler
Network
None
None
None
None
None
None
None
None
Sampler Type
Single Channel
Single Channel
Multistage
Impactor
Single Channel
Single Channel
Two stage
impactor
Multistage
Impactor
Passive
Particle Size
Characteristics
MMAD <15 urn
"TSP"
8 stages MMAD
0.4-10 urn and above
"TSP"
"TSP" or MMAD <10
urn or MMAD <2.5 urn
MMAD <10 urn,
MMAD <2.5 urn
13 stages MMAD
0.007-10 urn
Not characterized
Inlet or
Fractionator
Type
Louvered Inlet
None
Inlet Cone
Isokinetic
Sampling Head
None, PM-io
Impactor, or PM2.s
Impactor
PM-io Size
Selective Inlet
Berner Impactor
Mesh screen
Collection
Medium
Teflon
Teflon
Aluminum
Teflon
Teflon
Nuclepore
Aluminum or
polycarbonate
SEM stub
Typical
Flow Rate
16.67 L/min
16.67 L/min
28.3 L/min
20 L/min
5 L/min
17 L/min
10 or 30 L/min
0 L/min
Reference
Kenny et al.
(2005)
Wang et al.
(2005b);
Wanjura et al.
(2QQ5)
Mercer et al.
(1970)
Jerez et al.
(2006)
Chen et al.
(2011 a)
IAEA (1993)
Keskinen et al.
(1992)
Leith et al.
(2007);
Wagner and
Leith (2001)
a Monitors that are not FRM sampling methods but are stated by their manufacturer to capture TSP are called "manufacturer-designated TSP inlets"
(denoted by "TSP").
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Collection medium and flow rate are two other key features of a sampling method. One
advantage of low volume sampling is its suitability for collection of samples for XRF
analysis. Because Pb in PM2 5 is analyzed by XRF in the CSN and IMPROVE networks,
sampling methods that employ Teflon filters suitable for XRF analysis have been
developed for these networks. In practice, this restricts sampling for airborne Pb to low
volume samplers with a convenient filter size. This also holds true for the Pb-PMi0 FRM
sampling, which is also restricted to low volume PM10 samplers because XRF has been
designated as the FRM for Pb-PMi0 analysis. An additional practical advantage of
available low volume samplers over the existing high volume Pb-TSP FRM is that
established low volume PM2 5 and PMi0 sampling methods are not dependent on wind
direction. However, this has to do with sampler design rather than flow rate, and there are
high volume PMi0 sampling methods, including the PMi0 FRMs, that are also free of
wind direction bias. These would be suitable for Pb measurement with other analytical
methods, such as ICPMS, and could have a potential advantage of providing more
material in locations with very low concentrations.
Alternatives to TSP sampling have been developed that collect a particle size range that
extends beyond 10 jam. Early examples include samplers for "inhalable particulate
matter" that were designed to have cut points for 50% sampling efficiency of 15 jam
aerodynamic diameter. These included the Andersen Inhalable Particulate Sampler
(Model 7000, Thermo Electron, Smyrna, GA) and the Sierra Dichotomous Sampler
(Series 244, Sierra Instruments, Monterey, CA), which were evaluated and compared to
each other and to TSP sampling in both co-located field comparisons (Solomon et al..
1982) and wind tunnel studies (Watson et al., 1983). with the result that poor agreement
was observed for low or high wind speeds, or when much coarse particulate matter was
present. For example, the Dichotomous Sampler collected on average only 73 ± 18% as
much mass as the Inhalable Particulate Sampler, and differences were attributed to
differences in the efficiency of large particle collection (Solomon et al., 1982).
More recently, a variety of inlets have been developed for low volume "TSP" sampling.
The omnidirectional TEOM "TSP" Inlet (Model 10-002929, Rupprecht & Pataschnik
Co., Inc, Albany, NY) was designed to sample 100 jam particles in still air with the
suction velocity equal to the terminal velocity of a 100 jam diameter unit density sphere.
However, substantial PM mass loss was reported for this inlet and attributed to
anisokinetic sampling conditions that led to inefficient sampling of large particles (Jerez
et al., 2006). The inefficient sampling of larger particles by the "TSP" inlet was also
observed by Kenny et al. (2005). who carried out wind tunnel tests of 1) a commercially
available omnidirectional low volume (16.67 liters/minute) "TSP" inlet, and 2) a
louvered dichotomous inlet designed to select particles from a moving airstream and
transmit them to a downstream PM10 impactor. They observed that the "TSP" inlet
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exhibited low sampling efficiency even for larger particles within the PM10 range and
concluded that it is likely to give biased PM concentrations that vary with external winds
when large particles are present. However, for the louvered dichotomous inlet, Kenny et
al. (2005) reported high sampling efficiencies with little influence of wind speed across
the full PMio particle size range, and a 50% cutpoint at around 15 jam. For the full scale
"TSP," inlet sampling efficiencies for 46 jam particles were 17% for a 1 meter/second
wind speed and 28% at 2 meters/second.
Evaluation of other available "TSP" samplers also reveals variability among sampler
models. Jerez et al. (2006) compared the University of Illinois at Urbana Champaign
(UIUC) isokinetic "TSP" sampler with a tapered element oscillating microbalance
(TEOM) "TSP" sampler (Series 1400a, Rupprecht and Pataschnik Co., Inc., Albany, NY)
when measuring dust concentrations in swine and chicken houses in Illinois, Indiana,
Minnesota, and Texas between September to December, 2003. The TEOM measured
concentrations that were 26-117% of the UIUC sampler; for 86 of 90 measurements, the
TEOM measurements were lower than the UIUC measurements.
Other new approaches to high volume sampling include "saturation samplers" or low
volume (5 liters/minute) samplers designed for high spatial coverage of PMi0 and PM25.
Chen et al. (2011 a) provide an intercomparison among portable MiniVol (Airmetric,
Eugene, OR), Omni (BGI, Inc., Waltham, MA), and dichotomous samplers (Model 2025,
Rupprecht and Pataschnik, Albany, NY) and two FRMs (RAAS-100, Andersen, Smyrna,
GA; Partisol-FRM 2000, Rupprecht and Pataschnik, Albany, NY) for PM10 and PM2 5
measurements. Chen et al. (2011 a) observed R2 of 0.95-0.98 with average mass
concentrations within 4% among the PMi0 measurements. They observed R2 of 0.96-0.99
and average mass concentrations within 9% among the monitors for PM2 5
measurements. Hitzenberger et al. (2004) found more variability among portable PM2 5
and PMio monitors when performing intercomparisons in Melpitz, Germany. PM25 was
measured with stacked filter units (SFU) (developed at Ghent University), a TEOM
(Model 1400A, Rupprecht and Pataschnick, Albany, NY), a Digitel (Model DHA-80)
high volume sampler, an electrical low pressure impactor (ELPI) (Outdoor-ELPI, Dekati,
Ltd.), a TSP, and a MOUDI. PM10 was measured with SFUs, a "TSP" (manufacturer not
provided) with 30 L/min and 70 L/min Berner impactors, and a MOUDI (model and
manufacturer not provided). Based on Hitzenberger et al.'s (2004) reported average mass
concentrations, the PM25 samplers ranged from 73% (TEOM) to 180% (ELPI) of the
average. When trimming the extrema, the average PM25 concentrations measured within
10% of the overall average. For the PMio monitors, Hitzenberger et al.'s (2004) reported
average mass concentrations ranged from 83% (one SFU, MOUDI) to 123% (two SFUs)
of the overall average. When trimming the extrema, the average PMio concentrations
were within 13%.
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Passive samplers have also recently been used for capturing spatial variability of ambient
air Pb concentrations. Field testing of the Wagner and Leith Passive Sampler (University
of North Carolina) illustrated very good agreement with a cascade impactor (difference
within 5% for three sampling events) for PM25 in certain cases and very poor agreement
(difference of -51% and -110%) in others; Wagner and Leith (2001) attributed the poor
agreement to low PM2 5 mass events. PMi0 agreement was 14-65% and was attributed to
potential phenomena such as agglomeration on the passive sampler and water
evaporation from the impactor sample. However, when comparing the Wagner and Leith
Passive Sampler for PM10_2.5 with PM10.2.5 calculated by differencing measurements
obtained from co-located or dichotomous FRMs for PMi0 and PM25, Leith et al. (2007)
observed that Passive Sampler PM 10-2.5 measurements integrated over at least one week
were within one standard deviation of the PMi0-2.5 obtained from co-located or
dichotomous FRMs. Kumar et al. (2012) and Lagudu et al. (2011) illustrated how the
Wagner and Leith Passive Sampler can be coupled with computer-controlled scanning
electron microscopy (CCSEM) to produce concentrations of Pb-PMi0-25 in samples taken
across the cities of Syracuse, NY and Rochester, NY, respectively.
These results illustrate that alternative sampling options to TSP are available to capture
ambient air Pb concentrations for particles with a cutpoint of approximately 15 nm or
higher, and the state of the science for sampling in this particle size range is progressing,
including better documentation of performance and field intercomparison data. In
general, both the historical and recently available alternatives to the traditional TSP
sampler illustrate that samplers designed to collect particles up to a size range greater
than 10 nm have not performed or compared as well as samplers designed to collect
smaller particles, and the challenge of achieving good performance for collection of
particle size ranges greater than 10 nm has not been limited to the TSP sampler. This is
expected given the inherent difficulties associated with large particle sampling (Garland
and Nicholson. 1991). In spite of this, good performance by recently developed samplers,
including the louvered "TSP" inlet evaluated by Kenny et al. (2005) show promise for
good performance in collection of a particle size range extending beyond 10 nm. The
primary route of Pb exposure is hand-to-mouth contact with deposited Pb on soil or dust
having substantially larger size fractions compared with airborne Pb particles, as
described in Section 3.1. However, the relevant particle size distribution for ambient
sampling is smaller than the size distribution of the settled dust, as described in
Section 3.1.1.1. Particles larger than about 20 pirn are generally considered too large to be
transported for more than a few seconds under typical conditions; see Section 2.3.1.3. If
preliminary results concerning the "TSP" louvered inlet and other "TSP" alternatives are
verified, this may be very close to the practical limit for good sampling data quality. It
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follows that a size cut-off around 15 pirn may achieve both good data quality and
representative sampling in a limited area.
2.4.1.2 Sample Analysis: Federal Reference and Federal
Equivalence Methods
As described in Section 2.4.1.1. measurements for determinations of NAAQS compliance
must be made with FRMs or FEMs. As of October 12, 2011, one manual reference
method and 25 manual equivalent methods for sample analysis had been approved for Pb
(http://www.epa.gov/ttn/amtic/files/ambient/criteria/reference-equivalent-methods-
list.pdf). The FRM for Pb (Pb-TSP) was promulgated in 1979 and is based on flame
atomic absorption spectroscopy (AAS) (40 CFR Part 50, Appendix G). Ambient air
suspended in PM is collected on a glass fiber filter for 24 hours using a high volume air
sampler. Pb in PM is then solubilized by extraction with nitric acid (FŁNO3), facilitated
by heat, or by a mixture of FŁNO3 and hydrochloric acid (HC1) facilitated by
ultrasonication. The Pb content of the sample is analyzed by atomic absorption
spectrometry using an air-acetylene flame, using the 283.3 or 217.0 nm Pb absorption
line, and the optimum instrumental conditions recommended by the manufacturer.
Several FEMs have been approved based on a variety of principles of operation have
been approved, including inductively coupled plasma optical emission spectrometry, and
inductively-coupled plasma mass spectrometry (ICPMS).
Atomic Absorption Spectrometry
AAS is the basis for the existing FRM. Atomic absorption spectrometry was first
developed in the 19th century, and became widely used in the 1950s. More than 70
elements can be analyzed by AAS. Typically a liquid sample is nebulized into a flame
with sufficient heat for elements to be atomized. The liquid specified by the FRM is a
nitric acid extract of a glass fiber filter used for collection of suspended PM with a high
volume sampler. The atomized sample is then irradiated with visible light at a specific
wavelength to promote an electronic transition to a short-lived excited state, resulting in
absorption of the light. Elemental selectivity is achieved because light absorption is
specific to a particular electronic transition in a particular element. As a result, absorption
of light at a given wavelength generally corresponds to only one element. The flame is
irradiated with a known quantity of light and intensity of light is measured on the other
side of the flame to determine the extent of light absorption in the flame. Using the Beer-
Lambert law the concentration of the element is determined from the decrease in light
intensity due to sample absorption.
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A more sensitive variation of atomic absorption spectrometry for most elements is
graphite furnace atomic absorption spectrometry (GFAAS). Instead of introducing the
sample into a flame, the liquid sample is deposited in a graphite tube that is then heated to
vaporize and atomize the sample.
Inductively-Coupled Plasma Mass Spectrometry
Inductively coupled plasma mass spectrometry (ICPMS) is a sensitive method of
elemental analysis developed in the late 1980s. Argon (Ar) plasma (ionized gas) is
produced by transmitting radio frequency electromagnetic radiation into hot argon gas
with a coupling coil. Temperatures on the order of 10,000 K are achieved, which is
sufficient for ionization of elements. Liquid samples can be introduced into the plasma by
extracting samples in an acid solution or water and nebulizing dissolved elements.
Resulting ions are then separated by their mass to charge ratio with a quadrupole and
signals are quantified by comparison to calibration standards. While solid samples can be
introduced by laser ablation, nebulization of liquid extracts of PM collected on Teflon
filters is more typical. One major advantage of ICPMS over AAS is the ability to analyze
a suite of elements simultaneously. An additional advantage is low detection limits of
50-100 parts/trillion for Pb.
Inductively-Coupled Atomic Emission Spectroscopy
Inductively coupled atomic emission spectroscopy (ICP-AES) also generates ions from
elements with a hot Ar plasma, similar to ICPMS. Excited atoms and ions are produced,
and these emit electromagnetic radiation with frequencies characteristic of a particular
element. Intensity of emission is used to determine the concentration of an element in the
sample. Elements are extracted from filter samples and nebulized into the plasma.
Energy Dispersive X-ray Fluorescence
In energy dispersive X-ray fluorescence spectrometry a beam of X-ray photons from an
external excitation source is applied to a sample, causing ejection of inner shell electrons
from elements in the sample. Because inner shell electrons have higher electron binding
energies than outer shell electrons, the ejection of the inner shell electron induces an
energetically favorable electronic transition of an outer shell electron to replace the
ejected electron. The energy released as a result of this transition is in the form of
electromagnetic radiation, corresponding to the difference in electronic binding energies
before and after the transition. The energy released is typically in the X-ray portion of the
electromagnetic spectrum. The release of electromagnetic radiation as a result of an
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electronic transition is defined as fluorescence. Fluorescence energies associated with
electronic transitions depend on atomic structure, and vary between elements. As a result,
X-ray fluorescence energy is uniquely characteristic of an element, and X-ray intensity at
a given energy provides a quantitative measurement of elemental concentration in the
sample. The X-rays are detected by passing them through a semiconductor material,
resulting in an electrical current that depends on the energy of the X-ray.
2.4.1.3 Other Analysis Methods for Total Pb
Several other methods that have not been designated as FRM or FEM methods have also
been frequently used to obtain atmospheric Pb measurements. These include proton
induced x-ray emission (PIXE), X-ray photoelectron spectroscopy (XPS), and other
methods
PIXE
Proton-induced X-ray emission (PIXE) spectroscopy has been widely used to measure Pb
in atmospheric PM. In PIXE, a high-energy proton beam passes through the sample,
causing electrons to be excited from inner shells. The x-rays emitted when electronic
transition occur to replace the inner shell electrons are characteristic of an element and
can be used to identify it. Development of PIXE for analysis of airborne PM was
reviewed by Cahill et al. (1981). Numerous applications of PIXE to analysis of airborne
Pb-PM have been reported in the past five years (Cohen etal.. 2010; Waheed et al.. 2010;
Sanchez-Ccovllo et al.. 2009; Chan et al.. 2008; Johnson et al.. 2008; Cong et al.. 2007;
Ariola et al.. 2006; Johnson et al.. 2006; Wahlin et al.. 2006).
XPS
X-ray photoelectron spectroscopy (XPS), also called electron spectroscopy for chemical
analysis (ESCA) has been used to determine Pb concentrations on materials surfaces,
including atmospheric PM (Finlayson-Pitts and Pitts. 2000). A fixed frequency X-ray
beam causes inner shell electrons to be emitted and kinetic energy of ejected electrons is
measured. Binding energy characteristic of an element can be calculated from the
measured kinetic energy, allowing identification of the element. XPS can also provide
information about an element's chemical environment or oxidation states because of
chemical shifts in binding energy caused by differences in chemical environment. There
have been some recent applications of XPS to airborne PM, concluding that Pb was
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mostly in the form of Pb sulfate (Qi et al., 2006). XPS analysis is a surface technique that
is suitable only to a depth of 20-50A.
Other Total Pb Methods
Anodic stripping voltammetry, atomic emission spectroscopy, and colorimetry have also
been used for measurement of atmospheric Pb (Finlayson-Pitts and Pitts. 2000). In anodic
stripping voltammetry, metal ions are reduced to metallic form and concentrated as an
amalgam on a suitable electrode (e.g., a mercury (Hg) amalgam on a mercury (Hg)
electrode). This is followed by re-oxidation in solution, which requires "stripping" the
reduced metal from the electrode. Emission spectroscopy can be compared to the existing
FRM for Pb based on AAS. In AAS, radiation absorbed by non-excited atoms in the
vapor state is measured. In emission spectroscopy, radiation due to the transition of the
electron back to ground state after absorption is measured, and the energy of the
transition is used to uniquely identify an element in a sample. Colorimetric methods are
wet chemical methods based on addition of reagents to a Pb containing solution to
generate measurable light absorbing products. These methods are less sensitive than
ICPMS, XRF, and PIXE and their use is declining as more sensitive methods become
more widely used but have advantages regarding simplicity and cost.
2.4.1.4 Sequential Extraction
Sequential extraction has been widely used to further classify Pb for various purposes,
including bioavailability, mobility, and chemical speciation. In general the more easily
extractable Pb is considered more mobile in soil and is more bioavailable to organisms.
This approach has also been used widely in characterization of airborne PM. In its
original application (Tessier et al., 1979) metals extraction solvents were selected to
correspond to common species present in soil, and metals were classified as
exchangeable, bound to carbonates, bound to iron (Fe) and manganese (Mn) oxides,
bound to OM, and residual. Extraction was carried out with successively stronger
solutions, starting with magnesium chloride (MgCl2) for removal of exchangeable metals
and ending with hydrofluoric and perchloric acids for removal of residual metals. Pb was
one of the elements originally studied by Tessier et al. (1979) as well as one the elements
analyzed when Tessier's scheme was first applied to airborne PM (Fraser and Lum.
1983).
Tessier's scheme was modified and optimized for airborne PM over time (Fernandez
Espinosa et al., 2002) and additional extraction schemes were also developed (Chester et
al.. 1989). including the simplest case of two fractions corresponding to soluble and
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insoluble fractions (Falta et al.. 2008; Canepari et al.. 2006; Voutsa and Samara. 2002).
The variety of methods in current use was recently thoroughly reviewed by Smichowski
et al. (2005). With the recognition that biological processes involving deposited PM
metals were related to their solubility (U.S. EPA. 2009a). sequential extraction methods
or simpler schemes to divide metals into water and acid soluble fractions were
increasingly applied to PM samples to obtain data not just on total metal concentration
but also on water soluble concentration (Graney et al., 2004; Kyotani and Iwatsuki. 2002;
Wang et al.. 2002). Compared to other elements, a large fraction of total Pb is soluble
(Graney et al., 2004). Recent advances in this area have included application to size
fractionated PM (Dos Santos et al., 2009; Birmili et al., 2006). time resolved
measurements (Perrino et al.. 2010). and an extensive comparison of different
fractionation schemes (Canepari et al.. 2010). Sequential extraction with two or more
fractions is becoming more widely used for characterization of Pb-PM in a variety of
sources (Canepari et al.. 2008; Smichowski et al.. 2008; Poykio et al.. 2007; Sillanpaa et
al.. 2005) and locations (Perrino etal.. 2010; Dos Santos et al.. 2009; Cizmecioglu and
Muezzinoglu. 2008; Dahl et al.. 2008; Sato et al.. 2008; Annibaldi et al.. 2007; Richter et
al.. 2007; Al-Masri et al.. 2006; Canepari et al.. 2006; Fujiwara et al.. 2006; Wang et al..
2006c; Yadav and Raiamani. 2006; Gutierrez-Castillo et al.. 2005; Heal et al.. 2005).
leading to a better understanding of mobility characteristics of Pb in airborne PM.
2.4.1.5 Speciation Techniques
XAFS
There have been few attempts to speciate Pb in atmospheric PM. However, recently X-
ray absorption fine structure (XAFS) has been applied to PM and road dust to obtain Pb
speciation data from direct analysis of particle surfaces. In XAFS the absolute position of
the absorption edge can be used to determine the oxidation state of the absorbing atom,
and scattering events that dominate in the near edge region provide data on vacant orbital
energies, electronic configurations, and site symmetry of the absorbing atom that can be
used to determine the geometry of the atoms surrounding the absorbing atom. XAFS can
be divided into two spectral regions. X-ray absorption near edge structure (XANES) is
the region of the x-ray absorption spectrum up to 50 eV above the absorption edge
observed when an inner shell electron is electronically excited into unoccupied states, and
Extended X-ray Absorption Fine Structure (EXAFS) up to 1 keV above the absorption
edge. Both have been applied recently to Pb in PM. XANES spectra of Pb coordination
complexes with a wide range of environmentally relevant ligands have been reported
(Swarbrick et al.. 2009). XANES has been used to show that several different Pb species
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are probably present in urban airborne PM (Funasaka et al., 2008) and urban road dust
(Barrett et al.. 2010). XANES has been used to differentiate between Pb chromate,
Pb-sorbed minerals, Pb chloride, Pb oxide, Pb carbonate, Pb sulfide and Pb sulfate, which
are probably present in urban PM and road dust samples (Barrett et al.. 2010; Funasaka et
al., 2008; Tan et al.. 2006). XANES has also been used to investigate Pb in air samples
thought to be complexed with humic substances from soil (Pingitore et al.. 2009) and to
investigate the speciation of atmospheric Pb in soil after deposition (Guo et al., 2006b).
EXAFS has been applied to emission sources to show Pb from a sinter plant was mainly
carbonate (Sammut et al., 2010). XAFS has only been applied to airborne PM very
recently and shows promise for chemical speciation of airborne metals, including Pb.
GC-ICPMS and HPLC-ICPMS
Environmental analytical methods for organolead compounds prior to 2000 were
generally time consuming and costly, requiring extraction, derivatization, and detection
(Quevauviller. 2000). These have been thoroughly reviewed (Pyrzynska. 1996) and
method intercomparison studies have been conducted (Quevauviller. 2000). More
recently, speciation of organometallic compounds in environmental samples has usually
been carried out by coupling a chromatographic separation step with a mass
spectrometry-based multi-element detection system capable of analyzing a wide range of
elements along with Pb, and these approaches have also been recently reviewed (Hirner.
2006). Chromatographic systems in common use are gas chromatography (GC) and high
performance liquid chromatography (HPLC). Detection systems most commonly used are
an inductively coupled plasma mass spectrometer (ICP-MS [ICPMS]), electron impact
ionization mass spectrometry (El-MS), and electrospray ionization mass spectrometry
(ESI-MS) (Hirner. 2006). Using these techniques, organometallic species are separated
from each other based on differences in retention times on chromatographic columns, and
elemental Pb is determined by the ICPMS used as a detector downstream of the column
to measure elemental Pb in the pure compounds after chromatographic separation. Pb
speciation analysis has benefited from the development of HPLC-ICPMS in particular
(Quevauviller. 2000). Recent advances in metal speciation analysis in environmental
samples by HPLC-ICPMS have been extensively reviewed (Popp et al.. 2010). HPLC-
ICPMS has been used for analysis of Pb complexes with humic substances (Vogl and
Heumann. 1997). which could be relevant for resuspended soil and road dust. GC-ICPMS
has been more widely used for separation and analysis of methyl and ethyl Pb species in
atmospheric PM (Poperechna and Heumann. 2005; Jitaru et al., 2004; Leal-Granadillo et
al.. 2000).
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Pb-lsotope Ratio Analysis
Classifying Pb by its relative isotopic abundance has also proved useful for a variety of
purposes, including the determination of its geochemical origins in natural samples and
the relative contributions of coal burning, mining, smelting, and motor vehicle emissions
in polluted samples (Farmer et al.. 1996). Typically, isotopes of Pb (208Pb, 207Pb, 206Pb,
and 204Pb) are measured in a sample using mass spectrometry, and then ratios of the
isotopes are calculated to obtain a "signature." Isotopes of 208Pb, 207Pb, and 206Pb are
substantially more abundant than 204Pb, but they vary depending on the geologic
conditions under which the ore was produced through decay of different isotopes of
uranium and thorium (Cheng and Hu. 2010). Isotope ratio analysis was first applied to
airborne PM in 1965 to identify the impact of motor vehicle exhaust on marine and
terrestrial Pb deposition in the Los Angeles area (Chow and Johnstone. 1965). More
recently, high resolution ICPMS has also proved to be a sensitive tool for isotope ratio
analysis. High resolution ICPMS was first applied to geological samples (Walder and
Freedman. 1992). and has since been widely used for determination of Pb isotope ratios
in airborne PM samples. Pb isotope ratios have been measured in a number of recent
studies in a variety of locations to investigate the origin of airborne Pb (Knowlton and
Moran. 2010; Noble et al.. 2008; Hsu et al.. 2006; Widory. 2006). Shotyk and Krachler
(2010) also used Pb isotopes to demonstrate that the fate of Pb from runoff can be
different from Pb with different origins. They observed that humus PM impacted by
leaded on-road gasoline that are derived from soil surfaces are likely to be more easily
transferred to sediments than Pb of other origins, with substantial amounts retained by
lakes.
Recent studies have examined the use of Pb isotope ratios as a tool for source
apportionment. Duzgoren-Aydin and Weiss (2008) provide caveats for using isotope ratio
analyses. They point out that Pb isotope ratios may vary when Pb from several sources of
different geological origins are introduced to the same location. Duzgoren-Aydin (2007)
warned that the presence of a complex mixture of contaminants containing common Pb
isotopes can lead to an overestimation of the contribution of one source (e.g., soil
contaminated by Pb emissions from on-road gasoline) and an underestimate of another
source, such as that from industry. For this reason, Cheng and Hu (2010) suggest that Pb
isotope analysis only be used when the investigators are confident that the isotopic
signatures of various sources differ substantially. Pb recycling and international trading
may cause more blending of Pb from various sources, so that there is less heterogeneity
in the Pb isotopic signatures sampled. Additionally, Cheng and Hu (2010) point out that
the isotopic signature of Pb in air or soil may change over time with changing source
contributions, but historical Pb isotope data are lacking. Duzgoren-Aydin and Weiss
(2008) suggest the use of geographical information systems (GIS) mapping of Pb isotopic
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information to help distinguish potential sources based on location of sources in addition
to the sources' isotopic signature.
Gulson et al. (2007) examined the relationships between Pb isotope ratios and source
apportionment metrics at urban and rural sites in New South Wales, Australia. In this
study, Gulson et al. (2007) performed source apportionment with both principal
component analysis (PCA) and a neural network technique called the self-organizing map
(SOM) and compared results from each method with 206Pb/204Pb, 207Pb/206Pb, and
208pb/206pb obtained from PM sampieS5 although only 206Pb/204Pb results were presented
in detail. Wintertime "fingerprints" from both the PCA and SOM methods produced
similarly linear relationships with 206Pb/204Pb, with linearly decreasing relationships
between the isotope ratios and the "secondary industry," "smoke," "soil," and "seaspray"
source categories. However, the relationships of the isotope ratios with the SOM
fingerprints and PCA factors, respectively, were very similar. This finding may have
been due to the presence of elements such as black carbon and sulfur in several SOM
fingerprints and PCA factors. The authors suggest that this might be related to the
presence of several sources, which in combination result in a weak atmospheric signal.
Additionally, both PM2 5 and TSP samples were utilized for this study, and it was found
that similar results were obtained for either size cut. At the urban site, they observed that
the 206Pb/204Pb ratio decreased overtime with increasing contributions of industrial, soil,
smoke, and sea spray sources. For the most part, these sources were not substantial
contributions to Pb-PM25 for the rural site. As for the Tan et al. (2006) speciation study
described above, no notable differences were observed between the size fractions with
regard to isotopic signature.
2.4.1.6 Continuous Pb Monitoring
Development of high time resolution measurement capabilities has advantages for
determining peak exposure concentrations and diurnal exposure trends. High time
resolution samplers suitable for analysis after sampling by XRF and ICPMS have been
developed and applied. The eight-stage Davis Rotating Unit for Monitoring (DRUM)
impactor (Raabe et al.. 1988; Cahill et al., 1987) collects PM samples with a cascade
impactor on Mylar film substrate on a slowly rotating drum, with samples analyzed by
XRF. It has been used to measure size and time resolved Pb and other elements with a
time resolution of less than 6 hours using x-ray fluorescence (Cahill. 2003; Bench et al..
2002). The University of Maryland Semi-continuous Elements in Aerosol Sampler
(Kidwell and Ondov. 2004. 2001) uses direct steam injection to promote condensational
growth of sampled PM at a high flow rate, and accumulates resulting droplets in a slurry
by impaction. It has been successfully applied to measurement of Pb and other elements
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by AAS (Pancras et al.. 2006; Pancras et al.. 2005) with a 30-minute time resolution. This
approach is also suitable for ICPMS analysis. A gas converter apparatus has also been
developed to improve transfer of ions to the ICPMS, including Pb, and successfully
tested with outdoor air (Nishiguchi et al.. 2008). The continuous emissions monitor based
method (Xact) and X-ray based filter method (XFM) samples Pb and other metals
continuously and analyzes the samples using non-destructive XRF (Yanca et al.. 2006).
The Xact compared very well with theoretical concentrations based on wind tunnel
testing, but it has a very high (0.6 ng/m3) MDL, limiting its usefulness to areas with
higher concentrations.
Much of the recent progress in ambient aerosol instrumentation has been related to the
development and improvement of single particle mass spectrometry (Prather et al., 1994).
Preferential loss as a function of particle size is a concern with this method, but
considerable effort has been devoted to optimizing transfer from atmospheric pressure
down to time of flight operating pressures with minimal particle loss (Prather et al..
1994). This technique can also be considered as an effective method for real time Pb
measurement in PM, including size-resolved measurements from 0.1 to 4.0 (im (Silva and
Prather. 1997). Progress has continued in the development of single particle mass
spectrometry to quantify elements and organic ion fragments and a number of recent
applications that included (Snyder et al., 2009; Johnson et al., 2008; Bein et al., 2007;
Reinard et al.. 2007; Peknev et al.. 2006) or specifically targeted (Salcedo etal.. 2010;
Moffet et al., 2008a; Murphy et al., 2007) Pb measurements.
2.4.2 Network Design
Four national monitoring networks collect data on Pb concentrations in ambient air and
report it to the Air Quality System (AQS).: State and local agencies carry out the
monitoring at state and local monitoring stations (SLAMS) using FRMs and FEMs and
report data to these national networks, which have been established for various purposes.
Although these data may be used for other scientific purposes, the SLAMS network is
designed primarily with the goal of evaluating compliance with the Pb NAAQS. In
addition to FRM monitoring, Pb is also measured within the Chemical Speciation
Network (CSN), IMPROVE, and the NATTS networks as described in Section 2.4.2.2.
Measurements among these networks are not directly comparable in all cases because of
method differences, including the PM size range sampled (TSP, PMi0, or PM2 5).
1 The Air Quality System (AQS) is EPA's repository of ambient air quality data. AQS stores data from over 10,000
monitors, 5,000 of which are currently active (http://www.epa.gov/ttn/airs/airsaqs/').
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2.4.2.1 NAAQS Monitoring Network
Monitors in the SLAMS network include predominantly those sited in compliance with
regulatory requirements for the purposes of judging attainment with the NAAQS. For this
purpose, these sites employ FRM samplers coupled with FRM/FEM analysis methods. At
the time of the last review, there were approximately 250 sites operating in this network,
although analyses at the time indicated incomplete coverage of the larger stationary
sources of Pb (U.S. EPA. 2007h). As a result of the review, the Pb NAAQS monitoring
requirements were revised. These revisions, some aspects of which were finalized in
2008 and the remainder in December 2010, call for expanded monitoring at both source-
oriented and non-source-oriented sites (75 FR 81126, 40 CFR part 58, Appendix D,
Section 4.5 to Part 58). Source-oriented monitoring sites are required near sources of Pb
air emissions which are expected to or have been shown to contribute to ambient air Pb
concentrations in excess of the NAAQS. At a minimum there must be one source-
oriented site located to measure the maximum Pb concentration in ambient air resulting
from each non-airport Pb source estimated to emit Pb at a rate of 0.50 or more tons/year
and in locations near those airports at which activities associated with the use of leaded
aviation fuel are estimated to result in Pb emissions at a rate of 1.0 or more tons
per year.1 The emission monitoring threshold was established to ensure monitoring near
Pb air sources with the greatest potential to cause ambient air concentrations to exceed
the Pb NAAQS.2 The Pb NAAQS measurements required at source-oriented sites must
be as Pb-TSP but measurements at non-source oriented sites may be as Pb-TSP or
Pb-PM10(75FR81126).
Monitoring agencies are also required to conduct non-source-oriented Pb monitoring at
each National Core multipollutant monitoring network (NCore)3 site in a Core Based
Statistical Area (CBSA) with a population of 500,000 or more. While non-source-
oriented monitoring data can be used for purposes of NAAQS attainment designations,
the main objective for non-source-oriented monitoring is to gather information on
JThe requirement for monitoring near sources emitting 0.5 tons/year or more may be waived if it can be shown that
the source will not contribute to a maximum 3-month average Pb concentration in ambient air in excess of 50
percent of the NAAQS level based on historical monitoring data, modeling, or other means (40 CFR, part 58,
Appendix D, Section 4.5(a)(ii)).
2 EPA Regional Administrators may require additional monitoring beyond the minimum requirements where the
likelihood of Pb air quality violations is significant. Such locations may include those near additional industrial Pb
sources, recently closed industrial sources, airports where piston-engine aircraft emit Pb and other sources of re-
entrained Pb dust (40 CFR, part 58, Appendix D, Section 4.5(c).
3 NCore is a new network of multipollutant monitoring stations intended to meet multiple monitoring objectives.
The NCore stations are a subset of the SLAMS network are intended to support long-term trends analysis, model
evaluation, health and ecosystem studies, as well as NAAQS compliance. The complete NCore network consists of
approximately 60 urban and 20 rural stations, including some existing SLAMS sites that have been modified for
additional measurements. Each state will contain at least one NCore station, and 46 of the states plus
Washington, D.C., will have at least one urban station.
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neighborhood-scale Pb concentrations that are typical in urban areas so to better
understand ambient air-related Pb exposures for populations in these areas.
Spatial scales defined for Pb monitoring range from microscale to neighborhood scale,
with the most important spatial scales for source-oriented sites to effectively characterize
emissions from point sources being microscale and middle scale, and the most important
scale for non-source-oriented sites to characterize typical Pb concentrations in urban
areas being neighborhood scale [40 CFR Part 58, Appendix D, 4.5(d)]:
• Microscale: This scale is intended to typify areas in close proximity to Pb point
sources where it may represent an area impacted by the emissions plume with
dimensions ranging from several meters up to about 100 meters.
• Middle Scale: This scale is described as generally representing Pb air quality
levels in areas up to several city blocks in size with dimensions on the order of
approximately 100 meters to 0.5 km.
• Neighborhood Scale: This scale is to characterize concentrations throughout
some relatively uniform land use areas with dimensions in the 0.5 to 4.0 km
range. Where a neighborhood site is located away from immediate Pb sources,
the site may be very useful in representing typical air quality values for a larger
residential area, and therefore suitable for population exposure and trends
analyses.
Source oriented monitors near sources estimated to emit 1.0 tons/year Pb were required to
be operational by January 1, 2010, and the remainder of the newly required monitors,
including the non-source-oriented NCore sites, were required to be operational by
December 27, 2011 (75 FR 81126). Based on implementation of the December 2010 Pb
network requirements as of late 2012, the Pb NAAQS monitoring network consists of
approximately 260 required monitors. Figure 2-14 shows the geographic distribution of
Pb NAAQS monitors in the current Pb NAAQS monitoring network. This includes
monitors that previously existed and are still in operation, along with those that are newly
required.
With the December, 2010 regulations, EPA also required one year of Pb-TSP (FRM)
monitoring near 15 airports in order to gather additional information on the likelihood of
NAAQS exceedances near airports due to the combustion of leaded aviation gasoline
(75 FR 81126). These airports were selected based on three criteria: annual Pb inventory
between 0.50 tons/year and 1.0 tons/year, ambient air within 150 meters of the location of
maximum emissions (e.g., the end of the runway or run-up location), and airport
configuration and meteorological scenario that leads to a greater frequency of operations
from one runway. These characteristics were selected because they are expected,
collectively, to identify airports with the highest potential to have ambient Pb
concentrations approaching or exceeding the Pb NAAQS. Data from this monitoring
study will be used to assess the need for additional Pb monitoring at airports. These 15
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sites (Figure 2-15 and Table 2-4) were required to be operational no later than December
27, 2011. Evaluating the air quality impact of piston aircraft operations includes
consideration of the seasonal variation in activity by these aircraft. At some of the most
active general aviation airports in the country, spring and summer operations (by piston
aircraft) can increase as much as 73% over operations in the fall and winter, while at
other airports, piston aircraft activity is more consistent throughout the year.
Alaska
,0
Guam Hawaii
Start Date
a 2008 and earlier
' 2009 and later
Puerto Rico
Figure 2-14 Map of monitoring sites in current Pb NAAQS monitoring
network.
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Quality assured results of this study were not available in time for this assessment. Note that the two Santa Clara County, CA
airports are not distinguishable on the map.
Figure 2-15 Fifteen U.S. locations where a study is currently being performed
on airport Pb emissions.
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Table 2-4 List of 15 airports included in the airport study.
Airport County, State
Merrill Field Anchorage, AK
Pryor Field Regional Limestone, AL
Palo Alto Airport of Santa Clara County Santa Clara, CA
Reid-Hillview Santa Clara, CA
McClellan-Palomar San Diego, CA
Gillespie Field San Diego, CA
San Carlos San Mateo, CA
Nantucket Memorial Nantucket, MA
Oakland County International Oakland, Ml
Republic Suffolk, NY
Brookhaven Suffolk, NY
Stinson Municipal Bexar, TX
Northwest Regional Denton, TX
Harvey Field Snohomish, WA
Auburn Municipal King, WA
2.4.2.2 Other Pb Monitoring Networks
In addition to FRM monitoring, Pb is also measured within the Chemical Speciation
Network (CSN), Interagency Monitoring of Protected Visual Environments (IMPROVE),
and the National Air Toxics Trends Station (NATTS) networks. Pb in PM25 is monitored
as part of the CSN and IMPROVE networks, and Pb in PM10 as a part of the National Air
Toxics Trends (NATTS) networks (Figure 2-16 and Figure 2-17). These networks are
designed to meet different objectives than those of the Pb NAAQS monitoring network.
The purpose of the CSN is to monitor PM25 species to assist in understanding PM25
chemistry and for spatial and temporal analyses including annual, seasonal, and sub-
seasonal trends (http://www.epa.gov/ttn/amtic/specgen.html). The CSN consists of about
50 long-term trends sites (commonly referred to as the Speciation Trends Network or
STN sites) and about 150 supplemental sites, all operated by state and local monitoring
agencies. Higher spatial and temporal resolution of the CSN facilitates increased utility in
the scientific community, and the data from the CSN also assist states in formulating their
emission control strategies, even if the network is not compliance-oriented. Pb is one of
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33 elements in PM2 5 collected on Teflon filters every third day and analyzed by energy
dispersive XRF spectrometry.
In the IMPROVE networks, PM2 5 monitors, operated by the National Park Service,
largely with funding by EPA, are placed in "Class I" areas (including National Parks and
wilderness areas) and are mostly in rural locations. IMPROVE monitoring is intended to
establish current visibility conditions, track changes in visibility and determine causal
mechanisms of visibility impairment in 156 national parks and wilderness areas. There
are 110 formally designated IMPROVE sites and approximately 80 additional sites at
various urban and rural areas, informally treated as part of this network and operating
under IMPROVE protocols. At these sites, Pb in PM2 5 is determined by XRF (UC.
1995).
The NATTS network is designed to monitor concentrations of hazardous air pollutants
(HAPs). The NATTS is intended to provide model input, to observe long-term trends in
HAP concentrations, and to examine emission control strategies. The NATTS network
measures several inorganic HAPs in PMi0, along with several volatile organic
compounds (VOCs), carbonyls, and polycyclic aromatic hydrocarbons (PAHs). It is
operated by state and local agencies and has less extensive national coverage than the
other Pb monitoring networks. PMi0 is collected either by high volume sampling with a
quartz fiber filter or low volume sampling with a PTFE filter following EPA
Compendium Method IO-3.5 (U.S. EPA, 1999). Pb is one of seven core inorganic HAPs
collected on Teflon filters and analyzed by ICPMS. As of 2012, the network consisted of
25 monitoring stations that measure Pb-PMi0, including 20 urban and 5 rural stations
operating on a one in six day sampling frequency (Figure 2-17).
Pb monitoring is also conducted atNCore monitoring sites. Monitoring for Pb-PM25 is
currently being conducted at NCore sites as part of the larger CSN (see Section 2.4.2.1).
Non-source-oriented Pb monitoring required at the 51 NCore sites with a population of
500,000 or more (shown in Figure 2-17) is generally monitored in Pb-PM10. Methods for
Pb in PMio-2.5 are being developed as part of the PMi0.2 5 speciation pilot project and may
be implemented at some NCore sites in the future. As shown in Figure 2-17, there are
some cases where states have addressed their NATTS and NCore Pb monitoring needs
with the use of a single monitoring site or may have nearby sites. In addition to the
NATTS network and NCore sites, as of 2012, states were collecting Pb-PMi0 at an
additional 22 sites (most as part of the Urban Air Toxics Monitoring program).
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Alaska
0 250 500 1,000 Miles 0 70 140 280 Miles 0 245 490
CSN
IMPROVE
Puerto Rico &
Virgin Islands
Ł*
0 25 50 100 Miles
Figure 2-16 Pb-PM2.s monitoring sites for CSN and IMPROVE networks.
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Alaska
'*
<0i
Guam Hawaii
* NCore/NATTS
• NCore
* NATTS
Other
Puerto Rico
Note: Presented on this map are all NCore sites where FRM/FEM Pb monitoring is required along with other sites actively
monitoring Pb-PMio (by any method) based on having 2012 data in AQS as of September, 2012.
Figure 2-17 Pb-PMio monitoring sites for NATTS network.
2.5 Ambient Air Pb Concentrations
The following section synthesizes ambient air Pb concentration data obtained during the
years 2008-2010 with data from studies in the literature presenting Pb concentrations
under varied source influences. The 3-month average ambient air Pb concentrations
presented here were created using a simplified approach of the procedures detailed in
40 CFR part 50 Appendix R and, as such, cannot be directly compared to the Pb NAAQS
for determination of compliance with the Pb NAAQS. For the purpose of analyses within
this ISA, monitors were initially designated to be source-oriented if either (1) they were
designated in AQS as source-oriented, (2) they were located within one mile of a
0.5 ton/year or greater source as noted in the 2005 NEI (U.S. EPA. 2008a). or (3) they
were located within one mile of a 0.5 tons/year or greater source as noted in the 2008 NEI
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(U.S. EPA. 201 la). The remainder of Pb-TSP FRM monitors reporting to the AQS were
classified as non-source-oriented.1 For this analysis, 120 Pb-TSP FRM monitors were
considered source-oriented, while 184 were considered to be non-source-oriented.
However, the number of source-oriented and non-source-oriented monitors differed for
each analysis year because there were changes in monitor siting. Summary information is
presented within this section, and detailed data are included the Chapter 2 Appendix
(Section 2.8. Supplemental Material) to this chapter.
2.5.1 Spatial Distribution of Air Pb
2.5.1.1 Variability across the U.S.
This section presents nationwide Pb concentration data measured using source-oriented
and non-source-oriented Pb-TSP FRM monitors from 2008-2010 and PMi0 and PM2s
monitors for 2007-2009. The source and non-source-oriented Pb-TSP FRM monitors
present data pertaining to compliance with the current level of the NAAQS. Pb-PMi0
data obtained from the NATTS network and Pb-PM2 5 data from the CSN are presented
in the Chapter 2 Appendix (Section 2.8) to illustrate the nationwide distribution of Pb
concentration in different classes of particle size. This information is useful to develop a
sense of variability in Pb concentrations at a national scale.
Concentrations of Pb Measured using Pb-TSP Monitors (Source-Oriented
and Non-Source-Oriented Monitors)
Maximum 3-month average Pb concentrations2 were calculated for source-oriented
Pb-TSP monitors for 50 counties across the U.S. (1.6% of U.S. counties) during the
period 2008-2010. Figure 2-18 illustrates that the level of the NAAQS was exceeded in
twenty counties where source-oriented monitoring was performed. The mean exceeded
the level of the NAAQS and was skewed toward the 75th percentile of the distribution,
indicating that highest ambient air Pb concentrations are near a small subset of the
monitors. Upper 90th percentile ambient air Pb concentrations for 2008-2010 occurred in
1 Following this initial classification, staff from the EPA Regional offices tasked with acting as liaisons to the states
reviewed all monitors listed to fall within their Regions and reported any discrepancies between the initial
classification and ground observations of the sites made by EPA Regional or state staff. The source and non-source
monitor listing was edited accordingly. The definition of source-oriented monitoring is applied flexibly with input
from regions in this ISA because 2008 data were obtained before the latest monitor designation requirements were
implemented.
2 Maximum 3-month average Pb concentrations are calculated as the maximum 3-month average of 3 consecutive
monthly averages within the 2008-2010 time period.
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Pike Co., AL, Los Angeles Co., CA, Iron Co., MO, Jefferson Co., MO, and Collin Co.,
TX. Summary statistics for the monitor-specific one-month and three-month averages are
presented in Table 2-5. and detailed statistics for the one-month and three-month
averages are provided in Table 2-12. Table 2-14. Table 2-16. and Table 2-18 in the
Chapter 2 Appendix (Section 2.8).
Maximum 3-month average Pb concentrations were calculated for non-source-oriented
Pb-TSP monitors for 47 counties across the U.S. (1.5% of U.S. counties) during the
period 2008-2010. Figure 2-19 illustrates that the level of the NAAQS was never
exceeded at non-source-oriented monitors. Summary statistics are presented below in
Table 2-6. and detailed statistics for the one-month and three-month average and maxima
non-source-oriented Pb-TSP concentrations are provided in Table 2-13. Table 2-15.
Table 2-17. and Table 2-19 in the Chapter 2 Appendix (Section 2.8).
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Table 2-5 Summary data for source-oriented Pb monitors across the U.S.,
2008-2010.
Mean, |jg/m
Median, [jg/m3 95th%, [jg/m3 99th%, [jg/m3 Max,
Monthly
0.20
0.063
0.86
1.6
4.4
3-mo rolling avg 0.21
0.079
0.88
1.6
2.9
Concentration:
" >= 1.55 u.g/m3 (5 counties)
• 0.76 -1.54 u.g/m3 (2 counties)
0.16 - 0.75 ug/m3 (13 counties)
• 0.06-0.15 ug/m3 (13 counties)
<= .05 u.g/m3 (17 counties)
U no data
Figure 2-18 Highest county-level source-oriented Pb-TSP concentrations
(ug/m3), maximum 3-month average, 2008-2010.
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Table 2-6 Summary data for non-source-oriented Pb monitors across the U.S.,
2008-2010.
Mean, |jg/m
Median, [jg/m3 95th%, [jg/m3 99th%, [jg/m3 Max,
Monthly
0.012
0.010
0.040
0.052
0.14
3-mo rolling avg 0.012
0.010
0.037
0.048
0.073
Concentration:
• 0.06 - 0.15 u.g/m3 (4 counties)
<= .05 u.g/m3 (43 counties)
Q no data
Figure 2-19 Highest county-level non-source-oriented Pb-TSP concentrations
(ug/m3), maximum 3-month average, 2008-2010.
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2.5.1.2 Intra-urban Variability
Intra-urban variability is defined as the variation in Pb concentration across an urban
area. Because the source characteristics and size distribution of particle-bound Pb can
vary considerably in urban areas, spatial variability of Pb concentrations in urban areas
may also be high. Moreover, larger Pb-PM tends to settle quickly over short distances
after becoming airborne; short settling distances also contribute to high spatio-temporal
variability in ambient air Pb concentrations. Such variability may not be detected if one
or a small number of central site monitors is in use, so cities with multiple monitors are
used to characterize intra-urban variability.
Data for intra-urban variability in six U.S. counties are presented in Section 2.8.2 of the
Chapter 2 Appendix. When collectively reviewing the data from the six counties, it
became apparent that spatial and temporal variability of Pb-PM concentrations were
commonly high compared for example with PM2 5, which tends to have fairly
homogenous concentrations over urban areas because it is subject to secondary
formation. Variability was high for areas that included a Pb source, with high
concentrations downwind of the sources and low concentrations at areas far from sources.
When no large sources of Pb were present, variability of Pb concentrations were lower,
and more data were observed to lie below the MDL. For example, Los Angeles County,
CA data illustrated very high concentrations adjacent to a Pb recycling facility, but
non-source-oriented concentrations were well below the level of the NAAQS at all times,
including at sites near roads. As described in Section 2.3. PM size distribution influences
how far the particle will travel upon initial emission or resuspension before being
deposited. Meteorology, nature of the sources, distance from sources, and positioning of
sources with respect to the monitors all appeared to influence the level of concentration
variability across time and space.
Airborne Pb near Roads
Five monitors, described in Table 2-7. were selected from the TSP network to examine
Pb concentrations in the near road environment. These monitors were selected because
they are located in the vicinity of major roadways in urban areas with different
characteristics and because they each have long-term data available. Further, based on
reviews of emissions inventory information as well as satellite image searches, these sites
are not known to be near metals-related industrial sites. Time series of Pb-TSP monthly
concentration for all five monitors are shown in Figure 2-20. The annual average over the
two sites that were reporting data in 1980 was 0.90 (ig/m3. This Pb-TSP concentration
from 1980 likely reflected the influence of Pb emissions from leaded automobile gasoline
(see Figure 2-7 for annual national consumption of leaded motor vehicle gasoline). By
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1986, when all five monitors were reporting data, the annual average of Pb-TSP
concentration over all five monitors dropped to 0.18 (ig/m3. Over 2001-2010, the annual
average Pb-TSP concentration over all sites was 0.02 (ig/m3 with a standard deviation of
0.01 (ig/m3. The highest 2008-2010 design value was 0.04 (ig/m3, which occurred at the
Chicago site (17-031-6003) located less than 10 meters to Interstate 1-290 at a monitor
height of 2 meters AGL. The multi-site average was not substantially larger than the
maximum three-month rolling average of 0.012 (ig/m3 for non-source-oriented monitors
for the period 2008-2010, and the Pb-TSP concentration varied little over the period
2001-2010. Note that the monitor heights were 2-6 meters AGL, which may be higher
than the vertical distance likely traveled by some particles (depending on particle size)
following initial re suspension (see Section 2.3.1.3).
060374002
060651003
170310052
170316003
250250002
r
o o o o o o o
O O O
Note: Monitor IDs from Table 2-7: Los Angeles, CA: 06-037-4002 (dark red); Riverside, CA: 06-065-1003 (dark green);
Two from Cook, IL: [17-031-0052 (light red), and 17-031-6003 (light green)]; and Suffolk, MA: 25-025-0002 (blue).
Figure 2-20 Time series of monthly average Pb-TSP concentration at five
near-road monitors.
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Table 2-7 Sample of U.S. near-road Pb-TSP monitors.
County,
State Site ID
Los
Angeles, 06-037-4002
CA
Riverside, 06.065.1003
L//\
fL00k' 17-031-0052
[[00k' 17-031-6003
Suffolk, 25.025.0002
2008-2010
Design Monitor
Value Height Distance
Latitude Longitude (ug/m3)a'b (m AGL) from Roads
500 m to
Interstate I-405
33.82376 -118.18921 0.01 6 f»^0 m
to Long Beach
Blvd
Within 20 m of
33.94603 -117.40063 0.01 4 intersection of
Magnolia and
Arlington Ave.
Near to
intersection of
Interstates I-90
41.96548 -8,749928 0.02< 5 J^ ™
I-90, 200 m to
Interstate I-94,
70 m to railroad
Less than 10m
to Interstate
41.872202 -87.826165 0.04C 2 I-290 (Dwight
D. Eisenhower
Expressway)
95 mto
Interstate I-90,
42.348873 -71.097163 0.02C 5 jjside median
Commonwealth
Ave.
Surrounding
Area
High intensity
residential,
urban
High intensity
residential,
mixed use
urban
Located at
public utilities
water
pumping
station, high
density
residential
urban
Parking lot of
Circuit Court
of Cook
County, 75%
surrounded
by Concordia
Cemetary
High intensity
urban, mixed
use
residential
and
commercial
"The level of the 2008 NAAQS for Pb is 0.15 ug/m not to be exceeded in any 3-month period. The design value for the 2008 Pb
NAAQS is the maximum rolling 3-month Pb-TSP average within the 3-year design period.
bThe design values shown here are computed for the latest design value period using Federal Reference Method (FRM) or
equivalent data reported by States, Tribes, and local agencies to EPA's AQS as of 7/12/2011. Concentrations flagged by States,
Tribes, and local agencies as exceptional events (e.g., high winds, wildfires, volcanic eruptions, construction) and concurred by the
associated EPA Regional Office are not included in the calculation of these design values. Although the indicator for the 2008 Pb
NAAQS is Pb-TSP at "local conditions" (i.e., actual temperature and pressure; parameter 14129), 2008 Pb-TSP data reported in
"standard temperature and pressure" (i.e., 25 ° C, 760 mmHg; parameter 12128) are also considered valid for NAAQS comparisons
and related attainment/nonattainment determinations if the sampling and analysis methods that were utilized to collect that data
were consistent with previous or newly designated FRMs or FEMs and quality assurance requirements were met.
°Fewer than 36 rolling 3-month Pb-TSP average data are available at this site for this 3-year period; the value shown here is the
highest valid 3-month mean.
Airborne Pb near Airports
There have been only a few studies of air Pb concentrations near airports, but they have
generally demonstrated consistent results. Levin et al. (2008) summarized findings from
measurements at Buttonville Airport near Toronto, Canada that median air Pb-PM10
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levels were not substantially higher than average reported background levels (0.01 (ig/m3
versus 0.007 (ig/m3) (Conor Pacific Environmental Technologies Inc. 2000). although the
Buttonville analysis averaged upwind and downwind data. The maximum 24-hour
concentration measured in this 10-day study was 0.13 ug/m3. The Illinois report noted
that air Pb concentrations were elevated downwind of O'Hare airport compared with
upwind levels (Illinois Environmental Protection Agency. 2002). Carr et al. (2011) also
noted elevated Pb concentrations immediately downwind of the runway when studying
Pb concentrations at the Santa Monica Airport and surrounding neighborhood in Santa
Monica, CA in 2009. Twenty-four hr average concentrations at the downwind location
were higher than concentrations measured immediately upwind of the runway (winter:
0.040 (ig/m3 versus 0.0075 (ig/m3; summer: 0.049 (ig/m3 versus 0.0040 (ig/m3). Summer
measurements also included a residential neighborhood 100 meters further downwind,
which were still higher than upwind (0.033 (ig/m3 versus 0.004 (ig/m3). Modeling results
from Carr et al. (2011) also suggest that three-month average Pb concentrations above
local background extended beyond the airport property and that the preflight runup
check, taxi, and takeoff emissions were the most important contributors to Pb
concentrations. This airport had a Pb emissions inventory of 0.3 tons/year (U.S. EPA.
2011 a), which is below the threshold for airports for requiring consideration of Pb
NAAQS compliance monitoring; the monitoring regulations also specify the placement
of monitors in locations where available information indicates the potential for
exceedance of the NAAQS (see Section 2.4.2.1).
Airborne Pb at Urban and Rural Sites
A number of studies characterizing airborne Pb-bearing PM distribution at the
neighborhood scale suggest that spatial variability in Pb concentrations is related to local
sources. Yu et al. (2011) measured Pb-PMi0 concentration using a four-channel PM
sampler (Thermo Scientific) at four roof-top sites (10-13 meters AGL) within Paterson,
NJ: (1) background, (2) near-road (within 0.8 km of two major roads), (3) industrial
(within 1 km of three metal processing facilities), and (4) commercial. Coefficient of
variation (CV), defined as the standard deviation of site measurements divided by the
average) was 31.3%, with concentrations ranging of 5.61 ng/m3 (near road), 6.48 ng/m3
(industrial), and 6.58 ng/m3 (commercial), compared with 2.95 ng/m3 at the background
site. Harrison and Yin (2010) also noted that median urban background Pb concentrations
were elevated compared with rural background (urban: 13.9 ng/m3; rural: 8.0 ng/m3).
Martuzevicius et al. (2004) examined the spatial variability of Pb-PM2 5 samples obtained
in the greater Cincinnati, OH area at 6 urban, 4 suburban, and 1 rural site using Harvard
PM2 5 Impactors. They found that Pb-PM2 5 had a CV of 33.8%, compared with a CV for
PM25 of 11.3% over all sites. Average Pb-PM25 concentration among the sites varied
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from 1.79-28.4 ng/m3. Martuzevicius et al. (2004) suggested that differences between
mass and Pb spatial variability implied that Pb originated primarily from local sources.
Sabin et al. (2006a) measured Pb-PM with a Noll Rotary Impactor having an upper
cutpoint of 29 urn and found that urban concentrations ranged from 2.2 to 7.4 ng/m3 with
a CV of 40%. In contrast, a rural location had a concentration of 0.62 ng/m3. Li et al.
(2009a) collected PM2 5 with a Harvard Impactor and observed that Pb concentration in
PM2 5 samples was 2.2-3.0 times higher near a bus depot than next to a rural-suburban
road; in this study, the authors provided ratios but not actual concentrations. Ondov et al.
(2006) measured Pb-PM2 5 concentration at three Baltimore sites using an FRM. Average
Pb-PM25 concentrations at the different sites were 8.3 ng/m3, 7.2 ng/m3, and 1.9 ng/m3,
with the two higher concentration sites located within two miles of industrial facilities.
The industrial sites include a major steel plant; several chemical manufacturing plants;
and incinerators for municipal waste, medical waste, and sludge. Although these
concentrations are low, they agree with the body of literature to suggest that intra-urban
variability is most strongly related to source type, strength, and location.
2.5.2 Temporal Variability
The following sections present data for multi-year trends and seasonal variability of Pb
concentrations on a nationwide basis. The data presented here provide information on the
success of Pb reduction efforts over past decades as well as on areas for continued
attention with respect to Pb monitoring. The multi-year trends illustrate changes in air Pb
concentrations resulting from the phase-out of leaded gasoline for automobiles and
smaller reductions of industrial Pb usage. The seasonal variability plots demonstrate
changes in concentration within a given year, potentially related to climate or source
variation.
2.5.2.1 Multi-year Trends
Pb-TSP concentrations have declined substantially during the years 1980-2010. For
source and non-source monitors combined in the trends network, the annual average of
the maximum 3-month averages across 74 Pb-TSP monitors reporting air Pb
concentrations over the period from 1980-2010 has dropped by 89% from 1.3 ug/m3 in
1980 to 0.14 ug/m3 in 2010 (n = 31); see Figure 2-21. The median maximum 3-month
average concentration has declined by 97% from 0.87 ug/m3 in 1980 to 0.03 ug/m3 in
2010. The decline can be attributed to the phase-out of Pb antiknock agents in on-road
fuel and reductions in industrial use and processing of Pb, as described in Section 2.2.1.
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Average concentrations in these calculations are heavily influenced by the source-
oriented monitors in the network.
1978-2008 Level of the NAAQS
1990 1992 1994 1996 1998 2000 2002 2004 1006 2008 2010
Note: Annual average of maximum 3-month average Pb-TSP concentrations is shown by the solid black line, annual median of
maximum 3-month average concentrations is shown by the solid blue line, and the 10th and 90th percentiles are shown by the
dashed lines.
Figure 2-21 National trends in Pb concentration (ug/m3), 74 trends sites,
1990-2010.
2.5.2.2
Seasonal Variations
This section outlines seasonal variability among Pb monitors. Seasonal variation may
provide insight related to differential influences of sources and climate throughout a year.
Figure 2-22 illustrates average monthly trends in Pb-PM2 5 at four IMPROVE sites: Lake
County, CA (060333010), Bronx County, NY (360050110), Monroe County, NY
(360551007), Chittenden, VT (500070007). In each plot, some month-to-month
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variability is evident. Seasonal trends, with peaks in median,75th, and 95th percentiles in
the spring and fall, are apparent in the Bronx and Chittenden sites. Variability in the
median values is less pronounced. These sites do not illustrate national trends, but they do
collectively suggest that there can be seasonal variation in ambient air Pb concentration
within sites consistent with findings from the 1970s and 1980s (U.S. EPA. 1986a).
National trends in monthly concentrations, provided in the Chapter 2 Appendix (in Figure
2-58. Figure 2-59. Figure 2-60. and Figure 2-61). do not illustrate variability because
inter-site variability is averaged out.
D[
][
] [
] C
3E
3E
:
Legend: Top left panel: Lake County, CA 060333010, Top right panel: Bronx County, NY 360050110,
Bottom left panel: Monroe County, NY 360551007, and Bottom right panel: Chittenden, VT 500070007.
Notes: Data were not available for all years at all sites. IMPROVE sites were chosen where at least three years of data were
available during 2001 -2011. Boxplots are clipped at all but the Bronx site to improve illustration of the variability among the monthly
interquartile ranges.
Figure 2-22 Boxplots of average monthly Pb-PM2 5 concentrations measured
at four IMPROVE sites, 2001-2010.
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Other data regarding seasonal variability of ambient air Pb concentrations have been
limited. Laidlaw et al. (2012) also explored the seasonal variability of Pb-PM25 at four
cities (Birmingham, AL, Chicago, IL, Detroit, MI, and Pittsburgh, PA) using data from
the Interagency Monitoring of Protected Visual Environments (IMPROVE) network.
They observed a strong seasonal pattern with elevated Pb-PM2 5 levels in the summer
compared with the winter for all four cities. Likewise, Harrison and Yin (2010) observed
that winter background concentrations of Pb were 88% and 81% of summer background
concentrations for urban and rural settings, respectively. In addition to the data presented
above, monthly average Pb concentrations averaged across sites from the TSP, NATTS,
and CSN networks are provided in the Chapter 2 Appendix (Section 2.8.3).
2.5.3 Size Distribution of Pb-Bearing PM
The diverse nature of the main source types of ambient air Pb contributes to variations in
Pb-PM size distribution. Such variation in the size distribution, along with size-dependent
biases in Pb-TSP collection efficiency (Section 2.4.1.1). can lead to uncertainties in the
interpretation of results from Pb-PM measurements. Accordingly, depending on the
locations and magnitudes of nearby sources, ambient air Pb may be 1) mainly Pb in PMi0
and PM25, for which good sampler performance is well established, 2) Pb-PM with a size
distribution that ranges up to slightly larger than 10 (im, in which case the existing
Pb-TSP FRM could potentially be subject to wind related bias, or 3) a Pb-PM size range
that extends well above 10 (im, or too large to be efficiently collected even by an
improved Pb-TSP method. In the latter case, air sampling is likely to be less
representative of actual concentrations of Pb. The role of ambient air Pb size distribution
on human exposure, along with the role of the size distribution of Pb in soil and dust, is
described in Section 3.1.1.1.
Because atmospheric lifetime is dependent on particle size, as described in
Section 2.3.1.3 and in the U.S. EPA 2009 PM ISA (2009a). TSP sampling is likely to be
representative only on a very small spatial scale. Ultra-coarse particles have a sharp
concentration gradient with distance from source, because coarser particles have greater
settling velocities. Hence, concentrations of particles larger than 10 (im are likely to be
very spatially and temporally heterogeneous compared with finer particles (U.S. EPA.
2009a; Hinds. 1999). As a consequence, in locations near sources of ultra-course
particles, measurements may reflect true concentrations only in small areas in close
proximity to the monitor. This issue has been thoroughly discussed in the 2006 Pb AQCD
(U.S. EPA. 2006b). as well as in the 1977 Pb AQCD (U.S. EPA. 1977).
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Size-selective monitoring data from AQS and the literature are examined in this section.
Size distribution data enhance understanding of the relationship between sources and
characteristics of airborne Pb-bearing PM and hence inform monitoring strategies.
Several studies in the literature since the last review have been designed to characterize
the size distribution of Pb concentrations in the vicinity of sources. In the following
subsections, the currently available information is presented for locations in the vicinity
of industrial sources (active and closed), near roadways, and in other urban and rural
environments.
2.5.3.1 Co-located Monitoring Data Analysis
This section employs AQS data for Pb concentrations from co-located TSP, PMi0, and/or
PM25 monitors to analyze correlations and ratios of concentrations obtained from the
different monitors. These data were used because relationships among the monitors
provide information about the nature of Pb-bearing PM at different locations (e.g.,
whether the mode is in the fine or coarse fraction). Correlations indicate the extent to
which the size fractions vary together in time, and ratios signify the average proportion of
the smaller fraction to the larger fraction (e.g., the ratio of PM2 5 to PMi0 concentrations).
Estimation of the size distribution of Pb-bearing PM is possible at a limited number of
monitoring sites where monitors having different size-selective cut-points are co-located.
Data for correlations between concentrations at co-located monitors having different size
cuts and average ratios of concentrations from these co-located monitors are available per
co-location site in Table 2-26 in the Chapter 2 Appendix (Section 2.8). and a summary of
these data is provided in Table 2-8. To ensure quality of the comparisons, they were
limited to data from sites which had at least 30 pairs of co-located samples, with both
samples above the MDL and where both monitors reported data at STP.
The collective size cut comparison data illustrate that the correlations and concentration
ratios of Pb-TSP with the Pb-PMi0 and Pb-PM25 fractions are moderate, with less
correspondence of Pb-PM25 with Pb-TSP compared with Pb-PMi0. The findings indicate
that, on average, 81% of Pb-TSP is in the Pb-PM10 fraction, 50% is in the Pb-PM25
fraction, and 74% of the Pb-PMi0 is in the Pb-PM2 5 fraction (assuming no bias in the
Pb-PM measurements, which may not be a reasonable assumption based on Section
2.4.1). However, for co-located pairs of Pb-TSP with Pb-PMi0 or Pb-PM25, the ranges of
correlations and ratios were large, indicating substantial spatiotemporal variability. There
appeared to be little difference between urban and suburban correlations and
concentration ratios. For three co-located Pb-PM10 :Pb-TSP pairs in Wichita, KS, the
average concentration ratios greater than one were observed. This suggests that some
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portion of the particles captured by the PM10 sampler were not collected by the TSP
sampler, perhaps as a result of instrument biases, as discussed in Section 2.4.1. Likewise,
if such a bias is consistent across monitors, it is possible that, even for average ratios less
than one, the average ratios would be lower if the particles were sampled more efficiently
with the TSP monitor.
Table 2-8 Summary of comparison data for co-located ambient air Pb monitors.
Monitors3
Correlation
N
Standard
Average Deviation
Range
Average Ratio
Standard
Average Deviation
Range
All Sites
Pb-PM-io vs.
Pb-PM2.5 vs.
Pb-PM2.5 vs.
Pb-TSP
Pb-TSP
Pb-PM10
36
20
28
0.74
0.62
0.91
0,
0,
0,
.23
.28
.12
0.
0.
0.
13-0.99
11-0.96
50-0.99
0.81
0.50
0.74
0.19
0.12
0.06
0,
0,
0,
.38-1.28
.30-0.73
.59-0.90
Urban and City Center
Pb-PM-io vs.
Pb-PM2.5 vs.
Pb-PM2.5 vs.
Pb-TSP
Pb-TSP
Pb-PM-io
18
12
14
0.76
0.58
0.89
0,
0,
0,
.20
.31
.15
0.
0.
0.
40-0.99
11-0.96
50-0.99
0.80
0.50
0.74
0.12
0.09
0.06
0,
0,
0,
.57-0.99
.36-0.62
.69-0.90
Suburban
Pb-PM-io vs.
Pb-PM2.5 vs.
Pb-PM2.5 vs.
Pb-TSP
Pb-TSP
Pb-PM-io
18
8
12
0.73
0.69
0.92
0,
0,
0,
.26
.22
.08
0.
0.
0.
13-0.96
,31-0.91
,74-0.99
0.81
0.51
0.73
0.25
0.15
0.07
0,
0,
0,
.38-1.28
.30-0.73
.59-0.82
aNote: For comparability, they were limited to data from sites which had at least 30 pairs of co-located samples, with both samples
above the MDL and where both monitors reported data at STP. N: sample size, PM: participate matter, TSP: total suspended
particulate matter.
2.5.3.2 Studies of Pb-bearing PM Size Distribution in the
Literature
The size distribution of Pb-bearing PM has changed over time and by site. Table 2-9 is
reproduced from Cho et al. (2011). which reviewed studies of the size distribution of
Pb-bearing PM. Studies included in Cho et al. (2011) from the late 1960s to the early
1980s reported substantially higher Pb concentrations compared with current levels.
Traffic-related emissions produced higher contributions from Pb-PM2 5 compared with
industrial emissions. More recent studies from the 1990s and 2000s illustrated variability
in the size distribution regardless of whether the source was traffic or industrial. Cho et
al. (2011) concluded that the size distribution appears to have shifted after the 1980s,
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with the mode appearing to fall somewhere between 2.5 pirn and 10 pirn, compared with
previous estimations of a primary mode smaller than 2.5 pirn; however, Cho et al. (2011)
maintained that additional data are needed to improve characterization of the Pb-PM size
distribution. Metadata and size distribution data from cited studies are provided in Table
2-27 and Table 2-28. respectively, of the Chapter 2 Appendix (Section 2.8).
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Table 2-9 Summary of studies reporting Pb size distribution in the peer-reviewed literature.
Study
Lee et al. (1968)
Lee et al. (1972)
Dorn et al. (1976)a'b
Alpert and Hopke
(1981)
Holsen et al. (1993)
Sweet and Gatz
(1998)h
Location or
Site Type
Urban
Suburban
Chicago, IL
Cincinnati, OH
Denver, CO
Philadelphia, PA
St. Louis, MO
Washington, DC
Near a Pb smelter
Year
Winter
Summer
Control site
Year
Winter
Summer
Urban
Urban
Rural
Lake Michigan
Lake Erie
Lake Michigan
Lake Superior
Pb-TSP Pb-PMio
(ug/m3) (ug/m3)
-
-
: :
-
-
1.04 0.91
1.76 1.69
0.78 0.54
0.11 0.09
0.10 0.10
0.08 0.08
0.913
0.0257
0.0052
0.0112
0.0009
0.0014
0.0013
Pb-PM2.s Pb-PMi
(ug/m3) (ug/m3)
-
-
: :
-
-
0.47 0.27
0.84 0.46
0.32 0.18
0.06 0.04
0.07 0.04
0.04 0.03
0.720
0.0189
0.0043
0.0091
0.0070
0.0021
0.0031
Pb-PMio
Pb-TSP
0.93
1.00
0.99
0.98
0.98
0.99
0.88
0.96
0.69
0.83
0.93
0.94
-
-
-
-
-
Pb-PM25
Pb-TSP
0.83
0.88
0.88
0.87
0.81
0.90
0.45
0.48
0.41
0.52
0.70
0.51
0.79
-
-
-
-
Pb-PMi
Pb-TSP
0.75
0.65
0.59
0.72
0.70
0.70
0.62
0.74
0.26
0.26
0.23
0.32
0.36
0.34
-
-
-
-
-
Pb-PM25
Pb-PMio
0.89
0.88
0.89
0.89
0.83
0.91
0.51
0.50
0.59
0.63
0.75
0.54
-
0.69
0.92
0.81
-
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Table 2-9 (Continued): Summary of studies reporting Pb size distribution in the peer-reviewed literature.
Study
Singh et al. (2002)
Harrison et al.
(2003)
Lough et al. (2005)
Zereini et al. (2005)
Goforth and
Christoforou (2006)
Sabin et al. (2006b)
Wang et al.
(2006d)a
Dall'Osto et al.
(2QQ8)a'g
Bruggemann et al.
(2009)
Location or
Site Type
Traffic + Industrial
Receptor
9 m from a highway
Traffic Tunnel
Main street
Side street
Rural
Rural
10 m from hwy
Urban bkg
Traffic + Industrial
Near a large steelwork
site + major motorway
Curbside of a busy
street
Pb-TSP
(ug/m3)
_
-
0.0326 f
0.0126f
0.01 16f
0.0150
0.0200
0.0110
0.0045
0.0306
-
Pb-PMio
(ug/m3)
0.0069
0.0039
0.0274
-
-
0.0132
0.0091
0.0044
0.0290
0.0169
Pb-PM25 Pb-PMi Pb-PMio Pb-PM25 Pb-PMi
(ug/mj) (ug/mj) Pb-TSP Pb-TSP Pb-TSP
0.0059 0.0051 - - 0.67C
0.0021 0.0017 - - 0.41C
0.98 0.89 0.80
0.85d 0.39d 0.20d
- - - 0.45 c
- - - 0.60 c
- - - 0.64C
0.0061 - - 0.41
0.66
0.83
0.0031 0.0017 0.99 0.69 0.37
0.0245 0.0140 0.95 0.80 0.46
0.0154 0.0120 - - 0.71C
Pb-PM25
Pb-PMio
0.86
0.60
0.90
0.46d
0.1 T
0.59
0.78
0.82
-
0.69
0.84
0.86
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Table 2-9 (Continued): Summary of studies reporting Pb size distribution in the peer-reviewed literature.
Study
Zota et al. (2009)
Makkonen et al.
(2010)
Location or
Site Type
Near mine waste
Traffic + Mine waste
Upwind
Rural bkg
No wildfire
Wildfire
Pb-TSP Pb-PMio
(ug/m3) ("g/m3)
0.0114
0.0052
0.0030
0.0099
0.0153
Pb-PM25
(ug/m3)
0.0035
0.0022
0.0019
0.0055
Pb-PMi
(ug/m3)
-
0.0035
0.0097
Pb-PMio Pb-PM25 Pb-PMi
Pb-TSP Pb-TSP Pb-TSP
- - -
0.35C
0.64C
Pb-PM25
Pb-PMio
0.31
0.42
0.63
0.55
aTSP calculated as a sum of all size fractions.
bPM >0.43 urn.
CPM,:PMW.
Estimated from mass emissions distribution measured using MOUDIs.
eUniversity of Wisconsin samplers.
fPM >0.22 urn.
90.10um
-------
Airborne Pb near Metals Industries
Differences among size distributions have been noted for studies contrasting industrial
and background sites. Yi et al. (2006) collected Pb-PM size distribution in an industrial
area of Jersey City, NJ and contrasted it with the Pb-PM size distribution in suburban
New Brunswick, NJ, which is influenced only by traffic. Yi et al. (2006) sampled size
distribution for Pb-bearing particles with a MOUDI (cut point range: 0.18-18 (im) along
with a coarse particle rotary impactor (CPRI) collecting particles ranging in size from
14.4-100 (im. In the industrial area, 27% of Pb-PM were larger than PM10 (avg. Pb-TSP:
9.7 ng/m3), while in the suburban area 7% of Pb-PM were larger than PMi0 (avg.
Pb-TSP: 6.6 ng/m3). Singh et al. (2002) used a MOUDI to measure the mass distribution
of Pb-PMio in the coarse and fine PM size ranges (cut points range: 0.10-10 (im) for the
Downey site along the Alameda industrial corridor in Los Angeles, CA and a site
approximately 90 km downwind in Riverside, CA. At the industrial site, the Pb-PMi0
size distribution was unimodal with a concentration peak in the 100-350 nm size range.
The sum of the geometric mean concentrations in each size bin was 13 ng/m3 for the
Downey data. At the downwind site, a bimodal distribution was observed with peaks in
the 2.5-10 um bin and the 350 nm-1 urn bin. The sum of the geometric mean
concentrations in each size bin was 7 ng/m3 for the Riverside data. The authors suggested
that higher wind speeds in Riverside compared with the Downey site are effective in
resuspending larger particles from the ground to create a peak in the coarse mode of the
distribution.
Recent studies have indicated temporal variation in the size distribution given differences
among wind direction and industrial production. Bein et al. (2006) measured the size
distribution of PM containing Pb from the Pittsburgh Supersite using rapid single particle
mass spectrometry and a MOUDI. Source apportionment illustrated that Pb was
contained in a sub-population of particles of almost every major particle-containing class
in this study, emanating from point sources including fuel combustion, steel processing,
incinerators, foundries, battery manufacturing, and glass manufacturing (Pekney et al..
2006). Bein et al.'s (2006) measurements yielded different results on different days, with
a bimodal distribution with modes around 140 nm and 750 nm during an October, 2001
measurement and a single dominant mode around 800 nm during a March, 2002
measurement. Differences in the size distributions could have been related to differences
among wind speed, wind direction, and source contributions on the respective dates.
Weitkamp et al. (2005) used a HI-VOL sampler to measure Pb-bearing PM2s
concentrations across the river from a coke plant in the Pittsburgh, PA area and analyzed
the data with ICP-MS. Pb comprised 0.088% of the PM25 mass, and the mode of the size
distribution (measured overall but not specifically for Pb) was observed to shift between
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50 nm to 1 nm. Dall'Osto et al. (2008) used a MOUDI (cut points range: 0.196-18 (im) to
measure the size distribution of Pb emissions from a steel works facility in a coastal town
within the United Kingdom (U.K.). The size distribution was multimodal with a primary
mode around 1 urn, a secondary mode around 300 nm, and a very small additional mode
around 5 um. This multimodal distribution was thought to be associated with sintering
and steel working processes, from which Pb was emitted. Pekey et al. (2010) measured
PM25 and PMi0 concentrations in a heavily industrialized area of Kocaeli City, Turkey
using a low-volume PMi0 stack filter unit. They observed PM2 5/PMi0 ratios of 0.60
during summer and 0.45 during winter.
Airborne Pb Near Roadways
Traffic-induced turbulence may be a cause of resuspension of Pb-bearing particles from
deposited contemporaneous wheel weights, industrial emissions, or historic sources. Pb
mass in near-road PM is predominantly associated with the coarse mode (U.S. EPA.
2006b). The Pb fraction in resuspended dust generally ranges from 0.002 to 0.3%, with
the highest fractions observed for paved road dust and lowest for agricultural soil. Sabin
et al. (2006b) compared the size distribution of coarse Pb-PM measured using a Noll
Rotary Impactor at an urban background site and at a location 10 meters from the 1-405
Freeway in the southern California air basin; data from Sabin et al. (2006b) are displayed
in Figure 2-23. For both the urban background and near-road sites, the largest fraction
was from PM sampled below the 6 um cut point, but the near-road Pb-PM distribution
appeared bimodal with a mode in the largest size fraction. Sabin et al. (2006b) point out
that the freeway tends to be a source of very large particles that are dispersed via the
turbulent motion of the vehicular traffic. Song and Gao (2011) used an eight-stage
MOUDI (cut point range: 0.18-18 um) to measure roadside PM, 5 meters from the
New Jersey Turnpike in Carlstadt, NJ and speciated the samples. They observed a
bimodal distribution of the Pb concentration in summer and atrimodal distribution in
winter. 85% of the Pb-PM mass was measured as PM2 5 during the summer, and 68% was
measured as PM2 5 in the winter. Similarly, Zereini et al. (2005) observed that roughly
80% of particle-bound Pb measured with a MOUDI was smaller than 5.8 um for an urban
main street, and more than 90% were smaller than 5.8 um for a rural area included in that
study. However, in a study of automotive emissions in a traffic tunnel, Lough et al.
(2005) observed that 85% of Pb measured with a MOUDI was in the PM10, with just 39%
in the PM2 5 fraction and 20% in the PMi fraction. In a near-road study conducted in
Raleigh, NC with a 13-stage low-pressure impactor, Hays et al. (2011) note that the
proportion of Pb within particles in ultrafine, fine, and coarse size ranges was the same at
50 mg/kg; similar to Lough et al. (2005). mass concentrations were measured by Hays et
al. (2011) to be 0.4 ± 0.4 ng/m3, 1.4 ± 0.6 ng/m3, and 0.1 ± 0.02 ng/m3 for PM10_25,
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PM 2.5-o.i, and PM0.i, respectively. The Pb-PM10 samples from Hays et al. (2011) were
highly correlated with As samples (p = 0.7, p <0.0001); both Pb and As are found in
wheel weights (see Section 2.2.2.6). Hays et al. (2011) did not report correlations
between Pb and As for smaller size fractions, but they did state that the correlations for
other size fractions were lower compared with Pb-PM10 and As. Likewise, the Pb
samples were not well correlated with crustal elements in the coarse size distribution, so
it is more likely that resuspended Pb originated from contemporary roadway sources
rather than historic Pb on-road gasoline emissions. Chen et al. (201 Ob) measured Pb in
PMio-2.5, PM2.5_0.i, and PM0.i using a MOUDI at a roadside location and in atunnel in
Taipei, Taiwan in 2008. While roadside and tunnel concentrations of PMi0 and PM2.5
were roughly equivalent around 20-30 ng/m3, Pb in PM0.i was approximately 15 times
higher in the tunnel (during the hours 9:00 a.m. - 9:00 p.m.) than by the roadside (tunnel:
20 ng/m3; roadside: 1 ng/m3). The authors suggest that particle-bound Pb was emitted
from on-road gasoline and diesel engines. This could possibly be attributed to trace levels
of Pb in diesel fuel and lubricating oil. Birmili et al. (2006) compared concentrations of
Pb in PM measured with a Sierra-Anderson high volume cascade impactor at various
traffic and background sites in Birmingham, U.K. Birmili et al. (2006) captured PM at the
stage below a 0.5 urn cutpoint and on the 1.5-3.0 urn stage for near-road, in a traffic
tunnel, and remote and urban background sites. The highest concentrations were
measured in the tunnel, at 3.3 ng/m3 for Pb-PM0.5 and 10 ng/m3 for Pb-PMi 5.3 0. In
contrast, urban background was more enriched in the finer size fraction, with
concentrations of 5.4 ng/m3 for Pb-PM0s and 0.84 ng/m3 for Pb-PMi 5.3 0. Remote
background concentrations were on 0.16 ng/m3 for Pb-PM0s and 0.03 ng/m3 for
Pb-PM 1.5.3.0. Bruggemann et al. (2009) measured roadside distribution of Pb in PM in
Dresden, Germany using a 5-stage Berner-type low-pressure impactor to analyze the
effect of season and direction of the air mass. For all data combined as well as for data
broken down by season or by wind direction, it was found that the data followed a
unimodal distribution with a peak at the 0.42-1.0 urn size bin with roadside
measurements averaging 13-22 ng/m3, depending on wind direction. Evidence of Pb in
road dust related to near road ambient air Pb concentrations is described in Section 2.6.1
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Urban Background
Near-Road
<6 u.m
6-11 urn
11-20 urn
20-29 urn
>29 u.m
Source: Adapted, with permission of Elsevier Publishing, Sabin et al. (2006b).
Figure 2-23 Comparison of urban background and near-road size fractions of
Pb-bearing PM.
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Several studies have suggested that near-road ambient air Pb samples are derived from
sources other than from the road. Harrison et al. (2003) measured the distribution of Pb in
PM10 at a roadside sampler in Birmingham, U.K. using a MOUDI fitted with only
stages 1, 2, 4, and 8 with outpoints of 10 pirn, 2 um, 1 um, and 0.2 um. The size
distribution was unimodal with approximately 2% of the Pb mass (totaling 26.5 ng/m3)
above the 10 um cut point, 12% of the mass in the 2-10 pirn bin, 8% in the 1-2 um bin,
53% of the Pb mass in the 0.2-1 um bin, and 25% collected below the 0.2 um cut point.
Regression analysis against NOX concentration in the Harrison et al. (2003) paper
provided a weak indication that Pb-PM0 2 was associated with NOX (P = 0.067,
R2 = 0.38) as well as PM10 (P = 0.26, R2 = 0.35). Bruggemann et al. (2009) observed a
unimodal Pb size distribution with 51% of the mass in the 0.42-1.2 um size bin.
Observed Pb-PM10 concentration was 17 ng/m3. During winter, Pb concentrations were
more than twice as high as during the summer (winter: 24 ng/m3; summer: 10 ng/m3), and
they were also higher when winds blew from the east (0.42-1.2 um mode, east: 60 ng/m3;
west: 25 ng/m3). Bruggemann et al. (2009) suggested that this finding reflected coal
burning sources dominating Pb emissions rather than road dust resuspension during
winter. Wang et al. (2006d) used a nine-stage cascade impactor (cut point range:
0.43-11 um) to measure the Pb-PM size distribution in a heavily trafficked area of
Kanazawa, Japan with incineration and generation facilities nearby. They observed a
bimodal distribution with modes at the 0.65-1.1 um and the 3.3-4.7 um size bins.
Average concentration in the coarse mode was 2.1 ng/m3, while fine-mode average
concentration was 3.7 ng/m3. Wang et al.'s (2006d) source apportionment work in this
study suggested that the fine mode derives from incineration and combustion of oil and
coal.
Airborne Pb at Other Urban and Rural Sites
Spatial and temporal concentration variability is also reflected in varying Pb-PM size
distributions within and between cities. Martuzevicius et al. (2004) measured the size
distribution of Pb in Cincinnati, OH at the city center site using a MOUDI and showed it
to be bimodal with a primary peak at 0.56 um and a slightly smaller secondary peak at
5.6 um. Using high volume samplers, Moreno et al. (2008) measured Pb concentrations
in PM2 5 and PMi0 at urban, suburban, and rural sites around Mexico City, Mexico to
illustrate differences among the land use categories. At the urban site, PM2,5/PMi0 ratios
were 0.51 during the day and 0.57 at night (Pb-PMi0 was 59 ng/m3 and 162 ng/m3,
respectively). At the suburban site, Pb-PM25/Pb-PMi0 ratios were 0.63 during the day
and 0.81 at night (Pb-PMi0 was 24 ng/m3 and 42 ng/m3, respectively). Goforth and
Christoforou (2006) measured Pb-TSP and Pb-PM2 5 with a high volume cyclone
separator in rural Georgia and observed a Pb-PM2 5 concentration of 6 ng/m3 and a
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Pb-TSP concentration of 15 ng/m3. Makkonen et al. (2010) measured concentrations of
Pb-PMi, Pb-PM25, and Pb-PMi0 during a spate of wildfires in rural southeastern Finland
with high volume size-selective samplers. They found that the ratio of Pb-PM^Pb-PM^
varied substantially from day to day (examples provided of 64% on 8/14/07 and 35% on
8/25/07, with Pb-PM2 5/Pb-PM10 ratio of 51% on 8/25/07), and they attributed the highest
concentrations to long-range transport of wildfire emissions via southerly winds;
variability in concentration and ratios was related to shifting wind conditions.
2.5.4 Pb Concentrations in a Multipollutant Context
The correlations between Pb and copollutant concentrations were investigated because
correlation may indicate commonality of sources among the pollutants. For example,
correlation between Pb and SO2 may suggest common industrial sources. Correlation
between Pb and NO2 or CO may suggest roadway sources, such as trace Pb in unleaded
on-road gasoline or resuspension of material from pulverized wheel weights or
contaminated soil. Additionally, seasonality can influence correlations, potentially from
differences among sources or the contaminants' responses to climate differences.
Pb concentrations exhibit varying degrees of association with other criteria pollutant
concentrations. At most sites, Pb monitors are co-located with monitors for other criteria
pollutants, but monitoring the full suite of criteria pollutants at a single monitoring site is
rare. As a result, the number of observations for each copollutant varies. Pearson
correlations of monitored non-source Pb-TSP concentrations with concentrations of other
criteria pollutants are summarized in Figure 2-24 for 2008-2010 data for 46 Pb-TSP
monitors at which data were above MDL and more than 30 data pairs were measured at
each point. Seasonal co-pollutant measurement data from the literature are also provided
in the Chapter 2 Appendix (see Figure 2-63 through Figure 2-67).
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CO
PM10
NO2
PM2.5
S02
03
0
o oo
o
o oo o
O O O CTEBD
VlOGm-VKXMSiOOa
OQB9GQDO O O
OCD
-------
Zn —
EC -
PM2 SMass —
Cu -
OCmass —
OCadj —
Br —
K —
K+ —
Fe —
NH4+ —
Mn —
NH4NO3 —
NH4 2SO4 —
Ca -
Crustal —
Mg —
Se —
Au -
Si -
Sr -
Ti —
La —
V —
Al —
W -
Rb -
Mo —
Nb —
Ta —
Eu -
Sn -
Ce —
Hg —
Cs -
Sm —
Ir —
Hf -
h 1 , h 4i7m n
o i 1 i h 1 n
i 1 i i 1 no nr> n
i- 1 i i 1 mon«»
i 1 i i n
i- 1 i i- «tn nr> o n
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i ^ i i- iCXJD O
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i H I I 1 r>
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i 1 i \- 1 ess) n nno
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O •- "I It" - H
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o m o in o
^- o o o ^
Pearson R
Note: Correlations were calculated from available data when data were above MDL and there were at least 30 data pairs available
for comparison; organic carbon (OC) samples were blank-adjusted.
Figure 2-25 Pearson correlations of monitored Pb-PM2.s concentration with
copollutant concentrations, 2008-2010.
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Associations between Pb-PM2 5 and other species were generally low-to-moderate. The
strongest association was with zinc (Zn) (median R = 0.38). Elemental carbon (EC),
PM2 5 mass, copper (Cu), organic carbon (OC) mass, bromine (Br), and potassium (K)
also exhibited low-to-moderate associations with Pb-PM2 5 concentrations (median
R = 0.27 to 0.33). Such correlations may suggest some common sources affecting the
pollutants, as described in Section 2.2.2. For example, correlation with EC and OC mass
may be diagnostic indicators of some crustal, general combustion, wood burning,
industrial emission, and coal combustion processes. Piston-engine aircraft emit Pb as
PbBr2 so this source may explain the weak covariation in Pb and Br concentrations at the
CSN sites. At the same time, these species must have other disparate sources that drive
the Pearson correlations down.
A few recent studies have used speciation techniques to characterize Pb and other
components of PMi0, PM25, and PMi. Pingitore et al. (2009) used XAFS to speciate air
samples obtained near a defunct smelter in El Paso, TX, in 1999 and 2005 and found that
air Pb-TSP concentrations of 0.10 to 0.50 (ig/m3 could largely be attributed to Pb-humate.
Similarly, Laidlaw et al. (2012) observed statistically significant associations between
ambient air Pb-PM2 5 and ambient air soil in Pittsburgh, PA, Chicago, IL, Detroit, MI,
and Birmingham, AL (R2 = 0.31-0.49, p <0.01). Together, these results suggest a link
between soil resuspension and Pb-TSP levels
Murphy et al. (2008) studied weekly patterns of metals and other aerosol components
using data collected from 2000 to 2006 at IMPROVE sites. The authors concluded that
Pb concentrations were impacted by piston aircraft emissions. They reached this
conclusion because, in contrast to other species, Pb was elevated on weekends when there
is typically a peak in general aviation flights. The authors also note that Zn and Pb were
highly correlated in atmospheric samples, and they suggest that this is due to similar
sources (i.e., electric utility and industrial sources). Murphy et al. (2007) also carried out
a detailed study of the distribution of Pb in single atmospheric particles. During the fifth
Cloud and Aerosol Characterization Experiment in the Free Troposphere (CLACE 5)
campaign conducted at the Jungfraujoch High Altitude Research Station, Switzerland,
about 5% of analyzed aerosol particles in PMj contained Pb. Of these, 35% had a relative
signal for Pb greater than 5% of the total mass spectrum measured by an aerosol time of
flight mass spectrometer (ATOFMS). These "high Pb" particles also contained one or
more positive ions (e.g., of Na+, Mg2+, A13+, K+, Fe3+, Zn2+, Mo6+, Ag+, Ba2+). Sulfate
fragments were present in 99% of the negative ion spectra associated with high Pb
particles, and 50% also contained nitrite and nitrate. About 80% contained positive and/or
negative polarity organic fragments. The average aerodynamic diameter of the Pb-rich
particles (500 nm) was larger than the background aerosol (350 nm) but none had a
diameter less than 300 nm. Murphy et al. (2007) suggest that this mixture can be
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attributed to combined emissions from combustion (e.g., Pb and organics) and industry
(e.g., Pb sulfates).
2.5.5 Background Pb Concentrations
The 2006 and 1986 Pb AQCDs evaluated evidence on Pb emissions from natural sources,
such as volcanoes, sea-salt spray, biogenic sources, wild forest fires and wind-borne soil
particles in rural areas without elevated Pb soil concentration. The 1986 Pb AQCD
concluded that the natural particulate Pb concentration was probably lower than the
concentration of 0.076 ng/m3 reported at the South Pole and estimated 0.05 ng/m3 to be
the natural background concentration (U.S. EPA. 1986a). A 1980 National Academy of
Sciences (NAS) report estimated that average natural background levels of airborne Pb
might range from 0.02 to 0.5 ng/m3 (NAS. 1980).
Global transport can carry airborne Pb to remote areas with no industrial activity, thus it
is difficult to estimate a natural background concentration of Pb. Hong et al. (1994) found
that Pb concentrations in Greenland ice cores remained nearly constant (at about 0.55 pg
Pb/g ice) from about 7,760 years ago to about 3,000 years ago. Ratios of Pb to major
crustal elements were not enriched compared with natural background levels in this
section of the ice core suggesting that Pb was natural in origin, produced by rock and soil
dust. At about 2,500 years ago, Pb concentrations started to increase (to about 100
pg Pb/g snow averaged from 1930 to 1990) (Boutron et al., 1991) corresponding to an
enrichment of-200 times natural background levels. McConnell and Edwards (2008)
also noted elevated Pb levels in Greenland ice cores, with high correlation to black
carbon (BC), cadmium (Cd), sulfur (S), and thallium (Tl) during the period 1860-1940,
suggesting coal combustion sources from North America. Osterberg et al. (2008)
observed elevated Pb levels in a 1970-1998 ice core from Mt. Logan, Canada, indicating
elevated Pb levels corresponding to increased industrial activity in Asia during this time
period.
Measurements of Pb from IMPROVE sites and source apportionment modeling have
been used to assess the potential input from intercontinental transport. Liu et al. (2003)
used positive matrix factorization to attribute sources of Asian dust to the measurements
at two western IMPROVE sites at high elevations, Crater Lake (Oregon) and Lassen
Volcanic Park (California) from 1988 to 2000. Geometric mean Pb concentrations of 0.34
and 0.48 ng/m3 were found in the samples with only a few percent of these values
attributable to transport from Asia. No enrichment in Pb and other metals (As, Cr, Cu, Ni,
Pb, V and Zn) above reference Asian-dust material was found. Their results suggest
either that arriving air masses did not entrain contributions from Asian pollution sources
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or that these contributions were preferentially scrubbed out during transport. Large
enrichments in sulfur (S) were found, however, which might have been due to pollution
sources but also due to model artifacts. However, other studies have found some evidence
of trans-Pacific transport. Murphy et al. (2007) measured single Pb particles off the coast
of California (using a National Oceanic and Atmospheric Administration [NOAA]
aircraft elevated more than 2 km above ground level). Given the elevation of the
measurement and the timing of trans-Pacific plume events, the authors concluded that
these Pb-bearing PM2 5 originated in Asia. They also noted Pb/Zn ratios in PM2 5 at the
Mount Zirkel, CO IMPROVE site of 0.6 corresponding to measurements at Mauna Loa,
HI in spring, when measurements at other times of year produced Pb/Zn ratios of 0.3-0.4.
Ewing et al. (2010) used time series analysis of Pb isotope measurements to estimate
Asian and local contributions to Pb-PM2 5 concentrations measured at two observatories
near San Francisco, CA. They estimated a springtime contribution of Asian dust to
Pb-PM2.5 measurements. In both the Murphy et al. (2007) and Ewing et al. (2010)
studies, the authors conclude that the Asian contribution is still generally less than 1
ng/m3.
The use of data for PM25, PMi0.2 5, and PMi0 from monitoring sites in the East will
generally result in gross overestimates of background concentrations because
anthropogenic sources will cause extensive contamination. Intercontinental transport of
African dust contributes to PM and is observed mainly in the Southeast but is apparent on
an episodic basis elsewhere in the eastern U.S. [see e.g., 2004 PM AQCD (U.S. EPA.
2004) and 2009 PM ISA (U.S. EPA. 2009a)1. Data obtained at four eastern IMPROVE
sites ([1] Moosehorn National Wildlife Refuge, ME; [2] Acadia National Park, ME;
[3] Swanquarter, NC; and [4] Cape Remain National Wildlife Refuge, SC) from 2007 to
2009 indicate a median Pb-PM2 5 concentration of 1.0 ng/m3 with a 95th percentile value
of 2.5 ng/m3. As noted above, these sites are likely to be affected by upwind
anthropogenic sources within the U.S.
Rough estimates for the natural source of Pb in different size fractions of Pb-PM can be
made by multiplying the abundance of Pb in soils by the crustal component of PM in the
different size fractions. The mean abundance of Pb in surface rocks is -20 mg/kg (Potts
and Webb. 1992): the 2006 Pb AQCD (U.S. EPA. 2006b) reported Pb concentrations in
different types of rocks to range from 3.5 to 32 mg/kg (Reuer and Weiss. 2002). There is
substantial variation with location depending on composition, in particular on the
abundances of uranium (U) and thorium (Th), since Pb is produced mainly by radioactive
decay of these elements. The mean Pb concentration of 863 soil samples taken across the
U.S. at 2 meters depth is ~16 mg/kg; this value was derived by sampling residual Pb of
the weathered rocks on which they formed [Wedepohl (1978) and references therein].
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Concentrations of the Pb content of soils can be used with estimates of the crustal
component of PM25, PMi0-2.5 (which is mainly crustal), PMi0, and TSP produced by
wind erosion of natural surfaces to estimate contributions to Pb concentrations in these
size fractions. U.S. annual average PMi0 concentrations in some arid counties most
affected by windblown dust in the western U.S. are -20 ug/m3. If it is assumed that these
levels of PMio are entirely due to natural wind erosion without any anthropogenic
contribution and that the Pb concentration in all airborne size fractions if the same as the
Pb concentration in bulk soil or surface rock, an estimate of ~0.3 ng/m3 for the
contribution of wind erosion on natural surfaces to Pb in PM10 is obtained; however, it
must be observed that the natural contribution is probably lower than this estimate. An
assumed ratio 3.5 for TSP to PMi0 in dust storms, derived by Bacon et al. (2011).
indicates a contribution of-1 ng/m3 for Pb from natural sources in TSP. The more recent
estimate indicates that background airborne Pb concentrations are well below current
ambient concentrations. These estimates exceed estimates of natural background
presented in the 1986 AQCD (U.S. EPA. 1986a) and the National Academy of Sciences
Report (NAS, 1980) by a factor of 2 to 50. Hence, a plausible range of natural
background airborne Pb is 0.02 to 1 ng/m3.
2.6 Ambient Pb Concentrations in Non-Air Media and Biota
There have been some major recent research efforts to characterize geographic and
temporal trends in Pb concentrations across a variety of environmental media and biota.
In general these concentrations reflect the decreases observed in atmospheric Pb
concentrations due to reduced on-road Pb emissions.
The 2006 Pb AQCD (U.S. EPA. 2006b) describes several studies showing higher Pb
concentrations in plants grown in Pb contaminated soil related to mine spoils, smelting
operations, sludge amendment, contaminated irrigation water, and Pb containing agro-
chemicals. In general, metal accumulation occurs more readily for Pb salts applied to
soils than for the same quantity of metal in sewage sludge or fly ash. Root uptake is the
dominant means of accumulation, and it is strongly influenced by pH. Root vegetables
are the most strongly affected, and fruits and grains are the least susceptible. More Pb is
also generally found in roots than in other parts of the plant.
The 2006 Pb AQCD (U.S. EPA. 2006b) identified ingestion and water intake as major
routes of Pb exposure for aquatic organisms, and it identified food, drinking water, and
inhalation as major routes of exposure for livestock and terrestrial wildlife. The
2006 Pb AQCD (U.S. EPA. 2006b) reports data from the U.S. Geologic Service National
Water-Quality Assessment (NAWQA), which are updated every ten years. In the
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NAWQA survey, maxima concentrations in surface waters, sediments, and fish tissues
were 30 ug/L, 12,000 mg/kg, and 23 mg/kg, respectively, compared with median values
of 0.50 ug/L, 28 mg/kg, and 0.59 mg/kg. Some of the highest levels of Pb contamination
occur near major sources, like smelters, and fatal doses have been measured in tissue
from sheep and horses near sources. High levels in cattle have also been observed.
Wildlife in urban areas tend to contain higher Pb concentrations than in rural areas, and
higher Pb accumulations have been observed for aquatic organisms living in polluted
coastal zones than in the open sea. Ingestion of deposited Pb-PM on plant surfaces was
consistently observed to be more important than Pb accumulated from soil. Some
important variations between animals have been observed, and ruminants appear to be
less susceptible to Pb uptake than other animals. Uptake of Pb by lowest trophic levels,
including invertebrates, phytoplankton, krill, were described as the most important means
of introduction into food chains. Elevated Pb levels have been observed in aquatic
organisms that feed from sediments when the sediments contain appreciable Pb. In
shrimp, a substantial fraction of Pb can be absorbed from prey, and considerably more
accumulated Pb from food has been observed to be irreversibly retained than is the case
for dissolved Pb from water. These examples all illustrated that substantial Pb uptake by
livestock and wildlife readily occurs in Pb contaminated environments.
2.6.1 Soils
Several studies suggest that soil can act as a reservoir for contemporary and historical Pb
emissions. The importance of soil Pb to human exposure is described in Section 3.1. At
the same time, soils in remote or rural areas tend to have lower Pb concentrations. The
most extensive survey of background soil Pb concentration in the conterminous U.S. was
conducted between 1961 and 1976 and comprised 1,319 non-urban, undisturbed sample
locations, where 250 cm3 of soil was collected at a depth of 20 cm (Shacklette and
Boerngen. 1984). The lower detection limit was 10 mg/kg, and 14% of the 1,319 samples
were below it. The mean Pb concentration was 19.3 mg/kg, the median 15 mg/kg, and the
95th percentile was 50 mg/kg. Sixteen locations had Pb concentrations between 100 and
700 mg/kg. These results were in agreement with 3 previous surveys. When creating the
Ecological Soil Screening Level (Eco-SSL) Guidance and Documents, the U.S. EPA
(U.S. EPA. 2007d. 2003b) augmented these data with observations from an additional 13
studies conducted between 1982 and 1997, most of them limited to one state. The
resulting data were summarized using state means for each of the fifty states. Those
means ranged between 5 and 38.6 mg/kg, with an overall national mean of 18.9 mg/kg.
This is reasonably close to the values reported by Wedepohl (1978) and references
therein with a mean soil Pb concentration of roughly 16 mg/kg when samples were taken
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at 2-meter depths. Biasioli et al. (2006) contrasted urban and rural soils (tested at soil
depths of 0-20 cm) of the same alluvial composition near Torino, Italy to assess the
influence of anthropogenic inputs. The urban soils had a median Pb concentration of
117 mg/kg, while the median Pb concentration for rural soil was 19 mg/kg. Table 2-10
presents data from seven metropolitan areas (Cobb et al., 2006). Differences among the
intraurban concentration ranges illustrate a high level of spatial variability within
individual cities as well as high inter-urban variability. The rural New Orleans site
reported relatively low Pb soil concentrations, and the highest average Pb soil
concentrations were reported for pre-Katrina samples from the city of New Orleans,
although this was not the case for the post-Katrina samples.
Table 2-10 Soil concentrations in various cities, 1992-2005.
City
Baltimore, MD
Miami, FL
Mt. Pleasant, Ml
New Orleans, LA (pre-Katrina)
New Orleans, LA (post-Katrina)
New Orleans, LA (rural outskirts)
St. Louis, MO
Syracuse, NY
Avg Pb Concentration (mg/kg)a
275
320
784
95.9b
11
427
80b
Pb Concentration Range (mg/kg)a
1-10,900
25-1,612
100-840
31.7-5,195
9.84-1,880
4.8-17.3
35-1,860
20-800C
Dry weight basis;
bGeometric mean;
°Range represents 95% of collected values.
Source: Adapted with permission from the American Chemical Society, (Cobb et al., 2006).
In North American forest soils, Pb concentrations have decreased substantially since the
phase out of leaded motor vehicle gasoline. When sampling from the O horizon (often at
0-2 cm), Evans et al. (2005) observed Pb concentrations ranging from 60 to 200 mg/kg in
Vermont, Maine, and Quebec, with lower concentrations in Quebec than in southern
Vermont in 1979, but in 1996 concentrations had decreased to between 32 and 66 mg/kg
with no spatial trend. Johnson and Richter (2010) also observed a substantial decrease in
O-horizon (depth not specified) Pb concentrations in soil between 1978 and 2004 in West
Virginia, Maryland, Pennsylvania, New Jersey, New York, and Connecticut, with a
median change of-65%. However, elevation also appears to be an important factor in
determining whether appreciable decreases in Pb concentration have occurred since the
phase out of leaded gasoline (Kaste et al.. 2006). At sites above 800 meters in the
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northeastern U.S., O-horizon concentrations (depths not specified) ranged from 11 to
29 kg Pb/ha, and little change in Pb concentration was observed between 1980 and 2000.
In contrast, concentrations ranged from 10 to 20 kg Pb/ha at low elevation sites and
decreased to 2 to 10 kg Pb/ha by 2000. This difference was likely due to greater organic
turnover increasing Pb mobility at the lower elevations (Kaste et al.. 2006).
Soil Pb variability depends on the strength and prevalence of nearby sources. Joshi et al.
(2009) observed Pb dust concentrations to be highest at industrial sites (260 mg/kg)
followed by commercial sites (120 mg/kg) and residential sites (60 mg/kg) in Singapore.
Griffith et al. (2002) investigated spatial autocorrelation of soil Pb concentration at three
sites: urban Syracuse, NY (0-10 cm), rural Geul River, The Netherlands (0-5 cm), and an
abandoned Pb Superfund site in Murray, UT (0-5 cm). In both Syracuse and Geul River,
the soil Pb concentrations were not strongly correlated in space, with the exception of soil
obtained near roads, which exhibited less variability. The smelting and shooting areas of
the Superfund site were both demonstrated to have spatial clusters that were well
correlated. Later work on the spatial distribution of metals in Syracuse (sampling depth
not specified) produced similar results for that city (Griffith et al.. 2009). These studies
did not adjust for age of housing, although Griffith et al. (2009) did find that housing age
and Pb co-vary. An association between housing age and soil Pb would likely be
enhanced by such co-variation.
Emissions trends have shown that industrial activities are now one of the largest sources
of soil Pb following phase out of Pb in on-road gasoline. Pruvot et al. (2006) compared
urban and agricultural soils at depths of 0-25 cm near a closed Pb smelter with soils in
similar environments not exposed to smelter emissions in northern France. For samples
near the smelter, Pruvot et al. (2006) observed that median soil Pb levels in lawns were
roughly 2 times higher, while kitchen garden soil Pb concentrations were 10 times higher
and agricultural soil Pb was almost 15 times higher than soil not exposed to smelter
emissions. Bonnard and McKone (2009) reported surface soil Pb concentrations at depths
of 0-20 cm of 66-493 mg/kg outside homes of children living within 1 km of a Pb smelter
in France; air Pb levels reported by Bonnard and McKone (2009) for this town ranged
from 0.025-0.20 (ig/m3. The air samples Pingitore et al. (2009) obtained near a defunct
El Paso, TX smelter (described in Section 2.5.4) found that the air Pb-TSP concentrations
could largely be attributed to Pb-humate, which is created by sorption of Pb onto humic
substances in soil and can be resuspended. Spalinger et al. (2007) compared soil Pb
samples at depths of 0-2.5 cm from surrounding towns with those from the Bunker Hill
Superfund remediation site in Idaho. Median background soil-Pb concentration was
48 mg/kg, while the median soil-Pb concentration at Bunker Hill was 245 mg/kg.
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Recent studies of brownfield soils have shown variable Pb concentrations. Van
Herwijnen et al. (2007) measured soils at depths of 0-2 cm near a defunct Zn smelter in
Avonmouth, U.K. in areas termed low and high contamination by the authors. Total soil
Pb concentration in the low contamination area was 315 mg/kg, while soil Pb
concentration in the high contamination area was 1,688 mg/kg. Deng and Jennings (2006)
tested various Pb extraction methods on soils obtained from over 50 brownfield sites in
the greater Cleveland, OH area at depths of 0-5 cm. Comparison of twelve extraction
methods for three samples produced a range of 1,780-2,636 mg/kg for one sample,
283-491 mg/kg for a second sample, and 273-499 mg/kg for a third sample. Verstraete
and Van Meirvenne (2008) measured Pb in soils at a remediated brownfield site at depths
of 0-5 meters in Belgium and reported average Pb concentrations to be 188 mg/kg and
224 mg/kg in two sampling campaigns. Dermont et al. (2010) fractionated soil sampled at
depths of 0-150 cm by particle size class and measured the Pb concentration in each. Pb
concentrations by size bin were as follows: 125-250 (im: 1,132 mg/kg; 63-125 (im:
1,786 mg/kg; 38-63 (im: 1,712 mg/kg; 20-38 (im: 2,465 mg/kg; 0-20 (im: 3,596 mg/kg.
Hence, the highest concentration was in the smallest soil particle fraction. Bulk Pb
concentration over 0-250 (im particle sizes was 2,168 mg/kg.
Several studies explore the relationship between soil Pb concentration and land use.
Laidlaw and Filippelli (2008) displayed data for Indianapolis, IN showing the Pb
concentration at the soil surface (depths not specified) had a smoothed "bull's eye"
pattern. Cities generally have a similar pattern consisting of larger quantities of Pb
accumulated within the inner city and smaller quantities of Pb in outer cities (i.e., near the
outskirts or suburban areas) (Filippelli and Laidlaw, 2010). Similarly, Filippelli et al.
(2005) reported surface (depths not specified) soil Pb concentration distribution to have a
maximum at the center of Indianapolis, IN, around the location where two interstate
highways intersect, and to decrease with distance away from the center. However, the
spatial distribution of Pb was presumed to reflect contributions from historic sources of
on-road gasoline (Section 2.2.2.6) and Pb paint (Section 2.2.2.7). In this paper, soil Pb
concentrations were also shown to decrease with distance from roadways, but the levels
were roughly four times higher in urban areas compared with suburban areas. This is also
illustrated for urban scale Pb accumulation in New Orleans, LA during 1998-2000 in
Figure 2-26. Brown et al. (2008) also measured soil Pb concentration along three
transects of Lubbock, TX at depths of 0-2 cm and observed that soil Pb decreased with
increasing distance from the city center, which was the oldest part of the city.
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Median Soil Pb (mg/kg)
3-99
100-199
200-299
300-399
400 - 499
500-599
600 - 699
700 - 799
800-899
900-999
1000-1768
Note: At the urban scale, Pb quantities are largest within the inner-city residential communities that surround the Central Business
District where pavement and concrete cover the soil. Note the several orders of magnitude difference between the interior and the
exterior areas of the city. Note that the number on each census tract indicates the number of blood Pb samples taken from that tract
during the six years from which the study data were obtained.
Source: Reprinted with permission of Elsevier Publishing, Mielke et al. (2007a)
Figure 2-26 Map of median Pb content in soil in New Orleans, 1998-2000.
Mielke et al. (2008) compared soil Pb concentrations at depths of 0-2.5 cm for public and
private housing at the center and outer sections of New Orleans and found that median
and maximum soil Pb concentrations were substantially higher in the city center
compared with the outer portions of the city. This study also found that private residences
had higher soil Pb compared with public housing. In a separate study to examine surface
soil Pb loading and concentration on 25 properties at depths of 0-2.5 cm in 25
New Orleans properties that were previously identified as having median soil Pb
concentrations of at least 1,000 mg/kg, Mielke et al. (2007b) reported median and
maxima soil loading of roughly 25,000 and 265,000 (ig/m2, respectively. Median and
maxima surface soil Pb concentrations were observed to be 1,000 and 20,000 mg/kg,
respectively. Clark et al. (2006) performed isotopic analysis on urban garden soils at
depths of 0-10 cm and 30-40 cm in an area of Boston, MA with no large industrial
sources of Pb and estimated that 40-80% of the soil Pb could be attributed to Pb-based
paint while the remainder was attributed to historic Pb on-road gasoline emissions.
Additional discussion of historic sources of Pb is provided in Section 2.2.2.7'. Isotope
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ratios for paint and gasoline references used in the Clark et al. (2006) study were obtained
from Rabinowitz (1986).
Several studies have examined the effects of roadway attributes on Pb content in roadside
dust. Yesilonis et al. (2008) measured metal content in surface soil samples (0-10 cm) at
selected land parcels throughout Baltimore, MD, based on a stratified random sampling
design that accounted for land use factors. They compared soil metals within 100 meters
buffers of roadways and outside those buffers and found that median soil Pb
concentration inside the buffer was significantly higher than median soil Pb concentration
outside the buffers (outside: 38.7 kg/ha; inside: 134 kg/ha; p <0.0001). In an analysis of
the relationship between land use parameters and Pb concentration in soil in Los Angeles,
CA, Wu et al. (2010) observed that soil Pb concentration at depths of 0-2.5 cm was
higher near freeways and major traffic arteries compared with other locations. The
(square-root transformed) age of the building on a sampled land parcel, length of
highway within a 1,000-meter buffer, and length of local road within a 20-meter buffer in
which the sample was obtained were significant predictors of Pb. Home age within 30
meters of a soil sample and road length within 3,000 meters of a road sample were also
shown to be significant predictors of soil Pb concentration in areas not designated to be
near a freeway or major traffic artery. Wu et al. (2010) concluded that both historical
traffic and leaded paint contributed to Pb contamination in soils. However, Wu et al.
(2010) acknowledged uncertainty in historical roadway and traffic count data, which
introduces uncertainty into that conclusion. Study areas were classified as residential,
commercial, park, and industrial (not specific to Pb emissions), although the authors were
not able to distinguish the relative effects of each area on Pb content in roadside dust. Wu
et al. (2010) reported that the highest median measured concentrations of Pb content in
roadside dust were in residential freeway samples (112 mg/kg), followed by residential
arterial samples (98 mg/kg), and industrial freeway samples (90 mg/kg). Additional
sources of Pb to soil near roadways, such as traces of Pb in unleaded gasoline and
Pb-containing wheel weights (described in Section 2.2.2.6) were not considered in this
study. Amato et al. (2009) observed that deposited PM onto roadways, measured as dust
samples, in Barcelona, Spain was differentially enriched with Pb compared with dust
collected at a harbor area. Pb concentration in PM10 was highest at ring roads
(229 mg/kg) and in the city center (225 mg/kg), followed by demolition and construction
sites (177 mg/kg) and near a harbor (100 mg/kg). Roadside dust Pb concentration was
also found to vary with roadway activity by Preciado and Li (2006); average Pb dust
concentrations at a busy road were 90 mg/kg, compared with 56 mg/kg at a less busy
road. Preciado and Li (2006) also examined soil Pb depth to ascertain availability of soil
Pb for exposure. They observed peak soil Pb concentrations of 250-800 mg/kg at depths
of 0.12-0.23 meters, depending on the soil measurement location and roadway traffic.
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This finding may suggest that over time, historic emissions of Pb deposited to soil are
being covered by fresh soil and hence moving further down within the soil horizons.
Size distributions of Pb-containing dust settled on the ground contain larger particles than
the size distribution of ambient air Pb, described in Section 2.5.3. Preciado and Li (2006)
measured the size distribution of Pb-containing dust near the roadside, as seen in Figure
2-27. For the busy highway, the mass median dust diameter estimated from the figure
ranged from 34- 42 pirn, depending on distance from the road. For the lower traffic
highway, the mass median dust diameter estimated from the figure ranged from
64-99 urn.
JD
"CO
E
U)
73
ca
0.
-^0
iuu-
80-
60-
40-
20-
0^
Particle Size for HWY 1
dustfall
Particle Size for H_WY
17 dustfall
10 100
Particle Size (urn)
1000
Source: Reprinted with permission of Springer-Verlag Publishing; Preciado and Li (2006)
Figure 2-27 Size distribution of Pb-containing dust collected near busy
(HWY 1) and low traffic (HWY 17) highways.
Two recent studies focused on Pb from paint degradation by examining Pb dust loading
to hard surfaces located along transects of each of the five boroughs of New York City
(Caravanos et al., 2006a; Weiss et al., 2006). Caravanos et al. (2006a) used GIS to
examine Pb dust loadings on top of pedestrian traffic signals and observed "hot spots,"
defined by the authors as at least twice the Pb dust loading at adjacent samples near major
elevated bridges in upper Manhattan, the Bronx, and Queens. In Brooklyn and Staten
Island, areas with high dust loading were not clearly attributed to a source. "Low spots,"
defined by the authors as at least two times lower Pb dust loading compared with
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adjacent samples were observed in lower Manhattan, were thought to correspond with
intensive cleaning efforts that followed the September 11, 2001 World Trade Center
attack. Weiss et al. (2006) studied Pb concentrations of grit (granules of mixed
composition found to accumulate alongside street curbs) along the transects and found
that median Pb concentrations in grit under the elevated steel structures were 2.5-11.5
times higher than those obtained away from steel structures; 90th percentile values were
up to 30 times higher near steel structures compared with those further from these
structures.
Outdoor Pb dust has been also associated with demolition activities. Farfel et al. (2005.
2003) measured Pb dust within 100 meters of a demolition site before, immediately after,
and 1 month following the demolition. They found that the rate of Pb dust fall increased
by a factor of more than 40 during demolition (Farfel et al.. 2003). Immediately after
demolition, one demolition site had dust loadings increase by a factor of 200% for streets
(87,000 (ig/m2), 138% for alleys (65,000 (ig/m2), and 26% for sidewalks (23,000 (ig/m2)
compared with pre-demolition Pb dust levels. One month following demolition, Pb dust
levels dropped by a factor of 45% for the street (48,000 (ig/m2), compared with post-
demolition concentrations, 67% for alleys (21,000 (ig/m2), and 41% for sidewalks
(14,000 (ig/m2). At another demolition site, smaller increases were observed: 29% for
streets (29,000 (ig/m2), 18% for alleys (19,000 (ig/m2) and 18% for sidewalks
(22,000 (ig/m2). No values were reported for the 1-month follow-up for the second site
(Farfel et al.. 2005).
Pb can be elevated in soils located where ammunition is used for military or hunting
purposes. In a study of Pb content in sand used to cover a firing range, Lewis et al. (2010)
found that 93% of bullet mass was recovered in the top 0.3 meters of the sand, and 6.4%
was recovered at a depth of 0.3-0.45 meter. Pb oxides were observed to be the dominant
species in the contaminated sand. Berthelot et al. (2008) studied soil Pb concentrations in
grounds (0-15 cm) used for testing military tanks and munitions and measured soil Pb
levels to range from 250 to 2,000 mg/kg dry basis.
2.6.2 Sediments
The recently completed Western Airborne Contaminants Assessment Project (WACAP)
is the most comprehensive database, to date, on contaminant transport and depositional
effects on sensitive ecosystems in the U.S. (Landers et al.. 2010). The transport, fate, and
ecological impacts of semi-volatile compounds and metals from atmospheric sources
were assessed on ecosystem components collected from 2002-2007 in watersheds of
eight core national parks (Landers et al.. 2008). The goals of the study were to assess
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where these contaminants were accumulating in remote ecosystems in the western U.S.,
identify ecological receptors for the pollutants, and to determine the source of the air
masses most likely to have transported the contaminants to the parks. Pb was measured in
sediments, as well as snow, water, lichen, fish, and moose during the multiyear project,
and although Pb was not measured in air as a part of this study, routine monitoring find
particle Pb was monitored at IMPROVE sites in the majority of national parks included
in the study.
Pb concentrations in sediments from all lakes in which Pb was measured in the
conterminous 48 states exhibited higher Pb concentrations near the surface relative to
pre-industrial Pb levels measured at greater depth. This was not the case for other metals
measured, except for cadmium (Cd) and mercury (Hg). Sediments in most lakes exhibited
maximum concentrations between 1960 and 1980, followed by a decrease, as shown in
Figure 2-28. A clear decline in Pb concentrations in sediments after the discontinued use
of leaded on-road gasoline was observed at almost all WACAP locations, of nearly all
WACAP sites in the western U.S. Sediment Pb concentrations averaged over the year in
which they were obtained correlated moderately well with annual average Pb-TSP
concentrations from the AQS with R = 0.63 for 1980-2004, in which WACAP data were
available (NPS. 2011). Pb concentrations in sediments were much lower in Alaska, and
no such decline was observed. Pb in sediments was mainly attributed to on-road gasoline
use, but for some lakes a strong influence from other local sources of Pb to lake
sediments was shown to be important, including Pb mining, smelting, logging, and other
industrial activities. The reduction in sediment Pb concentrations shown in Figure 2-28
for recent years coincides with declines in air Pb concentrations following the phase-out
of Pb anti-knock agents in gasoline and reductions of air Pb emissions from industrial
activities. Elevated Pb deposition at the Glacier, Rocky Mountain, and Sequoia and Kings
Canyon National Park and Preserve sites was thought by Landers et al. (2008) to reflect
regional scale bioaccumulation of airborne contaminants in remote ecosystems in the
western U.S. Accumulation of contaminants was shown to vary geographically; Landers
et al. (2008) lists potentially influential factors causing variation in Pb deposition
including proximity to individual sources or source areas, primarily agriculture, mining,
and smelting operations. This finding was counter to the original working hypothesis that
most of the contaminants found in western parks would originate from eastern Europe
and Asia.
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o
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en
0
CD
CD —
CD
OD
CD
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f
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-------
trends, Balogh et al. (2010) observed that the mean Pb concentration peaked in the 1970s
then declined, with levels from the 1990s below 1930s levels.
Data from select regions of the U.S. illustrate that Pb concentrations in surface waters and
sediment are likely to be higher in urbanized areas compared with rural locations. Figure
2-29 illustrates such variability within a single watershed for the Apalachicola,
Chattahoochee, and Flint River Basin, which runs south from north of the greater Atlanta,
GA metropolitan area and drains into the Gulf of Mexico at the Apalachicola Bay in the
Florida panhandle. Sediment concentrations peaked near the Atlanta area and diminished
as distance from the Apalachicola Bay decreased. This observation suggests that rural
areas have lower Pb sediment levels compared with urban areas. Consistent with the
WACAP trends shown in Figure 2-28. the data also illustrated that Pb concentrations in
sediment have declined in the U.S. since 1975 (Figure 2-30). Note that Figure 2-30 does
not include data near Atlanta, so the urban peak cannot be seen here as in Figure 2-29.
100
90
80
| 7°
.§ 60
E
o
'g 50
I 40
S
S 30
20
10
0
100
ZOO 300 400 500
River km above Apalachicola Bay, FL
Downstream
600
700
SOO
—*—Pb in streambed-sediment and reservoir-core samples
X Pb background in streambed-sediment and baseline reservoir-core samples
Note: The background refers to concentrations from undeveloped geographic regions and baseline samples are obtained from the
bottom of the sediment core to minimize anthropogenic effects on the sample. Pb concentrations reported on a dry basis.
The lakes and reservoirs along the Apalachicola, Chattahoochee, and Flint River Basin (ACF) feed from north of the Atlanta, GA
metropolitan area into the Gulf of Mexico at Apalachicola Bay in the Florida panhandle.
Source: Reprinted with permission of the American Chemical Society, Callenderand Rice (2000).
Figure 2-29 Sediment core data (1992-1994) for the lakes and reservoirs along
the Apalachicola, Chattahoochee, and Flint River Basin (ACF).
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ISO
140
—«—1975-1980
-»-1980-1985
-A-1985-1990
-K- 1990-1995
1ZO
I
100
a so
I
5 eo
Ł
40
20
Downstream
100 200 300 400 500
River km above Apalachicola Bay, FL
600
700
soo
Note: The background refers to concentrations from undeveloped geographic regions and baseline samples are obtained from the
bottom of the sediment core to minimize anthropogenic effects on the sample. Pb concentrations reported on a dry basis. Sediment
samples were not obtained for various time periods in Atlanta, so the graph does not indicate a lack of elevated sediment Pb in
Atlanta.
Lakes and reservoirs along the Apalachicola, Chattahoochee, and Flint River Basin (ACF) feed from north of the Atlanta, GA
metropolitan area into the Gulf of Mexico at Apalachicola Bay in the Florida panhandle.
Source: Reprinted with permission of the American Chemical Society, Callenderand Rice (2000).
Figure 2-30 Sediment core data (1975-1995) for the lakes and reservoirs along
the Apalachicola, Chattahoochee, and Flint River Basin (ACF).
Many recent studies have illustrated the effects of natural disasters on Pb concentrations
in surface water and sediment in the wake of Hurricane Katrina, which made landfall on
August 29, 2005 in New Orleans, LA, and Hurricane Rita, which made landfall west of
New Orleans on September 23, 2005. Pardue et al. (2005) sampled floodwaters on
September 3 and September 7, 2005 following the hurricanes and observed that elevated
concentrations of Pb along with other trace elements and contaminants were not irregular
for stormwater but were important because human exposure to the stormwater was more
substantial for Hurricane Katrina than for a typical storm. Floodwater samples obtained
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throughout the city on September 18, 2005 and analyzed for Pb by Presley et al. (2006)
were below the limit of detection (0.04 (ig/mL). Likewise, Hou et al. (2006) measured
trace metal concentration in the water column of Lake Pontchartrain and at various
locations within New Orleans during the period September 19 through October 9, 2005
and found that almost all Pb concentrations were below the limit of detection
(0.0020 mg/kg). However, several studies noted no appreciable increase in Pb
concentration within Lake Pontchartrain soils and sediments (Abel et al., 2010; Abel et
al.. 2007: Schwab et al.. 2007: Cobb et al.. 2006: Presley et al.. 2006). Shi et al. (2010)
analyzed Lake Pontchartrain sediment samples using a factored approach and found that
most Pb was sequestered in carbonate-rich, Fe oxide-rich, and magnesium (Mg) oxide-
rich sediments in which it can be more readily mobilized and potentially more
bioaccessible. Zahran et al. (2010) and Presley et al. (2010) noted that soil Pb samples
obtained outside schools also tended to decrease in the wake of Hurricanes Katrina and
Rita, with some sites observing substantial increases and others noting dramatic
reductions. These studies suggest that floodwaters can change the spatial distribution of
Pb in soil and sediments to result in increased or reduced concentrations.
2.6.3 Rain
There are currently no routine measurements of Pb in precipitation in the U.S. Recent
results from locations outside the U.S. were consistent with decreasing rain water
concentrations described in the 2006 Pb AQCD, reflecting the elimination of Pb from
on-road gasoline in most countries. From the 2006 Pb AQCD (U.S. EPA. 2006b). volume
weighted Pb concentrations in precipitation collected in 1993-94 from Lake Superior,
Lake Michigan and Lake Erie ranged from ~0.7 to ~ 1.1 (ig/L (Sweet et al.. 1998). These
values fit well with the temporal trend reported in Watmough and Dillon (2007). who
calculated annual volume-weighted Pb concentrations to be 2.12, 1.17 and 0.58 (ig/L for
1989-1990, 1990-1991 and 2002-2003, respectively, in precipitation from a central
Ontario, Canada, forested watershed. A similar value of 0.41 (ig/L for 2002-03 for Plastic
Lake, Ontario, was reported in Landre et al. (2009). For the nearby Kawagama Lake,
Shotyk and Krachler (2010) gave Pb concentrations in unfiltered rainwater collected in
2008. For August and September 2008, the values were 0.45 and 0.22 (ig/L, respectively,
and so there had been little discernible change over the post-2000 period. In support, Pb
concentrations in snow pit samples collected in 2005 and 2009 collected 45 km northeast
of Kawagama Lake had not changed to any noticeable extent (0.13, 0.17, and 0.28 (ig/L
in 2005; 0.15 and 0.26 (ig/L in 2009) (Shotyk and Krachler. 2010).
There have also been a few recently published, long-term European studies of Pb
concentration in precipitation including Berg et al. (2008) and Farmer et al. (2010). Berg
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et al. (2008) compared the trends in Pb concentration in precipitation at Norwegian
background sites in relation to the decreasing European emissions of Pb over the period
1980-2005. The Birkenes site at the southern tip of Norway is most affected by long-
range transport of Pb from mainland Europe but there had been a 97% reduction in the
concentration of Pb in precipitation over the 26-year time period. This was similar to the
reductions of 95% and 92% found for the more northerly sites, Karvatn and
Jergul/Karasjok, respectively (Figure 2-31). A decline of-95% in Pb concentrations in
moss (often used as a biomonitor of Pb pollution) from the southernmost part of Norway,
collected every 5 years over the period 1977-2005, agreed well with the Birkenes
precipitation results (Berg et al., 2008). The reductions in Pb concentration in both
precipitation and moss appear to agree well with the reductions in emissions in Europe
(-85%) and Norway (-99%). Similar to the situation in the U.S., the greatest reductions
occurred prior to the late 1990s, and relatively minor reductions have occurred thereafter;
see Figure 2-31.
U)
-Q
0-
Birkenes
Karvatn
Jergul/Karasjok
1980
1985
1990
1995
2000
2005
Source: Reprinted with permission of Pergamon Press, Berg et al. (2008)
Figure 2-31 Trends in Pb concentration in precipitation from various sites in
Norway over the period 1980-2005.
Farmer et al. (2010) showed the trends in concentration of Pb in precipitation collected in
a remote part of northeastern Scotland over the period 1989-2007. The 2.6- and 3.0-fold
decline in mean concentration from 4.92 (ig/L (1989-1991) to 1.88 (ig/L (1999) and then
to 0.63 (ig/L (2006-2007) is qualitatively but not quantitatively in line with the sixfold
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decline in annual total U.K. emissions of Pb to the atmosphere over each of these time
periods. After leaded on-road gasoline was banned in the U.K. in 2000, the ratio of
rainwater Pb concentrations to Pb emissions (metric tons) appears to have stabilized to a
near-constant value of 0.009 (ig/L per metric ton. The concentrations in precipitation
reported in these studies are all at the lower end of the range reported in the
2006 Pb AQCD (U.S. EPA. 2006b). and similar to concentrations reported for those
studies conducted after the removal of Pb from on-road gasoline. Overall, recent studies
of wet deposition tended to confirm the conclusions of the 2006 Pb AQCD (U.S. EPA.
2006b) that wet deposition fluxes have greatly decreased since the removal of Pb from
on-road gasoline.
2.6.4 Snowpack
The location of Pb deposition impacts its further environmental transport. For example,
Pb deposited to some types of soil may be relatively immobile, while Pb deposited to
snow is likely to undergo further transport more easily when snow melts. Deposition to
snow was investigated in several studies. Measurements of Pb in snowmelt during the
WACAP study, showed that median Pb concentration ranged form 20-60 ng/L, with 95th
percentile values ranging from 30-130 ng/L; see Figure 2-32 (NPS. 2011). Measurements
in WACAP of Hg and particulate carbon deposition onto snow were thought to reflect
coal combustion, and Pb was not significantly correlated with Hg in terms of either
concentration or of calculated enrichment factors normalized to Al concentrations.
Shotyk and Krachler (2009) reported considerably higher concentrations at two North
American sites, Johnson and Parnell, in Ontario, Canada. Mean Pb concentration for
contemporary snow was 672 (Johnson, n = 6; Parnell, n = 3) ng/L. Shotyk et al. (2010)
presented additional values for Pb in contemporary snow samples in Simcoe County,
Ontario, and these were higher than for ground and surface waters. Luther Bog and Sifton
Bog snow had mean Pb concentrations of 747 and 798 ng/L, respectively. The relatively
high concentrations in snow were attributed to contamination with predominantly
anthropogenic Pb, although it was noted that the extent of contamination was
considerably lower than in past decades.
Seasonal patterns of heavy metal deposition to snow on Lambert Glacier basin, east
Antarctica, were determined by Hur et al. (2007). The snow pit samples covered the
period from austral spring 1998 to summer 2002 and Pb concentrations ranged from
1.29-9.6 pg/g with a mean value of 4.0 pg/g. This was similar to a mean value of 4.7 pg/g
(1965-1986) obtained by Planchon et al. (2003) for Coats Land, northwest Antarctica.
Estimated contributions to the Pb in Lambert Glacier basin snow were -1% from rock
and soil dust (based on Al concentrations) and -4.6% from volcanoes (based on the
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concentrations of nss-sulfate). There was almost negligible contribution from seaspray
(based on Na concentrations), and so it was suggested that a substantial part of the
measured Pb concentration must originate from anthropogenic sources. Highest Pb
concentrations were generally observed in spring/summer with an occasional peak in
winter. This contrasts with data for the Antarctic Peninsula, where highest concentrations
occurred during autumn/winter, and again with Coats Land, where high concentrations
were observed throughout the winter. These differences were attributed to spatial changes
in input mechanism of Pb aerosols arriving at different sites over Antarctica, which could
be due to their different source areas and transport pathways. Hur et al. (2007). however,
suggested that the good correlation between Pb and crustal metals in snow samples shows
that Pb pollutants and crustal PM are transported and deposited in Lambert Glacier basin
snow in a similar manner.
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I I I I I I I I I
DENA NOAT MORA NOCA OLYM SEKI GAAR ROMO GLAC
Note: DENA = Denali (Alaska), NOAT = Noatak (Alaska), MORA = Mount Rainier (Washington State), NOCA = North Cascades
(Washington State), OLYM = Olympic (Washington State), SEKI = Sequoia and Kings Canyon (California), GAAR = Gates of the
Arctic (Alaska), ROMO = Rocky Mountain (Colorado), GLAC = Glacier (Montana).
Source: WACAP Database (NFS. 2011)
Figure 2-32 Box plots illustrating Pb concentration in snow melt at nine
National Parks and Preserves.
Lee et al. (2008b) collected 42 snow samples during the period autumn 2004-summer
2005 from a 2.1-meter snow pit at a high-altitude site on the northeast slope of Mount
Everest, Himalayas. Pb concentrations ranged from 5-530 pg/g with a mean value of
77 pg/g. The mean value is clearly higher than the Hur et al. (2007) value for Antarctica
but is substantially lower than a mean concentration of 573 pg/g for snow from Mont
Blanc, France [1990-1991; Lee et al. (2008b)1. The mean Pb concentration for Mount
Everest snow was lower during the monsoon (28 pg/g) compared with the non-monsoon
periods (137 pg/g). From calculated enrichment factors (Pb/Alsnow:Pb/Alcmst),
anthropogenic inputs of Pb were partly important but soil and rock dust also contributed.
The low Pb concentrations during monsoon periods are thought to be attributable to low
levels of atmospheric loadings of crustal dusts. Lee et al. (2008b) noted that their
conclusions differ from those in Kang et al. (2007). who stated that anthropogenic
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contributions of Pb to Mount Everest snow were negligible because the Everest
concentrations were similar to those in Antarctica. Kang et al. (2007) did not take account
of the difference in accumulation rates at the two sites and had also used Pb
concentrations for Antarctic snow from a study by Ikegawa et al. (1999). Lee et al.
(2008b) suggested that these Pb concentrations were much higher than expected and that
their snow samples may have suffered from contamination during sampling and analysis.
2.6.5 Natural Waters
Monitoring data for streams, rivers, and lakes are summarized in periodic national
assessments of surface waters that are carried out periodically by EPA, and they include
measurement of major biological and chemical stressors. Human exposure to Pb in
drinking water is described in Section 3.1.3.3. Pb concentrations in natural waters also
may reflect deposition of Pb even in remote locations. WACAP data at five National
Parks and Preserves show median Pb concentrations in surface waters to range from 6 to
75 ng/L (NPS. 2011); see Figure 2-33. Four sites (Denali, Mt. Ranier, Glacier, and
Olympic National Parks) were in the lower range of 6 to 20 ng/L. One site (Noatak)
reported a single value of 75 ng/L. With the exception of the Noatak site, the WACAP
values were in line with measurements by Shotyk and Krachler (2007) of Pb
concentrations in six artesian flows in Simcoe County, near Elmvale, Ontario, Canada.
The values ranged from 0.9 to 18 ng/L with a median (n = 18) of 5.1 ng/L. These are
comparable with reports of a range of 0.3-8 ng/L for Lake Superior water samples (Field
and Sherrell 2003). Shotyk and Krachler (2007) also commented that such low
concentrations for ground and surface waters are not significantly different from those
(5.1 ± 1.4 ng/L) reported for Arctic ice from Devon Island, Canada, dating from
4,000-6,000 years ago. In a separate study, Shotyk and Krachler (2009) reported
concentrations of Pb in groundwater (from two locations, Johnson and Parnell), surface
water (Kawagama Lake [Ontario, Canada]) and contemporary snow (Johnson and
Parnell, as described in Section 2.6.4). The lowest mean dissolved Pb concentrations
were found for groundwater: 5.9 (Johnson, n = 11) and 3.4 (Parnell, n = 12) ng/L. For
lake water the mean Pb concentration was 57 (Kawagama Lake, n = 12) ng/L. The
extremely low concentrations of Pb in the groundwaters were attributed to natural
removal processes. Specifically, at the sampling location in Canada, there is an
abundance of clay minerals with high surface area and high cation exchange capacity and
these, combined with the elevated pH values (pH=8.0) resulting from flow through a
terrain rich in limestone and dolostone, provide optimal circumstances for the removal of
trace elements such as Pb from groundwater. Although such removal mechanisms have
not been demonstrated, the vast difference between Pb concentration in snow and that in
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the groundwaters, indicate that the removal process is very effective. Shotyk and
Krachler (2010) speculate that even at these very low Pb concentrations, much if not
most of the Pb is likely to be colloidal, as suggested by the 2006 Pb AQCD (U.S. EPA.
2006b). Finally, Shotyk et al. (2010) suggest that the pristine groundwaters from Simcoe
County, Canada, provide a useful reference level against which other water samples can
be compared.
Although Pb concentrations in Kawagama Lake (Ontario, Canada) water were
approaching "natural values," the 206Pb/207Pb ratios for the samples that had the lowest
dissolved Pb concentrations of 10, 10 and 6 ng/L were 1.16, 1.15 and 1.16, respectively.
These values are inconsistent with those expected for natural Pb (the clay fraction from
the lake sediments dating from the pre-industrial period had values of 1.19-1.21) and it
was concluded that most of the dissolved Pb in the lake water was of industrial origin.
Shotyk and Krachler (2010) found that the full range of isotope ratios for Kawagama
Lake water samples (Ontario, Canada) was 1.09 to 1.15; this was not only much lower
than the stream water values entering the lake but also lower than the values attributed to
leaded on-road gasoline in Canada (-1.15). The streamwater ratio values were ~1.16 to
1.17, while those for rainwater were as low as 1.09; in good agreement with the lower
lake water values. This means that there must be an additional atmospheric source of Pb,
which has a lower 206Pb/207Pb ratio than leaded on-road gasoline. Supporting evidence
came from contemporary samples such as near surface peat, rainwater and snow, all of
which confirmed a trend away from natural Pb (1.191 to 1.201) to lower 206Pb/207Pb
ratios. The local smelting activities (Sudbury) were unlikely to be the source of
anthropogenic Pb as Sudbury-derived emissions exhibit a typical 206Pb/207Pb ratio of
-1.15, similar to leaded on-road gasoline. Instead, it was suggested that long-range
transport of Pb from the smelter at Rouyn-Noranda (known as the "Capital of Metal,"
NW Quebec) may still be impacting on Kawagama Lake but no Pb isotope data was
quoted to support this supposition. Several studies, summarized in Mager (2012).
reported Pb concentrations in matched reference and mining-disturbed streams in
Missouri and the western U.S. They are summarized in Table 2-11.
2-136
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o
o
o
o
o
o
1=
O
O
o
o
CM
O
o
o
I
DENA
I
MORA
I
GLAC
I
OLYM
I
NOAT
Note: DENA = Denali (Alaska), MORA = Mount Rainier (Washington State), GLAC = Glacier (Montana), OLYM = Olympic
(Washington State), NOAT = Noatak (Alaska).
Source: WACAP Database (NFS. 2011)
Figure 2-33 Boxplots of Pb concentration in surface waters measured at five
National Parks and Preserves.
The range of Pb levels in various saltwater environments are available from several
studies although the values are not specific to the U.S. A range of 0.005-0.4 (ig Pb/L for
seawater was reported by Leland and Kuwabara (1985) to reflect localized anthropogenic
inputs in marine environments based on references from prior to 1980. A range of 0.01 to
27 (ig Pb/L was reported for non-U.S. marine and estuarine waters by a group of
publications dating from 1977 through 1990 that were summarized by Sadiq (1992).
Boyle et al. (2005) compared measurements of Pb in near-surface ocean waters from the
central north Pacific Ocean (north of Hawaii) from 1976 (Schaule and Patterson. 1981)
with those from 1999 and observed a roughly 2-fold decrease in Pb concentration within
the mixed layer (first 200 meters) from 13 ng/kg down to 7 ng/kg but no difference in
concentrations deeper than 200 meters. In combination with measurements from Schaule
and Patterson (1983) and Veron et al. (1993). Wu and Boyle (1997) observed a
consistently decreasing trend in Pb concentration in the north Atlantic Ocean north of
Bermuda, from 32 ng/kg in 1979 to 10 ng/kg in 1997. In general, Pb in seawater is higher
2-137
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in coastal areas and estuaries since these locations are closer to sources of Pb
contamination and loading from terrestrial systems (U.S. EPA. 2008b).
Table 2-11 Pb concentrations from stream food-webs; in mining-disturbed areas
of Missouri and the western U.S.
Area
Animas River, CO (Besser et al.
Reference Streams
Mining-disturbed areas
Boulder River, MT (Faraq et al.,
Reference Streams
Mining-disturbed areas
Coeur d'Alene River, ID (Clark.
Reference Streams
Mining-disturbed areas
Total Pb in water (ug/L)
,,2001):
<1.8
0.9-8.6
2007):
0.4 (colloidal)
0.1-44
2003: Faraq etal.. 1998):
2-20
6-2,000
Dissolved Pb (ug/L)
<0.2
<0.1-6.9
0.3-0.4
0.1-2
0.01-2
2-50
New Lead Belt, MO (Besser et al., 2007: Brumbauah et al., 2007):
Reference Streams
Mining-disturbed areas
Adapted with permission of Elsevier:
NR
NR
[Table 4.4 in Mager (2012)]
<0.01-1.6
0.02-1.7
2.6.6 Vegetation
The 2006 Pb AQCD (U.S. EPA. 2006b) presented data on Pb in vegetation. The main
conclusions were that Pb uptake was strongly affected by pH, and acidic soils are most
likely to have Pb in solution for absorption by plants. Additionally, the 2006 Pb AQCD
(U.S. EPA. 2006b) states that most Pb stored within vegetation is stored in roots rather
than fruits or shoots. Recent measurements from the WACAP study (NPS. 2011) have
shown some Pb storage in lichens. Median Pb concentrations ranged from 0.3 mg/kg in
Noatak National Park (Alaska) to 5 mg/kg in Glacier National Park (Montana), with
substantial variation in the Glacier and Olympic National Park (Washington State)
samples (Figure 2-34). Landers et al. (2008) state that lichen Pb concentrations have
decreased substantially from the 1980s and that metal concentrations were within
background levels for these remote Western sites.
2-138
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D
O
I I
NOAT DENA
GAAR
I I I
MORA OLYM SEKI
GLAC
Note: DENA = Denali (Alaska), GAAR = Gates of the Arctic (Alaska), GLAC = Glacier (Montana), MORA = Mount Rainier
(Washington State), NOAT = Noatak (Alaska), OLYM = Olympic (Washington State), SEKI = Sequoia and Kings Canyon
(California).
Source: WACAP Database (NFS. 2011)
Figure 2-34 Boxplots of Pb concentration in lichen measured at seven
National Parks and Preserves.
Mosses can be used effectively for monitoring trends in Pb deposition as demonstrated in
many studies (Harmens et al.. 2010; Harmens et al.. 2008). For example, Harmens et al.
(2008) showed that a 52% decrease in deposited Pb concentrations corresponded to a
57% decrease in Pb concentrations in moss. Farmer et al. (2010) showed that there was
good agreement between the 206Pb/207Pb ratio for precipitation and mosses collected in
northeast Scotland. A study in the Vosges Mountains (France) also found a ratio value of
1.158 for a moss sample and a surface soil litter value of 1.167 and concluded that 1.158
to 1.167 represented the current atmospheric baseline (Geagea et al.. 2008). For rural
northeast Scotland, a combination of sources is giving rise to a 206Pb/207Pb ratio of-1.15
in recent precipitation and mosses (Farmer et al.. 2010). Clearly, sources with a lower
ratio than coal (-1.20) must be contributing substantially to the overall emissions. Pb
from waste incineration has been implicated as a possible current source (cf typical
206Pb/207Pb ratios for Pb from European incineration plants are -1.14 to 1.15 [de la Cruz
et al. (2009) and references therein].
2-139
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Pb has been measured on vegetation near roads in recent years. Hasselbach et al. (2005)
measured Pb and other metals in mosses to assess deposition of metals along a haul road
leading from a port to the Red Dog Zn-Pb mine in Northwest Alaska. They observed that
moss concentrations of Pb decreased with increasing distance from the road, while
subsurface soils (average depth = 62 meters) did not vary with distance from the road.
The strong moss Pb gradient and constant subsurface soil Pb concentrations imply that Pb
concentrations in mosses were primarily attributed to deposition and did not have
appreciable contributions from soil. Throughout the study area, median moss Pb
concentration was 16.2mg/kg (dry basis), with a range of 1.1-912.5 mg/kg.
Concentrations along the port road also diminished with increasing distance from the
port, where ore loading operations take place. Hasselbach et al. (2005) attributed the
concentrations to ore dust generated during loading operations at the port and mine along
with fugitive dust escaping during truck transport. Maher et al. (2008) measured average
Pb loading onto tree leaves near highways to be 29 (ig/m2 (max: 81 (ig/m2) at elevations
ranging from 0.30 to 2.1 meters.
Trends in Pb concentration among flora have decreased in recent years. For example,
Franzaring et al. (2010) evaluated data from a 20-year biological monitoring study of Pb
concentration in permanent forest and grassland plots in Baden-Wurttemberg, southwest
Germany. Grassland and tree foliage samples were collected from 1985-2006. The
samples were not washed and so atmospheric deposition rather than uptake from the soil
probably predominates. For all foliage (beech and spruce), Pb concentrations have shown
large reductions overtime, particularly in the early 1990s. The Pb concentrations in the
grassland vegetation also decreased from the late 1980s to the early 1990s but the trend
thereafter was found to be statistically non-significant. The reduction corresponded to the
phase-out of leaded on-road gasoline in Germany. Similarly, Aznar et al. (2008a)
observed that the decline in Pb concentrations in the outer level of tree rings
corresponded with the decline in Cu smelter emissions in Gaspe Peninsula in Canada;
Figure 2-35. Both Pb concentrations and Pb isotope ratios declined with distance from the
smelter (Aznar et al.. 2008b; Aznar et al.. 2008a).
2-140
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Trends
icentrations
8
iŁ
800 -
700-
600 -
500 -
400-
300-
200-
100-
0 •
Humus
4
" (H)
a HI)
S!i?i
2002
0.8
0.6
i 0.4
0.2
0
Regional
pollution
Smelter
emissions
1950 1960 1970 1980 1990 2000
Tree rings
1950 1960 1970 1980 1990 2000
Sapwood-heartwood
boundary
Notes: Humus Pb concentration reported in units of mg/kg dry basis, and tree ring Pb concentration reported in units of ug/kg dry
basis.
Source: Reprinted with permission of Elsevier Publishing, Aznar et al. (2008a)
Figure 2-35 Trends in regional pollution near a copper (Cu) smelter in Canada
and Pb concentrations at the boundary of heartwood trees within
roughly 75 km of the smelter.
2-141
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2.6.7 Aquatic Bivalves
Data from invertebrate waterborne populations can serve as in indicator of Pb
contamination because animals such as mussels and oysters take in contaminants during
filter feeding. Kimbrough et al. (2008) surveyed Pb concentrations in mussels, zebra
mussels, and oysters along the coastlines of the continental U.S. In general, they observed
the highest concentrations of Pb in the vicinity of urban and industrial areas. Company et
al. (2008) measured Pb concentrations and Pb isotope ratios in bivalves along the
Guadiana River separating Spain and Portugal. Analysis of Pb isotope ratio data
suggested that high Pb concentrations were related to historical mining activities in the
region. Elevated Pb concentrations were also observed by Company et al. (2008) in the
vicinity of more populated areas. Couture et al. (2010) report data from a survey of the
isotopic ratios of Pb inMytilus edulis blue mussel, collected off the coast of France from
1985-2005. The results indicated that the likely source of Pb in mussel tissue is from
resuspension of contaminated sediments enriched with Pb runoff from wastewater
treatment plants, municipal waste incinerators, smelters and refineries rather than from
atmospheric deposition (Couture et al.. 2010).
2.6.8 Vertebrate Populations
Pb concentrations in fish fillet and liver were measured through the WACAP study in
eight National Parks and Preserves (NPS. 2011). For fish fillet, Pb concentrations ranged
from 0.0033-0.30 mg/kg dry basis, with a median of 0.016 mg/kg dry basis. Liver stores
were several times higher, with Pb concentrations ranging from 0.011-0.97 mg/kg dry
basis and a median of 0.096 mg/kg dry basis. Pb concentrations in moose meat and liver
were also measured at the Denali National Park and Preserve (Alaska) as part of WACAP
(NPS. 2011). Moose meat Pb concentrations ranged from 0.021-0.23 mg/kg dry basis
with a median of 0.037 mg/kg dry basis. Pb concentrations in moose liver ranged from
0.025-0.11 mg/kg dry basis with a median of 0.053 mg/kg dry basis. Boxplots of
measured Pb concentrations in fish fillet and liver are shown in Figure 2-36. and boxplots
of measured Pb concentrations for moose meat and liver are shown in Figure 2-37. For
fish and meat tissues, median and maximum Pb concentrations were substantially lower
than values reported in the 2006 Pb AQCD (U.S. EPA. 2006b). Similarly, in a study of
Pb levels in moose teeth from Isle Royale, MI, (Vucetich et al.. 2009) median and mean
Pb levels underwent a statistically significant decrease from the period 1952-1982 to
1983-2002 in both calves and adult moose. For 1952-1982, Pb concentrations were
relatively constant, and a linear decline (R2 = 0.86) was observed for 1983-2002. These
findings suggest an overall decline but still some Pb accumulation in fish and moose in
these remote locations occurring recently.
2-142
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1.0 -
0.8 -
CT
0.6 -
o
O
QJ
0-J
(n
0.2 -
o.o -
o
o
L^ ~"~ Q
1 1 1 1 1 1 1 1 1 1 1 1
LU
O v
S L1J
O Vl
CC +J ^
tj QJ Qt
Of = .Ł
DENA = Denali (Alaska), GAAR = Gates of the Arctic (Alaska), GLAC = Glacier (Montana), MORA = Mount Rainier (Washington
State), NOAT = Noatak (Alaska), OLYM = Olympic (Washington State), ROMO = Rocky Mountain (Colorado), SEKI = Sequoia and
Kings Canyon (California).
Note: Tissue concentration reported on a dry basis.
Source: WACAP Database (NFS, 2011)
Figure 2-36 Boxplots of Pb concentration in fish fillet and fish liver, measured
at eight National Parks and/or Preserves.
2-143
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1=1
I.---J
5
i=
o
O
-------
2.7 Summary and Conclusions
2.7.1 Sources of Atmospheric Pb
The 2006 Pb AQCD (U.S. EPA. 2006b) documented the decline in ambient air Pb
emissions following the ban on alkyl-Pb additives for on-road gasoline. Pb emissions
declined by 98% from 1970 to 1995 and then by an additional 77% from 1995 to 2008, at
which time national Pb emissions were 950 tons/year. As was the case for the 2008
NAAQS review, piston-engine aircraft emissions currently comprise the largest share
(58%) of total atmospheric Pb emissions nationally (U.S. EPA. 201 la). Other sources of
ambient air Pb, in approximate order of importance with regard to national totals, include
metal working and mining, fuel combustion, other industrial sources, roadway related
sources, and historic Pb. Although piston-engine aircraft collectively comprise the largest
emissions source, the highest emitting individual industrial sites produce more ambient
air Pb emissions than individual airports.
2.7.2 Fate and Transport of Pb
The atmosphere is the main environmental transport pathway for Pb, and on a global
scale atmospheric Pb is primarily associated with fine PM. Pb in fine PM is transported
long distances and found in remote areas. Atmospheric Pb deposition peaked in the
1970s, followed by a decline. Both wet and dry deposition are important removal
mechanisms for atmospheric Pb. Wet deposition is more important for fine Pb, and Pb
associated with coarse PM is usually removed by dry deposition. Local deposition fluxes
are much higher near industrial sources, and a substantial amount of emitted Pb is
deposited near sources, leading to increased soil Pb concentrations. After initial
deposition, Pb may potentially be resuspended and redeposited many times before
reaching a site where further transport is unlikely, especially for dry deposition.
In water, Pb is transported as free ions, soluble chelates, or on surfaces of Fe-rich and
organic-rich colloids. In most surface waters, atmospheric deposition is the largest source
of Pb, but urban runoff and industrial discharge are also considerable. A substantial
portion of Pb in runoff ultimately originates from atmospheric deposition, but substantial
amounts of Pb from vehicle wear and building materials can also be transported by runoff
waters without becoming airborne. Often the majority of Pb transport by runoff occurs at
the beginning of a rainfall event. Pb is rapidly dispersed in water, and highest
concentrations of Pb are observed near sources where Pb is deposited.
2-145
-------
Transport in surface waters is largely controlled by exchange with sediments. The cycling
of Pb between water and sediments is governed by chemical, biological, and mechanical
processes, which are affected by many factors. Organic matter in sediments has a high
capacity for accumulating trace elements like Pb. Binding of anoxic sediments to sulfides
is a particularly important process that affects Pb bioavailability. Pb is relatively stable in
sediments, with long residence times and limited mobility. However, Pb-containing
sediment particles can be remobilized into the water column. Resuspended Pb is largely
associated with OM or Fe and Mn particles. This resuspension of contaminated sediments
strongly influences the lifetime of Pb in water bodies and can be a more important Pb
source to the water column than atmospheric deposition. Resuspension of sediments
largely occurs during discrete events related to storms.
A complex variety of factors influence Pb retention in soil, including hydraulic
conductivity, solid composition, OM content, clay mineral content, microbial activity,
plant root channels, animal holes, geochemical reactions, colloid amounts, colloidal
surface charge, and pH. Leaf litter can be an important temporary sink for metals from
the soil around and below leaves, and decomposition of leaf litter can reintroduce
substantial amounts of Pb into soil "hot spots," where re-adsorption of Pb is favored. A
small fraction of Pb in soil is present as the free Pb2+ ion. The fraction of Pb in this form
is strongly dependent on soil pH.
In summary, environmental distribution of Pb occurs mainly through the atmosphere,
from where it is deposited into surface waters and soil. Pb associated with coarse PM
deposits to a great extent near sources, while fine Pb-PM can be transported long
distances. Surface waters act as an important reservoir, with half-lives of Pb in the water
column largely controlled by rates of deposition to and resuspension from bottom
sediments. Pb retention in soil depends on Pb speciation and a variety of factors intrinsic
to the soil.
2.7.3 Ambient Pb Monitoring
Since the publication of the 2006 Pb AQCD (U.S. EPA. 2006b) there has been little
progress in the state of the science regarding monitoring technology and monitor siting
criteria for representation of population exposures to airborne Pb and Pb of atmospheric
origin. Our understanding of sampling errors in the existing FRM, of possible alternatives
to existing Pb-TSP sampling technology, and of particle size ranges of Pb particles
occurring in different types of locations have changed little in that time. In addition to
monitors used historically for sampling Pb-PM, several single stage and multi-stage
impactors and inlets used for sampling PM are also potential options for monitoring Pb
2-146
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particles smaller than 15 pirn. Ambient air Pb deposits onto soil or dust. As described in
Section 3.1.1.1, the size distribution of dust and soil Pb particles is larger than the size
distribution of ambient air Pb particles.
The current Pb monitoring network design requirements include two types of monitoring
sites: source-oriented and non-source-oriented. Source-oriented monitoring sites are
required near sources of air Pb emissions which are expected to or have been shown to
contribute to ambient air Pb concentrations in excess of the NAAQS. Non-source-
oriented monitoring of Pb-TSP or Pb-PMi0 is also required atNCore sites in CBSAs with
a population of at least 500,000.
In addition to Pb-TSP monitoring for the purposes of judging attainment with the
NAAQS, Pb is also routinely measured in smaller PM fractions in the CSN, IMPROVE,
and the NATTS networks. While monitoring in multiple networks provides extensive
geographic coverage, measurements between networks are not directly comparable in all
cases because different PM size ranges are sampled in different networks. Depending on
monitoring network, Pb is monitored in TSP, PMi0, or PM25 using high-volume or
low-volume samplers.
2.7.4 Ambient Air Pb Concentrations
Ambient air Pb concentrations have declined drastically over the period 1980-2010. The
median annual maximum 3-month average concentration of Pb-TSP has dropped by 97%
from 0.87 ug/m3 in 1980 to 0.03 ug/m3 in 2010. The decline can be attributed to the
phase-out of Pb antiknock agents in on-road fuel and reductions in industrial use and
processing of Pb, as described in Section 2.2.1. The mean of maximum 3-month average
concentrations for source-oriented monitors was skewed toward the 75th percentile of the
data distribution and exceeded the level of the NAAQS, indicating that highest ambient
air Pb concentrations occur near a subset of source-oriented monitors. Studies in the peer-
reviewed literature have shown slightly elevated Pb concentrations downwind of
industrial sources and airports.
Spatial variability was observed in ratios and correlations of Pb within different size
fractions. Urban or suburban land types did not appear to affect sampled size
distributions. Studies in the peer-reviewed literature suggest that proximity to industrial
sources or some roadways can affect the Pb-PM size distribution. Pb concentrations
exhibit varying degrees of association with other criteria pollutant concentrations.
Overall, non-source Pb-TSP was moderately associated with CO, PM25, and PM10,
which may indicate some role of traffic in Pb exposure. Among trace metals speciated
from PM2 5, Pb was not associated with most pollutants; Pb did associate with Zn,
2-147
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although that association was low-to-moderate, suggesting mobile source emissions
contributing to the Pb. EC, Cu, OC, Br, and K concentrations also exhibited low-to-
moderate associations with Pb concentrations. Such correlations may suggest some
common sources affecting the pollutants. Finally, the evidence on natural background Pb
suggests a plausible background airborne Pb range of 0.02 to 1 ng/m3.
2.7.5 Ambient Pb Concentrations in Non-Air Media and Biota
Atmospheric deposition has led to measurable Pb concentrations observed in rain,
snowpack, soil, surface waters, sediments, agricultural plants, livestock, and wildlife
across the world, with highest concentrations near Pb sources, such as metal smelters.
Since the phase-out of Pb from on-road gasoline, concentrations in these media have
decreased to varying degrees. In rain, snowpack, and surface waters, Pb concentrations
have decreased considerably. Declining Pb concentrations in tree foliage, trunk sections,
and grasses have also been observed. In contrast, Pb is retained in soils and sediments,
where it provides a historical record of deposition and associated ambient concentrations.
In remote lakes, sediment profiles indicate higher Pb concentrations in near surface
sediment as compared to pre-industrial era sediment from greater depth and indicate peak
concentrations between 1960 and 1980, when leaded on-road gasoline was at peak use.
Concentrations of Pb in moss, lichens, peat, and aquatic bivalves have been used to
understand spatial and temporal distribution patterns of air Pb concentrations. Ingestion
and water intake are the major routes of Pb exposure for aquatic organisms, and food,
drinking water, and inhalation are major routes of exposure for livestock and terrestrial
wildlife. Overall, Pb concentrations have decreased substantially in media through which
Pb is rapidly transported, such as air and water. Substantial Pb remains in soil and
sediment sinks. In areas less affected by major local sources, the highest concentrations
are below the surface layers and reflect the previous use of Pb in on-road gasoline and
emissions reductions from other sources.
2-148
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2.8 Chapter 2 Appendix (Supplemental Material)
2.8.1 Variability across the U.S.
Table 2-12 Distribution of 1 -month average Pb-TSP concentrations (ug/m3) nationwide, source-oriented
monitors, 2008-2010.
state/ Site N:mo N
Year Season County State County name ID means sites
Mean
Min
1
5
10
25
50
75
90
95
99
max
Nationwide statistics
2008-2010 2,318
2008 548
2009 629
2010 1141
Winter 554
Spring 579
Summer 601
Fall 584
0.202
0.318
0.212
0.141
0.202
0.239
0.186
0.184
0.000
0.004
0.002
0.000
0.000
0.000
0.001
0.000
0.003
0.004
0.004
0.002
0.002
0.003
0.003
0.004
0.006
0.013
0.008
0.005
0.006
0.007
0.006
0.007
0.010
0.024
0.013
0.008
0.008
0.012
0.010
0.011
0.029
0.050
0.038
0.018
0.026
0.034
0.030
0.026
0.063
0.110
0.084
0.045
0.055
0.070
0.066
0.064
0.217
0.348
0.256
0.136
0.184
0.272
0.212
0.206
0.578
0.841
0.611
0.408
0.502
0.738
0.559
0.505
0.856
1.240
0.856
0.625
0.883
0.977
0.755
0.758
1.576
2.557
1.357
1.233
2.438
1.905
1.233
1.178
4.440
4.440
2.438
1.828
3.103
3.123
4.440
4.225
Nationwide statistics, pooled by site
2008-2010 111
2008 47
2009 54
2010 101
Winter 108
Spring 110
0.161
0.323
0.214
0.140
0.156
0.185
0.002
0.007
0.007
0.002
0.000
0.002
0.003
0.007
0.007
0.003
0.003
0.002
0.008
0.022
0.013
0.005
0.006
0.010
0.013
0.028
0.018
0.013
0.009
0.015
0.031
0.055
0.043
0.030
0.021
0.027
0.056
0.148
0.090
0.052
0.048
0.057
0.177
0.419
0.343
0.165
0.160
0.210
0.441
0.890
0.669
0.392
0.475
0.568
0.687
1.205
0.849
0.586
0.879
0.921
0.997
1.540
0.921
0.888
1.130
1.189
1.275
1.540
0.921
1.185
1.488
1.548
2-149
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Table 2-12 (Continued): Distribution of 1-month average Pb-TSP concentrations (ug/m3) nationwide, source-oriented
monitors, 2008-2010.
State/
Year Season County
Summer
Fall
State
Statistics for individual counties
01109
06037
12057
13015
13215
17031
17115
17119
17143
17195
17201
18035
18089
18097
18127
19155
20169
21019
21151
26067
27003
27037
27145
29093
AL
CA
FL
GA
GA
IL
IL
IL
IL
IL
IL
IN
IN
IN
IN
IA
KS
KY
KY
Ml
MN
MN
MN
MO
Site
County name ID
(2008-2010)
Pike
Los Angeles
Hillsborough
Bartow
Muscogee
Cook
Macon
Madison
Peoria
Whiteside
Winnebago
Delaware
Lake
Marion
Porter
Pottawattamie
Saline
Boyd
Madison
Ionia
Anoka
Dakota
Stearns
Iron
N:mo
means
32
131
81
12
12
11
12
36
24
12
11
59
57
70
12
12
11
7
12
12
12
36
12
171
N
sites
111
110
1
4
3
1
1
1
1
1
2
1
1
2
3
2
1
1
1
1
1
1
1
1
1
7
Mean
0.148
0.152
0.5252
0.2380
0.1755
0.0128
0.0361
0.1515
0.0800
0.1367
0.0119
0.0194
0.0339
0.2746
0.0309
0.0195
0.0125
0.1536
0.2020
0.0042
0.0255
0.1781
0.0157
0.1966
0.0028
0.3388
Min
0.002
0.002
0.054
0.018
0.007
0.007
0.004
0.028
0.018
0.018
0.010
0.010
0.010
0.034
0.004
0.003
0.004
0.025
0.043
0.002
0.004
0.016
0.003
0.037
0.000
0.007
1
0.003
0.004
0.054
0.019
0.007
0.007
0.004
0.028
0.018
0.018
0.010
0.010
0.010
0.034
0.004
0.003
0.004
0.025
0.043
0.002
0.004
0.016
0.003
0.037
0.000
0.008
5
0.006
0.009
0.083
0.026
0.017
0.007
0.004
0.028
0.018
0.022
0.010
0.010
0.010
0.040
0.007
0.005
0.004
0.025
0.043
0.002
0.004
0.016
0.003
0.048
0.000
0.014
10
0.012
0.013
0.164
0.034
0.020
0.008
0.010
0.028
0.025
0.024
0.010
0.012
0.014
0.049
0.008
0.005
0.005
0.026
0.044
0.002
0.008
0.023
0.005
0.058
0.000
0.018
25
0.025
0.034
0.252
0.047
0.053
0.008
0.013
0.050
0.035
0.037
0.010
0.012
0.020
0.080
0.012
0.008
0.007
0.063
0.083
0.004
0.013
0.054
0.007
0.084
0.000
0.033
50
0.050
0.062
0.402
0.085
0.104
0.014
0.027
0.074
0.074
0.068
0.010
0.015
0.024
0.128
0.020
0.012
0.009
0.164
0.133
0.004
0.017
0.169
0.011
0.137
0.003
0.093
75
0.153
0.168
0.798
0.246
0.187
0.016
0.043
0.196
0.118
0.175
0.012
0.024
0.032
0.241
0.035
0.025
0.021
0.257
0.320
0.004
0.022
0.279
0.021
0.259
0.005
0.518
90
0.430
0.421
1.053
0.602
0.530
0.017
0.058
0.304
0.144
0.304
0.016
0.036
0.050
0.427
0.052
0.046
0.024
0.276
0.457
0.007
0.032
0.361
0.022
0.424
0.006
0.850
95
0.696
0.616
1.117
0.905
0.567
0.019
0.140
0.580
0.168
0.363
0.023
0.040
0.118
1.011
0.079
0.050
0.026
0.282
0.488
0.007
0.121
0.414
0.054
0.572
0.008
1.110
99
0.882
1.081
1.277
2.501
1.007
0.019
0.140
0.580
0.168
0.836
0.024
0.040
0.118
4.440
0.298
0.125
0.026
0.282
0.488
0.007
0.121
0.414
0.054
0.738
0.008
2.557
max
1.031
1.189
1.277
2.880
1.007
0.019
0.140
0.580
0.168
0.836
0.024
0.040
0.118
4.440
0.298
0.125
0.026
0.282
0.488
0.007
0.121
0.414
0.054
0.738
0.008
4.225
2-150
-------
Table 2-12 (Continued): Distribution of 1-month average Pb-TSP concentrations (ug/m3) nationwide, source-oriented
monitors, 2008-2010.
State/
Year Season County
29099
29179
31053
31127
36071
39035
39051
39091
39101
39151
39155
40121
41071
42003
42007
42011
42045
42055
42063
42073
42079
42129
47093
47163
48085
48375
51770
State
MO
MO
NE
NE
NY
OH
OH
OH
OH
OH
OH
OK
OR
PA
PA
PA
PA
PA
PA
PA
PA
PA
TN
TN
TX
TX
VA
Site
County name ID
Jefferson
Reynolds
Dodge
Nemaha
Orange
Cuyahoga
Fulton
Logan
Marion
Stark
Trumbull
Pittsburg
Yamhill
Allegheny
Beaver
Berks
Delaware
Franklin
Indiana
Lawrence
Luzerne
Westmore-
land
Knox
Sullivan
Collin
Potter
Roanoke City
N:mo
means
453
48
9
8
105
72
34
102
10
11
8
11
12
24
54
117
12
11
12
8
10
12
48
120
108
6
12
N
sites
19
4
1
1
3
3
1
4
1
1
1
1
1
2
3
6
1
1
1
1
1
1
2
4
3
1
1
Mean
0.4795
0.0428
0.0515
0.0476
0.0281
0.0941
0.1462
0.0480
0.0358
0.0175
0.0075
0.0023
0.0157
0.0369
0.1130
0.0989
0.0452
0.0449
0.0454
0.0438
0.0953
0.0439
0.0165
0.0534
0.3062
0.0044
0.0412
Min
0.011
0.007
0.005
0.008
0.001
0.004
0.009
0.003
0.025
0.008
0.004
0.002
0.006
0.006
0.042
0.034
0.043
0.042
0.042
0.042
0.043
0.041
0.002
0.021
0.007
0.004
0.005
1
0.015
0.007
0.005
0.008
0.001
0.004
0.009
0.003
0.025
0.008
0.004
0.002
0.006
0.006
0.042
0.035
0.043
0.042
0.042
0.042
0.043
0.041
0.002
0.023
0.028
0.004
0.005
5
0.033
0.008
0.005
0.008
0.003
0.007
0.009
0.004
0.025
0.008
0.004
0.002
0.006
0.006
0.044
0.038
0.043
0.042
0.042
0.042
0.043
0.041
0.005
0.030
0.040
0.004
0.005
10
0.048
0.011
0.005
0.008
0.004
0.008
0.026
0.005
0.026
0.009
0.004
0.002
0.007
0.006
0.047
0.042
0.043
0.043
0.043
0.042
0.044
0.041
0.006
0.032
0.052
0.004
0.008
25
0.141
0.017
0.021
0.010
0.006
0.014
0.057
0.020
0.027
0.010
0.005
0.002
0.008
0.010
0.068
0.048
0.043
0.043
0.043
0.043
0.045
0.043
0.008
0.037
0.104
0.004
0.010
50
0.336
0.027
0.031
0.024
0.018
0.038
0.091
0.042
0.033
0.018
0.007
0.002
0.016
0.017
0.096
0.066
0.045
0.045
0.044
0.044
0.071
0.044
0.012
0.045
0.189
0.004
0.015
75
0.659
0.060
0.053
0.049
0.044
0.121
0.170
0.070
0.041
0.024
0.008
0.003
0.020
0.040
0.128
0.119
0.047
0.047
0.046
0.045
0.102
0.045
0.019
0.059
0.438
0.005
0.035
90
1.118
0.087
0.149
0.206
0.063
0.210
0.420
0.090
0.054
0.025
0.017
0.003
0.025
0.121
0.198
0.200
0.048
0.047
0.047
0.046
0.215
0.046
0.032
0.083
0.717
0.006
0.054
95
1.451
0.099
0.149
0.206
0.081
0.400
0.490
0.100
0.066
0.028
0.017
0.003
0.037
0.144
0.272
0.295
0.048
0.047
0.058
0.046
0.268
0.047
0.038
0.124
0.904
0.006
0.272
99
2.220
0.268
0.149
0.206
0.101
0.719
0.510
0.120
0.066
0.028
0.017
0.003
0.037
0.149
0.286
0.347
0.048
0.047
0.058
0.046
0.268
0.047
0.063
0.145
1.178
0.006
0.272
max
3.123
0.268
0.149
0.206
0.134
0.719
0.510
0.170
0.066
0.028
0.017
0.003
0.037
0.149
0.286
0.348
0.048
0.047
0.058
0.046
0.268
0.047
0.063
0.156
1.564
0.006
0.272
2-151
-------
Table 2-12 (Continued): Distribution of 1-month average Pb-TSP concentrations (ug/m3) nationwide, source-oriented
monitors, 2008-2010.
state/ Site
Year Season County State County name ID
55117 Wl Sheboygan
72013 PR Arecibo
(Puerto Rico)
N:mo
means
12
12
Statistics for individual sites where overall average monthly mean
11090003
060371 405
29093001 6
290930021
290990004
29099001 5
2909900203
29099002 1a
2909900223
29099900 1a
2909990053
4808500093
32
36
36
36
36
21
31
21
31
24
24
36
N
sites Mean
1 0.0802
1 0.1774
> national 90th
0.525
0.671
0.670
0.681
0.997
1.275
0.687
0.719
0.441
0.850
0.986
0.601
Min
0.001
0.038
percentile
0.054
0.100
0.166
0.082
0.256
0.340
0.191
0.084
0.140
0.186
0.155
0.137
1 5
0.001 0.001
0.038 0.038
(2008-2010)
0.054 0.083
0.100 0.188
0.166 0.186
0.082 0.084
0.256 0.307
0.340 0.421
0.191 0.195
0.084 0.141
0.140 0.171
0.186 0.208
0.155 0.250
0.137 0.138
10
0.003
0.064
0.164
0.235
0.219
0.095
0.408
0.646
0.297
0.359
0.208
0.319
0.330
0.185
25
0.007
0.102
0.252
0.285
0.330
0.194
0.598
0.756
0.368
0.572
0.303
0.449
0.558
0.420
50
0.054
0.178
0.402
0.359
0.466
0.650
0.918
1.118
0.620
0.666
0.409
0.845
0.864
0.579
75
0.136
0.264
0.798
0.771
0.726
0.879
1.236
1.349
0.808
0.876
0.599
1.071
1.487
0.757
90
0.182
0.290
1.053
2.086
0.974
1.437
1.690
2.440
1.111
1.164
0.683
1.382
1.802
1.101
95
0.279
0.310
1.117
2.501
2.435
2.438
1.905
3.103
1.280
1.168
0.754
1.558
1.828
1.178
99 max
0.279 0.279
0.310 0.310
1.277 1.277
2.880 2.880
4.225 4.225
2.557 2.557
2.416 2.416
3.123 3.123
2.220 2.220
1.553 1.553
0.861 0.861
1.623 1.623
1.985 1.985
1.564 1.564
"Sites listed in the bottom six rows of this table fall in the upper 90th percentile of the data pooled by site.
2-152
-------
Table 2-13 Distribution of 1 -month average Pb-TSP concentrations
monitors, 2008-2010.
Year Season State/ State C^ S.|
County name II
? meaTs N sites -*™
Min
1
(ug/m3) nationwide,
5
10
25
50
non-source-oriented
75
90
95
99
max
Nationwide statistics
2008-2010
2008
2009
2010
Winter
Spring
Summer
Fall
2290 0.0120
685 0.0126
768 0.0114
837 0.0120
556 0.0109
574 0.0122
584 0.0119
576 0.0129
0.000
0.000
0.000
0.000
0.000
0.000
0.000
0.000
0.000
0.000
0.000
0.000
0.000
0.000
0.000
0.000
0.000
0.000
0.000
0.000
0.000
0.000
0.000
0.000
0.002
0.002
0.002
0.000
0.001
0.002
0.002
0.002
0.004
0.005
0.004
0.004
0.004
0.004
0.005
0.005
0.010
0.010
0.010
0.009
0.008
0.009
0.010
0.010
0.015
0.015
0.014
0.016
0.013
0.015
0.016
0.016
0.026
0.029
0.023
0.026
0.022
0.028
0.026
0.026
0.040
0.040
0.040
0.036
0.038
0.040
0.040
0.040
0.052
0.052
0.048
0.054
0.056
0.052
0.050
0.053
0.136
0.066
0.128
0.136
0.087
0.128
0.057
0.136
Nationwide statistics, pooled by site
2008-2010
2008
2009
2010
Winter
Spring
Summer
Fall
88 0.0120
59 0.0125
66 0.0116
73 0.0119
88 0.0115
86 0.0119
88 0.0117
88 0.0130
0.000
0.001
0.000
0.000
0.000
0.000
0.000
0.000
0.000
0.001
0.000
0.000
0.000
0.000
0.000
0.000
0.001
0.002
0.001
0.001
0.001
0.001
0.000
0.001
0.002
0.003
0.002
0.001
0.001
0.002
0.001
0.003
0.005
0.006
0.004
0.005
0.004
0.004
0.005
0.005
0.011
0.010
0.010
0.010
0.009
0.009
0.010
0.011
0.016
0.016
0.014
0.018
0.016
0.016
0.016
0.017
0.024
0.024
0.024
0.023
0.025
0.027
0.026
0.028
0.033
0.043
0.032
0.028
0.038
0.032
0.034
0.031
0.046
0.051
0.050
0.046
0.048
0.059
0.043
0.054
0.046
0.051
0.050
0.046
0.048
0.059
0.043
0.054
Statistics for individual counties (2008-2010)
04013 AZ Maricopa
06025 CA Imperial
06037 CA Los Angeles
6 1 0.0218
33 1 0.0162
224 8 0.0098
0.009
0.004
0.000
0.009
0.004
0.000
0.009
0.006
0.000
0.009
0.009
0.002
0.014
0.011
0.006
0.021
0.015
0.010
0.028
0.019
0.012
0.038
0.025
0.017
0.038
0.032
0.020
0.038
0.035
0.038
0.038
0.035
0.044
2-153
-------
Table 2-13 (Continued): Distribution of 1-month average Pb-TSP concentrations (ug/m3) nationwide, non-source-oriented
monitors, 2008-2010.
state/
Year Season
County
06065
06071
08005
08031
13089
17031
17117
17119
17143
17163
18089
18097
18163
25025
26081
26163
27017
27037
27053
27075
27123
27137
27163
29097
29187
29189
State
CA
CA
CO
CO
GA
IL
IL
IL
IL
IL
IN
IN
IN
MA
Ml
Ml
MN
MN
MN
MN
MN
MN
MN
MO
MO
MO
County Site
name ID
Riverside
San
Bernardino
Arapahoe
Denver
DeKalb
Cook
Macoupin
Madison
Peoria
Saint Clair
Lake
Marion
Vanderburgh
Suffolk
Kent
Wayne
Carlton
Dakota
Hennepin
Lake
Ramsey
Saint Louis
Washington
Jasper
Saint
Francois
Saint Louis
N: mo
means
72
71
9
12
10
288
24
36
36
36
36
35
33
31
12
36
12
118
126
10
71
72
72
12
24
33
N sites
2
2
1
1
1
8
1
1
1
1
1
1
2
2
1
2
1
5
4
1
3
2
3
1
2
1
Mean
0.0077
0.0091
0.0120
0.0056
0.0033
0.0195
0.0101
0.0188
0.0105
0.0206
0.0150
0.0058
0.0045
0.0087
0.0053
0.0112
0.0000
0.0035
0.0032
0.0000
0.0062
0.0015
0.0016
0.0125
0.0327
0.0230
Min
0.000
0.001
0.004
0.003
0.002
0.010
0.010
0.010
0.010
0.010
0.005
0.002
0.001
0.004
0.003
0.003
0.000
0.000
0.000
0.000
0.000
0.000
0.000
0.007
0.008
0.005
1
0.000
0.001
0.004
0.003
0.002
0.010
0.010
0.010
0.010
0.010
0.005
0.002
0.001
0.004
0.003
0.003
0.000
0.000
0.000
0.000
0.000
0.000
0.000
0.007
0.008
0.005
5
0.003
0.003
0.004
0.003
0.002
0.010
0.010
0.010
0.010
0.010
0.005
0.002
0.001
0.004
0.003
0.003
0.000
0.000
0.000
0.000
0.000
0.000
0.000
0.007
0.009
0.005
10
0.004
0.004
0.004
0.004
0.002
0.010
0.010
0.010
0.010
0.012
0.005
0.003
0.002
0.005
0.003
0.004
0.000
0.000
0.000
0.000
0.000
0.000
0.000
0.007
0.009
0.005
25
0.006
0.007
0.007
0.005
0.003
0.012
0.010
0.012
0.010
0.014
0.008
0.004
0.003
0.007
0.005
0.005
0.000
0.000
0.000
0.000
0.002
0.000
0.000
0.009
0.018
0.006
50
0.008
0.010
0.012
0.005
0.003
0.016
0.010
0.016
0.010
0.018
0.014
0.005
0.004
0.008
0.005
0.009
0.000
0.002
0.002
0.000
0.004
0.000
0.000
0.012
0.032
0.008
75
0.010
0.012
0.016
0.006
0.004
0.025
0.010
0.020
0.010
0.026
0.019
0.008
0.005
0.010
0.006
0.015
0.000
0.005
0.005
0.000
0.008
0.002
0.003
0.017
0.039
0.050
90
0.010
0.014
0.018
0.008
0.005
0.034
0.010
0.032
0.013
0.032
0.030
0.010
0.006
0.013
0.008
0.021
0.000
0.008
0.006
0.000
0.013
0.004
0.004
0.018
0.054
0.050
95
0.012
0.014
0.018
0.008
0.006
0.040
0.010
0.053
0.013
0.038
0.033
0.012
0.010
0.016
0.008
0.023
0.000
0.010
0.008
0.000
0.020
0.006
0.005
0.019
0.080
0.050
99
0.014
0.022
0.018
0.008
0.006
0.060
0.012
0.066
0.014
0.054
0.049
0.013
0.010
0.020
0.008
0.032
0.000
0.017
0.010
0.000
0.028
0.010
0.006
0.019
0.089
0.066
max
0.014
0.022
0.018
0.008
0.006
0.070
0.012
0.066
0.014
0.054
0.049
0.013
0.010
0.020
0.008
0.032
0.000
0.036
0.044
0.000
0.028
0.010
0.006
0.019
0.089
0.066
2-154
-------
Table 2-13 (Continued): Distribution of 1-month average Pb-TSP concentrations (ug/m3) nationwide, non-source-oriented
monitors, 2008-2010.
state/
Year Season
County
36047
39017
39029
39035
39049
39143
39167
40115
42003
42021
42045
42101
42129
48061
48141
48201
48479
49035
51087
State
NY
OH
OH
OH
OH
OH
OH
OK
PA
PA
PA
PA
PA
TX
TX
TX
TX
UT
VA
County Site
name ID
Kings
Butler
Columbiana
Cuyahoga
Franklin
Sandusky
Washington
Ottawa
Allegheny
Cambria
Delaware
Philadelphia
Westmore-
land
Cameron
El Paso
Harris
Webb
Salt Lake
Henrico
N: mo
means
24
34
107
107
36
12
54
16
36
23
20
24
24
35
68
32
29
12
7
Statistics for individual sites where overall average monthly mean
170310022
170310026
170316003
291 8700063
291 8700073
42021 0808a
36
36
36
12
12
23
N sites
1
1
3
3
1
1
2
2
1
1
1
1
1
1
3
1
1
1
1
> national
Mean
0.0131
0.0055
0.0155
0.0143
0.0092
0.0048
0.0048
0.0124
0.0105
0.0463
0.0432
0.0210
0.0419
0.0041
0.0206
0.0053
0.0134
0.0173
0.0066
Min
0.010
0.002
0.004
0.004
0.004
0.003
0.002
0.003
0.000
0.040
0.040
0.011
0.037
0.002
0.014
0.003
0.004
0.003
0.003
90th percentile
0.0330
0.0282
0.0249
0.0383
0.0271
0.0463
0.012
0.014
0.012
0.009
0.008
0.040
1
0.010
0.002
0.004
0.004
0.004
0.003
0.002
0.003
0.000
0.040
0.040
0.011
0.037
0.002
0.014
0.003
0.004
0.003
0.003
(2008-2010)
0.012
0.014
0.012
0.009
0.008
0.040
5 10
0.010 0.010
0.003 0.004
0.006 0.007
0.006 0.007
0.005 0.005
0.003 0.003
0.002 0.003
0.003 0.005
0.000 0.000
0.040 0.040
0.040 0.040
0.011 0.012
0.040 0.040
0.003 0.003
0.014 0.014
0.003 0.004
0.005 0.006
0.003 0.006
0.003 0.003
0.014 0.016
0.014 0.018
0.014 0.018
0.009 0.015
0.008 0.009
0.040 0.040
25
0.011
0.004
0.008
0.009
0.007
0.004
0.003
0.006
0.004
0.040
0.040
0.014
0.040
0.003
0.015
0.004
0.008
0.009
0.003
0.020
0.020
0.020
0.024
0.013
0.040
50
0.012
0.005
0.011
0.012
0.009
0.005
0.005
0.013
0.009
0.040
0.043
0.020
0.040
0.004
0.017
0.005
0.011
0.011
0.004
0.033
0.028
0.026
0.035
0.026
0.040
75
0.014
0.007
0.018
0.017
0.011
0.006
0.006
0.017
0.015
0.044
0.046
0.027
0.042
0.005
0.019
0.006
0.018
0.024
0.005
0.040
0.034
0.031
0.042
0.035
0.044
90
0.018
0.008
0.027
0.024
0.013
0.006
0.007
0.021
0.019
0.054
0.047
0.033
0.050
0.006
0.029
0.007
0.026
0.040
0.024
0.056
0.044
0.033
0.080
0.052
0.054
95
0.020
0.009
0.034
0.030
0.014
0.007
0.008
0.025
0.024
0.058
0.048
0.033
0.050
0.007
0.056
0.008
0.028
0.043
0.024
0.062
0.048
0.038
0.089
0.054
0.058
99 max
0.020 0.020
0.009 0.009
0.065 0.136
0.041 0.041
0.016 0.016
0.007 0.007
0.010 0.010
0.025 0.025
0.053 0.053
0.128 0.128
0.048 0.048
0.039 0.039
0.053 0.053
0.009 0.009
0.087 0.087
0.010 0.010
0.035 0.035
0.043 0.043
0.024 0.024
0.070 0.070
0.052 0.052
0.040 0.040
0.089 0.089
0.054 0.054
0.128 0.128
2-155
-------
Table 2-13 (Continued): Distribution of 1-month average Pb-TSP concentrations (ug/m3) nationwide, non-source-oriented
monitors, 2008-2010.
Year Season *** State C°u"'y S"e
County name ID
4204500023
421 2900073
481 41 0002a
N: mo
means
20
24
23
N sites Mean
0.0432
0.0419
0.0236
Min
0.040
0.037
0.016
1
0.040
0.037
0.016
5
0.040
0.040
0.016
10
0.040
0.040
0.016
25
0.040
0.040
0.017
50
0.043
0.040
0.018
75
0.046
0.042
0.021
90
0.047
0.050
0.033
95
0.048
0.050
0.056
99
0.048
0.053
0.087
max
0.048
0.053
0.087
aSites listed in the bottom six rows of this table fall in the upper 90th percentile of the data pooled by site.
2-156
-------
Table 2-14 Distribution of 3-month moving average Pb-TSP concentrations
monitors, 2008-2010.
State/
Year Season State County name
County
SitelD mNeamn°s sites ^
Min
1
5
(ug/m3) nationwide
10
25
50
75
, source-oriented
90
95
99
max
Nationwide statistics9
2008-2010
2008
2009
2010
Winter
Spring
Summer
Fall
2,112 0.2134
537 0.3225
600 0.2177
975 0.1507
443 0.2366
535 0.2376
572 0.2022
562 0.1835
0.000
0.005
0.004
0.000
0.003
0.000
0.002
0.002
0.004
0.006
0.005
0.002
0.004
0.004
0.003
0.004
0.010
0.016
0.011
0.008
0.011
0.011
0.009
0.009
0.014
0.028
0.016
0.012
0.014
0.014
0.015
0.013
0.035
0.056
0.040
0.024
0.040
0.035
0.034
0.033
0.079
0.129
0.090
0.052
0.083
0.078
0.077
0.078
0.250
0.385
0.292
0.173
0.272
0.323
0.240
0.220
0.600
0.900
0.622
0.436
0.647
0.642
0.580
0.521
0.881
1.197
0.799
0.694
0.963
0.999
0.869
0.714
1.555
2.452
1.217
1.055
2.070
2.017
1.261
1.186
2.889
2.889
2.070
1.375
2.621
2.889
2.163
2.456
Nationwide statistics, pooled by site
2008-2010
2008
2009
2010
Winter
Spring
Summer
Fall
106 0.1671
47 0.3309
54 0.2203
96 0.1415
104 0.1700
101 0.1904
106 0.1597
105 0.1538
0.002
0.007
0.007
0.002
0.003
0.001
0.002
0.002
0.003
0.007
0.007
0.002
0.004
0.002
0.004
0.004
0.012
0.024
0.013
0.008
0.011
0.013
0.010
0.009
0.015
0.029
0.019
0.013
0.013
0.016
0.015
0.012
0.030
0.056
0.042
0.027
0.025
0.028
0.028
0.032
0.059
0.154
0.086
0.053
0.055
0.060
0.058
0.066
0.173
0.461
0.311
0.163
0.171
0.186
0.174
0.170
0.577
0.814
0.632
0.407
0.522
0.502
0.520
0.462
0.717
1.284
0.840
0.619
0.827
0.874
0.788
0.630
1.009
1.639
0.886
1.110
1.097
1.231
0.989
0.960
1.316
1.639
0.886
1.110
1.324
1.740
1.104
1.161
Statistics for individual counties (2008-2010)
01109 AL Pike
06037 CA Los Angeles
12057 FL Hillsborough
13015 GA Bartow
25 1 0.5771
131 4 0.2521
79 3 0.1940
11 1 0.0125
0.223
0.023
0.011
0.009
0.223
0.023
0.011
0.009
0.247
0.036
0.015
0.009
0.256
0.041
0.037
0.009
0.302
0.055
0.063
0.011
0.574
0.078
0.110
0.013
0.719
0.237
0.249
0.014
1.088
0.543
0.423
0.015
1.178
0.832
0.582
0.016
1.210
2.452
1.770
0.016
1.210
2.489
1.770
0.016
2-157
-------
Table 2-14 (Continued): Distribution of 3-month moving average Pb-TSP concentrations (ug/m ) nationwide, source-oriented
monitors, 2008-2010.
State/
Year Season
County
13215
17031
17115
17119
17143
17195
17201
18035
18089
18097
18127
19155
20169
21151
26067
27003
27037
27145
29093
29099
29179
31053
31127
36071
39035
39051
39091
State
GA
IL
IL
IL
IL
IL
IL
IN
IN
IN
IN
IA
KS
KY
Ml
MN
MN
MN
MO
MO
MO
NE
NE
NY
OH
OH
OH
County name
Muscogee
Cook
Macon
Madison
Peoria
Whiteside
Winnebago
Delaware
Lake
Marion
Porter
Pottawattamie
Saline
Madison
Ionia
Anoka
Dakota
Stearns
Iron
Jefferson
Reynolds
Dodge
Nemaha
Orange
Cuyahoga
Fulton
Logan
Site ID N: m°
Slte ID means
12
9
10
36
20
10
9
57
46
66
10
12
9
10
10
10
36
10
158
423
40
7
6
99
70
30
100
N
sites
1
1
1
1
2
1
1
2
2
2
1
1
1
1
1
1
1
1
6
19
4
1
1
3
3
1
4
Mean
0.0367
0.1364
0.0806
0.1346
0.0121
0.0191
0.0356
0.2866
0.0305
0.0198
0.0131
0.1581
0.2286
0.0212
0.1980
0.0161
0.2026
0.0032
0.3465
0.4925
0.0397
0.0474
0.0447
0.0271
0.0905
0.1609
0.0499
Min
0.014
0.068
0.048
0.027
0.010
0.012
0.019
0.053
0.007
0.005
0.007
0.034
0.096
0.013
0.106
0.006
0.068
0.000
0.010
0.023
0.012
0.019
0.019
0.003
0.006
0.025
0.004
1
0.014
0.068
0.048
0.027
0.010
0.012
0.019
0.053
0.007
0.005
0.007
0.034
0.096
0.013
0.106
0.006
0.068
0.000
0.011
0.033
0.012
0.019
0.019
0.003
0.006
0.025
0.004
5
0.014
0.068
0.048
0.035
0.010
0.012
0.019
0.059
0.011
0.006
0.007
0.034
0.096
0.013
0.106
0.006
0.072
0.000
0.019
0.050
0.014
0.019
0.019
0.004
0.010
0.027
0.004
10
0.020
0.068
0.052
0.036
0.010
0.014
0.019
0.073
0.012
0.007
0.007
0.067
0.096
0.014
0.110
0.008
0.088
0.001
0.022
0.071
0.015
0.019
0.019
0.005
0.011
0.046
0.006
25
0.022
0.109
0.067
0.063
0.011
0.016
0.021
0.090
0.016
0.011
0.007
0.113
0.107
0.015
0.128
0.010
0.104
0.002
0.033
0.187
0.017
0.020
0.024
0.007
0.021
0.054
0.033
50
0.031
0.135
0.080
0.113
0.012
0.019
0.027
0.159
0.027
0.014
0.013
0.153
0.231
0.017
0.212
0.013
0.216
0.004
0.142
0.385
0.031
0.060
0.032
0.027
0.050
0.092
0.047
75
0.052
0.150
0.088
0.207
0.014
0.022
0.057
0.246
0.036
0.025
0.017
0.220
0.324
0.024
0.259
0.022
0.248
0.004
0.549
0.723
0.057
0.067
0.075
0.037
0.122
0.254
0.072
90
0.066
0.241
0.117
0.283
0.015
0.025
0.063
0.495
0.040
0.036
0.020
0.246
0.421
0.037
0.273
0.029
0.357
0.005
0.901
0.989
0.087
0.072
0.087
0.068
0.221
0.354
0.090
95
0.070
0.241
0.123
0.341
0.016
0.025
0.063
1.867
0.057
0.043
0.022
0.263
0.421
0.049
0.284
0.031
0.415
0.005
1.167
1.186
0.089
0.072
0.087
0.075
0.287
0.453
0.095
99
0.070
0.241
0.123
0.416
0.016
0.025
0.063
2.163
0.129
0.079
0.022
0.263
0.421
0.049
0.284
0.031
0.429
0.005
2.076
2.017
0.100
0.072
0.087
0.086
0.531
0.567
0.100
max
0.070
0.241
0.123
0.416
0.016
0.025
0.063
2.163
0.129
0.079
0.022
0.263
0.421
0.049
0.284
0.031
0.429
0.005
2.456
2.889
0.100
0.072
0.087
0.086
0.531
0.567
0.100
2-158
-------
Table 2-14 (Continued): Distribution of 3-month moving average Pb-TSP concentrations (ug/m ) nationwide, source-oriented
monitors, 2008-2010.
State/
Year Season
County
39101
39151
39155
40121
41071
42003
42007
42011
42045
42055
42063
42079
42129
47093
47163
48085
51770
55117
72013
State
OH
OH
OH
OK
OR
PA
PA
PA
PA
PA
PA
PA
PA
TN
TN
TX
VA
Wl
PR
County name
Marion
Stark
Trumbull
Pittsburg
Yamhill
Allegheny
Beaver
Berks
Delaware
Franklin
Indiana
Luzerne
Westmoreland
Knox
Sullivan
Collin
Roanoke City
Sheboygan
Arecibo
(Puerto Rico)
Site ID N: m°
Slte ID means
8
9
6
9
10
20
41
105
10
7
10
6
10
44
118
108
10
10
10
N
sites
1
1
1
1
1
2
3
6
1
1
1
1
1
2
4
3
1
1
1
Mean
0.0379
0.0180
0.0080
0.0021
0.0166
0.0414
0.1160
0.0995
0.0447
0.0447
0.0447
0.1078
0.0434
0.0165
0.0554
0.3101
0.0466
0.0897
0.1725
Min
0.032
0.015
0.005
0.002
0.009
0.009
0.043
0.038
0.043
0.043
0.043
0.084
0.041
0.007
0.030
0.048
0.013
0.012
0.059
1
0.032
0.015
0.005
0.002
0.009
0.009
0.043
0.039
0.043
0.043
0.043
0.084
0.041
0.007
0.030
0.051
0.013
0.012
0.059
5
0.032
0.015
0.005
0.002
0.009
0.011
0.052
0.041
0.043
0.043
0.043
0.084
0.041
0.009
0.033
0.070
0.013
0.012
0.059
10
0.032
0.015
0.005
0.002
0.011
0.012
0.056
0.045
0.043
0.043
0.043
0.084
0.042
0.009
0.035
0.085
0.016
0.034
0.068
25
0.034
0.016
0.006
0.002
0.013
0.017
0.083
0.051
0.043
0.043
0.043
0.085
0.042
0.012
0.039
0.120
0.019
0.058
0.129
50
0.037
0.018
0.008
0.002
0.016
0.030
0.114
0.078
0.045
0.045
0.044
0.103
0.044
0.016
0.045
0.217
0.026
0.076
0.194
75
0.042
0.019
0.010
0.002
0.019
0.054
0.159
0.145
0.046
0.046
0.046
0.135
0.044
0.020
0.060
0.469
0.097
0.126
0.213
90
0.047
0.023
0.011
0.003
0.026
0.099
0.170
0.183
0.047
0.046
0.049
0.137
0.046
0.023
0.100
0.682
0.108
0.164
0.241
95
0.047
0.023
0.011
0.003
0.027
0.120
0.187
0.197
0.047
0.046
0.049
0.137
0.046
0.027
0.125
0.753
0.109
0.170
0.245
99
0.047
0.023
0.011
0.003
0.027
0.138
0.206
0.242
0.047
0.046
0.049
0.137
0.046
0.035
0.134
1.189
0.109
0.170
0.245
max
0.047
0.023
0.011
0.003
0.027
0.138
0.206
0.251
0.047
0.046
0.049
0.137
0.046
0.035
0.168
1.262
0.109
0.170
0.245
2-159
-------
Table 2-14 (Continued): Distribution of 3-month moving average Pb-TSP concentrations (ug/m ) nationwide, source-oriented
monitors, 2008-2010.
State/
Year Season State County name Site ID
County
Statistics for individual sites where overall average monthly mean >
011090003
060371405
290930016
290930021
290990004
29099001 5b
290990020"
290990021"
290999001"
290999005"
480850009"
N: mo N ,,
means sites Mean
national
25
36
36
36
36
21
29
21
22
22
36
90th percentile
0.5771
0.7174
0.6682
0.6950
1.0090
1.3162
0.6680
0.7317
0.8413
0.9875
0.6068
Min
1
5
10
25
50
75
90
95
99 max
(2008-2010)
0.223
0.188
0.207
0.173
0.640
0.612
0.452
0.429
0.587
0.612
0.196
0.223
0.188
0.207
0.173
0.640
0.612
0.452
0.429
0.587
0.612
0.196
0.247
0.234
0.258
0.192
0.655
0.632
0.471
0.435
0.592
0.630
0.268
0.256
0.237
0.313
0.218
0.699
0.743
0.482
0.507
0.600
0.644
0.335
0.302
0.309
0.418
0.346
0.775
0.921
0.555
0.547
0.699
0.783
0.469
0.574
0.476
0.543
0.689
0.913
1.074
0.651
0.685
0.845
0.995
0.585
0.719
0.791
0.634
0.954
1.081
1.258
0.754
0.900
0.963
1.220
0.704
1.088
2.178
1.167
1.214
1.555
2.621
0.891
0.999
1.061
1.271
0.965
1.178
2.452
2.076
1.275
2.011
2.634
0.943
1.013
1.100
1.278
1.189
1.210 1.210
2.489 2.489
2.456 2.456
1.937 1.937
2.017 2.017
2.889 2.889
0.989 0.989
1.141 1.141
1.204 1.204
1.375 1.375
1.262 1.262
aThe 3-month averages presented here were created using a simplified approach of the procedures detailed in 40 CFR part 50 appendix R and as such cannot be directly compared to
the Pb NAAQS for determination of compliance with the Pb NAAQS.
"Sites listed in the bottom six rows of this table fall in the upper 90th percentile of the data pooled by site.
2-160
-------
Table 2-15 Distribution of 3-month moving average Pb-TSP concentrations (ug/m3) nationwide, non-source-
oriented monitors, 2008-2010.
State/ N: mo N
Year Season County ST County name Site ID means sites Mean Min 1 5 10 25 50 75 90 95 99 max
Nationwide statistics
2008-2010
2008
2009
2010
Winter
Spring
Summer
Fall
2,164
663
727
774
494
548
565
557
0.0120
0.0130
0.0114
0.0118
0.0113
0.0119
0.0121
0.0126
0.000
0.000
0.000
0.000
0.000
0.000
0.000
0.000
0.000
0.000
0.000
0.000
0.001
0.000
0.000
0.000
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.002
0.002
0.002
0.001
0.002
0.002
0.002
0.002
0.005
0.005
0.004
0.005
0.005
0.005
0.005
0.005
0.010
0.011
0.009
0.010
0.009
0.009
0.010
0.011
0.015
0.016
0.014
0.016
0.014
0.015
0.016
0.017
0.025
0.027
0.024
0.025
0.023
0.025
0.026
0.027
0.037
0.040
0.038
0.035
0.037
0.036
0.037
0.037
0.048 0.073
0.050 0.055
0.043 0.073
0.047 0.057
0.050 0.055
0.050 0.073
0.046 0.053
0.048 0.057
Nationwide statistics, pooled by site
2008-2010
2008
2009
2010
Winter
Spring
Summer
Fall
86
59
65
71
84
83
86
86
0.0120
0.0127
0.0117
0.0118
0.0118
0.0118
0.0118
0.0126
0.000
0.001
0.001
0.000
0.000
0.000
0.000
0.000
0.000
0.001
0.001
0.000
0.000
0.000
0.000
0.000
0.001
0.002
0.001
0.001
0.001
0.001
0.001
0.001
0.002
0.003
0.003
0.001
0.002
0.002
0.002
0.002
0.005
0.005
0.004
0.005
0.005
0.004
0.005
0.005
0.010
0.011
0.010
0.010
0.010
0.010
0.009
0.011
0.016
0.016
0.014
0.017
0.015
0.015
0.016
0.016
0.024
0.024
0.026
0.022
0.025
0.025
0.023
0.026
0.034
0.043
0.031
0.028
0.036
0.034
0.037
0.030
0.046 0.046
0.050 0.050
0.049 0.049
0.045 0.045
0.048 0.048
0.059 0.059
0.043 0.043
0.046 0.046
Statistics for individual counties (2008-2010)
06025 CA Imperial
06037 CA Los Angeles
06065 CA Riverside
06071 CA San Bernardino
08005 CO Arapahoe
08031 CO Denver
13089 GA DeKalb
17031 IL Cook
31
218
72
69
7
10
8
287
1
8
2
2
1
1
1
8
0.0165
0.0100
0.0078
0.0091
0.0126
0.0054
0.0035
0.007
0.000
0.002
0.003
0.011
0.004
0.003
0.0196 0.010
2-161
0.007
0.000
0.002
0.003
0.011
0.004
0.003
0.010
0.008
0.002
0.004
0.005
0.011
0.004
0.003
0.010
0.011
0.004
0.005
0.006
0.011
0.004
0.003
0.010
0.013
0.006
0.007
0.007
0.011
0.005
0.003
0.012
0.017
0.009
0.008
0.009
0.013
0.006
0.004
0.017
0.021
0.013
0.010
0.011
0.014
0.006
0.004
0.025
0.023
0.016
0.011
0.013
0.014
0.006
0.004
0.033
0.023
0.020
0.011
0.014
0.014
0.006
0.004
0.038
0.023 0.023
0.028 0.035
0.011 0.011
0.017 0.017
0.014 0.014
0.006 0.006
0.004 0.004
0.047 0.051
-------
Table 2-15 (Continued): Distribution of 3-month moving average Pb-TSP concentrations (ug/m ) nationwide, non-source-
oriented monitors, 2008-2010.
Year Season
17117
17119
17143
17163
18089
18097
18163
25025
26081
26163
27017
27037
27053
27075
27123
27137
27163
29097
29187
29189
36047
39017
39029
39035
39049
39143
39167
40115
42003
42021
42045
42101
State/
County ST
IL
IL
IL
IL
IN
IN
IN
MA
Ml
Ml
MN
MN
MN
MN
MN
MN
MN
MO
MO
MO
NY
OH
OH
OH
OH
OH
OH
OK
PA
PA
PA
PA
County name Site ID
Macoupin
Madison
Peoria
Saint Clair
Lake
Marion
Vanderburgh
Suffolk
Kent
Wayne
Carlton
Dakota
Hennepin
Lake
Ramsey
Saint Louis
Washington
Jasper
Saint Francois
Saint Louis
Kings
Butler
Columbiana
Cuyahoga
Franklin
Sandusky
Washington
Ottawa
Allegheny
Cambria
Delaware
Philadelphia
N: mo
means
24
36
36
36
36
33
31
24
10
32
10
112
124
8
65
70
70
10
21
33
24
30
105
105
36
10
48
12
36
23
14
22
N
sites
1
1
1
1
1
1
2
2
1
2
1
5
4
1
3
2
3
1
2
1
1
1
3
3
1
1
2
2
1
1
1
1
Mean
0.0101
0.0188
0.0105
0.0204
0.0149
0.0056
0.0047
0.0093
0.0055
0.0119
0.0000
0.0036
0.0033
0.0000
0.0061
0.0016
0.0017
0.0135
0.0337
0.0243
0.0131
0.0055
0.0148
0.0144
0.0092
0.0052
0.0047
0.0128
0.0101
0.0459
0.0427
0.0214
Min
0.010
0.010
0.010
0.012
0.007
0.003
0.002
0.005
0.004
0.004
0.000
0.000
0.000
0.000
0.001
0.000
0.000
0.009
0.011
0.005
0.011
0.003
0.005
0.005
0.005
0.004
0.002
0.005
0.000
0.040
0.040
0.013
1
0.010
0.010
0.010
0.012
0.007
0.003
0.002
0.005
0.004
0.004
0.000
0.000
0.001
0.000
0.001
0.000
0.000
0.009
0.011
0.005
0.011
0.003
0.005
0.006
0.005
0.004
0.002
0.005
0.000
0.040
0.040
0.013
5
0.010
0.010
0.010
0.012
0.007
0.003
0.003
0.006
0.004
0.004
0.000
0.001
0.001
0.000
0.001
0.000
0.000
0.009
0.012
0.005
0.011
0.004
0.007
0.006
0.005
0.004
0.002
0.005
0.000
0.040
0.040
0.014
10
0.010
0.011
0.010
0.014
0.007
0.003
0.003
0.006
0.005
0.005
0.000
0.001
0.001
0.000
0.001
0.000
0.000
0.011
0.012
0.006
0.011
0.004
0.008
0.008
0.005
0.004
0.003
0.006
0.000
0.040
0.040
0.014
25
0.010
0.014
0.010
0.016
0.010
0.004
0.004
0.008
0.005
0.005
0.000
0.001
0.002
0.000
0.002
0.001
0.001
0.012
0.027
0.007
0.012
0.005
0.010
0.010
0.008
0.005
0.004
0.010
0.007
0.040
0.040
0.018
50
0.010
0.016
0.010
0.020
0.014
0.005
0.005
0.009
0.006
0.012
0.000
0.003
0.003
0.000
0.005
0.001
0.001
0.014
0.035
0.008
0.013
0.006
0.013
0.013
0.010
0.005
0.004
0.014
0.012
0.041
0.042
0.022
75
0.010
0.022
0.011
0.024
0.018
0.007
0.005
0.011
0.006
0.017
0.000
0.005
0.004
0.000
0.008
0.002
0.003
0.015
0.042
0.050
0.014
0.006
0.017
0.018
0.011
0.006
0.006
0.016
0.014
0.046
0.045
0.025
90
0.011
0.036
0.012
0.029
0.024
0.009
0.006
0.013
0.006
0.021
0.000
0.007
0.006
0.000
0.014
0.004
0.004
0.016
0.048
0.050
0.016
0.007
0.021
0.023
0.011
0.006
0.007
0.018
0.016
0.069
0.046
0.029
95
0.011
0.036
0.012
0.033
0.032
0.010
0.007
0.015
0.006
0.023
0.000
0.012
0.006
0.000
0.016
0.004
0.004
0.017
0.053
0.050
0.018
0.007
0.028
0.027
0.012
0.006
0.007
0.019
0.018
0.070
0.047
0.029
99
0.011
0.039
0.013
0.036
0.037
0.011
0.007
0.016
0.006
0.024
0.000
0.013
0.015
0.000
0.017
0.005
0.005
0.017
0.054
0.055
0.019
0.008
0.054
0.033
0.012
0.006
0.008
0.019
0.025
0.073
0.047
0.030
max
0.011
0.039
0.013
0.036
0.037
0.011
0.007
0.016
0.006
0.024
0.000
0.015
0.016
0.000
0.017
0.005
0.005
0.017
0.054
0.055
0.019
0.008
0.057
0.035
0.012
0.006
0.008
0.019
0.025
0.073
0.047
0.030
2-162
-------
Table 2-15 (Continued): Distribution of 3-month moving average Pb-TSP concentrations (ug/m ) nationwide, non-source-
oriented monitors, 2008-2010.
State/ M: mo
Year Season County ST County name Site ID means
42129 PA Westmoreland
48061 TX Cameron
48141 TX El Paso
48201 TX Harris
48479 TX Webb
49035 UT Salt Lake
24
33
56
30
23
10
Statistics for individual sites where overall average monthly mean
170310022
170310026
170316003
291870006"
291870007"
291892003"
420210808"
420450002"
421290007"
36
36
36
10
11
33
23
14
24
N
sites Mean
1 0.0417
1 0.0042
3 0.0212
1 0.0051
1 0.0121
1 0.0145
Min
0.037
0.002
0.014
0.004
0.006
0.007
> national 90th percentile
0.0335
0.0281
0.0245
0.0412
0.0268
0.0243
0.0459
0.0427
0.0417
0.016
0.018
0.015
0.017
0.011
0.005
0.040
0.040
0.037
1
0.037
0.002
0.014
0.004
0.006
0.007
5
0.040
0.003
0.014
0.004
0.007
0.007
10
0.040
0.003
0.015
0.004
0.007
0.007
25
0.040
0.004
0.016
0.005
0.008
0.008
50
0.041
0.004
0.018
0.005
0.010
0.011
75 90
0.043 0.046
0.005 0.005
0.023 0.038
0.006 0.006
0.016 0.021
0.016 0.032
95
0.047
0.006
0.040
0.007
0.022
0.036
99 max
0.048 0.048
0.006 0.006
0.040 0.040
0.007 0.007
0.026 0.026
0.036 0.036
(2008-2010)
0.016
0.018
0.015
0.017
0.011
0.005
0.040
0.040
0.037
0.018
0.019
0.015
0.017
0.011
0.005
0.040
0.040
0.040
0.026
0.022
0.017
0.026
0.012
0.006
0.040
0.040
0.040
0.028
0.023
0.020
0.035
0.012
0.007
0.040
0.040
0.040
0.032
0.026
0.025
0.043
0.028
0.008
0.041
0.042
0.041
0.038 0.047
0.032 0.038
0.028 0.031
0.048 0.054
0.035 0.036
0.050 0.050
0.046 0.069
0.045 0.046
0.043 0.046
0.048
0.043
0.035
0.054
0.041
0.050
0.070
0.047
0.047
0.051 0.051
0.046 0.046
0.036 0.036
0.054 0.054
0.041 0.041
0.055 0.055
0.073 0.073
0.047 0.047
0.048 0.048
aThe 3-month averages presented here were created using a simplified approach of the procedures detailed in 40 CFR part 50 appendix R and as such cannot be directly compared to
the Pb NAAQS for determination of compliance with the Pb NAAQS.
"Sites listed in the bottom six rows of this table fall in the upper 90th percentile of the data pooled by site.
2-163
-------
Table 2-16 Distribution of annual 1-month site maxima TSP Pb concentrations (ug/m3) nationwide, source-
oriented monitors, 2008-2010.
Year Site ID-year N (sites) Mean Min 1 5 10 25 50 75 90 95 99 max
Nationwide statistics
2008-2010 111 0.5003 0.003 0.006 0.016 0.032 0.066 0.156 0.575 1.530 2.416 4.225 4.440
2008 47 0.8138 0.012 0.012 0.052 0.057 0.096 0.320 0.850 2.557 3.123 4.440 4.440
2009 54 0.4486 0.016 0.016 0.022 0.050 0.090 0.170 0.618 1.280 1.623 2.438 2.438
2010 101 0.3105 0.003 0.006 0.008 0.024 0.054 0.142 0.347 0.854 1.117 1.576 1.828
Annual site max 1-month means >= national 90th percentile (2008-20010)
060371405-2008 2.8800
180350009-2008 4.4400
290930016-2008 4.2252
290930021-2008 2.5566
290930021-2009 2.4380
290990004-2008 2.4156
290990004-2009 1.5599
290990004-2010 1.5762
290990011-2008 1.5295
290990015a-2008 3.1228
290990020a-2008 2.2204
290990021 a-2008
290999001 a-2009
290999001 a-2010
290999005a-2009
290999005a-2010
480850009a-2008
1 .5528
1 .6228
1 .5576
1 .9850
1 .8278
1 .5640
aSites listed in the bottom eight rows of this table fall in the upper 90th percentile of the data pooled by site.
2-164
-------
Table 2-1 7
Year
Distribution of annual 1 -month site maxima TSP Pb
oriented monitors, 2008-2010.
Site ID - year N (sites) Mean Min 1 5
concentrations (ug/m3) nationwide, non-source-
10 25 50 75 90 95 99 max
Nationwide statistics
2008-2010
2008
2009
2010
Annual site max
88 0.0284 0.000 0.000 0.004
59 0.0232 0.004 0.004 0.005
66 0.0210 0.003 0.003 0.005
73 0.0233 0.000 0.000 0.002
0.006 0.010 0.020 0.041 0.057 0.070 0.136 0.136
0.006 0.010 0.016 0.033 0.053 0.058 0.066 0.066
0.006 0.008 0.014 0.026 0.040 0.056 0.128 0.128
0.004 0.008 0.015 0.029 0.049 0.065 0.136 0.136
1-month means >= national 90th percentile (2008-2010)
170310022-2009
170310022-2010
171193007-2008
291870006a-2010
291 892003a-2008
3902900 19a-20 10
390290022a-2010
42021 0808a-2008
42021 0808a-2009
481410002a-2010
481 41 0033a-2009
0.0700
0.0620
0.0660
0.0894
0.0660
0.1360
0.0652
0.0583
0.1280
0.0870
0.0570
"Sites listed in the bottom eight rows of this table fall in the upper 90th percentile of the data pooled by site.
2-165
-------
Table 2-18 Distribution of annual 3-month site maxima Pb-TSP concentrations (ug/m3) nationwide, source-
oriented monitors, 2008-2010.
Year
Site ID - year
N (sites) Mean Min
1 5 10 25 50 75 90 95 99 max
Nationwide statistics9
2008-2010
2007
2008
2009
Annual site max 3-month
means >= national
01 1 090003-2008
060371405-2008
120571066-2008
180350009-2008
29093001 6b-2008
29093001 6b-2009
29093002 1b-2009
290990004b-2008
29099001 5b-2008
29099900 1b-2009
290999005b-2009
290999005b-2010
480850009b-2008
106 0.3605 0.003
47 0.5831 0.009
54 0.3611 0.012
96 0.2112 0.003
90th percentile (2008-2010)
1.2100
2.4890
1 .7700
2.1630
2.4560
2.0700
1 .9370
2.0170
2.8890
1 .2040
1 .2580
1 .3750
1 .2620
0.005 0.016 0.023 0.047 0.109 0.378 1.204 1.937 2.489 2.889
0.009 0.038 0.043 0.085 0.242 0.815 2.017 2.456 2.889 2.889
0.012 0.017 0.035 0.060 0.121 0.467 1.079 1.258 2.070 2.070
0.003 0.011 0.021 0.046 0.091 0.262 0.630 0.865 1.375 1.375
aThe 3-month averages presented here were created using a simplified approach of the procedures detailed in 40 CFR part 50 appendix R and as such cannot be directly compared to
the Pb NAAQS for determination of compliance with the Pb NAAQS.
bSites listed in the bottom nine rows of this table fall in the upper 90th percentile of the data pooled by site.
2-166
-------
Table 2-19 Distribution of annual 3-month site maxima Pb-TSP concentrations (ug/m3) nationwide, non-source-
oriented monitors, 2008-2010.
Year
Site ID - year N (sites) Mean
Min
10
25
50
75
90
95
99
max
Nationwide
2008-2010
2008
2009
2010
Annual site
statistics9
max 3-month means >= national
1 7031 0022-2008
1 7031 0022-2009
170310022-2010
1 7031 0026b-2008
291870006b-2010
291 892003b-2008
3902900 19b-20 10
390290022b-2010
42021 0808b-2008
42021 0808b-2009
420450002b-2010
421 290007b-2008
86 0.0198 0.000 0.000
59 0.0176 0.002 0.002
65 0.0162 0.002 0.002
71 0.0171 0.000 0.000
90th percentile (2008-2010)
0.0480
0.0470
0.0510
0.0460
0.0540
0.0550
0.0570
0.0440
0.0490
0.0730
0.0470
0.0480
0.002 0.004 0.007 0.015 0.028 0.044 0.051 0.073 0.073
0.004 0.005 0.007 0.014 0.024 0.039 0.048 0.055 0.055
0.003 0.004 0.006 0.013 0.021 0.038 0.041 0.073 0.073
0.001 0.002 0.006 0.013 0.024 0.037 0.047 0.057 0.057
aThe 3-month averages presented here were created using a simplified approach of the procedures detailed in 40 CFR part 50 appendix R and as such cannot be directly
compared to the Pb NAAQS for determination of compliance with the Pb NAAQS.
b Sites listed in the bottom nine rows of this table fall in the upper 90th percentile of the data pooled by site.
2-167
-------
Concentrations of Pb Measured using PM10 Monitors (for Concentrations
and Trends)
Figure 2-38 displays maximum 3-month averages for Pb-PMi0 concentrations for 36
counties in which measurements were obtained. Among the 36 counties in which PM10
monitoring was conducted, only one county, Gila County, AZ, reported concentrations
above 0.076 ug/m3. Three other counties reported concentrations greater than
0.016 ug/m3: Wayne County, MI, Boyd County, KY, and the county of St. Louis City,
MO.
2007-2009 Pb-PM10County Maximum 3-Month Mean
Concentration:
• >= 0.076 ng/m3 (1 county)
• 0.016 - 0.075 wg/rn; [3 counties)
• 0.006 - 0.015 Hg/nv (17 counties)
<= .005 ng/m" [15 counties)
J no data
Figure 2-38 Highest county-level Pb-PM10 concentrations (ug/m ), maximum
3-month average, 2007-2009.
2-168
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Concentrations of Pb Measured using PM2.s Monitors (for Speciation
Concentrations and Trends)
Figure 2-39 displays maximum 3-month average county-level data for Pb in PM2 5
concentrations for 323 counties in which PM2 5 measurements were obtained for
speciation in the CSN and IMPROVE networks. The data presented here are not
compared to the NAAQS because PM2 5 monitors are not deployed for the purpose of
evaluating compliance for the NAAQS. Among the 323 counties in which PM25
monitoring was conducted, only eleven counties reported concentrations greater than
0.016 ug/m3: Jefferson, AL, San Bernardino, CA, Imperial, CA, Wayne, MI, Jefferson,
MO, Erie, NY, Lorain, OH, Allegheny, PA, Berks, PA, Davidson, TN, and El Paso, TX.
2007-2009 Pb-PM2 5 County Maximum 3-Month Mean
Concentration:
• 0.016 -0.075MŁ/m!(ll counties)
• 0.006 -0.015 u.g/m (71 counties)
<= .005 ne/m3(241 counties)
_J no data
Figure 2-39
Highest county-level Pb-PM2.s concentrations (ug/m3), maximum
3-month average, 2007-2009.
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2.8.2 Intra-urban Variability
Intra-urban variability in Pb concentrations reported to AQS was described in detail for
Los Angeles County, CA (Los Angeles), Hillsborough and Pinellas Counties, FL
(Tampa), Cook County, IL (Chicago), Jefferson County, MO (Herculaneum), Cuyahoga
County, OH (Cleveland), and Sullivan County, TN (Bristol) were selected for this
assessment to illustrate the variability in Pb concentrations measured across different
metropolitan regions with varying Pb source characteristics. Four of the counties
encompass large cities (Los Angeles, Tampa, Chicago, and Cleveland). All six counties
contain source-oriented monitors. Maps and wind roses (graphs representing wind
direction and wind speed at a location) are presented in this Chapter 2 Appendix for each
of the six urban areas. Additionally, annual and seasonal box plots of the Pb
concentration distributions and intra-monitor correlation tables are presented to illustrate
the level of variability throughout each urban area.
Maps of six areas (Los Angeles County, CA; Hillsborough/Pinellas Counties, FL; Cook
County, IL; Jefferson County, MO; Cuyahoga County, OH; and Sullivan County, TN) are
shown to illustrate the location of all Pb monitors meeting the inclusion criteria. Wind
roses for each season are also provided to help put the source concentration data in
context. Letters on the maps identify the individual monitor locations and correspond
with the letters provided in the accompanying concentration box plots and pair-wise
monitor comparison tables. The box plots for each monitor include the annual and
seasonal concentration median and interquartile range with whiskers extending from the
5th to the 95th percentile. Data from 2008-2010 were used to generate the box plots,
which are stratified by season as follows: 1 = winter (December-February), 2 = spring
(March-May), 3 = summer (June-August), and 4 = fall (September-November). The
comparison tables include the Pearson correlation coefficient (R), Spearman rank-ordered
correlation coefficient (p), the 90th percentile of the absolute difference in concentrations
(P90) in (ig/m3, the coefficient of divergence (COD) and the straight-line distance
between monitor pairs (d) in km. The COD provides an indication of the variability
across the monitoring sites within each county and is defined as follows:
Equation 2A-1
where Xy andXik represent the observed hourly concentrations for time period /' at sites j
and k, andp is the number of paired hourly observations. A COD of 0 indicates there are
2-170
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no differences between concentrations at paired sites (spatial homogeneity), while a COD
approaching 1 indicates extreme spatial heterogeneity.
In certain cases, the information contained in these figures and tables should be used with
some caution since many of the reported concentrations forthe years 2008-2010 are near
or below the analysis method's stated method detection limit (MDL). The MDL is
generally taken as 0.01 because it is the upper value of the range of MDLs reported for
atomic absorption (AA) and Emissions Spectra ICAP methods, which were the two
methods reported in the AQS to have been used for analysis of FRM samples (Rice.
2007). Generally, data are reported to the hundredth place, so this assumption is
reasonable. The approximate percentage of data below the MDL (to the nearest 5%) is
provided for each site along with box plots of seasonal Pb concentration at monitors
within each urban area studied.
Figure 2-40 illustrates Pb monitor locations within Los Angeles County, CA. Ten
monitors are located within Los Angeles County, five of which were source-oriented and
the other five were non-source-oriented monitors. Monitor A was located immediately
downwind of the Quemetco battery recycling facility in the City of Industry, CA. This
source was estimated to produce 0.32 tons of Pb/year (U.S. EPA. 2008c). Monitor C was
sited in a street canyon just upwind of the Exide Pb recycling facility, which was
estimated to produce 2.0 tons of Pb/year (U.S. EPA. 2008c). Monitor D was situated
slightly northwest of the same Pb recycling facility. It is still in relatively close proximity
but not downwind on most occasions. Monitor B was located 12 km downwind of the
Exide facility. Monitor E was located nearby the Trojan Battery recycling facility, which
emitted 0.79 tons Pb/year (U.S. EPA. 2008c). Location of the non-source-oriented
monitors varied. Monitor F was positioned on a rooftop 60 meters away from a 4-lane
arterial road and 100 meters from of a railroad. Monitor G was located on a rooftop
approximately 20 meters from an 8-lane arterial road, and monitor H was positioned at
the curbside of a four-lane road roughly 650 meters north of that road's junction with
Interstate 1-405. Monitor I was sited in a parking lot roughly 80 meters from a four-lane
road, and monitor J was located approximately 130 meters south of a 4-lane highway.
Figure 2-41 displays seasonal wind roses for Los Angeles County. In spring, summer,
and fall, the predominant winds come from the west-southwest. During winter, wind
direction varies with a portion from the west-southwest and the remainder from the east.
The highest winds during winter come more frequently from the west-southwest.
The maps shown in Figure 2-40 for source-oriented monitors A-E illustrate the different
conditions captured by the monitors; this informs analysis of the seasonal and year-round
concentrations reported in Figure 2-42. The average annual concentration at monitor A
was 0.074 ug/m3. The 95th percentile exceeded the level of the NAAQS in the spring
2-171
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(0.16 ug/m3) and summer (0.18 ug/m3). Monitor C reported the highest concentrations in
Los Angeles County, with a year-round mean of 0.68 ug/m3. Given the position of this
monitor with respect to the Exide facility, there is the potential for recirculation of
fugitive Pb emissions in the air sampled by that monitor. The average annual Pb
concentration at monitor D was 0.12 ug/m3, and the 75th percentile of year-round data
exceeded the level of the NAAQS; in spring, the 70th percentile exceeded 0.15 ug/m3.
Monitor B reported the lowest values among the source-oriented monitors with an
average annual concentration of 0.013 ug/m3. Note that 75% of reported values were
below the MDL for this site, and no data from this site exceeded the level of the NAAQS.
The annual average concentration at monitor E was 0.068 ug/m3, and the 95th percentile
of concentration was 0.17 ug/m3.
The non-source-oriented monitors located at sites F-J all recorded low concentrations,
with average values ranging from 0.004 to 0.018 ug/m3 (Figure 2-42). The highest
average year-round concentrations were recorded at site F. The 95th percentiles at these
sites ranged from 0.01 to 0.04 ug/m3. There is much less certainty in the data recorded at
the non-source-oriented sites, because 45-95% of the data from these monitors were
below the MDL. Additionally, only one of the non-source-oriented monitors (monitor H)
was positioned at roadside, and none of the non-source-oriented monitors were located at
the side of a major highway.
Intersampler correlations (Table 2-20). illustrate that Pb has high intra-urban spatial
variability. For the source-oriented monitors, the highest correlation (R = 0.59, p = 0.57)
occurred for monitors C and D, which covered the same site. Because monitor D was
slightly farther from the Exide source and slightly upstream of the predominant wind
direction, the signal it received from the source site was correspondingly lower. Hence,
the correlation between these sites was moderate despite their relatively close proximity.
In general, low or even negative correlations were observed between the source-oriented
and non-source-oriented monitors. The exception to this was the Spearman-ranked
correlation between source-oriented monitor B and non-source-oriented monitor F, with
p = 0.74. Pearson correlation was much lower for this pair (R = 0.33). Monitors B and F
are roughly 16 km apart, whereas monitor B is only 12 km from monitors D and C, 8 km
from monitor E, and 6 km from monitor A. It is possible that monitors B and F both
captured a source that was either longer in range or more ubiquitous and so would have
been obscured by the stronger source signals at sites A, C, D, and E. Comparisons
between the non-source-oriented monitors revealed moderate correlation between sites
(G to J [R = 0.29 to 0.71, p = 0.37 to 0.65]). Sites G, H, I and J are all located in the
southwestern quadrant of Los Angeles. It is possible that they are also exposed to a
ubiquitous source that produces a common signal at these four sites.
2-172
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• TSP Sourc* MonNora
• TSP Non-source
• Counly-trased Population Cenl
Inlet.
Majot Highways
o( Walei
Urban Areas
Note: Monitor locations are denoted by green markers, and source locations are denoted by red markers. Top: view of all Pb FRM
monitors in Los Angeles County. Bottom left: Close up of the industrial site near Monitors C and D. Bottom right: Close up of the
populated area captured by Monitor F.
Figure 2-40 Pb TSP monitor and source locations within Los Angeles County,
CA (06-037), 2007-2009.
2-173
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I
Note: Clockwise from top left: January, April, July, and October. Note that the wind percentages vary from month to month.
Source: NRCS (2011).
Figure 2-41 Wind roses for Los Angeles County, CA, from meteorological data
at the Los Angeles International Airport, 1961-1990.
2-174
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Site
SITE ID
MEAN
SD
OBS
% BELOW
MDL
Source
orientation
A
06-037-
1404
0.074
0.040
66
0
Source
B
06-037-
1602
0.013
0.017
112
75
Source
C
06-037-
1405
0.68
1.0
617
0
Source
D
06-037-
1406
0.12
0.092
242
0
Source
E
06-037-
1403
0.068
0.052
128
0
Source
F
06-037-
1103
0.018
0.011
121
45
Non-
source
G
06-037-
1301
0.015
0.012
108
65
Non-
source
H
06-037-
4002
0.0083
0.0068
120
85
Non-
source
1
06-037-
4004
0.0087
0.0069
117
85
Non-
source
J
06-037-
5005
0.0040
0.0064
109
95
Non-
source
E
15
01
C.
o
J.U -
2.9-
2.8 -
2.7
2.6-
2.5
2.4-
2.3 -
2.2-
2.1 -
2.0-
1.9-
1.7-
1.6
1.4-
1.3 -
1.1 -
1.0-
0.9-
0.8 -
0.7
0.6-
0.5 -
0.4-
0.3 -
0.2-
0.1 -
0.0 -
A
*lt*+
B
= , 4
C
D
m
E
iii = i
F
G
*i
H
I
J
Y1
234 Y1234 Y1234 Y1234 Y1234 Y1234 Y1234 Y1234 Y1234 Y1234
season
Figure 2-42 Box plots of annual and seasonal 24-h Pb TSP concentrations
(ug/m3) from source-oriented and non-source-oriented monitors
within Los Angeles County, CA (06-037), 2007-2009.
2-175
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Table
A
B
C
D
E
2-20 Comparisons between Pb TSP concentrations from source-oriented and non-source-oriented
monitors within Los Angeles County, CA (06-037), 2007-2009.
A
Source
Source R 1.00
p 1.00
P90 0.00
COD 0.00
Source R
P
P90
COD
Source R
P
P90
COD
Source R
P
P90
COD
Source R
P
P90
COD
B
C
Source Source
-0.04 0
0.16 0,
0.08 0
0.63 0
1.00 0,
1.00 0,
0.00 3
0.00 0
1
1
0,
0,
.14
.10
.49
.64
.06
.05
.59
.96
.00
.00
.00
.00
D
Source
0.10
0.08
0.10
0.31
0.17
0.05
0.25
0.84
0.59
0.57
1.76
0.68
1.00
1.00
0.00
0.00
E
Source
0.
0.
0.
0.
-0
0.
0.
0.
0.
0.
2.
0.
0.
0.
0.
0.
1.
1.
0.
0.
17
27
10
34
.06
07
10
71
08
03
14
77
18
12
17
42
00
00
00
00
F
Non-
Source
0.03
-0.15
0.08
0.57
0.33
0.74
0.02
0.46
0.12
-0.08
3.59
0.95
0.33
0.17
0.24
0.78
0.05
0.13
0.10
0.61
G
Non-
Source
0.00
0.00
0.06
0.57
0.29
0.12
0.02
0.48
0.24
0.26
4.22
0.96
0.09
0.11
0.25
0.80
0.07
0.06
0.10
0.64
H
Non-
Source
-0.08
0.14
0.08
0.79
0.40
0.28
0.01
0.61
0.28
0.28
3.59
0.98
0.32
0.24
0.25
0.89
0.00
0.24
0.11
0.78
I
Non-
Source
-0.07
-0.02
0.08
0.77
0.22
0.11
0.02
0.60
0.18
0.20
3.59
0.98
0.20
0.21
0.25
0.89
0.09
0.07
0.11
0.79
J
Non-
Source
-0.27
-0.09
0.08
0.85
0.20
0.10
0.02
0.81
0.08
0.13
3.92
0.99
0.03
0.07
0.25
0.95
-0.07
0.18
0.11
0.90
2-176
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Table 2-20 (Continued): Comparisons between Pb TSP concentrations from source-oriented and non-source-oriented
monitors within Los Angeles County, CA (06-037), 2007-2009.
A
Source
F Non-Source R
P
P90
COD
G Non-Source R
P
P90
COD
H Non-Source R
P
P90
COD
1 Non-Source R
P
P90
COD
J Non-Source R
P
P90
COD
B C D E F G H
Non- Non- Non-
Source Source Source Source
Source Source Source
1.00 0.10 0.43
1.00 0.02 0.19
0.00 0.02 0.02
0.00 0.39 0.61
1.00 0.71
1.00 0.65
0.00 0.01
0.00 0.54
1.00
1.00
0.00
0.00
I
Non-
Source
0.34
0.09
0.02
0.58
0.55
0.39
0.02
0.61
0.60
0.51
0.01
0.55
1.00
1.00
0.00
0.00
J
Non-
Source
0.21
0.09
0.02
0.82
0.54
0.38
0.02
0.85
0.51
0.40
0.01
0.77
0.29
0.37
0.01
0.78
1.00
1.00
0.00
0.00
Each comparison contains (in order): Pearson rank-order correlation (R), Spearman rank-order correlation (p), the difference between the 90th and 10th percentile data (P90), and the
coefficient of divergence (COD).
2-177
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Figure 2-43 illustrates Pb monitor locations within Hillsborough and Pinellas Counties in
FL, which comprise the greater Tampa-St. Petersburg metropolitan area. Two source-
oriented monitors (A and B) were located within Hillsborough County, and one non-
source-oriented monitor (C) was located in Pinellas County. Monitor A was located
360 meters north-northeast of the EnviroFocus Technologies battery recycling facility,
which produced 1.3 tons/year (U.S. EPA. 2008d). and monitor B was located 320 meters
southwest of the same facility. Monitor C was located next to a two-lane road in Pinellas
Park, FL.
Figure 2-44 displays seasonal wind roses for the Tampa-St. Petersburg metropolitan area.
These wind roses suggest shifting wind directions throughout the winter, spring, and
summer. During the winter, the highest winds came from the north and northeast with
little influence from the west and southwest. During spring and summer, easterly and
westerly winds were evident from the wind rose, with winds from the west being slightly
higher in wind speed. During autumn, winds came predominantly from the northeast with
little signal from the west or south.
Seasonal and year-round concentrations are reported for Hillsborough and Pinellas
Counties in Figure 2-45. The average annual concentration at monitor A was 0.15 ug/m3,
and the 95th percentile was 0.70 ug/m3. During winter, the 60th percentile of the data met
the level of the NAAQS. At this site, the highest concentrations occurred during summer,
which corresponded to the time when westerly winds were stronger. Concentration data
at monitor B were much higher, with an annual average of 0.45 ug/m3 and a 95th
percentile of 1.9 ug/m3. Annually, the 55th percentile exceeded the level of the NAAQS,
and in autumn the 45th percentile exceeded the NAAQS. The highest concentrations
occurred in autumn, coinciding with the time when winds blew from the northeast, when
monitor B was most often downwind of the battery recycling facility. The non-source-
oriented monitor C always reported concentrations of 0.0 ug/m3. This is likely related to
its location next to a quiet road in a small city.
Intersampler correlations, shown in Table 2-21. illustrate that Pb has high intra-urban
spatial variability. The source-oriented monitors were anticorrelated (R = -0.09,
p = -0.08). This was likely related to the fact that they were designated to monitor the
same source and were downwind of the source at different times.
2-178
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Top: view of all Pb FRM monitors in Hillsborough and Pinellas Counties.
Bottom: Close up of industrial site around monitors A and B.
Figure 2-43 Pb TSP monitor locations within Hillsborough and Pinellas
Counties, FL (12-057 and 12-103), 2007-2009.
2-179
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I
•
I
I
Note: Clockwise from top left: January, April, July, and October. Note that wind percentages vary from month to month.
Source: NRCS (2011).
Figure 2-44 Wind roses for Hillsborough/Pinellas Counties, FL, obtained from
meteorological data at Tampa International Airport, 1961-1990.
2-180
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Site
SITE ID
12-057-1073
12-057-1066
12-103-3005
MEAN
0.15
0.45
0.00
SD
0.27
1.08
0.00
DBS
154
155
58
% BELOW MDL
20
95
Source orientation
Ł
1
c
O
c
0)
u
c
O
u
Source
Source
Non-source
3.0 -
2.9 -
2.8 -
2.7-
2.6 -
2.5 -
2.4 -
2.3 -
2.2 -
2.1 -
2.0 -
1.9 -
1.8 -
1.7 -
1.6 -
1.5 -
1.4 -
1.3 -
1.2 -
1.1 -
1.0-
0.9 -
0.8 -
0.7 -
0.6 -
0.5 -
0.4-
0.3 -
0.2 -
0.1 -
o.o -
A
L_
B
C
Y1234 Y1234 Y1234
season
Figure 2-45 Box plots of annual and seasonal 24-h Pb TSP concentrations
(ug/m3) from source-oriented and non-source-oriented monitors
within Hillsborough and Pinellas Counties, FL (12-057 and
12-103), 2007-2009.
2-181
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Table 2-21 Correlations between Pb TSP concentrations from source-oriented
and non-source-oriented monitors within Hillsborough and Pinellas
Counties, FL (12-057 and 12-103), 2007-2009.
A Source
B Source
C Non-source
R
P
P90
COD
R
P
P90
COD
R
P
P90
COD
A B
Source Source
1.00 -0.09
1.00 -0.08
0.00 1.20
0.00 0.71
1.00
1.00
0.00
0.00
C
Non-source
0.50
1.00
2.20
1.00
1.00
1.00
0.00
0.00
Each comparison contains (in order): Pearson rank-order correlation (R), Spearman rank-order correlation (p), the difference
between the 90th and 10th percentile data (P90), and the coefficient of divergence (COD).
Figure 2-46 illustrates Pb monitor locations within Cook County, IL. Eight monitors were
located within Cook County, four of which were designated by the Illinois Environmental
Protection Agency (IEPA) in data reported to the AQS as source-oriented and the other
four were non-source-oriented monitors. Monitor A was situated within 10 km of 6
sources ranging in emissions from 0.14 to 1.08 tons/year (U.S. EPA. 2008a). Monitor A
was also sited in the median of Interstate I-90/I-94. Monitor B was located on the
northern roadside of Interstate 1-290, 5 meters from the closest lane of traffic and was
within 10 km of 2 Pb sources (0.41 and 1.08 tons/year) (U.S. EPA. 2008a). Monitor C
was also located within 10 km of 6 sources in Cook County and Lake County, IN; the
largest of those sources was 2.99 tons/year and was located 8 km southeast of monitor C
(U.S. EPA. 2008a). Monitor C was placed on the roof of a high school. Monitor D was
located roughly 60 meters west of Interstate 1-294 and adjacent to O'Hare International
Airport. Monitor E was located on the rooftop of a building rented for government offices
in Alsip, IL, a suburb south of Chicago. This location was roughly 1 km north of
Interstate 1-294 but not located on an arterial road; it was 9 km southeast of a
0.56 tons/year source (U.S. EPA. 2008a). Monitor F was sited in the parking lot of a
water pumping station, 100 meters north of Interstate 1-90 and 300 meters northwest of
the junction between Interstates 1-90 and 1-94. This site was 2 km north-northwest of a
0.10 tons/year source (U.S. EPA. 2008a). Monitor G was situated atop an elementary
2-182
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school in a residential neighborhood on the south side of Chicago, roughly 100 meters
south of a rail line and over 300 meters west of the closest arterial road. Although not
designated as a source monitor, monitor G was located 2 km southwest of facilities
emitting 0.30 and 0.41 tons/year (U.S. EPA. 2008a'). Monitor H was sited on the grounds
of the Northbrook Water Plant. Interstate 1-94 curves around this site and was
approximately 700 meters from the monitor to the east and around to the north. Figure
2-47 displays seasonal wind roses for Cook County. Wind patterns were quite variable
during each season for this area. During the winter, winds mostly came from the west,
with smaller contributions from the northwest, southwest, and south. In spring,
measurable winds were omni-directional, with the highest winds coming from the south
and northeast. Winds originated predominantly from the southwest and south during the
summer, with measurable contributions from the northeast as well. In autumn, wind flow
was predominantly from the south, but smaller contributions also came from the
southwest, west, and northwest.
Figure 2-48 presents seasonal box plots of Pb concentration at the eight monitors located
within Cook County. The maximum 95th percentile concentration on this plot was
0.14 ug/m3, so the scale of this box plot makes the variability in these data appear wider
than the data presented for Los Angeles County and Hillsborough/Pinellas Counties.
Monitor C was in closest proximity to the industrial steel facilities located in Lake
County, IN. The average of concentrations measured at monitor C was 0.031 ug/m3, with
a median of 0.02 ug/m3 and a maximum concentration of 0.31 ug/m3. In winter, the 95th
percentile of data was 0.14 ug/m3. The higher values could potentially be attributed to
transport of emissions; winds blow from the southeast roughly 10-15% of the time
throughout the year. No other monitors in Cook County reported values above the level
oftheNAAQS.
Three "near-road" monitors, A, B, and D can be compared with the other monitors to
consider the possibility of roadside resuspension of Pb dust from contemporaneous
sources, as discussed in Section 2.2.2.6. It would be expected that resuspension would
diminish with distance from the road. The 2 roadside monitors, A and B, reported
average concentrations of 0.030 ug/m3 and 0.024 ug/m3, respectively. The median
concentrations for monitors A and B were 0.02 ug/m3. Fifteen percent of data were below
the MDL for monitor A, and 25% were below the MDL for monitor B. Note that data
obtained from monitor A may reflect industrial emissions as well. Monitor D was located
roughly 60 meters from the closest interstate and 570 meters from the closest runway at
O'Hare International Airport. The average concentration at this site was 0.012 ug/m3, and
85% of data were below the MDL. Non-source monitors, E, F, G, and H had average
concentrations of 0.011-0.017 ug/m3. It is possible that the difference between Pb
2-183
-------
concentrations at monitors A and B and Pb concentrations at the other monitors was
related to proximity to the roadway, although this cannot be stated with certainty without
source apportionment data to confirm or refute the influence of industrial plumes from
Lake County, IN or local sources at each of the monitors.
Comparison among the monitor data demonstrates a high degree of spatial variability
(Table 2-22). None of the source-oriented monitors were well correlated with each other.
The highest correlation between source-oriented monitors occurred for monitors (A and
B [R = 0.32, p = 0.26]). This might have reflected more substantial differences related to
the additional influence of industrial sources nearby monitor A. Monitors (C and D) were
uncorrelated with each other and with monitors (A and B), likely because their exposure
to sources was substantially different. The source-oriented and non-source-oriented
monitors were generally not well correlated. The highest Spearman correlation occurred
between monitors D and H (p = 0.53), but Pearson correlation was much lower for this
pair (R = 0.19). Both were located on the north side of Cook County, but monitor H was
roughly 20 km northeast of monitor D. Winds blew from the southwest roughly 20-30%
of the time throughout the year and from the northeast 20-25% of the time between the
months of March and July, so the correlation may have been related to a common signal
transported across both sites. Monitors B and F (R= 0.52, p = 0.46) were also moderately
correlated. Monitor F is roughly 12 km northeast of monitor B, so the same common
wind influence for monitors D and H may have also caused the moderate correlation
between monitors (B and F). Monitor F was also moderately correlated with the other 3
non-source monitors (R = 0.42 to 0.54, p = 0.36 to 0.45), and the correlation between
monitors (E and G) was moderate (R = 0.65, p = 0.40). The data from monitor H did not
correlate well with those from monitors E and G. The non-source monitors were oriented
from north to south over a distance of roughly 50 km in the following order: monitor H,
monitor F, monitor G, and monitor E. The correlation pattern may have been related to
distance between samplers. Monitor H was located in the suburb of Northbrook, monitors
F and G were sited within the Chicago city limits, and monitor E was situated in a town
near the south side of Chicago. Differences among land use may have been related to the
lack of correlation of the monitor H data with those from monitors E and G. It is likely
that data from monitor F was at times better correlated with monitors E and G and at
other times with monitor H, since it had moderate correlation with all three other
non-source monitors.
2-184
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Legend
O TSP Source Monitors
• TSP Non-source Monitors
• City-based Population Cenler
* Courtly-based Population Cenlei
— Inierstates
Major Highways
Bodies of Water
Urban Aieas
Cook County. IL
«"-****
0 S 10
mom
Top: view of all Pb FRM monitors in Cook County.
Bottom left: Close up of the high traffic site around monitor A.
Bottom right: Close up of O'Hare International Airport adjacent to monitor D.
Figure 2-46 Pb TSP Monitor locations within Cook County, IL (17-031),
2007-2009.
2-185
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I
I
I
I
Note: Clockwise from the top left: January, April, July, and October. Note that the wind percentages vary from month to month.
Source: NRCS (2011)
Figure 2-47 Wind roses for Cook County, IL, obtained from meteorological
data at O'Hare International Airport, 1961-1990.
2-186
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Site
SITE ID
MEAN
SD
DBS
% BELOW
MDL
Source
orientation
0.15 -
0.14 -
0.13 -
0.12 -
0.11 -
0.10 -
sr 0.09 -
E
^ 0.08 -
° 0.07 -
2
c 0.06 -
Ol
o
§ 0.05 -
0.04 -
0.03 -
0.02 -
0.01 -
0.00 -
A
B
C
D
E F
G H
17-031-0026 17-031-6003 17-031-0022 17-031-3103 17-031-0001 17-031-0052 17-031-3301 17-031-4201
0.030
0.020
179
15
0.024
0.013
175
25
Source
Source
0.031
0.012
0.036 0.0062
177
25
168
85
Source Source
0.013 0.017
0.0078 0.0098
177 175
75 55
Non- Non-
source source
I
I
1
,
'I
A
I
I
B
1
C
I!
1
D
l.l
E
.nil
F
III
I
0.017 0.011
0.0097 0.0031
171 168
50 95
Non- Non-
source source
G
I
Illl
H
.i.l.
Y1234 Y1234 Y1234 Y1234 Y1234 Y1234 Y1234 Y1234
season
Figure 2-48 Box plots of annual and seasonal 24-h Pb TSP concentrations
(ug/m3) from source-oriented and non-source-oriented monitors
within Cook County, IL (17-031), 2007-2009.
2-187
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Table 2-22
Correlations between Pb TSP concentrations from source-oriented
and non-source-oriented monitors within Cook County, IL
2007-2009.
A Source
B Source
C Source
D Source
E Non-Source
F Non-Source
G Non-Source
H Non-Source
R
P
P90
COD
R
P
P90
COD
R
P
P90
COD
R
P
P90
COD
R
P
P90
COD
R
P
P90
COD
R
P
P90
COD
R
P
P90
COD
A B C D
Source Source Source Source
1.00 0.32 0.00 0.05
1.00 0.26 -0.01 0.08
0.00 0.03 0.06 0.04
0.00 0.29 0.38 0.43
1.00 0.14 0.07
1.00 0.05 0.10
0.00 0.04 0.03
0.00 0.33 0.36
1.00 0.01
1 .00 0.04
0.00 0.05
0.00 0.40
1.00
1.00
0.00
0.00
E
Non-
Source
0.17
0.06
0.04
0.41
0.54
0.32
0.03
0.34
0.24
0.16
0.05
0.39
0.18
0.21
0.01
0.19
1.00
1.00
0.00
0.00
F
Non-
Source
0.39
0.32
0.03
0.36
0.52
0.46
0.02
0.29
0.05
0.10
0.04
0.35
0.12
0.37
0.01
0.24
0.42
0.36
0.02
0.24
1.00
1.00
0.00
0.00
G
Non-
Source
0.34
0.18
0.03
0.36
0.60
0.35
0.02
0.30
0.19
0.17
0.05
0.35
0.08
0.07
0.02
0.28
0.65
0.40
0.01
0.24
0.54
0.41
0.01
0.24
1.00
1.00
0.00
0.00
(17-031),
H
Non-
Source
0.06
0.06
0.04
0.45
0.06
-0.01
0.03
0.40
-0.04
0.06
0.05
0.42
0.19
0.53
0.01
0.15
-0.01
0.07
0.01
0.20
0.42
0.45
0.02
0.26
0.01
0.05
0.02
0.27
1.00
1.00
0.00
0.00
Each comparison contains (in order):
between the 90th and 10th percentile
Pearson rank-order correlation (R), Spearman rank-order correlation (p), the difference
data (P90), and the coefficient of divergence (COD).
2-188
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Figure 2-49 illustrates Pb monitor locations with Jefferson County, MO. Ten source-
oriented monitors surrounded the Doe Run primary Pb smelter in Herculaneum, MO on
the west and northwestern sides. The largest distance between these monitors was
approximately 1.5 km. Monitor E located on the Doe Run facility roughly 20 meters west
of the nearest building. Monitors A, B, C, D, F, G, and H were all located approximately
200 meters west of the facility. Monitors D, E, and H were situated alongside service
roads to the facility. Monitor I was sited 100 meters north of the smelter, and monitor J
was located approximately 600 meters northwest of the facility. The Doe Run smelter
was the only active primary smelter in the U.S. at the time of this review, and the facility
was estimated to have emitted 41.1 tons Pb/year (U.S. EPA, 2008f). Figure 2-50 displays
seasonal wind roses for Jefferson County. During winter, predominant winds originated
from the northwest, with a smaller fraction of calmer winds originating in the south-
southeast. During the spring, the south-southeasterly winds became more prevalent with a
measurable fraction of stronger winds still originating in the north-northwest. In the
summer, winds were omni-directional and generally calmer. A slightly larger percentage
came from the south compared with other wind directions. Autumn winds were most
predominantly south-southeastern, with a smaller fraction from the west and northwest.
Figure 2-51 illustrates the seasonal distribution of concentrations at monitors A-J in
Jefferson County. The annual average concentrations ranged from 0.18 to 1.36 ug/m3
across the monitors. The maximum concentration was measured at monitor C to be
21.6 ug/m3, which was 144 times higher than the level of the standard. For this monitor,
the 25th percentile of the data was at the level of the standard. In general, median and
75th percentile concentrations were highest during the springtime and second highest
during the fall. These seasons coincide with periods when the southeastern winds were
stronger and more prevalent. Because the Doe Run facility had two 30-meter stacks
(Bennett. 2007). it is possible that the Pb measured at the closer monitors were due to
either fugitive emissions from the plant; or, if vehicles and ground equipment were
operated nearby, the previously-deposited emissions from the plant were resuspended.
Spatial variability among the monitors is lower than at many sites, because the monitors
are relatively close together and are located on one side of the same source (Table 2-23).
Correlations range substantially (R = -0.03 to 0.96, p = -0.04 to 0.96). High correlations
(R > 0.75, p > 0.75) occurred for monitors (A and C), (A and D), (C and D), (D and F),
(E and F), (G and H), and (I and J). Monitors (A and C), (A and D), (C and D), (D and F),
(E and F), and (G and H) are all within 250 meters of each other. For the highest
correlation (R = 0.96, p = 0.96, [for monitors E and F]), monitor F is 250 meters directly
east of monitor E. Low correlation (R < 0.25, p < 0.25) generally occurred when monitors
B, I, and J were compared with monitors A, C, D, E, F, G, and H. Monitors B, I, and J
were on the outskirts of the measurement area and so were likely oriented such that the
2-189
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southeasterly winds did not carry pollutants to these sites concurrently with the signal
recorded by the other monitors.
Note: All monitors surround the Doe Run industrial facility. Top: Map view of all monitors in Jefferson County. Bottom: Satellite view
of the monitors and the Doe Run facility.
Figure 2-49 Pb TSP Monitor locations within Jefferson County, MO (29-099),
2007-2009.
2-190
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i
I
i
•
Note: Clockwise from top left: January, April, July, and October. Note wind percentages vary from month to month.
Source: NRCS (2011)
Figure 2-50 Wind roses for Jefferson County, MO, obtained from
meteorological data at St. Louis/Lambert International Airport,
1961-1990.
2-191
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Site
SITE ID
MEAN
SD
OBS
% BELOW
MDL
Source
orientation
7.0-
A
29-099-
0022
0.43
0.54
622
0
Source
B
29-099-
0024
0.36
0.49
209
5
Source
C
29-099-
0015
1.36
1.97
1E3
0
Source
D E F G
29-099- 29-099- 29-099- 29-099-
0023 0004 0020 0021
0.39 1.12 0.69 0.75
0.54 1.67 1.01 1.25
632 1E3 575 953
0505
Source Source Source Source
I I I
H
29-099-
0005
0.29
0.59
351
25
Source
1
29-099-
0011
0.34
0.85
366
5
Source
J
29-099-
0013
0.18
0.33
177
15
Source
7.0 -
6.5-
6.0 -
5.5 -
5.0 -
4.5 -
4.0 -
3.5 -
3.0
2.5 -
2.0-
1.5-
1.0-
0.5 -
A
I.
II
!
i
B
ll
III
C
D
E
I
F
|
!
G
H
,
ll
I
!
I
,
[I
ii
j
i.i.i
Y1234 Y1234 Y1234 Y1234 Y1234 Y1234 Y1234 Y1234 Y1234 Y1234
season
Figure 2-51 Box plots of annual and seasonal 24-h Pb TSP concentrations
(ug/m3) from source-oriented and non-source-oriented monitors
within Jefferson County, MO (29-099), 2007-2009.
2-192
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Table 2-23 Correlations between Pb TSP concentrations from source-oriented
and non-source-oriented monitors within Jefferson County, MO
(29-099), 2007-2009.
A Source
B Source
C Source
D Source
E Source
F Source
G Source
R
P
P90
COD
R
P
P90
COD
R
P
P90
COD
R
P
P90
COD
R
P
P90
COD
R
P
P90
COD
R
P
P90
COD
A B C D
Source Source Source Source
1.00 0.66 0.80 0.84
1.00 0.59 0.80 0.83
0.00 0.71 1.55 0.42
0.00 0.46 0.48 0.30
1.00 0.54 0.40
1.00 0.53 0.43
0.00 1.86 0.87
0.00 0.58 0.51
1.00 0.86
1.00 0.86
0.00 1.56
0.00 0.50
1.00
1.00
0.00
0.00
E
Source
0.60
0.57
1.93
0.55
0.15
0.10
2.77
0.69
0.56
0.59
2.26
0.50
0.70
0.71
1.83
0.50
1.00
1.00
0.00
0.00
F
Source
0.65
0.64
1.14
0.45
0.15
0.14
1.96
0.62
0.72
0.72
1.26
0.46
0.80
0.80
1.02
0.36
0.96
0.96
0.86
0.35
1.00
1.00
0.00
0.00
G
Source
0.33
0.33
1.41
0.57
0.08
0.07
2.08
0.68
0.28
0.26
2.94
0.60
0.41
0.41
1.38
0.53
0.57
0.54
2.16
0.49
0.56
0.56
1.13
0.47
1.00
1.00
0.00
0.00
H
Source
0.32
0.35
0.74
0.64
0.16
0.22
0.94
0.68
0.32
0.27
2.65
0.74
0.48
0.56
0.76
0.61
0.53
0.46
2.50
0.66
0.56
0.54
1.51
0.63
0.85
0.87
1.53
0.61
I
Source
0.07
0.07
0.92
0.67
0.11
0.10
1.04
0.65
-0.03
-0.04
3.18
0.73
0.17
0.14
0.88
0.63
0.09
0.06
3.09
0.70
0.12
0.10
1.74
0.65
0.36
0.28
2.10
0.63
J
Source
0.05
0.05
0.78
0.69
0.01
0.09
0.91
0.65
-0.03
0.04
2.60
0.73
0.10
0.18
0.70
0.66
0.14
0.16
2.57
0.72
0.20
0.19
1.40
0.70
0.34
0.38
2.08
0.66
2-193
-------
Table 2-23 (Continued): Correlations between Pb TSP concentrations from source-oriented
and non-source-oriented monitors within Jefferson County, MO
(29-099), 2007-2009.
ABODE
H Source R
P
P90
COD
1 Source R
P
P90
COD
J Source R
P
P90
COD
F G H 1
1.00 0.24
1.00 0.20
0.00 0.89
0.00 0.67
1.00
1.00
0.00
0.00
J
0.33
0.30
0.56
0.65
0.87
0.79
0.62
0.48
1.00
1.00
0.00
0.00
Each comparison contains (in order): Pearson rank-order correlation (R), Spearman rank-order correlation (p), the difference
between the 90th and 10th percentile data (P90), and the coefficient of divergence (COD).
Figure 2-52 illustrates Pb monitor locations in Cuyahoga County, OH. Five monitors are
located within Cuyahoga County, three of which were designated by the Ohio EPA
(OEPA) as source-oriented and the other two were non-source-oriented monitors.
Monitors A, B, and C were all located within 1-10 km of six 0.1 tons/year source
facilities and one 0.2 tons/year source (U.S. EPA. 2008g). Additionally, monitor B was
located 30 meters north of the Ferro Corporation headquarters. This facility was stated in
the 2005 NEI to have no emissions, but it was thought by the OEPA to be the source of
exceedances at this monitor (U.S. EPA. 2008g). Monitor A was sited roughly 300 meters
south of the Ferro Corporation facility. Monitor C was located 2.2 km west-northwest of
the 0.5 tons/year Victory White Metal Co. facility. Monitor C was also roughly 250
meters southeast of Interstate 1-490. Monitors D and E were designated as non-source-
oriented monitors, although monitor D was just 600 meters further from the Victory
White Metal facility than was monitor C. Monitor D was sited on a residential street
located 50 meters north of Interstate 1-490. Monitor E was located on the rooftop of a
building within 20 meters of a four-lane arterial road. Figure 2-53 displays seasonal wind
roses for Cuyahoga County. During winter, summer, and autumn, the predominant winds
were from the southwest, with stronger winds recorded during the winter. In the spring,
the strongest winds still emanated from the south-southwest, but measurable winds were
also scattered from the northeast to the northwest.
Figure 2-54 illustrates the seasonal distribution of Pb concentration data at the five
monitoring sites. The influence of southern winds, along with close proximity to a
2-194
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potentially-emitting facility, could have caused the elevated concentrations observed at
monitor B (average: 0.10 ug/m3). The 80th percentile of data was at the level of the
NAAQS at this monitor, and during autumn the 60th percentile of data met the level of
the NAAQS. The maximum concentration during fall and for the monitor year-round was
0.22 ug/m3. Concentration data from all other monitors were below the level of the
NAAQS. For monitor A, the average concentration was 0.025 ug/m3, and the median
reached 0.04 ug/m3 during the summer. Maximum concentration at this monitor was
0.07 ug/m3. Concentrations at monitor C averaged 0.017 ug/m3, and those at monitors D
and E averaged 0.014 ug/m3 and 0.013 ug/m3, respectively. Maximum concentrations
reached 0.04 ug/m3 at all three monitors.
The level of spatial variability is illustrated by the intersampler correlations presented in
Table 2-24. Monitors A and B appear to be anticorrelated (R = -0.06, p = -0.13). If the
Ferro site was the dominant source in this area, then the anticorrelation was likely caused
by the positioning of monitors A and B on opposite sides of that facility. At any given
time, potential emissions from the Ferro plant may have affected monitors A and B at
distinct times. Monitors C, D, and E correlated moderately to well with each other
(R = 0.37 to 0.74, p = 0.67 to 0.77). Given that all 3 monitors are separated by roughly
2.8 km, it is possible that the relatively high correlations related to common sources, as
suggested in the previous paragraph. Little correlation was observed between the source-
oriented and non-source-oriented monitors.
2-195
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Legend
• TSP Source Monitors
• TSP Non-source Monitors
• City-based Population Center
• County-based Popu&lion Center 1 w
h***. «-^
Majoi Highways
^B Bodies ol Water
Urban Areas
Cuyahoga County. OH
0 5 10 3
Note: Top: view of all Pb FRM monitors in Cuyahoga County. Bottom left: Close up of industrial site around monitors A and B.
Bottom right: Close up of monitor D north of Interstate I-490.
Figure 2-52 Pb TSP Monitor locations within Cuyahoga County, OH (39-035),
2007-2009.
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I
I
I
I
Note: Clockwise from top left: January, April, July, and October. Note wind percentages vary from month to month.
Source: NRCS (2011)
Figure 2-53 Wind roses for Cuyahoga County, OH, obtained from
meteorological data at Cleveland/Hopkins International Airport,
1961-1990.
2-197
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Site
SITE ID
MEAN
SD
DBS
% BELOW MDL
Source orientation
0.25 -
0.24 -
0.23 -
0.22 -
0.21 ;
0.20 -
0.19 -
0.18 -
0.17 -
0.16 -
"E 0.15 -
"M 0.14 -
~ 0.13 -
•B 0.12 :
(D
Ł 0.11 -
g 0.10 -
I °-09:
0.08 -
0.07 -
0.06 -
0.05 -
0.04 -
0.03 -
0.02 -
0.01 -
o.oo -
A
B
39-035-0050 39-035-0049
0.025
0.018
36
20
Source
0.10
0.060
36
0
Source
A
|
i
I
B
C
D E
39-035-0061 39-035-0038 39-035-0042
0.017
0.010
36
30
Source
0.014 0.013
0.0072 0.0076
35 36
45 45
Non-source Non-source
C
^
I
D
i(
i,
E
Lit
Figure 2-54
Y1234 Y1234 Y1234 Y1234 Y1234
season
Box plots of annual and seasonal 24-h Pb TSP concentrations
(ug/m3) from source-oriented and non-source-oriented monitors
within Cuyahoga County, OH (39-035), 2007-2009.
2-198
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Table 2-24 Correlations between Pb TSP concentrations from source-oriented
and non-source-oriented monitors within Cuyahoga County, OH
(39-035), 2007-2009.
A Source
B Source
C Source
D Non-Source
E Non-Source
R
P
P90
COD
R
P
P90
COD
R
P
P90
COD
R
P
P90
COD
R
P
P90
COD
ABC
Source Source Source
1.00 -0.06 0.21
1.00 -0.13 0.24
0.00 0.18 0.05
0.00 0.64 0.33
1.00 0.26
1.00 0.31
0.00 0.18
0.00 0.69
1.00
1.00
0.00
0.00
D
Non-Source
0.17
0.19
0.04
0.35
0.43
0.24
0.19
0.71
0.74
0.77
0.01
0.17
1.00
1.00
0.00
0.00
E
Non-Source
0.24
0.21
0.05
0.37
0.11
0.34
0.19
0.73
0.51
0.67
0.01
0.18
0.37
0.67
0.01
0.17
1.00
1.00
0.00
0.00
Each comparison contains (in order): Pearson rank-order correlation (R), Spearman rank-order correlation (p), the difference
between the 90th and 10th percentile data (P90), and the coefficient of divergence (COD).
Figure 2-55 illustrates Pb monitor locations within Sullivan County, TN. Three source-
oriented monitors were situated around an Exide Pb recycling facility emitting
0.78 tons/year (U.S. EPA. 2008h). Monitors A and C are positioned along the facility's
service road and are approximately 100 meters and 200 meters away from the facility,
respectively. Monitor A is directly next to the road, and monitor C is roughly 15 meters
from the road. Monitor B is located in the facility's parking lot roughly 50 meters from
the closest building. The facility and all three monitors are approximately 1.5 km
northwest of the Bristol Motor Speedway and Dragway racetracks, which hosts a variety
of auto races each year, including NASCAR, KART, and drag racing. Although the
NASCAR circuit no longer uses tetraethyl Pb as an anti-knock agent in its fuel, some of
2-199
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the smaller racing circuits continue to do so. However, the speedway is rarely upwind of
the monitoring sites and so likely had minimal influence on the reported concentrations.
Figure 2-56 displays seasonal wind roses for Sullivan County. During winter and spring,
the predominant winds come from the southwest and west. In the summer, the percentage
of wind coming from the west and southwest is roughly equal to that for wind coming
from the east and northeast, although the easterly winds are calmer. During autumn,
winds come predominantly from the northeast and east, although these winds tend to be
calmer than those originating from the southwest and west.
The data presented in Figure 2-57 illustrates that concentrations above the level of the
NAAQS occurred frequently at the monitors. The average concentrations at monitors A,
B, and C were 0.11 ug/m3, 0.051 ug/m3, and 0.059 ug/m3, respectively. Median
concentrations were 0.08 ug/m3, 0.03 ug/m3, and 0.04 ug/m3, respectively. The 75th
percentile of year-round data at monitor A was at the level of the NAAQS, while the 95th
percentile of data were below the NAAQS level for monitors B and C. The maxima at
each monitor were 0.76 ug/m3, 0.26 ug/m3, and 0.43 ug/m3 for monitors A, B, and C. The
concentrations measured at monitor A tended to be higher because the predominant and
stronger winds came from the southwest, so in many cases monitor A was upwind of the
facility. It is possible that Pb that had either deposited or was stored in waste piles
became readily resuspended by traffic-related turbulence and was measured at monitor A
since that monitor was closest to the road. The slightly higher concentrations at monitor
A compared with those from monitor C are consistent with the southwestern winds.
Not surprisingly, the correlations of monitor A with monitors B and C (R = 0.06 to 0.14,
p = -0.04 to 0.13) were quite low (Table 2-25). The correlation between monitors B and
C was moderate (R = 0.31, p = 0.45). It makes sense that the correlation for these
monitors would be somewhat higher because they are both oriented to the east of the Pb
recycling facility, although monitor C is to the northeast and monitor B to the east-
southeast.
2-200
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Legend
• TSPSource Monitors
# City-based Population Canter
• County-based Population Center
Interstate*
Major Highways
Bodies of Water
Urban Areas
Sullivan County, TN
0 S 10 20 Kilometers
Note: Top: Map, bottom: Satellite image. Monitors A, B, and C surround the Exide Pb recycling facility. Just to the southeast is the
Bristol motor speedway.
Figure 2-55 Pb TSP Monitor locations within Sullivan County, TN (47-163),
2007-2009.
2-201
-------
i
•
S:
Source: NRCS (2011)
Note: Clockwise from top left: January, April, July, and October. Note that the wind percentages vary from month to month.
Figure 2-56 Wind roses for Sullivan County, TN, obtained from meteorological
data at Bristol/Tri City Airport, 1961-1990.
2-202
-------
Site A
B
C
SITE ID 47-163-3001 47-163-3002 47-163-3003
MEAN 0.11
SD 0.11
DBS 334
% BELOW MDL 0
Source orientation
0.44 -
0.42 -
0.40 -
0.38 -
0.36 -
0.34 -
0.32 -
0.30 -
0.28 -
°E 0.26 -
M o 24 -
7 0.22 -
0
'Ł 0-20 -
c 0.18 -
0)
c 0.16 -
0
'-' 0.14 -
0.12 -
0.10 -
0.08 -
0.06 -
0.04 -
0.02 -
o.oo -
Source
A
0.051
0.036
362
0
Source
0.059
0.047
345
0
Source
B
C
Figure 2-57
Y1234 Y1234 Y1234
season
Box plots of annual and seasonal 24-h Pb TSP concentrations
(ug/m3) from source-oriented monitors within Sullivan County, TN
(47-163), 2007-2009.
2-203
-------
Table 2-25 Correlations between Pb TSP concentrations from source-oriented
monitors within Sullivan County, TN (47-163), 2007-2009.
A
Source
A Source R 1.00
p 1.00
P90 0.00
COD 0.00
B Source R
P
P90
COD
C Source R
P
P90
COD
B
Source
0.06
-0.04
0.21
0.47
1.00
1.00
0.00
0.00
C
Source
0.14
0.13
0.19
0.43
0.31
0.45
0.06
0.23
1.00
1.00
0.00
0.00
Each comparison contains (in order): Pearson rank-order correlation (R), Spearman rank-order correlation (p), the difference
between the 90th and 10th percentile data (P90), and the coefficient of divergence (COD).
2.8.3 Seasonal Variation in Pb Concentrations
Monthly average Pb concentrations averaged over multiple sites and over 3 years from
2008-2010 are shown for Pb-TSP from source-oriented monitors (Figure 2-58). Pb-TSP
from non-source-oriented monitors (Figure 2-59). Pb-PMi0 (Figure 2-60). and Pb-PM2 5
(Figure 2-61). For source-oriented Pb-TSP (Figure 2-58). monthly average concentrations
were determined from between 146 and 154 samples in each month. For non-source-
oriented TSP (Figure 2-59). monthly average concentrations were determined from
between 141 and 151 samples in each month. A winter minimum was observed with
December, January, and February exhibiting the three lowest monthly averages. In both
cases, there is little seasonal variation. Minor variations in monthly averages are probably
driven by exceptional events. Monthly median concentrations are very similar for all
months.
2-204
-------
CO
2
-------
0.05
0.04-
0.03 -
o 0.02 -
0.01
0.00-
A
1
t
r
•}
i
*
c
Jan
Feb
Mar Apr May
Jun
Jul
Aug
Sep
Oct
Nov
Dec
Month
Note: Box and whisker plots are used for each month, with the box comprising the interquartile range of the data and the whiskers
comprising the range within the 5th to 95th percentiles. The median is noted by the red line, and the blue star denotes the mean.
Figure 2-59 Monthly non-source-oriented Pb-TSP average (ug/m3) over 12
months of the year, 2008-2010.
0.028 -
0.026 -
0.024 •
0.022 -
0.020 -
CO
j= 0.018-
O)
-3 0.016-
Ł 0.014-
•E 0.012-
0)
i 0.010-
o
0 0.008 -
0.006 •
0.004 -
0.002 -
0.000 -
T
-L
i
Jan
I
i
Feb
_3
r
==
1
Mar
3
f
->-
I
Apr
3
C
-L-
1
May
=^
Jun
— . —
r^
f-l
-L-
i
Jul
3
C
-L
i
Aug
3
C
-L-
i
Sep
•}
c
1
I
Oct
__^_
3
C
J-
I
Nov
3
c
-L-
Dec
Month
Note: Box and whisker plots are used for each month, with the box comprising the interquartile range of data and the whiskers
comprising the range from 5th to 95th percentiles. The median is noted by the red line, and the blue star denotes the mean.
Figure 2-60 Monthly Pb-PMi0 average (ug/m3) over 12 months of the year,
2007-2009.
2-206
-------
0.010-
0.009 -
0.008 -
ff 0.007-
E
g 0.006 -
•2 0.005 -
1 0.004 •
|
° 0.003 -
0.002 -
0.001 •
0.000 -
*
_l_
i
Jan
*
i
i
Feb
—
_L
Mar
3
C
J_
Apr
3
C
J_
i
May
-j
L-
*r-
_l_
i
Jun
3
C
_l_
i
Jul
__^
3
C
_l_
i
Aug
p
C
_L
i
Sep
J_
i
Oct
J
C
i
Nov
3
I
_L
Dec
Month
Note: Box and whisker plots are used for each month, with the box comprising the interquartile range of the data and whiskers
comprising the range from 5th to 95th percentiles. The median is noted by the red line, and the blue star denotes the mean.
Figure 2-61 Monthly Pb-PM2s average (ug/m3) over 12 months of the year,
2007-2009.
For both Pb-PM10 (Figure 2-60) and Pb-PM2 5, (Figure 2-61) there is also little seasonal
variation, with minor fluctuations in monthly averages probably driven by exceptional
events, and similar monthly median concentrations for all months. Pb-PMi0 monthly
average concentrations were determined from between 100 and 109 samples and
Pb-PM2.5 from between 866 and 1,034 samples each month.
2.8.4 Size Distribution of Pb-bearing PM
Table 2-26 presents data for co-located Pb-TSP, Pb-PMi0, and/or Pb-PM2 5 monitors.
Table 2-27 contains metadata for studies in Section 2.5.3 involving size distribution data,
and Table 2-28 contains the size distribution data for those studies. At times, the data
were extracted from figures in the original references.
2-207
-------
Table 2-26
Site ID*
060190008
060190008
060190008
060190008
060250004
060250005
060250005
060251003
060290004
060290004
060290014
060290014
Correlations and average of the concentration ratios for co-located monitors,
TSP versus PM2.5, and PM™ versus PM2.5.
CBSA
Fresno, CA
Fresno, CA
Fresno, CA
Fresno, CA
El Centra, CA
El Centra, CA
El Centra, CA
El Centra, CA
Bakersfield, CA
Bakersfield, CA
Bakersfield, CA
Bakersfield, CA
Land Type
Suburban
Suburban
Suburban
Suburban
Suburban
Suburban
Suburban
Urban and Center
City
Urban and
Center City
Urban and
Center City
Urban and
Center City
Urban and
Center City
Avg Avg
Years Corr Ratio Years Corr Ratio
PMi0:TSP PM2.5:TSP
9nni '-'•93 '-'•82 9nni Q-&2. 0.59
20o"' °'93 °'83 200* ~ °'80 °'56
1QQ6 - 1QQ6 -
2001 °'79 °'98 2001 °'77 °'73
^ 0.92 0.89 ^Ol" °'91 °'64
1992-
1994
}?9l~ °'94 °'43
1995- „„, „ ^r- 1994- „ „„ „ ^^
2000 °'94 °'75 2000 °'96 °'51
1QQE> - 1QQ4 -
2000 °'92 °'78 2000 °'92 °'53
TSP versus PM
Years Corr
PM2.5: PIV
1995 - 099
2001 U'aa
z- °™
z- •>•»
Ł?- "•«
1995 - 096
1996 °'96
1996 - 099
2001 U'aa
;s- »•«
z- «•-
%&- ^
10,
Avg
Ratio
lio
0.77
0.82
0.79
0.77
0.80
0.71
0.72
0.71
0.77
2-208
-------
Table 2-26 (Continued): Correlations and average of the concentration ratios for co-located monitors, TSP versus PMi0,
TSP versus PM2.s, and PMio versus PM2.s.
Site ID*
060290014
060290014
060292004
060310003
060310004
060370002
060374002
060374002
060390001
060631008
060658001
CBSA
Bakersfield, CA
Bakersfield, CA
Bakersfield, CA
Hanford-Corcoran,
CA
Hanford-Corcoran,
CA
Los Angeles-Long
Beach-Santa Ana,
CA
Los Angeles-Long
Beach-Santa Ana,
CA
Los Angeles-Long
Beach-Santa Ana,
CA
Madera-Chowchilla,
CA
NONE(PlumasCo.,
CA)
Riverside-San
Bernardino-Ontario,
f^ A
Land Type
Urban and
Center City
Urban and
Center City
Suburban
Unknown
Suburban
Suburban
Suburban
Suburban
Urban and
Center City
Unknown
Suburban
Avg Avg
Years Corr Ratio Years Corr Ratio Years
PM10:TSP PM2.5:TSP
2^o" °'47 °'80 2^o" °'28 °'60 2^o"
2^o" °'43 °'81 2^o" °'27 °'62 2^o"
1995-
2000
1995-
1998
1996-
2000
1995-
2000
%Ł- °™ °™ %Ł- °™ °™ ™-
%Ł- ™ ^ %Ł- ^ °"
1995-
1996
1997-
1999
1995- 1992- 1995-
1996 °'13 °'39 1996 °'31 °'33 1997
Corr
PM2.5: PM10
0.84
0.98
0.74
0.97
0.95
0.89
0.91
0.98
0.95
0.94
Avg
Ratio
0.80
0.72
0.74
0.83
0.77
0.59
0.62
0.90
0.72
0.67
2-209
-------
Table 2-26 (Continued): Correlations and average of the concentration ratios for co-located monitors, TSP versus PMi0,
TSP versus PM2.s, and PMio versus PM2.s.
Site ID*
060658001
060670010
060710014
060771002
060771002
060850004
060850004
060850004
060990002
060990002
060990005
060990005
061072002
CBSA
Riverside-San
Bernardino-Ontario,
CA
Sacramento— Arden-
Arcade-Roseville,
CA
Riverside-San
Bernardino-Ontario,
CA
Stockton, CA
Stockton, CA
San Jose-Sunnyvale-
Santa Clara, CA
San Jose-Sunnyvale-
Santa Clara, CA
San Jose-Sunnyvale-
Santa Clara, CA
Modesto, CA
Modesto, CA
Modesto, CA
Modesto, CA
Visalia-Porterville, CA
Land Type
Suburban
Urban and
Center City
Suburban
Urban and
Center City
Urban and
Center City
Urban and
Center City
Urban and
Center City
Urban and
Center City
Urban and
Center City
Urban and
Center City
Urban and
Center City
Urban and
Center City
Urban and
Center City
Avg Avg
Years Corr Ratio Years Corr Ratio Years
PM10:TSP PM2.5:TSP
^99?" °'93 °'72 \W~ °'86 °'46
1995-
2001
1996-
2000
"I QQ^ "I QQ9 "I QQ^
2000 0.70 0.84 ^DO °-59 °-56 2000
2000' °'91 °'74 20020~ °'76 °'48
1995- 1994- 1995-
2000 °'87 °'63 1997 °'11 °'36 2000
-IQQO _
'QQ^ 0.42 0.37
2^0 " °'54 °'39
1995- n QR n 7Q 1992- 0-24 0.61 1995-
"1 QQft 'J\* \j . i \j "IQQft "IQQft
1997-
1998
1998-
2000
1998-
2000
1995-
2000
Avg
Corr Ratio
PM2.5: PM10
0.99 0.75
0.78 0.73
0.94 0.71
0.95 0.69
0.64 0.80
0.50 0.73
0.99 0.71
0.97 0.71
0.99 0.70
2-210
-------
Table 2-26 (Continued): Correlations and average of the concentration ratios for co-located monitors, TSP versus PMi0,
TSP versus PM2.s, and PMio versus PM2.s.
Site ID*
170310022
170310052
171190010
171191007
171630010
180890023
201730007
201730008
201730009
201731012
201770007
202090015
202090020
270530053
300490719
CBSA
Chicago-Naperville-
Joliet, IL-IN-WI
Chicago-Naperville-
Joliet, IL-IN-WI
St. Louis, MO-IL
St. Louis, MO-IL
St. Louis, MO-IL
Chicago-Naperville-
Joliet, IL-IN-WI
Wichita, KS
Wichita, KS
Wichita, KS
Wichita, KS
Topeka, KS
Kansas City, MO-KS
Kansas City, MO-KS
Minneapolis-St. Paul-
Bloomington, MN-WI
Helena, MT
Land Type
Suburban
Suburban
Urban and
Center City
Urban and
Center City
Suburban
Urban and
Center City
Suburban
Suburban
Suburban
Suburban
Urban and
Center City
Urban and
Center City
Urban and
Center City
Urban and
Center City
Suburban
Years
Corr
Avg Avg Avg
Ratio Years Corr Ratio Years Corr Ratio
PM10:TSP PM2.5:TSP PM2.5: PM10
1992
1992
1992
1992
1992
2007-
2008
1990-
1997
1990-
1997
1990-
1997
1990-
1997
1990-
1997
1990-
1997
1990-
1997
1996-
2001
1990-
1991
0.81
0.84
0.40
0.92
0.96
0.73
0.18
0.34
0.54
0.85
0.56
0.75
0.99
0.54
0.81
0.94
0.86
0.96
0.91
0.91
0.86
1.28
1.12
1.05
0.89
0.99
0.80
0.81
0.57
0.48
300490719
Helena, MT
Suburban
1990
0.79 0.49
2-211
-------
Table 2-26 (Continued): Correlations and average of the concentration ratios for co-located monitors, TSP versus PMi0,
TSP versus PM2.s, and PMio versus PM2.s.
Site ID*
450430001
450790014
450791003
450791003
CBSA
Georgetown, SC
Columbia, SC
Columbia, SC
Columbia, SC
Land Type
Urban and
Center City
Suburban
Urban and
Center City
Urban and
Center City
Years
1990-
1991
1990
1990
1990
Avg Avg Avg
Corr Ratio Years Corr Ratio Years Corr Ratio
PM10:TSP PM2.5:TSP PM2.5: PM10
0.71 0.60
0.82 0.89
0.90 0.94
0.90 0.83
*Note: For comparability, comparisons were limited to samples from sites which had at least 30 pairs of co-located samples, with both samples above the MDL and where both
monitors reported data at STP.
2-212
-------
Table 2-27 Metadata for studies of Pb-PM size distribution.
Reference
Location
Nearest source
Proximity to source
Sampling dates
Yi et al. (2006)
Jersey City, NJ-
an urban/industrial area
New Brunswick, NJ-
a suburban area
Jersey City- near Manhattan, NJ
Turnpike, Hudson River- high gas/oil
consumption for industry/domestic
heating, heavy gasoline & diesel powered
vehicles and ship traffic from harbor.
New Brunswick- near NJ Parkway and
Garden State Parkway
Close
ASD measurements:
June 10-20, 2002
Bein et al. (2006)
Pittsburg, PA
Article does not describe sources; other
articles also use data from Pittsburgh Air
Quality Study (PAQS) and may have
more info on sources
July 2001-September
2002
Pekey et al. (2010)
Kocaeli, Turkey
Kocaeli is a very industrialized and
urbanized region in Turkey; sources
include a large refinery, a petrochemical
complex, a hazardous waste incinerator
and industry operations for textile,
machine, mine, metal, food, automotive,
paper, chemistry, wood, petroleum,
tanning, and coal sectors, plus heavy
traffic
Singh et al. (2002)
Downey, CA- a city in Los
Angeles County along
"Alameda corridor" joining
coastal area to downtown LA
Riverside, CA- an inland
county east of LA
Downey- a "source" site affected by fresh
PM emissions from nearby oil refineries,
industry, and heavy diesel emissions
Riverside- a "receptor" site affected by
aged PM emissions including high vehicle
emissions in LA
Downey: 10 km downwind of
refineries; 2-4 km from
Interstates 1-710 and I-605
Riverside- 70 km east of
downtown LA
Downey: September
2000-January 2001
Riverside: February
2001-June2001
Dall'Osto et al. (2008)
U.K. national air quality
monitoring station in
PortTalbot, U.K.
One of the U.K.'s largest integrated
steelworks; near major roadways; (Table
1 of this study describes the plants,
operations, emission types, and emission
components for steelwork sources)
Monitoring site next to
steelwork
April 24-May 5, 2006
2-213
-------
Table 2-27 (Continued): Metadata for studies of Pb-PM size distribution.
Reference
Location
Nearest source
Proximity to source
Sampling dates
Weitkamp et al. (2005)
Coking facility near Pittsburgh,
PA
Large coking facility that converts 6 million
tons of coal to 4 million tons of metallurgic
coke every year
Sampling site downwind,
directly across river from coke
facility(~400m)
August 22-September
5, 2002
Pekey et al. (2010)
Indoor/outdoor sample points
for 15 homes in Kocaeli,
Turkey
Kocaeli is a very industrialized and
urbanized region in Turkey; sources
include a large refinery, a petrochemical
complex, a hazardous waste incinerator
and industry operations for textile,
machine, mine, metal, food, automotive,
paper, chemistry, wood, petroleum,
tanning, and coal sectors, plus heavy
traffic
15 Kocaeli homes chosen as
representative sample; 10
close to high traffic roads, 5
near low/moderate traffic
roads
Summer: May 31-June
29, 2006
Winter: December 16-
January 20, 2007
Sabin et al. (2006b)
Near Interstate I-405 between
Sunset Blvd and Wilshire Blvd,
Los Angeles, CA
Heavy traffic on Interstate I-405 Freeway
UP: 150m upwind of Interstate
I-405 (background)
DW1: 10m downwind
DW2: 150m downwind
DW3: 450m downwind
April 13-May 1,2004
Song and Gao (2011) Carlstadt, NJ
Heavy traffic on NJ Turnpike
~5m from roadside of the
highway
Winter: December
2007-February 2008
Summer: July 2008
Zereini et al. (2005)
3 Sites in Frankfurt, Germany
with different traffic densities
Vehicle emissions
Site 1: next to main street with
32,500 cars/day
Site 2: next to side street with
<1,000 cars/day
Site 3: large garden 8km NW
of city
August 2001 to July
2002
Lough et al. (2005)
Two road traffic tunnels in
Milwaukee, Wl
Traffic emissions
5 m upwind from entrance
(inside tunnel); 15 m upwind
from tunnel exit
Summer: July 31-
August 28, 2000
Winter: December 13-
January 17, 2001
Hays et al. (2011)
20m downwind of Interstate
I-440 highway in Raleigh, NC
Traffic emissions from highway
20 m downwind
July 26-31 and August
3-10,2006
2-214
-------
Table 2-27 (Continued): Metadata for studies of Pb-PM size distribution.
Reference
Location
Nearest source
Proximity to source
Sampling dates
Chen et al. (201 Ob)
Roadside site and site inside
highway tunnel in Taipei,
Taiwan
Vehicle emissions
Roadside: sidewalk 4m from
road
Tunnel: relay station in tunnel
1.4km from outlet; 2 m from
traffic lane
January to December
2008
Birmili et al. (2006)
Remote background: Mace
Head atmospheric research
station in Connemara, Ireland
Urban background: University
of Birmingham campus, U.K.
Roadside: A38 Bristol Road,
Birmingham, U.K.
Road Tunnel: Queensway
Underpass in Birmingham,
U.K.
Traffic emissions
Remote background:
Urban background: at least
100m from road traffic
Roadside: 4 m from traffic
Road Tunnel: 30 m from
tunnel exit
Remote background:
August 8-28, 2002
Urban background:
April 23-October 7,
2002
Roadside: July 8-12,
2002
Road Tunnel: July 2,
2002
Bruggemann et al.
(2009)
Roadside in
Dresden, Germany
Traffic emissions from busy street (traffic
density-55,000 per day; 8% trucks),
tramline, railway station
Next to road, near tramline
crossing, 200 m to railway
station
September 2003-
August 2004
Harrison et al. (2003)
Roadside in
Birmingham, U.K.
Traffic emissions from A38
9m from road
October 26, 2000 to
January 17, 2001
Wang et al. (2006d)
Suburb of Kanazawa, Japan- a
western coastal city; the
largest in Hokuriku region of
Japan
Emissions from road traffic, nearby
incinerators and electricity generation
plants, and sea salt
Next to road; ~5km from
incinerators and electricity
generation plants; situated on
west coast
May-June 2003
Martuzevicius et al.
(2004)
9 Locations in Cincinnati, Ohio
metropolitan area
Vehicle emissions from Cincinnati
highway network; emission from industry
(233 facilities within municipal area limits)
11 Sites- varying distance to
linear and point sources;
distance to major highways
ranges from 210 m to 4,590 m
December 2001-
November2002
2-215
-------
Table 2-27 (Continued): Metadata for studies of Pb-PM size distribution.
Reference
Location
Nearest source
Proximity to source
Sampling dates
Moreno et al. (2008)
3 Sampling sites in Mexico
City (Mexico) Metropolitan
Area:
1 Site in the industrial center
(TO),
1 Site NE outside city limits
(T1), and
1 Rural site north of city (T2)
Urban pollution sources-traffic emissions,
industry
3 Sites with varying relation to
"Mexico urban plume" by
distance, wind direction
March 2006
Goforth and
Christoforou (2006)
Lake Hartwell, GA
(rural southeast U.S.)
February-March 2003
Makkonen et al. (2010)
Virolahti EMEP station,
Finland
European Route E18 (3,000 vehicles/day) 5 km
August 2007
Wojas and Almquist
(2007)
Oxford, OH and other towns in Transportation, manufacturing processes,
Greater Cincinnati region and coal-fired power plants
-80 km to northwest
Cincinnati
January to December
2005
Lin et al. (2005)
Roadside in city in southern
Taiwan
Traffic emissions (avg traffic
load = 72,000 vehicles/day), Pingtung
Industrial Park (146 factories), Nearby
crematory
10m from road, 2 km from
industrial park (146 factories,
[i.e., electron apparatus,
metal, and food
manufacturing]), 1 km from
crematory
February to April 2004
Csavina et al. (2011)
Main sampling site in
Winkelman, Arizona at Hayden
High School
Nearby an active mining and smelting site
in Hayden, Arizona
2km from mine tailings pile;
1km from smelting operations,
main smoke stack, and slag
pile
December 2008-
November2009
2-216
-------
Table 2-27 (Continued): Metadata for studies of Pb-PM size distribution.
Reference
Location
Nearest source
Proximity to source
Sampling dates
3 Sites in central Taiwan:
A school
(Hung-kuang),
Fang and Huang (2011) A wetland
(Gaomei in Taichung),
An industrial site
(Quan-xing)
Emissions from vehicle traffic and industry
Hung-kuang (HK): in
residential area 2 km from
major expressway
Gaomei in Taichung (GM):
300 hectare wetland with coal
combustion-based Taichung
Thermal Power Plant (located
along the coast of the west
side of the sampling site)
Quan-xing (QX): town with lots
of industry including metal
manufacturing, textiles,
petroleum and coal products
November 2010-
December2010
Lim et al. (2006)
7 Sites around Los Angeles,
CA, including 6 urban
watershed sites and one non-
urban coastal watershed.
Not stated
August, 2002-June,
2003
2-217
-------
Table 2-28 Size distribution data for various studies described in Table 2-27.
Reference Location Size Bin
0.18-0.32
0.32-0.56
0.56-1
1-1.8
1.8-3.2
3.2-5.6
Yi et al. (2006) Jersey City, NJ
V ' y y 5.6-10
10-14.4
14.4-19.9
19.9-26.1
26.1-36.1
36.1-100
0.18-0.32
0.32-0.56
0.56-1
1-1.8
1.8-3.2
Yietal.(2Qfi6) New Brunswick, NJ ^^
(Continued) 5.6-10
10-14.4
14.4-19.9
19.9-26.1
26.1-36.1
36.1-100
Concentration
0.001054
0.000668
0.000952
0.000852
0.000609
0.001229
0.001591
0.000948
0.000171
0.000693
0.000333
0.000447
0.001146
0.00078
0.001733
0.001083
0.000373
0.000446
0.000347
0.000182
2.02E-09
1.35E-05
1.62E-05
0.000152
2-218
-------
Table 2-28 (Continued): Size distribution data for various studies described in Table 3-27.
Reference Location Size Bin
1-1.8
1.8-3.2
Bein et al. (2006) Pittsburgh, PA
3.2-5.6
5.6-10
<0.1 pm
0.1-0.35
Downey, CA 0.35-1.0
1.0-2.5
2.5-10
<0.1 pm
0.1-0.35
Riverside, CA 0.35-1.0
1.0-2.5
2.5-10
0.1-0.196
0.196-0.356
0.356-0.57
0.57-1
Dall'Osto et al. (2008) Port Talbot, U.K. 1-1.8
1.8-3.1
3.1-6.2
6.2-9.9
9.9-18
<6
6-11
„ , . . . ,„„„„,, Los Angeles, CA, 10 m downwind of „„ „_
Sabm et al. (2006b) H 11-20
20-29
>29
Concentration
0.096608
0.314846
0.187393
0.239094
0.00133
0.00419
0.00334
0.00189
0.00175
0.0004
0.00089
0.0018
0.001
0.00297
0.000211
0.001871
0.005424
0.004935
0.010229
0.002216
0.001847
0.000807
0.000141
0.007953
0.004172
0.00013
0.000522
0.004563
2-219
-------
Table 2-28 (Continued): Size distribution data for various studies described in Table 3-27.
Reference Location Size Bin
Frankfurt, Germany <0.43
0.43-0.63
0.63-1.1
1.1-2.1
Zereini et al. (2005) 2.1-3.3
Main street; Frankfurt
3.3-4.7
4.7-5.8
5.8-9.0
>9.0
<0.43
0.43-0.63
0.63-1.1
1.1-2.1
z.ereini et ai. t^iuuo) ^-_i ±__ ± i— ._ . \_f — L i-* * /*, /*,
,„ ... ,, Side street, Frankfurt 2.1-3.3
(Continued)
3.3-4.7
4.7-5.8
5.8-9.0
>9.0
<0.43
0.43-0.63
0.63-1.1
1.1-2.1
Zereini et al. (2005) r-, , ,_ , ji-ir_.. o-ioo
,„ .. ,, v ' Rura background, Frankfurt 2.1-3.3
(Continued) a
3.3-4.7
4.7-5.8
5.8-9.0
>9.0
Concentration
0.005904
0.005332
0.004285
0.002857
0.002666
0.002857
0.001809
0.002476
0.004285
0.00332
0.002818
0.002239
0.001544
0.000849
0.000772
0.000386
0.00054
0.000733
0.003312
0.002442
0.002208
0.001405
0.000602
0.000502
0.000201
0.000201
0.000502
2-220
-------
Table 2-28 (Continued): Size distribution data for various studies described in Table 3-27.
Reference Location
Hays et al. (2011) Raleigh, NC
Chen et al. (201 Ob) Taipei, Taiwan tunnel
Dresden, Germany
Bruggemann et al. (2009)
Summer; Dresdeb
Bruggemann et al. (2009) .... , „ ,
(Continued) Winter; Dresden
Size Bin
0.03-0.06
0.06-0.108
0.108-0.17
0.17-0.26
0.26-0.4
0.4-0.65
0.65-1
1-1.6
1.6-2.5
2.5-4.4
4.4-6.8
6.8-10
10-18
<0.1
0.1-2.5
2.5-10
<0.1
0.1-2.5
2.5-10
0.05-0.14
0.14-0.42
0.42-1.2
1.2-3.5
3.5-10
0.05-0.14
0.14-0.42
0.42-1.2
1.2-3.5
3.5-10
Concentration
0.000186
0.000395
0.000732
0.001486
0.003593
0.007315
0.00423
0.002719
0.002701
0.003346
0.00123
0.000934
0.001265
0.018409
0.019773
0.030682
0.00125
0.020625
0.024375
0.001078
0.002874
0.004671
0.001617
0.000539
0.002335
0.00521
0.013293
0.003054
0.000539
2-221
-------
Table 2-28 (Continued): Size distribution data for various studies described in Table 3-27.
Reference Location
Harrison et al. (2003) Birmingham, U.K.
Wang et al. (2006d) Kanazawa, Japan
Cincinnati, OH
Martuzevicius et al. (2004)
Cycle VIII; Cincinnati
Martuzevicius et al. (2004) „ , IX, „. . ...
(Continued) Cycle IX; Cincinnatl
Size Bin
<0.2
0.2-1
1-2
2-10
>10
0.1-0.43
0.43-0.65
0.65-1.1
1.1-2.1
2.1-3.3
3.3-4.7
4.7-7
7-11
11-18
~0.1-0.18
0.18-0.32
0.32-0.56
0.56-1
1-1.8
1.8-3.2
3.2-5.6
5.6-10
~0.1-0.18
0.18-0.32
0.32-0.56
0.56-1
1-1.8
1.8-3.2
3.2-5.6
Concentration
0.00685
0.014923
0.002446
0.00318
0.000489
0.000792
0.000748
0.00118
0.00103
0.000393
0.000678
0.000375
0.000229
0.000125
0.000758
0.002045
0.003258
0.00447
0.005758
0.00697
0.008333
0.009545
0.000455
0.001591
0.002879
0.004091
0.005379
0.006667
0.007879
2-222
-------
Table 2-28 (Continued): Size distribution data for various studies described in Table 3-27.
Reference
Lim et al. (2006)
Note: Summer and Fall data not
provided because there is more
uncertainty in the sites for those
datasets.
Lim et al. (2006)
(Continued)
Note: Summer and Fall data not
provided because there is more
uncertainty in the sites for those
datasets.
Lim et al. (2006)
(Continued)
Note: Summer and Fall data not
provided because there is more
uncertainty in the sites for those
datasets.
Location
Los Angeles, CA
Los Angeles River Watershed #1
Winter
Spring
Los Angeles River Watershed #2
Winter
Spring
Los Angeles River Watershed #3
Winter
Spring
Size Bin
5.6-10
6-11
11-20
20-29
>29
6-11
11-20
20-29
>29
6-11
11-20
20-29
>29
6-11
11-20
20-29
>29
6-11
11-20
20-29
>29
6-11
11-20
20-29
>29
Concentration
0.009091
1.315
0.743
0.821
2.302
1.212
1.485
3.025
1.251
0.547
1.212
1.042
1.212
0.312
0.782
1.116
1.055
0.547
1.016
2.299
1.055
0.508
1.524
2-223
-------
Table 2-28 (Continued): Size distribution data for various studies described in Table 3-27.
Reference Location
Santa Ana River Watershed
Lim et al. (2006)
(Continued) Winter
Note: Summer and Fall data not
provided because there is more
uncertainty in the sites for those
datasets. sPrin9
Ballona Creek Watershed
Lim et al. (2006)
(Continued) Winter
Note: Summer and Fall data not
provided because there is more
uncertainty in the sites for those
datasets. sPrin9
Dominguez Creek Watershed
Lim et al. (2006)
(Continued) Winter
Note: Summer and Fall data not
provided because there is more
uncertainty in the sites for those
datasets. sPrin9
Size Bin
6-11
11-20
20-29
>29
6-11
11-20
20-29
>29
6-11
11-20
20-29
>29
6-11
11-20
20-29
>29
6-11
11-20
20-29
>29
6-11
11-20
20-29
>29
Concentration
-
0.097
0.235
0.195
0.185
0.235
0.039
0.156
1.263
1.016
0.313
0.664
5.064
1.29
0.312
2.58
2.315
1.368
0.547
0.625
0.683
0.469
0.078
0.508
2-224
-------
Table 2-28 (Continued): Size distribution data for various studies described in Table 3-27.
Reference Location Size Bin Concentration
Malibu Creek (non-urban) 6-11 0.201
Lim et al. (2006) 11'20 °-391
(Continued) Winter 20-29
>29 0.039
Note: Summer and Fall data not
uncertainty in the sites for those 11-20 0.156
datasets. Spring 20~29~
>29 0.117
2-225
-------
2.8.5
Pb Concentration in a Multipollutant Context
CO
PM1C
NO2
PM2.5
S02
03
o
0 OC
0
0 00 C
o c
CD o oco m
OCD CMXBD
i OOOQBD9D GOO O' O
(HDD* CMP 000
fjnri (DQiQQQOKD
K3OBOOQO
MOO Q OO O
-1
-0.5
0.5
Note: Correlations were calculated from available data when data were above MDL and there were at least 30 data pairs available
for comparison.
Correlations for individual sites are shown with black open circles, while median correlations are illustrated with a red square.
Figure 2-62 Spearman correlations of monitored non-source Pb-TSP
concentration with daily averages of copollutant concentrations,
2008-2010.
2-226
-------
Source, SO2
Non-Source, SO2
Source, PM2s
Non-Source, PM25
Source, PMio
Non-Source, PMio
Source, O3
Non-Source, O3
Source, NO2
Non-Source, NO2
Source, CO
Non-Source, CO
Source, SO2
Non-Source, SO2
Source, PM2s
Non-Source, PM25
Source, PMio
Non-Source, PM10
Source, O3
Non-Source, O3
Source, NO2
Non-Source, NO2
Source, CO
Non-Source, CO
US Summer
o oo
OCO QOBDXD ODOO O
O O ODO
COO O03QBD
00)
(3D O O o ooo
00
O OCCQEOBKCnxiD aUDODOCDO O
-1.0
-0.5 0.0 0.5
Spearman Correlation Coefficient
1.0
US Fall oo o
O O O GOOD 0(0)0 00 O 00 O 00 O OO
OOO O
O O O O (d^QE) ^^KDOQDOXMDGHCO OO CD )
o o
OOO OOD O O OBDOODODO
00 00
O OOGDQE> GOOD OOO QDOO O (ID ODDOQO 00 O '
OO
O O OQDOOOOIHDGDOOO CDdBKOD (3D
O O
O O OOOCDDSD O O dBDSDd) 00 (3D
1.0
-0.5 0.0 0.5
Spearman Correlation Coefficient
1.0
Note: Top panel: Summer; Bottom panel: Fall.
Figure 2-63
Seasonal correlations of monitored Pb-TSP concentration with
copollutant concentrations, 2007-2008.
2-227
-------
Non-Source, SO2 ' "
Source, PM2 5
Non-Source, PM2s
Non-Source, PM,0
Non-Source, O3
Non-Source, N02 • -
Non-Source, CO . -
US Winter
o
0 0
000 0
0
o o o o o
O 000 O (
0
0 0
0 0 00 0
-1.0
-0.5 0.0 0.5
Spearman Correlation Coefficient
1.0
Non-Source, S02 -
Source, PM2.s -
Non-Source, PM25 -
Non-Source, PMio
Non-Source, 03 .
Non-Source, NO2 .
Non-Source, CO .
US Spring
o o
QD O
-1.0
Note: Top panel: Winter; Bottom panel: Spring.
oo o
o oo
QD
-0.5 0.0 0.5
Spearman Correlation Coefficient
1.0
Figure 2-64 Seasonal correlations of monitored Pb-TSP concentration with
copollutant concentrations, 2009.
2-228
-------
Non-Source, SO2
Source, PM25
Non-Source, PM25
Non-Source, PMi0
Non-Source, O3
Non-Source, NO2
Non-Source, CO
US Summer
o o o
o
o o o
o
o ODD o
o
o o
00 O
o o
o
-1.0
-0.5 0.0 0.5
Spearman Correlation Coefficient
1.0
Non-Source, SO2
Source, PM25
Non-Source, PM25
Non-Source, PM10
Non-Source, O3
Non-Source, NO2
Non-Source, CO
US Fall
O 00
oo o o
o o o
o oo o o
o o
-1.0
Note: Top panel: Summer; Bottom panel: Fall.
Figure 2-65
-0.5 0.0 0.5
Spearman Correlation Coefficient
1.0
Seasonal correlations of monitored Pb-TSP concentration with
copollutant concentrations, 2009.
2-229
-------
Zn
K
Br
Cu
Fe
H03-
S
oc
EC
Crystal
Mi
Se
S04-2
NH4+
K+
Ca
Si
Ti
V
Cr
Ni
nvol H03-
Cl
Ha
Hg
vol H03-
Cd
Na+2
a
o
D
O
h
1
Q
0 | | , | 1
0 0 1 1 | | 1
o o oo o
-------
Zn -
Br —
Cu —
K -
Ca -
Nil -
Crystal —
Se -
H03- -
Ti -
K -
Mg
K+ -
NH4+ —
Hg
vol H03- -
Hi -
Cl -
As -
nvol H03- -
0 | 1 |
O | 1 |
i n |
o o i 1 I
o o o i i |
o o i H |
o o OD cm 1 |
O 1 ] |
o o i 1 i
0 1- j | =1
0 0 1 H 1 h
o 01 1 I 1- —
O O | 1 | |
OO 1 1 | |
~*- — J -*"
1 j 1
H III
J j *"
h - 1 [
o i i | |
i 1 | |
h H
1 1
^ H
| 1 O
~~l -1
f 1
1 1
3:::::::::^
1 o
1
J- 1
1 Q
1
1 0
1 O O
H
H
1 (DO
10
Zn -
K -
Cu
Br —
oc -
EC -
Fe -
vol H03- -
Mtl —
Ca -
Se —
Cr —
V —
nvol H03- -
Na —
Na+2 —
As -
Hi —
Hg -
Cd —
O O OO O 1 ]_ | f 1
oooi 1 | h 1
01- 111 H
o o o i i | | 1
O O OD O 1 -| | | 1
o o ooo 1 [ | 1 1 o
O OOOI 1 | | 1
i-Ol
° 1- \ 1 1 ~1 ~" H
O OD 1 1 | | 1 O
1 — • 1
rrt- J==T===l
' J '
" " I j I
"1 B . —
O OOI 1 | | ID
O 1 1 | | ICE O
dj | 1 | | 1 o
0 | 1 | \ 1 o
i- 1 i r ^
1 [ | | IQDO
O O 1 1 | | IO O
Top panel: Summer; Bottom panel: Fall.
Note: "nvol" = non-volatile, "vol" = volatile, and organic carbon (OC) samples were blank-adjusted.
Figure 2-67 Seasonal correlations of monitored Pb-PM2.s concentration with
copollutant concentrations, 2007-2009.
2-231
-------
Table 2-29 Copollutant
exposures
Adgate et al. (2007)
Location
PM2.5
Pb
S
Ca
Al
Na
Fe
Mg
K
Ti
Zn
Cu
Ni
Mn
Sb
Cd
V
La
Cs
Th
I-R (med)3'b
Personal
(median)0
Minnesota, U.S.
1.5
272.1
85.0
23.3
20.6
43.1
16.3
38.4
0.8
6.5
1.-0.15
2.4
0.21
0.12
0.12
0.05
0.00
0.00
0.00
3.2
351.6
174.1
58.6
31.9
78.6
27.5
47.5
1.4
9.6
4.9
1.8
2.3
0.30
0.14
0.16
0.11
0.00
0.00
for various trace metal studies.
Riediker et al. (2003) Pekey et al. (2010)
Vehicle
(range)0
Roadside
(range)0
New Jersey, U.S.
24,000
2-3
905-1,592
31-44
307-332
6-75
9-10
5-10
18-32
0
3-4
4-6
1
31,579
4-6
1,416-2,231
18-40
82-163
23-57
6-10
14-17
8-16
0
3
4-7
1
I -near industry
(range)3
Kocaeli, Turkey
24,400-29,800
34-85
435-489
309-452
53-60
44-58
160-215
29-39
51-88
21-58
2-3
28-32
3-5
Molnaretal. (2007)
I-R (median)3'13
l-School
(median)3
l-Pre-School
(median)3
Stockholm, Sweden
2.8
330
70
57
120
8.0
14
9.3
0.99
2.2
2.5
2.5
290
110
100
96
13
17
1.7
1.0
2.5
2.7
1.7
220
58
71
67
8.7
11
2.1
0.72
2.1
1.8
2-232
-------
Table 2-29 (Continued): Copollutant exposures for various trace metal studies.
Adgate et al. (2007)
Riedikeret al.
(2003) Pekey et al. (2010) Molnar et al. (2007)
Sc 0.00 0.01
Ag 0.07 0.08
Co 0.02 0.07
Cr 1.2 2.6
Si
Cl
Se
Rb
Sr
As
2
198-464
7-32
1
1
5-28
1
1 3-8 <1.1 1.3 1.1
338-672 387-401
3-9
1-2
1
1
1 1-2
Mo
Br
2.1 1.3 1.3
al: Indoor; Units: ng/m3
bR: Residential; Units: ng/m3
cUnits: ng/m3
2-233
-------
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CHAPTER 3 EXPOSURE, TOXICOKINETICS, AND
BIOMARKERS
3.1 Exposure Assessment
The purpose of this section is to present recent studies that provide insight about human
exposure to Pb through various pathways. Pb is considered to be a multimedia
contaminant with multiple pathways of exposure. The relative importance of various
media in affecting Pb exposure changes with source strength and location, location and
time activity of the exposed individuals, behavior of the exposed individuals, and risk
factors such as age and socioeconomic factors (risk factors are discussed in detail in
Chapter 3). Blood Pb and bone Pb biomarkers (discussed in Section 3.4 and Section 3.5).
are often used to indicate composite Pb exposure resulting from multiple media and
pathways of exposure.
The recent information provided here builds upon the conclusions of the 2006 Pb AQCD
(U.S. EPA. 2006b). which found that air Pb concentrations and blood Pb levels have
decreased substantially following the restrictions on Pb in on-road vehicle gasoline, Pb in
household paints, the use of Pb solder, and reductions in industrial Pb emissions that have
occurred since the late 1970s. Nevertheless, detectable quantities of Pb have still been
observed to be bioaccessible in various media types. It was reported in the
2006 Pb AQCD (U.S. EPA. 2006b) that airborne maximum quarterly Pb concentrations
in the U.S. were in the range of 0.03-0.05 ug/m3 for non-source-oriented monitors for the
years 2000-2004 and were 0.10-0.22 ug/m3 for source-oriented monitors during that time
period, while blood Pb levels reached a median of 1.70 ug/dL among children (1-5 years
of age) in 2001-2002. It was also observed that Pb exposures were associated with nearby
industrial Pb sources, presence of Pb-based paint, and Pb deposited onto food in several
of the studies described in the 2006 Pb AQCD.
3.1.1 Pathways for Pb Exposure
Pathways of Pb exposure are difficult to disentangle because Pb has multiple sources in
the environment and passes through various environmental media. These issues are
described in detail in Sections 2.2 and 2.3. Air-related pathways of Pb exposure are the
focus of this ISA. Pb can be emitted to air, soil, or water and then cycle through any or all
of these media. In addition to primary emission of particle-bound or gaseous Pb to the
atmosphere, Pb can be resuspended to the air from soil or dust. Additionally, Pb-bearing
PM can be deposited from the air to soil or water through wet and dry deposition. Air-
3-1
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related Pb exposures also include inhalation and ingestion of Pb-contaminated food,
water or other materials following atmospheric deposition of Pb; these exposures include
dust and soil via hand-to-mouth contact. In general, air-related pathways include those
pathways where Pb passes through ambient air on its path from a source to human
exposure. Some non-air-related exposures of Pb include ingestion of indoor Pb paint, Pb
in diet as a result of inadvertent additions during food processing, and Pb in drinking
water attributable to Pb in distribution systems, as well as other generally less prevalent
pathways.
The complicated nature of Pb exposure is illustrated in Figure 3-1. in which the Venn
diagram depicts how Pb can cycle through multiple environmental media prior to human
exposure. The "air/soil/water" arrows illustrate Pb exposures to plants, animals, and/or
humans via contact with Pb-containing media. The exposures are air-related if Pb passed
through the air compartment. When animals consume plant material or water exposed to
Pb that has at some point passed through the air compartment, and when human diet
includes animals, plants or drinking water exposed to Pb that has passed through the air
compartment, these are also considered air-related Pb exposures. As a result of the
multitude of possible air-related exposure scenarios and the related difficulty of
constructing Pb exposure histories, most studies of Pb exposure through air, water, and
soil can be informative to this review. Figure 3-1 also illustrates other exposures, such as
occupational exposures, contact with consumer goods in which Pb has been used, or
ingestion of Pb in drinking water conveyed through Pb pipes. Most Pb biomarker studies
do not indicate speciation or isotopic signature, and so exposures that are not related to
Pb in ambient air are also reviewed in this section because they can contribute to Pb body
burden. Many of the studies presented in the subsequent material focus on observations
of Pb exposure via one medium: air, water, soil and dust, diet, or occupation.
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Newly Emitted Pb
Historically Emitted Pb
OOTDOORSOIL
NDDUST
NATURAL WATERS
AND SEDIMENTS
Non-air Pb
eleases
•— --.
AIR
SOIL
WATER
1 — - -
**• —
PLANT
EXPOSURE
-+
^^.
AIR
SOIL
WATER
--. — — -
**• — — -— ,
r ANIMAL
t EXPOSURE]]"
AIR
SOIL
WATER
•——_—--
.-•— •
HUMAN
^EXPOSURE
( COSMETICS )
OYS et
Note: The Venn diagram is used to illustrate the passage of Pb through multiple environmental media compartments through which
exposure can occur.
Figure 3-1 Conceptual model of multimedia Pb exposure.
The relative importance of different sources or pathways of potential exposure to Pb in
the environment is often difficult to discern. Individual factors such as home
environment, location, and risk factors (described in more detail in Chapter 5 and Chapter
6) may influence exposures. The National Human Exposure Assessment Survey
(NHEXAS) study sampled Pb, as well as other pollutants and VOCs, in multiple
exposure media from subjects across six states in EPA Region 5 (Illinois, Indiana,
Michigan, Minnesota, Ohio, and Wisconsin) (Clayton et al.. 1999) as well as in Arizona
(O'Rourke et al.. 1999) and Maryland (Egeghy et al.. 2005). Results from NHEXAS
indicate that personal exposure concentrations of Pb are higher than indoor or outdoor
concentrations of Pb, perhaps suggesting a personal cloud effect; see Table 3-1. Pb levels
in windowsill dust were higher than Pb levels in surface dust collected from other
surfaces. Clayton et al. (1999) suggested that higher windowsill levels could be attributed
to the presence of Pb-based paint and/or to accumulation of infiltrated outdoor Pb-bearing
PM. Pb levels in food were higher than in beverages, and Pb levels in standing tap water
(also referred to as "first flush" or "first draw") were higher than Pb levels obtained after
allowing water to run for three minutes to flush out pipes. Layton and Beamer (2009)
estimated that 34-66% of Pb in floor dust was tracked in from outdoors and originated as
3-3
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ambient air Pb, based on 1992 levels in Sacramento; in 1992, phase-out of Pb usage in
gasoline was near complete, but industrial emissions were still higher than current levels;
see Section 2.2.
Table 3-1 Estimates of Pb measurements for EPA Region 5 from the
NHEXAS study.
Medium3
Personal air
(ng/m3)d
Indoor air
(ng/m3)d
Outdoor air
(ng/m3)d
Surface dust
(ng/cm2)
Surface dust
(mg/kg)
Window sill dust
(ng/cm2)
Window sill dust
(mg/kg)
Standing tap water
(H9/L)
Flushed tap water
(pg/L)
Solid food
(Mg/kg)
N
167
213
87
245
244
239
239
444
443
159
Percentage above
LODb (CLs)c
81.6
(71.3; 92.0)
49.8
(37.2; 62.3)
73.8
(56.3; 91.3)
92.1
(87.4; 96.8)
92.1
(87.4; 96.8)
95.8
(92.5; 99.0)
95.8
(92.5; 99.0)
98.8
(97.6; 100.0)
78.7
(70.7; 86.7)
100.0
(100.0; 100.0)
Mean (CLs)c
26.83
(17.60; 36.06)
14.37
(8.76; 19.98)
11.32
(8.16; 14.47)
514.43
(-336.6; 1365.5)
463.09
(188.15; 738.04)
1,822.6
(481.49;
3,163.6)
954.07
(506.70;
1,401.4)
3.92
(3.06; 4.79)
0.84
(0.60; 1.07)
10.47
(6.87; 14.07)
50th (CLs)c
13.01
(11.13; 18.13)
6.61
(4.99; 8.15)
8.50
(7.14; 10.35)
5.96
(3.37; 10.94)
120.12
(83.85; 160.59)
16.76
(10.44; 39.41)
191.43
(140.48;
256.65)
1.92
(1.49; 2.74)
0.33
(0.23; 0.49)
6.88
(6.44; 8.04)
90th (CLs)c
57.20
(31.18; 85.10)
18.50
(12.69; 30.31)
20.36
(12.60; 34.91)
84.23
(26.52; 442.63)
698.92
(411.84; 1,062.8)
439.73
(106.34; 4,436.2)
1,842.8
(1,151.3;
2,782.5)
9.34
(7.87; 12.35)
1.85
(1.21; 3.04)
14.88
(10.78; 19.08)
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Table 3-1 (Continued): Estimates of Pb measurements for EPA Region 5 from the
NHEXAS study.
Medium3
Beverages
(pg/kg)
Food + Beverages
(Hg/kg)
Food intake
(Hg/day)
Beverage intake
(pg/day)
Food + Beverage
intake (pg/day)
Blood
(Hg/dL)
N
160
156
159
160
156
165
Percentage above
LODb (CLs)c
91.5
(85.2; 97.8)
100.0
(100.0; 100.0)
100.0
(100.0; 100.0)
91.5
(85.2; 97.8)
100.0
(100.0; 100.0)
94.2
(88.2; 100.0)
Mean (CLs)c
1.42
(1.13; 1.72)
4.48
(2.94; 6.02)
7.96
(4.25; 11.68)
2.15
(1.66; 2.64)
10.20
(6.52; 13.89)
2.18
(1.78; 2.58)
50th (CLs)c
0.99
(0.84; 1.21)
3.10
(2.66; 3.52)
4.56
(3.68; 5.36)
1.41
(1.18; 1.60)
6.40
(5.21; 7.78)
1.61
(1.41; 2.17)
90th (CLs)c
2.47
(2.06; 3.59)
6.37
(4.89; 8.00)
12.61
(9.27; 16.38)
4.45
(3.15; 5.65)
16.05
(13.31; 18.85)
4.05
(3.24; 5.18)
Note: EPA Region 5 includes six states: Illinois, Indiana, Ohio, Michigan, Minnesota, and Wisconsin. Participants were enrolled
using a stratified, four-stage probability sampling design, and submitted questionnaire and physical measurements data. Summary
statistics (percentage above limit of detection (LOD), mean, median, 90th percentile) were computed using weighted sample data
analysis. The estimates apply to the larger Region 5 target population (all non-institutionalized residents residing in households).
Estimates for indoor air, outdoor air, dust media, and water media apply to the target population of Region 5 households; estimates
for other media apply to the target population of Region 5 residents.
""Percentage of the target population of residents (or households) estimated to have Pb levels above limit of detection (LOD).
°The lower and upper bounds of the 95% confidence limits (CL) are provided.
dPM50.
Source: Reprinted with permission of Nature Publishing Group, Clayton et al. (1999)
3.1.1.1 Particle Size Distributions for Airborne-Pb, Dust-Pb, and
Soil-Pb
The size distribution of ingestible dust particles differs from the size distribution of
inhalable ambient air Pb particles and therefore cannot be directly compared. The
inhalability of airborne PM is a gradually decreasing function of particle size. Inhalability
criteria established from experimental data, obtained at wind speeds of 1-8
meters/second, describe PM inhalability of 77% for particles <10 jam (dae, aerodynamic
diameter). Inhalability of particles ranging in size from 40 to 100 (im dae is 50%; above
100 (im, inhalability data are lacking (Soderholm. 1989; ACGIH. 1985). The particles
that are not inhaled may settle to surfaces, making them available for subsequent
ingestion. The size distribution of soil and house dust particles tends to be much larger
than airborne PM. Que Hee et al. (1985) and U.S. EPA (1990^) observed that 50% or
more of the mass of house dust tends to be comprised of particles smaller than 150 nm.
Gulson et al. (1995b) observed that the mode of the Pb house dust size distribution was in
the 38-53 urn range; they did not report the overall house dust size distribution. Given the
house dust Pb size distributions documented, dust Pb brought into homes with foot traffic
3-5
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may be aerosolized but is likely to stay airborne for only a few seconds, since particles
larger than PM25 tend to settle from the air quickly; see Section 2.3.1.3 and Section
3.1.3.1. Siciliano et al. (2009) observed different size distributions for different types of
soils: agricultural sites had median soil Pb of 34 pirn, and brownfields had median soil Pb
of 105 nm. These observations of larger particle sizes for soil and dust Pb support the
notion that exposure to Pb in dusts and soils would occur by ingestion rather than
inhalation following resuspension.
The main pathway for Pb ingestion by children is by hand to mouth contact (Lanphear et
al.. 1998). In a playground environment in London, U.K., Duggan et al. (1985) reported
that hand to mouth transfer was effectively limited to particles smaller than 10 um, even
when the soil itself exhibited a much larger particle size distribution. More recently,
Yamamoto et al. (2006) reported for a cohort of children in Kanagawa Prefecture, Japan
(greater Tokyo area) that the mode of size distributions of particles adhering to children's
hands was 39 ± 26 (im, with the upper tail ranging from 200-300 (im. Kissel et al. (1996)
measured three size fractions of soil adhered to a hand via a hand press: < 150 um,
150-250 pirn, and > 250 pirn and observed that, when soil was dry (<2% moisture
content), 43%-69% of the soil was in the smallest fraction. When the moisture content
was higher than 2%, 28-81% of the adhered soil was in larger than 250 pirn. Percentage
and mass adhered per area (mg/cm2) depended on soil type, with wet sand and loamy
sand adhering more to hands than sandy loam or silt loam. For dry soil, silt loam mass
produced the largest adherence in terms of mass per area. Differences among the size
distribution results may be related to differences in the soil type, soil moisture levels
between the locations, and/or to differences between the analytical methods used to
measure size distribution; Duggan et al. (1985) used optical microscopy of the dust
wipes, while Yamamoto et al. (2006) used a laser scattering device measuring sampled
particles suspended in an aqueous solution. Siciliano et al. (2009) regressed adhered
average soil size on hands bulk soil particle size and found a log-log relationship with
(3 = 0.66 using both brownfield and agricultural soils; the proportion of soil adhered
depended on organic content.
Several studies have found that Pb is enriched in the smaller particle sizes of the soil and
house dust. Davies and White (1981) observed that enrichment decreased linearly with
increasing dust size bin, with dust particles smaller than 64 urn having a Pb concentration
of 76.1 mg/kg and particles in the 1,000-2,000 urn size range having a Pb concentration
of 16.4 mg/kg. Sheets and Bergquist (1999) also found that Pb content decreased with
increasing particle size. More recently, Ljung et al. (2006) investigated childhood
exposures to trace metals on playgrounds in Uppsala, Sweden and observed that the Pb
content in soil in the <50 urn size fraction was 1.5 times higher than that in the >4 mm or
3-6
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50-100 pirn size fractions. Sheppard et al. (1995) measured enrichment in different types
of soils (sand and clay) and found that enrichment was substantially higher in the sand.
Studies focusing on particle size distributions of house dust adhered to the hands are
lacking. Ingestion of house dust has been reported to be the major source of Pb intake
during early childhood (Lanphear et al., 2002). If a similar particle size distribution holds
for household dust, then ingestion of indoor Pb of atmospheric origin could also be
strongly dependent on dust particle size. Therefore, larger particles of atmospheric origin,
which may not be considered relevant for exposure by inhalation exposure, are still
relevant for Pb exposure by ingestion. However, no studies in the literature have
presented information on the relative contributions of Pb from different PM size fractions
to blood Pb concentrations.
It should be noted that different measurement techniques are used for different
environmental media. For example, ambient air Pb-PM size distribution is measured by
one of the non-FRM instruments such as a MOUDI, described in Section 2.4. and its
measurement is subject to errors specific to the technique. Dust and soil size distribution
are typically measured with graduated sieves, and errors associated with these methods
occur more often in the smaller size fractions that are subject to agglomeration and
clogging if the particle shape is nonspherical.
3.1.1.2 Estimating Pb Exposure in the Integrated Exposure
Uptake Biokinetic (IEUBK) Model
Several studies have used a combination of measured values and default model values to
represent exposures and determine their relative contributions to blood Pb. For example,
Cornelis et al. (2006) used the Integrated Exposure Uptake Biokinetic model (IEUBK),
described in detail in the 2006 Pb AQCD (U.S. EPA. 2006b) to model children's
exposures to Pb emissions from a non-ferrous smelter in Hoboken, Belgium. In deriving
the model input (annual averages) for ambient air Pb concentration, as well as soil and
dust, they employed weighting based on children's time spent in different locations in the
study area and air, soil and indoor dust measurements in those areas. In their results for
the area of the smelter, the ingestion of dust and soil pathways accounted for more than
70% of the exposure, while the inhalation pathway accounted for less than 2%. Similarly,
Carrizales et al. (2006) analyzed exposures to children living near a copper (Cu) smelter
in San Luis Potosi, Mexico. They employed the IEUBK default options for assignment of
Pb dust concentration as 70% of the soil Pb concentration, while air Pb concentration was
assigned based on measurements by the Mexican government. Based on these
assumptions, they attributed 87% of blood Pb to soil and dust exposure. These studies did
3-7
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not estimate the air Pb contribution to the soil/dust Pb concentrations and consequently
did not estimate the portion of the ingestion pathway that derives from ambient air Pb.
Appendix I of the 2007 Pb Risk Assessment (U.S. EPA. 2007f) provides estimates of the
contribution of various pathways to the blood Pb of children simulated in several case
studies. Simulations provided estimates of contributions from outdoor ambient air Pb by
inhalation and by ingestion of indoor dust, including the fraction of indoor dust Pb
associated with recent penetration of ambient air Pb into the residence. Although ambient
air Pb may also contribute to Pb ingestion through other pathways (i.e., diet, soil), data
and tools to support a simulation of the linkage between air Pb concentrations and
concentrations in other media were limited. Accordingly, Pb concentrations pertaining to
other pathways (e.g., diet, outdoor soil, the component of indoor dust Pb other than that
derived from Pb recently in ambient air) were held constant across the different air
quality scenarios simulated. Table 3-2 provides estimates for the General Urban Case
Study in the 2007 Pb Risk Assessment (U.S. EPA. 2007f). The General Urban Case
Study, unlike the various location-specific case studies, was not based on any specific
urban location and reflected several simplifying assumptions including uniform ambient
air Pb levels across the simulated, hypothetical study area and a uniform study
population.
-------
Table 3-2 Predicted concurrent blood Pb levels and source contributions for
children in their seventh year of life.
Air Pba
(ug/m3)
0.05
0.149
0.2
0.87 h
Median
Blood Pbb
(ug/dL)
1.7(5.7)'
1 .9 (6.5)
2.0 (6.9)
2.1 (7.2)
Pathway Contribution (%)
Ingestion
Diet0
32
28
26
25
Outdoor
Soil/Dust
44
38
36
33
Indoor Dust
Otherd
11
6
5
4
Pb recently in
aire
12.6
28.3
32.7
37.2
Inhalation
Ambient
Air
0.1
0.5
0.7
0.9
"Concentrations are maximum calendar quarter averages of Pb in air with exception of 0.05 ug/m which is a maximum monthly
average
bAverage of blood Pb concentrations at 75 and 81 months, assuming exposure concentrations were constant through 7 years of life
""Includes food and drinking water, the Pb concentrations for which were held constant across the different air quality scenarios, as
were Pb concentrations for outdoor soil/dust.
dlncludes indoor dust with Pb contributions from sources other than Pb recently in the air (e.g., indoor paint, outdoor soil/dust, and
additional sources including historical air Pb)
Includes contributions associated with outdoor ambient air Pb from ingestion of indoor dust predicted to be associated with outdoor
ambient air Pb levels
Values in parentheses are the 95th percentile blood Pb for a geometric standard deviation of 2.1
9Mean of the maximum quarterly average concentrations of Pb in TSP (for period 2003 to 2005) among monitor locations in urban
areas having more than one million residents
h95th percentile of the maximum quarterly average concentration of Pb in TSP (for period 2003 to 2005) among monitor locations in
urban areas having more than one million residents
Source: Based on General Urban Case Study (Hybrid Dust Model) in Appendix I, 2007 Pb Risk Assessment (U.S. EPA. 2007fl.
3.1.2 Environmental Exposure Assessment Methodologies
A number of monitoring and modeling techniques have been employed for exposure
assessment. These are detailed in either the 2006 Pb AQCD (U.S. EPA. 2006b) or in the
subsequent Risk and Exposure Assessment performed as part of the same NAAQS
review (U.S. EPA. 2007g). Some of these methods are briefly described here to provide a
context for the exposure studies described in Section 3.1.3. Blood Pb sampling is
described in detail in Section 3.3.2.
Data collection to assess Pb exposure pathways may involve air, soil, and dust samples.
Methods used for digesting air Pb samples are described in Section 2.4. as are ambient air
Pb monitoring techniques. Factors affecting collection of ambient air Pb samples are
described in detail in Section 2.4. For the monitors in the FRM network, the primary role
is compliance assessment. Accordingly, this network includes monitors in locations near
sources of air Pb emissions which are expected to or have been shown to contribute to
ambient air Pb concentrations in excess of the Pb NAAQS. In such locations, Pb may be
3-9
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associated with relatively larger size particles, contributing to air Pb concentration
gradients with distance from the source and greater deposition in the near-source
locations. The FRM network also includes non-source-oriented monitors for which the
main objective is to gather information on neighborhood-scale Pb concentrations that are
typical in urban areas so to better understand ambient air-related Pb exposures for
populations in these areas. This part of the Pb NAAQS network, was required to be
operational as of December 27, 2011. These monitor locations are distributed across a
broad geographic area, representing approximately 63 large urban areas which contain
approximately half of the total U.S. population (based on recently published 2010 Census
Bureau data). In lieu of more detailed analysis of population proximity for these newly
established monitors, population counts were calculated near previously existing
monitors for which data are presented in Section 2.5. For the monitors in that limited
dataset, among the total population of 311,127,619 people in the 2010 Census (ESRI.
2011). 181,100 (0.06%) lived within 1 km of a source-oriented monitor, while 918,351
(0.30%) lived within 1 km of a non-source-oriented monitor.
Dust sampling has not changed drastically since it was first proposed by Sayre et al.
(1974). in which a disposable paper towel was soaked in 20% denatured alcohol and
1:750 benzalkonium chloride and then used to wipe a 1 ft2 sampling area in a systematic
fashion. Que Hee et al. (1985) and Sterling et al. (1999) compared wipe testing with
vacuum methods. Sampling efficiency for the first attempt varied between 53-76% with
vacuum pump flow rate and tube type and was 52% for the wipe method for the Que Hee
et al. (1985) study, with 100% efficiency after five consecutive samples were obtained.
Sterling et al. (1999) observed that two of three vacuuming methods had significantly
higher geometric mean collection (vacuum 1: 94.3 (ig/ft2; vacuum 2: 23.5 (ig/ft2)
compared with dust wipes (5.6 (ig/ft2).
Models may also be used in exposure assessment. For example, two dispersion models,
the American Meteorological Society/Environmental Protection Agency Regulatory
Model (AERMOD), and Industrial Source Complex-Plume Rise Model Enhancements
(ISC-PRIME) were employed to model dispersion of Pb emissions from specific
industrial facilities (Cimorelli et al.. 2005; Perry etal. 2005; EPRI. 1997). and to
estimate ambient air Pb concentrations at some of the case studies included in the 2007
Risk and Exposure Assessment (U.S. EPA. 2007g). These models assume plume
dispersion follows a Gaussian distribution from a point source. For the two point source
case studies included in the 2007 risk assessment, the plume models were used to track
emissions to ambient air near homes located within a few miles of emitting facilities.
However, dispersion models can also be used to track long distance transport of Pb
emissions, as performed by Krell and Roeckner (1988) to model the dispersion and
deposition of Pb and Cd from European nations into the North Sea.
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Several models estimate blood Pb levels resulting from estimated exposure to Pb in
environmental media. These models, which are described in detail in the 2006 Pb AQCD
(U.S. EPA. 2006b) include the IEUBK model, and the EPA All Ages Lead Model
(AALM), which combines and expands the thorough exposure and absorption modules of
the IEUBK model with the comprehensive biokinetic model of Leggett (1993). As of the
writing of this assessment, the AALM is still in development.
The Stochastic Human Exposure and Dose (SHEDS) and NORMTOX models also are
capable of modeling metals exposures through various routes including inhalation,
ingestion, and dermal exposure (Loos M 2010; Burke et al.. 2002). Pb exposure
modeling can also be accomplished using the Modeling Environment for Total Risk
(MENTOR) framework, in which airborne Pb levels could be modeled using AQS,
dispersion modeling, or chemical transport modeling, while human exposure is modeled
with SHEDS or a similar exposure model (Georgopoulos and Lioy. 2006). Additionally,
housing data and time-activity data from the Consolidated Human Activity Database
(CHAD) are incorporated into MENTOR to develop refined estimates of Pb exposure and
tissue burden. However, a literature search did not produce any Pb exposure studies using
the SHEDS, NORMTOX, or MENTOR modeling systems. In general, these models take
input for several environmental Pb exposure media including soil, dust, food and water,
outdoor air, and indoor air. The models are designed to evaluate different exposure
scenarios based on specification of particular conditions.
3.1.3 Exposure Studies
3.1.3.1 Airborne Pb Exposure
Limited personal exposure monitoring data for airborne Pb were available for the
2006 Pb AQCD (U.S. EPA. 2006b). As described above, the NHEXAS study showed
personal air Pb concentrations to be significantly higher than indoor or outdoor air Pb
concentrations (Clayton et al.. 1999). Indoor air Pb concentration was moderately
correlated with floor dust and residential yard soil Pb concentration (Rabinowitz et al..
1985). Egeghy et al. (2005) performed multivariate fixed effects analysis of the
NHEXAS-Maryland data and found that Pb levels measured in indoor air were
significantly associated with log-transformed outdoor air Pb levels, ambient temperature,
number of hours in which windows were open, whether homes were built before 1950,
and frequency of fireplace usage (Table 3-3).
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Table 3-3 Estimates of fixed effects multivariate modeling of Pb levels
measured during the NHEXAS-MD study.
Pb in Indoor Air
Fixed Effect pa p-value
Intercept -0.50 0.0051
Pb in Dust Dermal Pb Blood Pb
pa p-value pa p-value pa p-value
6.22 <0.0001 6.23 <0.0001 0.02 0.91
Outdoor Pb concentration" 0.51 <0.0001
Average weekly temperature (°F) 0.01 0.046
Open window periods (hr) 0.01 0.035
-0.03 0.0082
House pets (yes) -0.15 0.078
Air filter use (yes) -0.28 0.087
Home age (<1 950) 0.25 0.025
Fireplace (frequency of use) 0.11 0.045
Pb concentration in soilb
Interior Pb paint chipping/peeling
(yes)
Cement at primary entryway (yes)
Indoor pesticide usage last 6 mo
(yes)
Electrostatic air filter usage (yes)
Sex of participants (male)
Ethnic minority participants (yes)
Washing hands after lawn mowing
(no)
Gasoline power- equipment usage
(yes)
Bathing or showering activities
(yes)
Dust level indoors (scale: 1-3)
Residing near commercial areas
(yes)
Age of participants (yr)
Number cigarettes smoked (count)
Burning wood or trash (days)
Showering frequency (avg # days)
Work outside home (yes)
Health status (good)
Adherence to high fiber diet (yes)
Gas or charcoal grill usage (yes)
-0.12 0.088
0.96 0.029
0.46 0.0054
0.27 0.037
0.43 0.091
1 .97 0.0064
-0.78 0.0003
-0.91 0.062
0.41 0.0012 0.43 <0.0001
0.41 0.0063
1.04 0.0010
0.61 0.0072
-0.43 0.019
0.22 0.019
0.32 0.0087
0.02 <0.0001
0.03 <0.0001
0.58 0.0099
-0.29 0.0064
-0.26 <0.0001
0.23 0.0009
-0.15 0.040
-0.17 0.0002
"Estimates of fixed effects in final multiple regression analysis models for Pb in the Maryland investigation data in the National
Human Exposure Assessment Survey (NHEXAS-MD).
bLog transform.
Source: Reprinted with permission of Nature Publishing Group, Egeghy et al. (2005).
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Some recent studies have shown that the ratio of indoor to outdoor Pb-PM varies from
site to site depending on factors including infiltration, indoor and outdoor Pb sources, and
meteorology. Adgate et al. (2007) measured the concentrations of several trace elements
in personal, indoor, and outdoor air samples of PM25 and found that average personal
Pb-PM2 5 concentration was roughly three times higher than outdoor air Pb-PM2 5
concentration and two times higher than indoor Pb-PM25 concentration (Table 3-4).
Another study of indoor and outdoor air concentrations of Pb was carried out by Molnar
et al. (2007). PM25 trace element concentrations were determined in homes, preschools
and schools in Stockholm, Sweden. In all sampled locations, Pb-PM2 5 concentrations
were higher in the outdoor environment than in the proximal indoor environment. The
indoor/outdoor ratios for Pb-PM25 suggest an outdoor Pb-PM25 net infiltration of-0.6
for these buildings. Outdoor air Pb concentrations did not differ between the central and
more rural locations. Indoor air Pb concentrations were higher in spring than in winter,
which the authors attributed to greater resuspension of elements that had accumulated in
road dust over the winter period and increased road wear on days with dry surfaces.
Pekey et al. (2010) measured indoor and outdoor trace element composition of PM2 5 and
PMio in Kocaeli, an industrial region of Turkey, and found that average airborne Pb
concentrations were higher outdoors than indoors for both PM2 5 and PMi0 during
summer and for PMi0 during winter, but that indoor Pb concentration was higher than
outdoor Pb concentration for PM2 5 during winter. The indoor-to-outdoor ratio of
airborne Pb varied by environment; it tended to be less than one, but the ratio varied from
one microenvironment to another. In a pilot study in Windsor, Ontario, Rasmussen et al.
(2007) observed that the concentration of Pb in PM2 5 from a personal exposure sample
was roughly 40% higher than the concentration of Pb in outdoor PM2 5 and 150% higher
than Pb in indoor PM2 5. The three studies that included personal samples recorded
measurements that were consistently higher than indoor or outdoor levels, and outdoor
concentrations were higher than indoor concentrations.
Domestic wood burning is a potential source of Pb compounds (Section 2.2.2.5). Alves et
al. (2011) measured trace metals in woodstove and fireplace emissions and found that
PM25 contained Pb, with concentrations from wood burning ranging from 3.3-12.2 (ig/g
and 2.89-30.3 for woodstoves and fireplaces, respectively. When burning briquettes, the
PM2 5 measurements showed Pb enrichment above all other metal elements other than
potassium (woodstove: 1361 (ig/g; fireplace: 616 (ig/g). Molnar et al. (2005) measured
trace element concentration in indoor and personal exposure PM2 5 samples for homes in
which wood is burned and in a reference group where no wood burning occurs in the
home. For both indoor and personal samples, Molnar et al. (2005) observed that Pb
concentrations were higher for the wood burning group and nearly statistically significant
for the personal exposure samples (indoor concentration: 6.0 (ig/m3 versus 4.3 (ig/m3,
p = 0.26; personal exposure: 4.6 (ig/m3 versus 3.0 (ig/m3, p = 0.06).
3-13
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Indoor activity has been associated with resuspension of settled dust, which could cause
airborne contact with particle-bound Pb. Qian et al. (2008) estimated a PMi0
resuspension rate of 1.4xlO"4/hr for one person walking across a carpeted floor.
Measurements of submicron particles illustrated a roughly two-fold increase of airborne
particle concentration for particles smaller than 1.8 (im for activity versus low activity
periods, with maximum concentrations reaching 4-11 times the maximum value during
low activity periods. For PM10, average concentration was 2.5 times higher than
background levels during activity periods, while peak concentration was 4.5 times higher.
Qian and Ferro (2008) observed that resuspension rates depend on particle size, floor
material, and ventilation position. Increases in walking speed and weight of the walker
did not consistently produce increases in resuspension. 5-10 (im particles produced a
higher resuspension rate compared with smaller particles. Newly carpeted areas produced
significantly higher resuspension rates than vinyl floors. Zhang et al. (2008) modeled and
conducted experiments of particle dispersion from walking and observed that human
activity did affect resuspension. They found that larger particles were more readily
detached from the carpet by walking motion, but that smaller particles are more easily
resuspended once detached. Hunt and Johnson (2012) studied the duration and spatial
extent of resuspension of 0.3-5.0 (im particles following walking by a soiled shoe.
0.3-0.5 (im particle concentration remained increased over a time period of 23 min, while
1-5 (im particles declined in concentration over the same time period. Experiments and
computational fluid dynamics simulations by Eisner et al. (2010) for a mechanical foot
moving on carpeting suggested that the rotating motion of the moving foot on the carpet
induced rotating air movement beneath the foot that re-entrained the particles.
Several of the studies can be used to develop an understanding of how personal exposure
to PM-bound Pb varies with other exposures. Molnar et al. (2007) reported Spearman
correlations of Pb with PM25 andNO2 in three outdoor microenvironments (residence,
school, and preschool) and found that Pb and other trace metals were generally well
correlated with PM25 (r = 0.72 to 0.85), but Pb was only statistically significantly
correlated with NO2 in one of the three outdoor microenvironments (r = 0.24 [residence
and school], versus 0.75 [preschool]). Pb was attributed by Molnar et al. (2007) to long
range transport. Table 3-4 illustrates that measured Pb concentrations (summarized in
Table 2-29 of the Chapter 2 Appendix [Section 2.81) are typically well below the level of
the NAAQS. The higher personal air concentrations occurred in a heavily industrialized
area of Kocaeli, Turkey with an incinerator and several industrial facilities including
metal processing, cement, petroleum refining, and agriculture processing. Otherwise,
concentrations were all between 0.002 and 0.006 ug/m3. The proportion of Pb compared
with other trace metals varied with location and component. It was typically several times
lower than S as well as crustal elements such as Ca2+ and Fe. In the industrial area of
Kocaeli, Pb comprised a greater proportion of the PM compared with other areas.
3-14
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Table 3-4
Study
Clayton et al.
(1999)
Adgate et al.
(2QQ7)
Molnar et al.
(2007)
Tovalin-
Ahumada et
al. (2007)
Pekey et al.
(2010)
Rasmussen et
al. (2007)
Comparison of personal, indoor, and outdoor Pb-PM measurements
from several studies.
Location
IL, IN, Ml, MN,
OH, Wl
Minneapolis-
St. Paul, MN
Stockholm,
Sweden
Mexico City,
Mexico
Puebla,
Mexico
Kocaeli,
Turkey
Windsor,
Ontario,
Canada
3.1.3.2
The 2006 Pb
Pb Metric
Med. Pb-PMso
(ng/m3)
Avg. Pb-PM2.s
(ng/m3)
Avg. Pb-PM2.s
(ng/m3)
Med.
Pb-PM25
(ng/m3)
Med.
Pb-PM25
(ng/m3)
Avg. Pb-PM 2 s
(ng/m3)
Avg. Pb-PM 10
(ng/m3)
Med.
Pb-PM2.5
(mg/kg)
Sampling
Period Personal Pb
July, 1995-
May, 1997
Spring,
Summer, 6.2
Fall, 1999
December,
2003-
July, 2004
April-May,
2002
April-May,
2002
May-June,
2006,
December,
2006-
January 2007
May-June,
2006,
December,
2006-
January 2007
April, 2004 311
Exposure to Pb in Soil and Dust
AOCD (U.S. EPA. 2006b) lists indoor Pb dust
Indoor Pb Outdoor Pb
6.6 8.5
3.4 2.0
Homes: 3.4 Homes: 4.5
Schools: 2.5 Schools: 4.6
Preschools: Preschools:
1.8 2.6
26 56
4 4
Summer: 34 Summer: 47
Winter: 85 Winter: 72
Summer: 57 Summer: 78
Winter: 125 Winter: 159
124 221
infiltrated from outdoors as a
potential source of exposure to Pb soil and dust. Thus, outdoor soil Pb may present an
inhalation exposure if resuspended indoors or an ingestion exposure during hand-to-
mouth contact. A detailed description of studies of outdoor soil Pb concentration is
provided in Section 2.6.1. Indoor measurements can reflect infiltrated Pb as well as Pb
dust derived from debrided paint, consumer products, or soil that has been transported
into the home via foot traffic. Table 3-5 presents indoor dust Pb concentrations for
2006-2011 observational studies in which indoor dust Pb was measured.
3-15
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Table 3-5
Reference
Caravanos et
al. (2006b)
Khoder et al.
(2010)
Brattin and
Griffin (2011)
Yu et al. (2006)
Turner and
Simmonds
(2006)
Gaitens et al.
(2009)
Wilson et al.
(2007)
Zota et al.
(2011)
Spalinger et al.
(2QQ7)
Measurements of indoor dust Pb concentration from 2006-2011
studies.
Study Location
New York City, New York
Giza, Egypt (extensive leaded
gasoline use; industrial area)
Eureka, Utah near Eureka Mills
Superfund Site
Denver, CO, nearVBI70
Superfund Site
East Helena, MT, near East
Helena Superfund Site
Syracuse, New York
Birmingham, Plymouth,
and 2 rural sites, U.K.
U.S. (nationwide)
Milwaukee, Wisconsin
Ottawa County, Oklahoma
(area surrounding the Tar
Creek Superfund Site)
Rural towns, Idaho
Bunker Hill, Idaho Superfund
site
Metric (units)
Weekly dust
loading (ug/m2)
Weekly dust
loading (ug/m2)
Dust
concentration
(mg/kg)
Dust
concentration
range (mg/kg)
Dust
concentration
(mg/kg)
Dust loading
(ug/m2)
Dust
concentration
(ug/m2)
Dust
concentration
(mg/kg)
Dust
concentration
(mg/kg)
Dust
concentration
(mg/kg)
Sample Site
Glass plate next to
open window of
academic building
Glass plate in second-
floor living room of
apartments
Indoor home site (not
specified)
Indoor home site (not
specified)
Indoor home site (not
specified)
Floor
Floor
Smooth floor
Rough floor
Smooth windowsill
Rough windowsill
Central perimeter
Entry
Window
Indoor (site not
specified)
Vacuum
Floor
Vacuum
Floor
Indoor Pb
Concentration
Median: 52
Median: 408
160-2000
11-660
68-1000
Range: 209-1770
Median: 178
Median: 1.7
Avg.: 4.4
Median: 5.6
Avg.: 16
Median: 2.5
Avg.: 190
Median: 55
Avg.: 480
Avg.: 107
Avg.: 140
Avg.: 151
Avg.: 109
Median: 63
Max.: 881
Median: 120
Max: 830
Median: 95
Max: 1,300
Median: 470
Max: 2,000
Median: 290
Max: 4,600
3-16
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Several studies suggested the infiltration of Pb dust into buildings. For example,
Caravanos et al. (2006b) collected dust on glass plates at an interior location near an open
window, a sheltered exterior location, and an open exterior location for a two-year period
in Manhattan, NY. Median weekly dust loading was reported to be 52 (ig/m2 for the
indoor site, 153 (ig/m2 for the unsheltered outdoor site, and 347 (ig/m2 for the sheltered
outdoor site. This paper demonstrated the likely role of outdoor Pb in influencing indoor
dust Pb loading and indicated that under quiescent conditions (e.g., no cleaning) near an
open second-story window, the indoor dust Pb level might exceed EPA's hazard level for
interior floor dust of 430 ug/m2 (40 ug/ft2). Khoder et al. (2010) used the same
methodology to study Pb dust deposition in residential households in the town of Giza,
Egypt, located between two industrial areas and where leaded gasoline is still in use; the
investigators reported a median weekly deposition rate of 408 (ig/m2 and an exterior
median deposition rate of 2,600 (ig/m2. In the latter study, Pb deposition rate correlated
with total dust deposition rate (R=0.92), Cd deposition rate (R=0.95), and Ni deposition
rate (R=0.90). Statistically significant differences in Pb deposition rates were observed
between old and new homes (p <0.01) in the Khoder et al. (2010) study, although the
only quantitative information provided regarding home age stated that the oldest home
was 22 years old when the study was performed in 2007. Khoder et al. (2010) found no
statistically significant difference between Pb loadings when segregating the data by
proximity to roadways. Recently, Brattin and Griffin (2011) performed linear regressions
of dust Pb on soil Pb based on data collected previously for outdoor soil Pb and indoor
dust Pb near mining and/or smelting Superfund sites in Utah, Colorado, and Montana
(U.S. EPA. 2005f: SRC. 2002: U.S. EPA. 2001). They observed that the dust Pb
concentration was 4-35% of outdoor soil Pb. Excluding outliers on the regression, dust
Pb concentration ranged from 160-2,000 mg/kg, 11-660 mg/kg, and 68-1,000 mg/kg at
three sites.
Correlations between indoor and outdoor Pb content in dust can be partially explained
with speciation. Beauchemin et al. (2011) used XANES to speciate in-home paint
samples to assess the contributions of indoor paint and outdoor material to indoor dust Pb
concentrations. In indoor dust samples of particles <150 (im in size, Pb oxide, Pb sulfate,
and Pb carbonate were measured. These materials commonly were used in white paint. In
the size fraction of particles <36 (im, half of the measured Pb was associated with
Fe-oxyhydroxides such as ferrihydrite and goethite and presumably adsorbed onto these
species. This finding suggested that a mix of indoor and outdoor sources may affect the
composition of dust in the smaller size fraction in houses with leaded paint.
Residual Pb dust contamination following cleaning activities has been documented. For
instance, Hunt et al. (2008) estimated Pb deposition and concentration from experiments
in which Herculaneum, MO yard soil samples that had been dried, ground, and sieved
3-17
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were tracked onto a tile test surface and then repeatedly cleaned until visual inspection of
the tiles uncovered no surface discoloration. Cleaning resulted in a 5-6 fold decrease in
residual Pb, with 7,100 (ig/m2 measured after multiple walks across the sample floor prior
to cleaning. Yu et al. (2006) analyzed dust samples from 50 homes in northern
New Jersey (typically of older housing stock, although the study does not specify housing
age). The investigators found that total Pb concentration in carpet dust ranged from 209
to 1,770 mg/kg dust. Wilson et al. (2007) studied Pb dust samples from homes in
Milwaukee, WI, and in resident children with and without elevated blood Pb > 10 (ig/dL.
They found that Pb dust samples obtained from the floor were always significantly higher
in residences of children with elevated blood Pb, with the exception of samples from the
bathroom floor. Windowsill dust was not significantly higher in residences of children
with elevated blood Pb. Residual Pb dust in homes is a potential exposure source for
small children who use touch to explore their environments.
Pb dust on floors, windowsills, and other accessible surfaces is related to several
demographic, socioeconomic, and housing conditions. Gaitens et al. (2009) used
NHANES data from 1999 through 2004 to examine Pb in dust in homes of children ages
12-60 months. Floor Pb dust loading value was modeled against several survey covariates
and was significantly associated with several covariates but with mixed sign (p <0.05).
Floor Pb dust was positively associated with windowsill Pb dust loading, being of non-
Hispanic black race/ethnicity, and presence of smokers in the home. Floor Pb dust was
negatively associated with presence of carpeting, poverty-to-income ratio, and living in a
home built after 1950. It was nearly significantly and positively associated (p = 0.056)
with renovations made to pre-1950 homes. Windowsill Pb dust level was also
significantly associated (p <0.05) with several covariates. It was positively associated
with being of non-Hispanic black race/ethnicity, negatively associated with living in a
home built after 1950, positively associated with not smooth and cleanable window
surface condition, positively associated with presence of smokers in the home, and
positively associated with deterioration of indoor paint. It was nearly statistically
significantly and positively associated (p = 0.076) with deterioration of outdoor paint
when homes were built prior to 1950. Dust Pb loading was found by Egeghy et al. (2005)
to be significantly and positively associated with the log-transform of soil Pb
concentration, cement content in the home entryway, frequency of fireplace usage, and
homes built before 1950. Dust Pb loading was significantly and negatively associated
with indoor pesticide use and number of hours in which windows were open (Table 3-3).
Building demolition and renovation activities can create dust from interior and exterior
paints with Pb content. Mielke and Gonzales (2008) measured Pb content in paint chips
from paint applied prior to 1992 and found that median Pb levels were 420 mg/kg for
interior paint and 77,000 mg/kg for exterior paint. Maximum levels were 63,000 mg/kg
3-18
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and 120,000 mg/kg for interior and exterior paint, respectively. Mielke et al. (2001)
compared dust samples from two New Orleans houses that were prepared for painting.
One home was power sanded without any confinement or control of removed material,
while the other was hand-scraped with containment and collection of paint chips.
Immediately after sanding, Pb dust samples ranged from <32 to 300,000 ng/m2 at the
sanded house. Pb dust samples from the scraped house ranged from 11 to 30,000 ng/m2.
Dust Pb concentrations have also been reported for homes in the vicinity of historic and
active metals mining and smelting sources. As described in Section 2.6.1. soil Pb has
been found to be elevated near source of ambient air Pb. Near an active smelter in Port
Pirie, Australia, median hand dust Pb loadings increased with age among a cohort of
fourteen children followed over age 0-36 months (2-5 months: 54 (ig/m2, >15 months:
336 (ig/m2) (Simon et al.. 2007). Zota et al. (2011) studied Pb dust and indoor Pb-PM2 5
concentration in Ottawa County, OK near the Tar Creek Superfund Site, in which a
metals mine had closed. Statistically significant correlations among outdoor soil Pb
concentration, indoor dust Pb concentration, indoor dust Pb loading, and indoor air
Pb-PM25 concentrations were observed (r = 0.25-0.65), with an average dust Pb
concentration of 109 mg/kg, dust Pb loading of 54 (ig/m2, soil Pb concentration of
201 mg/kg, and indoor Pb-PM2 5 concentration of 1 ng/m3. House dust Pb concentrations
were found to increase significantly with residential proximity to two chat (i.e., dry
mining waste) sources and to decrease with distance to the street and presence of central
air conditioning. Spalinger et al. (2007) measured Pb in dust in homes in a 34 km2 area
surrounding a designated Superfund site where a Pb and Zn smelter formerly operated at
Bunker Hill, ID. During spring of 1999, vacuum and floor mat samples were taken from
homes in three towns within the 34 km2 area and five "background" towns further from
the Superfund site. For the background towns, Pb concentration in vacuum dust had a
median of 120 mg/kg, and Pb concentration in floor dust had a median of 95 mg/kg. The
median Pb dust loading rate was measured to be 40 (ig/m2 per day. In contrast, Pb in
vacuum dust and floor mats for the towns contained within the Bunker Hill Superfund
site had a median Pb concentration of 470 mg/kg and 290 mg/kg, respectively. The
median Pb loading rate for indoor dust in houses in these towns was 300 (ig/m2 per day.
These results suggest that those living in close proximity to large Pb and Zn smelters or
mines that are now Superfund sites are at much greater risk of exposure to Pb dust
compared to the general population.
Pb exposure has been reported on children's playgrounds. Mielke et al. (2011 a) reported
median soil Pb concentration of 558 mg/kg on playground soils at eleven New Orleans
day care or community centers. Following remediation efforts to cover playground soil
with clean soil, median concentration dropped to 4.1 mg/kg. Duggan et al. (1985)
reported on the concentration and size distribution of wipe samples on the hands of 368
3-19
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pre-school children from eleven schools in London, U.K.. Hand Pb residue (PbH) values
were modeled as linear (p <0.05) and power functions (p <0.001) of Pb dust; linear slope
was 0.0064 (ig hand Pb residue per mg/kg Pb dust. Given that the Duggan et al. (1985)
study was performed when Pb additives were used in gasoline, dust Pb concentration
values are not reported here. As described in Section 3.1.1.1, exposure to Pb in soil and
dust may be related to size distribution of the soil or dust particles, with higher Pb
enrichment in the smaller particles.
3.1.3.3 Dietary Pb Exposure
This subsection covers several dietary Pb exposures from a diverse set of sources.
Included among those are drinking water, fish and meat, agriculture, urban gardening,
dietary supplements, tobacco, cultural food sources, and breastfeeding. The breadth of
dietary Pb exposures is illustrated in Figure 3-2. which illustrates the data obtained in the
2008 FDA Total Diet Study market basket survey (FDA. 2008). Among the highest Pb
concentrations were those for noodles, baby food carrots, baby food oatmeal, Swiss
cheese, beef tacos from a Mexican restaurant, and fruit-flavored cereal. Possible sources
of Pb in food samples include introduction during processing or preparation with drinking
water contaminated with Pb, deposition of Pb onto raw materials for each food, and Pb
exposure in livestock that produce dairy or meat ingredients. Manton et al. (2005) used
Pb isotope ratios to estimate sources of dietary Pb among a cohort of mothers and
children from Omaha, NE using a combination of food samples, hand wipes, house dust
wipes, and aerosol samples collected between 1990 and 1997. Drinking water Pb was not
included in this study. The authors cited results from Egan et al. (2002) that imported
vegetables contributed 55% of Pb dietary intake for infants, 30% for 2-6 year old
children, and 20% for 25-30 year old women. Imported candy contributed 10% of Pb
dietary intake for 2-6 year old children and 9% for 25-30 year old women. Isotopic data
from Manton et al. (2005) suggested that, with the exception of children age 0-12
months, house dust is a large contributor to dietary Pb. The pattern of certain Pb-isotope
ratios observed in the diet of children 0-12 months are suggested to derive from Ca salts
in limestone that may have been used in dietary supplements in baby formula. The
contribution of ambient air Pb to dietary Pb samples was not statistically significant for
this urban exposure study.
3-20
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k 1
INQQQIGS ~~
BF(carrot) -
BF(oatmeal) —
Cheese —
Taco —
Cereal —
Egg'n'chs —
Peas —
Milk(2%) -
3F(gmbean) —
Cabbage —
BF(lamb) -
Beans —
Rice —
BF(applej) -
Coffee —
Strawberries —
Squash —
Potato —
Chicken —
Biscuits —
BF(cobbler) -
BF(beef) -
HI 1 1
1 1
i I i H
ED
r-|~~|H
H
h[J}--H
111
r»0
EH
J]
il 1 1 I
ill i
m
h|~~fH
ulM-i
1
O 1- 1 | | H
EH
i i i i i i i
0.00 0.02 0.04 0.06 0.08 0.10 0.12 0.14
Concentration (mg/kg)
Note: from the 2008 FDA Total Diet Study. "BF" denotes baby food.
Source Data: (FDA. 2008)
Figure 3-2
Market basket survey results for Pb concentration in foods.
3-21
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Drinking Water
Pb concentrations in drinking water vary substantially. For example, Shotyk and Krachler
(2009) measured the Pb concentration in tap water, commercially bottled tap water and
bottled natural water. They found that, in many cases, tap water contained less Pb than
bottled water. Excluding bottled water in glass containers because Pb can be leached
from the glass, the median Pb concentration in the bottled water samples was 8.5 ng/L
(range < 1 to 761 ng/L). Pb in drinking water supplies can derive from atmospheric
deposition onto surface waters, runoff of atmospheric deposition as described in Section
2.3. or via corrosion of Pb in the distribution network exacerbated by contact with acidic
disinfection byproducts, as described in the following paragraphs.
It is now recognized that environmental nanoparticles (NPs) (-1-100 nm) can play a key
role in determining the chemical characteristics of treated drinking water as well as
natural waters (Wigginton et al.. 2007). An important question is whether or not NPs
from source waters affect the quality of drinking water. For example, if Fe-oxide NPs are
not removed during the flocculation/coagulation stage of the treatment process, they may
become effective transporters of contaminants such as Pb, particularly if these
contaminants are leached from piping in the distribution system.
Corrosion byproducts can influence Pb concentrations in drinking water. Schock et al.
(2008) characterized Pb pipe scales from 91 pipes made available from 26 different
municipal water systems from across the northern U.S. They found a wide range of
elements including Cu, Zn and V as well as Al, Fe and Mn. Interestingly, V was present
at nearly one percent levels in pipes from many geographically diverse systems. In a
separate study, Gerke et al. (2009) identified the corrosion product, vanadinite
(Pb5(VO4)3Cl) in Pb pipe corrosion byproducts collected from 15 Pb or Pb-lined pipes
representing 8 different municipal drinking water distribution systems in the Northeastern
and Midwest regions of the U.S. Vanadinite was most frequently found in the surface
layers of the corrosion products. The vanadate ion, VO43", essentially replaces the
phosphate ion in pyromorphite and hydroxyapatite structures. It is not known whether the
application of orthophosphate as a corrosion inhibitor would destabilize vanadinite, but
this substitution would have implications for V release into drinking water. The stability
of vanadinite in the presence of monochloramine is also not known, and its stability
might have implications for both Pb and V release into drinking water.
In recent years, drinking water treatment plants in many municipalities have switched
from using chlorine to other disinfecting agents because their disinfection byproducts
may be less carcinogenic. However, chloramines are more acidic than chlorine and can
increase Pb solubility (Raab et al.. 1991) and increase Pb concentrations in tap water. For
example, after observing elevated Pb concentrations in drinking water samples, Kim and
3-22
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Herrera (2010) observed Pb oxide corrosion scales occurring after using acidic alum as a
disinfection agent. Edwards and Dudi (2004) observed a red-brown particle-bound Pb in
Washington, B.C. water that could be confused with innocuous Fe. The source of the
particle-bound Pb was not known but was thought to originate from the source water. The
high Pb concentrations were attributed to leaching of Pb from Pb-bearing pipes promoted
by breakdown products of disinfection agents (Edwards and Dudi. 2004). Maas et al.
(2007) tested the effect of fluoridation and chlorine-based (chlorine and chloramines)
disinfection agents on Pb leaching from plumbing soldered with Pb. When using chlorine
disinfection agents alone, the Pb concentration in water samples doubled during the first
week of application (from 100 to 200 ppb) but then decreased over time. When adding
fluorosilicic acid and ammonia, the Pb concentration spiked to 900 ppb and increased
further overtime. However, Macek et al. (2006) regressed blood Pb among children ages
1-16 years on fluoride treatment, adjusted for several demographic and socioeconomic
factors, and found no association when all data were combined into one model; when
stratifying by housing age, Macek et al. (2006) found statistically significant odds ratios
for those living in housing built before 1946 or for housing age unknown. Similarly,
Lasheen et al. (2008) observed Pb leaching from pipes in Egypt when exposed to an acid
of pH = 6. Exposure to basic solutions actually resulted in reduction of Pb concentration
in the drinking water. Leaching of Pb from pipes following disinfection with acidic
agents can lead to increased Pb exposure; Miranda et al. (2007b) observed a statistically
significant association between blood Pb levels among children living in Wayne County,
NC and use of chloramines (p <0.001) in a log-linear model, although the study did not
control for the presence of Pb paint in the dwellings, so it is difficult to distinguish the
influence of Pb pipes from Pb in paint on blood Pb levels.
Several chemical mechanisms may contribute to release of Pb during use of chloramine
disinfection agents. Edwards and Dudi (2004) hypothesized that Pb leaching occurs when
chloramines cause the breakdown of brass alloys and solder containing Pb. After
observing that nitrification also leads to increased Pb concentrations in water, they also
proposed that chloramines may trigger nitrification and hence cause decreasing pH,
alkalinity, and dissolved oxygen that leads to corrosion after observing that nitrification
also leads to increased Pb concentrations in water. However, Zhang et al. (2009b) found
no evidence that nitrification brought about significant leaching of Pb from Pb pipes.
Lytle et al. (2009) suggested that a lack of increased Pb(II) concentrations in drinking
water following a change from free chlorine to chloramine disinfection is attributed to the
formation of the Pb(II) mineral hydroxypyromorphite (Pb5(PO4)3OH) instead of Pb(IV)
oxide. Xie et al. (2010) further investigated the mechanisms by which Pb(II) release is
affected by chloramines. Two opposing mechanisms were proposed: Pb(IV)O2 reduction
by an intermediate species from decomposition of monochloramine; and increasing redox
potential which decreases the thermodynamic driving force for reduction. They suggest
3-23
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that the contact time of monochloramine with PbO2 and the C12:N ratio in
monochloramine formation will determine which mechanism is more important. Free
chlorine can control Pb concentrations from dissolution under flowing conditions but for
long stagnation periods, Pb concentrations can exceed the action level: 4-10 days were
required for Pb concentrations to exceed 15 (ig/L (for relatively high loadings of PbO2, of
1 g/L). Thus, under less extreme conditions, it was concluded that chloramination was
unlikely to have a major effect on the release of Pb into drinking water.
Agriculture
The 2006 Pb AQCD (2006b) states that surface deposition "represents a significant
contribution to the total Pb in and on the plant", while uptake through a plant's roots can
also contribute to a plant's Pb concentration. Consequently, Pb content in plants may
contribute to human dietary exposure. Uptake of Pb by plants growing in contaminated
soil has been repeatedly demonstrated in some species during controlled potted plant
experiments (Del Rio-Celestino et al.. 2006). In this study, most species retained Pb in
the roots with little mobilization to the shoots of the plants. However, certain species of
grasses were able to mobilize Pb from the roots to the shoots of the plant; these specific
species could lead to human exposures through consumption of grazing animals. Lima et
al. (2009) conducted similar greenhouse experiments with several vegetable crops grown
in soil contaminated by Pb-containing residue from battery recycling waste. In this study,
carrots had high bioaccumulation, measured as the percent of Pb concentration measured
in the plant compared with the Pb concentration in the soil, with little translocation of the
Pb to the shoots. Conversely, beets, cabbages, sweet peppers, and collard greens had low
bioaccumulation but moderate to high translocation. Okra, tomatoes, and eggplants had
moderate bioaccumulation and moderate to high translocation. Sesli et al. (2008) also
noted uptake of Pb within wild mushrooms. Vandenhove et al. (2009) reviewed
bioaccumulation data for plant groupings and found that grasses had the highest uptake,
followed by leafy vegetables and root crops grown in sandy soils; see Table 3-6. These
references also suggested high transfer from roots to shoots among root crops, with
shoots having roughly four times higher Pb bioaccumulation than roots.
Sources of atmospheric Pb can lead to vegetable contamination. For example, Uzu et al.
(2010) found that Pb deposition from smelter emissions caused a linear increase in Pb
concentrations of 7.0 mg/kg per day (R2=0.96) in lettuce plants cultivated in the
courtyard of a smelter. They reported that lettuce grown 250-400 meters from the smelter
had concentrations that were 10-20 times lower, which is consistent with findings
described in Section 2.3 that deposition of Pb containing material drops off with distance
from a source. Pb contamination of crops may also occur through piston-engine aircraft
Pb emissions during aerial application of fertilizers and pesticides. In 2010, the U.S.
3-24
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Federal Aviation Administration (FAA) recorded 396,000 hours of flight time for aerial
application. This term encompasses crop and timber production including seeding
cropland and fertilizer and pesticide application. It is estimated that 86% of these flight-
hours involved piston engine aircraft utilizing leaded fuel (FAA. 2010).
Some land use and soil characteristics have been shown to increase bioaccessibility of Pb
in soil, which could then lead to plant contamination. Fernandez et al. (2010; 2008; 2007)
measured Pb from atmospheric deposition in two adjacent plots of land having the same
soil composition but different uses: one was pasture land and one was agricultural. In the
arable land, size distributions of soil particle-bound Pb, were uniformly distributed. In
pasture land, size distributions of soil particle-bound Pb were bimodal with peaks around
2-20 urn and 50-100 um (Fernandez et al.. 2010). For the agricultural plot, Pb
concentration was constant around 70 mg/kg in samples taken over the first 30 cm of soil,
at which time it dropped below 10 mg/kg at soil depths between 35 and 100 cm. In
contrast, Pb concentration in pasture land peaked at a depth of 10 cm at a concentration
of roughly 70 mg/kg and then dropped off gradually to approach zero concentration at a
depth of approximately 50 cm. The sharp change in concentration for the arable land was
attributed to a combination of plowing the soil and use of fertilizers to increase the
acidity of the soil and solubility of Pb into the soil (Fernandez et al.. 2007). They found
that the surface layer was acidic (pH: 3.37-4.09), as was the subsurface layer (pH:
3.65-4.38). Jin et al. (2005) examined how soil characteristics affect Pb contamination of
crops by testing soil Pb, bioaccessibility of soil Pb (determined by CaCl2 extraction), and
Pb in tea samples from tea gardens. They observed that the Pb concentration in tea leaves
was proportional to the bioaccessible Pb in soil.
There is some evidence that Pb contamination of crops can originate with treatment of
crops. For example, compost produced from wastewater sludge has the potential to add
Pb to crops. Cai et al. (2007) demonstrated that production of compost from sludge
enriched the Pb content by 15-43% compared with the Pb content in sludge prior to
composting. Chen et al. (2008b) observed that the median concentration of Pb in
California crop soil samples was 16.2 mg/kg (range: 6.0-62.2 mg/kg). Chen et al. (2008a)
further observed that in three of the seven California agricultural regions sampled,
concentrations of Pb increased following addition of fertilizer, but the increase was less
than that for phosphorous (P) and Zn indicators of fertilizer. In four regions, there was no
increase of Pb at all. Furthermore, Tu et al. (2000) observed a decrease in Pb fraction
with increasing P application. Nziguheba and Smolders (2008) also surveyed phosphate-
based fertilizers sold in European markets to determine the contribution of these
fertilizers to heavy metal concentrations in agricultural products. They reported a median
fertilizer Pb concentration of 2.1 mg/kg based on total weight of the fertilizer, with a 95th
percentile concentration of 7.5 mg/kg. Across Europe, Nziguheba and Smolders (2008)
3-25
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estimated that the amount of Pb applied via fertilizers to be only 2.6% of that resulting
from atmospheric deposition.
Although Pb in on-road vehicle gasoline has been phased out in the U.S., if imported
crops are produced in countries that still use Pb antiknock agents in on-road gasoline,
they have the potential to introduce dietary Pb to U.S. consumers. For example, high
concentrations of Pb have been found in chocolate from beans grown in Nigeria, during
the time when leaded gasoline was still legally sold. Rankin et al. (2005) observed that
the ratios of 207Pb to 206Pb and 208Pb to 207Pb were similar to those of Pb in gasoline.
Although this study showed that Pb concentration in the shelled cocoa beans was low
(~1 ng/g), manufactured cocoa powder and baking chocolate had Pb concentrations
similar to those of the cocoa bean shells, on the order of 200 ng/g, and Pb concentration
in chocolate products was roughly 50 ng/g (Rankin et al.. 2005). It is possible that the
increases were attributed to contamination of the cocoa by the shells during storage or
manufacture, but the authors note that more research is needed to verify the source of
contamination.
Findings from Pb uptake studies have implications for urban gardening if urban soils may
be contaminated with Pb. For instance, Clark et al. (2006) tested the soil in 103 urban
gardens in two Boston neighborhoods. Using isotopic analysis, they found that Pb-based
paint contributed 40-80% of Pb in the urban garden soil samples, with the rest coming
from historical Pb emissions. Furthermore, Clark et al. (2006) estimated that Pb
consumption from urban gardens can be equivalent to 10-25% of the exposure to Pb from
drinking water for children living in the Boston neighborhoods studied. Because soil Pb
levels in urban areas will depend on surrounding sources (Pruvot et al.. 2006). Pb
exposures in urban garden vegetables will vary.
3-26
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Table 3-6 Pb bioaccumulation data for various plants. Bioaccumulation is
expressed as percent of Pb concentration in the plant to the Pb
concentration in the soil.
Plant Group
All
Cereals
Maize
Rice
Leafy Vegetables
Non-Leafy
Vegetables
Legumes
Root Crops
Tubers
Fruits
Grasses
Natural Pastures
Leguminous
Fodder
All Cereals
Pastures/Grasses
Fodder
Plant
Compartment
Grain
Straw
Grain
Straw
Grain
Fruits
Shoots
Pods
Shoots
Roots
Shoots
Tubers
Fruits
Leaves
Soil
All
All
All
All
All
All
Sand
Loam
Clay
All
All
All
Sand
Loam
Clay
All
All
Sand
Loam
All
All
Sand
Loam
All
All
All
All
All
All
Sand
Loam
Clay
All
All
Sand
Clay
n
210
9
4
9
3
2
31
4
3
7
5
2
17
3
5
4
1
27
5
5
12
30
5
17
5
1
17
34
1
20
5
8
6
51
24
4
4
GM
2.0%
1.0%
2.3%
0.12%
0.28%
8.0%
7.3%
82%
2.8%
1.5%
0.53%
0.27%
0.14%
0.080%
1.5%
6.4%
2.3%
6.3%
0.15%
0.64%
0.052%
0.77%
31%
92%
0.43%
0.61%
0.17%
0.90%
14%
2.5%
4.5%
0.82%
GSD
14
3.6
3.5
2.3
6.6
13
1.5
1.0
4.1
26
12
3.2
4.4
1.0
16
1.6
4.7
15
7.4
3.5
2.4
2.6
1.8
4.8
4.7
5.3
3.9
4.0
4.2
12
2.3
5.7
AM
63%
1.8%
3.8%
0.17%
0.85%
2.2%
210%
7.8%
82%
5.1%
78%
0.88%
34%
0.42%
0.42%
0.33%
0.080%
41%
7.0%
0.50%
250%
9.1%
1.2%
0.073%
1.0%
25%
36%
23%
1.6%
1.1%
1.3%
0.53%
1.8%
27%
130%
5.6%
2.7%
SD
290%
1.6%
4.0%
0.14%
1.3%
1.4%
610%
3.3%
3.5%
4.8%
170%
0.42%
120%
0.34%
0.34%
0.47%
98%
3.4%
0.68%
570%
48%
1.6%
0.062%
0.60%
22%
29%
1.4%
1.3%
1.1%
1.8%
27%
420%
4.0%
4.6%
Min
0.015%
0.19%
0.51%
0.052%
0.060%
1 .2%
0.32%
4.9%
79%
0.41%
0.15%
0.58%
0.046%
0.065%
0.065%
0.046%
0.024%
4.2%
0.024%
0.30%
0.015%
0.16%
0.015%
0.15%
11%
0.22%
0.052%
0.052%
0.059%
0.22%
0.22%
0.060%
1 .6%
0.16%
Max
2,500%
4.8%
9.6%
0.38%
2.3%
3.2%
2,500%
11%
86%
12%
390%
1.17%
490%
0.89%
0.89%
1.0%
330%
12%
1.7%
16%
260%
3.9%
0.23%
1.7%
100%
100%
4.8%
3.2%
3.2%
4.8%
100%
1,600%
11%
9.6%
Source: Reprinted with permission of Elsevier Publishers, Vandenhove et al. (2009)
3-27
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Game
Atmospheric sources of Pb have also been shown to contaminate game meat, thus
potentially posing a risk of Pb exposure. In Pb mining or smelting areas, several studies
have documented Pb concentrations in game [e.g., (Nwude et al., 2010; Reglero et al.,
2009b)1.
Potential Pb exposure through consumption of animals exposed to or killed with Pb shot
has also been well documented (Hunt et al., 2009; Tsuji et al., 2009; Tsuji et al., 2008;
Hunt et al.. 2006). For example, Martinez-Haro et al. (2010) observed Pb in the feces of
mallards that ingested gunshot of 34-13,930 mg/kg with a median of 1,104 mg/kg, while
mallards that did not ingest gunshot had feces Pb levels <12.5 mg/kg. Mateo et al. (2011)
studied Pb bioaccessibility as a function of cooking method for breast meat from
partridges killed with gunshot. They observed that preparation in cold or hot vinegar
increased bioaccessibility compared with total Pb in the samples.
Fish
Pb content in fish could also lead to human exposure to Pb (U.S. EPA. 2006b. 1986a).
Ghosh et al. (2007) demonstrated in laboratory experiments that exposure to Pb in water
can lead to linearly increasing Pb levels in the kidneys, liver, gills, skeleton, and muscle
offish. Several studies have documented the potential for human Pb exposure through
fish and seafood. Welt et al. (2003) conducted a survey of individuals who fished in
Bayou St. John, Louisiana in conjunction with sampling Pb content in sediment. They
found that median sediment Pb concentrations ranged from 43 to 330 mg/kg in different
locations, while maximum sediment Pb concentrations ranged from 580 to 6,500 mg/kg.
In total, 65% of the surveyed individuals fished for food from the Bayou, with 86%
consuming fish from the Bayou each week. In a study of the effect of coal mining on
levels of metals in fish (measured as blood Pb) in northeastern Oklahoma, Schmitt et al.
(2005) found that fish blood Pb levels varied with respect to species offish, but blood Pb
levels were higher in fish in areas close to mining activities. Similarly, Besser et al.
(2008) observed higher levels of fish blood Pb, close to mining activities in southeastern
Missouri. In a related study offish species in the same region of Missouri, fish blood Pb
levels were found to be statistically significantly higher in sites within 10 km downstream
of active Pb-Zn mines (p <0.01) compared with fish located further from the mines
(Schmitt et al.. 2007a). and elevated fish blood Pb levels were again noted near a Pb-Zn
mine (Schmitt et al.. 2009). It was noted that the Ozark streams where these studies were
performed were often used for recreational fishing.
3-28
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Breast Milk
Studies of breastfeeding women suggest that infants may be exposed to Pb in breast milk.
Ettinger et al. (2004a) observed in a 1994-1995 study of Mexico City women that at
1 month postpartum, 88 women breastfeeding exclusively (with mean blood Pb level of
9.4 (ig/dL) had breast milk Pb concentrations of 1.4 ± 1.1 (ig/L, and 165 women
breastfeeding partially (with mean blood Pb level of 9.5 (ig/dL) had breast milk Pb
concentrations of 1.5 ± 1.2 (ig/L. During the same time period, Ettinger et al. (2006)
studied breastfeeding women in Mexico City over a child's first year of life and sampled
Pb concentration in breast milk at 1, 4, and 7 mo post-partum. They observed that mean
breast milk concentrations dropped from 1.4 (ig/L at 1 mo (mean maternal blood
Pb = 9.3 (ig/dL) to a mean of 1.2 (ig/L at 4 mo (mean maternal blood Pb = 9.0 (ig/dL) to
0.9 (ig/L at 7 mo (mean maternal blood Pb = 8.1 (ig/dL); this reduction was statistically
significant (p <0.00001). Among the 310 women included in the study, 181 had previous
pregnancies. In one study of nursing mothers living in Port Pirie, Australia near a Pb
smelter, 10 of the 11 mothers had breast milk concentrations <5 (ig/L (Simon et al..
2007). The authors hypothesized that breast milk concentration was too low to be a major
contributor to blood Pb level in these infants relative to other factors such as hand loading
of Pb. However, one mother with a blood Pb level of 25 (ig/dL had a breast milk Pb level
of 28 (ig/L (Simon et al.. 2007).
In summary, several sources of dietary Pb can originate from atmospheric Pb emissions,
including drinking water, vegetables, game, fish, and breast milk. Drinking water Pb
levels are affected by source strength and proximity, runoff, and water treatment
processes and chemicals. Among plants grown for agriculture, Pb content is highest in
grasses, followed by leafy vegetables, then root vegetables. Pb in soil or dust can also
collect on the surfaces of vegetables. Pb contamination of vegetables depends on a
number of factors, including presence of nearby sources of atmospheric Pb, soil type and
chemistry, land use, and land treatment. Other sources of Pb, such as international
consumer products or historic emissions, also have the potential to introduce Pb into the
U.S. diet. Pb contamination through the food chain potentially leads to elevated Pb levels
in meat. Likewise, Pb contamination of surface waters can lead to elevated levels of Pb in
fish used for consumption. Breastfeeding also presents a potential Pb exposure to
newborn babies, and exposure drops off as the mothers nurse and as the babies age and
add more food to their diet.
3-29
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3.1.3.4 Occupational
Occupational environments have the potential to expose individuals to Pb. Some modern
day occupational exposures are briefly discussed below in the context of understanding
potential exposures that are not attributed to ambient air. For example, Miller et al.
(2010) obtained personal and area samples of particle-borne Pb in a precious metals
refinery; year of the study was not reported. It was not stated explicitly, but it is likely
that Miller et al. (2010) measured the PM as TSP because the Occupational Safety and
Health Administration (OSHA) permissible exposure limit (PEL) for Pb is based on TSP
rather than a smaller size cut, and the OSHA PEL was used for comparison.
Concentrations measured by personal samples ranged from 2 to 6 (ig/m3, and
concentrations from area samples ranged from 4 to 14 (ig/m3. The OSHA PEL is 5 (ig/m3.
In steel production, sintering was found to be the largest source of airborne Pb exposure
in a survey of operations (Sammut et al.. 2010). with Pb enrichment in PM reported to be
20,000 mg/kg. Although total PM concentration was not reported by the authors, the PM
was reported to have 75% of its particulate mass at below the 2.5 (im diameter size.
Operations involving Pb-containing materials in various industries are a source of
occupational Pb exposure, in addition to a residential exposure. Rodrigues et al. (2010)
reported exposures to airborne Pb among New England painters, who regularly use
electric grinders to prepare surfaces for painting. Two-week averaged airborne Pb
concentrations, sampled with an Institute of Medicine inhalable PM sampler designed to
capture PM smaller than 100 (im, were reported to be 59 (ig/m3, with a maximum daily
value of 210 (ig/m3. The Pb concentrations reported here were corrected by the National
Institute for Occupational Safety and Health (NIOSH) respirator protection factors,
although the respirator protection factors were not reported by Rodrigues et al. (2010).
Information on the air Pb-blood Pb relationship can be found in Section 3.5.1.
3.1.3.5 Exposure to Pb from Consumer Products
Pb is present in varying amounts in several consumer products including alternative
medicines, candies, cosmetics, pottery, tobacco, toys, and vitamins (Table 3-7). Several
of these categories suggest children may incur regular exposures. Pb concentrations were
reported to range from non-detectable levels up to 77% by mass, for the case of one
medicinal product. Exposure to these products, which originate in a range of different
countries, can account for substantial influence on Pb body burden (Miodovnik and
Landrigan. 2009; Levin et al.. 2008).
3-30
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Table 3-7 Pb content in various consumer products.
Product
Category
Alternative
and
Traditional
Medicines
Product
Cissus quadrangularis, Caulophyllum
thalictroides, Turnera diffuse, Centella
asiatica, Hoodia gordonii, Sutherlandia
frutescens, Curcuma longa, fucoxanthin,
Euterpe oleracea
(dietary supplements claimed to be from
Hoodia gordonii)
Malva sylvestris
Yugmijihwang-tang, Bojungigki-tang,
Sibjeondaebo-tang, Kuibi-tang,
Ojeogsan
Lemongrass, licorice, holy basil, cloves,
ginger
Location of
Purchase
U.S.
(Mississippi)3
Turkey
Korea
India
Pb Content (units)
Not detected (N.D.)b
to 4.21 mg/kg
1.1-2.0 mg/kg
7.9x10-6to2.5xirj5
mg/kg body weight
per day
Average:
Lemongrass &
Reference
Avula et al.
(2010)
Hicsomnez et
al. (2009)
Kim et al.
(2009a)
Naithani and
Kakkar (2006)
Holy Basil Leaves:
6.1 mg/kg;
Licorice Stolons:
6.1 mg/kg,
Clove Dried Flower
Buds: 7.8 mg/kg,
Ginger Rhizome:
5.8 mg/kg
B-Success 28, Operation Sweep, Aloe Nigeria 925-27,000 ug
Vera Plus Bitter Aloes, Zarausmacine,
Virgy-Virgy Computer Worm-Expeller,
Dorasine Powder, Sexual Energy,
U&DEE Infection Cleansing Powder,
U&DEE Sweet Bitter, Natural Power
Stone, Chama Black Stone, Portugal
Antiseptic Soap, Edysol Antiseptic Soap,
H-Nal, M-Reg, Veins Flocher, Diabor, C-
Candi, C-Cysta, Firas, D-Diab, P-Pile,
Infecta, Ribacin Forte, Aloe Vera Cure
Formula
Shell of Hen's Egg India 14 mg/kg
Berberis (B. aristata, B. chitria, India Berberis:
B. asiatica, B. lyceum), Daruharidra Roots' 3 1-24 7 mg/h
Obi et al.
(2006)
Sharma et al.
(2009)
Srivastava et al.
(2006)
Stems: 8.0-23.8 mg/kg
Daruharidra:
16.9-49.8 mg/kg
Greta powder
U.S.
(California)
770,000 ppm
CDC (2002)
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Table 3-7 (Continued): Pb content in various consumer products.
Product
Category
Candy
Cosmetics
Pottery
Tobacco
Toys
Vitamins
Product
Tamarind Candy
Tamarind Candy
Lipsticks
Eye Shadows
Foods prepared in Pb-glazed pottery
Smokeless Tobacco
Cigarette Tobacco (210Pb
concentrations)
Red and yellow painted toy vehicles and
tracks
535 PVC and non-PVC toys from day
care centers
Soft plastic toys
Toy necklace
Soft plastic toys
Vitamins for young children, older
children, and pregnant or lactating
women
Location of
Purchase
U.S.
(Oklahoma)
U.S.
(California)
U.S.
Nigeria
Mexico
U.K.
Pakistan
Brazil
U.S.
(Nevada)
India
U.S.
Nigeria
U.S.
Pb Content (units)
Product: 0.15-3.61
mg/kg
Stems: 0.36-2.5 mg/kg
Wrappers: 459-27,125
mg/kg
Product: 0.2-0.3 mg/kg
Stems: 400 mg/kg
Wrappers:
16,000-21, 000 mg/kg
Average: 1.07 mg/kg
N.D. to 55 mg/kg
N.D. to 3, 100 mg/kg
0.15-1.56 mg/kg
Activity cone.: 7-20
Bq/kg
500-6,000 mg/kg
PVC: avg. 325 mg/kg
Non-PVC: avg. 89
mg/kg Yellow: 216
mg/kg
Non-yellow: 94 mg/kg
Average (by city):
21 -280 mg/kg
388,000 mg/kg
2. 5-1 ,445 mg/kg
Average:
Young children:
2.9 ug/day
Older children:
1.8 ug/day
Pregnant and lactating
women: 4.9 ug/day
Reference
Lynch et al.
(2000)
CDC (2002)
Hepp et al.
(2009)
Omolaoye et al.
(201 Oa)
Villalobos et al.
(2009)
McNeill et al.
(2006)
Tahir and
Alaamer (2008)
Godoi et al.
(2009)
Greenway and
Gerstenberger
(2010)
Kumar and
Pastore (2007)
Meyer et al.
(2008)
Omolaoye et al.
(2010b)
Mindak et al.
(2008)
aHoodia gordonii, from Eastern Cape, South Africa Euterpe oleracea from Ninole Orchard, Ninole, Hawaii
*Note that the country of origin is not provided because it was not published in the references cited.
3-32
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3.2 Kinetics
This section summarizes the empirical basis for understanding Pb toxicokinetics in
humans. The large amount of empirical information on Pb biokinetics in humans and
animal models has been integrated into mechanistic biokinetics models (U.S. EPA.
2006b). These models support predictions about the kinetics of Pb in blood and other
selected tissues based on the empirically-based information about Pb biokinetics. In
Section 3.3 (and Section 3.2.2.1). Pb biokinetics is described from the context of model
predictions.
The discussion of Pb toxicokinetics emphasizes inorganic Pb since this comprises the
dominant forms of Pb to which humans in the U.S. are currently exposed as a result of
releases of Pb to the atmosphere and historic surface deposition of atmospheric Pb (see
Section 2.2.2). The toxicokinetics of organic Pb is only briefly described and a more
extensive discussion can be found in the 2006 Pb AQCD. Human exposures to organic
Pb could occur in occupational settings (e.g., during manufacturing of tetraethyl Pb or
aviation fuels); however, environmental exposures to organic Pb compounds rarely occur
in the U.S. other than in the limited circumstances of those involved in fueling piston-
driven aircraft that use leaded aviation gasoline.
3.2.1 Absorption
The major exposure routes of Pb in humans are inhalation and ingestion. Therefore, these
exposure routes are important in the discussion of Pb absorption (see Section 3.2.1.1 and
Section 3.2.1.2). The term "absorption" refers to the fraction of the amount of Pb ingested
or inhaled that is absorbed from the respiratory or gastrointestinal tract. The term
bioavailability, as it is used in this section, refers to the fraction of the amount of Pb
ingested or inhaled that enters the systemic circulation. If properly measured (e.g., time-
integrated blood Pb), under most conditions Pb bioavailability is equivalent (or nearly
equivalent) to Pb absorption. The time-integrated blood Pb (i.e., the integral of blood Pb
over time) provides a useful measure of bioavailability because it reflects both recent Pb
absorption as well as contributions from Pb sequestered in soft tissue and bone.
Bioaccessibility is a measure of the physiological solubility of Pb in the respiratory or
gastrointestinal tract. Pb must become bioaccessible in order for absorption to occur.
Processes that contribute to bioaccessibility include physical transformation of Pb
particles and dissolution of Pb compounds into forms that can be absorbed (e.g., Pb2+).
Bioaccessibility is typically assessed by measuring the fraction of Pb in a sample that can
3-33
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be extracted into a physiological or physiological-like solution (e.g., gastric juice or
solution similar to gastric juice).
The 2006 Pb AQCD (U.S. EPA. 2006c) also presented dermal absorption of inorganic
and organic Pb compounds, which is generally considered to be much less than by
inhalation or ingestion. A study published subsequent to the 2006 Pb AQCD measured
rates of absorption of Pb in skin patches harvested from nude mice (PanetaL 2010).
Following application of 12 mg Pb as Pb acetate or Pb nitrate, the absorption rate
(measured over a 10-hour observation period) was approximately 0.02 (ig Pb/cm2 per
hour. Absorbed Pb was detected in liver and kidney of nude mice following a 120-hr
occluded dermal application of approximately 14 mg Pb as either Pb acetate or Pb nitrate.
Uptake of Pb into the skin at the site of application was greater when Pb acetate was
applied to the skin compared to Pb nitrate; however, liver and kidney Pb concentrations
observed at the conclusion of the study (120 hours following the application of Pb) were
not different for the two Pb compound. No additional information provides evidence of
dermal absorption being a major exposure route of environmental Pb.
3.2.1.1 Inhalation
Systemic absorption of Pb deposited in the respiratory tract is influenced by particle size
and solubility, as well as by the pattern of regional deposition within the respiratory tract.
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 (im) that are primarily
deposited in the ciliated airways (nasopharyngeal and tracheobronchial regions) can be
transferred by mucociliary transport into the esophagus and swallowed, thus being
absorbed via the gut.
Inhaled Pb lodging deep in the respiratory tract seems to be absorbed equally and totally,
regardless of chemical form (Morrow et al.. 1980; Chamberlain et al.. 1978; Rabinowitz
et al.. 1977). Absorption half-times (ti/2) have been estimated for radon decay progeny in
adults who inhaled aerosols of Pb and bismuth isotopes generated from decay of 220Rn or
222Rn. The absorption half-time for Pb from the respiratory tract to blood was estimated
to be approximately 10 hours in subjects who inhaled aerosols having an activity median
particle diameter of approximately 160 nm (range 50-500 nm) (Marsh and Birchall.
1999). and approximately 68 min for aerosols having diameters of approximately 0.3-
3 nm (Butterweck et al.. 2002). Given the submicron particle size of the exposure, these
rates are thought to represent, primarily, absorption from the bronchiolar and alveolar
regions of the respiratory tract.
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Several studies have quantified the bioaccessibility of Pb in atmospheric PM, based on
various in vitro extraction methods. In a study of PMi0 and PM25 samples from
downtown Vienna, Austria, Falta et al. (2008) used synthetic gastric juice to investigate
the bioaccessibility of metals including Pb. The rationale was that inhaled particles in the
2.5-10 (im size range are mostly deposited in the tracheal and bronchial regions of the
lung from where they are transported within hours by mucociliary clearance, i.e., they are
mainly swallowed. In contrast, the <2.5 (im particles are deposited in the pulmonary
alveoli where they can stay for months to years. The study aimed to determine the
bioaccessibility of the 2.5-10 (im PM. It is important to note that they do not isolate the
2.5-10 (im size range; instead, they infer the characteristics from the difference between
the PM2 5 and PMi0 fractions. The Pb concentrations associated with the two fractions
were almost identical, as was the percentage extracted by synthetic gastric juice (86% and
83% Pb for PM2 5 and PMi0 fractions, respectively). The mean daily bioavailable mass
was calculated to be 16 ng for the PM2 5.10 size range. Since the quantitative clearance of
these particles to the stomach was assumed, this value represents an upper estimate for
the amount of bioavailable Pb. Niu et al. (2010) determined the bioaccessibility of Pb in
fine (100-1,000 nm) and ultrafme-sized (<100 nm) urban airborne PM from two sites
within the city of Ottawa, Canada. For all size fractions, the median Pb concentrations for
particles smaller than 10 um were 8,800 and 7,800 mg/kg for the two different locations.
The bioaccessibility was based on ammonium acetate extractability and it was found that,
within the fine and ultrafine-size ranges, 13-28% Pb was extracted. The Falta et al.
(2008) and Niu et al. (2010) results illustrate that different extraction techniques result in
different bioaccessible fractions. The main finding from Niu et al. (2010) was that the
highest values (-28% and -19% for the two different locations) were found for the
<57 nm particles, with percent bioaccessibility decreasing with increasing particle size.
This result indicated that Pb was potentially most bioaccessible in the ultrafine-size
range.
A recent study by Barrett et al. (2010) investigated the solid phase speciation of Pb in
urban road dust in Manchester, U.K., and considered the health implications of inhalation
and ingestion of such material. Human exposure via inhalation is likely to involve only
the finest grained fractions (up to 10 (im) and unfortunately this study characterized only
the <38 (im fraction. Pb-goethite and PbCrO4 comprised the largest fractions, 45% and
21% respectively, of Pb in the <38 (im fraction. These forms tend to be less bioaccessible
if ingested compared with PbO or Pb acetate because they are less soluble.
The above considerations indicate that the relationship between air Pb exposure and
blood Pb will depend on numerous exposure variables (e.g., particle size, solubility,
exposure frequency and duration) and physiological variables (age, activity level,
transport and absorption in the respiratory tract, blood Pb kinetics). For a detailed
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discussion of factors affecting particle deposition and retention in the human respiratory
tract the reader is referred to Chapter 4 of 2009 PM ISA (U.S. EPA. 2009a). Section 4.2.4
of that document specifically addresses biological factors affecting particle deposition
such as activity level and age with an emphasis on children. Mechanistic models provide
one means for integrating these variables into predictions of blood Pb - air Pb
relationships; although, predictions are subject to simplifications and generalizations
made in constructing the models. As an example, the ICRP (Pounds and Leggett 1998;
ICRP. 1994; Leggett 1993) model (see Section 3.3 for a brief description) can be used to
predict blood Pb - air Pb slopes for specific direct Pb inhalation exposure scenarios. For a
long-term continuous (24 hours/day) exposure of a typical adult male engaged in light
exercise (ventilation rate 20-22 m3/day) to Pb-bearing particles having a 1 (im uniform
particle size, the predicted blood Pb - air Pb slopes range from 0.7 (ig/dL per (ig/m3 (for
low solubility particles; e.g., Pb oxide) to 3 (ig/dL per (ig/m3 (for highly soluble Pb;
e.g., Pb salts). These slopes were calculated by running ICRP model simulations with
varying air concentrations (0.1 - 6 (ig/m3) to achieve a range of blood Pb concentrations
up to 10 (ig/dL, starting with a baseline of 1.6 (ig/dL, and estimating the linear slope of
the relationship between blood Pb concentration and air Pb. Empirical estimates of blood
Pb - air Pb slopes for various populations, derived from epidemiological studies, are
summarized in Section 3.5.1.
Organic Pb
Alkyl Pb compounds can exist in ambient air as vapors. Inhaled tetraalkyl Pb vapor is
nearly completely absorbed following deposition in the respiratory tract. As reported in
the 2006 Pb AQCD (U.S. EPA. 2006c). a single exposure to vapors of radioactive (203Pb)
tetraethyl Pb resulted in 37% initially deposited in the respiratory tract, of which -20%
was exhaled in the subsequent 48 hours (Heard 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 hours (Heard et al..
1979).
Estimation of bioavailability of tetraethyl Pb following combustion is re levant to some
aviation exposures (e.g., persons exposed to leaded gasoline used in piston-engine
aircraft). Chamberlain et al. (1975) suggested that 35% of inhaled combustion products of
tetraethyl 203Pb fuel [likely to have been a mixture dominated by inorganic Pb halides, but
may also have included alkly Pb species (U.S. EPA. 2006b)1 are deposited and then
retained in adult lungs with a half-life of 6 hours. Fifty percent of that 203Pb was
detectable in the blood within 50 hours of inhalation, and the rest was found to deposit in
bone or tissue. Chamberlain et al. (1975) estimated that continuous inhalation of Pb in
3-36
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engine exhaust from fuel containing tetraethyllead at a concentration of 1 ug/m3 for a
period of months could produce a 1 ug/dL increment in blood Pb.
3.2.1.2 Ingestion
The extent and rate of GI absorption of ingested inorganic Pb are influenced by
physiological states of the exposed individual (e.g., age, fasting, nutritional calcium
(Ca2+) and iron (Fe) status, pregnancy) and physicochemical characteristics of the
Pb-bearing material ingested (e.g., particle size, mineralogy, solubility). Pb absorption in
humans may be a capacity-limited process, in which case the percentage of ingested Pb
that is absorbed may decrease with increasing rate of Pb intake. Numerous observations
of nonlinear relationships between blood Pb concentration and Pb intake in humans
provide support for the likely existence of a saturable absorption mechanism or some
other capacity-limited process in the distribution of Pb in humans (Sherlock and Quinn.
1986; Sherlock et al.. 1984; Pococketal.. 1983; Sherlock et al. 1982). While evidence
for capacity-limited processes at the level of the intestinal epithelium is compelling, the
dose at which absorption becomes appreciably limited in humans is not known.
In adults, estimates of absorption of ingested water-soluble Pb compounds
(e.g., Pb chloride, Pb nitrate, Pb acetate) range from 3 to 10% in fed subjects (Maddaloni
etal.. 1998: Watson etal.. 1986: James etal.. 1985: Heard and Chamberlain. 1982:
Rabinowitz et al.. 1980). The absence of food in the GI tract increases absorption of
water-soluble Pb in adults. Reported estimates of soluble Pb absorption range from 26 to
70% in fasted adults (Maddaloni et al.. 1998: James etal.. 1985: Blake etal.. 1983: Heard
and Chamberlain. 1982: Rabinowitz et al.. 1980). Reported fed:fasted ratios for soluble
Pb absorption in adults range from 0.04 to 0.2 (James etal.. 1985: Blake etal.. 1983:
Heard and Chamberlain. 1982: Rabinowitz et al.. 1980).
Limited evidence demonstrates that GI absorption of water-soluble Pb is higher in
children than in adults. Estimates derived from dietary balance studies conducted in
infants and children (ages 2 weeks to 8 years) indicate that -40-50% of ingested Pb is
absorbed (Ziegler et al.. 1978: Alexander etal.. 1974). Experimental studies provide
further evidence for greater absorption of Pb from the gut in young animals compared to
adult animals (Aungst et al.. 1981: Kostial etal.. 1978: Pounds etal.. 1978: Forbes and
Reina. 1972). The mechanisms for an apparent age difference in GI absorption of Pb have
not been completely elucidated and may include both physiological and dietary factors
(Mushak. 1991). To further investigate the effects of the presence of food in the GI tract
on Pb absorption, children (3-5 years old) who ate breakfast had lower blood Pb levels
compared to children who did not eat breakfast (Liu etal.. 201 la). This difference
3-37
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persisted after controlling for nutritional variables (blood iron [Fe], calcium [Ca2+],
copper [Cu], magnesium [Mg], and zinc [Zn]). This observation may be explained by
lower GI absorption of Pb ingested with or in close temporal proximity to meals. Direct
evidence for meals lowering GI absorption of Pb has also been reported for adults
(Maddaloni et al.. 1998; James et al. 1985).
Nutritional interactions of Pb with dietary elements (e.g., Fe, Ca2+, Zn) are complex. Pb
competes with other elements for transport and binding sites that can result in
adjustments of homeostatic regulators to absorb and retain needed elements.
Additionally, low levels of macronutrients may alter Pb bioaccessibility in the GI tract.
Genetic variation in absorption and metabolism may modify all of the above.
Children who are Fe-deficient have higher blood Pb concentrations than similarly
exposed Fe-replete children, suggesting that Fe deficiency may result in higher Pb
absorption or, possibly, other changes in Pb biokinetics that contribute to altered blood
Pb concentrations (Schell et al., 2004; Marcus and Schwartz. 1987; Mahaffey and
Annest. 1986). Studies conducted in animal models have provided direct evidence for
interactions between Fe deficiency and increased Pb absorption, perhaps by enhancing
binding of Pb to Fe-binding proteins in the intestine (Bannon et al.. 2003; Morrison and
Quarterman. 1987; Barton etal. 1978b). An analysis of data from a sample 448 woman
(age 20-55 years) did not find a significant association between Fe body stores (indicated
from serum ferritin concentration) and blood Pb concentrations, although depleted irons
stores (serum ferritin of <12 (ig/L) was associated with higher blood concentrations of
Cd, cobalt (Co) and manganese (Mn) higher (Meltzer et al., 2010).The effects of Fe
nutritional status on blood Pb include changes in blood Pb concentrations in association
with genetic variation in genes involved in Fe metabolism. For example, genetic variants
in the hemochromatosis (HFE) and transferrin genes are associated with higher blood Pb
concentrations in children (Hopkins et al., 2008). In contrast, HFE gene variants are
associated with lower bone and blood Pb levels in elderly men (Wright et al.. 2004).
Several studies have suggested that dietary Ca2+ may have a protective role against Pb by
decreasing absorption of Pb in the GI tract and by decreasing the mobilization of Pb from
bone stores to blood. In experimental studies of adults, absorption of a single dose of Pb
(100-300 (ig Pb chloride) was lower when the Pb was ingested together with
Ca2+ carbonate (0.2 g Ca2+ carbonate) than when the Pb was ingested without additional
Ca2+ (Blake and Mann. 1983; Heard and Chamberlain. 1982). A similar effect of Ca2+
occurs in rats (Barton et al.. 1978a). Similarly, an inverse relationship was observed
between dietary Ca2+ intake and blood Pb concentration in children, suggesting that
children who are Ca2+-deficient may absorb more Pb than Ca2+-replete children (Elias et
al.. 2007; Schell et al.. 2004; Mahaffev et al.. 1986; Ziegleretal.. 1978). These
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observations suggest that Ca2+ and Pb share and may compete for common binding and
transport mechanisms in the small intestine which are regulated in response to dietary
Ca2+ and Ca2+-body stores (Fullmer and Rosen, 1990; Bronneretal.. 1986). However,
animal studies have also shown that multiple aspects of Pb toxicokinetics are affected by
Ca2+ nutritional status. For example, feeding rats a Ca2+ deficient diet is associated with
increased Pb absorption, decreased whole body Pb clearance, and increased volume of
distribution of Pb (Aungst and Fung, 1985). These studies suggest that associations
between Ca2+ nutrition and blood Pb that have been observed in human populations may
not be solely attributable to effects of Ca2+ nutrition on Pb absorption. Other potential
mechanisms by which Ca2+ nutrition may affect blood Pb and Pb biokinetics include
effects on bone mineral metabolism and renal function.
Blood Pb concentrations in young children have also been shown to increase in
association with lower dietary Zn levels (Schell et al., 2004). Mechanisms for how Zn
affects blood Pb concentration, i.e., whether it involves changes in absorption or changes
in distribution and/or elimination of Pb, have not been determined.
Dissolution of Pb from the soil/mineralogical matrix in the stomach appears to be the
major process that renders soil Pb bioaccessible for absorption in the GI tract. Absorption
of Pb has been shown to vary depending upon the Pb mineralogy and physical
characteristics of the Pb in the soil (e.g., encapsulated or exposed) and size of the
Pb-bearing grains. GI absorption of larger Pb-containing particles (MOO urn) tends to be
lower than smaller particles (Healy etal.. 1992; Barltrop and Meek. 1979). Absorption of
Pb in soils and dust has been most extensively studied in the in vivo swine model. Gastric
function of swine is thought to be sufficiently similar to that of humans to justify use of
swine as a model for assessing factors that may affect GI absorption of Pb from soils in
humans (Juhasz et al. 2009: U.S. EPA. 2007b: Casteel et al.. 2006: Casteel et al.. 1997:
Weis and Lavelle. 1991). Other practical advantages of the swine model over rodent
models have been described, and include: absence of coprophagia; ease with which Pb
dosing can be administered and controlled; and higher absorption fraction of soluble Pb
(e.g., Pb acetate) in swine, which is more similar to humans than rats (Smith et al..
2009a). The swine studies measure blood and/or tissue Pb (e.g., kidney, liver, bone)
concentrations following oral dosing of swine with either soil or with a highly water
soluble and fully bioaccessible form of Pb (e.g., Pb acetate). A comparison of the internal
concentrations of Pb under these two conditions provides a measure of the bioavailability
(i.e., absorption) of Pb in soil relative to that of Pb acetate, which is typically referred to
as relative bioavailability (RBA). Relative bioavailability measured in the swine assay is
equivalent to the ratio of the absorbed fraction (AF) of ingested dose of soil Pb to that of
water-soluble Pb acetate (e.g., RBA = AFSoiiPb/AFpbaCetate).
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Collectively, published studies conducted in swine have provided 39 estimates of Pb
RBA for 38 different soil or "soil-like" test materials (Bannon et al.. 2009; Smith et al..
2009a; Casteel et al.. 2006; Marschner et al., 2006). The mean of RBA estimates from 25
soils is 0.49 (± 0.29[SD]), median is 0.51, and 5th to 95th percentile range is 0.12 to
-0.89. RBA estimates for soils collected from 8 firing ranges were approximately 1.0
(Bannon et al.. 2009). The relatively high RBA for the firing range soils may reflect the
high abundance of relatively un-encapsulated Pb carbonate (30-90% abundance) and Pb
oxide (1-60%) in these soils. Similarly, a soil sample (low Pb concentration) mixed with
a NIST paint standard (55% Pb carbonate, 44% Pb oxide) also had a relatively high
bioavailability (0.72) (Casteel et al., 2006). Samples of smelter slag, or soils in which the
dominant source of Pb was smelter slag, had relatively low RBA (0.14 - 0.40, n = 3), as
did a sample from a mine tailings pile (RBA = 0.06), and a sample of finely ground
galena mixed with soil (Casteel et al.. 2006).
Based on data for 18 soil materials assayed in swine, RBA of Pb mineral phases were
categorized into "low" (<0.25 [25%]), "medium" (0.25-0.75 [25 to 75%]), and "high"
(>0.75 [75%]) categories (Casteel et al.. 2006). Figure 3-3 shows some of the materials
that fall into these three categories. Mineral phases observed in mineralogical wastes can
be expected to change overtime (i.e., weathering), which could change the RBA over
time. The above observations in swine are supported by various studies conducted in rats
that have found RBA of Pb in soils to vary considerably and to be less than 1.0 (Smith et
al.. 2009a. 2008; Freeman et al.. 1996; Freeman et al.. 1994; Freeman et al.. 1992).
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Group
Note: based on results from juvenile swine assays.
Source: Casteel et al. (2006).
Figure 3-3 Estimated relative bioavailability (RBA, compared to Pb acetate)
of ingested Pb in mineral groups.
Drexler and Brattin (2007) developed an in vitro bioaccessibility (IVBA) assay for soil
Pb that utilizes extraction fluid comprised of glycine, deionized water, and hydrochloric
acid at a pH of 1.50 that is combined with sieved test material (<250 urn) for 1 hour. The
assay was tested for predicting in vivo RBA of 18 soil-like test materials that were
assayed in a juvenile swine assay (Casteel et al., 2006). A regression model relating
IVBA and RBA was derived based on these data (Equation 3-1):
RBA = (0.878 X IVBA) - 0.028
Equation 3-1
where RBA and IVBA are expressed as fractions (i.e., not as percent). The weighted r2
for the relationship (weighted for error in the IVBA and RBA estimates) was 0.924
(p <0.001). The IVBA assay reported in Drexler and Brattin (2007) has been identified by
the U.S. EPA as a validated method for predicting RBA of Pb in soils for use in risk
3-41
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assessment (U.S. EPA. 2007e). A review of soil Pb RBA estimates made using the IVBA
assay described above and Equation 3-1 identified 270 estimates of Pb RBA in soils
obtained from 11 hazardous waste sites. The mean for the site-wide RBA estimates
(n = 11 sites) was 0.57 (SD 0.15), median was 0.63, and 5th to 95th percentile range was
0.34 to 0.71.
Equation 3-1 cannot be reliably extrapolated to other in vitro assays that have been
developed for estimating Pb bioaccessibility without validation against in vivo RBA
measurements made on the same test materials. Comparisons of outcomes among
different in vitro assays applied to the same soil test materials have found considerable
variability in IVBA estimates (Juhaszetal.. 2011; Smith etal.. 2011; Saikat et al.. 2007;
Van de Wiele et al.. 2007). This variability has been attributed to differences in assay
conditions, including pH, liquid:soil ratios, inclusion or absence of food material, and
differences in methods used to separate dissolved and particle-bound Pb
(e.g., centrifugation versus filtration). Smith et al. (2011) found that algorithms for
predicting RBA, based on two different IVBA assays, did not yield similar predictions of
RBA when applied to the same material. Given the dependence of IVBA outcomes on
assay conditions, in vitro assays used to predict in vivo RBA should be evaluated against
in vivo RBA estimates to quantitatively assess uncertainty in RBA predictions (U.S.
EPA. 2007e).
Absorption of Pb in house dusts has not been rigorously evaluated quantitatively in
humans or in experimental animal models. The RBA for paint Pb mixed with soil was
reported to be approximately 0.72 (95% CI: 0.44, 0.98) in juvenile swine, suggesting that
paint Pb dust reaching the gastrointestinal tract maybe highly bioavailable (Casteel et al..
2006). The same material yielded a bioaccessibility value (based on IVBA assay) of 0.75
(Drexler and Brattin. 2007). which corresponds to a predicted RBA of 0.63, based on
Equation 3-1. A review of indoor Pb RBA estimates made using the IVBA assay and
Equation 3-1 identified 100 estimates of Pb RBA in dusts obtained from two hazardous
waste sites. Mean Pb RBAs for the Herculaneum site were 0.47 (SD 0.07, 10 samples)
for indoor dust and 0.69 (SD 0.03, 12 samples) for soil. At the Omaha site, mean Pb
RBAs were 0.73 (SD 0.10, 90 samples) for indoor dust and 0.70 (SD 0.10, 45 samples)
for soil. Yu et al. (2006) applied an IVBA method to estimate bioaccessibility of Pb in
house dust samples collected from 15 urban homes. Homes were selected for inclusion in
this study based on reporting to the state department of health of at least on child with a
blood Pb concentration >15 (ig/dL and Pb paint dust may have contributed to indoor dust
Pb. The mean IVBA was 0.65 (SD 0.08, age: 52.5 to 77.2 months).
The above results, and the IVBA assays used in studies of interior dust, have not been
evaluated against in vivo RBA estimates for dust samples. Although, expectations are
3-42
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that a validated IVBA methodology for soil would perform well for predicting RBA of
interior dust, this validation has not actually been experimentally confirmed. Factors that
may affect in vivo predictions of RBA of interior dust Pb could include particle size
distribution of interior dust Pb and the composition of the dust matrix, which may be
quite different from that of soil.
Other estimates of bioaccessibility of Pb in house dusts have been reported, based on
results from in vitro extraction assays that have not been validated for predicting in vivo
bioavailability. Bioaccessibility assays that sequentially extract soil at gastric pH
followed by intestinal pH tend to show higher bioaccessibility of soil and dust Pb when
incubated at gastric conditions (Juhasz et al.. 2011; Lu et al.. 2011; Smith et al.. 2011;
Rousseletal.. 2010; Yu et al., 2006). Yu et al. (2006) dissolved Pb dust, obtained from
vacuuming carpet samples into simulated gastric and intestinal acids (also
Section 3.1.3.2). The carpet samples were obtained from homes located in northern
New Jersey. Pb concentration in carpet ranged from 209 to 1,770 mg/kg dust, with
52-77% of Pb dissolving in simulated gastric acid and 5-32% dissolving in simulated
intestinal acids. In a similar test in the U.K., Turner and Simmonds (2006) observed
median Pb dust concentrations of 178 mg/kg with approximately 80% bioaccessibility in
simulated gastric acid. Jin et al. (2005) observed that bioaccessibility of Pb in soil was
proportional to the soil acidity and organic matter content of the soil.
3.2.2 Distribution and Metabolism
A simple conceptual representation of Pb distribution is that it contains a fast turnover
pool, comprising mainly soft tissue, and a slow pool, comprising mainly skeletal tissues
(Rabinowitz et al., 1976). The highest soft tissue concentrations in adults occur in liver
and kidney cortex (Gerhardsson et al.. 1995; Oldereid et al.. 1993; Gerhardsson et al..
1986; Barry. 1975; Gross et al., 1975). Pb in blood (i.e., plasma) exchanges with both of
these compartments.
3.2.2.1 Blood
Blood comprises -1% of total Pb body burden. Pb in blood is found primarily (>99%) in
the RBCs (Smith et al.. 2002; Manton etal. 2001; Bergdahl et al.. 1999; Bergdahl et al..
1998; Hernandez-Avila et al.. 1998; Bergdahl et al.. 1997a; Schutzetal. 1996).
5-aminolevulinic acid dehydratase (ALAD) is the primary binding ligand for Pb in
erythrocytes (Bergdahl et al.. 1998; Xieetal.. 1998; Bergdahl et al.. 1997a; Sakai et al..
1982). Two other Pb-binding proteins have been identified in the RBC, a 45 kDa protein
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(Kmax 700 (ig/dL; Kd 5.5 (ig/L) and a smaller protein band having a molecular weight of
<10 kDa (Bergdahl et al.. 1998: Bergdahl et al.. 1997a: Bergdahl et al.. 1996). Of the
three principal Pb-binding proteins identified in RBCs, ALAD has the strongest affinity
for Pb (Bergdahl etal.. 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 et al.. 1998;
Bergdahl et al.. 1996; Sakaietal.. 1982). Pb binding to ALAD is saturable; the binding
capacity was estimated to be -850 (ig/dL RBCs (or ~40 (ig/dL whole blood) and the
apparent dissociation constant has been estimated to be -1.5 (ig/L (Bergdahl et al.. 1998).
Hematocrit is somewhat higher in the neonate at birth (51%) than in later infancy (35% at
6 months), which may lead to a decrease in the total binding capacity of blood over the
first 6 months of life that results in a redistribution of Pb among other tissues (Simon et
al.. 2007).
Saturable binding to RBC proteins contributes to an increase in the plasma/blood Pb ratio
with increasing blood Pb concentration and curvature to the blood Pb-plasma Pb
relationship (Rentschler etal.. 2012; Kang et al., 2009; Jin et al., 2008; Barbosaet al.,
2006a; Smith et al.. 2002; Manton etal.. 2001: Bergdahl et al.. 1999; Bergdahl et al..
1998; Bergdahl et al.. 1997b; DeSilva. 1981). An example of this is shown in Figure 3-4.
Saturable binding of Pb to RBC proteins has several important consequences. As blood
Pb increases and the higher affinity binding sites for Pb in RBCs become saturated, a
larger fraction of the blood Pb is available in plasma to distribute to brain and other
Pb-responsive tissues. This change in distribution of Pb contributes to a curvature in the
relationship between Pb intake (at constant absorption fraction) and blood Pb
concentration. Plasma Pb also exhibits faster kinetics. Following exposures of 5 adults
that resulted in relatively high blood Pb concentrations (56-110 (ig/dL), the initial (fast-
phase) elimination half-time for plasma Pb (38 ± 20 [SD] days) was approximately half
that of blood (81 ± 25 days) (Rentschler etal.. 2012).
Typically, at blood Pb concentrations <100 (ig/dL, only a small fraction (<1%) of blood
Pb is found in plasma (Marcus. 1985; Manton and Cook. 1984; DeSilva. 1981). However,
as previously noted, plasma Pb may be the more biologically labile and lexicologically
effective fraction of the circulating Pb. Approximately 40-75% of Pb in the plasma is
bound to proteins, of which albumin appears to be the dominant ligand (Al-Modhefer et
al.. 1991; Ong and Lee. 1980a). Pb in serum that is not bound to protein exists largely as
complexes with low molecular weight sulfhydryl compounds (e.g., cysteine,
homocysteine) and other ligands (Al-Modhefer et al.. 1991).
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oAdults • Children
0
20 40 60
Blood Pb (Mg/dL)
80
100
Source: Adapted with permission of Elsevier Publishing and the Finland Institute of Occupational Health, Bergdahl et al. (1999:
1997b).
Figure 3-4 Plot of blood and plasma Pb concentrations measured in adults
and children.
As shown in Figure 3-4. the limited binding capacity of Pb binding proteins in RBCs
produces a curvilinear relationship between blood and plasma Pb concentration. The
limited binding capacity of RBC binding proteins also confers, or at least contributes, to a
curvilinear relationship between Pb intake and blood Pb concentration. A curvilinear
relationship between Pb intake and blood Pb concentration has been observed in children
(Sherlock and Quinn. 1986; Lacey et al., 1985; Ryu et al.. 1983). As shown in Figure 3-5.
the relationship becomes pseudo-linear at relatively low daily Pb intakes
(i.e., <10 ug/kg per day) and at blood Pb concentrations <25 (ig/dL.
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O)
.0
O
O
CO
0
100 200 300
Pb Intake (|jg/day)
400
Data represent mean and standard errors for intake; the line is the regression model (blood Pb = 3.9 + 2.43 (Pb intake [|jg/week]1'3).
Source: Adapted with permission of Taylor & Francis Publishing, Sherlock and Quinn (1986).
Figure 3-5 Relationship between Pb intake and blood Pb concentration in
infants (N = 105, age 13 weeks, formula-fed).
Figure 3-6 shows the predicted relationship between quasi-steady state blood and plasma
Pb concentrations in a 4-year old child using the ICRP model [(Pounds and Leggett.
1998; ICRP. 1994; Leggett. 1993). see Section 33, for a brief description of the ICRP
model]. The abrupt inflection point that occurs at approximately 25 (ig/dL blood Pb is an
artifact of the numerical approach to simulate the saturation of binding using
discontinuous first-order rate constants for uptake and exit of Pb from the RBC. A
continuous function of binding sites and affinity, using empirical estimates of both
parameters, yield a similar but continuous curvature in the relationship (Bergdahl et al..
1998; O'Flaherty. 1995). Nevertheless, either approach predicts a pseudo-linear
relationship at blood Pb concentrations below approximately 25 (ig/dL which, in this
model, corresponds to an intake of approximately 100 (ig/day (absorption rate
~ 30 (ig/day) (upper panel). An important consequence of the limited Pb-binding capacity
of RBC proteins is that the plasma Pb concentration will continue to grow at a linear rate
above the saturation point for RBC protein binding. One implication of limited binding
capacity is that a larger fraction of the Pb in blood will become available to distribute to
brain and other Pb-responsive tissues as blood Pb increases. This process could
potentially contribute to non-linearity in dose-response relationships in studies in which
blood Pb is the used as the internal dose metric.
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0.60
0.50 -
3) 0.40 -
°- 0.30 -
ra
E
ra 0.20 -
o.
0.10 -
0.00
10 20 30 40
Blood Pb (ng/dL)
50
60
50 -
IT
1) 40 -
^
si
Ł 30 H
o
m 20 -
10 -
Blood
•Plasma
1.0
0.8
0)
tn
0.6 3
0)
TJ
0-4
0.2
0.0
100 200 300
Intake (tig/day)
400
Note: Model simulations are for a 4-year old having from birth a constant Pb intake of between 1 and 400 ug/day. Simulation based
on ICRP Pb biokinetics model (Leggett. 1993).
Figure 3-6 Simulation of quasi-steady state blood and plasma Pb
concentrations in a child (age 4 years) associated with varying Pb
ingestion rates.
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Studies conducted in swine provide additional evidence in support of RBC binding
kinetics influencing distribution of Pb to tissues. In these studies, the relationship
between the ingested dose of Pb and tissue Pb concentrations (e.g., liver, kidney, bone)
was linear, whereas, the relationship between dose and blood Pb was curvilinear with the
slope decreasing as the dose increased (Casteel et al.. 2006). Saturable binding of Pb to
RBC proteins also contributes to a curvilinear relationship between blood Pb and both
plasma Pb and urinary Pb, whereas Pb in plasma and urine are linearly related (Bergdahl
etal.. 1997b).
3.2.2.2 Bone
The dominant compartment for Pb in the body is in bone. In human adults, more than
90% of the total body burden of Pb is found in the bones, whereas bone Pb accounts for
just under 60% of the body burden in infants less than a year old and just over 70% of the
body burden in older children (Barry. 1975). Bone is comprised of two main types,
cortical (or compact) and trabecular (or spongy or cancellous). The proportion of cortical
to trabecular bone in the human body varies by age, but on average is about 80 to 20
percent (O'Flahertv. 1998; Leggett. 1993; ICRP. 1973).
The exchange of Pb from plasma to the bone surface is a rapid process (i.e., adult
ti/2 =0.19 and 0.23 hours for trabecular and cortical bone, respectively) (Leggett. 1993).
Some Pb diffuses from the bone surface to deeper bone regions (adult ti/2 = 150 days)
where it is relatively inert (in adults) and part of a "nonexchangeable" (removed only
through bone resorption/remodeling) pool of Pb in bone (Leggett. 1993).
Pb distribution in bone includes uptake into cells that populate bone (e.g., osteoblasts,
osteoclasts, osteocytes) and exchanges with proteins and minerals in the extracellular
matrix (Pounds etal.. 1991). Pb forms highly stable complexes with phosphate and can
replace calcium in the calcium-phosphate salt, hydroxyapatite, which comprises the
primary crystalline matrix of bone (Meirer et al.. 2011; Bres etal., 1986; Miyake. 1986;
Verbeeck et al.. 1981). Several intracellular kinetic pools of Pb have been described in
isolated cultures of osteoblasts and osteoclasts which appear to reflect physiological
compartmentalization within the cell, including membranes, mitochondria, soluble
intracellular binding proteins, mineralized Pb (i.e., hydroxyapatite) and inclusion bodies
(Long etal.. 1990; Pounds and Rosen. 1986; Rosen. 1983V Approximately 70-80% of Pb
taken up into isolated primary cultures of osteoblasts or osteocytes is associated with
mitochondria and mineralized Pb (Pounds et al.. 1991).
Pb accumulates in bone regions having the most active calcification at the time of
exposure. Pb accumulation is thought to occur predominantly in trabecular bone during
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childhood and in both cortical and trabecular bone in adulthood (Aufderheide and
Wittmers. 1992). Early Pb uptake in children is greater in trabecular bone due to its larger
surface area and higher metabolic rate. With continued exposure, Pb concentrations in
bone may increase with age throughout the lifetime beginning in childhood, indicative of
a relatively slow turnover of Pb in adult bone (Park et al.. 2009b; Barry and Connolly.
1981: Barry. 1975: Gross etal.. 1975: Schroeder and Tipton. 1968V The cortical and
trabecular bones have distinct rates of turnover and Pb release. For example, tibia has a
turnover rate of about 2% per year whereas trabecular bone has a turnover rate of more
than 8% per year in adults (Rabinowitz. 1991).
A high bone formation rate in early childhood results in the rapid uptake of circulating Pb
into mineralizing bone; however, bone Pb is also recycled to other tissue compartments
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 (60-65%)
(O'Flahertv. 1995: Leggett. 1993: ICRP. 1973). However, some Pb accumulated in bone
does persist into later life. McNeill et al. (2000) compared tibia Pb levels and cumulative
blood Pb indices in a population of 19- to 29-year-olds who had been highly exposed to
Pb in childhood from the Bunker Hill, Idaho smelter; they concluded that Pb from
exposure in early childhood had persisted in the bone matrix until adulthood.
Additional discussion of the Pb in bone and its mobilization are provided in other
sections of this chapter. Maternal mobilization of Pb from the bone to the fetus is
discussed in Section 3.2.2.4. The relationship between Pb in blood and bone is discussed
in Section 3.3.5.
3.2.2.3 Soft Tissues
Most of the Pb in soft tissue is in liver and kidney (Gerhardsson et al.. 1995; Oldereid et
al.. 1993: Gerhardsson et al.. 1986: Barry. 1981. 1975: Gross etal.. 1975). Presumably,
the Pb in these soft tissues (i.e., kidney, liver, and brain) exists predominantly bound to
protein. High affinity cytosolic Pb-binding proteins have been identified in rat kidney and
brain (DuVal and Fowler. 1989; Fowler. 1989). The Pb-binding proteins in rat are
cleavage products of a2\a globulin, a member of the protein superfamily known as
retinol-binding proteins that are generally observed only in male rats (Fowler and DuVal.
1991). Other high-affinity Pb-binding proteins (Kd ~14 nM) have been isolated in human
kidney, two of which have been identified as a 5 kDa peptide, thymosin 4 and a 9 kDa
peptide, acyl-CoA binding protein (Smith etal.. 1998). Pb also binds to metallothionein,
but does not appear to be a significant inducer of the protein in comparison with the
inducers Cd and Zn (Waalkes and Klaassen. 1985: Eaton et al.. 1980).
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The liver and kidneys rapidly accumulate systemic Pb (ti/2 = 0.21 and 0.41 hours,
respectively), which amounts to 10-15% and 15-20% of intravenously injected Pb,
respectively (Leggett 1993). A linear relationship in dose-tissue Pb concentrations for
kidney and liver has been demonstrated in swine, dogs, and rats (Smith et al.. 2008;
Casteel et al.. 2006; Casteel et al., 1997; Azaretal.. 1973). 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 (Treble and Thompson.
1997; Barry. 1975). reflecting a faster turnover of Pb in soft tissue relative to bone.
3.2.2.4 Fetus
Evidence for maternal-to-fetal transfer of Pb in humans is derived from cord blood Pb to
maternal blood Pb ratios (i.e., cord blood Pb concentration divided by mother's blood Pb
concentration). Group mean ratios range from about 0.7 to 1.0 at the time of delivery for
mean maternal blood Pb levels ranging from 1.7 to 8.6 (ig/dL (Amaral et al.. 2010;
Kordas et al.. 2009; Patel and Prabhu. 2009; Carbone et al.. 1998; Gover. 1990; Graziano
et al.. 1990). The relationship for individual mother-child pairs is variable, but well
correlated (Pearson r = 0.79); in a predominantly young, low income, urban population
(N = 159), factors associated with higher cord compared to maternal blood Pb level
included maternal elevated blood pressure and alcohol consumption, while factors
associated with relatively lower ratios of cord blood Pb to maternal blood Pb included
maternal increased hemoglobin levels and sickle cell trait (Harville et al.. 2005). Calcium
intake and physical activity were not associated with differences between cord Pb and
maternal blood Pb. Consistent with other studies, the ratio of mean cord blood Pb (1.64
ug/dL) to mean maternal blood Pb (1.93 ug/dL) was 0.85. The similarity of isotopic
ratios in maternal blood and in blood and urine of newly-born infants provides further
evidence for placental transfer of Pb to the fetus (Gulson et al.. 1999).
Transplacental transfer of Pb may be facilitated by an increase in the plasma/blood Pb
concentration ratio during pregnancy (Montenegro et al.. 2008; Lamadrid-Figueroa et al..
2006). Maternal-to-fetal transfer of Pb appears to be related partly to the mobilization of
Pb from the maternal skeleton. Evidence for transfer of maternal bone Pb to the fetus has
been provided by stable Pb isotope studies in cynomolgus monkeys exposed during
pregnancy. Approximately 7-39% of the maternal Pb burden transferred to the fetus was
derived from the maternal skeleton, with the remainder derived from contemporaneous
exposure (O'Flaherty. 1998; Franklin et al.. 1997). The upper value in the range (39%)
represented the one monkey with historical Pb exposure, but received only small amounts
of environmental Pb exposure during pregnancy; for the monkeys that received high
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doses of Pb during pregnancy, the range was lower (7-25%) (O'Flaherty. 1998; Franklin
etal.. 1997).
3.2.2.5 Organic Pb
Information on the distribution of Pb in humans following exposures to organic Pb is
extremely limited. However, as reported in the 2006 Pb AQCD (U.S. EPA, 2006c). the
available evidence demonstrates near complete absorption following inhalation of
tetraalkyl Pb vapor and subsequent transformation to trialkyl Pb metabolites. One hour
following brief inhalation exposures to 203Pb tetraethyl or tetramethyl Pb (1 mg/m3),
-50% of the 203Pb body burden was associated with liver and 5% with kidney; the
remaining 203Pb was widely distributed throughout the body (Heard etal.. 1979). The
kinetics of 203Pb in blood showed an initial declining phase during the first 4 hours
(tetramethyl Pb) or 10 hours (tetraethyl Pb) after the exposure, followed by a
reappearance of radioactivity back into the blood after ~20 hours. The high level of
radioactivity initially in the plasma indicates the presence of tetraalkyl/trialkyl Pb. The
subsequent rise in blood radioactivity, however, probably represents water-soluble
inorganic Pb and trialkyl and dialkyl Pb compounds that were formed from the metabolic
conversion of the volatile parent compounds (Heard etal.. 1979).
Alkyl Pb compounds undergo oxidative dealkylation catalyzed by cytochrome P450 in
liver and, possibly, in other tissues. Trialkyl Pb metabolites have been found in the liver,
kidney, and brain following exposure to the tetraalkyl compounds in workers
(Bolanowska et al.. 1967); these metabolites have also been detected in brain tissue of
nonoccupational subjects (Nielsen et al.. 1978).
3.2.3 Elimination
The rapid-phase (30-40 days) of Pb excretion amounts to 50-60% of the absorbed
fraction (Chamberlain et al.. 1978; Rabinowitz et al.. 1976; Kehoe. 196la. b, c).
Absorbed Pb is excreted primarily in urine and feces, with sweat, saliva, hair, nails, and
breast milk being minor routes of excretion (Kehoe. 1987; Chamberlain et al.. 1978;
Rabinowitz et al.. 1976; Griffin et al.. 1975; Hurshetal.. 1969; Hursh and Suomela.
1968).
Approximately 30% of intravenously injected Pb in humans (40-50% in beagles and
baboons) is excreted via urine and feces during the first 20 days following administration
(Leggett 1993). The kinetics of urinary excretion following a single dose of Pb is similar
to that of blood (Chamberlain et al.. 1978). likely due to the fact that Pb in urine derives
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largely from Pb in plasma. Evidence for this is the observation that urinary Pb excretion
is strongly correlated with the rate of glomerular filtration of Pb (Araki et al.. 1986) and
plasma Pb concentration (Rentschler et al., 2012; Bergdahl et al., 1997b) (i.e., glomerular
filtration rate x plasma Pb concentration), and both relationships are linear. While the
relationship between urinary Pb excretion and plasma Pb concentration is linear, the
plasma Pb relationship to blood Pb concentration is curvilinear (as described in Section
3.2.2.1 and demonstrated in Figure 3-6). This relationship contributes to an increase in
the renal clearance of Pb from blood with increasing blood Pb concentrations
(Chamberlain. 1983). Similarly, a linear relationship between plasma Pb concentration
and urinary excretion rate predicts a linear relationship between Pb intake (at constant
absorption fraction) and urinary Pb excretion rate, whereas the relationship with blood Pb
concentration would be expected to be curvilinear (Section 3.3.6).
Estimates of urinary filtration of Pb from plasma range from 13-22 L/day, with a mean of
18 L/day (Araki etal.. 1986: Manton and Cook. 1984: Manton and Malloy. 1983:
Chamberlain et al.. 1978). which corresponds to half-time for transfer of Pb from plasma
to urine of 0.1-0.16 days for a 70-kg adult who has a plasma volume of ~3 L. The rate of
urinary excretion of Pb was less than the rate of glomerular filtration of ultrafilterable Pb,
suggesting that urinary Pb is the result of incomplete renal tubular reabsorption of Pb in
the glomerular filtrate (Araki etal.. 1986): although, net tubular secretion of Pb has been
demonstrated in animals (Victery etal.. 1979: Vander et al.. 1977). On the other hand,
estimates of blood-to-urine clearance range from 0.03-0.3 L/day with a mean of
0.18 L/day (Diamond. 1992: Araki etal.. 1990: Bergeretal.. 1990: Kosteretal. 1989:
Manton and Malloy, 1983: Ryu etal.. 1983: Chamberlain et al.. 1978: Rabinowitz et al..
1973). consistent with a plasma Pb to blood Pb concentration ratio of-0.005-0.01 L/day
(Klotzback et al.. 2003). Based on the above differences, urinary excretion of Pb can be
expected to reflect the concentration of Pb in plasma and variables that affect delivery of
Pb from plasma to urine (e.g., glomerular filtration and other transfer processes in the
kidney).
The value for fecal:urinary excretion ratio (-0.5) was observed during days 2-14
following intravenous injection of Pb in humans (Chamberlain et al.. 1978: Booker et al..
1969: Hurshetal.. 1969). This ratio is slightly higher (0.7-0.8) with inhalation of
submicron Pb-bearing PM due to ciliary clearance and subsequent ingestion. The transfer
of Pb from blood plasma to the small intestine by biliary secretion in the liver is rapid
(adult ti/2 = 10 days), and accounts for 70% of the total plasma clearance (O'Flaherty.
1995).
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Organic Pb
Pb absorbed after inhalation of tetraethyl and tetramethyl Pb is excreted in exhaled air,
urine, and feces (Heard et al.. 1979). Fecal:urinary excretion ratios were 1.8 following
exposure to tetraethyl Pb and 1.0 following exposure to tetramethyl Pb (Heard et al.,
1979). Occupational monitoring studies of workers exposed to tetraethyl Pb showed that
tetraethyl Pb is excreted in the urine as diethyl Pb, ethyl Pb, and inorganic Pb (Vural and
Duydu. 1995; Zhang et al.. 1994; Turlakiewicz and Chmielnicka. 1985).
3.3 Pb Biomarkers
This section describes the biological measurements of Pb and their interpretation as
indicators of exposure or body burden.
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., instantaneous, cumulative over lifetime, or cumulative over a circumscribed age
range). In recent studies of Pb and health, the exposure biomarkers most frequently used
are Pb in blood and bone. 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.
As an exposure biomarker, blood Pb concentration has other limitations. Only about 1%
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 ALAD, with the balance in plasma. From a toxicological perspective, the
unbound fraction is likely to be the most important subcompartment of blood Pb because
it distributes into soft tissues. The concentration of Pb in plasma is much lower than in
whole blood (<1%). The greater relative abundance of Pb in whole blood makes its
measurement much easier (and more affordable) than 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.
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Moreover, binding Pb in red blood cells 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.
Another limitation of blood Pb as an exposure biomarker is that the kinetics of Pb in
blood is relatively fast compared to the kinetics of Pb in bone, and therefore, of the whole
body burden. Thus, a high blood Pb concentration measured at any given time 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 changes with time/age depends on exposure history due to re-
equilibration of Pb stored in the various body pools.
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 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 reflects a long exposure averaging time.
Mechanistic models are used throughout the section as a means to describe basic
concepts that derive from the wealth of information on Pb toxicokinetics. Although
predictions from models are inherently uncertain, models can serve to illustrate expected
interrelationships between Pb intake and tissue distribution that are important in
interpreting human clinical and epidemiologic studies. Thus, models serve as the only
means available for synthesizing the extensive, but incomplete, knowledge of Pb
biokinetics into a holistic representation of Pb biokinetics. Furthermore, models can also
be used to make predictions about biokinetics relationships that have not been thoroughly
evaluated in experiments or epidemiologic studies. In this way, models can serve as
heuristic tools for shaping data collection to improve understanding of Pb biokinetics.
Mechanistic toxicokinetics models can make predictions about hypothetical populations
and exposure scenarios. When a model is run as a single simulation, the output represents
average outcomes from what is in reality a distribution of possible outcomes that would
be expected in the population (or in any single individual) where intra-individual and
inter-individual variability in exposure and toxicokinetics exist. More realistic predictions
for the population can be developed by running a series of model simulations in which
ranges (i.e., distributions) of parameter values are considered that may better represent
the population of interest. In this section, only single simulations are used to demonstrate
relationships between various biomarkers (e.g., blood Pb and bone Pb) that would apply
to a population having "typical" or "average" exposure and toxicokinetics. These single
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simulations are used for illustrative purposes to describe general concepts and patterns.
Variability would be expected in real populations.
Numerous mechanistic models of Pb biokinetics in humans have been proposed, and
these are described in the 2006 Pb AQCD (U.S. EPA. 2006b) and in the supporting
literature cited in that report. In this section, for simplicity and for internal consistency,
discussion is limited to predictions from a single model, the ICRP Pb biokinetics model
(Pounds and Leggett. 1998: ICRP. 1994: Leggett. 1993). The ICRP model consists of a
systemic biokinetics model (Leggett, 1993) and a human respiratory tract model (ICRP.
1994). The Leggett model simulates age-dependent kinetics of tissue distribution and
excretion of Pb ingestion and inhalation intakes. This model was originally developed for
the purpose of supporting radiation dosimetry predictions and it has been used to develop
cancer risk coefficients for internal radiation exposures to Pb and other alkaline earth
elements that have biokinetics similar to those of calcium (ICRP. 1993). Although the
ICRP model has not been validated by U.S. EPA as a regulatory model for Pb risk
assessment, it has been applied in Pb risk assessment (Abrahams et al.. 2006: Lorenzana
et al.. 2005: Khoury and Diamond. 2003). Portions of the model have been incorporated
into an AALM that is being developed by EPA (2005a). In addition to the above
considerations regarding previous applications of the ICRP model, the model was
selected for use in the ISA because it has several useful features for predicting exposure-
body burden relationships. The model simulates blood Pb and tissue Pb concentration
dynamics associated with the uptake and elimination phases of exposures of > 1 day in
duration; and it simulates age-dependent and particle size-dependent deposition and
clearance of inhaled Pb in the respiratory tract. These types of simulations can only be
approximated with the U.S. EPA IEUBK Model for Pb in children because it simulates
exposures in time steps of 1 year (i.e., age-year average exposures); lumps the simulation
of deposition, mechanical clearance, and absorption of inhaled Pb into a single absorption
term representing the combined processes of gastrointestinal and respiratory tract
absorption of inhaled Pb; simulates steady state blood Pb concentrations and was does
not allow access to the underlying simulations of tissue Pb concentrations which serve as
intermediate variables in the model for predicting steady state blood Pb concentrations.
Other models have been developed that allow simulations of tissue Pb concentrations
(e.g., O'Flahertv. 1995: Leggett. 1993) and comparisons of these models have been
previously described (Maddaloni et al.. 2005).
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Pb biokinetics in adolescents is poorly characterized by all existing Pb biokinetics
models. Individuals undergo rapid changes in sexual development, growth, food and
water intake, bone growth and turnover, behavior, etc. during adolescence. There is a
paucity of experimental measurements of Pb biomarkers during this time developmental
window. The individual biological and kinetic parameters for adolescents are largely
interpolated rather than based on solid experimental and toxicological measurements.
These deficiencies limit the validity of model predictions in this age group.
3.3.1 Bone-Pb Measurements
For Pb measurements in bone, the most commonly examined 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. 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. Recent studies favor measurement of the
patella for estimating trabecular bone Pb, because it has more bone mass and may afford
better measurement precision than the calcaneus.
Bone Pb measurements are typically expressed in units of ug Pb per g bone mineral. This
convention may potentially introduce variability into the bone Pb measurements related
to variation in bone density. Typically, potential associations between bone density and
bone Pb concentration are not evaluated in epidemiologic studies (Theppeang et al..
2008a; Hu et al., 2007a). An important consequence of expressing bone Pb measures
relative to bone mineral content is that lower bone mineral density is associated with
greater measurement uncertainty in bone Pb. This uncertainty can have important
implications for studies in older women for whom low bone mineral density is more
common than in other populations including men and younger adults.
Methods of direct analysis of bone tissue samples include flame atomic absorption
spectrometry (AAS), anode stripping voltammetry (ASV), inductively coupled plasma
atomic emission spectroscopy (ICP-AES), inductively coupled plasma mass spectrometry
(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). Non-invasive, in vivo measurements of bone Pb is achieved with XRF. The
upsurge in popularity of the XRF method has paralleled a decline in the use of the other
methods. More information on the precision, accuracy, and variability in bone Pb
measurements can be found in the 2006 Pb AQCD (U.S. EPA. 2006b).
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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 the most
widely used, as there have been relatively few developments in L-shell devices since the
early 1990s. However, Nie et al. (2011 a) recently reported on the use of a new portable
L-shell device for human in vivo Pb measurements. Advances in L-shell device
technology resulted in much higher sensitivity than previous L-shell devices. The new
L-shell device showed sensitivity similar to that of K-shell methods (detection limit was
approximately 8 (ig/g bone mineral with 2 mm of soft tissue overlay targeted bone) and a
high correlation with results obtained from K-shell methods (intraclass
correlation = 0.65). Behinaen et al. (2011) described application of a 4-detector system
("clover leaf array ") for the K-shell method that provided higher precision and lower
minimum detection limits (MDL) for tibia and calcaneus Pb measurements (3.25 and
4.78 (ig/g bone mineral, respectively) compared to measurements made with single
detectors (8-12 (ig/g and 14-15 (ig/g, respectively).
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., (Todd et al.. 2002: Toddetal.. 2001: Aro et al.. 2000: Toddetal.. 2000)1. 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 overtime. 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 4!/2 month period, Todd et al. (2000) also 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.
In the epidemiologic 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 former 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 standard deviation (SD) of the background; or 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, as
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demonstrated by Bellinger et al. Q994a), Gerhardsson et al. (1993). and Christoffersson
et al. (1986).
With the epidemiologic approach, all values are used (including negative values) to
determine the minimum detection limit of an instrument that results in extremely low
detection limits. Rosen and Pounds (1998) noted that this approach yields population
bone Pb averages that were artificially low. However, not including values that are
negative or below the detection limit, or assigning these values a fixed number is also of
concern. Using the epidemiologic approach of retaining all point estimates of measured
bone Pb concentrations provided the least amount of bias and the greatest efficiency in
comparing the mean or median levels of bone Pb of different populations (Kim et al..
1995).
3.3.2 Blood-Pb Measurements
Analytical methods for measuring Pb in blood include AAS, graphite furnace atomic
absorption spectrometry (GFAAS), ASV, ICP-AES, and ICP-MS. GFAAS and ASV are
generally considered to be the methods of choice (Flegal and Smith. 1995). Limits of
detection for Pb using AAS are on the order of 5-10 (ig/dL for flame AAS measurements
and approximately 0.1 (ig/dL for flameless AAS measurements (Flegal and Smith. 1995;
NIOSH. 1994). A detection limit of 0.005 (ig/dL has been achieved for Pb in blood
samples analyzed by GFAAS.
For measurement of Pb in plasma, ICP-MS provides sufficient sensitivity (Schutz et al..
1996). While the technique has been applied to assessing Pb exposures in adults, ICP-MS
has not received widespread use in epidemiologic studies. Maternal blood and cord blood
samples have been utilized to assess blood Pb concentrations of neonates (e.g., Amaral et
al..2010;Carboneetal.. 1998).
The primary binding ligand for Pb in RBC, ALAD, is encoded by a single gene in
humans that is polymorphic in two alleles (ALAD1 and ALAD2) (Scinicariello et al..
2007). Since the ALAD1 and ALAD 2 alleles can be co-dominantly expressed, 3
different genotypes (ALAD 1-1, ALAD 1-2, and ALAD 2-2) are possible. The ALAD
1-1 genotype is the most common. Scinicariello et al. (2010) tested genotypes in civilian,
non-institutionalized U.S. individuals that participated as part of NHANES III from
1988-1994 and found that 15.6% of non-Hispanic whites, 2.6% non-Hispanic blacks, and
8.8% Mexican Americans carried the ALAD2 allele.
The 2006 Pb AQCD (U.S. EPA. 2006c) reported that many studies have shown that, with
similar exposures to Pb, individuals with the ALAD-2 allele have higher blood Pb levels
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than those without (Kim et al.. 2004; Perez-Bravo et al.. 2004; Bergdahl et al.. 1997b;
Smith etal.. 1995a; Wetmur. 1994; Wettnur et al.. 199 Ib; Astrinetal.. 1987). More
recent meta analyses provide further support for ALAD2 carriers having higher blood Pb
levels than ALAD1-1 homozygotes (Scinicariello et al.. 2007; Zhao et al.. 2007). The
mechanism for this association may be higher Pb binding affinity of ALAD2. Although,
this interpretation would be consistent with the structural differences that result in greater
electronegativity of ALAD1 compared to ALAD2 (Wetmur. 1994; Wetmur etal.. 199 la).
measurements of Pb binding affinity to ALAD1 and ALAD2 (i.e., Pb2+ displacement of
Zn2+ binding to recombinant ALAD1 and ALAD2) have not revealed differences in Pb
binding affinity (Jaffe et al.. 2000). In a meta-analysis of 24 studies, Scinicariello et al.
(2007). observed the greatest differences for ALAD2 compared to ALAD1 in highly
exposed adults with little difference among environmentally-exposed adults; large
differences were also observed for children at low exposures. However, there are few
studies that evaluated children and the largest study contributing to the meta analysis may
have been influenced by selection bias (Scinicariello et al.. 2007). Individual studies find
similar results in occupationally-exposed adults, with blood Pb levels being higher in
individuals with ALAD2 alleles (Mivaki et al.. 2009; Shaik and Jamil. 2009). A
subsequent meta analysis of adult data from NHANES III did not find any differences in
blood Pb level between all carriers of either the ALAD 1-1 or ALAD 1-2/2-2 allele
(Scinicariello et al.. 2010). Other studies provide further support for no blood Pb
differences among ALAD1 and ALAD2 carriers (Sobin et al.. 2009; Rabstein et al..
2008; Montenegro et al.. 2006; Wananukul et al.. 2006) or lower blood Pb levels for
individuals with ALAD 1-2/2-2 (Krieg et al.. 2009; Chia et al.. 2006). The contribution of
ALAD alleles to blood Pb levels is inconclusive and the underlying causes of
discrepancies among studies remain to be elucidated.
Genetic polymorphism in the gene that encodes for peptide transporter 2 (PEPT2) has
been associated with variability in blood Pb concentrations in children (Sobin et al..
2009). PEPT2 expression in the brain and renal proximal tubule has been associated with
transport of di- and tri-peptides and may function in the transport of 5-ALA into brain
and renal tubular reabsorption of peptides. The PRPT2*2 polymorphism was associated
with increased blood Pb concentrations in a sample of 116 children of Mexican-
American/Hispanic heritage (age 4-12 years, mean blood Pb concentration 3-6 (ig/dL).
Analyses of serial blood Pb concentrations measured in longitudinal epidemiologic
studies found relatively strong correlations (e.g., r = 0.5-0.8) among each child's
individual blood Pb concentrations measured after 6-12 months of age (Schnaas et al..
2000; Dietrich et al.. 1993a; McMichael et al.. 1988; Otto etal.. 1985; Rabinowitz et al..
1984). These observations suggest that, in general, exposure characteristics of an
individual child (e.g., exposure levels and/or exposure behaviors) tend to be relatively
3-59
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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 concept is illustrated in Figure 3-7. Two
hypothetical children are simulated. Child A has a relatively constant Pb intake from
birth, whereas Child B has the same Pb intake as Child A for the first two years of life,
then a 1-year elevated intake beginning at age 24 months (Figure 3-7. upper panel) that
returns to the same intake as Child A at 36 months. The absorption fraction is assumed to
be the same for both children. Blood Pb samples 1 and 5 for Child A and B, or 2 and 4
for Child B, 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
overtime. 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 3-7. lower
panel). Time-integrated measurements of Pb concentration may provide a means for
accounting for some of these factors and, thereby, provide a better measure of long-term
Pb exposure.
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0 12 24 36 48 60 72 84
Age (months)
O)
o>
40
30
20 ^
m
0
0 12 24 36 48 60
Age (months)
72 84
Note: Child A and Child B had a constant basal Pb intake (10 |jg/day) from birth; in addition to the basal intake of 10 |jg/day, Child B
experienced an elevated intake of 5.5 |jg/kg per day (i.e., an average of 88 |jg/day) for 1 year beginning at 24 months of age (upper
panel). Blood Pb measurements 1 and 5 for Child A and B, or 2 and 4 for Child B, will yield similar blood Pb 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 blood Pb measurements can provide information about the magnitude of change in exposure, but not necessarily the
temporal history of the change (lower panel). Child C and D had a constant basal Pb intake (10 ug/day) from birth. In addition to the
basal intake of 10 ug/day, Child C experienced an elevated intake of 13 ug/kg per day (i.e., an average of 160 ug/day) for 1 year
beginning at 12 months of age, whereas, Child D experienced the same exposure as Child B. Simulation based on ICRP Pb
biokinetics model (Leggett. 1993).
Figure 3-7 Simulation of temporal relationships between Pb exposure and
blood Pb concentration in children.
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3.3.3 Urine-Pb Measurements
Standard methods that have been reported for urine Pb analysis 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.
The concentration of Pb in urine is a function of the urinary Pb excretion (Section 3.2.3)
and the urine flow rate. Urine flow rate requires collection of a timed urine sample, which
is often problematic in epidemiologic studies. Collection of untimed ("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 (Fukui et al., 1999; Araki et al., 1990). 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). 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 (Araki etal., 1990).
Urinary Pb concentration reflects, mainly, the concentration of Pb in the blood. As such,
urinary concentrations reflect both recent and past exposures to Pb (see Section 3.3.5). A
single urinary Pb measurement cannot distinguish between a long-term low level of
exposure or a higher acute exposure. Urinary Pb measurements would be expected to
correlate with concurrent blood Pb (see Section 3.3.6 for additional discussion of the
relationship between blood and urine Pb). Chiang et al. (2008) reported a significant, but
relatively weak correlation between urinary Pb levels (ng/dg creatinine) and individual
Pb intakes ((ig/day) estimated in a group of 10- to 12-year-old children (|3: 0.053,
R = 0.320, p = 0.02, N = 57). A contributing factor to the relatively weak correlation may
have been the temporal displacement between the urine sampling and measurements used
to estimate intake, which may have been as long as 6 months for some children.
Thus, 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 for the same reasons that blood
Pb is not a good indicator of body burden. On the other hand, long-term average
measurements of urinary Pb can be expected to be a better index of body burden (Figure
3-62
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10
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25 30 35 40 45 50 55 60 65 70
Age (year)
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Age (year)
Note: A change in Pb uptake results in a relatively rapid change in urinary excretion of Pb, to a new quasi-steady state, and a
relatively small change in body burden (upper panel). Baseline ingestion was 20 ug/day from age 0 to 30 yrs, then intake increased
to 120 ug/day from age 30 to 50 with a subsequent decrease in intake to the baseline of 20 ug/day at age 50. The long-term
average urinary Pb excretion more closely tracks the pattern of change in body burden (lower panel). Simulation based on ICRP Pb
biokinetics model (Leggett. 1993).
Figure 3-8 Simulation of relationship between urinary Pb excretion and body
burden in adults.
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3.3.4 Pb in Other Biomarkers
There was extensive discussion in the 2006 Pb AQCD (U.S. EPA. 2006c) regarding the
utility of other Pb biomarkers as indicators of exposure or body burden. Due to the fact
that most epidemiologic studies continue to use blood Pb or bone Pb, and other potential
biomarkers (i.e., teeth, hair, and saliva) have not been established to the same extent as
blood or bone Pb, only summaries are provided below.
3.3.4.1 Teeth
Tooth Pb is a minor contributor to the total body burden of Pb. As teeth accumulate Pb,
tooth Pb levels are generally considered an estimate of cumulative Pb exposure. The
tooth Pb-blood Pb relationship is more complex than the bone Pb-blood Pb relationship
because of differences in tooth type, location, and analytical method. 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 (Gulson et
al.. 1997; Rabinowitz et al., 1993). Limited studies have demonstrated moderate-to-high
correlations between tooth Pb levels and blood Pb levels (Rabinowitz. 1995; Rabinowitz
etal.. 1989).
Teeth are composed of several tissues formed pre- and postnatal. Therefore, if a child's
Pb exposure during the years of tooth formation varied widely, different amounts of Pb
would be deposited at different rates (Rabinowitz et al., 1993). This difference may allow
investigators to elucidate the history of Pb exposure in a child. Robbins et al. (2010)
found a significant association between environmental Pb measures that correlated with
leaded gasoline use and tooth enamel Pb in permanent teeth. Costa de Almeida et al.
(2007) discerned differences between tooth enamel Pb concentration in biopsy samples
from children who lived in areas having higher or lower levels of Pb contamination.
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., (Tvinnereim et
al.. 1997; Fosse et al.. 1995). 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 6-7 years of age.
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3.3.4.2 Hair
The 2006 Pb AQCD (U.S. EPA. 2006c) discussed applications of hair Pb measurements
for assessing Pb body burden or exposure and noted methodological limitations
(e.g., external contamination) and lack of a strong empirical basis for relating hair Pb
levels to body burden or exposure. No new methodological or conceptual advances
regarding hair Pb measurements have occurred since 2006, and widespread application of
hair Pb measurements in epidemiologic studies has not occurred.
Pb is incorporated into human hair and hair roots (Bos et al., 1985; Rabinowitz et al.,
1976) and has been explored as a noninvasive approach for estimating Pb body burden
rWilhelm et al.. 2002; Gerhardsson et al.. 1995; Wilhelm et al.. 1989). 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 blood Pb levels
less than 10-12 (ig/dL (Rodrigues et al., 2008; Campbell and Toribara. 2001; Esteban et
al.. 1999; Draschetal.. 1997). Temporal relationships between Pb exposure and hair Pb
levels, and kinetics of deposition and retention of Pb in hair have not been evaluated.
Although hair Pb measurements have been used in some epidemiologic studies (Shah et
al.. 2011; U.S. EPA. 2006b). an empirical basis for interpreting hair Pb measurements in
terms of body burden or exposure has not been firmly established.
3.3.4.3 Saliva
A growing body of literature on the utility of measurements of salivary Pb has developed
since the completion of the 2006 Pb AQCD (U.S. EPA. 2006b). Earlier reports suggested
a relatively strong correlation between salivary Pb concentration and blood Pb
concentration (Omokhodion and Crockford. 1991; Brodeur et al.. 1983; P'an. 1981);
however, more recent assessments have shown relatively weak or inconsistent
associations (Costa de Almeida et al.. 2011; Costa de Almeida et al.. 2010; Costa de
Almeida et al., 2009; Barbosa et al., 2006b; Nriagu et al.. 2006). The differences in these
outcomes may reflect differences in blood Pb concentrations, exposure history and/or
dental health (i.e., transfer of Pb between dentin and saliva) and possibly methods for
determining Pb in saliva. Barbosa et al. (2006b) found a significant but relatively weak
correlation (logfblood PB] versus logfsaliva Pb], r = 0.277, p = 0.008) in a sample of
adults, ages 18-60 years (N = 88). The correlation was similar for salivary and plasma Pb.
Nriagu et al. (2006) found also found a relatively weak association (R2 = 0.026) between
blood Pb ((ig/dL) and salivary Pb ((ig/L) in a sample of adults who resided in Detroit, MI
(N = 904). Costa de Almeida et al. (2009) found a significant correlation between
3-65
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salivary and blood Pb concentrations in children in a Pb-contaminated region in Sao
Paulo State, Brazil (r = 0.76. p = 0.04, N = 7) prior to site remediation; however, the
correlation degenerated (r = 0.03, p = 0.94, N = 9) following remediation. Nevertheless,
salivary Pb concentrations in the group of children who lived in the contaminated area
were significantly elevated compared to a reference population. It is possible, that
salivary Pb measurements may be a useful non-invasive biomarker for detecting elevated
Pb exposure; however, it is not clear based on currently available data, if salivary Pb
measurements would be a more reliable measure of exposure than blood Pb
measurements.
3.3.4.4 Serum 8-ALA and ALAD
The association between blood Pb and blood ALAD activity and serum 5-aminolevulinic
acid (5-ALA) levels was recognized decades ago as having potential use as a biomarker
of Pb exposure (Mitchell et al., 1977; Hernberg et al., 1970). More recently reference
values for blood ALAD activity ratio (the ratio of ALAD activity in the blood sample to
that measured after fully activating the enzyme in the sample) have been reported
(Gultepe et al.. 2009). Inhibition of erythrocyte ALAD by Pb results in a rise in the
plasma concentration of the ALAD substrate 5-ALA. The 5-ALA biomarker can be
measured in serum and has been used as a surrogate for Pb measurements in studies in
which whole blood samples or adequately prepared plasma or serum samples were not
available for Pb measurements (OpleretaL 2008; Opler et al.. 2004).
3.3.5 Relationship between Pb in Blood and Pb in Bone
The kinetics of elimination of Pb from the body reflects the existence of multiple pools of
Pb in the body that have different elimination kinetics. The dominant washout phase of
Pb from the blood, exhibited shortly after a change in exposure occurs, has a half-life of
-20-30 days (Leggett. 1993; Rabinowitz et al.. 1976). Studies of a limited number of
adults (four individuals with hip or knee replacement, a married couple, and 10 female
Australian immigrants) in which the Pb exposure was from historical environmental
sources (i.e., minimal current Pb exposure relative to past Pb exposure) have found that
bone Pb stores can contribute as much as 40-70% to blood Pb (Smith etal.. 1996; Gulson
et al.. 1995a; Manton. 1985). Bone Pb burdens in adults are slowly lost by diffusion
(heteroionic exchange) as well as by bone resorption (O'Flahertv. 1995). Half-times for
the release of Pb in bone are dependent on age and intensity of exposure. Bone
compartments are much more labile in infants and children than in adults as reflected by
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half-times for movement of Pb from bone into the plasma (e.g., cortical ti/2 = 0.23 years
at birth, 1.2 years at 5 years of age, 3.7 years at 15 years of age, and 23 years in adults;
trabecular ti/2 = 0.23 years at birth, 1.0 years at 5 years of age, 2.0 years at 15 years of
age, and 3.9 years in adults) (Leggett. 1993). Slow transfer rates for the movement of Pb
from nonexchangeable bone pools to the plasma are the dominant transfer process
determining long-term accumulation and elimination of bone Pb burden.
When blood Pb concentrations are monitored in individuals over periods of years
following a cessation or decrease in exposure, the decrease in blood Pb concentration
exhibits complex kinetics that can be disaggregated into components having faster and
slower rates. The slower rates of clearance of Pb from the blood over months and years
following the cessation or reduction in exposures is thought to primarily reflect
elimination of Pb stores in bone. Nilsson et al. (1991) reported a tri-exponential decay in
the blood Pb concentrations of 14 individuals having a median occupational exposure
period of 26 years. Thirteen of these 14 individuals had been temporarily removed from
work because of excessive exposures (blood levels > 70 ug/dL or high urinary
5-aminolevulinic acid levels). Representing 22% of blood Pb, the fast compartment had a
clearance half time of 34 days. The intermediate compartment, 27% of blood Pb, had a
clearance half time of 1.12 year. And, the slow compartment, 50% of blood Pb, had a
clearance half time of 13 years. The authors attributed the fast, intermediate, and slow
compartment clearance to elimination of Pb from blood and some soft tissues, from
trabecular bone, and cortical bone, respectively. Rentschler et al. (2012) also observed a
slow terminal phase of Pb elimination from blood in five adults who had Pb poisoning
due to either occupational or non-occupational exposures that ranged from approximately
1 month to 12 years and resulted in blood Pb concentrations of 70-110 (ig/dL. In this
study, the blood Pb monitoring period extended from 1 to 74 days following cessation of
exposure to approximately 800 days following the diagnosis of poisoning; however, it
was not of sufficient duration to estimate the terminal half-time. When the terminal half-
time estimated by Nilsson et al. (1991) was used (13 years) to fit data for these Pb
poisoning cases to a two-component exponential decay model, the initial faster phase
represented approximately 80% of the blood Pb and the half-time was estimated to range
from 60 to 120 days. The relatively longer fast phase half-time reported by Rentschler et
al. (2012) compared to Nilsson et al. (1991) may reflect the relatively high blood Pb
concentrations in these poisoning cases that resulted in temporary anemia and subsequent
reestablishment of a normal erythrocyte levels. Also, the use of a two-compartment
model, with an assumed slow half-time of 13 years, as well as uncertainty about the
actual time of cessation of exposure may have prevented discerning a third, faster
elimination compartment in these data.
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Children who have been removed from a relatively brief exposure to elevated
environmental Pb also exhibit faster slow-phase kinetics than children removed from
exposures that lasted several years, with half-times of 10 and 20-38 months, respectively
(Manton et al.. 2000). Rothenberg et al. (1998) also showed that exposures in the first 6
months of life could contribute to elevated blood Pb levels through at least 3 years
relative to children with lower early life exposures, despite similar environmental
exposures at later time points. In both adults and children, the longer half-times measured
under the latter conditions reflect the contribution of bone Pb stores to blood Pb
following a change in exposure.
The longer half-life of Pb in bone compared to blood Pb, allows a more cumulative
measure of long-term Pb exposure. Pb in adult bone can serve to maintain blood Pb levels
long after external exposure has ceased (Fleming et al.. 1997; Inskip etal. 1996; Smith et
al.. 1996; Kehoe. 1987; O'Flaherty et al., 1982). even for exposures that occurred during
childhood (McNeill et al.. 2000). The more widespread use of in vivo XRF Pb
measurements in bone and indirect measurements of bone processes with stable Pb
isotopes have enhanced the use of bone Pb as a biomarker of Pb body burden.
Several studies have found a stronger relationship between patella Pb and blood Pb than
tibia Pb and blood Pb (Park et al.. 2009b; Huetal.. 1998; Hernandez-Avila et al.. 1996;
Hu et al.. 1996a). Hu et al. (1998) suggest that trabecular bone is the predominant bone
type providing Pb back into circulation under steady-state and pathologic conditions. The
stronger relationship between blood Pb and trabecular Pb compared with cortical bone is
probably associated with the larger surface area of trabecular bone allowing for more Pb
to bind via ion exchange mechanisms and more rapid turnover making it more sensitive
to changing patterns of exposure.
Relationships between Pb in blood and bone in children and adults are discussed in
greater detail below (Section 3.3.5.1. and Section 3.3.5.2). In these discussions,
simulations based on a biokinetics model are shown to illustrate general patterns in the
relationships between bone Pb and blood Pb that can be predicted based on the current
understanding of Pb biokinetics in children and adults. However, these simulations reflect
assumptions in the model and may not accurately represent the observed blood Pb
kinetics in individuals or variability in blood Pb kinetics observed in specific populations.
The simulations include two metrics of blood Pb, the blood Pb concentration at each time
point in the simulation and the time-integrated blood Pb for the period preceding each
time point in the simulation (also referred to as the cumulative blood Pb index [CBLI]).
The time-integrated blood Pb metric has been used to estimate long-term average and
cumulative absorbed Pb doses in epidemiologic studies (e.g., Nie et al.. 20lib; Healev et
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al.. 2008; Hu et al. 2007a: Chuang et al.. 2000; McNeill et al.. 2000; Fleming et al..
1997: Roelsetal. 1995: Gerhardsson et al.. 1993: Armstrong et al.. 1992).
3.3.5.1 Children
As mentioned in Section 3.2.2.2. bone growth in children will contribute to accumulation
of Pb in bone, which will comprise most of the Pb body burden. As a result, Pb in bone
will more closely reflect Pb body burden than blood Pb. However, changes in blood Pb
concentration in children (i.e., associated with changing exposures) are thought to more
closely parallel changes in total body burden. Figure 3-9 shows a biokinetics model
simulation of the temporal profile of Pb in blood and bone in a child who experiences a
period of constant Pb intake (from age 2-5) via ingestion (|ig Pb/day) followed by an
abrupt decline in intake. The figure illustrates several important general concepts about
the relationship between Pb in blood and bone. While blood Pb approaches a quasi-steady
state after a period of a few months with a constant rate of Pb intake (as demonstrated by
the vertical dashed line), Pb continues to accumulate in bone with continued Pb intake
after the quasi-steady state is achieved in blood. The model also predicts that the rate of
release of Pb from bone after a reduction in exposure is faster than in adults. This
difference has been attributed to accelerated growth-related bone mineral turnover in
children, which is the primary mechanism for release of Pb that has been incorporated
into the bone mineral matrix.
Empirical evidence in support of this conclusion comes from longitudinal studies in
which relatively high correlations were found between concurrent (r = 0.75) or average
lifetime (obtained at 6-month intervals from birth to age 10 or 12) blood Pb
concentrations (r = 0.85) and tibia bone Pb concentrations (measured by XRF) in a
sample of children in which the group mean concurrent blood Pb concentration exceeded
20 (ig/dL; the correlations was much weaker (r <0.15) among the group of children with
a mean concurrent blood Pb concentration <10 (ig/dL (Wasserman et al.. 2003).
Time-integrated blood Pb metrics display rates of change in response to the exposure
event that more closely approximate the slower kinetics of total skeletal bone Pb and
body burden, than the kinetics of blood Pb concentration, with notable differences
(Figure 3-9). The time-integrated blood Pb concentration is a cumulative function and
increases throughout childhood; however, the slope of the increase is higher during the
exposure event than prior to or following the event. Following cessation of the enhanced
exposure period, the time-integrated blood Pb and body burden diverge. This result is
expected, as the time-integrated blood Pb curve is a cumulative function which cannot
decrease over time and bone Pb levels will decrease with reduction in exposure.
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The time-integrated blood Pb concentration will be a better reflection of the total amount
of Pb that has been absorbed, than the body burden at any given time. The time-
integrated blood Pb concentration will also reflect cumulative Pb absorption, and
cumulative exposure if the absorption fraction is constant. This concept is illustrated in
the hypothetical simulations of an exposure event experienced by a child (Figure 3-10).
This pattern is similar for adults.
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10
10
50
246
Age (year)
10
Note: Blood Pb concentration is thought to parallel body burden more closely in children than in adults, due to more rapid turnover of
bone and bone-Pb stores in children (upper panel). Baseline Pb intake is 3.2 ug/day from birth until age 2, followed by a period of
increased intake (38.2 ug/day) from age 2 until age 5, with a return to baseline intake of 3.2 ug/day at age 5. The time-integrated
blood Pb concentration increases overtime (lower panel). Simulation based on ICRP Pb biokinetics model (Leggett. 1993).
Figure 3-9 Simulation of relationship between blood Pb concentration and
body burden in children, with an elevated constant Pb intake from
age 2 to 5 years.
3-71
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Note: The simulations include a 3-year period of elevated constant Pb intake during ages 2-5 years. Baseline Pb intake is
3.2 ug/day from birth until age 2, followed by a period of increased intake (38.2 ug/day) from age 2 until age 5, with a return to
baseline intake of 3.2 ug/day at age 5. The time-integrated blood Pb concentration closely parallels cumulative Pb absorption.
Simulation based on ICRP Pb biokinetics model (Leggett. 1993).
Figure 3-10 Simulation of relationship between time-integrated blood Pb
concentration and cumulative Pb absorption in children.
3.3.5.2
Adults
In adults, where a relatively large fraction of the body burden residing in bone has a
slower 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 over a much longer period, largely as a consequence of
continued accumulation of Pb in bone. This pattern is illustrated in Figure 3-11 that
depicts a hypothetical simulation of an exposure consisting of a 20-year period of daily
ingestion of Pb in an adult. The exposure shown in the simulations gives rise to a
relatively rapid increase in blood Pb concentration from a baseline of approximately
2 (ig/dL, to a new quasi-steady state of approximately 9 (ig/dL, achieved in -75-100 days
(i.e., approximately 3-4 times the blood elimination half-life). In contrast, the body
burden exhibits a steady increase across the full exposure period of 70 years. Following
cessation of the enhanced exposure period, blood Pb concentration declines relatively
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rapidly compared to the slower decline in body burden. Careful examination of the
simulation shown in Figure 3-11 reveals that the accumulation and elimination phases of
blood Pb kinetics are not symmetrical; elimination is slower than accumulation as a result
of the gradual release of bone Pb stores to blood. This response, known as the prolonged
terminal elimination phase of Pb from blood, has been observed in retired Pb workers and
in workers who continued to work after improved industrial hygiene standards reduced
their exposures. Based on this hypothetical simulation, a blood Pb concentration
measured 1 year following cessation of a period of increased Pb uptake would be
elevated by only a relatively small amount over the baseline measured prior to the
exposure (3 (ig/dL versus the 2 (ig/dL), whereas, the body burden would remain elevated.
These simulations in Figure 3-11 illustrate how a single blood Pb concentration
measurement, or a series of measurements taken over a short-time span, could be a
relatively poor index of Pb body burden.
The simulation shown in Figure 3-11 represents an exposure that resulted in a quasi-
steady state blood Pb concentration of approximately 10 (ig/dL. Exposures that achieve
higher blood Pb concentrations, more indicative of poisoning or historic occupational
exposures will result in a more prolonged elevation of blood Pb concentration following
cessation of the enhanced exposure period. Figure 3-12 shows a model simulation of an
adult exposed to Pb, that results in a quasi-steady state blood Pb concentration of
approximately 90 (ig/dL. In this case, the blood Pb concentration remains substantially
elevated 1 year following the exposure event (42 (ig/dL versus 2 (ig/dL) and 20 years
following the exposure event (11 (ig/dL).
The drop in blood Pb concentrations following cessation of elevated exposure in Figure
3-11 and Figure 3-12 are well described (r = 0.996) by a tri-exponential decay function
having the half-times of 30 days, 5 months, and 8 years for the fast, intermediate, and
slow compartments, respectively. For the low level of Pb intake illustrated in Figure 3-11,
the fast, intermediate, and slow clearance compartments represent 66%, 19%, and 15% of
blood Pb, respectively. For the high level of Pb intake illustrated in Figure 3-12, the fast,
intermediate, and slow clearance compartments represent 35%, 19%, and 46% of blood
Pb, respectively. The higher exposure resulted in more accumulation of Pb in bone
relative to the lower exposure scenario. This bone accumulation is reflected in the blood
Pb clearance kinetics by a larger slow compartment and smaller fast compartment for the
high exposure in Figure 3-12 relative to the lower exposure in Figure 3-11.
One important potential implication of the profoundly different kinetics of Pb in blood
and bone is that, for a constant Pb exposure, Pb in bone will increase with increasing
duration of exposure and, therefore, with age. In contrast, blood Pb concentration will
achieve a quasi-steady state. As a result, the relationship between blood Pb and bone Pb
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will diverge with increasing exposure duration and age. This divergence can impart
different degrees of age-confounding when either blood Pb or bone Pb is used as an
internal dose metric in dose-response models. In a review of epidemiologic studies that
evaluated the associations between blood Pb, bone Pb and cognitive function, the
association was stronger for bone Pb than blood Pb (particularly for longitudinal studies)
for older individuals with environmental Pb exposures and low blood Pb levels (Shih et
al., 2007). In contrast, occupational workers with high current Pb exposures had the
strongest associations for blood Pb levels with cognitive function, thus providing
evidence for this divergence (Shih et al., 2007).
The aforementioned expectation for an increase in bone Pb and body burden with age
applies to scenarios of constant exposure but not necessarily to real world populations in
which individual and population exposures have changed overtime. Longitudinal studies
of blood and bone Pb trends have not always found strong dependence on age (Nie et al..
2009; Kimetal.. 1997). Kim et al. (1997) found that bone Pb levels increased with
increasing age in elderly adults (age 52-83) years), only when the data were analyzed
cross-sectionally. When analyzed longitudinally, the trend for individual patella Pb was a
23% decrease over a 3-year period (approximate ti/2 of 8 years), whereas tibia Pb levels
did not change with over the same period. Therefore, although older individuals tended to
have higher bone Pb levels, the 3-year temporal trend for individuals was a loss of Pb
from the more labile Pb stores in trabecular bone. Nie et al. (2011 a) observed that
longitudinal observations of blood and bone Pb in elderly adults did not show a
significant age effect on the association between blood Pb and bone Pb (patella and tibia),
when the sample population (N=776) was stratified into age tertiles (mean age 62, 69 or
77 years). Nie et al. (2009) did find that regressed function bone Pb and appeared to level
off at bone Pb levels >20 (ig/g bone mineral.
3-74
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10
O)
~ "H
CL
i "^
2 -
20
CD
« 1
O
•^
CD
O
10 <
CD
c.
a
(D
3
-5 •=•
CD
25 30 35 40 45 50 55 60 65 70
Age (year)
350
.0
Q.
300 -
•o 250 -
^ > 200 -
||15°-
•T 100 -
E
'Ł 50 H
- 15
C>-i 10
- 5
20
03
o
D
(D
O
CD
O
<
03
c.
a
(D
3
CD
25 30 35 40 45 50 55 60 65 70
Age (year)
Note: A constant baseline intake of 20 |jg/day from age 0-30 results in a quasi-steady state blood Pb concentration and body
burden. An increase in Pb intake to 120 ug/day from age 30 to 50 gives rise to a relatively rapid increase in blood Pb, to a new
quasi-steady state, and a slower increase in body burden (upper panel). At age 50, intake returns to the baseline of 20 ug/day.
Following the long period of elevated Pb intake, there is a rapid decline in blood Pb from 9 to 3 ug/dL over the first year and a more
gradual decline in blood Pb to less than 2 ug/dL by age 60. The time-integrated blood Pb concentration increases over the lifetime,
with a greater rate of increase during periods of higher Pb uptake (lower panel). Simulation based on ICRP Pb biokinetics model
(Leggett. 1993).
Figure 3-11 Simulation of relationship between blood Pb concentration, bone
Pb and body burden in adults with relatively low Pb intake.
3-75
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600
25 30 35 40 45 50 55 60 65 70
Age (year)
Ł1
0.
2500
2000 -
600
F 1500 -
25
|1
Q)
1000 -
500 -
Blood
— • -Bone ..
Body
25 30 35 40 45 50 55 60 65 70
Age (year)
Note: A constant baseline intake of 20 |jg/day from age 0-30 results in a quasi-steady state blood Pb concentration and body
burden. An increase in Pb intake to 6020 ug/day from age 30 to 50 gives rise to a relatively rapid increase in blood Pb, to a new
quasi-steady state, and a slower increase in body burden (upper panel). At age 50, intake returns to the baseline of 20 ug/day.
Following the long period of high Pb intake, there is a rapid decline in blood Pb from 90 to 40 ug/dL over the first a year followed by
a more gradual decline in blood Pb to 20 ug/dL by age 60 and 10 ug/dL at age 70. The time-integrated blood Pb concentration
increases over the lifetime, with a greater rate of increase during periods of higher Pb uptake (lower panel). Simulation based on
ICRP Pb biokinetics model (Leggett. 1993).
Figure 3-12 Simulation of relationship between blood Pb concentration, bone
Pb and body burden in adults with relatively high Pb intake.
3-76
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Tibia bone Pb is correlated with time-integrated blood Pb concentration (i.e., CBLI).
McNeill et al. (2000) compared tibia Pb levels and cumulative blood Pb indices in a
population of 19- to 29-year-olds who had been highly exposed to Pb in childhood from
the Bunker Hill, Idaho smelter. They concluded that Pb from exposure in early childhood
had persisted in the bone matrix until adulthood. The bone Pb/CBLI slopes from various
studies range from 0.022 to 0.067 ug/g bone mineral per ug-year/dL (Healev et al.. 2008;
Hu et al.. 2007a). Because the CBLI is a cumulative function which cannot decrease over
time, CBLI and bone Pb would be expected to diverge following cessation of exposure,
as bone Pb levels decrease. This divergence was observed as a lower bone Pb/CBLI slope
in retired Pb workers compared to active workers and in worker populations whose
exposures declined overtime as a result of improved industrial hygiene (Fleming et al..
1997; Gerhardsson et al.. 1993).
Although differences in kinetics of blood and bone Pb degrade the predictive value of
blood Pb as a metric of Pb body burden, within a population that has similar exposure
histories and age demographics, blood and bone Pb may show relatively strong
associations. A recent analysis of a subset of data from the Normative Aging Study (an
all male cohort) showed that cross-sectional measurements of blood Pb concentration
accounted for approximately 9% (tibia) to 13% (patella) of the variability in bone Pb
levels. Inclusion of age in the regression model accounted for an additional 7-10% of the
variability in bone Pb (Park et al.. 2009b).
Mobilization of Pb from Bone in Adulthood
In addition to changes in exposure (e.g., declines in exposure discussed in prior sections),
there are physiological processes during different life circumstances that can increase the
contribution of bone Pb to blood Pb. These life circumstances include times of
physiological stress associated with enhanced bone remodeling such as during pregnancy
and lactation (Hertz-Picciotto et al.. 2000; Silbergeld. 1991; Manton. 1985). menopause
or in the elderly (Silbergeld et al.. 1988). extended bed rest (Markowitz and Weinberger.
1990). hyperparathyroidism (Kessler et al.. 1999) and severe weight loss (Riedt et al..
2009).
During pregnancy, bone Pb can serve as a Pb source as maternal bone is resorbed for the
production of the fetal skeleton (Gulson et al.. 2003; Gulson et al.. 1999; Franklin et al..
1997; Gulson et al.. 1997). Increased blood Pb during pregnancy has been demonstrated
in numerous studies and these changes have been characterized as a "U-shaped" pattern
of lower blood Pb concentrations during the second trimester compared to the first and
third trimesters (Lamadrid-Figueroa et al.. 2006; Gulson et al.. 2004a: Hertz-Picciotto et
al.. 2000; Gulson etal.. 1997; Lagerkvist et al.. 1996; Schuhmacher et al.. 1996;
3-77
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Rothenberg et al., 1994a). The U-shaped relationship reflects the relatively higher impact
of hemodilution in the second trimester versus the rate of bone Pb resorption
accompanying Ca2+ releases for establishing the fetal skeleton. In the third trimester, fetal
skeletal growth on calcium demand is greater, and Pb released from maternal skeleton
offsets hemodilution. Gulson et al. Q998b) reported that, during pregnancy, blood Pb
concentrations in the first immigrant Australian cohort (N = 15) increased by an average
of about 20% compared to non-pregnant migrant controls (N = 7). 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%. Interestingly, the percent change in blood Pb
concentration was significantly greater during the post-pregnancy period than during the
second and third trimesters. This is consistent with Hansen et al. (20 lib) that
demonstrated the greatest blood Pb levels at 6 weeks postpartum compared to the second
trimester in 211 Norwegian women. Increased calcium demands of lactation (relative to
pregnancy) may contribute to the greater change in blood Pb observed post pregnancy
compared to the second and third trimesters. The contribution of skeletal Pb to blood Pb
during the post-pregnancy period remained essentially constant at the increased level of
Pb mobilization.
Gulson et al. (2004a) observed that calcium supplementation was found to delay
increased mobilization of Pb from bone during pregnancy and halved the flux of Pb
release from bone during late pregnancy and postpartum. In another study, women whose
daily Ca2+ intake was 850 mg per day showed lower amounts of bone resorption during
late pregnancy and postpartum than those whose intake was 560 mg calcium per day
(Manton et al.. 2003). Similarly, calcium supplementation (1,200 mg/day) in pregnant
Mexican women resulted in an 11% reduction in blood Pb level compared to placebo and
a 24% average reduction for the most compliant women (Ettinger et al.. 2009). When
considering baseline blood Pb levels in women who were more compliant in taking
calcium supplementation, the reductions were similar for those <5 (ig/dL and those
> 5 (ig/dL (14% and 17%, respectively). This result is in contrast to a study of women
who had blood Pb concentrations <5 (ig/dL, where calcium supplementation had no
effect on blood Pb concentrations (Gulson et al.. 2006b). These investigators attributed
their results to changes in bone resorption with decoupling of trabecular and cortical bone
sites.
Miranda et al. (2010) studied blood Pb level among pregnant women aged 18-44 years
old. The older age segments in the study presumably had greater historic Pb exposures
and associated stored Pb than the younger age segments. Compared with the blood Pb
levels of a reference group in the 25-29 years old age category, pregnant women
> 30 years old had significant odds of having higher blood Pb levels (aged 30-34:
3-78
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OR = 2.39, p <0.001; aged 35-39: OR = 2.98, p <0.001; aged 40-44: OR = 7.69,
p <0.001). Similarly, younger women had less chance of having higher blood Pb levels
compared with the reference group (aged 18-19: OR = 0.60, p = 0.179; aged 20-24:
OR = 0.54, p = 0.015). These findings indicate that maternal blood Pb levels are more
likely the result of Pb mobilization of bone stores from historic exposures as opposed to
contemporaneous exposures.
Blood Pb levels increase during lactation due to alterations in 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). In another Mexico
City study, Ettinger et al. (2006; 2004b) concluded that an interquartile increase in patella
Pb was associated with a 14% increase in breast milk Pb, whereas for tibia Pb the
increase was ~5%. Breast milk:maternal blood Pb concentration ratios are generally <0.1,
although values of 0.9 have been reported (Kovashiki et al., 2010; Ettinger et al., 2006;
Gulsonet al.. 1998a). Dietary intake of polyunsaturated fatty acids (PUFA) has been
shown to weaken the association between Pb levels in patella and breast milk, perhaps
indicating decreased transfer of Pb from bone to breast milk with PUFA consumption
(Arora et al., 2008). Breast milk as a source of infant Pb exposure was also discussed in
Section 3.1.3.3 on dietary Pb exposure.
The Pb content in some bones (i.e., mid femur and pelvic bone) plateau at middle age and
then decreases at older ages (Drasch etal.. 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). Two studies indicate that the endogenous release rate in
postmenopausal women ranges from 0.13-0.14 (ig/dL in blood per (ig/g bone and is
nearly double the rate found in premenopausal women (0.07-0.08 (ig/dL per (ig/g bone)
(Popovic et al.. 2005; Garrido Latorre et al.. 2003). An analysis of data on blood Pb
concentrations and markers of bone formation (serum alkaline phosphatase) and
resorption (urinary cross-linked N-telopeptides, NTx) in a sample of U.S. found that
blood Pb concentrations were higher in women (pre- or post-menopausal) who exhibited
the highest bone formation or resorption activities (Jackson et al.. 2010). Calcium or
vitamin D supplementation decreased the blood Pb concentrations in the highest bone
formation and resorption tertiles of the population of post-menopausal women.
Significant associations between increasing NTx and increasing blood Pb levels
(i.e., increased intercept of regression model relating the change in blood Pb per change
in bone Pb) have also been observed in elderly males (Nie et al., 2009).
3-79
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Studies of the effect of hormone replacement therapy on bone Pb mobilization have
yielded conflicting results (Popovic et al.. 2005; Berkowitz et al.. 2004; Garrido Latorre
et al., 2003; Korrick et al., 2002; Webber et al.. 1995). In women with severe weight loss
(28% of BMI in 6 months) sufficient to increase bone turnover, increased blood Pb levels
of approximately 2.1 (ig/dL (250%) were reported, and these blood Pb increases were
associated with biomarkers of increased bone turnover (e.g., urinary pyridinoline cross-
links) (Riedt et al.. 2009).
3.3.6 Relationship between Pb in Blood and Pb in Soft Tissues
Figure 3-13 shows simulations of blood and soft tissues Pb (including brain) for the same
exposure scenarios previously displayed. Pb uptake and elimination in soft tissues is
much faster than bone. As a result, following cessation of a period of elevated exposure,
Pb in soft tissues is more quickly returned to blood. The terminal elimination phase from
soft tissue mimics that of blood, and it is similarly influenced by the contribution of bone
Pb returned to blood and being redistributed to soft tissue.
Information on Pb levels in human brain is limited to autopsy data. These data indicate
brain/blood Pb ratios of approximately 0.5 in infancy which remain relatively constant
over the lifetime (range 0.3 to 1.1) (Barry. 1981. 1975). The simulation of brain Pb
shown in Figure 3-14 reflects general concepts derived from observations made in
non-human primates, dogs and rodents. These observations suggest that peak Pb levels in
the brain are reached 6 months following a bolus exposure and within two months
approximately 80% of steady state brain Pb levels are reached (Leggett. 1993). There is a
relatively slow elimination of Pb from brain (ti/2 ~ 2 years) compared to other soft tissues
(Leggett 1993). This slow elimination rate is reflected in the slower elimination phase
kinetics is shown in Figure 3-14. Although in this model, brain Pb to blood Pb transfer
half-times are assumed to be the same in children and adults, uptake kinetics are assumed
to be faster during infancy and childhood, which achieves a higher fraction of the soft
tissue burden in brain, consistent with higher brain/body mass relationships. The uptake
half times predicted by Leggett (1993) vary from 0.9 to 3.7 days, depending on age.
Brain Pb kinetics represented in the simulations are simple outcomes of modeling
assumptions and cannot currently be verified with available observations in humans.
3-80
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10
8 -
O)
^ 6 H
.a
Q.
•o
m
2 -
, Blood
"" \
•Soft Tissue
1.0
-- 0.8
0.0
4 6
Age (year)
10
10
0.0
25 30 35 40 45 50 55 60 65 70
Age (year)
Note: For the child simulation (upper panel), baseline Pb intake is 3.2 ug/day from birth until age 2, followed by a period of increased
intake to 38.2 ug/day from age 2 until 5, with a return to baseline intake at age 5. For the adult simulation (lower panel), baseline
intake is 20 ug/day from age 0-30, followed by a 20-year period of increased intake to 120 ug/day from age 30 to 50, with a return to
baseline intake at age 50. Simulation based on ICRP Pb biokinetics model (Leggett. 1993).
Figure 3-13 Simulation of blood and soft tissue (including brain) Pb in
children and adults who experience a period of increased Pb
intake.
3-81
-------
10
„ 8
en
6
CL
•D
8 4H
CO
2 -
m
Blood
— Brain
4 6
Age (year)
30
25
o
20 sr
03
15 I
T3
tr
10 g
5
10
25 30 35 40 45 50 55 60 65 70
Age (year)
Note: For the child simulation (upper panel), baseline Pb intake is 3.2 ug/day from birth until age 2, followed by a period of increased
intake to 38.2 ug/day from age 2 until 5, with a return to baseline intake at age 5. For the adult simulation (lower panel), baseline
intake is 20 ug/day from age 0-30, followed by a 20-year period of increased intake to 120 ug/day from age 30 to 50, with a return to
baseline intake at age 50. Simulation based on ICRP Pb biokinetics model (Leggett. 1993).
Figure 3-14 Simulation of blood and brain Pb in children and adults who
experience a period of increased Pb intake.
3-82
-------
Urinary filtering and excretion of Pb is associated with plasma Pb concentrations. Given
the curvilinear relationship between blood Pb and plasma Pb, a secondary expectation is
for a curvilinear relationship between blood Pb and urinary Pb excretion that may
become evident only at relatively high blood Pb concentrations (e.g., >25 (ig/dL). Figure
3-15 shows these relationships predicted from the model. In this case, the exposure
scenario shown is for an adult (age 40 years) at a quasi-steady state blood Pb
concentration; the same relationships hold for children (Leggett 1993). At lower blood
Pb concentrations (<25 (ig/dL), urinary Pb excretion is predicted to closely parallel
plasma Pb concentration for any given blood Pb level (Figure 3-15. top panel). It follows
from this that, similar to blood Pb, urinary Pb will respond much more rapidly to an
abrupt change in Pb exposure than will bone Pb. One important implication of this
relationship is that, as described previously for blood Pb, the relationships between
urinary Pb and bone Pb will diverge with increasing exposure duration and age, even if
exposure remains constant. Furthermore, following an abrupt cessation of exposure, urine
Pb will quickly decrease while bone Pb will remain elevated (Figure 3-15. lower panel).
3-83
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80
20 30 40
Blood Pb (ng/dL)
0.00
50
IZ. •
10 •
"> j"
WTJ
5 o> 8
0 3.
3.—
""""^
Ł °- 6 •
a>^
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2
n
Blood
^x"**1!'.1- • • • • J •'
f**'*'
* •
t
1 •
*
/
r
*
r
*
^— Urine
. Bone
•
•
•
• .
V
^-*— „
• ^u
- 15
CD
0
- 10 m
(Q
*™*
- 5
n
25 30 35 40 45 50 55 60 65 70
Age (year)
Note: Upper panel, model simulations are for a 40-year old having a constant intake from birth of between 1 and 1,000 ug/day. For
the lower panel, baseline intake is 20 ug/day from age 0-30, followed by a 20-year period of increased intake to 120 ug/day from
age 30 to 50, with a return to baseline intake at age 50. Simulation based on ICRP Pb biokinetics model (Leggett, 1993).
Figure 3-15 Relationship between Pb in urine, plasma, blood and bone.
3-84
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3.4 Studies of Pb Biomarker Levels
3.4.1 Pb in Blood
Overall, trends in blood Pb levels have been decreasing among U.S. residents over the
past 35 years. Blood Pb concentrations in the U.S. general population have been
monitored in the NHANES. Analyses of these data show a progressive downward trend
in blood Pb concentrations during the period 1976-2010, with the most dramatic declines
coincident with the phase out of leaded gasoline and reductions in point source Pb
emissions described in Section 2.2 (Pirkle et al., 1998; Brody et al., 1994; Pirkle et al.,
1994; Schwartz and Pitcher. 1989). The temporal trend for the period 1988-2010 is
shown in Figure 3-16. Summary statistics from the most recent publicly available data
(1999-2010) are presented in Table 3-8 (CDC. 2013). The geometric mean Pb
concentration among children 1-5 years of age, based on the sample collected during the
period 2009-2010, was 1.17 (ig/dL (95% CI: 1.08, 1.26), which was decreased from
2007-2008 (1.51 ng/dL, 95% CI: 1.37, 1.66). Figure 3-17 uses NHANES data to illustrate
temporal trends in the distribution of blood Pb levels among U.S. children aged 1-5 years.
For 2005-2010, the 95th percentile of blood Pb levels for children aged 1-5 years was less
than 5 (ig/dL. The geometric mean blood Pb concentration among adults > 20 years of
age was 1.23 (ig/dL (95% CI: 1.19, 1.28) for the sample collected during the period
2009-2010 (CDC. 2013). Based on these same data, the geometric mean for all males
(aged > 1 year) was 1.31 (ig/dL (95% CI: 1.25, 1.36), and for females (aged > 1 year) was
0.97 (ig/dL (95% CI: 0.93, 1.01).
There has been a steep decline in mean blood Pb levels from 1975 through 2010 among
all birth cohorts from 1975 to 2010. Figure 3-18 illustrates results of an analysis of data
from the NHANES II (1976-1980), NHANES III (1988-1994), and continuous NHANES
(1999-2008) surveys. In this analysis, participants were grouped by year of birth,
ascertained by survey time period and reported age. The geometric mean of participant
blood Pb was plotted for each age cohort as a function of year in which the survey was
performed. For all cohorts, blood Pb generally decreases with age during childhood until
adolescence; following adolescence (in the early 20s), blood Pb generally levels off or
even increases with age. It is possible that bone growth in young people and occupational
exposure for adults influences the shape of these curves. For the 1960 to 1970 birth
cohort, the mean blood Pb is the highest of the cohorts in the 1970s, but beginning in
1993 the mean blood Pb is one of the lowest of the cohorts. This interaction between time
and cohort may be due to the faster release of Pb from bone in younger people
(Rabinowitz. 1991). This interaction is also apparent for some of the other more recently
born cohorts. In comparison, the slopes of blood Pb overtime are nearly parallel among
3-85
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the cohorts born before 1960. This observation suggests that the time-cohort-interaction
diminishes among older people. Also, the leveling of the blood Pb in the 2000s could be
due to aging of the birth cohort and consequent slowing of their Pb release from bone.
When race/ethnicity groups were compared for years 1999-2004, geometric means (GM)
of blood Pb levels in children were highest in the ethnicity category non-Hispanic black
(GM 2.8, 95% CI: 2.5, 3.0) compared to the categories Mexican-American (GM 1.9, 95%
CI: 1.7, 2.0) and non-Hispanic white (GM 1.7, 95% CI: 1.6, 1.8) (Jones et al.. 2009a).
Figure 3-19 demonstrates the change in percent of children (aged 1-5 years) with various
blood Pb levels by race/ethnicity between the survey during 1988-1991 and that during
1999-2004. When these data for children aged 1-5 years were aggregated for all survey
years from 1988 to 2004, residence in older housing, poverty, age, and being
non-Hispanic black were significant predictors of higher Pb levels (Jones et al.. 2009a).
O)
_i
ro
•o
4 -
3 -
1 -
Children 1-5 yrs
Children 6-11 yrs
Teens 12-19 yrs
Adults>20 yrs
88-91 91-94 99-00 01-02 03-04 05-06 07-08 09-10
Survey Period
Note: Shown are geometric means and 95% CIs based on data from NHANES III Phase 1 (Brodyet al., 1994: Pirkleetal., 1994):
NHANES III Phase 2 (Pirkleetal.. 1998): and NHANES IV (CDC. 2011 a). Data for adults during the period 1988-1994 are for ages
20-49 years, and > 20 years for the period 1999-2008.
Figure 3-16 Temporal trend in blood Pb concentration.
3-86
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Table 3-8 Blood Pb concentrations in the U.S. population.
Survey Stratum
Period
Geometric Mean (ug/dL) 95% Confidence Interval Number of Subjects
All
1-5yr
6-1 1 yr
1999-2000
2001-2002
2003-2004
2005-2006
2007-2008
2009-2010
1999-2000
2001-2002
2003-2004
2005-2006
2007-2008
2009-2010
1999-2000
2001-2002
2003-2004
2005-2006
1.66
1.45
1.43
1.29
1.27
1.12
2.23
1.70
1.77
1.46
1.51
1.17
1.51
1.25
1.25
1.02
1.60, 1.72
1.39, 1.51
1.36, 1.50
1.23, 1.36
1.21, 1.34
1.08, 1.16
1.96,2.53
1.55, 1.87
1.60, 1.95
1.36, 1.57
1.37, 1.66
1.08, 1.26
1.36, 1.66
1.14, 1.36
1.12, 1.39
0.95, 1.01
7,970
8,945
8,373
8,407
8,266
8,793
723
898
911
968
817
836
905
1,044
856
934
2007-2008
2009-2010
12-1 9 yr 1999-2000
2001-2002
2003-2004
2005-2006
2007-2008
2009-2010
0.99
0.84
1.10
0.94
0.95
0.80
0.80
0.68
0.91, 1.07
0.79, 0.89
1.04, 1.17
0.90, 0.99
0.88, 1.02
0.75, 0.85
0.74, 0.86
0.64, 0.73
1,011
1,009
2,135
2,231
2,081
1,996
1,074
1,183
>20yr
1999-2000
2001-2002
2003-2004
2005-2006
2007-2008
2009-2010
1.75
1.56
1.52
1.41
1.38
1.23
1.68, 1.81
1.49, 1.62
1.45, 1.60
1.34, 1.48
1.31, 1.46
1.19, 1.28
4,207
4,772
4,525
4,509
5,364
5,765
3-87
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Table 3-8 (Continued): Blood Pb concentrations in the U.S. population.
Survey Stratum
Period
Geometric Mean (ug/dL) 95% Confidence Interval Number of Subjects
Males
Females
Mexican - Americans
Non-Hispanic blacks
Non-Hispanic whites
1999-2000
2001-2002
2003-2004
2005-2006
2007-2008
2009-2010
1999-2000
2001-2002
2003-2004
2005-2006
2007-2008
2009-2010
1999-2000
2001-2002
2003-2004
2005-2006
2007-2008
2009-2010
1999-2000
2001-2002
2003-2004
2005-2006
2007-2008
2009-2010
1999-2000
2001-2002
2003-2004
2005-2006
2007-2008
2009-2010
2.01
1.78
1.69
1.52
1.47
1.31
1.37
1.19
1.22
1.11
1.11
0.97
1.83
1.46
1.55
1.29
1.25
1.14
1.87
1.65
1.69
1.39
1.39
1.24
1.62
1.43
1.37
1.28
1.24
1.10
1.93,2.09
1.71, 1.86
1.62, 1.75
1.42, 1.62
1.39, 1.56
1.25, 1.36
1.32, 1.43
1.14, 1.25
1.14, 1.31
1.05, 1.17
1.06, 1.16
0.93, 1.01
1.75, 1.91
1.34, 1.60
1.43, 1.69
1.21, 1.38
1.15, 1.36
1.03, 1.28
1.75,2.00
1.52, 1.80
1.52, 1.89
1.26, 1.53
1.30, 1.48
1.18, 1.30
1.55, 1.69
1.37, 1.48
1.32, 1.43
1.19, 1.37
1.16, 1.33
1.04, 1.16
3,913
4,339
4,132
4,092
4,147
4,366
4,057
4,606
4,241
4,315
4,119
4.427
2,742
2,268
2,085
2,236
1,712
1,966
1,842
2,219
2,293
2,193
1,746
1,593
2,716
3,806
3,478
3,310
3,461
3,760
Age strata correspond to the NHANES study design.
Source: Adapted from data from the NHANES (CDC.
2013).
-------
70 -
10 -
I I
4- 1
Note: Top: all data. Bottom: data for subjects having blood Pb levels less than 15 ug/dL. Dark lines represent the median of the
data. Boxes represent the interquartile range of the data, and whiskers represent the 5th -95th percentile range of the data. Outliers
beyond the 95 percentile are shown with open circles.
Source: Adapted from data from the NHANES (NCHS. 2010)
Figure 3-17 Box plots of blood Pb levels among U.S. children (1-5 years old)
from the NHANES survey, 1988-2010.
3-89
-------
o
04
-Q
"S^r
Sco
E
o
1^
o
0
O)
cohort birth yrs
1900 to 1930<
1930 to 1940 <
1940 to 1950 J
1950 to 1960i
1960 to 1970 <
1970to1975<
1975 to 1930J
1980 to 1935 i
1985 to 1990
1990 to 1995
1995 to 2000
2000 to 2005
2005 to 2008 <
1970
1930
1990
exam year
2000
2010
Note: The means of logged blood Pb were weighted to represent national averages. Data were from the publicly available
NHANES II, NHANES III for 1988-1991 and 1992-1994, and the continuous NHANES in 1999-2000, 2003-2004, 2005-2006,
2007-2008. Continuous NHANES data from 2001-2002 and 2009-2010 are not included because there were only 551 blood Pb
samples in each of those data sets. The year plotted for exam year was the reported exam year for NHANES II, the middle year of
each of the phases of NHANES III, and the second year of each of the continuous NHANES.
Source: Adapted from data from the NHANES (NCHS. 2010)
Figure 3-18 Blood Pb cohort means versus year of exam.
3-90
-------
10
1999 - 2004
0
<1 l-<2.5 2.5-<5 5-<7.5 7.5 - <10 > 10
Blood Pb Level (ug/dL)
...+.. Non-Hispanic black^^^—Mexican American —^— Non-Hispanic white
Source: Data used with permission of the American Academy of Pediatrics, Jones et al. (2009a)
Figure 3-19 Percent distribution of blood Pb levels by race/ethnicity among
U.S. children (1-5 years) from the NHANES survey, 1988-1991 (top)
and 1999-2004 (bottom).
3-91
-------
In agreement with the 1986 AQCD (U.S. EPA. 1986a). several studies have shown
seasonal variation in blood Pb concentrations in children (e.g., Havlena et al.. 2009;
Gulson et al.. 2008; Kemp et al.. 2007; Laidlaw et al.. 2005; Haley and Talbot. 2004;
Johnson and Bretsch. 2002; Yiin et al.. 2000; Johnson et al.. 1996). with elevated
concentrations during the warm season and lower levels in the cold season. Seasonal
dynamics of blood Pb concentrations in children appear to be caused at least in part by
seasonal patterns in access of children to soils and soil properties (e.g., moisture content)
that may contribute to seasonal variation in entrainment of soil and dust Pb into breathing
zone air (Laidlaw et al.. 2012; Laidlaw et al.. 2005; Johnson and Bretsch. 2002). Seasonal
variation in blood Pb concentrations occur, and summertime blood Pb concentrations
demonstrated the strongest association with soil Pb concentrations (Johnson and Bretsch.
2002). Laidlaw et al. (2012) observed that air Pb in the PM25 fraction and PM25
attributed to soil were elevated in the warm season compared with the cold season. Yiin
et al. (2000) also found that geometric mean for blood Pb, floor Pb loading and
concentration were statistically significantly higher in the hot season compared with the
cold season, although carpet dust Pb loading was statistically significantly higher in the
cold season compared with the hot season. However, regression of blood Pb on floor and
windowsill dust with and without adjustment for the hot, warm, and cool seasons showed
no statistically significant effect of the seasons directly on blood Pb. Meteorological
factors appear to contribute to blood Pb seasonality, and the tendency of children to play
outside during the warm months may also allow for more contact with soil Pb. Laidlaw et
al. (2005) analyzed the temporal relationships between child blood Pb concentrations and
various atmospheric variables in three cities (Indianapolis, IN: 1999-2002; Syracuse, NY:
1994-1998; New Orleans, LA: 1998-2003). Blood Pb data was obtained from public
health screening programs conducted in the three cities. Blood Pb samples were
dominated by children <5 years of age and age distribution varied across the three cities.
The temporal variation in blood Pb concentrations in each city was predicted by
multivariate regression models that included the following significant variables: PM10,
wind speed, air temperature, and soil moisture; as well as dummy variables accounting
for temporal displacement of the effects of each independent variable on blood Pb.
Laidlaw et al. (2005) reported R2 values for the regression models, but did not report the
actual regression coefficients. The R2 values were as follows: Indianapolis 0.87
(p = 0.004); Syracuse 0.61 (p = 0.0012); New Orleans 0.59 (p O.OOOOl).
Studies have examined the change in blood Pb with changes in potential Pb sources.
Gulson et al. (2004b) observed that children living near a Zn-Pb smelter in Australia had
blood Pb levels ranging from 10 to 42 (ig/dL, with 55-100% of Pb attributed to the
smelter based on isotope ratio analysis. Rubio-Andrade et al. (2011) followed a cohort of
6-8 year old children living within 3.5 km of a Mexican smelter at 0, 6, 12, and
60 months after environmental intervention including removal of 100,000 kg of
3-92
-------
Pb-containing dust from roads and homes using high efficiency vacuums. Soil Pb was
concurrently obtained but not reported at 6, 12, or 60 months. Median blood Pb level at
initiation of the study was 10.1 ug/dL for the 598 initial participants (average age: 7.2 y),
and median soil Pb was 3,300 mg/kg at the start of the study. After 60 months, median
blood Pb level was 4.4 ug/dL for the remaining 232 participants (average age: 12.2 y),
and median soil Pb concentration was 370 mg/kg at that time. Bonnard and McKone
(2009) modeled blood Pb of French children ages 21-74 months living within a village
containing a Pb smelter and estimated blood Pb levels of 3.2-10.9 ug/dL. Lanphear et al.
(1998) noted that the probability of children having blood Pb > 10 ug/dL increases both
with exterior soil Pb content and interior Pb dust loading. Mielke et al. (20lib) noted
significant increases in percentages of children younger than 7-years old with blood Pb
level > 10 ug/dL for those living in inner city New Orleans housing developments
(22.9%) compared with children living in communities located on the city outskirts
(9.1%). At the same time, median soil Pb was significantly higher in the inner city
(438 mg/kg) compared with the city outskirts (117 mg/kg).
For infants <1 year old, very little data are available on blood Pb levels. Simon et al.
(2007) followed a cohort of 13 children living near an Australian smelter from birth
through 36 months. In general, except for children born with low blood Pb levels of ~1 to
2 ug/dL, immediately after birth blood Pb levels fell for 1-2 months to approximately
47% of birth blood Pb level. After this initial fall, all infants' blood Pb levels rose with
age until approximately 12 months old for children living in a high risk area and until
approximately 18 months for children living in a low risk area (Simon et al.. 2007).
Median blood Pb level among the children was 1.9 ug/dL at 2 months and increased to
13.6 ug/dL at 16 months. Geometric mean hand-Pb loading of the child and the mother
were significant contributors to the area under the curve for infant blood Pb, with 46%
(infant hand loading) and 60% (mother hand loading) of the variance being explained by
these variables, respectively; geometric mean of the mothers' blood Pb explained 46% of
the variance (Simon et al.. 2007). Across all the data, there was a good correlation
between child blood Pb level and child hand Pb loading (R2 = 0.70). In another study
(Carbone et al.. 1998). blood Pb levels of 15 infants aged 6-12 months were statistically
significantly lower than their neonatal cord blood Pb levels (2.24 ug/dL versus
4.87 ug/dL). Additionally, 3 infants born with blood Pb levels of greater than 7 ug/dL
were followed for a week, there was a dramatic drop in the blood Pb of from an average
of 7.6 ug/dL on Day 1 to 2.4 ug/dL on Day 7 (Carbone et al.. 1998).
Pb body burden has been reported among individuals known to consume wild game
hunted with Pb shot. For example, fifty men from Nuuk, Greenland participated in a
study in which they recorded their diet and produced blood samples (Johansen et al..
2006). Men who regularly ate hunted sea birds had an average blood Pb concentration of
3-93
-------
12.8 (ig/dL, in contrast with those who did not and had an average blood Pb
concentration of 1.5 (ig/dL. Umbilical cord blood was collected from a cohort of Inuit
newborns from northern Quebec, where the Inuit population consumes game killed with
Pb shot (Levesque et al.. 2003). The geometric mean cord blood Pb level was
0.19 junol/L [3.9 jig/dL], with arange of 0.01-1.31 nmol/L [0.2-27 jig/dL]; the Canadian
level of concern for cord blood Pb is 0.48 (imol/L [10 jig/dL]. The authors contrasted the
finding that 7% of Inuit newborns had cord blood Pb concentration > 0.48 (imol/L
[10 (ig/dL] in contrast with 0.16% of the Caucasian population in southern Quebec.
Recent studies have sought to characterize human exposure to Pb from piston-engine
aircraft emissions. Section 2.2.2.1 describes a study by Carr et al. (2011) in which Pb
concentrations, both modeled and monitored, extended beyond airport property. Miranda
et al. (2011) used GIS to study the association between blood Pb level and distance from
airports in six North Carolina Counties. They observed that the trend in blood Pb level
decreases monotonically with distance class from the airports, with subjects within 500
meters of the airports having significantly increased blood Pb levels (|3 = 0.043, 95% CI:
(0.006,0.080), p <0.05) compared with the general population for a given county after
controlling for proportion of black, Hispanic, percent receiving public assistance, and
household median income at the census block group level and including dummy variables
for season during which the children were screened for blood Pb. In this study, children
living within 500 meters of an airport had blood Pb levels that were, on average, 4.4%
higher than those at distance. Note that the authors did not include Pb emissions in their
model.
Trends in blood Pb levels have been accompanied by changes in Pb isotope ratios for
blood Pb. Isotopic ratios, described in Sections 2.2 and 2.3 as a tool for source
apportionment, have been used to associate blood Pb measurements with anthropogenic
sources of Pb in the environment. Changes in Pb isotopic ratios in blood samples reflect
the changing influence of sources of Pb following the phase-out of tetraethyl Pb
antiknock agents in automotive gasoline and changes in Pb usage in paints and other
industrial and consumer products (Gulson et al.. 2008; Ranft et al.. 2008; Gulson et al..
2006a; Ranft et al.. 2006). Gulson et al. (2006a) illustrated how a linear increase in the
isotopic ratio 206Pb/204Pb occurred in concert with a decrease in blood Pb levels among
selected study populations in Australia during the period 1990-2000 (Figure 3-20).
Gulson et al. (2006a) point out that the isotopic signature of 206Pb/204Pb derived from
Australian mines (median -16.8) differs from that of European and Asian mines, where
206Pb/204Pb varies between -17.4 and -18.1. Liang et al. (2010) also examined the trends
in blood Pb level over the period 1990 to 2006 in Shanghai and saw a reduction
corresponding to the phase out of Pb in gasoline. A plot of 208Pb/206Pb to 207Pb/206Pb for
blood and environmental samples showed overlap between the isotopic signature for coal
3-94
-------
combustion ash and that measured in blood. This result suggests a growing influence of
Pb from coal ash in Shanghai in the absence of Pb in automobile emissions. Oulhote et al.
(2011) examined Pb isotope ratios in blood Pb samples of 125 French children aged
6 mo-6 years. The study found that Pb isotope ratios could be used to attribute Pb
exposure to one source for 32% of children and to eliminate an unlikely source of Pb
exposure in 30% of children.
17.60-
Q.
*fr
O
Q.
ID
O
CM
17.20H
16.80-
16.40-
R Sq Linear = 0.53
source
O Adelaide
A BK Adult
' females
D BK children
A Broken Hill
V Hobart
< PbinCa
> Port Pirie
I Sydney
' children
(a)
1990 1992 1994 1996 1998 2000
Year
12.5-
10.0-
OQ
.Q
Q.
7.5-
5.0-
2.5-
0.0-
source
O Adelaide
O BK Adult females
D BK children
A Broken Hill
V Hobart
< Pb in Ca
C> Port Pirie
-(- Sydney children
(b)
1990 1992 1994 1996 1998 2000
Year
Source: Reprinted with permission of Academic Press, Gulson et al. (2006a).
Figure 3-20 Trends in 206Pb/204Pb isotope ratio in blood Pb (a) and trends in
blood Pb levels (b) among Australian study populations of
children during the period 1990-2000.
3-95
-------
3.4.2 Pb in Bone
An extensive national database (i.e., NHANES) is available for blood Pb concentrations
in children and adults, as described in Section 3.4.1. Bone Pb concentrations are less well
characterized. Table 3-9 and Table 3-10 are compilations of data from epidemiologic
studies that provided bone Pb concentrations by K-XRF and/or variability in
concentrations among individuals without reported occupational exposure and those with
occupational exposures, respectively. In non-occupationally exposed individuals, typical
group mean tibia bone Pb concentrations ranged from 10 to 30 ug/g. Patella bone Pb
levels are typically higher than tibia bone Pb levels in the studies considered (Table 3-9).
For example, in the Normative Aging Study, patella bone Pb concentrations were
approximately 32 ug/g, whereas tibia bone Pb concentrations were about 22 ug/g.
Occupationally exposed individuals generally had greater bone Pb concentrations than
seen in control groups (i.e., unexposed). Bone Pb data in Table 3-10 for occupationally
exposed individuals were also generally higher compared to non-occupationally exposed
individuals (Table 3-9).
Table 3-9 Epidemiologic studies that provide bone Pb measurements for non-
occupationally exposed populations.
Reference
Bandeen-
Roche et al.
(2009)
Bellinger et
al. (1994a)
Prior Pb
Study Methods Exposure
Cohort: Cumulative
Baltimore Memory Study
cohort
Age (yrs): 50-70
N:1,140
Location: Baltimore, MD
Study Period: 2001-2005
Cohort: Not reported Cumulative
Age (yrs): 5-8 (recruited);
19-20 (follow-up)
N: 79
Location: Boston, MA
Study Period: 1989-1990
Bone Pb Bone Pb Cone.
biomarker (ug/g)
Tibia Mean ± SD
Tibia: 18.8 ±
11.6
Tibia Mean (Range):
Patella Tibia: 5.4 (3-16)
Patella: 9.2
(4-18)
Distribution
of Bone Pb
(ug/g)
Not reported
High
exposure:
>24
Low
exposure:
<8.7
3-96
-------
Table 3-9 (Continued): Epidemiologic studies that provide bone Pb measurements for non-
occupationally exposed populations.
Reference
Cheng et al.
(2001)
Coon et al.
(2006)
Elmarsafawy
et al. (2006)
Prior Pb
Study Methods Exposure
Cohort: Cumulative
Normative Aging
Study cohort
Age (yrs): Mean ± SD:
Normotensive:
65.49 ±7.17
Borderline hypertension:
68.3 ±7.79
Definite hypertension:
67.93 ±6.79
N: 833 males
Location: Boston, MA
Study Period:
8/1/1991-12/31/1997
Cohort: Cumulative
Participants from Henry Ford
Health System (HFHS)
Age (yrs): > 50; Mean: 69.9
N: 121 cases; 414 controls
Location: Southeastern
Michigan
Study Period:
1995-1999 (participants
received primary health care
services)
Cohort: Not reported
Normative Aging Study
Age (yrs): Not reported
N: 471 elderly males
Location: Greater Boston
area, MA
Study Period:
6/1991-12/1994
Bone Pb Bone Pb Cone.
biomarker (ug/g)
Tibia Mean ± SD
Patella Tibia:
Normotensive:
20.27 ± 11.55
Borderline
hypertension:
23.46 ± 15.02
Definite
hypertension:
22.69 ± 14.71
Patella:
Normotensive:
28.95 ± 18.01
Borderline
hypertension:
33.73 ±21. 76
Definite
hypertension:
32.72 ± 19.55
Tibia Mean ± SD:
Calcaneus Tibia: 12.5 ± 7.8
Calcaneus: 20.5
± 10.2
Tibia Mean± SD:
Patella Tibia: 21. 6 ±1
2.0
Patella: 31. 7 ±
18.3
Distribution
of Bone Pb
(ug/g)
Lowest
quintile: Tibia:
8.5
Patella: 12.0
Highest
quintile:
Tibia: 36.0
Patella: 53.0
Tibia
Q1: 0-5.91
Q2:
5.92-10.40
Q3:
10.41-15.50
Q4: > 15.51
Calcaneus
Q1: 0-11.70
Q2:
11.71-19.07
Q3:
19.08-25.28
Q4: > 25.29
Not reported
3-97
-------
Table 3-9 (Continued): Epidemiologic studies that provide bone Pb measurements for non-
occupationally exposed populations.
Reference
Glass et al.
(2009)
Hsiehetal.
(2009b)
Huetal.
(1996a)
[As reported
in Navas-
Acien et al.,
(2008)]
Study Methods
Cohort:
Baltimore Memory Study
Age (yrs): Mean: 59.4;
Range: 50-70
N: 1,001
Location: Baltimore, MD
Study Period: 2001-2005
Cohort: Not reported
Age (yrs): Mean: Control:
46.06
N: 18 controls
Location: Not reported
Study Period: Not reported
Cohort:
Normative Aging Study
Age (yrs): 48-92; Mean ±
SD: 66.6 ±7.2
N: 590 males
Location: Boston, MA
Study Period:
8/1991-12/1994
Prior Pb Bone Pb Bone Pb Cone.
Exposure biomarker (ug/g)
Cumulative Tibia Mean ± SD:
(lifetime) Tibia: 1 8.8 ±
11.1
Control group for Tibia Mean ± SD
occupational Patella Tibia Control:
exposure group 18. 51 ±22.40
Patella Control:
7.14 ±9.81
Cumulative Tibia Mean ± SD:
Patella Tibia: 21. 8 ±
12.1
Patella: 32.1 ±
18.7
Range:
Tibia: <1-96
Patella: 1-142
Distribution
of Bone Pb
(ug/g)
NPH Scale:
Lowest tertile:
Mean Tibia
level: 16.3 ±
11.0
Middle tertile:
Mean Tibia
level: 19.3±
10.7
Highest tertile:
Mean Tibia
level: 20.3 ±
11.4
Not reported
Figures 1 and
2 show both
types of bone
Pb levels
increasing
with age
3-98
-------
Table 3-9 (Continued): Epidemiologic studies that provide bone Pb measurements for non-
occupationally exposed populations.
Prior Pb Bone Pb Bone Pb Cone.
Reference Study Methods Exposure biomarker (ug/g)
Jain et al. Cohort: Not reported Tibia Mean ± SD
(2007) VA-Normative Aging Study Patella Tibia:
Age (yrs): Not reported Non-Cases: 21 .4
N: 837 males ±13.6
Location: Greater Boston, Cases: 24.2 ±
MA 15.9
Study Period:
9/1/1991-12/31/2001 Patella:
Non-cases:
30.6±19.7
Cases: 36. 8 ±
20.8
Range:
Tibia:
Non-cases:
-3-126
Cases: -5-75
Patella:
Non-cases:
-10-165
Cases: 5-101
Distribution
of Bone Pb
(ug/g)
Mean ± SD
(Range):
Tibia:
Non-cases:
Tertilel: 10.2
±3.8 (-3-1 5)
Tertile2: 19.1
±2.3(16-23)
Tertile 3: 35.5
± 14.4
(24-126)
Cases:
Tertilel: 10.1
±5.3 (-5-1 5)
Tertile 2: 19.8
±2.2(16-23)
Tertile 3: 39.5
± 14.9(25-75)
Patella:
Non-cases:
Tertile 1:
13.9±4.9
(-10-20)
Tertile 2:
27.1±4.1
(21-34)
Tertile 3:
52.5± 20.7
(35-165)
Cases:
Tertilel:
15.3±4.3
(5-19)
Tertile 2: 25.7
±3.8(21-33)
Tertile 3: 53.3
± 17.3
(35-101)
3-99
-------
Table 3-9 (Continued): Epidemiologic studies that provide bone Pb measurements for non-
occupationally exposed populations.
Reference
Kamel et al.
(2002):
Kamel et al.
/or\r\c \-
(2005).
Kamel et al.
(2008)
V *- ^ ^ w /
Khaliletal.
(2009a)
Study Methods
Cohort: Not reported
Age (yrs): 30-80
N: 256 controls (Bone
samples collected from 41
controls)
Location: New England
(Boston, MA)
Study Period: 1993-1996
Cohort:
1982 Lead
Occupational Study
Age (yrs): Control mean: 55
N: 51 controls
Prior Pb
Exposure
Cumulative
Control group for
occupational
exposure group
Control group for
occupational
exposure group
Bone Pb Bone Pb Cone.
biomarker (ug/g)
Tibia Mean ± SE
Patella Tibia Controls:
11.1 ± 1.6
Patella Controls:
16.7 ±2.0
Tibia Median (IQR)
Tibia Control: 12
(-8-32)
Distribution
of Bone Pb
(ug/g)
Controls
Tibia: N (%)
-7-7: 14 (34)
8-14: 12(29)
15-61: 15(37)
Patella: N (%)
-4-9: 14 (34)
10-20: 14(34)
21-107: 13
(32)
Not reported
Location: Eastern
Pennsylvania
Study Period: 1982-2004
Korrick et al.
(1999) [As
reported in
Navas-Acien
etal.,
(2008)]
Cohort: Nonoccupationally Tibia
Nurses' Health Study exposed Patella
Age (yrs): Combined: 47-74;
Mean ± SD:
Combined: 58. 7 ± 7.2;
Cases: 61.1 ± 7.1;
High controls: 61.1 ± 7.2;
Low controls: 58.7 ±7.1
N: 284 females; (89 cases;
195 controls)
Location: Boston, MA
Study Period: 7/1993-7/1995
Mean ± SD Patella:
Tibia: 10th
Combined: 13.3 percentile: 6
± 9.0 90th
Cases: 13.0± percentile: 31
9.4
High controls:
14.7± 10
Low controls:
12.7 ±8.1
Patella:
Combined: 17.3
± 11.1
Cases: 19.5±
12.9
High controls:
17.2 ±9
Low controls:
15.8 ± 10.6
Range
Tibia Combined:
-5-69
Patella
Combined: -5-87
3-100
-------
Table 3-9 (Continued): Epidemiologic studies that provide bone Pb measurements for non-
occupationally exposed populations.
Reference
Lee et al.
(2001 a) [As
reported in
Navas-Acien
etal.,
(2QQ8)]
Martin et al.
(2006)
Needleman
et al. (2002)
Osterberg et
al. (1997)
[As reported
in Shih et
al., (2007)1
Study Methods
Cohort: Not reported
Age (yrs):
22.0-60.2
Mean ± SD: Controls:
34.5 ± 9.1
N: 135 controls
Location: South Korea
Study Period:
10/24/1997-8/19/1999
Cohort:
Baltimore Memory Study
Age (yrs): 50-70; Mean: 59.4
N:964
Location: Baltimore, MD
Study Period: 5/2001-9/2002
(1st study visit)
8/2002-3/2004 (2nd study
visit -tibia Pb measured)
Cohort: Not reported
Age (yrs):
12-18; Mean age ± SD:
African American cases:
15.8± 1.4
African American controls:
15.5± 1.1;
White cases: 15.7± 1.3;
White controls: 1 5.8 ± 1.1
N: 194 male youth cases;
146 male youth controls
Location: Allegheny County,
PA (cases); Pittsburgh, PA
(controls)
Study Period: 4/1996-8/1998
Cohort: Not reported
Age (yrs): Median: 41.5
N: 19 male controls
Location: Not reported
Study Period: Not reported
Prior Pb
Exposure
Control group for
occupational
exposure group
Cumulative
(lifetime)
Not reported
Control group for
occupational
exposure group
Bone Pb Bone Pb Cone.
biomarker (ug/g)
Tibia Mean ± SD
Tibia Controls:
5.8 ±7.0
Range
Tibia Controls:
-11-27
Tibia Mean ± SD
Tibia: 18.8 ±
12.4
Tibia Mean ± SD
Tibia Cases
(ppm):
All subjects: 11.0
±32.7
Africsn
American: 9.0 ±
33.6
White: 20 ± 27.5
Tibia Controls
(ppm):
All subjects: 1.5
±32.1
African
American: -1.4 ±
31.9
White: 3.5 ±32.6
Finger Median (range)
bone Finger Bone
Controls:
4 (-19-1 8)
Distribution
of Bone Pb
(ug/g)
Not reported
Tibia IQR:
11.9-24.8
Table 4 of the
paper
distributes
bone Pb by
> 25 or <25
for race, two
parental
figures, and
parent
occupation
Not reported
3-101
-------
Table 3-9 (Continued): Epidemiologic studies that provide bone Pb measurements for non-
occupationally exposed populations.
Reference
Park et al.
(2006)
Park et al.
(2009a)
Prior Pb
Study Methods Exposure
Cohort: Not reported
Normative Aging Study
Age (yrs): Mean: 72.9 ± 6.5
N: 413 males
Location: Greater Boston,
MA
Study Period:
11/14/2000-12/22/2004;
(HRV measurements taken);
1991-2002 (bone Pb
measurements taken)
Cohort: Not reported
Normative Aging Study
Age (yrs): Mean: 67.3 ± 7.2
N: 613 males
Location: Greater Boston,
|\/1 A
MA
Study Period: 8/1991 -
12/1995
Bone Pb Bone Pb Cone.
biomarker (ug/g)
Tibia Median (IQR)
Patella
Tibia: 19.0
(11-28)
Patella: 23.0
(15-34)
Estimated
Patella3 16.3
(10.4-25.8)
Tibia Median (IQR)
Patella
Tibia: 19(14-27)
Patella: 26
(18-37)
Distribution
of Bone Pb
(ug/g)
Median (IQR)
for No. of
metabolic
abnormalities:
Tibia:
0: 18.5
(10.5-23)
1: 19(11-28)
2: 19(12-26)
Patella:
0:22
(13.5-32)
1:25(16-36)
2:20(15-32)
Estimated
Patella:
0: 16.3
(10.8-24.8)
1: 17.1
(11-29.3)
2: 15.1
(9.4-22.1)
Table 1 of the
paper
distributes
tibia and
patella Pb by
genotype;
Table 2 of the
paper
distributes
tibia and
patella Pb by
number of
gene variants
3-102
-------
Table 3-9 (Continued): Epidemiologic studies that provide bone Pb measurements for non-
occupationally exposed populations.
Reference
Park et al.
(2010)
Payton et al.
(1998)
Peters et al.
(2007)
Rajan et al.
(2007)
Rajan et al.
(2008)
Study Methods
Cohort:
VA Normative Aging
Study cohort
Age (yrs): Mean: 64.9 (at
bone Pb measurement)
N: 448 males
Location: Eastern
Massachusetts
Study Period: 1991-1996
Cohort:
VA Normative Aging
Study cohort
Age (yrs): Mean: 66.8
N: 141 males
Location: Boston, MA
Study Period: 4/1993-3/1994
Cohort:
Normative Aging Study
cohort
Age (yrs): Mean: 66.9
N: 513 male cases
Location: Boston, MA
Study Period: 1991-1996
Cohort:
VA Normative Aging
Study Cohort
Age (yrs): Mean: 67.5 (at
bone scan)
N: 1075 males
Location: Boston, MA
Study Period: 1991-2002
Cohort:
VA Normative Aging
Study Cohort
Age (yrs): > 45
N: 720 males
Location: Boston, MA
Study Period: 1993-2001
Prior Pb
Exposure
Cumulative
(chronic
exposure)
Not reported
Cumulative
Not reported
Current and
cumulative
Bone Pb
biomarker
Tibia
Patella
Tibia
Patella
Tibia
Patella
Tibia
Patella
Tibia
Patella
Bone Pb Cone.
(Hg/g)
Mean ± SD
Tibia: 22.5 ±
14.2
Patella: 32.5 ±
20.4
Mean ± SD
Tibia: 22.5 ±
12.2
Patella: 31. 7 ±
19.2
Mean ± SD
Tibia: 21. 5 ±
13.4
Patella: 31. 5 ±
19.3
Mean ± SD
Tibia: 22.1 ±
13.8
Patella: 31. 4 ±
19.6
Mean ± SD
ALAD 1-1
Tibia: 21. 9 ±
13.8
Patella: 29.3 ±
19.1
ALAD 1-2/2-2
Tibia: 21. 2 ±
11.6
Patella: 27.9 ±
17.3
Distribution
of Bone Pb
(ug/g)
Tibia IQR: 15
Patella IQR:
21
Table 2 of the
paper
provides age-
adjusted
mean bone
Pb levels
(age, race,
education,
smoking
[pack-yr],
occupational
noise, noise
notch, BMI,
hypertension,
diabetes)
Not reported
Not reported
Not reported
Not reported
3-103
-------
Table 3-9 (Continued): Epidemiologic studies that provide bone Pb measurements for non-
occupationally exposed populations.
Reference
Rhodes et
al. (2003)
Roels et al.
(1994)
Rothenberg
etal.
(2002a) [as
reported in
Navas-Acien
et al. (2008)1
Shih etal.,
(2006)
Prior Pb
Study Methods Exposure
Cohort: Not reported
VA Normative Aging
Study Cohort
Age (yrs): Mean: 67.1
N: 526 males
Location: Boston, MA
Study Period:
1/1/1991-12/31/1995
Cohort: Not reported Control group for
Age (yrs)- 30-60 occupational
N: 68 males exposure group
Location: Belgium
Study Period: Not reported
Cohort: Not reported Not reported
Age (yrs): 15-44; Mean ±
SD: 31.0 ±7.7
N: 720 females
Location: Los Angeles, CA
Study Period: 6/1995-5/2001
Cohort: Not reported
Baltimore Memory Study
cohort
Bone Pb Bone Pb Cone.
biomarker (ug/g)
Tibia Mean ± SD
Patella Tibia: 21. 9 ±
13.5
Patella: 32.1 ±
19.8
Tibia Geometric Mean
(Range)
Tibia Controls:
Normotensive:
21.7
(<1 5.2-69.3)
Hypertensive:
20.2
(<1 5.2-52.9)
Total: 21. 4
(<1 5.2-69.3)
Tibia Mean ± SD
Calcaneus Tibia: 8.0 ± 11.4
Calcaneus: 10.7
± 11.9
Tibia Mean ± SD:
Tibia: 18.7 ±
11.2
Distribution
of Bone Pb
(ug/g)
No. of
participants
(%)
Tibia:
<1-15: 173
(33)
16-24: 186
(35)
25-126: 167
(32)
Patella:
<1-22: 189
(36)
23-35: 165
(31)
36-165: 172
(33)
Not reported
Tibia
quartiles:
Q1: -33.7-0.9
Q2: 1.0-8.0
Q3: 8.1-16.1
Q4: 16.2-42.5
Calcaneus
quartiles:
Q1: -30.6-3.0
Q2: 3.1-10.0
Q3: 10.1-18.7
Q4: 18.8-49.0
Not reported
Age (yrs): Mean: 59.39
N:985
Location: Baltimore, MD
Study Period: Not reported
3-104
-------
Table 3-9 (Continued): Epidemiologic studies that provide bone Pb measurements for non-
occupationally exposed populations.
Reference Study Methods
Prior Pb
Exposure
Bone Pb
biomarker
Bone Pb Cone.
(M9/9)
Distribution
of Bone Pb
(ug/g)
Stokes et al. Cohort: Not reported
H998) [as Age (yrs): 19-29 (in 1994);
reported in Mean ± SD: Cases: 24.3 ±
Shih et al. 3.13 Control: 24.2 ± 3.02
(2007)] Cases: 9 months-9 yr
(during 1/1/1974-12/31/1975)
N: 257 cases; 276 controls
Location: Silver Valley, ID;
Spokane, WA
Study Period:
7/10/1994-8/7/1994
Cumulative
(lifelong)
Environmental
(resided near Pb
smelter during
childhood)
Tibia Mean (Range):
Tibia Cases: 4.6
(-28.9-37)
Tibia Controls:
0.6 (-46.4-17.4)
Tibia
No. of Cases:
<1 ug/g:
31.5%
1-5 ug/g:
24.4%
5-10 ug/g:
22.3%
>10ug/g:
21.8%
No. of
Controls:
50.4%
1-5 ug/g:
25.6%
5-10ug/g:
19.4%
>10ug/g:
4.7%
Mean ± SD
Tibia
concentration
by age group:
Cases:
19-21: 1.47±
8.35
22-24: 4.48 ±
7.45
25-27: 4.82 ±
8.92
28-30: 6.64 ±
9.53
Controls:
19-21: 1.27±
6.60
22-24: -0.61 ±
6.19
25-27: 0.60 ±
8.60
28-30: 1.74±
6.42
Van Cohort: Not reported
Wijngaarden Age (yrs): Mean: 61.5
et al. (2009) N. 47
Location: Rochester, NY
Study Period: Not reported
Cumulative
Tibia
Calcaneus
Mean ± SD
Tibia: 2.0 ±5.2
Calcaneus: 6.1 ±
8.5
Not reported
3-105
-------
Table 3-9 (Continued): Epidemiologic studies that provide bone Pb measurements for non-
occupationally exposed populations.
Reference
Wasserman
et al. (2003)
Weisskopf et
al. (2004),
[as reported
in Shih et al.
(2QQ7)]
Weisskopf et
al. (2007a)
Prior Pb
Study Methods Exposure
Cohort: Cumulative
Yugoslavia Prospective Study (lifetime)
of Environmental Pb Environmental
Exposure (Rb smelter,
Age (yrs): 10-12 refinery, battery
N: 167 children plant)
Location: Kosovska,
Mitrovica, Kosovo,
Yugoslavia;
Pristina, Kosovo, Yugoslavia
Study Period:
5/1985-12/1986
(mother's enrollment);
1986-1999 (follow-up through
age 12yr);
Tibia Pb measured 11-13 yr
1 j
old
Cohort: Environmental
Normative Aging Study
Age (yrs): Mean ± SD: 67.4
±6.6
N: 466 males
Location: Boston, MA
Study Period: 1991-2002
Cohort: Not reported
VA Normative Aging Study
cohort
Age (yrs): Mean:
Lowest Patella quintile: 73.2;
Highest Patella quintile: 80.7
N: 31 males
Location: Boston, MA
Study Period: Bone Pb
measured: 1994-1999 Scans
performed: 2002-2004
Bone Pb Bone Pb Cone.
biomarker (ug/g)
Tibia Mean ± SD:
Tibia
Pristina: 1.36 ±
6.5
Mitrovica: 39.09
± 24.55
Tibia Median (IQR)
Patella Tibia: 19 (12,26)
Patella: 23 (15,
35)
Tibia Median (IQR)
Patella Tibia
Lowest quintile:
13(9-17)
Highest quintile:
41 (38-59)
Patella
Lowest quintile:
9 (5-15)
\s y \s i \s j
Highest quintile:
63 (43-86)
Distribution
of Bone Pb
(ug/g)
Tibia
quartiles:
Q1:
-14.4-1.85
Q2: 1.85-10.5
Q3: 10.5-35
Q4: 35-193.5
Table 3 of the
paper
distributes
tibia Pb by
sex, ethnicity,
address at
birth relative
to factory, and
maternal
education
Tibia IQR: 14
Patella IQR:
20
Table 3 of the
paper shows
mean Pb
levels across
categorical
variables (yr
of education,
smoking
status,
computer
experience,
first language
English)
Not reported
3-106
-------
Table 3-9 (Continued): Epidemiologic studies that provide bone Pb measurements for non-
occupationally exposed populations.
Reference
Weisskopf et
al. (2007b)
Weisskopf et
al. (2009)
Weisskopf et
al. (2010)
Prior Pb
Study Methods Exposure
Cohort: Concurrent and
VA Normative Aging Study cumulative
cohort
Age (yrs): Mean: 68.7
N: 1,089 males
Location: Boston, MA
Study Period: 1993-2001
Cohort: Cumulative
Normative Aging Study; (95%
white)
Age (yrs): Mean ± SD (at
Patella baseline); Tertile 1:
65.2 ±7.1;
Tertile 2: 66.5 ±6.5
Tertile 3: 70.2 ± 7.2
N: 868 males
Location: Greater Boston
area, MA
Study Period: 1991-1999
Cohort: Cumulative
BUMC, BWH, BIDMC,
HVMA, Normative Aging
Study (NAS), Harvard
Cooperative Program on
Aging (HCPOA)
Age (yrs): Mean:
Cases: 66.5; Controls: 69.4
N: 330 cases; 308 controls
Location: Boston, MA
Study Period: 2003-2007
1991 -1999 (NAS patients
bone Pb measured)
Bone Pb Bone Pb Cone.
biomarker (ug/g)
Tibia Median (IQR)
Patella Tibia: 20 (13-28)
Patella: 25
(17-37)
Tibia Mean ± SD
Patella Tibia: 21. 8 ±
13.6
Patella: 31. 2 ±
19.4
Tibia Mean ± SD:
Patella Tibia: 1 0.7 ±
12.1
Patella: 13.6±
15.9
Distribution
of Bone Pb
(ug/g)
Table 1 of the
paper shows
distribution of
Pb
biomarkers by
categories of
covariates
(age,
education,
smoking
status, alcohol
intake,
physical
activity,
computer
experience,
first language
English)
Patella
tertiles:
1: <22
2: 22-35
3: >35
Tibia
quartiles:
Q1: <3.1
Q2: 3.5-9.6
Q3: 10.0-17.0
Q4: >17.3
Patella
quartiles:
Q1: <2.7
Q2: 3.5-11.0
Q3: 11.3-20.9
Q4: >20.9
3-107
-------
Table 3-9 (Continued): Epidemiologic studies that provide bone Pb measurements for non-
occupationally exposed populations.
Reference
Weuve et al.
(2006)
Weuve et al.
(2009)
Wright et al.
(2003) [as
reported in
Shihetal.
(2QQ7)]
Prior Pb
Study Methods Exposure
Cohort: Cumulative
VA Normative Aging Study
cohort
Age (yrs): > 45
N: 720 males
Location: Boston, MA
Study Period: 1991
(measuring bone Pb levels)
End date not reported
Cohort: Recent and
Nurses' Health Study cohort cumulative
Age (yrs): 47-74
N: 587 females
Location: Boston, MA
Study Period: 1995-2005
Cohort: Environmental
Normative Aging Study
Age (yrs): Mean ± SD: 68.2
±6.9
N: 736 males
Location: Boston, MA
Study Period: 1991-1997
Bone Pb Bone Pb Cone.
biomarker (ug/g)
Tibia Median (1st-3rd
Patella quartile):
Tibia: 19(13-28)
Patella: 27
(18-39)
Tibia Mean ± SD:
Patella Tibia: 10.5 ± 9.7
Patella: 12. 6 ±
11.6
Tibia Mean ± SD:
Patella
Tibia: 22.4 ±
15.3
Patella: 29.5 ±
21.2
Distribution
of Bone Pb
(ug/g)
Table 1 of the
paper shows
distribution of
mean Pb bio-
marker levels
L-
DV
i_fjr
characteristics
of participants
(age,
education,
computer
experience,
smoking
status, alcohol
consumption,
fertile of Ca2+
intake, fertile
of physical
activity,
diabetes)
Not reported
Tibia:
Difference in
mean from
Lowest-
highest
quartile: 34.2
Patella:
Difference in
mean from
lowest-highest
quartile: 47
3-108
-------
Table 3-10 Epidemiologic studies that provide bone Pb measurements for
occupationally exposed populations.
Reference
Bleecker
etal.
(1997) [as
reported in
Shih et al.
(2007)]
Bleecker
etal.
(2007b)
Caffo et al.
(2008)
Dorsey et
al. (2006)
Study Methods
Cohort: Canada Lead Study
Age (yrs): Cumulative:
24-64
Younger: 24-43
Older: 44-64
Mean ± SD:
Cumulative: 44.1 ± 8.36
Younger: 37.2 ± 4.57
Older: 50.9 ±4.86
N: 80 males
Location: Canada
Study Period: Not Reported
Cohort: Not reported
Age (yrs): Mean: 39.7
N:61
Location: Northern Canada
Study Period: Not Reported
Cohort: Not reported
Age (yrs): Mean: 60.39
N: 513 males
Location: Delaware and
New Jersey, U.S.
Study Period: 1994-1997
(Phase 1 recruitment);
2001-2003 (Phase 2
recruitment)
Cohort: Not reported
Age (yrs): Mean: 43.4
N:652
Location: Korea
Study Period:
10/24/1997-8/19/1999
(enrolled)
Prior Pb
Exposure
Occupational
(Pb smelter
workers)
Occupational
(primary Pb
smelter workers)
Cumulative
Occupational
(Former
organolead
manufacturing
workers)
Occupational (Pb
workers)
Bone Pb
Bone Pb Concentration
biomarker (ug/g)
Tibia Mean ± SD
(Tibia):
Cumulative: 41.0
± 24.44
Younger: 35 ±
24.11
Older: 46.9 ±
23.59
Range (Tibia):
Cumulative:
-12-90
Younger: -12-80
Older: 3-90
Tibia Mean:
Tibia: 38.6
Tibia Mean ± SD:
Peak Tibia: 23.99
± 18.46
Tibia Mean ± SD:
Patella Tibia: 33.5 ± 43.4
Patella: 75.1 ±
101.1
Distribution of
Bone Pb
(ug/g)
Not reported
Not reported
Not reported
Not reported
3-109
-------
Table 3-10 (Continued): Epidemiologic studies that provide bone Pb measurements for
occupationally exposed populations.
Reference
Glenn et
al. (2003)
[as
reported in
Navas-
Acien et
al. (2008)1
Glenn et
al. (2006)
Hanninen
etal.
(1998) [as
reported in
Shih etal.
(2QQ7)]
Study Methods
Cohort: Not reported
Age (yrs): 40-70; Mean:
55.8 (baseline)
N: 496 males
Location: Eastern U.S.
Study Period:
6/1 994-6/1 996 (enrolled);
6/1998 (follow-up period
ended)
Cohort: Not reported
Age (yrs): 0-36.2 (baseline);
Mean ± SD: 41. 4 ± 9.5
(baseline)
N: 575; (76% male; 24%
female)
Location: South Korea
Study Period:
10/1997-6/2001
Cohort: Not reported
Age (yrs): Mean ± SD:
Male: 43; Female: 48
Blood Pb (max) <2.4
umol/L:
41. 7 ±9.3
Blood Pb (max) >2.4 umol/L:
46.6 ± 6.2
N: 54; (43 males, 11
females)
Location: Helsinki, Finland
Study Period: Not reported
Prior Pb
Exposure
Occupational
(Chemical
manufacturing
facility; inorganic
and organic Pb)
Cumulative and
recent
Occupational
(Pb-using
facilities)
Occupational (Pb
acid battery
factory workers)
Bone Pb Distribution of
Bone Pb Concentration Bone Pb
biomarker (ug/g) (M9/9)
Tibia Mean ± SD: Not reported
Tibia: 14.7 ±9.4
(at yr 3)
Peak Tibia: 24.3
± 18.1
Range:
Tibia: -1.6-52 (at
year 3)
Peak Tibia:
-2.2-118.8
Tibia Mean ± SD: Not reported
Tibia: 38.4 ±42.9
Tibia-Women:
Visit 1:28.2±19.7
Visit2:22.8±20.9
Tibia-Men:
Visit 1: 41. 7±47.6
Visit 2: 37.1±48.1
Tibia Mean ± SD: Not reported
Calcaneus
Tibia:
Blood Pb (max)
< 2.4 umol/L: 19.8
± 13.7
Blood Pb (max)
>2.4 umol/L: 35.3
± 16.6
Calcaneus:
Blood Pb (max)
< 2.4 umol/L: 78.6
±62.4
Blood Pb (max)
>2.4 umol/L:
100.4 ±43.1
3-110
-------
Table 3-10 (Continued): Epidemiologic studies that provide bone Pb measurements for
occupationally exposed populations.
Reference
Hsieh et
al. (2009b)
Kamel et
al. (2002):
Kamel et
al. (2005):
Kamel et
al. (2008)
Study Methods
Cohort: Not reported
Age (yrs): Mean:
Cases: 45.71
Controls: 46.06
N: 22 cases; 18 controls
Location: Not Reported
Study Period: Not reported
Cohort: Not reported
Age (yrs): 30-80
N: 109 cases; 256 controls;
(Bone samples collected
from
104 cases and 41 controls)
Location: New England
(Boston, MA)
Study Period: 1993-1996
Bone Pb
Prior Pb Bone Pb Concentration
Exposure biomarker (ug/g)
Occupational Tibia Mean ± SD
(Pb paint factory Patella Tibia
workers) Case: 61. 55 ±
30.21
Control: 18.51 ±
22.40
Patella
Case: 66.29 ±
19.48
Control: 7.14 ±
9.81
Cumulative Tibia Mean ± SE
Occupational (Pb Patella Tibia
fumes, dust, or Cases: 14.9 ± 1.6
particles) _ . , „„ „
K ' Controls: 11.1 ±
1.6
Patella
Cases: 20.5 ± 2.1
Controls: 16.7 ±
2.0
Distribution of
Bone Pb
(ug/g)
Not reported
Cases
Tibia Pb: N (%)
-7-7: 21 (20)
8-14: 35(34)
15-61:48(46)
Patella Pb: N
-4-9: 27 (26)
10-20:40(38)
21-107: 37(36)
Controls
Tibia Pb: N (%)
-7-7: 14(34)
8-14: 12(29)
15-61: 15(37)
Patella Pb: N
-4-9: 14(34)
10-20: 14(34)
21-107: 13(32)
Khalilet Cohort: 1982 Pb
al. (2009a) Occupational Study cohort
Age (yrs): Mean:
Cases: 54
Controls: 55
N: 83 cases; 51 controls
Location: Eastern
Pennsylvania
Study Period: 1982-2004
Occupational (Pb
battery plant
workers>
Tibia
Median (IQR)
Tibia
Controls: 12
(-8-32)
Not reported
3-111
-------
Table 3-10 (Continued): Epidemiologic studies that provide bone Pb measurements for
occupationally exposed populations.
Reference
Osterberg
etal.
(1997) [as
reported in
Shihetal.
(2007)]
Roels et
al. (1994)
Study Methods
Cohort: Not reported
Age(yrs): Median: 41.5
N: 38 male cases; 19 male
controls
Location: Not reported
Study Period: Not Reported
Cohort: Not reported
Age (yrs): 30-60
N: 76 male cases; 68 male
controls
Location: Belgium
Study Period: Not Reported
Prior Pb
Exposure
Occupational
(secondary Pb
smelter -
inorganic Pb)
Occupational (Pb
smelter workers)
Mean case
exposure: 18 yr
(range: 6 to 36 yr)
Bone Pb
Bone Pb Concentration
biomarker (ug/g)
Finger Median
bone Finger Bone:
High Cases: 32
Low cases: 16
Control: 4
Range
Finger Bone:
High Cases:
17-101
Low cases: -7-49
Control: -19-18
Tibia Geometric Mean
(Range)
Tibia Cases:
Normotensive:
64.0(19.6-167.1)
Hypertensive:
69.0(21.7-162.3)
Total: 65.8
(19.6-167.1)
Distribution of
Bone Pb
(ug/g)
Not reported
Not reported
Schwartz
etal.
(2000b)
[as
reported in
Shih etal.,
(2007)]
Cohort: U.S. Organolead
Study
Age (yrs): Mean ± SD:
Cases: 55.6 ± 7.4
Controls: 58.6 ± 7.0
N: 535 male cases
118 male controls
Location: Eastern U.S.
Occupational Tibia
(tetraethyl and
tetramethyl Pb
manufacturing
facility)
Tibia Controls:
Normotensive:
21.7(<15.2-69.3)
Hypertensive:
20.2(<15.2-52.9)
Total: 21. 4
(<1 5.2-69.3)
Mean ± SD
Current Tibia:
Cases: 14.4 ± 9.3
Peak Tibia:
Cases: 22.6 ±
16.5
Not reported
Study Period:
6/1994-10/1997 (enrolled);
Completed 2-4 annual
follow-up visits; Tibia Pb
taken in 3rd year
3-112
-------
Table 3-10 (Continued): Epidemiologic studies that provide bone Pb measurements for
occupationally exposed populations.
Reference
Schwartz
etal.
(2000c)
[as
reported in
Navas-
Acien et
al. (2008)1
Schwartz
etal.
(2001);
Lee et al.
(2001 a)
\*-"" ' " /
Schwartz
etal.
(2005)
Study Methods
Cohort: Not reported
Age (yrs): 41. 7-73.7
(Combined)
Mean ± SD:
Combined: 57.6 ± 7.6
Hypertensive: 60.2 ± 6.9
Nonhypertensive: 56.6 ± 7.5
N: 543 males
Location: Eastern U.S.
Study Period: 1995
(recruited); 1996-1997 (Tibia
Pb taken during the 3rd yr)
Cohort: Not reported
Age (yrs): Mean:
Exposed: 40.4
Control: 34.5
N: 803 cases; 135 controls
Location: South Korea
Study Period:
10/24/1997-8/19/1999
Cohort: Not reported
Age (yrs): Mean at 1st visit:
41.4
N:576
Location: South Korea
Study Period:
10/1997-6/2001
Prior Pb
Exposure
Occupational
(former
organolead
manufacturing
workers)
Occupational
(battery
manufacturing,
secondary
smelting, Pb
oxide
manufacturing,
car radiator
manufacturing)
Occupational
(current and
former Pb
workers)
Bone Pb
Bone Pb Concentration
biomarker (ug/g)
Tibia Mean ± SD
Tibia:
Combined: 14.4 ±
9.3
Hypertensive:
15.4 ±9.1
Nonhypertensive:
14.0 ± 9.3
Range Tibia:
Combined:
-1.6-52
Tibia Mean ± SD
Tibia
Cases: 37.1 ±
A f\ O
40.3
Control: 5.8 ± 7.0
Range:
Tibia
Cases: -7-338
Controls: -11 -27
Tibia Mean ± SD
Tibia: 38.4 ± 43
Distribution of
Bone Pb
(ug/g)
Not reported
Not reported
Tibia:
25th percentile
atV1: 14.4
75th percentile
at V1: 47.1
3-113
-------
Table 3-10 (Continued): Epidemiologic studies that provide bone Pb measurements for
occupationally exposed populations.
Reference
Stewart et
al. (1999)
[as
reported in
Shihetal.,
(2007)]
Stewart et
al. (2006)
Weaver et
al. (2008)
Study Methods
Cohort: U.S. Organolead
Study
Age (yrs): 40-70 (in 1995)
38% > 60 yrs
Mean: 58
N: 534 males
Location: Eastern U.S.
Study Period: Not Reported
Cohort: Not reported
Age (yrs): Mean: 56.1
N: 532 males
Location: Eastern U.S.
Study Period: 1994-1997;
2001-2003
Cohort: Not reported
Age (yrs): Mean ± SD: 43.3
±9.8
N. ceo
. uoz.
Location: South Korea
Study Period:
12/1999-6/2001
Prior Pb
Exposure
Occupational
(tetraethyl and
tetramethyl Pb
manufacturing
facility)
Cumulative
Occupational
(Organolead
workers - not
occupationally
exposed to Pb at
time of
enrollment)
Occupational
(Current and
former Pb
workers; plants
produced Pb
batteries, Pb
oxide, Pb crystal,
or radiators, or
were secondary
Pb smelters)
Bone Pb
Bone Pb Concentration
biomarker (ug/g)
Tibia Mean ± SD
Tibia:
Current: 14.4±
9.3
Peak: 23.7 ± 17.4
Range: Tibia
Current: -1.6-52
Peak: -2.2-105.9
Tibia Mean ± SD
Current Tibia:
14.5 ±9.6
Peak Tibia: 23.9
± 18.3
Patella Mean ± SD
Patella: 37.5 ±
41.8
Distribution of
Bone Pb
(ug/g)
Current Tibia
Pb: N (%)
<5: 77(14.2)
5-9.99: 113
(20.8)
10-14.99: 119
(21 9)
\*- /
15-19.99: 117
(21.5)
>20: 118
(21.7)
Peak Tibia Pb:
N (%)
<5:49(9.1)
5-9.99: 64
(11.8)
10-14.99: 70
(12.9)
15-19.99: 87
(16.1)
20-24.99: 79
(14.6)
25-29.99: 55
(10.2)
>30: 137
(26.1)
Not reported
Not reported
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3.4.3 Pb in Urine
Urine Pb concentrations in the U.S. general population have been monitored in the
NHANES. Data from the most recent survey (CDC. 2011 a) are shown in Table 3-11. The
geometric mean for the entire sample for the period 2007-2008 (N = 2,627) was
0.52 (ig/g creatinine (95% CI: 0.48, 0.55). The geometric means for males (N = 1,327)
and females (N = 1,300) were 0.50 (ig/g creatinine (95% CI: 0.47, 0.53) and 0.53 (ig/g
creatinine (95% CI: 0.49, 0.57), respectively.
Table 3-11 Urine Pb concentrations
Geometric
Survey Stratum Period (ug Pb/g
All 1999-2000
2001-2002
2003-2004
2005-2006
2007-2008
6-1 1 yr 1999-2000
2001-2002
2003-2004
2005-2006
2007-2008
12-1 9 yr 1999-2000
2001-2002
2003-2004
2005-2006
2007-2008
>20yr 1999-2000
2001-2002
2003-2004
2005-2006
2007-2008
Males 1999-2000
2001-2002
in the U.S.
Mean
CR)a
0.721
0.639
0.632
0.546
0.515
1.170
0.918
0.926
0.628
0.644
0.496
0.404
0.432
0.363
0.301
0.720
0.658
0.641
0.573
0.546
0.720
0.639
population.
95% Confidence
Interval
0.700, 0.742
0.603, 0.677
0.603, 0.662
0.502, 0.573
0.483, 0.549
0.975, 1.410
0.841, 1.000
0.812, 1.060
0.563, 0.701
0.543, 0.763
0.460, 0.535
0.380, 0.428
0.404, 0.461
0.333, 0.395
0.270, 0.336
0.683, 0.758
0.617, 0.703
0.606, 0.679
0.548, 0.600
0.513, 0.580
0.679, 0.763
0.607, 0.673
Number of Subjects
2,465
2,689
2,558
2,576
2,627
340
368
290
355
394
719
762
725
701
376
1,406
1,559
1,543
1,520
1,857
1,227
1,334
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Table 3-11 (Continued): Urine Pb concentrations in the U.S. population.
Survey Stratum
Females
Mexican -
Americans
Non-Hispanic
blacks
Non-Hispanic
whites
Period
2003-2004
2005-2006
2007-2008
1999-2000
2001-2002
2003-2004
2005-2006
2007-2008
1999-2000
2001-2002
2003-2004
2005-2006
2007-2008
1999-2000
2001-2002
2003-2004
2005-2006
2007-2008
1999-2000
2001-2002
2003-2004
2005-2006
2007-2008
Geometric Mean
(ug Pb/g CR)a
0.615
0.551
0.502
0.722
0.639
0.648
0.541
0.527
0.940
0.810
0.755
0.686
0.614
0.722
0.644
0.609
0.483
0.452
0.696
0.615
0.623
0.541
0.506
95% Confidence
Interval
0.588, 0.644
0.522, 0.582
0.471, 0.534
0.681, 0.765
0.594, 0.688
0.601, 0.698
0.507, 0.577
0.489, 0.568
0.876, 1.010
0.731, 0.898
0.681, 0.838
0.638, 0.737
0.521, 0.722
0.659, 0.790
0.559, 0.742
0.529, 0.701
0.459, 0.508
0.414, 0.492
0.668, 0.725
0.579, 0.654
0.592, 0.655
0.500, 0.585
0.466, 0.550
Number of Subjects
1,281
1,271
1,327
1,238
1,355
1,277
1,305
1,300
884
682
618
652
515
568
667
723
692
589
822
1,132
1,074
1,041
1,095
aValues are in |jg Pb/g creatinine (CR)
Source: Based on data from the NHANES (CDC, 2011 a)
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3.4.4
Pb in Teeth
The influence of historical Pb exposures was recently studied by Robbins et al. (2010).
Tooth enamel samples from 127 subjects born between 1936 and 1993 were analyzed for
Pb concentration and Pb isotope ratios of the tooth enamel and compared with those
parameters for sediment cores and estimates of Pb emissions from gasoline during the
years when 50% enamel formation was estimated to occur. They found that the log-
transform of tooth enamel concentration was significantly predicted by the log-transform
of Lake Erie sediment core data obtained by Graney et al. (1995) (p <0.00001) and by the
log-transform of U.S. consumption of Pb in gasoline (p <0.00001); see Figure 3-21.
Additionally, Robbins et al. (2010) found that 207Pb/206Pb was significantly predicted by
the 207Pb/206Pb observed in the Lake Erie sediment cores obtained by Graney et al. (1995)
(p O.OOOl) and for this study (p <0.0002).
100
50
25
'•&•
o-'
...A
1930 1940 1950
1960 1970
Year
1980 1990
Note: The lines and symbols on the plot represent Pb in study participant teeth (solid line), newly obtained Pb sediment Lake Erie
cores (open triangles), Pb in previously obtained Lake Erie sediment [open circles, Graney et al. (1995)1. and U.S. gasoline usage
(closed circles). All values are normalized by the peak observation for that parameter.
Source: Reprinted with permission of Elsevier Publishing, Robbins et al. (2010).
Figure 3-21 Comparison of relative temporal changes in tooth enamel Pb
concentration.
Several Brazillian studies have found increased levels of Pb in teeth in areas where Pb
sources are present. For example, Costa de Almeida et al. (2007) reported Pb
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concentration in tooth enamel among 4-6 year old kindergarteners in Sao Paulo, Brazil to
be significantly higher for children living near a Pb-acid battery processing plant in the
Baruru neighborhood compared with 4-6 year old children in other parts of the city
(non-exposed median: 206 mg/kg, N = 247; exposed median: 786 mg/kg, N = 26;
p <0.0001). Subsequent analysis revealed that 55% of 4-6 year old children from Baruru
had tooth enamel Pb concentrations greater than 600 mg/kg, forming a significant
comparison with other neighborhoods having 0-33% of 4-6 year old children with tooth
enamel Pb greater than 600 mg/kg (p <0.0001) (de Almeida et al. 2008). The authors did
not describe controlling for additional factors, such as socioeconomics or housing
conditions. Arruda-Neto et al. (2009) studied Pb in tooth samples among Sao Paulo
children to compare exposures of children age 4-12 years, living near a dam with heavy
metal sediments with those of children ages 4-13 years, living in a control area thought to
have few exposures. They observed a significant comparison (near dam: avg 1.28 ± 0.11
mg/kg, N = 50; control region: avg 0.91 mg/kg, N = 24). In a related study of Pb
measures in teeth among the general population ages 7-60 years, Arruda-Neto et
al.(2010) observed that 10-year old children had the highest teeth Pb concentrations,
which were 115% of the teeth Pb concentrations in 7-year olds. Twenty-year old subjects
had teeth Pb concentrations at roughly 50% of the 7-year olds' teeth Pb concentrations.
Tooth Pb concentrations stayed fairly constant throughout adulthood but then dropped to
just above 30% among 65-year old subjects. Note that the authors did not clarify if
average or median values were presented, nor did they adjust for potentially confounding
factors.
3.5 Empirical Models of Pb Exposure-Blood Pb Relationships
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., air Pb concentration, surface dust Pb concentration). Structural equation
modeling links several regression models together to estimate the influence of
determinants on the internal dose metric. Regression models can provide estimates of the
rate of change of blood or bone Pb concentration in response to an incremental change in
exposure level (i.e., slope factor). One strength of regression models for this purpose is
that they are empirically verified within the domain of observation and have quantitative
estimates of uncertainty 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 and are typically not generalizable
to other populations. Regression models also frequently exclude numerous parameters
that are known to influence human Pb exposures (e.g., soil and dust ingestion rates) and
3-118
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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 can be problematic.
A variety of factors may potentially affect estimates of blood Pb-air Pb slope factors.
Simultaneous changes in other (non-air) sources of Pb exposure can affect the
relationship indicated for air Pb. For example, remedial programs (e.g., community and
home-based dust control and education) may be responsible for partial blood Pb
reduction seen in some studies. The effect of remedial programs may lead to an
overestimation of declines in blood Pb due to changes in air Pb and a corresponding
positive bias in blood Pb-air Pb slopes. However, model adjustment for remedial
programs and other factors (e.g., soil Pb concentrations) may also cause a negative bias in
blood Pb-air Pb slopes. A tendency overtime for children with lower blood Pb levels to
not return for follow-up testing has been reported. The follow-up of children with higher
blood Pb levels would likely lead to an underestimation of reductions in blood Pb
following reductions in air Pb and cause a negative bias in blood Pb-air Pb slopes.
Another factor is the extent to which all the air Pb exposure pathways are captured by the
data set and its analysis. For example, some pathways (such as exposure through the diet
or surface soils) may respond more slowly to changes in air Pb than others (such as
inhalation). Additionally, some studies may include adjustments for variables that also
reflect an influence from air Pb (e.g., SES or soil Pb). Studies may also vary in the ages
of subjects, which given age-related changes in blood Pb can also influence estimates.
Many studies have utilized TSP measurements of air Pb concentrations. The sampling
efficiency of TSP samplers is affected by particle size distribution, wind speed, and wind
direction as described in Section 2.4.1. For example, especially for larger particles
(aerodynamic diameter of 20 urn or more), TSP sampling efficiency decreases with
increasing wind speed. Such effects on TSP sampling efficiency can, in areas where such
large particles are a substantial portion of airborne Pb, lead to uncertainties in the
comparability of air Pb concentrations between samples within a study and across
studies. A uniformly low bias in air Pb concentrations in a study could positively bias
estimated blood Pb-air Pb slopes for that study. Moreover, variability in TSP samples is
likely to result from temporal variation in wind speed, wind direction, and source
strength; see Sections 2.3 and 2.5. Such temporal variability would tend to increase
uncertainty and reduce the statistical strength of the relationship between air Pb and
blood Pb but may not necessarily affect the slope of this relationship. A number of factors
including those described above cause uncertainty in the magnitude of estimated blood
Pb-air Pb slope factors and may lead to both positive and negative biases in the estimates
from individual studies.
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3.5.1 Air Pb-Blood Pb Relationships in Children
The 1986 Pb AQCD (U.S. EPA. 1986a) described epidemiological studies of
relationships between air Pb and blood Pb. Of the studies examined, the aggregate blood
Pb-air Pb slope factor (when considering both air Pb and Pb in other media derived from
air Pb) was estimated to be approximately double the slope estimated from the
contribution due to inhaled air alone (U.S. EPA. 1986a).
Much of the pertinent earlier literature (e.g., prior to 1984) on children's blood Pb levels
was summarized by Brunekreef (1984). Based on meta-analysis of data from studies of
urban or industrial-urban populations in 18 different locations, Brunekreef (1984)
estimated the blood Pb-air Pb slope for children to be 0.3485 ln[(ig/dL blood Pb] per
ln[(ig/m3 air Pb] (R2 = 0.69; see Figure 3-22). This slope corresponds to an increase of
4.6 (ig/dL blood Pb per (ig/m3 air Pb at an air Pb concentration of 1.5 (ig/m3 for all
groups included in the analysis. The 1.5 (ig/m3 value is the median of the air Pb
concentrations that match the blood Pb concentrations in 96 different child populations in
Figure 3 of Brunekreef et al. (1984). taken from the Appendix to the same paper. When
the analysis was limited to child populations whose mean blood Pb concentrations were
<20 (ig/dL (N=43), the slope was 0.2159 (R2=0.33), which corresponds to an increase of
4.8 (ig/dL blood Pb per (ig/m3 air Pb at the median air concentration (0.54 (ig/m3).
Newer studies that provide estimates for the blood Pb-air Pb slope factor are described in
the sections that follow. Those studies that have at least three data points are included, as
fewer than that contributes little to the understanding of the shape of the blood Pb-air Pb
relationship. A tabular summary of the major outcomes is provided in Table 3-12. In
some studies, the blood Pb-air Pb relationship was described with a nonlinear regression
function, in which the blood Pb-air Pb slope factor varied with air Pb concentration.
Studies also varied with regard to the use of simple or multivariate regression and, for the
latter, with regard to variables included. In Table 3-12. with the exception of Ranft et al.
(2008). slopes corresponding to a central estimate of the air Pb concentrations are
provided, to represent each study. These were calculated by evaluating each regression
function at ± 0.01 (ig/m3 from the central estimate of the air Pb concentration. Air Pb
concentration ranges and central estimates varied across studies, making it difficult to
interpret comparisons based solely on the central estimates of the slope factors.
Therefore, Figure 3-23 depicts the relationship between the blood Pb-air Pb slope factor
as a function of air Pb concentration for the range of air Pb concentrations evaluated in
those studies that provided the regression equation (the central estimate is also shown).
Figure 3-23 provides a more informative picture of the extent to which slope estimates
vary (and overlap) within and between studies. The Ranft et al. (2008) study includes a
separate term for soil Pb, so the blood Pb-air Pb slope factor presented for that study
3-120
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underestimates the slope factor that would reflect all air-related pathways, since soil Pb
encompasses deposited ambient air Pb. A few studies used a log-log model that predict
an increase in the blood Pb-air Pb slope factor with decreasing air Pb concentration, and
the remainder of the studies used linear models that predict a constant blood Pb-air Pb
slope factor across all air Pb concentrations.
Table 3-12 Summary of estimated slopes for blood Pb to air Pb slope factors
in humans.
Reference
Study Methods
Model Description
Blood Pb-
AirPb
Slope3
Children Populations -Air
Location: Various countries
Years: 1974-1983
Brunekreef Subjects: Children (varying age ranges;
(1984) N>190,000)
Analysis: Meta analysis of 96 child populations
from 18 study locations
Model: Log-Log
Blood Pb: 5-76 ug/dL
(mean range for study
populations)
AirPb:0.1-24ug/m3
(mean range for study
locations)
All children:
4.6(1.5)b
Children
<20 ug/dL:
4.8 (0.54)c
Hayes et al.
(1994)
Location: Chicago, IL
Years: 1974-1988
Subjects: 0.5-5 yr (N = 9,604)
Analysis: Regression of quarterly median
blood Pb and quarterly mean air Pb
Model: Log-Log
Blood Pb: 10-28ug/dL
(quarterly median range)
Air Pb: 0.05-1.2 ug/m3
(quarterly mean range)
8.2 (0.62)d
Location: Trail, BC
Years: 1996-2001
Subjects: 0.5-5 yr, 1996-2000; 0.5-3 yr, 2001
(Estimated N = 220-460 per yr, based on
Hilts (2003) 292-536 eligible children per yr with 75-85%
participation)
Analysis: Regression of blood Pb screening
and community air Pb following upgrading of a
local smelter
Model: Linear
Blood Pb: 4.7-11.5 ug/dL
(annual geometric mean
range)
Air Pb: 0.03-1.1 ug/m3
(annual geometric mean
range)
7.0(0.48)e
Schwartz and
Pitcher (1989),
U.S. EPA
(1986a)
Location: Chicago, IL
Years: 1976-1980
Subjects: Black children, 0-5 yr (N = 5,476)
Analysis: Multivariate regression of blood Pb
with mass of Pb in gasoline (derived from
gasoline consumption data and Pb
concentrations in gasoline for the U.S.)
Model: Linear
Blood Pb: 18-27 ug/dL(mean
range)'
Air Pb: 0.36-1.22 ug/m3
(annual maximum quarterly
8.6(0.75)g
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Table 3-12 (Continued): Summary of estimated slopes for blood Pb to air Pb slope factors
in humans.
Reference
Study Methods
Model Description
Blood Pb-
AirPb
Slope3
Tripathi et al.
(2001)
Location: Mumbai, India (multiple residential
locations)
Years: 1984-1996
Subjects: 6-10 yr (N = 544)
Analysis: Regression of residential location-
specific average blood Pb and air Pb data
Model: Linear
Blood Pb: 8.6-14.4 ug/dL
(GM range for residential
locations)
AirPb:0.10-1.18|jg/m3
(GM range for residential
locations)
3.6 (0.45)'
Children Populations - Air and Soilj
Location: Germany
Years: 1983-2000 (blood Pb and air Pb),
Ranft et al. 2000-2001 (soil Pb)
(2008) Subjects: 6-11 yr (N = 843)
Analysis: Pooled multivariate regression of
5 cross-sectional studies
Model: Log-Linear
Blood Pb: 2.2-13.6 ug/dL
(5th-95th percentile)
Air Pb: 0.03-0.47 ug/m3
(5th-95th percentile)
3.2, 6.4k
Mixed Child-Adult Populations
Schwartz and
Pitcher (1989),
U.S. EPA
(1986a)
Location: U.S.
Years: 1976-1980
Subjects: NHANES II, 0.5-74 yr, whites
(N = 9,987)
Analysis: Multivariate regression of blood Pb
with mass of Pb in gasoline (derived from
gasoline consumption data and Pb
concentrations in gasoline for the U.S.)
Model: Linear
Blood Pb: 11-18ug/dLg
(mean range)'
Air Pb: 0.36-1.22 ug/m3
(annual maximum quarterly
mean)h
9.3(0.75)'
a Slope is predicted change in blood Pb (|jg/dL per |jg/m3) evaluated at ± 0.01 ug/m from central estimate of air Pb for the study
(shown in parentheses), with the exception of Ranft et al. (2008) in which the slope from the paper is provided because the
regression equation was not available. The central estimate for Brunekreef (1984) is the median of air Pb concentrations since it
was a meta-analysis; for all other studies the mean is presented. For multiple regression models, this is derived based only on air
Pb coefficient and intercept. Depending on extent to which other variables modeled also represent air Pb, this method may
underestimate the slope attributable to air pathways. In single regression models, the extent to which non-modeled factors,
unrelated to air Pb exposures, exert an impact on blood Pb that covaries with air Pb may lead to the slope presented here to over
represent the role of air Pb.
b In(PbB) = In(PbA) x 0.3485 + 2.853
0 In(PbB) = In(PbA) x 0.2159 + 2.620
dln(PbB) = ln(PbA)x0.24 + 3.17
e PbB = PbA x 7.0, see Table 3-13 for more information.
f Observed blood Pb values not provided; data are for regressed adjusted blood Pb.
9 PbB = PbA x 8.6
h Based on air Pb data for U.S. (1986 Pb AQCD) as a surrogate for Chicago.
iPbB = PbAx3.6
' Study that considered air Pb and soil Pb where the air Pb-blood Pb relationship was adjusted for soil Pb.
k Slope provided in paper with background blood Pb level of 1.5 and 3 ug/dL, respectively, and GMR of 2.55 for ambient air.
' PbB = PbA x 9.63
GM, geometric mean; GSD, geometric standard deviation; PbB, blood Pb concentration (ug/dL); PbA, air Pb concentration (ug/m3)
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60
50 -
J 40 -
Ł 30 "
T3
O
° 20 -\
CQ
10 -\
0
•All children
<20
0
5 10 15
AirPb (|jg/m3)
20
25
Note: The regression model is: (ln[|jg/dL blood Pb] = 0.3485-ln[|jg/m air Pb] + 2.85) for all children (n=96 subject groups) and
(ln[ug/dL blood Pb] = 0.2159-ln[ug/m3air Pb] + 2.62) when the sample was restricted to populations that had blood Pb
concentrations <20 ug/dL (n=44 subject groups).
Data provided from Brunekreef (1984).
Figure 3-22 Predicted relationship between air Pb and blood Pb based on a
meta-analysis of 18 studies.
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40
35
0)30
o>
-25
O)
-520
0)
Q.
O
co 15
CQ
5
0
0.0
.43.
0.5
1.0
AirPb(jjg/m3)
1.5
ABrunekreef84<20
AHayes94
*Hilts03
oSchwartz89(Chi)
OSchwartz89(US)
DTripathiOl
2.0
Note: Slopes are calculated for a change in air Pb (±0.01 ug/m3) over ranges of air Pb concentrations reported in each study (lines).
The air Pb axis is truncated at 2 ug/m3; the actual range for the Brunkreef et al. (1984) study was 0.1-6.4 ug/dL per ug/m3. The slope
axis has been truncated at 40; the actual range for the Hayes et al. (1994) study was 5-56 ug/dL per ug/m3 (the high end of the
range was estimated for the minimum annual average air Pb of 0.05 ug/m3). The two estimates for Schwartz and Pitcher (1989)
represent data for U.S. and Chicago. Models are log-log (solid lines) and linear (dotted lines). Symbols show the slope at the central
estimate of air Pb (e.g., median for Brunekreef and mean for the other studies).
Figure 3-23 Blood Pb - air Pb slopes (ug/dL per ug/m3) predicted from
epidemiologic studies.
Hilts (2003) reported child blood Pb and air Pb trends for the city of Trail, British
Columbia, over a period preceding and following installation of a new smelter process in
1997 which resulted in lower air Pb concentrations. Blood Pb data were obtained from
annual (1989-2001) surveys of children 6-60 months of age who lived within 4 km from
the smelter (N: 292-536 eligible per year, 75-85% participation). Air Pb concentrations
were obtained from high volume suspended particulate samplers placed within 2 km of
the smelter that operated 24 hours every 6th day. Data on Pb levels in air, residential soil,
interior dust, and blood for three sampling periods are summarized in Table 3-13. Based
on these data, blood Pb decreased 6.5 (ig/dL per 1 (ig/m3 air Pb and by 0.068 (ig/dL per
mg/kg soil Pb (based on linear regression with air or soil Pb as the sole independent
variable) for the entire period from 1996-2001. When considering a 9-month weighted
mean of 0.13 (ig/m3) for 2001 (3 months when the smelter was closed, 0.03 (ig/m3;
6 months when it was open, 0.18 (ig/m3), the slope is 7.0. Several uncertainties apply to
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these estimates. Potential mismatching of air Pb concentrations (often termed
misclassification) with individual blood Pb levels may have occurred as a result of air Pb
being measured within 2 km of the smelter, whereas, the blood Pb data included children
who resided >2 km from the smelter. The regression estimates were based on group mean
estimates for three sampling dates, rather than on the individual blood Pb estimates,
which included repeated measures on an unreported fraction of the sample. The limited
number of data pairs (three) constrained parameter estimates to simple regression
coefficients. Other important factors probably contributed to blood Pb declines in this
population that may have been correlated with air, soil and dust Pb levels. These factors
include aggressive public education and exposure intervention programs (Hilts et al.,
1998; Hilts. 1996). Therefore, the coefficients shown in Table 3-13 are likely to
overestimate the influence of air, dust, or soil Pb on blood Pb concentrations at this site.
Table 3-13 Environmental Pb levels and blood Pb levels in children in Trail,
British Columbia.
Date
Age (yrs)
Blood Pb (Hg/dL)
Air Pb (|jg/m3)
Soil Pb (mg/kg)
Interior Dust Pb (mg/kg)
1996a
0.5-5
11.5
1.1
844
758
1999
0.5-5
5.9
0.3
756
583
2001
0.5-3
4.7
0.13b
750 c
580 c
Regression Coefficient11
—
—
7.01 ± 0.009 (R2=1.00, p=0.001)
0.069 ± 0.008 (R2=0.99, p=0.069)
0.035 ± 0.005 (R2=0.98, p=0.097)
A new smelter process began operation in 1997. Values for air, soil and dust Pb are annual geometric means; values for blood Pb
are annual geometric means. Regression coefficients are for simple linear regression of each exposure variable on blood Pb.
a Values for air Pb, soil Pb, and interior dust Pb are actually for period of 1994-1996.
bNine month time-weighted average of 0.03 ug Pb/m3for 3 months and 0.18 ug Pb/m3for 6 months.
0 Values assumed by study authors.
d Slope for change in blood Pb d
Source: Data from Hilts (2003).
d Slope for change in blood Pb due to change in air, soil, and dust concentrations .
Ranft et al. (2008) reported a meta-analysis of five cross-sectional surveys of air and soil
Pb levels and blood Pb concentrations in children living in Duisburg, Germany. The
analysis included observations on 843 children (6-11 years of age) made during the
period 1983-2000. Children recruited in 1983 were an average of 9.1 yrs of age, whereas
children recruited in later years of the study averaged 6.3 to 6.4 yrs of age. The 1983 air
Pb concentrations were based on two monitoring stations, while a combination of
dispersion modeling and monitoring data was used in the later years to estimate Pb in PM
in a 200 meter by 200 meter grid that encompassed the city. Pb in surface soil (0-10 cm)
was measured at 145 locations in the city in 2000 and 2001. Air and soil Pb
concentrations were assigned to each participant by spatial interpolation from the
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sampling grid data to each home residence. The 5th-95th percentile ranges were
0.025-0.465 (ig Pb/m3 for air and 72-877 mg Pb/kg for soil. The results of multivariate
regression analyses were reported in terms of the relative increase (the geometric mean
blood Pb ratio, GMR) for an increase in air or soil Pb from the 5th to the 95th percentile
value. In a multivariate regression model (equation not provided) that included air and
soil Pb in the same model and adjusted for covariates, the GMR values were: 2.55 per
0.44 (ig/m3 increase in air Pb (95% CI: 2.40, 2.71, R2=0.484, p <0.001) and 1.30 per
800 mg/kg soil Pb (95% CI: 1.19, 1.43, R2 = 0.017, p <0.001). Based on the values for
R2, the regression model accounted for approximately 59% of the total variance in blood
Pb and, of this, 83% was attributed to air Pb. Values for GMR for soil Pb ranged from
1.41 to 2.89, with most recent blood Pb data (from the year 2000) yielding a value of 1.63
per 800 mg/kg increase in soil Pb. The GMR values can be converted to regression slopes
(slope = [starting blood Pbxln(GMR)]/[95th - 5th percentile air or soil Pb]) for
calculating equivalent airblood Pb ratios. The model predicts an increase of 3.2 (ig/dL
blood Pb per 1 (ig/m3 increase in air Pb at the median air Pb concentration for the study
(0.1 (ig/m3) and assuming a background blood Pb concentration of 1.5 (ig/dL. Based on
the GMR estimate of 1.63 for soil Pb, a 1,000 mg/kg increase in soil Pb would be
associated with an increase in blood Pb of 0.9 (ig/dL per mg/kg soil at the median soil Pb
concentration of 206 mg/kg and assuming a background blood Pb concentration of
1.5 (ig/dL. The degree of confounding of the GMR and estimates resulting from the air
and soil Pb correlation was not reported, although the correlation coefficient for the two
variables was 0.136 for the whole data set and 0.703 when data collected in 1983 was
omitted. Because the model also included Pb levels in soil, the blood Pb-air Pb ratio may
be underestimated since some of the Pb in soil was likely derived from air. The blood
Pb-air Pb slope does not include the portion of the soil/dust Pb ingestion pathway that
derives from air Pb, such as recently airborne Pb deposited to soil and dust which remains
available for inhalation and ingestion.
To estimate the blood Pb-air Pb ratio that included all air-related pathways, data for
median of blood Pb and air Pb among the cohort of children studied were extracted from
Table 2 in Ranft et al. (2008) for each of the five study years. The median blood Pb and
air Pb were used in regressions employing linear, log-log, and log-linear (i.e., similar to
authors' approach with In [blood Pb] against air Pb) fits. The linear model obtained was:
PbB = 12.2*PbA + 3.0 (R2 = 0.96); i.e., the linear regression produced a constant slope of
12 (ig/dL per (ig/m3. The log-log model was: In(PbB) = 0.36xm(PbA) + 2.4 (R2 = 0.90),
resulting in an inverse curve for dPbB/dPbA, versus PbA with a slope of 17 (ig/dL
per(ig/m3atPbA = 0.1 (ig/m3. The log-linear model was: In(PbB) = 2.2xPbA+ 1.2
(R2 = 0.90), resulting in an exponential curve of dPbB/dPbA, versus PbA, with a slope of
8.7 (ig/dL per (ig/m3 at PbA = 0.1 (ig/m3. Geometric mean data for blood Pb and air Pb
3-126
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are also available from Figure 1 in Ranft et al. (2008) and provide similar regression
coefficients and R2 values to those described above using median values.
Schnaas et al. (2004) analyzed data on blood Pb and air Pb concentrations during and
after the phase out of leaded gasoline use in Mexico (1986-1997) in children as part of a
prospective study conducted in Mexico City. The sample included 321 children born
during the period 1987 through 1992. Repeated blood Pb measurements were made on
each child at 6-month intervals up to age 10 years. Air Pb measurements (annual average
of quarterly means) were derived from three area monitors which represented distinct
study zones. Children were assigned to study zones based on their current address.
Associations between lifetime (across the first 10 years of life) blood Pb concentration,
air Pb concentration (mean annual for each calendar year of study) and other variables
(e.g., age, year of birth, family use of glazed pottery) were evaluated using multivariate
regression models. The largest air Pb coefficient occurred in the cohort born in 1987, who
experienced the largest decline in air Pb (from 2.8 to about 0.25 (ig/m3); the air Pb
coefficient for this group of children was 0.213 (95% CI: 0.114-0.312) In [(ig/dL blood]
per ln[(ig/m3 air]. The smallest, statistically significant air Pb coefficient occurred for the
1990 birth year cohort, who experienced a decline in air Pb from 1.5 to about 0.1 (ig/m3.
The air Pb coefficient for the 1990 cohort was 0.116 (95% CI: 0.035-0.196). Based on
these air Pb coefficients, children in the 1987 and 1990 cohorts were estimated to have
24% and 12% decreases in lifetime (across the first 10 years of life) blood Pb levels,
respectively, per natural log decrease in air Pb. Table 3-14 provides predicted blood Pb
and blood Pb-air Pb slopes as a function of age for the 1987 and 1990 cohorts. The values
in Table 3-14 are for children having complete datasets that lived in the Merced (study
region having medium air Pb concentrations) of Mexico City in medium SES families
(for the study population) that did not use clay pottery. Higher estimated blood Pb and
blood Pb-air Pb slopes than those in Table 3-14. would be predicted for low SES families
living in Xalostoc (study region having highest air Pb) that use clay pottery (e.g., 2-year-
olds predicted blood Pb of 11-14 ug/dL and blood Pb-air Pb slope of 5.7-7.2 ug/dL
per ug/m3). Conversely, lower estimates would be predicted for high SES families living
in Pedregal (study region having lowest air Pb) that did not use clay pottery (e.g., 2-year-
olds predicted blood Pb of 7-9 ug/dL and blood Pb-air Pb slope of 3.8-4.8 ug/dL
per ug/m3). The effect of air Pb on blood Pb may have been underestimated in this study
due to inclusion of location and SES terms in the regression model. It was specifically
noted by the authors that air Pb differed significantly between the locations and the
poorer residential areas were usually the more industrialized areas with higher pollution.
Hence, the inclusion of these terms may have accounted for some of the variance in blood
Pb attributable to air Pb.
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Table 3-14 Predicted blood Pb levels and blood-air slopes for Mexico City
children (1987 and 1990 cohorts).
Age (in years)
1
2
3
4
5
6
7
8
9
10
Blood Pb
7.4
8.9
8.4
7.9
7.7
6.9
6.8
6.1
5.8
5.6
-8.5a
-10.2
-9.7
-9.1
-8.8
-7.9
-7.8
-7.1
-6.7
-6.5
Blood Pb-Air Pb Slope (pg/dL
per |jg/m3)
2.1
2.6
2.4
2.3
2.2
2.0
2.0
1.8
1.7
1.6
-4.5a
-5.5
-5.2
-4.9
-4.7
-4.2
-4.2
-3.8
-3.6
-3.5
aValues are for 1990 and 1987 cohorts, respectively, at an air Pb concentration of 0.4 |jg/m3 which is the median and geometric
mean of the annual air Pb concentrations over the course of the study based on data in Figure 1 of Schnaas et al. (2004)
Source: Based on Table 4 of Schnaas et al. (2004)
For an approach that considers all the Schnaas et al. (2004) cohorts simultaneously, data
for annual geometric mean of blood Pb and air Pb were extracted from Figure 1 in
Schnaas et al. (2004). However, in employing this approach, blood Pb is confounded by
age and year because in the early years of the study, only younger children were available
and in the later years of the study, only older children contributed data. The extracted
values of the geometric mean of blood Pb and mean air Pb were used in regressions
employing linear and log-log models for comparison to other studies. The linear model
obtained was: PbB = 2.50xPbA + 5.61 (R2 = 0.84), i.e., the linear model produced a
constant slope of 2.50 (ig/dL per (ig/m3. However, inspection of the graph (not shown
here) suggested a bi-linear fit. Regression of the data over the interval 0.1-0.4 (ig/m3
produced a slope of 9.0 (ig/dL per (ig/m3 (R2 = 0.83), and regression of the data over the
interval 0.4-2.8 (ig/m3 produced a slope of 1.52 (ig/dL per (ig/m3 (R2 = 0.83). The log-log
model was: In(PbB) = 0.26xm(PbA) + 2.20 (R2 = 0.94), resulting in an inverse curve for
dPbB/dPb, versus PbA, with a slope of 4.5 (ig/dL per (ig/m3 at PbA = 0.4 (ig/m3.
Schwartz and Pitcher (1989) reported a multivariate regression analysis of associations
between U.S. gasoline Pb consumption (i.e., sales) and blood Pb concentrations in the
U.S. population during the period 1976-1980 when use of Pb in gasoline was being
phased out. Although this analysis did not directly derive a slope for the air Pb-blood Pb
relationships, other analyses have shown a strong correlation between U.S. gasoline Pb
consumption and ambient air Pb levels during this same period (U.S. EPA. 1986a).
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Therefore, it is possible to infer an air Pb-blood Pb relationship from these data. Two
sources of blood Pb data were used in Schwartz and Pitcher (1989): NHANES II
provided measurements for U.S. individuals 6 months to 74 years of age (N = 9,987
subjects) between February 1976 and February 1980, and the Chicago blood Pb screening
program provided measurements in black children aged birth to 5 years (N = 5476
subjects) for the period from 1976 to mid-1980. Observed blood Pb levels were not
provided. Gasoline Pb consumption for the U.S. was estimated as the product of monthly
gasoline sales and quarterly estimates of Pb concentrations in gasoline reported to U.S.
EPA. Based on the NHANES blood Pb data for whites, the regression coefficient for
blood Pb on the previous month's gasoline Pb usage, adjusted for age, race, sex, income,
degree of urbanization, nutrient intake, smoking, alcohol consumption, occupational
exposure, and other significant covariates was 2.14 (ig/dL blood per 100 metric tons of
gasoline Pb/day (SE=0.19, p=0.0000); the authors reported that the results for blacks
were essentially identical. Based on the Chicago blood Pb data, the age-adjusted
regression coefficient was 16.12 ((ig/dL per 1,000 metric tons gasoline Pb/quarter
[SE=1.37, p=0.0001]). When the coefficient was scaled by the ratio of Chicago's
gasoline use to the nation's and converted to units of 100 metric tons per day, the
gasoline Pb coefficient was 1.97 (ig/dL blood per 100 metric tons of gasoline Pb/day),
which is similar to the coefficient reported for the NHANES cohort. U.S. EPA (1986a)
reported data on gasoline Pb consumption (sales) and ambient air Pb levels in the U.S.
during the period 1976-1984 (Table 3-15). Based on these data, air Pb concentrations
decreased in association with gasoline Pb consumption. The linear regression coefficient
for the air Pb decrease was 0.23 (ig/m3 per 100 metric tons gasoline Pb/day (SE = 0.02,
R2 = 0.95, p <0.0001). If this regression coefficient is used to convert the blood Pb slopes
from Schwartz and Pitcher (1989). the corresponding air Pb-blood Pb slopes would be
9.3 and 8.6 (ig/dL per (ig/m3, based on the NHANES and Chicago data, respectively
(e.g., 2.14/0.23 = 9.3 and 1.97/0.23=8.6).
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Table 3-1 5
Date
1976
1977
1978
1979
1980
1981
1982
1983
1984
U.S. gasoline Pb consumption
Total Gasoline Pb
(103 metric tons/year)
171.4
168.9
153.0
129.4
78.8
60.7
59.9
52.3
46.0
and air Pb levels.
Total Gasoline Pb
(102 metric tons/day)3
4.70
4.63
4.19
3.53
2.16
1.66
1.64
1.43
1.26
AirPb
(M9/m3)b
1.22
1.20
1.13
0.74
0.66
0.51
0.53
0.40
0.36
The linear regression coefficient is 0.23 |jg/m3 air per 100 metric tons/day (SE= 0.020, R2 = 0.95, p <0.0001)
"Conversion factor is 10/365 days/year.
bAnnual mean of per site maximum quarterly means (1984 Trend Report, Available online:
http://www.epa.aov/air/airtrends/pdfs/Trends Report 1984.pdf
Source: Table 5-5, U.S. EPA 1986 Pb AQCD Q986a).
Tripathi et al. (2001) reported child blood Pb and air Pb for the city and suburbs of
Mumbai, India over the period 1984-1996. Pb-free petroleum was introduced in India
beginning in late 1996, which was outside the period of this study. Blood Pb data were
obtained from children 6-10 years of age (N = 544) who lived in 13 locations within the
Mumbai area. Air Pb concentrations were measured from high volume PM samplers
(with the majority of Pb in the respirable size range) placed at a height of 1.6 meters that
operated 24 hours. Data on Pb concentrations in air and blood are summarized in Table
3-16. An additional 16 children from two regions of Mumbai were excluded from the
analysis because of their high blood Pb levels (geometric means: 69.2 and 20.8 (ig/dL)
and proximity to industrial Pb sources with high air Pb concentrations (geometric means:
41.2 and 6.7 (ig/m3). Based on the data from residential locations presented in Table 3-16.
blood Pb increased 3.6 (ig/dL per 1 (ig/m3 air Pb (based on linear regression with air Pb
as the sole independent variable). Several uncertainties apply to these estimates,
including potential exposure misclassification since the mean air Pb concentration was
used for each suburb over the entire study period. In addition, the regression estimates
were based on group mean blood Pb estimates for the 13 sampling locations, rather than
on the individual blood Pb estimates. Ingestion of Pb-containing food was estimated in
this study, but was not considered in the regression equation for estimating blood Pb,
despite the author's conclusion that the ingestion route is important for the intake of Pb
by children in Mumbai.
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Table 3-16 Air Pb concentrations and blood Pb levels in children in Mumbai,
India.
Blood Pb (Hg/dL)
Location
Borivilli
Byculla
Deonar
Goregaon
Govandi
Jogeshwari
Khar
Parel
Sion
Thans (SS)
Vile Parle
Colaba
Vakola
N
12
117
46
21
20
20
17
168
34
37
19
12
21
GM
10.4
11.0
9.5
9.1
8.9
8.6
9.0
10.4
9.6
12.0
9.1
9.2
14.4
GSD
1.67
1.99
2.29
1.30
1.42
1.32
1.53
1.91
1.49
1.86
1.46
1.86
1.64
N
10
30
93
24
10
24
22
37
96
4
7
9
7
Air Pb (|jg/m3)
GM
0.32
0.99
0.11
0.35
0.10
0.11
0.18
0.44
0.39
1.18
0.37
0.14
1.12
GSD
1.51
1.73
3.21
1.77
1.52
2.47
3.15
1.48
1.75
1.04
1.34
1.63
1.12
The linear regression coefficient is 3.62 |jg/dL blood per |jg/m3 air (SE= 0.61, R2 = 0.76, p <0.001).
GM, geometric mean; GSD, geometric standard deviation; N, number of subjects.
Source: Data are from Tripathi et al. (2001).
Hayes et al. (1994) analyzed data collected as part of the Chicago, IL blood Pb screening
program for the period 1974-1988, following the phase-out of leaded gasoline. The data
included 9,604 blood Pb measurements in children (age: 6 months to 5 years) and
quarterly average air Pb concentrations measured at 12 monitoring stations in Cook
County, IL. Annual median blood Pb levels declined from 30 (ig/dL in 1968 to 12 (ig/dL
in 1988. During most of the years of the study, blood Pb measurements at or below
10 (ig/dL were recorded as 10 (ig/dL because of concerns over measurement accuracy of
the instrument below these levels. Quarterly median blood Pb levels declined in
association with quarterly mean air Pb concentrations. The regression model predicted a
slope of 0.24 In [(ig/dL blood] per ln[(ig/m3 air], as illustrated in Figure 3-24. This slope
corresponds to an increase of 8.2 (ig/dL blood Pb per (ig/m3 at the average annual mean
air Pb concentration of 0.62 (ig/m3. As shown in Figure 3-24. with decreasing air Pb
concentration, the slope increases. The study reports a slope of 5.6 associated with
ambient air Pb levels near 1 ug/m3 and a slope of 16 for ambient air Pb levels in the range
of 0.25 ug/m3, indicating a pattern of higher ratios with lower ambient air Pb and blood
Pb levels.
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30
25 -
20 -
O)
T3
I 10 H
5 -
0
0.0 0.2 0.4
0.6 0.8 1.0
AirPb (|jg/m3)
1.2 1.4 1.6
Note: The regression model is: (ln[|jg/dL blood Pb] = 0.24-ln[|jg/m air Pb] + 3.17).
Modified from Hayes et al. (1994).
Figure 3-24 Predicted relationship between air Pb and blood Pb based on data
from Chicago, IL in children age 0-5 years (1974-1988).
The evidence on the quantitative relationship between air Pb and blood Pb is now, as in
the past, limited by the circumstances in which the data are collected. These estimates are
generally developed from studies of populations in various Pb exposure circumstances.
The 1986 Pb AQCD (U.S. EPA. 1986a) discussed the studies available at that time that
addressed the relationship between air Pb and blood Pb, recognizing that there is
significant variability in air-to-blood ratios for different populations exposed to Pb
through different air-related exposure pathways and at different exposure levels. The
1986 Pb AQCD noted that ratios derived from studies involving higher blood and air Pb
levels are generally smaller than ratios from studies involving lower blood and air Pb
levels [see the 1986 Pb AQCD, Chapter 11, pp 99 (U.S. EPA. 1986a)1. In consideration
of this factor, slopes in the range of 3 to 5 for children generally reflected study
populations with blood Pb levels in the range of approximately 10-30 ug/dL [see
Chapter 11, pp 100 of the 1986 Pb AQCD, Table 11-36, from (Brunekreef. 1984)1. much
higher than those common in today's population. The slope of 3.6 from Tripathi et al.
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(2001) is consistent with this observation, given that the blood Pb levels were at the lower
end of this range (i.e., 10-15 ug/dL).
There are fewer studies that evaluate the air Pb-blood Pb relationship in conditions that
are more reflective of the current state. Hilts (2003) is one such study that provides
insight because the blood Pb and air Pb levels were relatively lower than those studies
mentioned above; the slope reported was 6.5, but could be as high as 7.0. Similarly,
Hayes et al. (1994) demonstrates greater slopes observed with decreasing air Pb
concentrations. These studies provide evidence that air-to-blood slopes relevant for
today's population of children would likely extend higher than the 3 to 5 range identified
in the 1986 Pb AQCD (U.S. EPA. 1986a). Due to the limited evidence, there is increased
uncertainty in projecting the magnitude of the air Pb-blood Pb relationship to ambient air
Pb concentrations below 0.2 ug/m3. In the 2008 final rule for the Pb NAAQS (73 FR
66964), with recognition of uncertainty and variability in the absolute value of an air-to-
blood relationship, the air-to-blood slopes of 5, 7, and 10 (ig/dL per (ig/m3 were utilized
in evaluating air-related IQ loss of children. Figure 3-25 illustrates the impact of these
air-to-blood slopes on the estimated change in air-related blood Pb as a function of
change in airPb.
1
.Q
Q.
-o
O
_0
CO
0)
M
re
1.5 n
1 -
0.5
Slope
(ng/dLperng/m3)
•10
7
5
0
0.05 0.1
Change in Air Pb (u,g/m3)
0.15
Figure 3-25 Effect of air-to-blood slope on estimated change in air-related
blood Pb with change in air Pb.
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3.5.2 Air Pb-Blood Pb Relationships in Occupational Cohorts
At the time of the 1986 Pb AQCD, there was a great deal of information on blood Pb
responses to air Pb exposures of workers in Pb-related occupations (U.S. EPA, 1986a).
Almost all such exposures were at air Pb exposures far in excess of typical non-
occupational exposures and typically did not account for other potential sources of Pb
exposure. The air Pb-blood Pb slopes in these studies were generally much less
(i.e., 0.03-0.2 (ig/dL per (ig/m3; pg 11-106) than those observed in children when
considering aggregate air Pb contributions (i.e., 3-5 (ig/dL per (ig/m3; pg 11-106). In
addition, the air Pb concentrations in occupational studies are typically collected at much
shorter durations (e.g., over an 8-hr workday) compared to ambient air Pb monitoring,
making it difficult to draw comparisons between occupationally and non-occupationally
exposed populations. Therefore, only a few occupational studies are presented below to
demonstrate that more recent air Pb and blood Pb levels remain much higher in these
studies compared to those conducted in the general population.
Rodrigues et al. (2010) examined factors contributing to variability in blood Pb
concentration in New England bridge painters, who regularly use electric grinders to
prepare surfaces for painting. The study included 84 adults (83 males, 1 female) who
were observed during a 2-week period in 1994 or 1995. The geometric mean air Pb
concentration obtained from personal PM samplers worn over the workday was 58 (ig/m3
(GSD 2.8), with a maximum daily value of 210 (ig/m3. Hand wipe samples were
collected and analyzed for Pb (GM = 793 (ig, GSD 3.7). Blood Pb samples were
collected at the beginning of the 2-week period (GM = 16.1 (ig/dL, GSD 1.7) and at the
end of the period (GM=18.2 (ig/dL, GD=1.6). Associations between exposure variables
and blood Pb concentrations were explored with multivariate regression models. When
the model excluded hand-wipe data, the regression coefficient for the relationship
between ln[blood Pb concentration (jig/dL)] and ln[air Pb ((ig/m3)] was 0.11 (SE = 0.05,
p = 0.03). This corresponds to a slope of 0.009 (ig/dL per (ig/m3 at the geometric mean
air Pb concentration for the study. A second regression model included hand wipe Pb
(n = 54) and yielded a regression coefficient of 0.05 (SE = 0.07, p = 0.45), which
corresponds to a slope of 0.02 (ig/dL per (ig/m3 at the geometric mean air Pb
concentration for the study.
Two other studies that examined the air Pb-blood Pb relationship in occupational settings
at higher air Pb concentrations (geometric mean of 82 and 111 (ig/m3) for Pb battery and
crystal workers, respectively (Pierre et al.. 2002; Lai etal. 1997). Blood Pb levels for the
Pb battery workers averaged 56.9 (ig/dL (SD 25.3) and for the crystal workers was
21.9 (ig/dL. Both studies employed log-log regression models, resulting in slopes of 0.49
(ig/dL per (ig/m3 (Pierre et al.. 2002) and 0.08 (ig/dL per (ig/m3 (Laietal.. 1997V
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3.5.3 Environmental Pb-Blood Pb Relationships
Empirically-based relationships between blood Pb levels and Pb intakes and/or Pb
concentrations 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 etal., 2005; U.S. EPA. 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 x ISF
Equation 3-2
PbB = E x AF x USF
Equation 3-3
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 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 tm for elimination of Pb in blood).
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The U.S. EPA Adult Lead Methodology (ALM) is an example of a slope factor model
that has had extensive regulatory use in the EPA Superfund program for assessing health
risks to adults associated with non-residential exposures to Pb in contaminated soils
(Maddaloni et al.. 2005; U.S. EPA. 1996a). The model was developed to predict maternal
and fetal blood Pb concentrations that might occur in relation to maternal exposures to
contaminated soils. The model assumes an uptake slope factor of 0.4 (ig/dL blood
per (ig/day Pb uptake. Additional discussion of slope factor models that have been used
or proposed for regulatory use can be found in the 2006 Pb AQCD (U.S. EPA. 2006b).
Previous studies included in the 2006 Pb AQCD (U.S. EPA. 2006b) explored the
relationship between blood Pb in children and environmental Pb concentrations. In a
pooled analysis of 12 epidemiologic studies, interior dust Pb loading, exterior soil/dust
Pb, age, mouthing behavior, and race were all statistically significant variables included
in the regression model for blood Pb concentration (Lanphear et al., 1998). Significant
interactions were found for age and dust Pb loading, mouthing behavior and exterior
soil/dust level, and SES and water Pb level. In a meta-analysis of 11 epidemiologic
studies, among children the most common exposure pathway influencing blood Pb
concentration in structural equation modeling was exterior soil, operating through its
effect on interior dust Pb and hand Pb (Succop et al.. 1998). Similar to Lanphear et al.
(1998). in the linear regression model, interior dust Pb loading had the strongest
relationships with blood Pb concentration. Individual studies conducted in Rochester,
NY, Cincinnati, OH, and Baltimore, MD report similar relationships between children's
blood Pb and interior dust concentrations (Lanphear and Roghmann. 1997; U.S. EPA.
1996b: Bornschein et al.. 1985).
Dixon et al. (2009) reported a multivariate analysis of associations between
environmental Pb concentrations and blood Pb concentrations, based on data collected in
the NHANES (1999-2004). The analyses included 2,155 children, age 12-60 months. The
population-weighted geometric mean blood Pb concentration was 2.03 (ig/dL
(GSD 1.03). A linear model applied to these data yielded an R2 of 40% (Table 3-17). The
regression coefficient for the relationship between ln[blood Pb concentration ((ig/dL)]
and ln[floor dust Pb concentration ((ig/ft2)] was 0.386 (SE 0.089) for "not smooth and
cleanable" surfaces (e.g., high-pile carpets) and 0.205 (SE 0.032) for "smooth and
cleanable" surfaces (e.g., uncarpeted or low-pile carpets). These coefficients correspond
to a 2.4-fold or 1.6-fold increase in blood Pb concentration, respectively, for a 10-fold
increase in floor dust Pb concentration.
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Table 3-17 Linear model relating environmental Pb
concentration in children.
Variables
Intercept
Age (in years)
Year of construction
PIR
Race/ethnicity
Country of birth
Floor surface/condition *
log floor PbD
Floor surface/condition
x (log floor PbD)2
Floor surface/condition
x (log floor PbD)3
Log windowsill PbD
Overall p-value Levels3
0.172
<0.001 Age
Age2
Age3
Age4
0.014 Intercept for missing
1990-present
1978-1989
1960-1977
1950-1959
1940-1949
Before 1 940
<0.001 Intercept for missing
Slope
<0.001 Non-Hispanic white
Non-Hispanic black
Hispanic
Other
0.002 Missing
U.S.b
Mexico
Elsewhere
<0.001 Intercept for missing
Not smooth and
cleanable
Smooth and cleanable or
carpeted
Not smooth and
cleanable
Smooth and cleanable or
carpeted
Uncarpeted not smooth
and cleanable
Smooth and cleanable or
carpeted
0.002 Intercept for missing
Slope
exposure and blood Pb
Estimate (SE)
-0.517(0.373)
2.620 (0.628)
-1.353(0.354)
0.273 (0.083)
-0.019(0.007)
-0.121 (0.052)
-0.198(0.058)
-0.196(0.060)
-0.174(0.056)
-0.207 (0.065)
-0.012(0.072)
0.000
0.053 (0.065
-0.053(0.012)
0.000
0.247 (0.035
-0.035 (0.030)
0.128(0.070)
-0.077(0.219)
0.000
0.353(0.097)
0.154(0.121
0.178(0.094)
0.386 (0.089)
0.205 (0.032)
0.023(0.015)
0.027 (0.008)
-0.020(0.014)
-0.009 (0.004)
0.053 (0.040
0.041 (0.011
p-Value
0.172
<0.001
<0.001
0.002
0.008
0.024
0.001
0.002
0.003
0.003
0.870
0.420
<0.001
<0.001
0.251
0.073
0.728
<0.001
0.209
0.065
<0.001
<0.001
0.124
0.001
0.159
0.012
0.186
<0.001
-------
Table 3-17 (Continued): Linear model relating environmental Pb exposure and blood Pb
concentration in children.
Variables Overall p-value
Home-apartment type <0.001
Anyone smoke inside the 0.015
home
Log cotinine 0.004
concentration (ng/dL)
Window, cabinet, or wall 0.045
renovation in a pre-1978
home
Levels3
Intercept for missing
Mobile home or trailer
One family house,
detached
One family house,
attached
Apartment (1-9 units)
Apartment (> 10 units)
Missing
Yes
No
Intercept for missing
Slope
Missing
Yes
No
Estimate (SE)
-0.064 (0.097
0.127(0.067)
-0.025 (0.046)
0.000
0.069 (0.060)
-0.133(0.056)
0.138(0.140)
0.100(0.040)
0.000
-0.150(0.063)
0.039(0.012)
-0.008(0.061)
0.097 (0.047)
0.000
p-Value
0.511
0.066
0.596
0.256
0.022
0.331
0.015
-
0.023
0.002
0.896
0.045
""Children N = 2,155 (age 10-60 months); R2 = 40%
blncludes the 50 states and the District of Columbia
Source: Dixon et al. (2009).
Mielke et al. (2007a) analyzed blood Pb and soil Pb concentration data collected as part
of a blood Pb screening program in New Orleans, Louisiana (2000-2005). The data set
included 55,551 blood Pb measurements for children 0-6 years of age and 5,467 soil Pb
measurements. Blood Pb and soil Pb concentrations were matched at the level of census
tracts. The association between blood Pb concentration and soil Pb concentration was
evaluated using nonparametric permutation methods. The resulting best-fit model was:
PbB = 2.038 +(0.172 xPbS05)
Equation 3-4
where PbB is the median blood Pb concentration and PbS is the median soil Pb
concentration. Although the overall association between blood Pb and soil Pb was strong
(R2=0.528), there was considerable scatter in the data. For example, at the median soil Pb
levels of 100 and 500 mg/kg, median blood Pb ranged from 2 to 8 (ig/dL and 3 to
12 (ig/dL, respectively. The resulting curvilinear relationship predicts a twofold increase
in blood Pb concentration for an increase in soil Pb concentration from 100 to 1,000 ppm
(Figure 3-26).
3-138
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O)
.Q
CL
•a
o
_o
m
Blood=2.038+0.172xSoil°-5
400 800 1200
SoilPb(ppm)
1600
2000
Note: The data set included 55,551 blood Pb measurements for children 0-6 years of age and 5,467 soil Pb measurements. Blood
Pb and soil Pb concentrations were matched at the level of census tracts (Mielke et al.. 2007a).
Figure 3-26 Predicted relationship between soil Pb concentration and blood
Pb concentration in children based on data collected in
New Orleans, Louisiana: 2000-2005.
In a subsequent re-analysis of the New Orleans (2000-2005) data, individual child blood
Pb observations were matched to census tract soil concentrations (Zahran etal. 2011).
This analysis confirmed the association between blood Pb and both soil Pb and age
reported in Mielke et al. (2007a). Regression coefficients for soil Pb (random effects
generalized least squares regression) ranged from 0.217 to 0.214 (per soil Pb05), which is
equivalent to approximately a 2-fold increase in blood Pb concentration for an increase in
soil Pb concentration from 100 to 1,000 ppm.
Several studies have linked elevated blood Pb levels to residential soil exposures for
populations living nearby industrial or mining facilities. Gulson et al. (2009) studied the
blood Pb and isotopic Pb ratios of children younger than 5-years old and adults older than
18-years old living in the vicinity of a mine producing Magellan Pb ore in western
Australia. They observed a median blood Pb level of 6.6 (ig/dL for the children, with
isotopic ratios indicating contributions from the mine ranging from 27 to 93%. A weak
but significant linear association between blood Pb level and percent Magellan Pb was
observed (R2 = 0.12, p = 0.018). Among children with blood Pb levels over 9 (ig/dL and
among adults, the isotopic ratios revealed Pb exposures from a variety of sources.
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Garavan et al. (2008) measured soil Pb and blood Pb levels among children aged 1-month
to 17.7-years old in an Irish town near a coal mine. The blood Pb measurements were
instituted as part of a screening and community education program given that the
presence of Pb had been documented in the environment. Garavan et al. (2008) found that
over 3 years of the screening period, median blood Pb levels reduced by roughly 22%
from 2.7 to 2.1 (ig/dL.
An extensive discussion of the relationships between environmental Pb levels and blood
Pb concentrations in children at the Bunker Hill Superfund Site location, a former Pb
mining and smelting site, was provided in the 2006 Pb AQCD (U.S. EPA. 2006c). In the
most recent analysis (TerraGraphics Environmental Engineering. 2004) of the data on
environmental Pb levels and child blood Pb concentrations (1988-2002), blood Pb
concentrations (annual GM) ranged from 2.6 to 9.9 (ig/dL. Environmental Pb levels
(e.g., dust, soil, paint Pb levels) data were collected at -3,000 residences, with interior
dust Pb concentrations (annual GM) ranging from -400 to 4,200 mg/kg and yard soil Pb
concentration (annual GM) ranging from -150 to 2,300 mg/kg. Several multivariate
regression models relating environmental Pb levels and blood Pb concentration were
explored; the model having the highest R2 (0.26) is shown in Table 3-18. The model
predicts significant associations between blood Pb concentration, age, interior dust, yard
soil, neighborhood soil (geometric mean soil Pb concentration for areas within 200 ft of
the residence), and community soil Pb concentration (community GM). Based on the
standardized regression coefficients, the community soil Pb concentration had the largest
effect on blood Pb concentration, followed by neighborhood soil Pb concentration,
interior dust Pb concentration, and yard soil Pb concentration (Table 3-18). The model
predicted a 1.8 (ig/dL decrease in blood Pb concentration in association with a decrease
in community soil Pb concentration from 2,000 to 1,000 mg/kg. The same decrease in
neighborhood soil Pb concentration, interior dust Pb concentration, or yard soil Pb
concentration was predicted to result in a 0.8, 0.5, or 0.2 (ig/dL decrease in blood Pb
concentration, respectively. Note that the soil Pb component of the model was similar to
that derived by Lewin et al. (1999). in which a model of blood Pb as a function of soil Pb
among 0-6 year old children living near one of four industrial sites was given as
PbB = 0.24381n(PbS) + 0.2758.
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Table 3-18 General linear model relating blood Pb concentration in children and
environmental Pb levels—Bunker Hill Superfund Site.
Parameter
Intercept
Age (yr)
ln(interior dust Pb); (mg/kg)
ln(yard soil Pb); (mg/kg)
GM soil Pb within 200 ft of residence
(mg/kg)
GM community soil Pb (mg/kg)
Coefficient
-0.1801
-0.4075
0.7288
0.2555
0.0008
0.0018
P-value
0.7916
<0.0001
<0.0001
0.0002
<0.0001
<0.0001
Standardized
Coefficient
0.00000
-0.2497
0.1515
0.0777
0.1380
0.2250
R2 = 0.264; p <0.0001; based on data from Bunker Hill Superfund Site, collected over the period 1988-2002.
GM: geometric mean; In: natural log.
Source: TerraGraphics (2004).
Malcoe et al. (2002) analyzed 1997 data on blood Pb and environmental Pb
concentrations in a representative sample of Native American and white children
(N = 224, age 1-6 years) who resided in a former Pb mining region in Ottawa County,
OK. The data set included measurements of blood Pb, yard soil Pb, residential interior
dust Pb loading, first-draw water Pb, paint Pb assessment and other behavioral (i.e., hand-
to-mouth activity, hygiene rating) and demographic variables (i.e., hygiene rating,
poverty level, caregiver education). A multivariate regression model accounted for 34%
of the observed variability in blood Pb. Yard soil Pb and interior dust Pb loading
accounted for 10% and 3% of the blood Pb variability, respectfully. The regression model
predicted a slope of 0.74 (ig/dL blood Pb per ln[(ig/g soil Pb] and a slope of 0.45 (ig/dL
blood Pb per ln[(ig/ft2] dust Pb loading.
3.6 Biokinetic Models of Pb Exposure-Blood Pb Relationships
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 are more complex than
regression models; this added complexity introduces challenges in terms of their
mathematical solution and empirical verification. However, by incorporating parameters
that can be expected to vary spatially or temporally, or across individuals or populations,
mechanistic models can be extrapolated to a wide range of exposure scenarios, including
those that may be outside of the domain of paired predictor-outcome data used to develop
the model. Exposure-intake models, a type of mechanistic models, are highly simplified
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mathematical representations of relationships between levels of Pb in environmental
media and human Pb intakes (e.g., (ig Pb ingested per day). These models include
parameters representing processes of Pb transfer between environmental media (e.g., air
to surface dust) and to humans, including rates of human contact with the media and
intakes of the media (e.g., g soil ingested per day). Intake-biokinetic models provide the
analogous mathematical representation of relationships between Pb intakes and Pb levels
in body tissues (e.g., blood Pb concentration). Biokinetic models include parameters that
represent processes of Pb transfer (a) from portals of entry into the body and (b) from
blood to tissues and excreta. Linked together, exposure-intake and intake-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, concentration, 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
epidemiologic-based information on blood Pb-response relationships to the quantification
and characterization of human health risk. These models are also critical for assessing the
potential impacts of public health programs directed at mitigation of Pb exposure or of
remediation of contaminated sites.
However, these models are not without their limitations. 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 elimination. The large
number of parameters complicates the assessment of confidence in individual parameter
values, many of which cannot be directly measured. Statistical procedures can be used to
evaluate the degree to which model outputs conform to "real-world" observations and
values of influential parameters can be statistically estimated to achieve good agreement
with observations. Still, uncertainty may be expected to remain about parameters in
complex exposure-biokinetic models. Such uncertainties need to be identified and their
impacts on model predictions quantified (i.e., sensitivity analysis or probabilistic
methods).
Modeling of human Pb exposures and biokinetics has advanced considerably during the
past several decades, although there have been relatively few developments since the
2006 Pb AQCD was published. Still in use are the Integrated Exposure Uptake Biokinetic
(IEUBK) Model for Lead in Children (U.S. EPA. 1994) and models that simulate Pb
biokinetics in humans from birth through adulthood (O'Flahertv. 1995; Leggett. 1993;
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O'Flaherty. 1993). The EPA AALM is still in development. A complete and extensive
discussion of these models can be found in the 2006 Pb AQCD (U.S. EPA. 2006b).
3.7 Summary and Conclusions
3.7.1 Exposure
Exposure data considered in this assessment build upon the conclusions of the
2006 Pb AQCD (2006b), which found that air Pb concentrations in the U.S. and
associated biomarkers of exposure to Pb have decreased substantially following
reductions in industrial point sources of Pb, and restrictions on Pb in gasoline, house-hold
paints, and solder. Pb exposure is difficult to assess because Pb has multiple sources in
the environment and passes through various media. The atmosphere is the main
environmental transport pathway for Pb, and, on a global scale, atmospheric Pb is
primarily associated with fine particulate matter, which can deposit to soil and water. In
addition to primary emission of particle-bearing or gaseous Pb to the atmosphere, Pb can
be suspended to the air from soil or dust. Air-related pathways of Pb exposure are the
focus of this assessment. In addition to inhalation of Pb from ambient air, air-related Pb
exposure pathways include inhalation and ingestion of Pb from indoor dust and/or
outdoor soil that originated from recent or historic ambient air (e.g., air Pb that has
penetrated into the residence either via the air or tracking of soil), ingestion of Pb in
drinking water contaminated from atmospheric deposition onto surface waters or from
indirect surface runoff of deposition of ambient air Pb, and ingestion of Pb in dietary
sources after uptake by plants or grazing animals. Non-air-related Pb exposures may
include occupational exposures, hand-to-mouth contact with Pb-containing consumer
goods, hand-to-mouth contact with dust or chips of peeling Pb-containing paint, or
ingestion of Pb in drinking water conveyed through Pb pipes. Pb can cycle through
multiple media prior to human exposure. Given the multitude of possible air-related
exposure scenarios and the related difficulty of constructing Pb exposure histories, most
studies of Pb exposure through air, water, and soil can be informative to this review.
Other exposures, such as occupational exposures, contact with consumer goods in which
Pb has been used, or ingestion of Pb in drinking water conveyed through Pb pipes may
also contribute to Pb body burden.
A number of monitoring and modeling techniques have been employed for environmental
Pb exposure assessment. Environmental Pb concentration data can be collected from
ambient air Pb monitors, soil Pb samples, dust Pb samples, and dietary Pb samples to
estimate human exposure. Exposure estimation error depends in part on the collection
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efficiency of these methods; collection efficiency for ambient air Pb FRM samplers is
described in Section 2.4. Models, such as the Integrated Exposure Uptake Biokinetic
(IEUBK) model, simulate human exposure to Pb from multiple sources and through
various routes including inhalation, ingestion, and dermal exposure. IEUBK model inputs
include soil Pb concentration, air Pb concentration, dietary Pb intake including drinking
water, Pb dust ingestion, human activity, and biokinetic factors. Measurements and/or
assumptions can be utilized when formulating the model inputs; errors in measurements
and assumptions thus have the potential to propagate through the exposure models.
Section 3.1 presents data illustrating potential exposure pathways. Soil can act as a
reservoir for deposited Pb emissions, and exposure to soil contaminated with deposited
Pb can occur through resuspended PM as well as hand-to-mouth contact, which is the
main pathway of childhood exposure to Pb. Airborne particles containing Pb tend to be
small (much of the distribution <10 nm) compared with Pb in soil or dust particles
(-50 nm to several hundred nm); Pb deposition to soil is described in Section 2.3. Hence,
hand-to-mouth contact with Pb-bearing soil or dust and/or tracking Pb contaminated soil
or dust into homes are more common means for human exposure to Pb. Infiltration of Pb
dust into indoor environments has been observed, and Pb dust has been shown to persist
in indoor environments even after repeated cleanings. Measurements of particle-bound
inhalation Pb exposures reported in this assessment have shown that personal inhalation
exposure measurements for air Pb concentration are typically higher than indoor or
outdoor ambient air Pb concentrations.
3.7.2 Toxicokinetics
The majority of Pb in the body is found in bone (roughly 90% in adults, 70% in children);
only about 1% of Pb is found in the blood. Pb in blood is primarily (-99%) bound to red
blood cells (RBCs). It has been suggested that the small fraction of Pb in plasma (<1%)
may be the more biologically labile and lexicologically active fraction of the circulating
Pb. The relationship between Pb in blood and plasma is pseudo-linear at relatively low
daily Pb intakes (i.e., <10 ug/kg per day) and at blood Pb concentrations <25 (ig/dL, and
becomes curvilinear at higher blood Pb concentrations due to saturable binding to RBC
proteins. As blood Pb level increases and the higher affinity binding sites for Pb in RBCs
become saturated, a larger fraction of the blood Pb is available in plasma to distribute to
brain and other Pb-responsive tissues.
The burden of Pb in the body may be viewed as divided between a dominant slow
(i.e., uptake and elimination) compartment (bone) and smaller fast compartment(s) (soft
tissues). Pb uptake and elimination in soft tissues is much faster than in bone. Pb
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accumulates in bone regions undergoing the most active calcification at the time of
exposure. During infancy and childhood, bone calcification is most active in trabecular
bone (e.g., patella); whereas, in adulthood, calcification occurs at sites of remodeling in
cortical (e.g., tibia) and trabecular bone (Aufderheide and Wittmers. 1992). A high bone
formation rate in early childhood results in the rapid uptake of circulating Pb into
mineralizing bone; however, in early childhood bone Pb is also recycled to other tissue
compartments or excreted in accordance with a high bone resorption rate (O'Flaherty.
1995). Thus, much of the Pb acquired early in life is not permanently fixed in the bone.
The exchange of Pb from plasma to the bone surface is a relatively rapid process. Pb in
bone becomes distributed in trabecular and the more dense cortical bone. The proportion
of cortical to trabecular bone in the human body varies by age, but on average is about
80% cortical to 20% trabecular. Of the bone types, trabecular bone is more reflective of
recent exposures than is cortical bone due to the slow turnover rate and lower blood
perfusion of cortical bone. Some Pb diffuses to kinetically deeper bone regions where it
is relatively inert, particularly in adults. These bone compartments are much more labile
in infants and children than in adults as reflected by half-times for movement of Pb from
bone into to the plasma (e.g., cortical half-time = 0.23 years at birth, 3.7 years at 15 years
of age, and 23 years in adults; trabecular half-time = 0.23 years at birth, 2.0 years at
15 years of age, and 3.8 years in adults) (Leggett 1993).
Evidence for maternal-to-fetal transfer of Pb in humans is derived from cord blood to
maternal blood Pb ratios. Group mean ratios range from about 0.7 to 1.0 at the time of
delivery for mean maternal blood Pb levels ranging from 1.7 to 8.6 (ig/dL. Transplacental
transfer of Pb may be facilitated by an increase in the plasma/blood Pb concentration
ratio during pregnancy. Maternal-to-fetal transfer of Pb appears to be related partly to the
mobilization of Pb from the maternal skeleton.
The dominant elimination phase of Pb kinetics in the blood, exhibited shortly after a
change in exposure occurs, has a half-life of-20-30 days. An abrupt change in Pb uptake
gives rise to a relatively rapid change in blood Pb, to a new quasi-steady state, achieved
in -75-100 days (i.e., 3-4 times the blood elimination half-life). A slower phase of Pb
clearance from the blood may become evident with longer observation periods following
a decrease in exposure due to the gradual redistribution of Pb among bone and other
compartments.
3.7.3 Pb Biomarkers
Overall, trends in blood Pb levels have been decreasing among U.S. children and adults
over the past 35 years (Section 3.4). The median blood Pb level for the entire U.S.
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population is 1.1 ug/dL and the 95th percentile blood Pb level was 3.3 ug/dL, based on
the 2009-2010 NHANES data (CDC. 2013). Among children aged 1-5 years, the median
and 95th percentiles were slightly higher at 1.2 ug/dL and 3.4 ug/dL, respectively.
Blood Pb is dependent on both the recent exposure history of the individual, as well as
the long-term exposure history that determines body burden and Pb in bone. The
contribution of bone Pb to blood Pb changes depending on the duration and intensity of
the exposure, age, and various other physiological stressors that may affect bone
remodeling (e.g., nutritional status, pregnancy, menopause, extended bed rest,
hyperparathyroidism) beyond that which normally and continuously occurs. In children,
largely due to faster exchange of Pb to and from bone, blood Pb is both an index of recent
exposure and potentially an index of body burden. In adults and children, where exposure
to Pb has effectively ceased or greatly decreased, a slow decline in blood Pb
concentrations over the period of years is most likely due to the gradual release of Pb
from bone. Bone Pb is an index of cumulative exposure and body burden. Even bone
compartments should be recognized as reflective of differing exposure periods with Pb in
trabecular bone exchanging more rapidly than Pb in cortical bone with the blood. This
difference in the compartments makes Pb in cortical bone a better marker of cumulative
exposure and Pb in trabecular bone more likely to be correlated with blood Pb, even in
adults.
Sampling frequency is an important consideration when evaluating blood Pb and bone Pb
levels in epidemiologic studies, particularly when the exposure is not well characterized.
It is difficult to determine what blood Pb is reflecting in cross-sectional studies that
sample blood Pb once, whether recent exposure or movement of Pb from bone into blood
from historical exposures. In contrast, cross-sectional studies of bone Pb and longitudinal
samples of blood Pb concentrations overtime provide more of an index of cumulative
exposure and are more reflective of average Pb body burdens overtime. The degree to
which repeated sampling will reflect the actual long-term time-weighted average blood
Pb concentration depends on the sampling frequency in relation to variability in
exposure. High variability in Pb exposures can produce episodic (or periodic) oscillations
in blood Pb concentration that may not be captured with low sampling frequencies.
Furthermore, similar blood Pb concentrations in two individuals (or populations),
regardless of their age, do not necessarily translate to similar body burdens or similar
exposure histories.
The concentration of Pb in urine follows blood Pb concentration, in that it mainly reflects
the exposure history of the previous few months and therefore, is likely a relatively poor
index of Pb body burden. There is added complexity with Pb in urine because
concentration is also dependent upon urine flow rate, which requires timed urine samples
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that is often not feasible in epidemiologic studies. Other biomarkers have been utilized to
a lesser extent (e.g., Pb in teeth).
3.7.4 Air Lead-Blood Lead Relationships
The 1986 Pb AQCD (U.S. EPA. 1986a) described epidemiological studies of
relationships between air Pb and blood Pb. Much of the pertinent earlier literature for
child populations described in the 1986 Pb AQCD was also included in a meta-analysis
by Brunekreef (1984). Based on the studies available at that time, the 1986 Pb AQCD
concluded that "the blood Pb versus air Pb slope (3 is much smaller at high blood and air
levels." This is to say that the slope (3 was much smaller for occupational exposures
where high blood Pb levels (>40 ug/dL) and high air Pb levels (much greater than
10 ug/m3) prevailed relative to lower environmental exposures which showed lower
blood Pb and air Pb concentrations (<30 ug/dL and <3 ug/m3). For those environmental
exposures, it was concluded that the relationship between blood Pb and air Pb "... for
direct inhalation appears to be approximately linear in the range of normal ambient
exposures (0.1-2.0 ug/m3)" (Chapter 1, pp 98 of the 1986 Pb AQCD). In addition to the
meta-analysis of Brunekreef (1984). more recent studies have provided data from which
estimates of the blood Pb-air Pb slope can be derived for children (Table 3-12). The range
of estimates from these studies is 4-9 ug/dL per ug/m3, which encompasses the estimate
from the Brunekreef (1984) meta-analysis. Most studies have described the blood Pb-air
Pb relationship as either log-log (Schnaas et al.. 2004; Haves et al.. 1994; Brunekreef,
1984). which predicts an increase in the blood Pb-air Pb slope with decreasing air Pb
concentration or linear (Hilts. 2003; Tripathi et al.. 2001; Schwartz and Pitcher. 1989).
which predicts a constant blood Pb-air Pb slope regardless of air Pb concentrations. These
differences may simply reflect model selection by the investigators; alternative models
are not reported in these studies. The blood Pb-air Pb slope may also be affected in some
studies by the inclusion of parameters (e.g., soil Pb) that may account for some of the
variance in blood Pb attributable to air Pb. Other factors that likely contribute to the
derived blood Pb-air Pb slope include differences in the populations examined and Pb
sources, which varied among individual studies.
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CHAPTER 4 INTEGRATED HEALTH EFFECTS OF
LEAD EXPOSURE
4.1 Introduction
This chapter summarizes, integrates, and evaluates the evidence for the broad spectrum of
health effects associated with exposure to Pb. The chapter begins (Section 4.2) with a
discussion of the evidence for the modes of action that mediate the health effects of Pb,
including those modes of action that are shared by all of the health effects evaluated in
this ISA and those modes of action that are specific to particular endpoints. Subsequent
sections comprise evaluations of the epidemiologic and toxicological evidence for the
effects of Pb exposure on health outcomes related to nervous system effects (Section 4.3).
cardiovascular effects (Section 4.4). renal effects (Section 4.5). immune effects
(Section 4.6). hematological effects (Section 4.7). and reproductive and developmental
effects (Section 4.8). Section 4.9 reviews the evidence for the effects of Pb on other
noncancer health outcomes, for which the cumulative bodies of evidence are smaller,
including those related to the hepatic system (Section 4.9.1). gastrointestinal system
(Section 4.9.2). endocrine system (Section 4.9.3). bone and teeth (Section 4.9.4). ocular
health (Section 4.9.5). and respiratory system (Section 4.9.6). Chapter 4 concludes with a
discussion of the evidence for the cancer effects of Pb (Section 4.10).
Individual sections for major outcome categories (e.g., nervous system, cardiovascular,
renal) begin with a summary of conclusions from the 2006 Pb AQCD (U.S. EPA. 2006c)
followed by an evaluation of recent (i.e., published since the completion of the
2006 Pb AQCD) studies that is intended to build upon evidence from previous reviews.
Within each of these sections, results are organized into smaller groups of related
endpoints (e.g., cognitive function, externalizing behaviors, neurodegenerative diseases)
then specific scientific discipline (i.e., epidemiology, toxicology). This chapter evaluates
evidence for both short- and long-term Pb exposures, which are defined as up to
four weeks and greater than four weeks, respectively, in animal toxicological studies and
up to one year and greater than one year, respectively, in epidemiologic studies
(Section 1.1).
Sections for each of the major outcome categories (e.g., nervous system, cardiovascular,
renal effects) conclude with an integrated summary of the assessment of evidence and
conclusions regarding causality. A determination of causality was made for a group of
related endpoints within a major outcome category (e.g., cognitive function,
hypertension) by evaluating the evidence for each endpoint group independently with the
causal framework (described in the Preamble to this ISA).
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Judgments regarding causality were made by evaluating evidence for consistency of
findings across multiple studies and the extent to which chance, confounding (i.e., bias
due to a correlation with Pb biomarker level and causal association with the outcome),
and other biases could be ruled out with reasonable confidence. Such evidence included
high quality epidemiologic studies with representative population-based groups,
prospective versus cross-sectional or ecologic design; rigorous statistical analysis
(i.e., multivariate regression) with assessment of potential confounding factors;
information on the concentration-response relationship; and supporting toxicological
evidence. Epidemiologic studies varied in the extent to which they considered potential
confounding. Because no single study considered all potential confounding factors, and
not all potential confounding factors were examined in the collective body of evidence,
residual confounding by unmeasured factors is possible. Residual confounding also is
possible by factors measured with error. However, the evidence was evaluated for the
extent to which there was examination of factors well documented in the literature to be
associated with Pb exposure and health outcomes and supporting toxicological evidence
to help minimize the undue influence of confounding bias. The biological plausibility
provided by the coherence of evidence between toxicology and epidemiology and across
a spectrum of related endpoints, including evidence for modes of action, was used as
support to address uncertainties in the epidemiologic evidence due to biases from factors
such as selective publication, recruitment or participation of subjects; reverse causality;
or confounding.
Judgments regarding causality also were made by evaluating the evidence for Pb
exposure routes and the full range of concentrations in animal toxicological and
epidemiologic studies considered relevant to this ISA, i.e., blood Pb levels up to 30
(ig/dL for most outcomes as described in Section 1.1. Studies that examined higher Pb
concentrations were evaluated particularly to inform mode of action. For epidemiologic
studies of adults that examined concurrent blood Pb levels, it is recognized that the
associations observed with health effects likely are influenced by higher past blood Pb
levels and Pb exposures (Section 3.4.1). Some data are available to inform past air Pb
concentrations and blood Pb levels in the U.S. As discussed in Section 2.2. peak U.S. use
of Pb anti-knock additives in automobile gasoline and peak industrial emissions occurred
between 1968 and 1972, and Pb in gasoline was finally banned from use in 1996. In
addition to these changes, the annual average of the maximum 3-month average air Pb
concentration for trends monitors across the U.S. decreased from 1.3 (ig/m3 in 1980 to
0.14 (ig/m3 in 2010 (Section 2.5). The mean 2010 3-month rolling average for non-source
monitors (i.e., monitors not in close proximity to specific sources) was an order of
magnitude lower than the 2010 trends site average. Studies of blood Pb during the time
period of peak air Pb concentrations in the U.S. (1968-1980) show higher U.S. blood Pb
levels than those in 2009-2010 (Section 3.4). Among subjects ages 1-74 years examined
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in U.S. NHANES II (1976-1980) by Pirkle et al. (1994). the geometric mean blood Pb
level was 12.8 ug/dL. Studies that examined blood Pb levels in various age groups
reported geometric means in adults (ages 20-79 years in 1968-1980) of 13.1-16.2 ug/dL
(Pirkle et al.. 1994; Tepper and Levin. 1975) and in children (ages 1 month-6 years in
1970-1980) of 15.0-30.4 ug/dL (Pirkle et al.. 1994; Billick et al.. 1979; Fineetal.. 1972).
These data indicate that for U.S. adults examined in the studies reviewed in this ISA, past
blood Pb levels likely were within the range considered relevant for this ISA.
4.2 Modes of Action
4.2.1 Introduction
The diverse health effects associated with Pb exposure are dependent on multiple factors,
including the concentration and duration of exposure, the particular Pb compounds
constituting the exposure, and which tissues are affected. Pb exposure is linked to
downstream health effects by various modes of action. A mode of action (MOA) is the
common set of biochemical, physiological, or behavioral responses (i.e., empirically
observable precursor steps) that can cumulatively result in the formation of negative
health outcomes. Although the effects of Pb exposure appear to be mediated through
multiple modes of action, alteration of cellular ion status (including disruption of calcium
homeostasis, altered ion transport mechanisms, and perturbed protein function through
displacement of metal cofactors) seems to be the major unifying mode of action
underlying all subsequent modes of action (Figure 4-1). This section draws information
from all of the subsequent health effects sections in Chapter 4, and identifies the major
modes of action operating at the molecular, cellular, and tissue/organ level. In turn, the
individual health effect sections bridge these MOA effects to those observed on the
organismal level. Each of the individual health effect sections includes a more detailed
description of the mechanisms specific to the individual health effect. Accordingly, this
section differs in structure and content from other health effects sections as it does not
primarily focus on the literature published since the 2006 Pb AQCD, but rather
incorporates recent information with earlier studies (which together represent the current
state of the science) on the possible modes of action of Pb. Higher concentrations of Pb
are often utilized in mode of action studies. This section includes some studies that are
conducted at concentrations greater than one order of magnitude above the upper end of
the blood Pb distribution of the general U.S. population when it is likely that the mode of
action does not differ at higher concentrations.
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Protein
Binding
(4.2.3)
kf
Oxidative
Stress
(4.2.4)
^\
/
Cell Death
and
Genotoxicity
(4.2.7)
Inflammation
(4.2.5)
Endocrine
Disruption
(4.2.6)
Note: The subsections where these MOAs are discussed are indicated in parentheses.
(Section 4.2.2: Section 4.2.3: Section 4.2.4: Section 4.2.5: Section 4.2.6: and Section 4.2.7).
Figure 4-1 Schematic representation of the relationships between the
various MOAs by which Pb exposure exerts its health effects.
4.2.2 Altered Ion Status
Physiologically-relevant metal ions (e.g., Ca2+, Mg, Zn, Fe) are known to have a
multitude of functions in biological systems, including roles as charge carriers,
intermediates in enzymatically-catalyzed reactions, and structural elements in the proper
maintenance of tertiary protein conformations (Garza et al.. 2006). It is through
disruption of these biological functions that Pb exerts its negative actions, ultimately
interfering with such tightly regulated processes as cell signaling, intracellular ion
homeostasis, ion transport, energy metabolism, and enzymatic function.
4.2.2.1 Disruption of Ca2+ Homeostasis
Calcium (Ca2+) is one of the most important carriers of cell signals and regulates virtually
all aspects of cell function, including energy metabolism, signal transduction, hormonal
regulation, cellular motility, and apoptosis (Carafoli. 2005). Ca2+ homeostasis is
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maintained through a tightly regulated balance of cellular transport and intracellular
storage (Pentvala et al.. 2010). Disruption of Ca2+ homeostasis by Pb has been observed
in a number of different cell types and cell-free environments, indicating that this is a
major mode of action for Pb-induced toxicity on a cellular level.
Ca2+ homeostasis is particularly important in bone cells, as the skeletal system serves as
the major dynamic reservoir of Ca2+ in the body (Wiemann et al., 1999; Long et al.,
1992). Bone cells also are unique in that they exist in a microenvironment that is high in
Ca2+, and potentially high in Pb concentrations. This may increase their relative exposure
to Pb and thus Pb-induced effects (Long etal. 1992). A series of studies from the
laboratory of Long, Dowd, and Rosen have indicated that exposure of cultured
osteoblastic bone cells to Pb alters intracellular Ca2+ levels ([Ca2+]j). Exposure of
osteoblasts to 1, 5, or 25 (iM Pb for 40-300 minutes resulted in prolonged increases in
[Ca2+]j of 36, 50 and 120% over baseline, respectively (Schanne et al., 1997; Schanne et
al.. 1989). Long et al. (1992) observed that exposure of osteoblasts to either 400 ng
parathyroid hormone (PTH)/mL culture medium for 1 hour or 25 (iM Pb for 20 hours
increased [Ca2+]j. Pb-exposed cells pretreated with PTH increased [Ca2+]j above
concentrations observed in either single exposure (Pb alone or PTH alone), indicating
that Pb may disrupt the ability of bone cells to respond to normal hormonal control. A
similar increase in [Ca2+]j was also observed when bone cells were co-treated with
epidermal growth factor (EOF, 50 ng/mL) plus Pb (5 (iM), versus EOF alone (Long and
Rosen. 1992). Pb-induced increases in [Ca2+]j were blocked by a protein kinase C (PKC)
inhibitor, indicating that PKC activation may serve as one mechanism by which Pb
perturbs [Ca2+]j (Schanne etal., 1997). Schirrmacher et al. (1998) also observed
alterations in Ca2+ homeostasis in osteoblasts exposed to 5 (iM Pb for 50 minutes due to
potential disruption of Ca2+ATPases. However, Wiemann et al. (1999) demonstrated that
exposure to 5 or 12.5 (iM Pb inhibited the Ca2+-release-activated calcium influx of Ca2+
independently of any inhibitory effect on Ca2+ATPases.
Ca2+ homeostasis has also been shown to be disturbed in erythrocytes exposed to Pb
(Quintanar-Escorza et al.. 2010; Quintanar-Escorza et al.. 2007; Shin et al.. 2007). In
blood samples taken from Pb-exposed workers (mean [SD] blood Pb level: 74.4
[21.9] (ig/dL), the [Ca2+]j was approximately 2.5-fold higher than that seen in
nonexposed workers (mean [SD] blood Pb level: 9.9 [2.0] (ig/dL) (Quintanar-Escorza et
al.. 2007). The increase in [Ca2+]j was associated with higher osmotic fragility and
modifications in erythrocyte shape. In a separate investigation, when erythrocytes from
10 healthy volunteers were exposed (in vitro) at concentrations of 0.2 to 6.0 (iM Pb for
24 or 120 hours, concentration-related increases in [Ca2+]j were observed across all
concentrations for both durations of exposure (Quintanar-Escorza et al.. 2010).
Subsequent exposures of erythrocytes to either 0.4 or 4.0 (iM Pb [corresponding to 10 or
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80 (ig/dL in exposed workers (Quintanar-Escorza et al.. 2007)1 for 12-120 hours resulted
in duration-related increases with durations >12 hours. Osmotic fragility (measured as
percent hemolysis) was increased in erythrocytes exposed to 0.4 (iM Pb for 24 hours.
Co-incubation with a vitamin E analog mitigated these effects, indicating that the
increase in [Ca2+]j is dependent on the oxidative state of the erythrocytes. Shin et al.
(2007) observed that incubation of human erythrocytes with 5 (iM Pb for 1 hour resulted
in a 30-fold increase in [Ca2+]j in vitro, inducing the pro-coagulant activity of exposed
erythrocytes. Induction of pro-coagulant activity in erythrocytes could lead to thrombus
formation and negatively contribute to overall cardiovascular health; whereas increased
osmotic fragility could substantially reduce erythrocyte life span and ultimately lead to
anemic conditions.
Similar to effects seen in erythrocytes, Pb has been observed to interfere with Ca2+
homeostasis in platelets and white blood cells. Dowd and Gupta (1991) observed that
1 (iM Pb (for 3.5 hours) was the lowest exposure concentration to result in increases in
[Ca2+]j in human platelets (in vitro). The observed increase in [Ca2+]j levels was
attributed to the increased influx of external Ca2+, possibly through ligand-gated Ca2+
channels. In mouse splenic lymphocytes, 1 (iM Pb was the lowest exposure concentration
found to increase [Ca2+]1) with incubation periods of 10 minutes or greater (Li et al..
2008c). These increases in [Ca2+]j appeared to be reversible as [Ca2+]j returned to
baseline after one hour. Pretreatment with a calmodulin antagonist slightly mitigated the
effects of Pb exposure, indicating a role for calmodulin in disruption of Ca2+ homeostasis
by Pb exposure in lymphocytes. In rat tail arteries exposed to 1.2 (iM Pb acetate for 1
hour, [Ca2+]j increased over controls, possibly through increased transmembrane influx of
external Ca2+ (Piccinini et al.. 1977).
Exposure of the microsomal fraction (prepared from rat brain cells) to as little as 0.25
Pb for 2 minutes resulted in increased release of Ca2+ into the culture medium (Pentyala
et al.. 2010). Further, Pb exposure also decreased the activity of microsomal Ca2+ATPase,
thus decreasing the sequestration of Ca2+ into microsomes. The results of this study
suggest that Pb-induced disruption of microsomal release and re-uptake of Ca2+ may alter
Ca2+ homeostasis, ultimately leading to altered signal transduction and neuronal
dysfunction. However, Ferguson et al. (2000) observed that [Ca2+]j was decreased in rat
hippocampal neurons in response to exposure to 0. 1 (iM Pb for 1-48 hours; although the
observed decreases were not time-dependent. The decrease in [Ca2+]j was shown to be
due to increased efflux of Ca2+ out of the neuron via a calmodulin-regulated mechanism,
possibly through stimulated Ca2+ efflux via Ca2+ATPase.
Pb exposure has been shown to disrupt [Ca2+]j levels in multiple cell types including
osteoblasts, erythrocytes, platelets, and neuronal cells. This alteration in Ca2+ homeostasis
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could potentially affect cell signaling and disrupt the normal physiological function of
these cells.
4.2.2.2 Disruption of Ion Transport Mechanisms
As described above, deregulation of Ca2+ homeostasis can result in negative effects in
multiple organ systems. Under normal conditions in the life cycle of most cells, cytosolic
concentrations of free Ca2+ fluctuate between approximately 100 to 200 nM and Ca2+ that
has entered the cell must be removed in order to maintain normal homeostatic
concentrations (Carafoli. 2005). An important component in the maintenance of Ca2+
homeostasis is transmembrane transport of Ca2+ ions via Ca2+ATPase and voltage-gated
Ca2+ channels (Carafoli. 2005). Pb has been shown to disrupt the normal movement of
Ca2+ ions, as well as other physiologically important ions through interactions with these
transport mechanisms.
Multiple studies have reported alterations in the activity of Na+/K+ATPase, Ca2+ATPase,
and Mg2+ATPases after Pb exposure in animal models. Decreases in the activity of all
three ATPases were observed in the kidneys and livers of rats exposed to 750 ppm Pb in
drinking water for 11 weeks (mean [SD] blood Pb level: 55.6 [6.3] (ig/dL) (Kharoubi et
al.. 2008a) and in erythrocytes from rats exposed to 2,000 ppm Pb in drinking water for
5 weeks (mean [SD] blood Pb level: 97.56 [11.8] jig/dL) (Sivaprasad et al.. 2003).
Increases in lipid peroxidation were seen in both studies, and the decrements in ATPase
activities may be explained by generation of free radicals in Pb-exposed animals. A
decrease in the activity of Na+/K+ATPase was observed in rabbit kidney membranes
exposed to 0.01 to 10 (iM Pb, possibly due to Pb inhibiting the hydrolytic cleavage of
phosphorylated intermediates in the K-related branch of the pump (Gramigni et al..
2009). Similar decreases in Na /K+ATPase activity were observed in brain synaptosomes
isolated from rats that were exposed to 200 ppm Pb in drinking water for 3 months (blood
Pb level: 37.8 ng/dL) (Rafalowska et al.. 1996) or 15 mg Pb/kg injected (i.p.) for 7 days
(blood Pb level: 112.5 jig/dL) (Struzynska et al.. 1997a). Inhibition ofNa+/K+ATPase
activity was also observed in primary cerebellar granule neuronal cultures obtained from
rat pups that were pre- and post-natally (to PND8) exposed to Pb (1,000 ppm Pb acetate
in dams' drinking water, resulting in blood Pb level of 4 (ig/dL) (Baranowska-Bosiacka
et al.. 20 lib). The activity of Ca2+ATPase in the sarcoplasmic reticulum of rabbits
exposed to 0.01 (iM Pb was similarly decreased (Hechtenberg and Beyersmann. 1991).
The inhibitory effect of Pb was diminished in the presence of high Mg-ATP
concentrations. The activity of generic ATPase was reported to be altered in the testes of
rat pups exposed to 300 ppm (mg/L) Pb acetate, both during lactation and in drinking
water after weaning to the age of 6, 8, 10, or 12 weeks (Liu et al.. 2008). In pregnant rats
4-7
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fed a Pb-depleted (20 ± 5 (ig/kg) or control (1 mg/kg) diet during gestation and lactation,
no difference was observed in the activity of Na+/K+ATPase and Ca2+/Mg2+ATPase in the
parental generation (Eder et al., 1990). However, the offspring (exposed viaplacental and
lactational transfer of Pb) of Pb-depleted rats displayed decreased activities in both
enzymes compared with offspring of rats with higher Pb exposures. An increase in the
Na+/K+ATPase activity was observed in rats treated (i.p.) with 20 mg/kg Pb for
14 consecutive days (Jehan and Motlag. 1995). Co-exposure of Pb with Zn and Cu
greatly attenuated the increase in ATPase activity. Although the precise mechanism was
not investigated, Navarro-Moreno et al. (2009) reported that Ca2+ uptake was diminished
in proximal renal tubule cells in rats chronically exposed to 500 ppm Pb in drinking water
for 7 months (mean [SD] blood Pb level: 43.0 [7.6] (ig/dL).
In vitro studies of ATPase activities in human erythrocyte ghosts have also shown that Pb
affects the transport of metal ions across membranes. Calderon-Salinas et al. (1999a)
observed that 1-5 x 103 \iM Pb and Ca2+ were capable of inhibiting the passive transport
of each other in human erythrocyte ghosts incubated with both cations. Subsequent
inhibition experiments indicated that both cations share the same electrogenic transport
pathway (Sakuma et al., 1984). Further study by this group (Calderon-Salinas et al.,
1999b) demonstrated that Pb can noncompetitively block the transport of Ca2+ by
inhibiting the activity of Ca2+/Mg2+ATPase at concentrations of 1-5 x 103 (iM. Mas-Oliva
(1989) demonstrated that the activity of Ca2+/Mg2+ATPase in human erythrocyte ghosts
was inhibited by incubation with 0.1-100 (iM Pb. The inhibitory action was most likely
due to direct reaction with sulfhydryl groups on the ATPase enzyme at Pb concentrations
greater than 1 (iM, but due to the action of Pb on calmodulin at lower concentrations.
Grabowska and Guminska (1996) observed that 10 (ig/dL was the lowest
Pb concentration to decrease the activity of Na+/K+ATPase in erythrocyte ghosts; activity
of Ca2+/Mg2+ATPase was less sensitive to Pb exposure, and Mg2+ATPase activity was not
affected.
Effects on ATPase activity are also observed in association with blood Pb levels in
human populations. In a study investigating ATPase activities in Pb-exposed workers in
Nigeria, Abam et al. (2008) observed that the activity of erythrocyte membrane-bound
Ca2+/Mg2+ATPase was decreased by roughly 50% in all occupational groups (range of
mean [SD] blood Pb level across nine occupational groups: 28.75 [11.31] to 42.07
[12.01] (ig/dL) compared to nonexposed controls (mean [SD] blood Pb level: 12.34
[2.44] in males and 16.85 [6.01] (ig/dL in females). Higher membrane concentrations of
Ca2+ and Mg2+ were also observed, indicating that Pb prevented the efflux of those
cations from the cell, most likely by substituting for those metals in the active site of the
ATPase. In a study of 247 mother-newborn pairs, Campagna et al. (2000) observed that
newborn cord) blood Pb (geometric mean [5th-95th percentile]: 4.8 [2.8-9.2] (ig/dL) was
-------
negatively and significantly associated with maternal blood Ca2+ pump activities;
however, newborn cord blood Pb was not significantly associated with newborn (cord)
blood Ca2+ pump activities. Newborn hair Pb (geometric mean [5th-95th percentile]: 1.1
[0.1-8.0] (ig/g) was negatively and significantly associated with both maternal and
newborn cord blood Ca2+ pump activities. In a population of 81 newborns, Huel et al.
(2008) found that newborn hair and newborn cord blood Pb levels (mean [SD] newborn
hair Pb and newborn cord blood Pb levels: 1.22 [1.41] (ig/g and 3.54 [1.72] (ig/dL) were
negatively associated with Ca2+ATPase activity in plasma membranes of erythrocytes
isolated from newborn cord blood; newborn hair Pb levels were more strongly associated
with newborn cord Ca2+ pump activity than were newborn cord blood Pb levels.
Pb has also been shown to disrupt cation transport mechanisms through direct action on
voltage-gated cation channels. Audesirk and Audesirk (1993. 1991) demonstrated that
extracellular free Pb inhibits the action of multiple voltage-gated Ca2+ channels, with free
Pb IC50 (half maximal inhibitory concentration) values of 0.7 (iM for L-type channels
and 1.3 (iM for T-type channels in neuroblastoma cells maintained in culture media, and
IC50 values as low as 0.03 (iM for L-type channels in cultured hippocampal neurons. Sun
and Suszkiw (1995) corroborated the inhibitory action of extracellular Pb on voltage-
gated Ca2+ channels, demonstrating an IC50 value of 0.3 (iM in bovine adrenal chromaffin
cells. The observed disruption of the voltage-gated Ca2+ channels most likely reflects
competition between Pb and Ca2+ for the extracellular Ca2+ binding domain of the
channel. Research by other laboratories supported these findings: Pb inhibited the action
of multiple Ca2+ channels in human embryonic kidney cells transfected with L-, N-, and
R-type channels (IC50 values of 0.38 (iM, 1.31 (iM, and 0.10 (iM, respectively) (Peng et
al.. 2002) and P-type channels in cultured hippocampal neurons at concentrations up to
3 (iM (Ujihara et al.. 1995). However, in bovine adrenal chromaffin cells, intracellular Pb
was observed to enhance Ca2+ currents through attenuation of the Ca2+ dependent
deactivation of Ca2+ channels at an EC50 (half maximal effective concentration) value of
200 (JVI, possibly through blocking the intracellular Ca2+ binding domain, or through
Ca2+ dependent dephosphorylation of the channel (Sun and Suszkiw. 1995). Recently, Pb
has also been shown to enter cells (HEK293, HeLa, and PC12 cell lines) through store-
operated Ca2+ channels (Chiu et al.. 2009; Chang et al.. 2008b). In particular, the Orail-
STIM1 complex was shown to be critical in the entry of Pb ions into cells, and increased
Pb permeation was directly related to decreased [Ca2+]j concentrations at exposure
concentrations as low as 0.1
Pb also has been found to disrupt the action of Ca2+-dependent K+ channels. Alvarez et al.
(1986) observed that Pb promoted the efflux of K+ from inside-out erythrocyte vesicles in
a concentration-dependent manner at concentrations of 1-300 (iM, either through action
on a Mg2+ modulatory site or through direct interaction with the Ca2+ binding site. Fehlau
4-9
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et al. (1989) also demonstrated Pb-induced activation of the K+ channel in erythrocytes.
However, Pb only activated the K+ channels at concentrations below 10 (iM; higher
concentrations of Pb completely inhibited channel activity, indicating the modulation of
K+ permeability is due to concentration-dependent alterations in channel gating. Silken et
al. (2001) observed that Pb activated K+ channels in erythrocytes from the marine teleost
Scorpaena porcus in a concentration-dependent manner after a 20-minute incubation;
minor loss of K+ was seen at Pb concentrations of 1-2 (iM, whereas exposure to
20-50 (iM Pb resulted in approximately 70% K+ loss. Results from competitive and
inhibitory binding assays suggest that Pb directly activates K+ channels in S. porcus.
Disruption of Neurotransmitter Release
Pb has been shown to inhibit the evoked release of neurotransmitters by inhibiting Ca2+
transport through voltage-gated channels in in vitro experiments (Cooper and Manalis.
1984; Suszkiw et al.. 1984). However, in these same experiments, concentrations of Pb
> 5 (iM were also observed to actually increase the spontaneous release of
neurotransmitters. Subsequent research by other groups affirmed that Pb demonstrates
Ca2+-mimetic properties in enhancing neurotransmitter release from cells in the absence
of Ca2+ and Ca2+-induced depolarization. Tomsig and Suszkiw (1993) reported that Pb
exposure induced the release of norepinephrine (NE) from bovine adrenal chromaffin
cells, and was considerably more potent (as measured by half-maximal metal-dependent
release [K05]) than was Ca2+ (K05 of 4.6 x 10'3 jiM for Pb versus 2.4 jiM for Ca2+).
Activation of PKC was observed to enhance the Pb-induced release of NE (Tomsig and
Suszkiw. 1995). Westerink and Vijverberg (2002) observed that Pb acted as a high
affinity substitute for Ca2+, and triggered enhanced catecholamine release from PC 12
cells at 10 (iM in intact cells and 0.03 (iM in permeabilized cells. The suppression of
Ca2+-evoked release of neurotransmitters combined with the ability of Pb to enhance
spontaneous releases could result in higher noise observed in the synaptic transmission of
nerve impulses in Pb-exposed animals.
In rats exposed to Pb at concentrations of 1,000-10,000 ppm in drinking water beginning
at gestational days GDI5-GDI6 and continuing to postnatal days PND120, decreases in
total K+-stimulated hippocampal gamma aminobutyric acid (GABA) release were seen at
exposure levels of 1,000-5,000 ppm (range of mean [SD] blood Pb levels: 26.8 [1.3] -
61.8 [2.9] (ig/dL) (Lasley and Gilbert. 2002). Maximal effects were observed at
2,000 ppm Pb in drinking water, but effects were less evident at 5,000 ppm and were
absent at 10,000 ppm. In the absence of Ca2+, K+-induced GABA release was increased
with the two highest Pb exposure concentrations, suggesting a Pb-induced enhancement
of K+-evoked release of GABA. The authors suggest that this pattern of response
indicates that Pb is a potent suppressor of K+-evoked release at low concentrations, but a
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Ca2+ mimic in regard to independently inducing exocytosis and evoking neurotransmitter
release at higher concentrations (Laslev and Gilbert. 2002). Suszkiw (2004) reports that
augmentation of spontaneous release of neurotransmitters may involve Pb-induced
activation of CaMKII-dependent phosphorylation of synapsin I or direct activation of
synaptotagmin I. Further, Suszkiw (2004) suggests that unlike the intracellularly
mediated effects of Pb on spontaneous release of neurotransmitters, Pb-induced inhibition
of evoked transmitter releases is largely due to extracellular blockage of the voltage-gated
Ca2+ channels.
In summary, Pb has been shown to disrupt ion transport mechanisms in toxicological and
epidemiologic studies. Specific mechanisms disrupted include various cation-specific
ATPases and voltage-gated cation channels. Alterations in ion transport functions have
also been shown to disrupt neurotransmitter release in both in vivo and in vitro
experiments.
4.2.2.3 Displacement of Metal Ions and Perturbed Protein
Function
The binding of metal ions to proteins causes specific changes in protein shape, and these
conformational changes may alter specific cellular function of many proteins (Kirberger
and Yang. 2008). Metal binding sites on proteins are generally ion-specific and are
influenced by multiple factors, including binding geometries, ligand preferences, ionic
radius, and metal coordination numbers (Kirberger and Yang. 2008; Garza etal.. 2006).
The coordination chemistry that normally regulates metal-protein binding makes many
proteins particularly susceptible to perturbation from Pb, as it is able to function with
flexible coordination numbers and can bind multiple ligands (Kirberger and Yang. 2008;
Garza et al. 2006). However, due to differences in its physical properties, Pb induces
abnormal conformational changes when it binds to proteins (Kirberger and Yang. 2008;
Bitto et al.. 2006; Garza et al.. 2006; Magyar et al.. 2005). and these structural changes
elicit altered protein function. It is known that [Ca2+]j is an important second messenger
in cell signaling pathways, and operates by binding directly to and activating proteins
such as calmodulin and PKC (Goldstein. 1993). Alterations in the functions of both of
these proteins due to direct interaction with Pb have been well documented in the
literature.
PKC is a family of serine/threonine protein kinases critical for cell signaling and
important for cellular processes, including growth and differentiation (Goldstein. 1993).
PKC contains a "C2" Ca2+-binding domain and requires binding of the cation, as well as
the presence of diacylglycerol and phospholipids, for proper cellular activity (Garza et
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al.. 2006). Markovac and Goldstein (1988b) observed that, in the absence of Ca2+,
exposure to 10"6 (iM concentrations of Pb for 5 minutes directly activated PKC purified
from rat brains. The activation of PKC by Pb was more potent than was Ca2+-dependent
activation by five orders of magnitude. Long et al. (1994) affirmed these findings,
reporting that Pb had a Kact 4,800 times smaller than that of Ca2+ (5.5 x 10"5 (iM versus
25 (iM, following a 3 minute exposure). However, Ca2+ had a higher maximal activation
of PKC than did Pb. This possibly indicates the presence of multiple Ca2+-binding sites
on the protein, and that Pb may bind the first site more efficiently than does Ca2+, but not
subsequent sites. Tomsig and Suszkiw (1995) further demonstrated the ability of Pb to
activate PKC in bovine adrenal chromaffin cells incubated with 10"6 (iM concentrations
of Pb for 10 minutes but also reported that activation of PKC by Pb was only partial
(approximately 40% of the maximum activity induced by Ca2+) and tended to decrease at
concentrations >1 x 10"3 (iM.
Contrary to the above findings, Markovac and Goldstein (1988a) observed that Pb and
Ca2+ activated PKC at equivalent concentrations and efficacies when broken cell
preparations of rat brain microvessels were incubated with either cation for 45 minutes.
However, when PKC activation was investigated in whole vessel preparations, no
activation was observed, but PKC did become redistributed from the cytosolic to the
particulate fraction after centrifugation. This suggests that Pb redistributes PKC at (iM
concentrations, but does not activate the protein in brain microvessels. In human
erythrocytes exposed to Pb acetate for 60 minutes, the amount of PKC found in
erythrocyte membranes and total PKC activity was increased at concentrations greater
than 0.1 (iM (Belloni-Olivi et al., 1996). The observation that neither Ca2+ nor
diacylglycerol concentrations were increased due to Pb exposure, indicates that
Pb-induced activation of PKC is due to direct interaction with the protein. Pb-induced
alterations in PKC have also been observed in other tissues, including increased activity
in rabbit mesenteric arteries at 10"6 (iM concentrations of Pb (Watts et al.. 1995; Chai and
Webb. 1988) and human erythrocytes from Pb-exposed workers (range of blood Pb
levels: 5.4 to 69.3 (ig/dL) (Hwang et al., 2002). and decreased activity in mouse
macrophages and the rat brain cortex at (iM concentrations (Murakami et al.. 1993; Lison
etal.. 1990).
Calmodulin is another important protein essential for proper Ca2+-dependent cell
signaling. Calmodulin contains an "EF-hand" Ca2+ binding domain, which is dependent
on the cation for proper activity (Garza et al., 2006). Calmodulin regulates events as
diverse as cellular structural integrity, gene expression, and maintenance of membrane
potential (Vetter and Leclerc. 2003; Saimi and Kung. 2002). Habermann et al. (1983)
observed that exposure to Pb altered numerous cellular functions of calmodulin,
including activation of calmodulin-dependent phosphodiesterase activity after 10 minutes
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incubation (minimal activation at 0.1 (JVI, EC50 = 0.5-1.0 (JVI), stimulation of brain
membrane phosphorylation at Pb concentrations greater than 0.4 (iM after 1 minute
incubation, and increased binding of calmodulin to brain membranes at Pb concentrations
greater than 1 (iM after 10 minutes incubation. Habermann et al. (1983) reported that the
affinity of Pb for Ca2+-binding sites on calmodulin was approximate to that of Ca2+ itself
(Kd ~20 (JVI), whereas Richardt et al. (1986) observed that Pb was slightly more potent
than Ca2+ was at binding calmodulin (IC50 = 11 and 26 (JVI, respectively). Both studies
indicated that Pb was much more effective at binding to calmodulin than was any other
metal cation investigated (e.g., Hg, Cd, Fe). Kern et al. (2000) observed that Pb was more
potent in binding to, and affecting conformational changes in, calmodulin compared to
Ca2+ (EC50 values of 4-5.5 x 10'4 jiM [threshold = 1 x 10'4 jiM] and 0.45-0.5 jiM
[threshold = 0.1 (iM], respectively). Pb, in the absence of Ca2+, was also observed to
activate calmodulin-dependent cyclic nucleotide phosphodiesterase activity at much
lower concentrations compared to Ca2+ (EC50 value 4.3 x 10"4 (iM [threshold = 3 x
10"4 (iM] versus EC50 1.2 x 10"3 (iM [threshold = 0.2 (iM; 50 minute incubation]). When
incubated with physiological concentrations of Ca2+, Pb induced phosphodiesterase
activity at concentrations as low as 5 x io~5 (iM. Pb activated calcineurin, a
Ca2+-dependent phosphatase with widespread distribution in the brain and immune
system, at threshold concentrations as low as 2 x 10"5 (iM in the presence of Ca2+
(incubation time = 30 minutes), but inhibited its activity at concentrations greater than 2
x 10"4 (iM (Kern and Audesirk. 2000). Thus, 10"6 (iM concentrations of intracellular Pb
appear to amplify the activity of calmodulin and thus can be expected to alter intracellular
Ca2+ signaling in exposed cells (Kern et al.. 2000). Mas-Oliva (1989) observed that
low-exposure (<1 (iM, 20 minute incubation) stimulatory effects of Pb exposure on the
activity of Ca2+/Mg2+ATPase was due to Pb binding to calmodulin and subsequent
activation of the ion pore. Ferguson et al. (2000) observed that exposure of rat
hippocampal neurons to Pb for 1 to 48 hours resulted in increased activation of a
calmodulin-dependent Ca2+ extrusion mechanism.
Pb has also been observed to alter the activity of other proteins that rely on Ca2+ binding
for normal cellular function. Osteocalcin is a matrix protein important in bone resorption,
osteoclast differentiation, and bone growth; and has three Ca2+-binding sites (Dowd et al..
2001). Incubation of osteocalcin in solution with Ca2+ and Pb resulted in the competitive
displacement of Ca2+ by Pb (Dowdet al., 1994). Pb was found to bind to osteocalcin
more than 1,000 times more tightly than was Ca2+ (Kd = 1.6 x 10"2 (iM versus 7.0 (iM,
respectively), and analysis with nuclear magnetic resonance (NMR) indicated that Pb
induced similar, though slightly different, secondary structures in osteocalcin, compared
to Ca2+. The authors hypothesized that the observed difference in Pb-bound osteocalcin
structure may explain previous findings in the literature that Pb exposure reduced
osteocalcin adsorption to hydroxyapatite (Dowd et al., 1994). Further research by this
4-13
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group also found that Pb binded osteocalcin approximately 10,000-times more tightly
than did Ca2+ (Kd = 8.5 x 10~2 jiM versus 1.25 x 103 jiM, respectively) (Dowd et al..
2001). However, the authors reported that Pb exposure actually caused increased
hydroxyapatite adsorption at concentrations 2-3 orders of magnitude lower than that seen
with Ca2+. Additionally, Pb can displace Ca2+ in numerous other Ca2+-binding proteins,
such as proteins important in muscle contractions, renal Ca2+ transport and
neurotransmission, including troponin C, parvalbumin, CaBP I and II, phospholipase A2,
and synaptotagmin I, at concentrations as low as the 10"3 (iM range (Bouton etal.. 2001;
Osterode and Ulberth. 2000; Richardt et al.. 1986).
Pb can displace metal cations other than Ca2+ that are requisite for protein function. One
of the most researched targets for molecular toxicity of Pb is the second enzyme in the
heme synthetic pathway, aminolevulinic acid dehydratase (ALAD). ALAD contains four
Zn-binding sites, and all four need to be occupied to confer full enzymatic activity
(Simons. 1995). ALAD has been identified as the major protein binding target for Pb in
human erythrocytes (Bergdahl et al., 1997a). and inhibition of the enzyme is found in the
erythrocytes of Pb-exposed workers and adolescents in association with blood Pb levels
(blood Pb levels >10 ng/dL) (Ahamed et al.. 2006; Ademuviwa et al.. 2005b). in lysed
human erythrocytes exposed to Pb in vitro for 60 minutes (K; = 7 x 10"6 (iM) (Simons.
1995). and in erythrocytes of rats exposed to 25 mg/kg Pb once a week for 4 weeks
(mean [SD] blood Pb level: 6.56 [0.98] (ig/dL) (Lee et al.. 2005V Additional experiments
indicated that lower concentrations of Zn result in greater inhibition of enzyme activity
by Pb, suggesting a competitive inhibition between Zn and Pb at a single site (Simons.
1995).
Zn-binding domains are also found in transcription factors and proteins necessary for
gene expression, including GATA proteins and transcription factors TFIIIA, Spl, and
Erg-1 (Ghering etal.. 2005; Huang et al.. 2004; Reddv and Zawia. 2000; Hanas et al..
1999; Zawia etal.. 1998). Pb was found to form tight complexes with the cysteine
residues in GATA proteins (Pb stoichiometric stability constants of [CF|3iPb] = 6.4
(± 2.0)x 109 M'1 for single C-terminal GATA Zn finger from chicken and DF(32pb2 = 6.3
(± 6.3) x 1019 M"2 for double-GATA Zn finger from human), and was able to displace
bound Zn from the protein under physiologically relevant conditions (Ghering et al..
2005). Once Pb was bound to GATA proteins, they displayed decreased ability to bind to
DNA (Pb concentrations > 1.25 (iM) and activate transcription. Pb at a minimum
concentration of 10 (iM also binds to the Zn domain of TFIIIA, inhibiting its ability to
bind DNA (Huang et al.. 2004; Hanas etal.. 1999). Huang et al. (2004) also reported that
exposure to Pb caused the dissociation of TFIIIA-DNA adducts and using NMR
spectroscopy, found that altered TFIIIA activity was the result of a Pb-induced abnormal
protein conformation.
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Pb exposure modulated the DNA-binding profiles of the transcription factors Spl and
Erg-1 in rat pups exposed to 2,000 ppm Pb acetate via lactation, resulting in a shift in
DNA-binding toward early development (i.e., the first week following birth) (Reddy and
Zawia. 2000; Zawiaet al.. 1998). The shifts in Spl DNA-binding profiles were shown to
be associated with abnormal expression of genes related to myelin formation
(Section 4.2.7.5). Further mechanistic research utilizing a synthetic peptide containing a
Zn finger motif demonstrated that Pb can bind the histidine and cysteine residues of the
Zn finger motif, thus displacing Zn and resulting in an increase in the DNA-binding
efficiency of the synthetic peptide (Razmiafshari et al., 2001; Razmiafshari and Zawia.
2000). However, in DNA-binding assays utilizing recombinant Spl (which has three Zn
finger motifs, opposed to only one in the synthetic peptide), 37 (iM Pb was the lowest
concentration observed to abolish the DNA-binding capabilities of Spl (Razmiafshari
and Zawia. 2000).
Pb has also been reported to competitively inhibit Mg binding and thus inhibit the
activities of adenine and hypoxanthine/guanine phosphoribosyltransferase in erythrocyte
lysates from rats exposed to 1,000 ppm Pb in drinking water for 9 months (mean [SD]
blood Pb level: 7.01 [1.64] (ig/dL) and in in vitro human erythrocyte lysates exposed to
0.1 (iM Pb for as little as 5 minutes (Baranowska-Bosiacka et al.. 2009) and also inhibit
cGMP phosphodiesterase at 10"6 (iM concentrations in homogenized bovine retinas
(Srivastava et al.. 1995). Pb was also reported to inhibit pyrimidine 5'-nucleotidase
through competitive inhibition of Mg binding, resulting in conformational changes and
improper amino acid positioning in the active site (Bitto et al.. 2006).
In summary, Pb has been shown to displace metal cations from the active sites of
multiple enzymes and proteins, and thus to alter the functions of those proteins in
occupationally-exposed humans with blood Pb levels of 5.4-69.3 (ig/dL, in adult rodents
with blood Pb levels of 6.5 (ig/dL (exposure 4 weeks), in suckling rats exposed to
2,000 ppm Pb via lactation, and in cell-free and cellular in vitro experiments conducted at
exposure concentrations ranging from 10"6 (iM to 1 (iM. These alterations in protein
function have implications for numerous cellular and physiological processes, including
cell signaling, growth and differentiation, gene expression, energy metabolism, and
biosynthetic pathways. Table 4-1 provides a list of enzymes and proteins whose function
may be perturbed by Pb exposure.
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Table 4-1 Enzymes and proteins potentially affected by exposure to Pb and the
metal cation cofactors necessary for their proper physiological
activity.
Metalloprotein/Enzyme
Direction of Action3
Metal Cation; Reference
Enzymes
Aminolevulinic acid
dehydratase
Erythrocyte
phosphoribosyltransferase
Zn; Simons (1995)
Ferrochelatase
Superoxide dismutase
IT
Catalase J,f
Glutathione peroxidase J,f
Guanylate cyclase
cGMP phosphodiesterase J,
NAD synthase J,
NAD(P)H oxidase f
Pyrimidine 5'-nucleotidase
Fe (2Fe-2S Cluster);
Crooks et al. (2010)
Mn, Cu, Zn, Fe;
Antonyuk et al. (2009),
Borgstahl et al. (1992)
Fe (Heme); Putnam et al. (2000)
Se; Rotruck et al. (1973)
Fe (Heme);
Boerrigter and Burnett (2009)
Mg, Zn; Ke (2004)
Mg; Hara et al. (2003)
Ca2+; Leseney et al. (1999)
Mg, Ca2+;
Bitto et al. (2006).
Amici et al. (1997).
Paglia and Valentine (1975)
Mg2+;
Dengetal. (2010),
Arnold and Kelley (1978)
Ion Channels/
Transport
ATPase
Mitochondrial
transmembrane pore
IT
2+
Ca+, Mg, Na/K; Technische
Universitat Braunschweig (2011)
Ca2+; He et al. (2000)
Signal
Transduction
Pb Binding
DNA Binding
Calcium-dependent
potassium channel
Protein kinase C
Calmodulin
Metallothionein
GATA transcriptional
factors
T
IT
T
T
1
Ca2+;
Silkin et al. (2001),
Alvarez et al. (1986)
Ca2+; Garza et al. (2006)
Ca2+; Garza et al. (2006)
Zn, Cu; Yu et al. (2009)
Zn;
Hanas et al. (1999),
Huang et al. (2004)
af indicates increased activity; J, indicates decreased activity; J,f indicates activity can be alternatively increased or decreased.
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4.2.2.4 Mitochondrial Abnormality
Alterations in mitochondrial function, including disruptions in ion transport,
ultrastructural changes, altered energy metabolism, and perturbed enzyme activities due
to Pb exposure are well documented in the scientific literature. Exposure of rats to Pb in
feed (10,000 ppm Pb for 4, 6, 8, 10, 12, or 20 weeks) or drinking water (300 ppm for
8 weeks, 500 ppm for 7 months, or 10,000 ppm Pb for 9 months) resulted in gross
ultrastructural changes in renal tubule mitochondria and epididymal mitochondria
characterized as a general swollen appearance with frequent rupture of the outer
membrane, distorted cristae, loss of cristae, frequent inner compartment vacuolization,
observation of small inclusion bodies, and fusion with adjacent mitochondria (Wang et
al..2010e: Marchlewicz et al.. 2009; Navarre-Moreno et al.. 2009; Gover. 1968; Gover et
al.. 1968).
Transmembrane mitochondrial ion transport mechanisms have been found to be
perturbed by exposure to Pb. Pb inhibits the uptake of Ca2+ into mitochondria (Parr and
Harris. 1976). while simultaneously stimulating the efflux of Ca2+ out of the organelle
(Simons. 1993a). thus disrupting intracellular/mitochondrial Ca2+ homeostasis. Pb
exposure has also been shown to decrease the mitochondrial transmembrane potential in
primary cerebellar granule neuronal cultures from rats exposed to 1,000 ppm Pb in
drinking water throughout gestation and lactation (Baranowska-Bosiacka et al., 201 Ib).
astroglia incubated with 0.1 or 1.0 (iM Pb for 14 days (Legare etal.. 1993). proximal
tubule cells exposed to 0.25, 0.5, and 1.0 (iM for 12 hours (Wang et al.. 2009c). and
retinal rod photoreceptor cells incubated with 0.01 to 10 (iM for 15 minutes (He et al..
2000). Further research indicated that Pb-induced mitochondrial swelling and decreased
membrane potential is the result of the opening of a mitochondrial transmembrane pore
(MTP), possibly by Pb directly binding to the metal (Ca2+)-binding site on the matrix side
of the pore (Bragadin et al.. 2007; He et al.. 2000). Opening of the MTP is the first step of
the mitochondrial-regulated apoptotic cascade pathway in many cells (Rana. 2008;
Lidsky and Schneider. 2003). He et al. (2000) additionally observed other indicators of
apoptosis including, cytochrome c release from mitochondria, and caspase-9 and -3
activation following exposure of retinal rod cells to Pb. Induction of mitochondrially-
regulated apoptosis via stimulation of the caspase cascade following exposure to Pb has
also been observed in rat hepatic oval cells (Agarwal et al.. 2009).
Altered Energy Metabolism
Pb has been reported to alter normal cellular bioenergetics. In mitochondria isolated from
the kidneys of rats exposed to 10,000 ppm Pb in feed for 6 weeks, the rate of oxygen
uptake during ADP-activated (state 3) respiration was lower compared to controls (Gover
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etal.. 1968). The rate of ATP formation from exposed mitochondria was observed to be
approximately 50% that of control mitochondria. A decrease in state 3 respiration and
respiratory control ratios (state 3/state 4 [succinate or pyruvate/malate-activated]) was
also observed in kidney mitochondria from rats exposed continuously from conception to
six or nine months of age (i.e., gestationally, lactationally, and via drinking water after
weaning) to 50 or 250 ppm Pb (Fowler etal.. 1980). Pb-induced decreases in ATP and
adenylate energy charge (AEC) were observed concurrently with increases in ADP,
AMP, and adenosine in adult rats exposed to 10,000 ppm Pb in drinking water for
9 months (Marchlewicz et al., 2009). Similarly, ATP and AEC were decreased, and AMP
increased, in primary cerebellar granule neuronal cultures from rats exposed to
1,000 ppm Pb in drinking water throughout gestation and lactation (Baranowska-
Bosiacka et al., 201 Ib). One (iM Pb (48 hours) was the lowest concentration observed to
decrease cellular ATP levels in NGF-differentiated PC- 12 cells, and these changes were
correlated with a Pb-induced decrease in the expression of the mitochondrial voltage-
dependent anion channel, which maintains cellular ATP levels in neurons (Prins et al..
2010). Dowd et al. (1990) reported that oxidative phosphorylation was decreased up to
74% after exposure of osteoblasts to 10 (iM Pb. Parr and Harris (1976) reported that Pb
inhibited both coupled and uncoupled respiratory oxygen use in mitochondria, and that
Pb prevented pyruvate, but not malate, uptake. Mitochondrial levels of ATP were
diminished after Pb exposure, and the authors compared the effects of Pb on the energy
supply to the actions of classic respiratory inhibitors, low temperature, and chemical
uncouplers. Bragadin et al. (1998) supported this view by demonstrating that alkylated Pb
compounds acted as a chemical uncoupler of respiration by abolishing the proton gradient
necessary for oxidative phosphorylation. Further, the enzymatic activities of complex I
and IV of the respiratory chain have been shown to be decreased in the peroneous longus
muscle of rats exposed to 250 ppm Pb (or 5 ppm thallium) in drinking water for 90 days
(Mendez-Armenta et al.. 201 1). Contrary to the above findings, Rafalowska et al. (1996)
reported that, although ATP levels did decrease in the forebrain synaptosomes prepared
from rats exposed to 200 ppm Pb in water for 3 months, this chronic exposure to Pb did
not inhibit oxidative phosphorylation in the synaptosomal mitochondria. Similar effects
with regard to the activity of the mitochondrial oxidative chain were observed in rats
injected with 15 mg Pb/kg (i.p.) daily for seven days, as reported by Struzynksa et al.
(1997a). although ATP levels were reported to increase after exposure to Pb.
Pb has also been shown to decrease glycolysis in osteoblasts exposed to 10 (iM Pb and in
human erythrocytes exposed (in vitro) to 30 (ig/dL Pb (Grabowska and Guminska. 1996;
Dowd etal.. 1990). Contrary to these findings, Antonowicz et al. (1990) observed higher
levels of glycolytic enzymes in erythrocytes obtained from Pb workers directly exposed
to Pb, compared to workers exposed to lower concentrations of Pb (mean blood Pb
levels: 82. 1 versus 39.9 (ig/dL), and suggested that Pb activated anaerobic glycolysis. In
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vitro exposure of human umbilical cord erythrocytes to 100-200 (ig/dL Pb for 20 hours
was observed to lower the cellular pools of adenine and guanine nucleotide pools,
including NAD and NADPH (Baranowska-Bosiacka and Hlvnczak. 2003). These
decreases in nucleotide pools were accompanied by an increase in purine degradation
products (adenosine, etc.). Similar decreases in cellular nucleotide pools were observed
when rats were exposed to 10,000 ppm Pb in drinking water for four weeks
(Baranowska-Bosiacka and Hlvnczak. 2004). In erythrocytes, nucleotides are synthesized
via salvage pathways such as the adenine pathway, which requires adenine
phosphoribosyltransferase. The activity of this enzyme was inhibited by exposure to Pb
in human and rat erythrocytes (see above for concentration and duration) (Baranowska-
Bosiacka etal.. 2009).
Disruptions in erythrocyte energy metabolism have been observed in adults
occupationally exposed to Pb. Nikolova and Kavaldzhieva (1991) reported higher ratios
of ATP/ADP in Pb-exposed workers with an average duration of exposure of 8.4 years
(blood Pb not reported) than in unexposed controls. Morita et al. (1997) evaluated the
effect of Pb on NAD synthetase in the erythrocytes of Pb-exposed workers (mean [SD]
blood Pb level: 34.6 [20.7] (ig/dL) and observed an apparent concentration-dependent
decrease in NAD synthetase activity with increasing blood Pb level. The blood Pb level
associated with 50% inhibition of NAD synthetase, which requires a Mg2+ cation for
activity (Hara et al.. 2003). was 43 (ig/dL.
Altered Heme Synthesis
Exposure to Pb is demonstrated to inhibit two key enzymes in the biosynthetic pathway
of heme: porphobilinogen synthase (i.e., 5-aminolevulinic acid dehydratase), a
cytoplasmic enzyme requiring Zn for enzymatic activity that condenses two molecules of
aminolevulinic acid into porphobilinogen, and ferrochelatase, a mitochondrial iron (Fe)-
sulfur (S) containing enzyme that incorporates Fe2+ into protoporphyrin IX to create
heme. Farant and Wigfield (1990, 1987) observed that Pb inhibits the activity of
porphobilinogen synthase in rabbit and human erythrocytes, and that the effect on the
enzyme was dependent on the affinity of Pb for thiol groups at the enzyme's active site.
Taketani et al. (1985) examined the activity of Pb on ferrochelatase in rat liver
mitochondria and observed that 10 (iM Pb (30 minute incubation) reduced NAD(P)H-
dependent heme synthesis by half when ferric, but not ferrous, iron was used. Pb inhibits
the insertion of Fe2+ into the protoporphyrin ring and instead, Zn is inserted into the ring
creating Zn-protoporphyrin (ZPP). While not directly measuring the activity of
ferrochelatase, numerous studies have shown that blood Pb levels are associated with
increased erythrocyte ZPP levels in humans (mean blood Pb levels ranging from 21.92 to
53.63 (ig/dL) (Mohammad et al.. 2008; Counter et al.. 2007; Patil et al.. 2006b:
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Ademuyiwa et al.. 2005b) and in animals (blood Pb level: 24.7 (ig/dL) (Rendon-Ramirez
et al.. 2007).
In summary, Pb has been shown to disrupt mitochondrial function including
transmembrane mitochondrial ion transport mechanisms and has been shown to perturb
normal cell bioenergetics. These effects have not only been demonstrated in in vitro
toxicological studies but also exposed worker populations.
4.2.3 Protein Binding
Evidence indicates that Pb binds to proteins within cells through interactions with side
group moieties (e.g., thiol residues) to form inclusion bodies and can thereby potentially
disrupt cellular function (Sections 4.2.2.3 and 4.2.2.4). However, some proteins are also
able to bind Pb and protect against its negative effects through sequestration. The ability
of Pb to bind proteins was first reported by Blackman (1936): the formation of
intranuclear inclusion bodies were observed in the liver and kidneys of children exposed
to high levels of Pb. Since that time, further research has been conducted into
characterizing the composition of intranuclear inclusion bodies and identifying specific
Pb-binding proteins.
4.2.3.1 Intranuclear and Cytoplasmic Inclusion Bodies
Goyer (1968) and Goyer et al. (1968) observed the formation of intranuclear inclusion
bodies in the renal tubules of rats fed 10,000 ppm Pb in food for up to 20 weeks. The
observation of inclusion bodies was accompanied by altered mitochondrial structure and
reduced rates of oxidative phosphorylation. Pb has further been observed to form
cytoplasmic inclusion bodies preceding the formation of the intranuclear bodies, and to
be concentrated within the subsequently induced intranuclear inclusion bodies following
i.p. injection, drinking water, and dietary exposures (Navarre-Moreno et al.. 2009;
Oskarsson and Fowler. 1985; Fowler et al., 1980; McLachlin et al., 1980; Choie and
Richter. 1972; Goyer etal.. 1970b; Goyer etal.. 1970a). Inclusion bodies have also been
observed in the mitochondria of kidneys and the perinuclear space in the neurons of rats
exposed to 500 ppm Pb acetate in drinking water for 60 days or 7 months (Navarro-
Moreno et al., 2009; Deveci. 2006). Intranuclear and cytoplasmic inclusions have also
been found in organs other than the kidney, including liver, lung, and glial cells (Singh et
al., 1999; Gover and Rhvne. 1973). Pb found within nuclei has also been shown to bind
to the nuclear membrane and histone fractions (Sabbioni and Marafante. 1976).
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Upon denaturing intranuclear inclusion bodies with strong denaturing agents, Moore et
al. (1973) observed that proteins included in the bodies were rich in aspartic and glutamic
acid, glycine, and cysteine. Further work by Moore and Goyer (1974) characterized the
protein as a 27.5 kilodalton (kDa) protein that migrates as a single band on
polyacrylamide gel electrophoresis. In contrast with the findings of Moore and Goyer,
Shelton and Egle (1982) identified a 32 kDa protein with an isoelectric point of 6.3 from
the kidneys of rats exposed to 10,000 ppm Pb acetate in feed or 7,500 ppm in drinking
water. This protein, dubbed p32/6.3, was not found in control rats, indicating that the
protein was induced by Pb exposure. This finding was in agreement with studies that
indicated the formation of intranuclear inclusion bodies required protein synthesis
(McLachlin et al.. 1980; Choie etal. 1975). In addition to its presence in kidneys of
Pb-exposed animals, p32/6.3 has been observed to be present and highly conserved in the
brains of rats, mice, dogs, chickens, and humans (Egle and Shelton. 1986). Exposure of
neuroblastoma cells to 50 or 100 (iM Pb glutamate for 1 or 3 days increased the
abundance of p32/6.3 (Klann and Shelton. 1989). Shelton et al. (1990) determined that
p32/6.3 was enriched in the basal ganglia, diencephalon, hippocampus, cerebellum,
brainstem, spinal cord, and cerebral cortex, and that it contained a high percentage of
glycine, aspartic, and glutamic acid residues. Selvin-Testa et al. (1991) and Harry et al.
(1996) reported that pre- and post-natal exposure of rats to 2,000-10,000 ppm Pb in
drinking water increased the levels of another brain protein, glial fibrillary acidic protein,
in developing astrocytes; and that this increase may be indicative of a demand for
astrocytes to sequester Pb.
4.2.3.2 Cytosolic Pb Binding Proteins
Numerous studies have also identified cytosolic Pb-binding proteins. Two binding
proteins, with molecular weights (MW) of 11.5 and 63 kDa, were identified by Oskarsson
et al. (1982) in the kidney postmitochondrial cytosolic fraction prepared from adult male
rats after i.p. injection with 50 mg Pb acetate/kg, followed by i.p. injection of 50 (iCi
203Pb acetate 6 days later. The two proteins were also found in the brain, but not the liver
or lung. Mistry et al. (1985) identified three proteins (MW = 11.5, 63, and >200 kDa) in
rat kidney cytosol, two of which, the 11.5 and 63 kDa proteins, were able to translocate
into the nucleus. The 11.5 kDa kidney protein was also able to reverse Pb binding to
ALAD through chelation of Pb and donation of a Zn cation to ALAD (Goering and
Fowler. 1985. 1984). Cd and Zn, but not Ca2+ or Fe, prevented the binding of Pb to the 63
and 11.5 kDa cytosolic proteins, which agrees with previous observations that Cd is able
to reduce total kidney Pb and prevent the formation of intranuclear inclusion bodies
(Mistry et al.. 1986; Mahaffev et al.. 1981; Mahaffev and Fowler. 1977). Additional
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cytosolic Pb-binding proteins have been identified in the kidneys of Pb-exposed rats and
humans, including the cleavage product of a2-microglobulin, acyl-CoA binding protein
(MW = 9 kDa), and thymosin |34 (MW = 5 kDa) (Smith etal.. 1998; Fowler and DuVal.
1991).
Cytosolic Pb-binding proteins distinct from kidney proteins have also been identified in
the brain of Pb-exposed rats, and in human brain homogenates exposed to Pb in vitro
(Ouintanilla-Vega et al. 1995: DuVal and Fowler. 1989: Goering et al.. 1986). One
protein (MW =12 kDa) was shown to alleviate hepatic ALAD inhibition due to Pb
exposure through competitive binding with Pb and donation of Zn to ALAD. Cytosolic
Pb-binding proteins have been shown to be high in glutamic acid, aspartic acid, and
cysteine residues (Fowler et al.. 1993: DuVal and Fowler, 1989). Some evidence exists
that cytosolic Pb-binding proteins directly target Pb and compartmentalize intracellular
Pb as a protective measure against toxicity (Oian et al.. 2005: Oian et al.. 2000).
4.2.3.3 Erythrocytic Pb Binding Proteins
The majority (94%) of Pb in whole blood is found in erythrocytes (Ong and Lee. 1980a).
Originally, the major Pb-binding protein in erythrocytes was identified as hemoglobin
(Cohen et al.. 2000: Lolin and O'Gorman, 1988: Ong and Lee, 1980a. b; Raghavan and
Gonick. 1977). However, Bergdahl et al. (1997b) observed the principal Pb-binding
protein to be 240 kDa and identified it as ALAD. Two smaller Pb-binding proteins were
observed, but not identified (MW = 45 and <10 kDa). ALAD levels are inducible by Pb
exposure; the total concentration of the enzyme, but not the activity, is higher in both Pb-
exposed workers (blood Pb = 30-75 (ig/dL) and rats (Pb exposure = 2.5 x 10"4 (iM in
drinking water) (Boudene et al., 1984: FujitaetaL 1982: FujitaetaL 1981).
ALAD is a polymorphic gene with three genotypes: ALAD 1-1, ALAD 1-2, or ALAD
2-2. Carriers of the ALAD-2 allele have been shown to have higher blood Pb levels than
carriers of the homozygous ALAD-1 allele (Scinicariello et al., 2007: Zhao et al., 2007:
Kim et al.. 2004: Perez-Bravo et al.. 2004: Smith etal.. 1995a: Wetmur. 1994: Wetmur et
al.. 1991b: Astrin etal.. 1987). Some recent studies, however, either observed lower
blood Pb levels in carriers of the ALAD-2 allele or no difference in Pb levels among the
different allele carriers (Scinicariello et al., 2010: Krieg et al., 2009: Chen et al., 2008c:
Chia et al.. 2007: Chia et al.. 2006: Wananukul et al.. 2006).
The ALAD-2 protein binds Pb more tightly than the ALAD-1 form: in Pb-exposed
workers carrying the ALAD-2 gene, 84% of blood Pb was bound to ALAD versus 81%
in carriers of the ALAD-1 gene (p = 0.03) (Bergdahl et al.. 1997a). This higher affinity
for Pb in ALAD-2 carriers may sequester Pb and prevent its bioavailability for reaction
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with other enzymes or cellular components. This is supported by the observation that
carriers of the ALAD-2 gene have higher levels of hemoglobin (Scinicariello et al..
2007). decreased plasma levulinic acid (Schwartz et al., 1997b). decreased levels of Zn
protoporphyrin (Scinicariello et al.. 2007; Kim et al.. 2004). lower cortical bone Pb
(Smith etal.. 1995b). and lower amounts of dimercaptosuccinic acid (DMSA)-chelatable
Pb (Scinicariello et al.. 2007; Schwartz et al.. 2000a: Schwartz et al.. 1997a). However,
the findings, that ALAD-2 polymorphisms reduced the bioavailability of Pb, are
somewhat equivocal. Wu et al. (2003a) observed that ALAD-2 carriers had lower blood
Pb level (5.8 ± 4.2 (ig/dL) than carriers of the homozygous ALAD-1 gene (blood Pb
level = 6.3 ±4.1 (ig/dL), and that ALAD-2 carriers demonstrated decreased renal
function at lower patellar Pb concentrations than those associated with decreased renal
function in homozygous ALAD-1 carriers. This potentially indicates that ALAD-2
carriers have enhanced Pb bioavailability. Weaver et al. (2003b) observed that ALAD-2
polymorphisms were associated with higher DMSA-chelatable Pb concentrations, when
normalized to creatinine levels. Further, Montenegro et al. (2006) observed that
compared with individuals with the ALAD 1-1 genotype, individuals with the ALAD
1-2/2-2 genotypes had a higher amount of Pb in the plasma (0.44 (ig/L versus 0.89 (ig/L,
respectively) and in the percent plasma/blood ratio (0.48% versus 1.45%, respectively).
This potentially suggests that individuals with the ALAD 1-2/2-2 genotypes are at
increased risk of Pb-induced health effects due to lower amounts of Pb sequestration by
erythrocyte ALAD, although this study did not specifically investigate the clinical
implications of ALAD polymorphisms.
ALAD has the estimated capacity to bind Pb at 85 (ig/dL in erythrocytes and 40 (ig/dL in
whole blood (Bergdahl et al.. 1998). The 45 kDa and <10 kDa Pb-binding proteins bound
approximately 12-26% and <1% of the blood Pb, respectively. At blood Pb
concentrations greater than 40 (ig/dL, greater binding to these components would likely
be observed. Bergdahl et al. (1998) tentatively identified the 45 kDa protein as
pyrimidine-5'-nucleotidase and the <10 kDa protein as acyl-CoA binding protein. Smith
et al. (1998) previously identified acyl-CoA binding protein as a Pb-binding protein
found in the kidney.
4.2.3.4 Metallothionein
In adults occupationally exposed to Pb, the presence of an inducible, low-molecular
weight (approximately 10 kDa) Pb-binding protein was identified in multiple early
studies (Gonick et al.. 1985; Raghavan et al.. 1981. 1980; Raghavan and Gonick. 1977).
The presence of this low molecular weight protein seemed to have a protective effect, as
workers who exhibited toxicity at lower blood Pb concentrations were observed to have
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lowered expression of this protein or low levels of Pb bound to it (Raghavan et al.. 1981.
1980). The presence of low molecular weight Pb-binding proteins in exposed workers
was corroborated by Lolin and O'Gorman (1988) and Church et al. Q993a, b). Further
Lolin and O'Gorman (1988) reported that the observed protein was only present when
blood Pb levels were greater than 39 (ig/dL, in agreement with the Pb-binding capacity of
ALAD, identified by Bergdahl et al. (1998). Xie et al. (1998) confirmed this, observing
the presence of a second low molecular weight protein with greater affinity than ALAD,
only at higher blood Pb levels. Church et al. (1993a. b) observed the presence of a 6-7
kDa protein in the blood of two Pb workers (blood Pb >160 (ig/dL); approximately 67%
of Pb was bound to the protein in the blood of the asymptomatic worker, whereas only
22% of the Pb was bound to it in the symptomatic (tremor, ataxia, extremity numbness)
worker. The reported protein was rich in cysteine residues and tentatively identified as
metallothionein.
Metallothionein is a low-MW metal-binding protein, most often binding Zn or Cu, that is
rich in cysteine residues and plays an important role in protecting against heavy metal
toxicity, maintaining trace element homeostasis, and scavenging free radicals (Yu et al..
2009). Exposure to Pb acetate induced the production of Pb- and Zn-metallothionein in
mice treated via i.p. or intravenous (i.v.) injection at 30 mg/kg (Maitani et al.. 1986). in
mice treated via i.p. injection at 300 (imol/kg (Yu et al.. 2009). or in rats treated via i.p.
injection at 24 (imol/lOOg (Ikebuchi et al.. 1986). The induced Pb-metallothionein
consisted of 28% half-cysteine and reacted with an antibody for Zn-metallothionein II
(Ikebuchi et al.. 1986). In contrast, exposure of rats to Pb via drinking water (200 or
300 ppm) failed to induce metallothionein in the kidneys or intestines (Wang et al..
2009b: Jamieson et al.. 2007). Goering and Fowler (1987a. b) observed that pretreatment
of rats with Zn before injection with Pb resulted in Pb and Zn co-eluting with Zn-
thionein, and that Zn-thionein I and II were able to bind Pb in vitro (Goering and Fowler.
1987a. b). Further, Goering and Fowler (1987a) found that kidney and liver Zn-thionein
decreased binding of Pb to liver ALAD and was able to donate Zn to ALAD, thus
attenuating the inhibition of ALAD due to Pb exposure. These findings are in agreement
with Goering et al. (1986) and DuVal and Fowler (1989) who demonstrated that rat brain
Pb-binding proteins attenuated Pb-induced inhibition of ALAD.
Metallothionein has been reported to be important in the amelioration of Pb-induced
toxicity effects. Liu et al. (1991) reported that Zn-metallothionein reduced Pb-induced
membrane leakage and loss of K+ in cultured hepatocytes incubated with 600-3,600 (iM
Pb. Metallothionein-null mice exposed to 1,000, 2,000, or 4,000 ppm Pb for 20 weeks
suffered renal toxicity described as nephromegaly and decreased renal function compared
to Pb-treated wild-type mice (Qu et al.. 2002). Interestingly, metallothionein-null mice
were unable to form intranuclear inclusion bodies and accumulated less renal Pb than did
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the wild-type mice (Ou et al., 2002). Increased metallothionein levels were induced by Pb
exposure in non-null mice. Exposure to Pb (1,000, 2,000, or 4,000 ppm), both for
104 weeks as adults and from GD8 to early adulthood, resulted in increased preneoplastic
lesions and carcinogenicity in the testes, bladder, and kidneys of metallothionein-null rats
compared to wild type mice (TokaretaL 2010; Waalkes et al., 2004). Inclusion bodies
were not observed in null mice. The authors concluded that metallothionein is important
in the formation of inclusion bodies and in the mitigation of Pb-induced toxic effects, and
that those with polymorphisms in metallothionein coding genes that are associated with
reduced inducibility may be at greater susceptibility to Pb. In support of this theory, Chen
et al. (2010a) observed that Pb-exposed workers with a mutant metallothionein allele had
higher blood Pb levels than did homozygous carriers of the normal allele (means 24.17
and 21.27 versus 17.03 (ig/dL), and had larger blood Pb-associated changes in systolic
BP and serum renal function parameters.
In summary, a number of proteins have been identified that can bind and sequester Pb
including ALAD and metallothionein. Additionally, evidence suggests that certain
polymorphisms that alter the binding capability or inducibility of these proteins can
increase the risk of Pb-induced health effects.
4.2.4 Oxidative Stress
Oxidative stress occurs when free radicals or reactive oxygen species (ROS) exceed the
capacity of antioxidant defense mechanisms. This oxidative imbalance results in
uncontained ROS, such as superoxide (O2~), hydroxyl radical (OH), and hydrogen
peroxide (H2O2), which can attack and denature functional/structural molecules and,
thereby, promote tissue damage, cytotoxicity, and dysfunction. Pb exposure has been
shown to cause oxidative damage to the heart, liver, kidney, reproductive organs, brain,
and erythrocytes, which may be responsible for a number of Pb-induced health effects
(Salawu et al.. 2009: Shan et al.. 2009: Vaziri. 2008b: Gonick et al.. 1997: Sandhir and
Gill 1995: Khalil-Manesh et al.. 1994: Khalil-Manesh et al.. 1992a). The origin of ROS
produced after Pb exposure is likely a multipathway process, resulting from oxidation of
5-aminolevulinic acid (ALA), membrane and lipid oxidation, NAD(P)H oxidase
activation, and antioxidant enzyme depletion, as discussed below. Some of these
processes result from the disruption of functional metal ions within oxidative stress
enzymes, such as superoxide dismutase (SOD), catalase (CAT), and glutathione
peroxidase (GPx). Interestingly, Pb exposure in many species of plants, invertebrates, and
vertebrates discussed in Chapter 6 (Ecological Effects of Lead) results in upregulation of
antioxidant enzymes and increased lipid peroxidation. Oxidative stress is a common
mode of action for a number of other metals (e.g., Cd, Mn, As, Co, Cr) that are often
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found with Pb and by which possible interactions with Pb have been suggested to occur
(Jomova and Valko. 2011; Jomova et al.. 2011; Matovic et al.. 2011; HaMai and Bondv.
2004). Not all of these co-occurring metals directly produce ROS or redox cycle, but
instead may suppress the free radical scavenging ability of the organism thus leading to
oxidative stress.
4.2.4.1 5-ALA Oxidation
The majority of Pb present in the blood accumulates in erythrocytes where it enters
through passive carrier-mediated mechanisms including a vanadate-sensitive Ca2+ pump.
Once Pb enters erythrocytes, it is predominantly found in the protein-bound form, with
hemoglobin and 5-ALAD both identified as targets (Bergdahl et al.. 1997a). Through its
sulfhydryl and metal ion disrupting properties, Pb incorporates with and inhibits a
number of enzymes in the heme biosynthetic process, including 5-ALA synthetase,
5-ALAD, and ferrochelatase. Pb has been shown to be able to disrupt the Zn ions
requisite for the activity of 5-ALAD, the rate limiting step in heme synthesis, leading to
enzyme inhibition at 10"6 (iM concentrations (Simons. 1995). Additionally, blood Pb
levels (mean: 7 (ig/dL) have been associated with inhibited activity of 5-ALAD in
humans, and the correlation was found in the group with blood Pb levels > 5 (ig/dL
(Ahamed et al.. 2005; Sakai and Morita. 1996). A negative correlation (r = -0.6) was
found between blood Pb levels in adolescents (range of blood Pb levels: 4 to 20 (ig/dL)
and blood 5-ALAD activity (Ahamed et al.. 2006). This inhibition of 5-ALAD results in
the accumulation of 5-ALA in blood and urine, where 5-ALA undergoes tautomerization
and autoxidation. Oxidized 5-ALA generates ROS through reduction of ferricytochrome
c and electron transfer from oxyHb, metHb, and other ferric and ferrous iron complexes
(Hermes-Lima et al.. 1991; Monteiro et al., 1991). The autoxidation of 5-ALA produces
O2~, OH, H2O2, and an ALA radical (Monteiro et al.. 1989; Monteiro et al.. 1986).
4.2.4.2 Membrane and Lipid Peroxidation
A large number of studies in humans and experimental animals have indicated that
exposure to Pb can lead to membrane and lipid peroxidation. It is possible that ROS
produced from 5-ALA oxidation, as described above, interacts with and disrupts
membrane lipids (Oteizaet al., 1995; Bechara et al.. 1993). Additionally, Pb has the
capacity to stimulate ferrous ion initiated membrane lipid peroxidation serving as a
catalyst for these events (Adonavlo and Oteiza. 1999; Quinlan etal.. 1988). The extent of
peroxidation of lipids varies based on the number of double bonds present in unsaturated
fatty acids, since double bonds weaken the C-H bonds on the adjacent carbon, making
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hydrogen (H) removal easier (Yiin and Lin. 1995). After the essential unsaturated fatty
acid solutions were incubated with Pb (4-12 (ig/dL, 24 hours), the production of
malondialdehyde (MDA), a marker of oxidative stress and lipid oxidation end product,
increased relative to the number of double bonds of the fatty acid (Yiin and Lin. 1995). In
the absence of Fe2+, Pb has not been shown to promote lipid peroxidation; however, it
may accelerate peroxidation by H2O2 (Quinlan et al.. 1988). This could be due to altering
membrane structure, restricting phospholipid movement, and facilitating the propagation
of peroxidation.
Pb induces changes in the fatty acid composition of a membrane, which could lead to
oxidative damage. Exposure to Pb (>62.5 ppm in drinking water, 3 weeks) in chicks
promoted an increase in arachidonic acid (20:4) as a percentage of total fatty acids, and
decreased the relative proportion of shorter chain fatty acids (linoleic acid, 18:2) (Lawton
and Donaldson. 1991). It is possible that Pb depressed the desaturation of saturated fatty
acids to the corresponding monoenoic fatty acids, while stimulating elongation and
desaturation of linoleic acid to arachidonic acid. Since fatty acid chain length and
unsaturation are related to the oxidative potential, changes in fatty acid membrane
composition may result in enhanced lipid peroxidation. In addition, changes in fatty
acids, thus membrane composition, can result in altered membrane fluidity (Donaldson
and Knowles. 1993). Changes in membrane fluidity will disturb the conformation of the
active sites of membrane associated enzymes, disrupt metabolic regulation, and alter
membrane permeability and function.
A number of recent studies report increased measures of lipid peroxidation in various
organs, tissues, and species in association with Pb. Occupational Pb exposure resulting in
elevated blood Pb levels (means >8 (ig/dL) reported in various countries provides
evidence of lipid peroxidation, including increased plasma MDA levels (Ergurhan-Ilhan
et al.. 2008; Khan et al.. 2008; Mohammad et al.. 2008; Ouintanar-Escorza et al.. 2007;
Patil et al.. 2006a; Patil et al.. 2006b). One study found a correlation between the MDA
levels and blood Pb levels even in the unexposed workers, although they had blood Pb
levels higher than the mean blood Pb level of the current U.S. population
(i.e., <12 (ig/dL) (Ouintanar-Escorza et al.. 2007). Other studies found evidence of
increased lipid peroxidation among the general population, including children, with
elevated blood Pb levels (means >10 (ig/dL) (Ahamed et al., 2008; Ahamed et al., 2006;
Jin et al.. 2006). In adolescents, Ahamed et al. (2006) found blood MDA levels to be
positively correlated (r = 0.7) with concurrent blood Pb levels ranging between 4 and
20 (ig/dL. Similar results have been shown after Pb exposure in animal studies (Abdel
Moneimetal. 20lib; Pandvaet al.. 2010; Dogru et al.. 2008; Yu et al.. 2008; Adegbesan
and Adenuga. 2007; Lee et al.. 2005). Enhanced lipid peroxidation has been found in Pb
treated (50 (ig, 1-4 hours) rat brain homogenates (Rehman et al.. 1995). rat proximal
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tubule cells (0.5-1 (JVI, 12 hours) (Wang etal. 20lib), and in specific brain regions,
hippocampus and cerebellum, after Pb exposure (500 ppm, 8 weeks) to rats (Bennet et al.
2007). Overall, there was a correlation between the blood Pb level and measures of lipid
peroxidation often measured by MDA levels.
In summary, studies in humans and animals provide evidence for increased lipid and
membrane oxidation following Pb exposure. Interestingly, many species of plants,
invertebrates, and other vertebrates also exhibit increased lipid peroxidation with Pb
exposure (Sections 6.3.12.6 and 6.4.12.6). The increase in lipid peroxidation following
Pb exposure observed across species and kingdoms demonstrate an evolutionarily
conserved oxidative response following Pb exposure.
4.2.4.3 NAD(P)H Oxidase Activation
NAD(P)H oxidase is a membrane bound enzyme that requires Ca2+ in order to catalyze
the production of O2 from NAD(P)H and molecular oxygen (Lesenev et al.. 1999). Two
studies provide evidence for increased activation of NAD(P)H oxidase that may
contribute to the production of ROS after Pb exposure (Ni et al.. 2004; Vaziri et al..
2003). Vaziri et al. (2003) found increased protein expression of the NAD(P)H subunit
gp91phox in the brain, heart, and renal cortex of Pb-treated rats (100 ppm in drinking
water, 12 weeks). This upregulation was present in Pb-treated (1-10 ppm) human
coronary artery endothelial cells, but not vascular smooth muscle cells (VSMC), which
do not express the protein (Ni et al.. 2004). It is possible that NAD(P)H oxidase serves as
a potential source of ROS in cells that express this protein.
4.2.4.4 Antioxidant Enzyme Disruption
Oxidative stress can result not only from the increased production of ROS, but also from
the decreased activity of antioxidant defense enzymes. Pb has been shown to alter the
function of several antioxidant enzymes, including SOD, CAT, glucose-6-phosphate
dehydrogenase (G6PD), and the enzymes involved in glutathione metabolism, GPx,
glutathione-S-transferase (GST), and glutathione reductase (GR). These changes in the
antioxidant defense system could be due to the high affinity of Pb for sulfhydryl groups
contained within proteins and its metal ion mimicry. However changes could also be a
consequence of increased oxidative damage by Pb.
Studies of the effects of Pb exposure on the activities of SOD and CAT give divergent
results. These metalloprotein enzymes require various essential trace elements for proper
structure and function, making them a target for Pb toxicity. CAT is a heme containing
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protein that requires Fe ions to function (Putnam et al.. 2000). SOD exists in multiple
isoforms that require Cu and Zn (SOD1 and SODS) (Antonyuk et al.. 2009) or Mn
(SOD2) (Borgstahl et al.. 1992). A number of studies have found decreased activity of
these enzymes (Conterato et al.. 2013; Pandvaetal.. 2010; Ergurhan-Ilhan et al.. 2008;
Mohammad et al.. 2008; Yu et al.. 2008; Patil et al.. 2006a: Patil et al.. 2006b). whereas
others have observed increased activity following Pb exposure (Ahamed et al.. 2008; Lee
et al.. 2005). The heterogeneity in species examined, (i.e., humans, rodents, boars), and
Pb exposure duration and metrics may account for some of the heterogeneity in reported
results. These heterogeneities in study designs limit the evaluation of whether a nonlinear
concentration-response relationship could explain the conflicting results regarding the
directionality of effects observed. Pb exposure (500 ppm Pb acetate, 1, 4, and 8 weeks) in
adult male rats showed that SOD and CAT activity responded differently depending on
the brain region analyzed and length of exposure (Bennet et al.. 2007). Another study
found that the brain had consistently decreased SOD activity, irrespective of Pb dose in
prenatally-exposed animals (0.3 and 3.0 ppm, blood Pb level 20.4 and 24.5 (ig/dL);
however, hepatic SOD activity increased at low level Pb administration and decreased
after high level exposure (Uzbekov et al.. 2007). It is possible that the increased activity
of the SOD and CAT proteins is due to activation by ROS, while decreased enzyme
activity is the result of metal ion substitution by Pb, causing enzyme inactivation.
Glutathione is a tripeptide antioxidant containing a cysteine with a reactive thiol group
that can act nonenzymatically as a direct antioxidant or as a cofactor in enzymatic
detoxification reactions by GST. Glutathione will donate an electron while in its reduced
state (GSH), which leads to conversion to the oxidized form, glutathione disulfide
(GSSG). Pb binds to the thiol and can both interfere with the antioxidant capacity of
GSH, and can decrease levels of GSH. Short-term administration of Pb in vitro (0.1 (iM)
and biomarkers of Pb exposure in humans (18 (ig/dL mean blood Pb level) have been
associated with decreased blood and organ GSH and cysteine content, which may be due
to increased GSH efflux from tissues (Pandvaetal.. 2010; Pillai etal.. 2010; Ahamed et
al.. 2009; Ahamed et al.. 2008; Flora et al.. 2007; Ahamed etal.. 2005; Chettv et al..
2005; Nakagawa. 1991. 1989). Long-term Pb exposure has elicited a compensatory
upregulation in the biosynthesis of GSH in the attempt to overcome Pb toxicity, thus
often manifesting as an increase in Pb-induced GSH in animals (Daggett et al.. 1998;
Corongiu and Milia. 1982; Hsu. 1981) and occupationally exposed adults (Conterato et
al.. 2013; SRC. 2002). However, other studies have found that long-term Pb exposure,
resulting in mean blood Pb levels between 6.6 and 22 (ig/dL, causes the depletion of GSH
in animals (Lee et al.. 2005; Ercal etal.. 1996) and occupationally exposed adults
(Mohammad et al.. 2008). Thus, the duration of Pb exposure is important to consider
when measuring GSH levels.
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Glutathione reductase is able to reduce GSSG back to GSH. Therefore, an increased
GSSG/GSH ratio is considered to be indicative of oxidative stress. Epidemiologic studies
have found higher blood Pb levels to be associated with increases in the GSSG/GSH ratio
(Mohammad et al.. 2008; Ercaletal.. 1996; Sandhir and Gill. 1995). In one study, this
association was observed in a population of children with a mean blood Pb level below
10 (ig/dL (Diouf et al.. 2006). Studies have found mixed effects on GR activation. GR
possesses a disulfide at its active site that is a target for inhibition by Pb. Studies in
animals and cells have reported decreased (Bokara et al.. 2009; Sandhir and Gill. 1995;
Sandhiretal.. 1994). increased (Sobekova et al.. 2009; Howard. 1974). and no change
(Hsu. 1981) in GR activity after Pb exposure. This could be because the effect of Pb on
GR varies depending on sex (Sobekova et al.. 2009). and organ or organ region (Bokara
et al.. 2009). Additionally, heterogeneity in species examined, (i.e., humans, rodents), and
Pb exposure duration and metrics may account for some of the heterogeneity in reported
results. These heterogeneities in study designs make it difficult to compare exposures and
doses across studies and limit the evaluation of whether a nonlinear concentration-
response relationship could explain the conflicting results regarding the directionality of
effects observed.
GSH is used as a cofactor for peroxide reduction and detoxification of xenobiotics by the
enzymes GPx and GST. GPx requires Se (selenium) for peroxide decomposition (Rotruck
et al.. 1973). whereas GST functions via a sulfhydryl group. Evidence indicates that by
reducing the uptake of Se, depleting cellular GSH, and disrupting protein thiols, Pb
decreases the activity of GPx and GST (Pillai etal.. 2010: Yu et al.. 2008: Lee et al..
2005: Nakagawa. 1991: Schrauzer. 1987). Similar to other antioxidant enzymes,
compensatory upregulation of these enzymes was observed after Pb exposure in animals
and in Pb-exposed workers (painters with a mean blood Pb level of 5.4 (ig/dL (Conterato
etal.. 2013: Bokara et al.. 2009: Ergurhan-Ilhan et al.. 2008: Conterato et al.. 2007:
Daggett et al.. 1998). However, in another study, these enzymes were not able to
compensate for the increased Pb-induced ROS, further contributing to the oxidative
environment (Farmand et al.. 2005).
Recently, y-glutamyltransferase (GGT) within its normal range has been regarded as an
early and sensitive marker of oxidative stress. This may be because cellular GGT
metabolizes extracellular GSH to be used in intracellular GSH synthesis. Thus, cellular
GGT acts as an antioxidant enzyme by increasing the intracellular GSH pool. However,
the reasons for the association between GGT and oxidative stress have not been fully
realized (Lee et al.. 2004). In one study, occupational Pb exposure (mean blood Pb level
of 29.1 (ig/dL) was associated with increased serum GGT levels (Khan et al.. 2008).
Interestingly, higher blood Pb level was similarly associated with higher serum GGT
levels in a sample of the U.S. adult population (NHANES III) (Lee et al.. 2006a). In this
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study of nonoccupationally-exposed individuals, a concentration-dependent relationship
was observed with blood Pb levels <7 (ig/dL.
In summary, Pb has been shown to alter the function of several antioxidant enzymes,
including SOD, CAT, G6PD, and the enzymes involved in glutathione metabolism, GPx,
GST, and GR in human populations and experimental animal models, although not
consistently in the same direction. Interestingly, many species of plants, invertebrates,
and other vertebrates also exhibit altered antioxidant enzyme activity following Pb
exposure (Sections 6.3.12.6. 6.4.12.6. and 6.4.21.6). Alteration of these enzymes may
lead to further oxidative stress following Pb exposure.
4.2.4.5 Nitric Oxide Signaling
NO (nitric oxide or nitric oxide radical), also known as endothelium-derived relaxing
factor, is a potent endogenous signaling molecule involved in vasodilation. Short- and
long-term Pb exposures in animals have been found to decrease the biologically active
NO, not through reduction in NO-production capacity (Vaziri and Ding. 2001; Vaziri et
al.. 1999a). but as a result of inactivation and sequestration of NO by ROS (Malvezzi et
al.. 2001; Vaziri et al.. 1999b). Endogenous NO can interact with ROS, specifically O2~,
produced following exposure to Pb to form the highly cytotoxic reactive nitrogen species,
peroxynitrite (ONOO"). This reactive compound can damage cellular DNA and proteins,
resulting in the formation of nitrotyrosine among other products. Overabundance of
nitrotyrosine in plasma and tissues is present after exposure to Pb (Vaziri etal.. 1999b).
NO is also produced by macrophages in the defense against certain infectious agents,
including bacteria. Studies have indicated that Pb exposure can significantly reduce
production of NO in human (Pineda-Zavaleta et al.. 2004) and animal (Lee etal.. 200 Ib;
Tian and Lawrence. 1995) immune cells, possibly leading to reduced host resistance
(Tian and Lawrence. 1996).
Production of NO is catalyzed by a family of enzymes called nitric oxide synthases
(NOS), including endothelial NOS (eNOS), neuronal NOS (nNOS), and inducible NOS
(iNOS), which require a heme prosthetic group and a Zn cation for enzymatic activity
(Messerschmidt et al.. 2001). Paradoxically, the reduction in NO availability in vascular
tissue following Pb exposure is accompanied by statistically significant upregulation in
NOS isotypes (Vaziri and Ding. 2001: Vaziri etal.. 1999a: Gonicketal. 1997). A direct
inhibitory action of Pb on NOS enzymatic activity has been rejected (Vaziri et al..
1999a). Instead, the upregulation of NOS occurs as compensation for the decreased NO
resulting from ROS inactivation (Vaziri et al.. 2005; Vaziri and Ding. 2001; Vaziri and
Wang. 1999).
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Soluble Guanylate Synthase
Many biological actions of NO, such as vasorelaxation, are mediated by cyclic guanosine
monophosphate (cGMP), which is produced by soluble guanylate cyclase (sGC) from the
substrate guanosine triphosphate. Soluble guanylate cyclase is a heterodimer requiring
one molecule of heme for enzymatic activity (Boerrigter and Burnett. 2009). In VSMC,
sGC serves as the NO receptor. Marked reduction in plasma concentrations and urinary
excretion of cGMP is observed after Pb exposure to rats [5 ppm Pb in drinking water for
30 days (Marques et al., 2001) and 100 ppm Pb acetate in drinking water for 3 months,
resulting in a mean blood Pb level of 29.4 (ig/dL (Khalil-Manesh et al.. 1993b)] (Marques
et al., 2001; Khalil-Manesh et al.. 1993b). In addition, Pb exposure downregulated the
protein abundance of sGC in vascular tissue (Tarmand et al.. 2005; Courtois et al.. 2003;
Marques et al.. 2001). This downregulation in sGC was prevented by antioxidant therapy
(ascorbic acid) suggesting that oxidative stress also plays a role in Pb-induced
downregulation of sGC (no change in blood Pb level was observed after ascorbic acid
treatment) (Marques et al.. 2001).
4.2.5 Inflammation
Misregulated inflammation represents one of the major hallmarks of Pb-induced immune
effects. It is important to note that this can manifest in any tissue where immune cell
mobilization and tissue insult occurs. Enhanced inflammation and tissue damage occurs
through the modulation of inflammatory cell function and production of pro-
inflammatory cytokines and metabolites. Overproduction of ROS and an apparent
depletion of antioxidant protective enzymes and factors (e.g., Se) accompany this
immunomodulation (Chettv et al.. 2005).
Traditional immune-mediated inflammation can be seen with bronchial
hyperresponsiveness, asthma, and respiratory infections, some of which have been
associated with exposure to Pb. But it is important to recognize that any tissue or organ
can be affected by immune-mediated inflammatory dysfunction given the distribution of
immune cells as both permanent residents and infiltrating cell populations (Mudipalli.
2007; Carmignani et al.. 2000). Pb has been associated with multiple indicators of
inflammation in multiple cell types. Pb has also induced renal tubulointerstitial
inflammation (100 ppm exposure for 14 weeks) (Rodriguez-Iturbe et al., 2005)
(24.6 (ig/dL blood Pb level, 150 ppm for 16 weeks) (Roncal et al.. 2007). Renal
tubulointerstitial inflammation has been coupled with activation of the redox sensitive
nuclear transcription factor kappa B (NFKB) and lymphocyte and macrophage infiltration
in rats (100 ppm for 14 weeks resulting in mean blood Pb levels ranging 23.7 (ig/dL)
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(Bravo et al.. 2007). These events could occur in response to the oxidative environment
arising from Pb exposure, since Pb-induced inflammation and NFxB activation can be
ameliorated by antioxidant therapy (Rodriguez-Iturbe et al.. 2004). Pb spheres implanted
in the brains of rats produced neutrophil-driven inflammation with apoptosis and
indications of neurodegeneration (Nakao et al.. 2010).
Inflammation can be mediated by the production of chemical messengers such as
prostaglandins (PG). Pb exposure has been associated with increased arachidonic acid
(AA) metabolism, thus elevating the production of PGE2, PGF2, and thromboxane in
occupationally-exposed humans (mean blood Pb level 48 (ig/dL) (Cardenas et al.. 1993)
and animal and cell models (e.g., 0.01 (iM, 48 hours) (Chetty et al.. 2005; Flohe et al..
2002; Knowles and Donaldson. 1997; Lee and Battles. 1994). Dietary Pb exposure of
animals (500 ppm, 19 days) can increase the percentage of cell membrane AA, the
precursor of cyclooxygenase and lipoxygenase metabolism to PGs and leukotrienes
(Knowles and Donaldson. 1990). Additionally, Pb (1 (iM) may promote the release of
AA via activation of phospholipase A2, as shown in isolated VSMC (Dorman and
Freeman. 2002).
Inflammation may be the result of increased pro-inflammatory signaling or may stimulate
these signaling pathways. Pb can elevate the expression of the pro-inflammatory
transcription factors NFxB and activator protein-1 (AP-1), as well as the AP-1
component c-Jun (Korashy and El-Kadi. 2008; Korashy and Ei-Kadi. 2008; Bravo et al.,
2007; Ramesh et al.. 1999; Pvatt et al.. 1996). Pb exposure (25 (iM) to dendritic cells
stimulated phosphorylation of the Erk/MAPK pathway, but not p38, STATS or 5, or
CREB (Gao et al.. 2007)
4.2.5.1 Cytokine Production
There are three modes by which Pb has been shown to affect immune cytokine
production. First, Pb can act on macrophages to elevate the production of pro-
inflammatory cytokines such as TNF-a and interleukin (IL)-6 (Cheng et al.. 2006; Chen
etal.. 1999; Miller et al.. 1998; Dentener et al.. 1989). This can result in local tissue
damage during the course of immune responses affecting such targets as the liver.
Second, Pb can skew the ratio of IL-12/IL-10 such that T-derived lymphocyte helper
(Th)l responses are suppressed and Th2 responses are promoted (Gao et al.. 2007)
possibly by affecting dendritic cells. Third, when acquired immune responses occur
following exposure to Pb, Thl lymphocyte production of cytokines is suppressed
(e.g., IFN-y) (Lynes et al., 2006; Heo et al., 1996); in contrast, Th2 cytokines such as
IL-4, IL-5, and IL-6 are elevated (Gao et al.. 2007; Chen et al.. 2004; Kim and Lawrence.
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2000). The combination of these three modes of cytokine changes induced by Pb can
create a hyperinflammatory state among innate immune cells and skew acquired
immunity toward Th2 responses.
lavicoli et al. (2006b) reported that low blood Pb concentrations produced significant
changes in cytokine levels in mice. At a low dietary Pb concentration (0.11 ppm, blood
Pb level of 1.6 (ig/dL), IL-2 and IFN-y were decreased compared to the controls
(0.02 ppm, 0.8 (ig/dL), indicating a suppressed Thl response. As the dietary and blood Pb
concentrations increased (resulting in blood Pb levels 12-61 (ig/dL), a Th2 phenotype
was observed with suppressed IFN-y and IL-2 and elevated IL-4 production. These
findings support the notion that the immune system is remarkably sensitive to Pb-induced
functional alterations and that nonlinear effects may occur at low Pb exposures. TGF-|3
production was also altered by Pb exposure to transfected mouse limb bud mesenchymal
stem cells (1 (JVI, 3 days) (Zuscik et al.. 2007). IL-2 is one of the more variable cytokines
with respect to Pb-induced changes. Depending upon the protocol it can be slightly
elevated in production or unchanged. Recently, Gao et al. (2007) found that Pb-treated
dendritic cells (25 (iM) promoted a slight but statistically significant increase in IL-2
production among lymphocytes. Proinflammatory cytokines have been measured in other
organs and cell types after Pb exposure. Elevation of IL-lp and TNF-a were observed in
the hippocampus of rats after Pb treatment (15 ppm, i.p., daily for 2 weeks, blood Pb
level of 30.8 (ig/dL) and increased IL-6 was found in the forebrain (Struzynska et al..
2007).
Consistent with animal studies, epidemiologic studies also found higher concurrent blood
Pb levels in children and occupationally-exposed adults to be associated with a shift
toward production of Th2 cytokines relative to Thl cytokines. The evidence in children
was based on comparisons of serum cytokine levels among groups with different blood
Pb levels without consideration of potential confounding factors. Among children ages
9 months to 6 years in Missouri, Lutz et al. (1999) found that children with concurrent
blood Pb levels 15-19 (ig/dL had higher serum levels of IL-4 and IgE (Section 4.6.3) than
did children with lower blood Pb levels. These results were consistent with the mode of
action for IL-4 to activate B cells to induce B cell class switching to IgE. In another
study, concurrent blood Pb levels did not differ by residence in old versus new homes or
by urban versus rural residence (means: 3.2-3.8 (ig/dL) but were higher among children
living near an oil refinery, in particular, among children with known respiratory allergies
(mean: 8.8 (ig/dL) (Hsiao etal. 2011). This latter group of children also had the lowest
serum levels of IFN-y and highest levels of IL-4. There was no direct comparison of
cytokine levels between blood Pb level groups in the population overall; however,
cytokine levels were similar between healthy and allergy groups in the other Pb source
groups that had similar blood Pb levels. Thus, the differences in cytokine levels between
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healthy and allergic children living near the oil refinery may have been influenced by
differences in their blood Pb levels or other factors related to residence near an oil
refinery.
Evidence of association between blood Pb levels and cytokine levels in
nonoccupationally-exposed adults was unclear. Among healthy adult university students
in Incheon, Korea, Kim et al. (2007) found associations of concurrent blood Pb level with
serum levels of TNF-a and IL-6 that were larger among male students with blood Pb
levels 2.51-10.47 (ig/dL. Notably, the relative contributions of lower recent versus higher
past Pb exposures to these cytokine effects is not known. In models that adjusted for age,
sex, BMI, and smoking status, a 1 (ig/dL increase in blood Pb level was associated with a
23% increase (95% CI: 4, 55%) in log of TNF-a and a 26% increase (95% CI: 0, 55%) in
log of IL-6. The association between levels of blood Pb and plasma TNF-a was greater
among men who were GSTM1 null or had the TNF-a GG genotype. For the association
between blood Pb level and plasma IL-6, the effect estimate was slightly elevated in
TNF-a GG genotype but not elevated in the GSTM1 positive group. The effects of Pb on
several physiological systems have been hypothesized to be mediated by the generation
of ROS (Daggett et al.. 1998). Thus, the null variant of GSTM1, which is associated with
reduced elimination of ROS, may increase the risk of Pb-associated immune effects. The
results for the TNF-a polymorphism are difficult to interpret. The GG genotype is
associated with lower expression of TNF-a, and the literature is mixed with respect to
which variant increases risk of inflammation-related conditions. A study of adults in Italy
did not provide quantitative results and only reported a lack of statistically significant
correlation between blood Pb levels and Th2 or Thl cytokine levels in men (Boscolo et
al.. 1999) and women (Boscolo et al.. 2000).
Results from studies of occupationally-exposed adults also suggested that Pb exposure
may be associated with decreases in Thl cytokines and increases in Th2 cytokines;
however, analyses were mostly limited to comparisons of levels among different
occupational groups with different mean blood Pb levels (Di Lorenzo et al., 2007;
Valentino et al.. 2007; Yiicesov et al.. 1997a) without consideration for potential
confounding factors including other occupational exposures. An exception was a study of
male foundry workers, pottery workers, and unexposed workers by Valentino et al.
(2007). Although quantitative regression results were not provided, higher blood Pb level
was associated with higher IL-10 and TNF-a with adjustment for age, BMI, smoking, and
alcohol consumption. In analyses of blood Pb groups, levels of IL-2, IL-10, and IL-6 also
increased from the lowest to highest blood Pb group. In contrast with most other studies,
both exposed worker groups had lower IL-4 levels compared with controls. In a similar
analysis, DiLorenzo et al. (2007) separated exposed workers into intermediate
(9.1-29.4 (ig/dL) and high (29.4-81.1 (ig/dL) blood Pb level groups, with unexposed
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workers comprising the low exposure group (blood Pb levels 1-11 (ig/dL). Mean TNF-a
levels showed a monotonic increase from the low to high blood Pb group. Levels of
granulocyte colony-stimulating factor (G-CSF) did not differ between the intermediate
and high blood Pb groups among the Pb recyclers; however, G-CSF levels were higher in
the Pb recyclers than in the unexposed controls. Furthermore, among all subjects, blood
Pb showed a strong, positive correlation with G-CSF. Yucesoy et al. Q997a) found lower
serum levels of the Thl cytokines, IL-1(3 and IFN-y, in workers (mean blood Pb level of
59.4 (ig/dL) compared with controls (mean blood Pb level of 4.8 (ig/dL); however, levels
of the Th2 cytokines, IL-2 and TNF-a levels, were similar between groups. As most
occupationally-exposed cohorts represent populations highly exposed to Pb (with mean
blood Pb levels >22 (ig/dL), effects observed within these cohorts may not be
generalizable to the population as a whole.
In summary, animal, general population, and occupational studies suggest that exposure
to Pb increases the production of pro-inflammatory cytokines, skews the ratio of Thl and
Th2 cytokines to favor Th2 responses, and suppresses Thl lymphocyte cytokine
production.
4.2.6 Endocrine Disruption
4.2.6.1 Hypothalamic-Pituitary-Gonadal Axis
Evidence indicates that Pb is a potent endocrine disrupting chemical found to be
associated with reproductive and developmental effects in both male and female animal
models (see Section 4.8). Pb may act both at multiple points along the hypothalamic-
pituitary-gonadal (HPG) axis and directly at gonadal sites. The HPG axis functions in a
closely regulated manner to produce circulating sex steroids and growth factors required
for normal growth and development. Long-term Pb exposure in animals (21-35 day
dosing at 8 mg/kg) has been shown to alter serum levels of follicle-stimulating hormone
(FSH), luteinizing hormone (LH), testosterone, and estradiol (Biswas and Ghosh. 2006;
Rubio et al.. 2006). Similar changes in serum HPG hormones have been observed after
high-level Pb exposure in animals, resulting in blood Pb levels >20 (ig/dL (Dearth et al..
2002; Ronisetal.. 1998b: Foster. 1992; Sokol and Berman. 1991). Increases in serum LH
and FSH have been associated with increasing concurrent blood Pb levels in adult women
from the NHANES cohort (Krieg. 2007). The change in HPG hormones likely occurs
through the inhibition of LH secretion and the reduction in the expression of the
steroidogenic acute regulatory protein (StAR) (Huang and Liu, 2004; Srivastava et al..
2004; Huang et al.. 2002; Ronis et al.. 1996). StAR expression is the rate-limiting step
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essential in maintaining gonadotropin-stimulated steroidogenesis, which results in the
formation of testosterone and estradiol. Prenatal and lactational Pb exposure (via daily
gavage of dams beginning 30 days prior to breeding and resulting in 3 (ig/dL blood Pb in
the female rat offspring at PND31) was found to decrease basal StAR synthesis, but not
gonadotropin-stimulated StAR synthesis, suggesting that Pb may not directly affect
ovarian responsiveness to gonadotropin stimulation (Srivastava et al.. 2004). Instead, Pb
may act at the hypothalamic-pituitary level to alter LH secretion, which is necessary to
drive StAR production and subsequent sex hormone synthesis. Release of LH and FSH
from the pituitary is controlled by gonadotropin-releasing hormone (GnRH). Pb exposure
(10 (iM, 90 min) in rat brain median eminence cells can block GnRH release (Bratton et
al.. 1994). Pb may also interfere with release of pituitary hormones through interference
with cation-dependent secondary messenger systems that mediate hormone release and
storage.
Endocrine disruption may also be a result of altered hormone binding to endocrine
receptors. Prenatal and postnatal Pb exposure (20 ppm in drinking water) to rats was able
to decrease the number of estrogen, LH, and FSH receptors found in the uterus or ovaries
and receptor binding affinity (Wiebe etal., 1988; Wiebe and Barr. 1988). Altered
hormone binding ability may be due to the ion binding properties of Pb, resulting in
changes in receptor tertiary structure that will disrupt ligand binding. In addition,
Pb-induced changes in hormone levels that act as inducing agents for receptor synthesis
may affect the number of hormone receptors produced.
Some of these endocrine disrupting effects of Pb have been related to the generation of
ROS. Treatment with antioxidants has been able to counteract a number of the endocrine
disrupting effects of Pb, including apoptosis and decreased sperm motility and production
(Salawu et al.. 2009: Shan et al.. 2009: Madhavi et al.. 2007: Rubio et al.. 2006: Wang et
al.. 2006a: Hsu etal.. 1998b). Direct generation of ROS in epididymal spermatozoa was
observed after Pb treatment in rats (i.p. 20 or 50 ppm, 6 weeks) (Hsu et al.. 1998a). In
addition, testicular lipid peroxidation has been observed in Pb-treated rats (i.p.
0.025 ppm, 15 days) (Pandva et al.. 2012). Lipid peroxidation in the seminal plasma was
significantly increased in a group of Pb-exposed workers with high blood Pb levels
(>40 (ig/dL) than in unexposed controls (Kasperczyk et al.. 2008).
The liver is often associated with the HPG axis due in part to its production of insulin-
like growth factor 1 (IGF-1). Children with higher concurrent blood Pb levels (>4 (ig/dL)
(Huseman et al.. 1992) and Pb-exposed animals (blood Pb level of 14 (ig/dL) (Pine et al..
2006: Dearth et al., 2002) have shown a decrease in plasma IGF-1, which may be the
result of decreased translation or secretion of IGF-1 (Dearth et al.. 2002). IGF-1 release
was also inhibited by Pb treatment in gonadal cells (46 ppm Pb exposure) (Kolesarova et
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al.. 2010). IGF-1 also induces LH-releasing hormone release, such that IGF-1 decrements
may explain decreased LH and estradiol levels. IGF-1 production is stimulated by growth
hormone (GH) secreted from the pituitary gland and could be the result of GH depletion.
A number of studies have revealed that Pb exposure affects the dynamics of growth (see
Section 4.8.1). Decreased growth after Pb exposure could be the result of Pb-induced
decreased GH levels (Berry et al.. 2002; Camoratto et al.. 1993; Huseman et al.. 1992;
Huseman et al.. 1987). This decrease in GH could be a result of decreased release of GH
releasing hormone (GHRH) from the hypothalamus or disrupted GHRH binding to its
receptor, which has been reported in vitro after Pb treatment (IC50 free Pb in solution 5.2
x 10"5 (iM, 30 minutes) (Lau et al.. 1991). GH secretion may also be altered from
decreased testosterone resulting from Pb exposure.
4.2.6.2 Hypothalamic-Pituitary-Thyroid Axis
The evidence for the effects of Pb exposure on the hypothalamic-pituitary-thyroid (HPT)
axis is mixed. Pb exposure impacts a variety of components in the thyroid hormone
system. A number of occupational studies (blood Pb levels >7.3 (ig/dL) have shown that
elevated blood Pb is associated with lower thyroxine (T4) (and free T4 levels) without
alteration in triiodothyronine (T3), suggesting that long-term Pb exposure may depress
thyroid function in workers (Dundar et al.. 2006; Tuppurainen et al.. 1988; Robins et al..
1983). However, animal studies on thyroid hormones have shown mixed results.
Pb-exposed cows (blood Pb levels >51 (ig/dL) were reported to have an increase in
plasma T3 and T4 levels (Swarup et al.. 2007). whereas mice and chickens manifested
decreased serum T3 concentrations after Pb exposure, which was accompanied by
increased lipid peroxidation (Chaurasia et al.. 1998; Chaurasia and Kar. 1997). Both
decreased serum T3 and increased lipid peroxidation were restored by vitamin E
treatment, suggesting the disruption of thyroid hormone homeostasis could be a result of
altered membrane architecture and oxidative stress; however, no data were provided to
exclude changes in Pb kinetics as the mechanism of protection (Chaurasia and Kar.
1997).
Decreased T4 and T3 may be the result of altered pituitary release of thyroid stimulating
hormone (TSH). However, several studies have reported higher TSH levels in high-level
Pb-exposed workers (blood Pb levels >39 (ig/dL) (Lopez et al.. 2000; Singh et al.. 2000;
Gustafson et al.. 1989). which would result in increased T4 levels. Overall, results on the
effects of Pb on the HPT axis are inconclusive.
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4.2.7 Cell Death and Genotoxicity
A number of studies have attempted to characterize the genotoxicity of inorganic Pb in
human populations, laboratory animals, and cell cultures. Endpoints investigated include
DNA damage (single- and double-strand breaks, DNA-adduct formation), mutagenicity,
clastogenicity (sister chromatid exchange, micronucleus formation, chromosomal
aberrations), and epigenetic changes (changes in gene expression, DNA methylation,
mitogenesis). It is important to note that numerous studies have utilized exposure to
Pb chromate to investigate genotoxicity endpoints; some studies have specifically
attributed the observed increases in DNA damage and clastogenicity to the chromate ion
while others have not. Due to the uncertainty regarding whether observed genotoxic
effects are due to chromate or Pb in studies using this form of inorganic Pb, only studies
utilizing other forms of inorganic Pb (e.g., Pb nitrate, acetate, chloride, sulfate) are
discussed below. Overall, evidence indicates that in vitro or in vivo exposure to various
Pb compounds can increase risk of genotoxic effects, including DNA damage,
clastogenicity, and mutagenicity.
4.2.7.1 DNA Damage
A number of studies in human populations have observed associations between indicators
of Pb exposure and increased DNA damage, as measured as DNA strand breaks. Most of
these associations have been observed in occupationally-exposed populations (Grover et
al..201Q: Minozzo etal.. 2010: Shaik and Jamil. 2009: Danadevi et al.. 2003: Hengstler
etal.. 2003: Palus etal., 2003: Fracasso et al.. 2002: de Restrepo et al., 2000). Evidence
overall was equivocal in regard to how blood Pb levels correlated with DNA damage:
Fracasso et al. (2002) observed that DNA damage increased with increasing blood Pb
levels (blood Pb levels, <25, 25-35, and >35 (ig/dL), whereas Palus et al. (2003) (mean
blood Pb level: 50.4 (ig/dL [range: 28.2 to 65.5 ng/dL]) and Minozzo et al. (2010) (mean
[SD]: 59.43 (ig/dL [28.34]) observed no correlation. Hengstler et al. (2003) examined
workers exposed to Pb, Cd, and cobalt (Co) and observed that neither blood (mean: 4.4
[IQR: 2.84-13.6] (ig/dL) nor air Pb levels (mean: 3.0 [IQR: 1.6-50.0] (ig/m3) were
associated with DNA damage when examined alone, but that blood Pb influenced the
occurrence of single strand DNA breaks when included in a multiple regression model
along with Cd in air and blood and Co in air.
A few studies were found that investigated Pb-induced DNA damage resulting from
nonoccupational exposures. Mendez-Gomez (2008) observed that children attending
grade schools at close and intermediate distances to a Pb smelter had mean (range) blood
Pb levels of 28.6 (11.4 to 47.5) and 19.5 (11.3 to 49.2) ng/dL, respectively, compared to
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blood Pb level of 4.6 (0.1 to 8.7) (ig/dL for children living distant to the smelter. DNA
damage in lymphocytes was higher in children living nearest to the smelter, compared to
the children at the intermediate distance, but was not different from children living
farthest away from the smelter. Multivariate analysis (which considered children urinary
As levels, highest in children farthest from the smelter), revealed no statistically
significant associations between DNA damage and blood Pb level. Further, DNA repair
ability was also observed to be unrelated to blood Pb levels. Alternatively, Yanez et al.
(2003) observed that children living close to a mining complex (mean [range] blood Pb
level: 11.6 [3.0 to 19.5] (ig/dL) did have higher levels of DNA damage compared to
control children who lived further away from the mining facility (mean [range] blood Pb
level: 8.3 [3.0 to 25.0] (ig/dL).
Pb-induced DNA damage was observed in multiple animal studies. In mice exposed to Pb
(blood Pb level of 0.68 ug/dL) via inhalation for up to 4 weeks, differential levels of
DNA damage were observed in different organ systems, with only the lung and the liver
demonstrating statistically greater DNA damage compared to the respective organ
controls after acute exposure (Valverde et al.. 2002). Statistically elevated levels of DNA
damage were observed in the kidneys, lungs, liver, brain, nasal cavity, bone marrow, and
leukocytes of mice exposed to Pb over a period of 4 weeks, although variability was high
in all groups. The magnitude of the DNA damage was characterized as weak and did not
increase with increasing durations of exposure. In mice given Pb nitrate (0.7 to
89.6 mg/kg) by gavage for 24, 48, or 72 hours, or 1 or 2 weeks, single strand DNA breaks
in white blood cells were observed but did not increase with increasing Pb concentration
(Devi et al.. 2000). The three highest concentrations had responses that were similar in
magnitude to each other and were actually lower than the responses to the lower
concentrations tested. Xu et al. (2008) exposed mice to 10-100 mg/kg Pb acetate via
gavage for four weeks and observed a concentration-dependent increase in DNA single
strand breaks in white blood cells that was statistically significant at 50 and 100 mg/kg.
The authors characterized the observed DNA damage as severe. Pb nitrate induced DNA
damage in primary spermatozoa in rats (blood Pb levels of 19.5 and 21.9 (ig/dL) over that
in control rats (Nava-Hernandez et al.. 2009). The level of DNA damage was not
concentration dependent and was comparable in both exposure groups. Narayana and Al-
Bader (2011) observed no increase in DNA damage in the livers of rats exposed to 5,000
or 10,000 ppm Pb nitrate in drinking water for 60 days. Interestingly, although the results
were not statistically significant and were highly variable within exposure groups, DNA
fragmentation appeared to be lower in the exposed animals.
Studies investigating Pb-induced DNA damage in human cell cultures were
contradictory. Pb acetate did not induce DNA strand breaks in human HeLa cells when
exposed in vitro to 500 (iM Pb acetate for 20-25 hours or 100 (iM for 0.5-4 hours
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(Hartwig et al., 1990; Snyder and Lachmann. 1989). Pb nitrate, administered to
lymphoma cells in vitro at 1,000-10,000 (iM for 6 hours, did not result in any
DNA-protein crosslinks (Costa etal.. 1996). Pb acetate was observed by Wozniak and
Blasiak (2003) to result in DNA single and double strand breaks in primary human
lymphocytes exposed in vitro to 1-100 (iM for 1 hour, although the pattern of damage
was peculiar. DNA damage was greater in cells exposed to 1 or 10 (iM, compared to
those exposed to 100 (iM. DNA-protein crosslinks were only observed in the 100 (iM
exposure group, suggesting that the decreased strand breaks observed in the high
exposure group may be a result of increased crosslinking in this group. Pasha Shaik et al.
(2006) also observed DNA damage in human lymphocytes exposed in vitro to
2,100-3,300 (iM Pb nitrate for 2 hours. Although there was a concentration-dependent
increase in DNA damage from 2,100-3,300 (iM, no statistics were reported and no
unexposed control group was included, making it difficult to interpret these results.
Gastaldo et al. (2007) observed that in vitro exposure of human endothelial cells to
1-1,000 uM Pb nitrate for 24 hours resulted in a concentration-dependent increase in
DNA double strand breaks.
Studies in animal cell lines collectively were equally as ambiguous as those using human
cell lines. Zelikoff et al. (1988) and Roy and Rossman (1992) reported that Pb acetate
(concentration not stated and 1,000 uM, respectively) did not induce single or double
DNA strand breaks or DNA-protein or DNA-DNA crosslinks in CHV79 cells. However,
both Xu et al. (2006) and Kermani et al. (2008) reported Pb acetate-induced DNA
damage in undifferentiated PC12 cells exposed to 0.1, 1, or 10 (iM for 24 hours; and in
bone marrow mesenchymal stem cells exposed to 60 (iM for 48 hours, respectively.
Wedrychowski et al. (1986) reported that DNA-protein crosslinks were induced in a
concentration-dependent manner in hepatoma cells exposed to 50-5,000 (iM Pb nitrate
for 4 hours. Pb acetate and Pb nitrate increased the incidence of nick translation in
CHV79 cells when a bacterial DNA polymerase was added.
Pb exposure has also been shown to inhibit DNA repair mechanisms. Pb acetate did not
induce single strand DNA breaks in HeLa cells exposed to 500 uM for 20-25 hours
(Hartwig etal., 1990). However, exposure to both Pb acetate and UV light resulted in
increased persistence of UV-induced strand breaks, compared with exposure to UV light
alone. Similar effects were seen in hamster V79 cells: UV-induced mutation rates and
SCE frequency was exacerbated by co-incubation with Pb acetate. Taken together, these
data suggest that Pb acetate interferes with normal DNA repair mechanisms triggered by
UV exposure alone. Pb nitrate was observed to affect different DNA double strand break
repair pathways in human endothelial cells exposed in vitro to 100 uM for 24 hours.
Exposure to Pb inhibited nonhomologous end joining repair, but increased two other
repair pathways, MRE11-dependent and Rad51-related repair (Gastaldo et al.. 2007).
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Interestingly, in contrast to the above studies, exposure of lung carcinoma cells to 100,
300, or 500 uM Pb acetate for 24 hours resulted in an increase in nucleotide excision
repair efficiency (Li et al., 2008a). Roy and Rossman (1992) observed an increase in UV-
induced mutagenicity when CHV79 cells were co-exposed to 400 uM Pb acetate (a
nonmutagenic concentration of Pb acetate), indicating an inhibition of DNA repair.
Treatment of Chinese hamster ovary cells to 0.5-500 uM Pb acetate resulted in a
concentration-dependent accumulation of apurinic/apyrimidinic site incision activity,
indicating that DNA repair was diminished (McNeill et al.. 2007).
4.2.7.2 Mutagenicity
Only one human study was found that investigated Pb-induced mutagenicity. Van
Larebeke et al. (2004) investigated the frequency of mutations in the hypoxanthine
phosphoribosyltransferase (HPRT) gene in Flemish women without occupational
exposures to Pb or to a number of other heavy metals and organic contaminants. Higher
blood Pb level (1 Oth-90th percentile: 1.6 to 5.2 (ig/dL) was associated with greater HPRT
mutation frequency than found in the total population. Also, women with high blood Pb
levels (i.e., greater than the population median, not reported) demonstrated a greater
mutation frequency compared to women with lower blood Pb levels.
Pb-induced mutagenicity was investigated in a few studies using human cell cultures. Ye
(1993) exposed human keratinocytes to 100 (iM to 1 x 105 (iM Pb acetate for 2-24 hours.
This study did not measure HPRT mutations directly, but rather measured the amount of
tritium (3H) incorporated into DNA as an indicator of mutation. In the presence of
6-thioguanine, tritium incorporation was increased in exposed cells, indicating weak
mutagenicity. Hwua and Yang (1998) reported that Pb acetate was not mutagenic in
human foreskin fibroblasts exposed to 500-2,000 (iM for 24 hours. Pb acetate remained
nonmutagenic in the presence of 3-aminotriazole, a catalase inhibitor, indicating that
oxidative metabolism did not play a part in potential mutagenicity of Pb. Exposure to
Pb acetate alone did not induce mutagenicity in lung carcinoma cells (100-500 (iM for 24
hours) or fibroblasts (300-500 (iM for 24 hours) (Li et al.. 2008a: Wang et al.. 2008c).
However, pretreatment with PKC inhibitors before Pb treatment did result in statistically
significant increases in mutagenicity in both cell lines.
Results from investigations into Pb-induced mutagenicity using animal cell lines were as
equivocal as were the findings from human cell line studies, although the mixed findings
may be reflective of specific Pb compounds used. Pb acetate was observed to be
nonmutagenic (HPRT assay) in CHV79 cells exposed to 1-25 (iM of the compound for
24 hours (Hartwig et al.. 1990). but elicited a mutagenic response in CHV79 cells (gpt
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assay) exposed to 1,700 (iM for 5 days (Roy and Rossman. 1992). Pb acetate was
observed to be nonmutagenic (HPRT assay) in Chinese hamster ovary cells exposed to
5 (jJVI for 6 hours (McNeill et al.. 2007). The implication of mutagenicity in the latter
study is complicated by the concurrent observation of severe cytotoxicity at the same
concentration. Pb nitrate was alternatively found to be nonmutagenic in CHV79 cells (gpt
assay) exposed to 0.5-2,000 (iM for 5 days (Roy and Rossman. 1992) but mutagenic in
the same cell line (HPRT assay) exposed to 50-5,000 (iM for 5 days (Zelikoff et al..
1988). However, mutagenicity was only observed at 500 (JVI, and was higher than that
observed at higher Pb concentrations. Pb sulfate was also observed to be mutagenic in
CHV79 cells (HPRT assay) exposed to 100-1,000 (iM for 24 hours, but as with
Pb nitrate, it was not concentration-dependent (Zelikoff etal. 1988). Pb chloride was the
only Pb compound tested in animal cell lines that was consistently mutagenic: three
studies from the same laboratory observed concentration-dependent mutagenicity in the
gpt assay in Chinese hamster ovary cells exposed to 0.1-1 (iM Pb chloride for one hour
(Ariza and Williams. 1999: Arizaetal.. 1998: Ariza and Williams. 1996V
4.2.7.3 Clastogenicity
Clastogenicity is the ability of a compound to induce chromosomal damage, and is
commonly observed as sister chromatid exchange (SCE), micronuclei formation, or
incidence of chromosomal aberrations (i.e., breaks or gaps in chromosomes). Pb has been
shown to increase sister chromatid exchange, micronuclei formation, and chromosomal
aberrations in human populations, exposed animal models, and in vitro experiments.
Sister Chromatid Exchange
An association between blood Pb levels (means: 10.48 - 86.9 (ig/dL) and sister chromatid
exchange (SCE) was observed in a number of occupational studies (Wiwanitkit et al..
2008: Duvdu etal.. 2005: Palus etal.. 2003: Duvdu etal.. 2001: Pinto et al.. 2000:
Bilban. 1998: Anwar and Kamal. 1988: Huang etal.. 1988). In most studies that
attempted to investigate the concentration-response relationship in workers, no
association was observed between increasing blood Pb levels and the number of SCE
(Palus etal.. 2003: Duvduet al.. 2001: Pinto et al.. 2000V However, Huang et al. (1988)
did observe increased SCE in exposed workers in the two highest blood Pb groups (52.1
and 86.9 (ig/dL), with a statistically significant association observed in the 86.9 (ig/dL
group. Pinto et al. (2000) did report an association with duration of exposure (range of
years exposed: 1.6-40). Two studies reported no correlation between occupational
exposure to Pb and number of SCE (Rajah and Ahuja. 1996: Rajah and Ahuja. 1995).
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Mielzynska et al. (2006) found no association between blood Pb level and SCEs in
children in Poland. Children had an average blood Pb level of 7.69 (ig/dL and 7.87
SCEs/cell.
Pb exposure has been observed to induce SCEs in multiple laboratory animal studies. In
mice treated with up to 100 mg/kg Pb acetate i.p., Pb induced SCEs with 50 and
100 mg/kg (Fahmy. 1999). Pb nitrate, also administered i.p., induced the formation of
increased SCE levels in a concentration-dependent manner (10-40 mg/kg) in the bone
marrow of exposed mice (Dhir et al., 1993). Nayak et al. (1989b) treated pregnant mice
with 100-200 mg/kg Pb nitrate via i.v. injection and observed an increase in the number
of SCE in dams at 150 and 200 mg/kg; no increases in SCE levels were observed in the
fetuses. Tapisso et al. (2009) treated rats with 21.5 mg/kg Pb acetate (l/10th the LD50)
via i.p. injection on alternating days for 11 or21 days, foratotal of 5 or 10 treatments.
Induction of SCE in the bone marrow of exposed rats was increased over controls in a
statistically significant duration-dependent manner. It is important to note that all of these
studies utilized an injection route of exposure that may not be relevant to routes of
exposure in the human population (e.g., air, drinking water exposure).
Few studies were found that investigated SCE formation due to Pb exposure in human
cell lines. Statistically significant, concentration-dependent increases in SCE were
observed in human lymphocytes obtained from a single donor when incubated with 1,5,
10, or 50 (iM Pb nitrate (Ustundag and Duydu. 2007). Melatonin and N-acetylcysteine
were reported to ameliorate these effects, indicating Pb may induce increases in SCE
levels through increased oxidative stress. Pb chloride was also observed to increase SCE
levels in human lymphocytes exposed to 3 or 5 ppm (Turkez et al.. 2011).
Evidence from studies investigating SCE in rodent cells was more equivocal than that in
human cells. Pb sulfate, acetate, and nitrate were found not to induce SCE in CHV79
cells (Hartwig et al.. 1990; Zelikoff et al.. 1988). Both of these studies only examined
25-30 cells per concentration, reducing their power to detect Pb-induced increases in SCE
levels. Cai and Arenaz (1998). on the other hand, used 100 cells per treatment and
observed that exposure to 0.05-1 (iM Pb nitrate for 3-12 hours resulted in a weak,
concentration-dependent increase in SCE levels in Chinese hamster ovary cells. Lin et al.
(1994) also observed a concentration-dependent increase in SCE levels in Chinese
hamster cells exposed to 3-30 (iM Pb nitrate for 2 hours.
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Micronucleus Formation
Pb-induced micronucleus formation was observed in numerous occupational studies
(Groveretal..201Q: Khanetal.. 2010b: Minozzo etal.. 2010: Shaik and Jamil. 2009:
Minozzo et al.. 2004: Palus etal.. 2003: Vaglenov etal.. 2001: Pinto et al.. 2000: Bilban.
1998: Vaglenov et al.. 1998). Pinto et al. (2000) observed increased micronuclei in
exposed workers with an average blood Pb level of 10.48 (ig/dL compared with
unexposed controls. In studies investigating the correlation between blood Pb levels and
micronucleus formation, no association was observed (Minozzo et al.. 2010: Minozzo et
al.. 2004: Palus etal.. 2003: Pinto et al.. 2000). although Pinto et al. (2000). Grover et al.
(2010). and Minozzo et al. (2010) did report an association between micronuclei
formation and duration of exposure. Mielzynska et al. (2006) investigated micronucleus
formation in a nonworker population and reported a statistically significant positive
correlation between blood Pb levels and micronuclei frequency in children in Poland.
Children, with an average blood Pb level of 7.69 (ig/dL, were observed to have 4.44
micronucleated cells per 1,000 cells analyzed. Children with blood Pb levels greater than
10 (ig/dL had significantly more micronucleated cells than did children with blood Pb
levels less than 10 (ig/dL.
Micronucleus formation in response to Pb exposure has been observed in rodent animal
studies. Celik et al. (2005) observed that exposure of female rats to Pb acetate (140, 250,
or 500 mg/kg once per week for 10 weeks) resulted in statistically significant increases in
numbers of micronucleated polychromatic erythrocytes (PCEs) compared to controls.
Similarly, Alghazal et al. (2008b) exposed rats to Pb acetate (100 ppm daily for 125 days)
and observed statistically significant increases in micronucleated PCEs in both sexes.
Tapisso et al. (2009) treated rats with Pb acetate (21.5 mg/kg; l/10th the LD50) via i.p.
injection on alternating days for 11 or 21 days, for a total of 5 or 10 exposures. Formation
of micronuclei in the bone marrow of exposed rats was increased over formation in
controls in a significant duration-dependent manner. Two further studies investigated
formation of micronuclei in the bone marrow of exposed mice: Roy et al. (1992) treated
mice with Pb nitrate (10 or 20 mg/kg, i.p.) and observed a concentration-dependent
increase in micronuclei, whereas Jagetia and Aruna (1998) observed an increase in
micronuclei in mice treated with Pb nitrate (0.625-80 mg/kg, i.p.), though the increase
was not concentration-dependent. Mice exposed to Pb acetate (0.1 (ig/L via drinking
water, a more environmentally relevant route of exposure, for 90 days) had statistically
significant increases in micronucleated PCEs (Marques et al.. 2006).
A few studies were found that reported increased micronucleus formation in human cell
lines treated with Pb. Concentration-dependent micronucleus formation was observed in
human lymphocytes when exposed in vitro to either 1, 5, 10, or 50 (iM Pb nitrate or 3 or
5 ppm Pb chloride (Turkez etal.. 2011: Ustundag and Duydu. 2007). Gastaldo et al.
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(2007) also observed a concentration-dependent increase in micronuclei in human
endothelial cells exposed in vitro to 1-1,000 (iM Pb nitrate for 24 hours. Animal cell
culture studies investigating micronuclei formation produced contrasting results. One
study observed that micronuclei were not induced in Chinese hamster cells exposed to
3-30 (iM Pb nitrate for 2 hours (Lin et al.. 1994). whereas the other observed that
Pb acetate induced a concentration-dependent increase in Chinese hamster cells when
administered at 0.03-10 (iM for 18 hours (Bonacker et al., 2005).
Chromosomal Aberrations
Chromosomal aberrations (e.g., chromosome breaks, nucleoplasmic bridges, di- and a-
centric chromosomes, and rings) were examined in a number of occupational studies
(Groveretal..201Q: Shaik and Jamil. 2009; Pinto et al.. 2000; Bilban. 1998; De et al..
1995; Huang etal.. 1988). No correlation was observed between increasing blood Pb
level and the number of chromosomal aberrations, although an association was observed
between duration of exposure and chromosomal damage (Grover et al.. 2010; Pinto et al..
2000). Other studies reported no association between occupational exposure to Pb and
chromosomal aberrations (Anwar and Kamal. 1988; Andreae. 1983). Smejkalova (1990)
observed greater chromosomal damage and aberrations in children living in a heavily
Pb-contaminated area of Czechoslovakia compared with children living in an area with
less contamination, although the difference between the two areas was not statistically
significant. Blood Pb levels were comparable between children living in the
Pb-contaminated area and children living in the less contaminated area (low 30s versus
high 20s (ig/dL, respectively), indicating there may not be enough of a dose contrast to
detect a significant difference in aberration rates.
The majority of animal studies investigating Pb-induced genotoxicity focused on the
capacity of Pb to produce chromosomal damage. Fahmy (1999) treated mice with
Pb acetate (25-400 mg/kg i.p.), either as a single dose or repeatedly for 3, 5, or 7 days.
Chromosomal damage was observed to increase in bone marrow cells (100-400 mg/kg)
and spermatocytes (50-400 mg/kg) in a concentration-dependent manner after both
dosing regimens. Pb nitrate was also observed to produce concentration-dependent
chromosomal damage in mice treated i.p. to a single dosage of 5, 10, or 20 mg/kg (Dhir
et al.. 1992b). In a similar experiment, Dhir et al. (1990) treated mice with Pb nitrate (10,
20, or 40 mg/kg) and saw an increase in chromosomal aberrations, although there was no
concentration-dependent response as the response was similar in all concentrations tested.
Nayak et al. (1989b) treated pregnant mice with Pb nitrate (100, 150, or 200 mg/kg via
i.v. injection) and observed no chromosomal gaps or breaks in dams or fetuses but did
report some karyotypic chromosomal damage and weak aneuploidy at the low dose. In a
similar experiment, low levels of chromosomal aberrations were observed in dams and
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fetuses injected with Pb nitrate (12.5, 50, or 75 mg/kg), but there was no concentration-
dependent response reported and few cells were analyzed (Navaketal.. 1989a). In rats
given Pb acetate (2.5 mg/100 g body weight, i.p. daily for 5-15 days or 10-20 mg/100 g
once and analyzed after 15 days), Pb-induced chromosomal aberrations were observed
(Chakraborty et al.. 1987). The above studies all are limited by the use of a route of
exposure that may not be relevant to human environmental exposures. However, studies
utilizing oral exposures also observed increases in chromosomal damage. Aboul-Ela
(2002) exposed mice to Pb acetate (200 or 400 mg/kg by gavage for 5 days) and reported
that chromosomal damage was present in the bone marrow cells and spermatocytes of
animals exposed to both concentrations. Dhir et al. (1992a) also observed a
concentration-dependent increase in chromosomal damage in mice exposed via gavage,
albeit at much lower concentrations: either 5 or 10 mg/kg. Nehez et al. (2000) observed a
Pb-induced increase in aneuploidy and percent of cells with damage after exposure to
10 mg/kg administered by gavage 5 days a week for 4 weeks. In the only study that
investigated dietary exposure, El-Ashmawy et al. (2006) exposed mice to 5,000 ppm
Pb acetate in feed, and observed an increase in abnormal cells and frequency of
chromosomal damage.
In the few studies that investigated the capacity of Pb to induce chromosomal damage in
human cell lines, Pb exposure did not induce chromosomal damage. Wise et al. (2005;
2004) observed that Pb glutamate was not mutagenic in human lung cells exposed in vitro
to 250-2,000 (iM for 24 hours. Pasha Shaik et al. (2006) observed that Pb nitrate did not
increase chromosomal aberrations in primary lymphocytes (obtained from healthy
volunteers) when incubated with 1,200 or 2,000 (iM for 2 hours. Studies utilizing animal
cell lines generally supported the finding of no Pb-induced chromosomal damage in
human cell lines. Pb nitrate was found to induce no chromosomal damage in Chinese
hamster ovary cells exposed to 500-2,000 (iM for 24 hours (Wiseetal.. 1994). 3-30 (iM
for 2 hours (Linetal.. 1994). or 0.05-1 (JVI for 3-12 hours (Cai and Arenaz. 1998). Wise
et al. (1994) did observe increased chromosomal damage in Chinese hamster ovary cells
exposed to 1,000 (iM Pb glutamate for 24 hours but did not see any damage in cells
exposed to higher concentrations (up to 2,000
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4.2.7.4 Epigenetic Effects
Epigenetic effects are heritable changes in gene expression resulting without changes in
the underlying DNA sequence. A prime example of an epigenetic effect is the abnormal
methylation of DNA, which could lead to altered gene expression and cell proliferation
and differentiation. Possible indications of Pb-induced epigenetic changes include
alterations in methylation patterns in exposed rats, and alterations in mitogenesis and cell
proliferation in exposed humans and animals, as well as human and animal cell cultures.
DNA Methylation
A single i.v. injection of Pb nitrate (75 umol/kg) resulted in global hypomethylation of
hepatic DNA in rats (Kanduc et al.. 1991). The observed hypomethylation in the liver
was associated with an increase in cell proliferation. A few additional studies in humans
observed that higher bone Pb levels were associated with lower global DNA methylation
patterns in adults and cord blood of newborns (Wright et al.. 2010; Pilsner et al.. 2009).
Hypomethylation specifically is associated with increased gene expression. Changes in
DNA methylation patterns could potentially lead to dysregulation of gene expression and
altered tissue differentiation.
Mitogenesis
Conflicting results have been reported regarding Pb-induced effects on mitogenesis, with
both increased and decreased cell growth and mitogenesis. A discernible pattern of
effects is difficult to detect when analyzing effects across human, in vivo animal, and in
vitro studies. Only a few studies have investigated the epigenetic effects of Pb exposure
in human populations indirectly by examining mitogenesis or the induction of cell
proliferation, which can be a consequence of epigenetic changes. These studies (Minozzo
et al.. 2010; Minozzo et al.. 2004; Rajah and Ahuja. 1995) reported reduced mitogenesis
in two groups of Pb-exposed workers compared with unexposed controls (mean blood Pb
levels: 35.4 (ig/dL, 59.4 (ig/dL, and not reported, respectively). The observation of
decreased cell division in exposed workers may indicate that cells suffered DNA damage
and died during division, or that division was delayed to allow for DNA repair to occur. It
is also possible that Pb exerts an aneugenic effect and arrests the cell cycle.
Many studies have investigated the ability of Pb to induce mitogenesis in animal models,
and have consistently shown that Pb nitrate can stimulate DNA synthesis and cell
proliferation in the liver of animals treated with 100 (iM Pb per kg body weight, via i.v.
injection (Nakajima et al.. 1995; Coni etal.. 1992; Ledda-Columbano et al.. 1992;
Columbano et al., 1990; Columbano et al., 1987). Shinozuka et al. (1996) observed that
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Pb-induced hepatocellular proliferation was similar in magnitude to that induced by
TNF-a at 100 (iM/kg; and Pb was observed to induce TNF-a in glial and nerve cells in
mice (and NF-KB, TNF-a, and iNOS in rat liver cells) from mice treated with Pb at
12.5 mg/kg and 100 (imol/kg, respectively (Cheng et al.. 2002; Menegazzi et al. 1997).
The only study that examined Pb exposure via inhalation (Pb acetate, 10,000 (iM for
4 weeks) found increased cellular proliferation in murine lungs (Fortoul et al.. 2005).
Extensive research has been conducted investigating the potential effects of Pb on
mitogenesis in human and animal cell cultures. In human cell cultures, Pb acetate
inhibited cell growth in hepatoma cells (0.1-100 (iM for 2-6 days) (Heiman and Tonner.
1995) and primary oligodendrocyte progenitor cells (1 (iM for 24 hours) (Deng and
Poretz. 2002) but had no observable effects on growth in glioma cells (0.01-10 (iM for
12-72 hours) (Liu et al.. 2000). Pb glutamate had no effect on cell growth in human lung
cells in vitro, but did increase the mitotic index (250-1,000 (iM exposure for 24 hours)
(Wise et al.. 2005). The increase in the mitotic index was attributed to an arrest of the cell
cycle at M-phase, and was not attributed to an actual increase of cell growth and
proliferation. Gastaldo et al. (2007) also reported S and G2 cell cycle arrests in human
endothelial cells following exposure to 100 uM Pb nitrate for 24 hours. Conflicting
results with regard to DNA synthesis were reported, with a concentration-dependent
inhibition of DNA synthesis reported in hepatoma cells (1-100 (iM for 72 hours) (Heiman
and Tonner. 1995). but an induction of synthesis observed in astrocytoma cells (1-50 (iM
for 24 hours) (Lu et al.. 2002).
In rat fibroblasts and epithelial cells, Pb acetate, Pb chloride, Pb oxide, and Pb sulfate
were all observed to inhibit cell growth (10-1,000 (iM for 1-7 days and 0.078-320 (iM for
48 hours, respectively) (lavicoli et al.. 2001; Apostoli et al.. 2000). lavicoli et al. (2001)
observed that in addition to inhibiting cell growth in rat fibroblasts, Pb acetate caused
GS/M and S-phase arrest. Pb acetate decreased cell proliferation in mouse bone marrow
mesenchymal stem cells when administered at 20-100 uM for 48 hours (Kermani et al..
2008). Pb nitrate was alternatively reported to increase (Lin et al.. 1994) and decrease
(Cai and Arenaz. 1998) the mitotic index in Chinese hamster ovary cells exposed to 1 uM
Pb nitrate. Lin et al. (1994) did not consider cell cycle arrest when measuring the mitotic
index and did not observe a decrease at higher concentrations; in fact, the highest
concentration tested, 30 uM, had a mitotic index equal to that in the untreated control
cells.
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4.2.7.5 Gene Expression
A few animal studies have investigated the ability of Pb exposure to alter gene expression
in regard to phase I and II metabolizing enzymes. Suzuki et al. (1996) treated rats with
Pb acetate or Pb nitrate (100 ug/kg via i.p. injection) and observed an induction of GST-P
with both Pb compounds. The induction of GST-P by Pb was observed to occur on the
transcriptional level and to be dependent on the direct activation of the cis-element GPEI
enhancer. Degawa et al. (1993) reported that Pb nitrate (20, 50, or 100 umol/kg, via i.v.)
selectively inhibited CYP1A2 levels. Pb was shown to inhibit CYP1A2 not by direct
enzyme inhibition, but rather to decrease the amount of CYP1A2 mRNA. In contrast,
Korashy and El Kadi (2004) observed that exposure of murine hepatoma cells to
Pb nitrate (10-100 uM for 24 hours) increased the amount of CYP1A1 mRNA while not
influencing the activity of the enzyme. NAD(P)H:quinone oxidoreductase and GST Ya
activities and mRNA levels were increased after exposure to Pb. Incubation of primary
human bronchial epithelial cells with Pb acetate (500 ug/L for 72 hours) resulted in the
up-regulation of multiple genes associated with cytochrome P450 activity, glutathione
metabolism, the pentose phosphate pathway, and amino acid metabolism (Glahn et al..
2008).
Additional animal studies provide further evidence that exposure to Pb compounds can
perturb gene expression. Zawia and Harry (1995) investigated whether the observed
Pb-induced disruption of myelin formation in rat pups exposed postnatally was due to
altered gene expression. In pups exposed to 2,000 ppm Pb acetate via lactation from
PND1-PND20, the expression of proteolipid protein, a major structural constituent of
myelin, was elevated (statistically significant) at PND20, compared to controls. The
expression of another structural element of myelin (myelin basic protein) was similarly
elevated in exposed animals, although not significantly so. The expression of both genes
returned to control levels 5 days following the termination of exposure. These data
suggest that altered gene expression in structural myelin proteins due to Pb exposure may
be responsible for observed alterations in abnormal conduction of nerve impulses. Long
et al. (2011) investigated the Pb-induced increase in ABCC5, an ATP-binding cassette
transporter, in embryonic and adult zebrafish. In the initial in vitro portion of the study,
exposure of zebrafish fibroblasts to 20 (iM Pb nitrate for 24 hours significantly increased
the induction of ABBC5 mRNA 2.68-fold over controls. Similar levels of induction were
observed when embryonic zebrafish were exposed to 5 (iM for 24 to 96 hours;
specifically, induction of ABCC5 was seen in the livers of developing embryos. In adult
fish, induction of ABCC5 was observed in the brains, intestines, and kidneys of exposed
fish, but a decrease was found in their livers. Induction of ABCC5 in adult fish was
observed to attenuate the toxicity of Cd (but not Hg or As); however, in developing
embryos, the attenuation of Pb-induced toxicity was not investigated. These findings
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indicate that increased expression of ABCC5 due to heavy metal exposure may play a
part in cellular defense mechanisms.
4.2.7.6 Apoptosis
Occupational exposure to Pb and induction of apoptosis in various cell types was
investigated in a few studies. The study that directly measured apoptosis reported that
exposure to Pb increased apoptosis of lymphocytes compared to nonexposed controls
(Minozzo et al.. 2010). whereas the others reported that two early indicators of apoptosis,
karyorrhexis and karyolysis, were elevated in occupationally exposed workers (Grover et
al.. 2010; Khan et al.. 2010b). Pb nitrate was also observed to induce apoptosis in the
liver of exposed animals (Columbano et al.. 1996; Nakajima et al.. 1995). Apoptosis was
observed in rat fibroblasts exposed in vitro to Pb acetate and rat alveolar macrophages
exposed to Pb nitrate davicoli et al.. 2001; Shabani and Rabbani. 2000). Observation of
Pb-induced apoptosis may represent the dysregulation of genetically-controlled cell
processes and tissue homeostasis.
4.2.8 Summary
The diverse health effects of Pb are mediated through multiple, interconnected modes of
action. Each of the modes of action discussed here has the potential to contribute to the
development of a number of Pb-induced health effects (Table 4-2). While this section
draws from earlier literature as well as newer lines of evidence, the inclusion of recent
evidence does not qualitatively change the previous conclusions regarding individual
modes of action. Rather, the more recent evidence agrees with, and thus strengthens these
conclusions. Evidence for the majority of these modes of action is observed with
concurrent blood Pb levels in humans ranging between 2 and 17 (ig/dL, with supporting
evidence from animal and in vitro assays. As many of these studies examined adults, with
likely higher past than current Pb exposures, uncertainty exists as to the Pb exposure
level, duration, frequency, and timing associated with these modes of action. The blood
Pb levels or in vitro concentrations presented in Table 4-2 reflect the current evidence for
these modes of action and are not intended to convey conclusions regarding specific
thresholds. Also, the data presented in this table do not inform the exposure frequency
and duration required to elicit a particular MOA.
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Table 4-2 MOAs, their related health effects, and information on concentrations
eliciting the MOAs.
Mode of Action
[Related Health Effects
(ISA Section )]
Concentrations or Doses (Conditions)3
Blood Pb
Dose
Altered Ion Status
[All Health Effects of Pb,
Chapter 4]
3.5 |jg/dL
(Mean in cord blood; association with
cord blood Ca2+ATPase pump activity)
Huel et al. (2008)
0.00005 uM free Pb
(In vitro; 30 minutes; calmodulin
activation assay)
Kern et al. (2000)
Protein Binding
[Renal (45),
Hematological Effects (4.7)1
17.0ug/dL
(Concurrent mean in adult workers with
wildtype metallothionein expression;
increased BP susceptibility)
Chen et al. (201 Oa)
50 uM Pb glutamate
(In vitro; 24 hours; increased nuclear
protein in neurological cell)
Klann and Shelton (1989)
Oxidative Stress
[All Heath Effects of Pb,
Chapter 4]
5.4 ug/dl_
(Concurrent mean in adult male
workers; decreased CAT activity in
blood)
Conterato et al. (2013)
0.1 uM Pb acetate
(In vitro; 48 hours; decreased cellular
GSH in neuroblastoma cells)
Chetty et al. (2005)
Inflammation
[Nervous System (4.3),
Cardiovascular (4.4),
Renal (4.5), Immune (4.6),
Respiratory (4.9.6),
Hepatic (4.9.1)1
Among males with concurrent blood Pb
>2.5ug/dL
(Increased serum TNF-a and blood
WBC count)
Kim et al. (2007)
0.01 uM Pb acetate
(In vitro; 48 hours; increased cellular
PGE2 in neuroblastoma cells)
Chetty et al. (2005)
Endocrine Disruption
[Reproductive and
Developmental Effects (4.8),
Endocrine System (4.9.3),
Bone and Teeth (4.9.4)1
1.7ug/dL
(Lowest level at which a relationship
could be detected in adult women with
both ovaries removed; increased serum
FSH)
Krieg (2007)
10 uM Pb nitrate
(In vitro; 30 minutes; displaced GHRH
binding to rat pituitary receptors)
Lau et al. (1991)
Cell Death/Genotoxicity
[Cancer (4.10),
Reproductive and
Developmental Effects (4.8).
Bone and Teeth (4.9.4)1
3.3 ug/dl_
(Concurrent median in adult women;
increased rate of HPRT mutation
frequency)
Van Larebeke et al. (2004)
0.03 uM Pb acetate
(In vitro; 18 hours; increased formation
of micronuclei)
Bonacker et al. (2005)
aThis table provides examples of studies that report effects with low Pb dosages or concentrations; they are not the full body of
evidence used to characterize the weight of the evidence. In addition, the levels cited are reflective of the data and methods
available and do not imply that these modes of action are not acting at lower Pb exposure or blood Pb levels or that these doses
represent the threshold of the effect. Additionally, the blood concentrations and doses (indicating Pb concentrations from in vitro
systems) refer to the concentrations and doses at which these modes of action were observed. While the individual modes of
action are related back to specific health effects sections (e.g., Nervous System, Cardiovascular), the concentrations and doses
given should not be interpreted as levels at which those specific health effects occur. Also, the data presented in this table do not
inform the exposure frequency and duration required to elicit a particular MOA.
The alteration of cellular ion status (including disruption of Ca2+ homeostasis, altered ion
transport mechanisms, and perturbed protein function through displacement of metal
cofactors) appears to be the major unifying mode of action underlying all subsequent
modes of action (Figure 4-1). Pb is well characterized to interfere with endogenous Ca2+
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homeostasis (necessary as a cell signal carrier mediating normal cellular functions).
[Ca2+]j has been shown to increase after Pb exposure in a number of cell types including
bone, erythrocytes, brain cells, and white blood cells, due to the increased flux of
extracellular Ca2+ into the cell. This disruption of ion transport is due in part to the
alteration of the activity of transport channels and proteins, such as Na+/K+ATPase and
voltage-gated Ca2+ channels. Pb can interfere with these proteins through direct
competition between Pb and the native metals present in the protein metal binding
domain or through disruption of proteins important in Ca2+-dependent cell signaling, such
as PKC or calmodulin.
Disruption of ion transport not only leads to altered Ca2+ homeostasis, but it can also
result in perturbed neurotransmitter function. Pb has been shown to displace metal ions,
(such as Zn2+, Mg2+, and Ca2+) from proteins due to the flexible coordination number of
Pb and multiple ligand binding ability, leading to abnormal conformational changes in
proteins and altered protein function. Evidence for this metal ion displacement and
protein perturbation has been shown at 10~6 (iM concentrations of Pb. Additional effects
of altered cellular ion status are the inhibition of heme synthesis and decreased cellular
energy production due to perturbation of mitochondrial function.
Although Pb can bind to proteins within cells through interactions with side group
moieties, thus potentially disrupting cellular function, protein binding of Pb may
represent a mechanism by which cells protect themselves against the toxic effects of Pb.
Intranuclear and intracytosolic inclusion body formation has been observed in the kidney,
liver, lung, and brain following Pb exposure. A number of unique Pb binding proteins
have been detected, constituting the observed inclusion bodies. The major Pb binding
protein in blood is ALAD with carriers of the ALAD-2 allele potentially exhibiting
higher Pb binding affinity. Additionally, metallothionein is an important protein in the
formation of inclusion bodies and mitigation of the toxic effects of Pb.
A second major mode of action of Pb is its role in the development of oxidative stress,
due in many instances to the antagonism of normal metal ion functions. The origin of
oxidative stress produced after Pb exposure is likely a multipathway process, resulting
from oxidation of 5-ALA, NAD(P)H oxidase activation, membrane and lipid
peroxidation, and antioxidant enzyme depletion. Through the inhibition of 5-ALAD (due
to displacement of Zn by Pb), accumulated 5-ALA goes through an auto-oxidation
process to produce ROS. Additionally, Pb can induce the production of ROS through the
activation of NAD(P)H oxidase. Pb-induced ROS can interact with membrane lipids to
cause a membrane and lipid peroxidation cascade. Enhanced lipid peroxidation can also
result from Pb potentiation of Fe2+ initiated lipid peroxidation and alteration of membrane
composition after Pb exposure. Increased Pb-induced ROS can also sequester and
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inactivate biologically active NO, leading to the increased production of the toxic product
nitrotyrosine, increased compensatory NOS, and decreased sGC protein. Pb-induced
oxidative stress not only can result from increased ROS production but also through the
alteration and reduction in activity of the antioxidant defense enzymes. The biological
actions of a number of these enzymes are antagonized due to the displacement of the
protein functional metal ions by Pb.
In a number of organ systems, Pb-induced oxidative stress is accompanied by
misregulated inflammation. Pb exposure can modulate inflammatory cell function,
production of pro-inflammatory cytokines and metabolites, inflammatory chemical
messengers, and pro-inflammatory signaling cascades. Cytokine production is skewed
toward the production of pro-inflammatory cytokines like TNF-a and IL-6 as well as
toward the promotion of a Th2 response and suppression of a Thl response accompanied
by a corresponding shift in the production of related cytokines.
Evidence indicates that Pb is a potent endocrine disrupting chemical. Pb can disrupt the
HPG axis evidenced by altered serum hormone levels, such as FSH, LH, testosterone,
and estradiol. Pb can interact with the hypothalamic-pituitary level hormone control
causing a decrease in pituitary hormones, alteration of growth dynamics due to decreased
IGF-1, inhibition of LH secretion, and reduction in StAR protein. Pb has also been shown
to alter hormone receptor binding likely due to interference of metal cations with
secondary messenger systems and receptor ligand binding and through generation of
ROS. Pb also may disrupt the HPT axis by alteration of a number of thyroid hormones,
possibly due to oxidative stress. However, the results of these studies investigating HPT
are mixed.
The association of Pb with increased genotoxicity and cell death has been investigated in
humans, animals, and cell models. Occupational Pb exposure in humans has been
associated with increased DNA damage; however, lower blood Pb and exposure levels
also have been associated with these effects in experimental animals and cells. While not
entirely consistent, a number of studies reported decreased repair processes following Pb
exposure. There is evidence of mutagenesis and clastogenicity in highly-exposed
humans; however, weak evidence has been shown in animals and cell based systems.
Human occupational studies provide limited evidence for micronucleus formation (blood
Pb levels >10 (ig/dL) and are supported by Pb-induced effects in both animal and cell
studies at higher exposure levels. Animal studies have also provided evidence for
Pb-induced chromosomal aberrations. The observed increases in clastogenicity may be
the result of increased oxidative damage to DNA due to Pb exposure, as co-exposures
with antioxidants ameliorate the observed toxicities. Limited evidence of epigenetic
effects is available, including abnormal DNA methylation, mitogenesis, and gene
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expression. Pb may alter gene expression by displacing Zn from multiple transcriptional
factors, thus perturbing their normal cellular activities. Consistently positive results have
provided evidence of increased apoptosis following Pb exposure.
Similar to Pb, other polyvalent metal ions (e.g., Cd, Cr, Be, Ba, Se, Sr, As, Al, Cu) have
demonstrated molecular mimicry and displacement of biological cations (Garza et al.
2006). In this manner, these metal ions share with Pb a common central mode of action of
disruption of ion status. Specifically, these metals have been shown to disrupt cellular
processes as diverse as Ca2+ homeostasis, cell signaling, neurotransmitter release, cation
membrane channel function, protein-DNA binding, and cellular membrane structure
(Pentyalaetal.. 2010; Huang et al.. 2004; Atchison. 2003; Jehan and Motlag. 1995;
Richardt et al., 1986; Cooper and Manalis. 1984; Habermann et al., 1983). Additionally,
presumably through their shared central mode of action, some of these metal ions also
display corresponding downstream modes of actions such as oxidative stress, apoptosis,
and genotoxicity (Jomova and Valko. 2011; Jomova et al.. 2011; Matovic et al.. 2011;
Agarwal et al.. 2009; Mendez-Gomez et al.. 2008; Rana. 2008; Hengstleretal.. 2003).
Overall, Pb-induced health effects can occur through a number of interconnected modes
of action that generally originate with the alteration of ion status.
4.3 Nervous System Effects
4.3.1 Introduction
The 2006 Pb AQCD concluded that the "overall weight of the available evidence
provides clear substantiation of neurocognitive decrements being associated in young
children with blood-Pb concentrations..." (U.S. EPA. 2006b). This conclusion was based
on evidence from several prospective and cross-sectional epidemiologic studies
conducted in diverse populations with adjustment for potential confounding by
socioeconomic status (SES) as well as parental intelligence, education, and caregiving
quality and stimulation. The association between blood Pb levels and cognitive function
decrements was further supported by an international pooled analysis of children, ages
4.8 to 10 years, participating in seven prospective studies (Boston, MA; Cincinnati, OH;
Rochester, NY; Cleveland, OH; Mexico City, Mexico; Port Pirie, Australia; and Kosovo,
Yugoslavia) (Lanphear et al.. 2005). Across all previously evaluated studies, associations
between blood Pb levels and decrements in full-scale intelligence quotient (FSIQ), infant
mental development, memory, learning, and executive function were found in children
ages 2 to 17 years with population mean blood Pb levels (measured at various lifestages
and time periods) 5-10 (ig/dL; however, several results indicated associations in groups
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of children (ages 2-10 years) with mean blood Pb levels in the lower range of 3-5 (ig/dL
(Bellinger. 2008: Canfield. 2008: Hornung. 2008: Tellez-Roio. 2008). Based on fewer
available studies, the 2006 Pb AQCD described evidence from prospective and cross-
sectional epidemiologic studies for associations of childhood blood Pb levels with
attention decrements in children ages 6-17 years and measures of conduct disorders in
children ages 4-17 years and young adults ages 18-21 years (U.S. EPA. 2006b).
Biological plausibility for epidemiologic evidence in children was provided by similarly
consistent toxicological findings for Pb-induced impairments in learning and behavior in
rodents and monkeys (U.S. EPA. 2006b). Pb exposure was not found consistently to
affect the memory of animals. In animals, learning impairments were demonstrated
largely as poorer performance in maze tests, shorter interresponse times on schedule
controlled behavior tasks, and response perseveration errors in discrimination reversal
tests. Some results from these tests also indicated Pb-induced increases in impulsivity.
Pb-induced impulsivity in animals also was demonstrated as increased response rates on
the Fixed Ratio (FR)/waiting for reward test. These effects on learning and behavioral
impairments in animals were found predominately with Pb exposures that resulted in
blood Pb levels 30-50 (ig/dL; however, some studies observed these impairments in
rodents (pre- and post-natal Pb exposure) and monkeys (postnatal Pb exposure) with
blood Pb levels 14-25 (ig/dL (Kuhlmann et al.. 1997: Altmann et al.. 1993: Rice and
Karpinski. 1988: Gilbert and Rice. 1987). Toxicological studies provided additional
biological plausibility for Pb-induced impairments in learning and behavior by
demonstrating effects on endpoints related to modes of action. Evidence for Pb affecting
neuronal development and function at the cellular and subcellular level (e.g., blood brain
barrier integrity, synaptic architecture during development, neurite outgrowth, glial
growth, neurotransmitter release, oxidative stress), provided biological plausibility for
associations observed in children between blood Pb levels and decrements in multiple
functional domains such as cognitive function, memory, motor function, mood, and
behavioral problems. Additional biological plausibility was provided by associations
observed of childhood blood Pb levels with changes indicative of neuronal damage and
altered brain physiology in small groups of children (Meng et al.. 2005: Trope et al..
2001) and young adults (Yuan et al., 2006: Cecil et al., 2005) assessed using magnetic
resonance imaging techniques.
A common finding across several different populations of children was a supralinear
concentration-response relationship between blood Pb level and cognitive function
decrements, i.e., a larger decrement in cognitive function per unit increase in blood Pb
level in children in the lower range of the study population blood Pb level distribution.
Particularly compelling evidence was provided by studies that examined prenatal or early
childhood blood Pb levels or considered peak blood Pb levels in school-aged children
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(Schnaas et al., 2006; Bellinger and Needleman. 2003; Canfield et al.. 2003a) or
examined concurrent blood Pb levels in young children age 2 years. (Tellez-Rojo et al..
2006). Most of these epidemiologic studies used a cut-point of 10 (ig/dL to define lower
and higher blood Pb levels. Consistent with these individual study findings, results from
the pooled analyses of seven cohorts indicated that a nonlinear relationship fit the data
better than a linear relationship (Lanphear et al.. 2005; Rothenberg and Rothenberg.
2005). Explanations for the supralinear concentration-response were not well
characterized.
Another area of focus was the comparison of various lifestages and time periods of Pb
exposure with respect to increased risk of neurodevelopmental decrements. Animal
studies clearly demonstrated that gestational Pb exposure with or without additional early
postnatal exposure resulted in neurodevelopmental impairments. However, gestational Pb
exposure was not necessary as neurodevelopmental effects in animals also occurred with
postnatal juvenile-only and lifetime Pb exposures. Epidemiologic studies observed
decrements in cognitive function in children ages 2 to 17 years in association with
prenatal, peak childhood, cumulative childhood, and concurrent blood Pb levels.
Although examined in few studies, tooth or bone Pb levels were associated with cognitive
function decrements and behavioral problems in children and adolescents rWasserman et
al.. 2003; Bellinger et al.. 1994b; Fergusson et al.. 1993; Needleman et al.. 1979).
potentially pointing to an effect of cumulative or early childhood Pb exposure. Among
studies of children ages 3-10 years that examined blood Pb levels measured at multiple
lifestages and time periods, several found that concurrent blood Pb was associated with a
similar magnitude or larger decrement in FSIQ than blood Pb levels measured earlier in
childhood or averaged over multiple years (Lanphear et al.. 2005; Canfield et al.. 2003a;
Wasserman et al., 1994; Dietrich et al.. 1993b). A common limitation of studies of Pb and
neurodevelopment in children is the high correlation among blood Pb levels at different
ages, making it difficult to identify an individual critical lifestage or duration of Pb
exposure associated with risk of neurodevelopmental decrements (Lanphear et al.. 2005).
Some evidence indicated the persistence of neurodevelopmental effects of Pb exposure,
by associations of biomarkers of earlier childhood Pb exposure (e.g., deciduous tooth Pb,
blood Pb prenatally or at age 2 or 6 years) with cognitive function decrements and
behavioral problems later in childhood or as young adults (Ris et al., 2004; Wasserman et
al.. 2003; Wasserman et al.. 2000a; Bellinger et al.. 1994a; Bellinger etal.. 1994b;
Fergusson et al.. 1993; Baghurst et al., 1992; Needleman et al., 1979). Persistence of
effects also was demonstrated by findings in some studies of rats and monkeys that
gestational and/or early postnatal Pb exposures were associated with impairments in
cognitive function and behavior in animals evaluated as adults (Kuhlmann et al.. 1997;
Altmann et al., 1993; Rice. 1992b. 1990). On the other hand, a few epidemiologic studies
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in children ages 5-7 years demonstrated lack of persistence with findings of higher FSIQ
in children who had decreases in blood Pb over time compared with children with no
change or an increase in blood Pb level (Wasserman et al., 2000a; Bellinger et al., 1990).
In epidemiologic studies of adults, a range of nervous system effects (e.g., decrements in
memory, attention, reaction time, visuomotor tasks and reasoning, alterations in visual or
brainstem evoked potentials, postural sway) were clearly indicated in Pb-exposed
workers with blood Pb levels in the range of 14 to 89 ug/dL dwataet al.. 2005; Bleecker
et al., 1997; Baker et al., 1979). In the smaller body of studies examining
nonoccupationally-exposed adults, poorer cognitive performance was associated with
bone Pb levels (Weisskopf et al.. 2004; Wright et al.. 2003) but less so with concurrent
blood Pb levels (Kriegetal.. 2005; Wright etal.. 2003; Nordberg et al.. 2000; Pavton et
al.. 1998; Muldoon et al.. 1996). These findings suggested the influence of past or
cumulative Pb exposures on cognitive function decrements in nonoccupationally-exposed
adults. With regard to neurodegenerative diseases, whereas a few studies of aged animals
found Pb-induced amyloid plaques, a pathology commonly found in the brains of adults
with Alzheimer's disease (Bashaetal.. 2005; Zawia and Basha. 2005). epidemiologic
studies did not indicate that Pb exposure was associated with Alzheimer's Disease in
adults. Blood and bone Pb levels were inconsistently associated with amyotrophic lateral
sclerosis (ALS) in adults in the general population; however, in some case-control
studies, history of occupational Pb exposure was more prevalent among ALS cases than
controls (Kamel et al.. 2002; Chancellor etal.. 1993). Associations were reported for
essential tremor and symptoms of anxiety and depression in adults, but each was
examined in only a few studies.
As discussed throughout this section, recent epidemiologic and toxicological studies
continued to demonstrate associations of Pb exposure and biomarkers of Pb exposure
with nervous system effects. The strongest evidence continued to be derived from
associations observed for Pb exposure in animals and blood Pb levels in children with
cognitive function decrements. Several recent studies in children expanded the evidence
for associations between concurrent blood Pb levels and attention decrements,
impulsivity, and hyperactivity. Recent epidemiologic studies in adults focused primarily
on cognitive function decrements but provided additional evidence for Pb-associated
psychopathological effects, ALS, Parkinson's disease, and essential tremor. Recent
toxicological evidence supported the effects of prenatal and postnatal Pb exposure on
learning, memory, and impulsivity in animals and examined interactions between Pb
exposure and stress. New or expanded areas of toxicological research related to Pb
exposure included neurofibrillary tangle formation and neurodegenerative effects after
early-life Pb exposures. Recent toxicological studies added to the large extant evidence
base for Pb-induced effects on endpoints describing modes of action, including
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neurotransmitters, synapses, glia, neurite outgrowth, the blood brain barrier, and
oxidative stress. The evidence detailed in the subsequent sections continue to enhance the
understanding of the spectrum of nervous system effects associated with Pb exposure.
4.3.2 Cognitive Function
In children, blood and tooth Pb levels have been linked with several measures of
cognitive function that are inter-related, which makes it difficult to assess whether Pb
exposure has an effect on a particular domain. Epidemiologic studies have assessed
cognitive function extensively by FSIQ in children ages 3 to 17 years. FSIQ has strong
psychometric properties (i.e., reliability, consistency, validity), is among the most
rigorously standardized cognitive function measures, is relatively stable in school-age,
and has been predictive of educational achievement and life success. In children ages
6 months to 3 years, mental development has been assessed with the Bayley Scales of
Infant Development. A large body of evidence also comprises associations of blood and
tooth Pb levels with memory and learning, executive function, language, and visuospatial
processing. These domains of cognitive function are related to intelligence, and several
are evaluated in the subtests of FSIQ. Further, some indices of memory, learning, and
executive function are more comparable to endpoints examined in tests with animals.
Fewer studies have examined academic performance and achievement; however, these
outcomes may provide information on the impact of Pb exposure on life success. In the
subsequent sections, the evidence for each of these categories of outcomes is reviewed
separately, to the extent possible, in order of strength of evidence as assessed by the
following parameters. Emphasis was placed on prospective epidemiologic studies with
repeated measurements of blood Pb levels and cognitive function and on studies that
examined blood Pb levels more similar to those of current U.S. children (i.e., <5 (ig/dL)
and children whose blood Pb levels were less influenced by higher past Pb exposures.
Studies of chelation in children generally were not included because the high pre-
chelation blood Pb levels may limit generalizability of results, and chelation itself has
been linked to neurodevelopmental effects. For animal studies, emphasis was placed on
studies examining Pb exposures considered relevant to this ISA, i.e., dietary exposures
producing blood Pb levels < 30 (ig/dL (Section 1.1).
Many factors have been shown to influence the cognitive function of children, including
SES, parental education, parental IQ, quality and stability of parental caregiving
environment (often measured as Home Observation for the Measurement of Environment
[HOMEITotsika and Sylva. 2004). nutritional status, and birth weight (Nation and
Gleaves. 2001; Wasserman and Factor-Litvak. 2001). These and other influences on
neurodevelopment often are correlated with blood Pb levels. Thus, due to their
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correlation with blood Pb level and causal association with outcome, these other risk
factors potentially may bias or confound the associations observed between blood Pb
level and indices of cognitive function. In the evaluation of the effects of Pb independent
from the effects of the other risk factors, greater weight was given to studies that more
extensively accounted for potential confounding in the study design or in statistical
analyses. A detailed evaluation of control for potential confounding in associations
between indicators of Pb exposure and neurodevelopmental effects is located in Section
4.3.13.
4.3.2.1 Full Scale IQ in Children
Evidence from Prospective Studies
Prospective cohort studies that were initiated in the 1980s addressed some limitations of
cross-sectional studies, including better characterizing the temporal sequence between
changes in blood Pb levels and cognitive function, examining the persistence of cognitive
function decrements to older ages, and comparing associations among blood Pb levels
measured at various lifestages or representing various time periods. Recruitment of
participants before or at birth without consideration of Pb exposure or maternal IQ, high
follow-up participation (>70%), and nonselective loss-to-follow-up in most studies
increase confidence that the observed associations are not due to selection bias.
Moreover, cooperation among investigators to adopt similar study protocols (e.g., similar
tests of IQ and consideration of similar potential confounding factors) strengthened
inferences regarding the consistency of associations with blood Pb level by facilitating
pooled analyses and by reducing sources of heterogeneity in evaluating results across
populations that varied in geographic location, proximity to Pb sources, blood Pb level
range, race/ethnicity, and SES.
Individual cohort studies of varying sample sizes (n = 112-537) conducted in several
different populations (e.g., Boston, MA; Cincinnati, OH; Rochester, NY; Cleveland, OH;
Mexico City, Mexico; Port Pirie, Australia; and Kosovo, Yugoslavia) were consistent in
demonstrating associations of higher blood Pb measured prenatally (maternal or
umbilical cord) or earlier in childhood or averaged over childhood with lower FSIQ
measured later in childhood, i.e., 4 to 17 years (Schnaas et al.. 2006; Ris et al.. 2004;
Canfield et al.. 2003a: Schnaas et al.. 2000; Factor-Litvak et al.. 1999; Tongetal. 1996;
Wasserman et al.. 1994; Dietrich et al.. 1993b: Baghurst et al.. 1992; Bellinger etal..
1992; Bellinger et al.. 1991; McMichael et al.. 1988) (Figure 4-2 and Table 4-3). Null or
weak associations were limited to a few cohorts, namely, the Cleveland and Sydney
cohorts (Greene et al.. 1992; Coonevetal.. 1991; 1989a: Ernhart et al.. 1988). In the
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prospective studies, lower FSIQ also was associated with higher concurrent blood Pb
levels (Figure 4-2 and Table 4-3) and tooth Pb levels. These latter results were based on
cross-sectional analyses; however, the pattern of associations consistently observed for
blood Pb levels measured at various lifestages or time periods does not strongly indicate
that reverse causation explains the FSIQ decrements observed in association with
concurrent blood Pb or tooth Pb levels.
Study Blood Pb Metric Blood Pb Mean (SD)
Analyzed (ug/dL)
Prospective Studies
Lanphearetal. (2005) Concurrent, peak < 7. 5 3.2
Canfield etal. (2003) Concurrent, peak< 10 3.3
Juskoetal. (2008)' Peak 11.4(7.3)
Bellingeretal. (1992) Age2 yr, peak< 10 3.8
Dietrich etal. (1993) Concurrent 11. 8 (6. 3) (Age 5)
Schnaas etal. (2006) Prenatal (maternal) 7. 8 (geometric)
Cooney etal.(1991)b Age3-5yravg 8.3(Age5yr)
Wasserman etal. (1997) Age 0 to 7yravg 16. 2 (geometric)
Tongetal. (1996) AgeOto 11-13yravg 14.0 (1.2) geometric
Kordasetal. (2011) Prenatal (cord) 6.6(3.3)
Concurrent 8.1 (3.6)
Greeneetal. (1992) Age2yr 15.6 (1.4) geometric
M in etal. (2009) Concurrent 7.0(4.1)
Age 4 yr
Age 4 yr
Cross-sectional Studies
Kim etal. (2009) Concurrent, low Mn 1.73(0.80)
Concurrent, high Mn
Fulton etal. (1987) Concurrent 11. 5 (range: 3. 3-34)
Royetal. (2011) Concurrent 11.5(5.3)
Blood Pb interval
examined3 (ug/dL)
1.3-6
0.5-8.4
2.1-10
5.5-10
Notre ported
6.0-10
11-19
3.2-11
10-24
3.0-10
0.9-2.8
5.6-10
FSIQ age (yr)
6 O
7 D
11-13 D
4.8 O
A •*•
6-9 »
-3.5 -3 -2.5 -2 -1.5 -1 -0.5 0
Change in FSIQ (95% Cl) per 1 ug/dL increase in various intervals
of blood Pb level
aSee Table 4-3 for explanation of the blood Pb level interval examined. Effect estimates were calculated for the lowest range
examined in the study or the 10th percentile of blood Pb level to a blood Pb level of 10 ug/dL.
bSufficient data were not available to calculate 95% Cl.
Note: Mn = manganese. Results are presented for most of the cohorts examined in the literature and generally are grouped
according to strength of study design, representativeness of the study population characteristics and blood Pb levels examined, and
extent of consideration for potential confounding. There is not necessarily a continuum of decreasing strength across studies.
Results usually are presented for the oldest age examined in cohorts. Multiple results from a cohort are grouped together. To
facilitate comparisons among effect estimates across studies with different distributions of blood Pb levels and model structures
(e.g., linear, log-linear), effect estimates are standardized to a 1 ug/dL increase for the lowest range of blood Pb levels examined in
the study or the interval from the 10th percentile of blood Pb level to 10 ug/dL. For populations with 10th percentiles near or above
10 ug/dL, the effect estimate was calculated for the 10th to 90th percentile of blood Pb level. The percentiles are estimated using
various methods and are only approximate values. Effect estimates are assumed to be linear within the blood Pb level interval
evaluated. The various tests used to measure FSIQ are scored on a similar scale (approximately 40-160 FSIQ points). Black
diamonds, blue circles, orange triangles, and gray squares represent effect estimates for concurrent, earlier childhood, prenatal, and
lifetime average blood Pb levels, respectively. The horizontal lines associated with point estimates represent 95% confidence
intervals (Cl).
Figure 4-2 Associations of blood Pb levels with full-scale IQ (FSIQ) in
children.
4-61
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Table 4-3 Additional characteristics and quantitative results for studies represented in Figure 4-2
Study
Prospective
Lanphear et
al. (2005)
Hornung
(2008)
Study Population and Methodological Details3
Studies
1,333 children pooled from Boston, Cincinnati, Cleveland, Mexico City, Port
Pirie, Rochester, and Yugoslavia cohorts, for each of which results are
described below.
Uniform analytic method applied to cohorts from diverse locations and
demographic characteristics. Blood Pb levels and FSIQ measured at different
ages. Several sensitivity analyses to examine heterogeneity of results by
cohort, model specification, and confounding. Log linear regression model
adjusted for HOME score, birth weight, maternal IQ, maternal education. Also
considered potential confounding by child sex, birth order, maternal age,
marital status, prenatal smoking status and alcohol use. FSIQ and covariates
measured with different instruments.
Blood Pb Metric
Analyzed (ug/dL)
Concurrent
Median: 9.7
Interval analyzed:
3.3 (5th percentile)-
10
FSIQ Testing
WISC-III,
WISC-R,
WPPSI,
\ A/I Of^ O
WISC-S
Ages
4.8-1 Oyr
Effect Estimate
(95% Cl)b
-0.82 (-1.1, -0.51)
103 children, subset with peak blood Pb levels <7.5 ug/dL
n = 13 Boston, 1 Cincinnati, 1 Cleveland, 8 Mexico City, 69 Rochester, 8
Yugoslavia, 0 Port Pirie
Linear regression model adjusted for HOME score, birth weight, maternal IQ,
maternal education.
Concurrent
Mean: 3.2
Interval analyzed
1.3-6.0 =
5th-95th percentiles
-2.9 (-5.2,-0.71)
Canfield et al.
(2003a)
Canfield
(2008)
Jusko et al.
(2008)
101 children born 1994-1995 followed from age 6 mo to 5 yr, Rochester, NY
Recruitment from study of dust control. 73% nonwhite. Moderate follow-up
participation but no selective attrition. Linear regression model adjusted for
child sex, Fe status, birth weight, maternal race, education, IQ, income, and
prenatal smoking status, HOME score.
174 children born 1994-1995 followed from age 6 mo to 6 yr, Rochester, NY
Same cohort as above. High follow-up participation. Higher follow-up of
nonwhite, higher peak blood Pb, lower maternal IQ. Generalized additive
model with LOESS adjusted for sex, birth weight, transferrin saturation,
maternal race, IQ, education, and prenatal smoking status, HOME score (6
yr), family income.
Concurrent, children
with peak<10
Mean: 3.3
Interval analyzed:
0.5-8.4 = range
Detection limit=1
Peak (6 mo-6 yr)
Mean (SD):
11 (7.3)
Interval analyzed:
2.1-10
Stanford- -1.8 (-3.0, -0.60)
Binet
Age 5 yr
WPPSI-R -1.2C
Age 6 yr
4-62
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Table 4-3 (Continued): Additional characteristics and quantitative results for studies represented in Figure 4-2
Study
Bellinger and
Needleman
(2003)
Bellinger
(2008)
U.S. EPA
(2006b)
Mazumdar et
al. (2011)
Dietrich et al.
(1993b)
Schnaas et al.
(2006)
Study Population and Methodological Details3
48 children followed from birth (1979-1981) to age 10 yr, Boston area, MA
Recruitment at birth hospital. Participation by 59% of original cohort but 88%
from age 5 yr. Participants had higher SES, HOME scores. 95% white. Linear
regression model adjusted for HOME score (age 10 and 5 yr), race, maternal
IQ, age, and marital status, SES, child sex, birth order, and stress events, #
residence changes before age 5 yr. Also considered potential confounding by
family stress, maternal age, psychiatric factors, child serum ferritin levels.
43 adults followed from birth (1979-1981) to age 28-30 yr, Boston, MA area
Same cohort as above. Small proportion of original cohort but no selective
attrition. 93% white. Linear regression model adjusted for maternal IQ. Also
found associations adjusted for maternal marital status at birth, education at
birth, prenatal smoking status, or alcohol use, HOME score (mean across
ages), subject sex, birth weight, birth order, gestational age, race, concussion
history, or current smoking status. Also considered potential confounding by
subject alcohol use.
231 children followed from birth (1979-1985) to age 6.5 yr, Cincinnati, OH
Recruitment at prenatal clinic. High follow-up participation. Participants had
slightly higher age 1 yr blood Pb levels. Primarily African-American. Linear
regression model adjusted for HOME score, maternal IQ and prenatal
smoking, child birth weight, birth length, and sex. Also considered potential
confounding by perinatal complications, prenatal maternal substance abuse,
and nutritional status.
150 children followed from birth (1987-1992) to age 10 yr, Mexico City,
Mexico.
Blood Pb Metric
Analyzed (ug/dL) FSIQ Testing
Early childhood (age WISC-R
2 yr), children with AaelOvr
peak<10
Mean: 3.8
Interval analyzed:
1-9.3 = range
Detection limit not
reported
Earlier childhood WASI
avg (age 6 mo- Aqe 28-30 yr
10yr):NR
Mean (SD), [age]
8.0 (5.3) [6 mo]
10(6.7)[1yr]
7.7 (4.0) [2 yr]
6.7 (3.6) [4 yr]
3.0(2.7)[10yr]
Concurrent: NR WISC-R
Age 4-5 yr avg Age 6. Syr
Mean (SD):
11.8(6.3)
Interval analyzed 5.5
(10th percentile)-10
Prenatal (maternal WISC-R
28-36 weeks) Ages 6-1 Oyr
Effect Estimate
(95% Cl)b
-1.6 (-2. 9, -0.2)
-1.1 (-2.3, 0.06)d
-0.33 (-0.60, -0.06)
-0.67 (-1.1, -0.28)
Recruitment at prenatal clinic. Low follow-up participation. Participants had
higher SES, FSIQ, higher blood Pb level before age 5 yr, lower at older ages.
Log linear mixed effects model adjusted for SES, maternal IQ, child sex, birth
weight, postnatal blood Pb, age of 1st FSIQ test, random slope for subject.
Also considered potential confounding by HOME score. Covariates assessed
in pregnancy or within age 6 mo.
Geometric mean
(5th-95th): 7.8
(2.5-25)
Interval analyzed 3.2
(10th percentile)-10
4-63
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Table 4-3 (Continued): Additional characteristics and quantitative results for studies represented in Figure 4-2
Study
Cooney et al.
(1991)
Wasserman
et al. (1997)
long et al.
(1996)
Kordas et al.
(2011)
Study Population and Methodological Details3
175 children followed from birth (1983) to age 7 yr, Sydney, Australia
Recruitment at birth hospital. Moderate follow-up participation but no selective
attrition. 100% white. Linear regression adjusted for maternal education and
IQ, paternal education and occupation, HOME score, child gestational age.
258 children followed prenatally (1984-1985) to age 7 yr, K. Mitrovica, Pristina,
Yugoslavia
50% subjects live near Pb sources. Moderate follow-up participation.
Participants had lower HOME score, maternal IQ, lower early childhood blood
Pb levels and lived in town without Pb sources. Log linear regression adjusted
for maternal age, education, and IQ, child age, sex, sibship size, and birth
weight, language spoken in home, HOME score.
375 children followed from birth (1979-1982) to age 11-13 yr, Port Pirie,
Australia
Residence near Pb smelter. Moderate follow-up participation. Participants had
higher parental occupational prestige. Regression model adjusted for maternal
IQ and age, parental occupational prestige, smoking, marital status, and
education, HOME score, family functioning score, family size, child sex, age,
school grade, birth weight, birth order, feeding method, breastfeeding duration,
life events, prolonged absences from school. Also considered potential
confounding by maternal psychopathology, child Fe status, medication use in
previous 2 weeks, length of residence in area.
186 children followed prenatally (1994-1995) to age 4 yr, Mexico City, Mexico
Recruitment at prenatal clinic. Low follow-up participation but no selective
attrition. Linear regression model adjusted for maternal age, education, IQ,
smoking status, and marital status, crowding in home, type of floor in home,
child sex, birth weight, gestational age. Did not consider potential confounding
by parental caregiving quality.
Blood Pb Metric
Analyzed (ug/dL) FSIQ Testing
Age 3-5 yr avg: NR WISC-R
Age 5 yr Age 7 yr
Mean (Max):
8.3 (27)
Lifetime avg WISC-III
(to age 7 yr) Age 7 yr
Geometric mean: 16
Interval analyzed 6.0
(10th percentile)-10
Lifetime avg WISC-R
(to age 11-1 Syr) Age 1 1-13 yr
Geometric mean
(GSD): 14(1.2)
Interval analyzed
11-19 =
10th-90th
percentiles
Prenatal (cord) McCarthy
Mean (SD): GCI
6.6(3.3) Age4yr
Interval analyzed
3.2-11 = 10th-90th
percentiles
Concurrent Mean
(SD):
8.1 (3.6)
Interval analyzed 4.3
(10th percentile)-10
Effect Estimate
(95% Cl)b
-0.07 (p > 0.05)c
-0.48 (-0.68, -0.27)
-0.30 (-0.60, -0.01)
-0.20 (-0.79, 0.39)
-0.60 (-0.99, -0.21)
4-64
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Table 4-3 (Continued): Additional characteristics and quantitative results for studies represented in Figure 4-2
Study
Greene et al.
(1992)
Min et al.
(2009)
Study Population and Methodological Details3
270 children followed from birth (yr not reported) to age 4 yr 10 mo, Cleveland,
OH
Recruitment at birth hospital. High prevalence of prenatal alcohol exposure.
High follow-up participation. Participants tended to be Black, prenatally
exposed to marijuana. Log linear regression adjusted for maternal IQ, weight,
street drug use, cigarettes/day, alcohol use, and age, parental education,
authoritarian scale, race, parity, gestation duration, date of first prenatal visit,
HOME score
267 children followed from birth (1994-1996) to age 11 yr, Cleveland, OH
Recruitment at birth hospital. 86% African-American with high prevalence of
prenatal drug and alcohol exposure. Moderate follow-up participation to age 4
yr, high retention to age 1 1 yr. Participants tended to be African-American and
had married mothers. Linear regression model adjusted for HOME score, head
circumference at birth (all ages), current caregiver vocabulary score, maternal
marital status, parity, child sex (age 4 yr), maternal vocabulary score at birth
(age 9 and 11 yr), race, average prenatal cocaine use (age 9 yr), prenatal 1st
trimester marijuana use (age 11 yr). Also considered potential confounding by
maternal education, Fe deficiency, maternal psychological distress, maternal
prenatal alcohol use.
Blood Pb Metric
Analyzed (ug/dL)
Early childhood (age
2yr)
Geometric mean
(GSD):
15.6(1.4)
Interval analyzed
10-24 =
10th-90th
percentiles
Age 4 yr
Mean (SD):
7.0(4.1)
Interval analyzed 3.0
(10th percentile)-10
FSIQ Testing
WPPSI
Age 4.8 yr
WPPSI,
Age 4 yr
concurrent
WISC-IV,
Aqe 9 yr
\\2 « ^f j
WISC-IV,
Age 1 1 yr
Effect Estimate
(95% Cl)b
-0.11 (-0.51, 0.29)
-0.50 (-0.89, -0.11)
-0.41 (-0.78, -0.04)
-0.54 (-0.91, -0.17)
4-65
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Table 4-3 (Continued): Additional characteristics and quantitative results for studies represented in Figure 4-2
Study
Study Population and Methodological Details3
Blood Pb Metric
Analyzed (ug/dL)
Effect Estimate
FSIQ Testing (95% Cl)b
Cross-sectional Studies
Fulton et al.
(1987)
Surkan et al.
(2QQ7)
Kim et al.
(2009b)
Roy et al.
(2011)
501 children, ages 6-9 yr, Edinburgh, Scotland
Recruitment at schools. High participation rate, representative of area
population. Log linear regression model adjusted for parental SES, education,
marital status, health, mental health, cigarettes smoked, vocabulary and
matrices test scores, involvement, interest, communication, and participation
with child, family size and structure, child age, sex, handedness, height,
gestation length, birth weight, medical history, # school absences, recent
school change, grade, birth order, time of day of test, people per room in
home, car/phone ownership, consumer goods ownership.
389 children, ages 6-10 years, Boston, MA, Farmington, ME
Recruitment from trial of amalgam fillings. High participation rate. Higher
participation of white children in Maine. Analysis of covariance adjusted for
caregiver IQ, child age, SES, race, birth weight. Also considered potential
confounding by site, sex, birth order, caregiver education and marital status,
parenting stress, and maternal utilization of prenatal and annual health care
but not parental caregiving quality.
261 children (born 1996-1999) ages 8-11 yr, Seoul, Seongnam, Ulsan, and
Yeoncheon, Korea
Recruitment at schools. Moderate participation rate. Log linear regression
model adjusted for maternal age, education and prenatal smoking status,
paternal education, yearly income, smoking exposure status after birth, child
age, sex, and birth weight. Did not consider potential confounding by parental
caregiving quality or IQ.
717 children ages 3-7 yr, Chennai, India
Recruitment at schools. High participation rate. Log linear model adjusted for
mid-arm circumference, age, sex, family income, parental education, family
size, school, classroom. Did not consider potential confounding by parental
caregiving quality.
Concurrent
Geometric mean
(range):
12(3.3-34)
Interval analyzed 5.6
(mean of 1st
decile)-10
Concurrent
Groupl: 1-2
Group 2: 3-4
Group 3: 5-10
Mean (SD):
2.2(1.6)
Concurrent
Mean (SD):
1.7(0.80)
Interval analyzed
0.9-2.8 = 10th-90th
percentiles
Concurrent
Mean (SD):
12(5.3)
Interval analyzed 5.9
(10th percentile)-10
BASC
Age 6-9 yr
WISC-III
Age 6-1 0 yr
KEDI-WISC
Ages 8-1 1 yr
Blood Mn
<1.4ug/dL
f-iy «•—
Blood Mn
>1.4ug/dL
Binet-Kamat
Ages 3-7 yr
-0.22 (-0.37, -0.06)
Reference
-0.12 (-3.3, 3.1)e
-6.0 (-11, -1.4)
-2.4 (-6.0, 1.1)
-3.2 (-6.1, -0.24)
-0.54 (-0.91, -0.17)
4-66
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Table 4-3 (Continued): Additional characteristics and quantitative results for studies represented in Figure 4-2
Study
Chiodo et al.
(2QQ7)
Chiodo et al.
(2004)
Study Population and Methodological Details3
495 children (born 1989-1991) age 7 yr, Detroit, Ml area
Recruitment at prenatal clinic. 100% African-American. High prevalence of
prenatal drug exposure. High participation rate. Linear regression model
adjusted for maternal concurrent psychopathology, IQ, prenatal cigarettes/day,
and prenatal use of marijuana, SES, HOME score, caretaker education and
marital status, # children in home, child sex. Also considered potential
confounding by child age, maternal age, custody, cocaine use, prenatal
alcohol use, and concurrent alcohol/week, cigarettes/day, and marijuana use.
237 children, age 7.5 yr, Detroit, Ml area
Recruitment at prenatal clinic. 100% African-American. High prevalence of
prenatal alcohol and drug exposure. Moderate participation rate. Log linear
regression model adjusted for SES, maternal education and vocabulary score,
# children <18 yr, HOME score, parity, family environment scale, child sex.
Also considered potential confounding by prenatal alcohol, marijuana,
smoking, or cocaine use, crowding, child age and life stress, caregiver life
stress, conflict tactics.
Blood Pb Metric
Analyzed (ug/dL)
Concurrent
Mean (SD):
5.0(3.0)
Interval analyzed
2.1-8.7= 10th-90th
percentiles
Concurrent
Mean (SD):
5.4(3.3)
Interval analyzed
2.2-9.5= 10th-90th
percentiles
Effect Estimate
FSIQ Testing (95% Cl)b
WPPSI -0.19 (-0.30, -0.08)e'f
Age 7 yr Standardized
regression coefficient
WISC-III -0.49 (-0.87, -0.12)e'f
Age 7.5 yr standardized
regression coefficient
WISC = Wechsler Intelligence Scale for Children, WPPSI = Wechsler Preschool and Primary Scale of Intelligence, WASI = Wechsler Abbreviated Scale of Intelligence, GCI = General
Cognitive Index, BASC = British Ability Scales Combined, KEDI = Korean Educational Development Institute.
aResults are presented for most of the cohorts examined in the literature and generally are grouped according to strength of study design, representativeness of the study population
characteristics and blood Pb levels examined, and extent of consideration for potential confounding. There is not necessarily a continuum of decreasing strength across studies.
bEffect estimates are standardized to a 1 ug/dL increase in blood Pb level within the lowest range examined in the study or 10th percentile to 10 ug/dL. For populations with 10th
percentiles above 10 ug/dL, effect estimates were calculated for the 10th-90th percentile interval of blood Pb level. Effect estimates are assumed to be linear within the evaluated
interval of blood Pb level. The percentiles are estimated using various methods and are only approximate values.
""Sufficient data were not available to calculate 95% Cl.
dResults not included in Figure 4-2 because FSIQ assessed in adults.
eResults not included in Figure 4-2 because blood Pb level analyzed as categorical variable or because standardized regression coefficient reported.
f95% CIs were constructed using a standard error that was estimated for the reported p-value of 0.01.
4-67
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In addition to better characterizing the temporal sequence between Pb exposure and
decrements in FSIQ, a common strength of most prospective studies was the adjustment
for several of the potential confounding factors noted above, including maternal IQ and
education, child sex and birth weight, SES, and parental caregiving quality (i.e., HOME
score) (Table 4-3). Although not considered as frequently, some studies also adjusted for
potential confounding by parental smoking, birth order, and nutritional factors. Multiple
testing of associations between blood Pb levels and FSIQ was common in prospective
studies that found and did not find associations between blood Pb level and FSIQ.
However, higher probability of finding associations due to chance alone does not appear
to unduly influence the evidence because in studies that found associations, there was a
consistent pattern of blood Pb-associated cognitive function decrements across the
various ages of blood Pb level and/or cognitive tests evaluated. Studies finding null or
weak associations also tended to show a consistent pattern across the various analyses
conducted.
Across cohort studies, blood Pb-associated FSIQ decrements were found in populations
with mean blood Pb levels of 5 to 10 (ig/dL. In analyses of the Boston cohort, Rochester
cohort, and data pooled across seven cohorts restricted to children in the lower range of
the blood Pb distribution (e.g., peak <10 (ig/dL), associations were observed in groups of
children with mean blood Pb levels of 3 to 4 (ig/dL (Bellinger. 2008; Canfield. 2008;
Hornung. 2008). Key evidence was provided by analyses of the Boston and Rochester
cohorts because they had lower blood Pb levels than other cohorts examined. Further, the
Rochester cohort study had extensive consideration for potential confounding by factors
such as sex, family income, maternal education, race, prenatal maternal smoking, birth
weight, maternal IQ, HOME score, and transferrin saturation which indicates Fe status
and compared unadjusted and covariate-adjusted results (Canfield et al., 2003a). At age 5
years, higher age 6-24 month average, peak, concurrent, and lifetime average blood Pb
levels (area under the curve calculated from repeat measures between age 6 months and
5 years) were associated with lower FSIQ, and while effect estimates in the covariate-
adjusted model were 40-45% smaller than estimated in the unadjusted models, they
remained statistically significant. While these results do not rule out potential residual
confounding, they support an independent effect of Pb exposure. A larger effect was
estimated for the 101 (59%) children with peak blood Pb levels < 10 (ig/dL, i.e., -1.8
FSIQ points (95% CI: -3.0, -0.60) per 1 (ig/dL in concurrent blood Pb level (U.S. EPA.
2006b). A larger effect also was estimated in the subset of the Boston cohort (n = 48,
32%) with peak blood Pb levels <10 (ig/dL, i.e., -1.6 FSIQ points at age 10 years (95%
CI: -2.9, -0.2) per 1 (ig/dL increase in age 2-year blood Pb level. The mean blood Pb
levels in these subsets of children were 3.3 (Rochester) and 3.8 (ig/dL (Boston)
(Bellinger. 2008; Canfield. 2008) closer to that of current U.S. children compared with
other prospective studies.
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Analyses of the Cincinnati and Port Pirie, Australia cohorts also indicated associations
between blood Pb level and FSIQ decrements with as extensive consideration for
potential confounding (Table 4-3) albeit in populations with higher blood Pb levels
(i.e., age 51-60 month mean: 11.8 (ig/dL, lifetime [to age 11-13 years] average geometric
mean: 14.0 (ig/dL) (Tong et al.. 1996; Dietrich et al.. 1993b). In contrast with other
studies, in the Cleveland cohort, associations of blood Pb level (ages 2 and 3 years) and
tooth Pb level with FSIQ (ages 3 and 4.8 years), were attenuated or were too imprecise to
be informative with adjustment for a large number of potential confounding factors,
including sociodemographic characteristics and maternal substance abuse (Greene et al..
1992; Ernhart et al., 1988). In analyses of the Cleveland cohort, HOME score accounted
for a large proportion of the variance in FSIQ and was the covariate that caused the
greatest attenuation of the effect estimates for Pb biomarkers. The association between
tooth Pb level and FSIQ at age 4.8 years was attenuated with additional adjustment for
HOME score but was estimated with similar precision (-3.0 points [95% CI: -6.4, 0.32]
per 1 (ig/g increase in tooth Pb level) (Greene and Ernhart. 1993). The few weak or null
associations do not weaken the otherwise strong evidence provided by other studies. A
far larger number of other prospective studies with similar population mean blood Pb
levels and important potential confounding factors considered (Table 4-3) found
associations, and the Cleveland cohort had high prevalence of maternal prenatal
substance abuse which may limit the external validity or representativeness of results.
Further, the blood Pb-FSIQ association in children was supported by a pooled analysis of
seven prospective studies (Lanphear et al., 2005), which included the Cleveland cohort,
as well as multiple meta-analyses that combined results across various prospective and
cross-sectional studies, including those from the Cleveland and Sydney cohorts (Pocock
et al.. 1994; Schwartz. 1994; Needleman and Gatsonis. 1990). The meta-analysis by
Schwartz (1994) demonstrated the robustness of evidence to potential publication bias.
The addition of eight hypothetical studies with a zero effect and with the average weight
of the eight published studies resulted in a 50% lower but still negative and precise
(p <0.001) blood Pb-FSIQ effect estimate.
The pooled analysis of seven prospective studies (Boston, MA; Cincinnati, OH;
Rochester, NY; Cleveland, OH; Mexico City, Mexico; Port Pirie, Australia; and Kosovo,
Yugoslavia, individual results reported in Figure 4-2 and Table 4-3) included individual-
level data from 1,333 children ages 4.8-10 years with a median concurrent blood Pb level
of 9.7 (ig/dL (5th-95th percentile: 2.5-33.2 (ig/dL) (Lanphear et al.. 2005). The pooled
analysis demonstrated a blood Pb-associated decrement in FSIQ in a diverse population
with application of a uniform statistical model across cohorts. In multivariate models that
adjusted for study site, maternal IQ, HOME score, birth weight, and maternal education,
higher concurrent, peak, lifetime average, and early childhood (age 6-24 months) blood
Pb levels were associated with lower FSIQ measured at age 4.8-10 years. Importantly, in
4-69
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this pooled analysis and in the meta-analyses, heterogeneity among studies in the tests
used to assess FSIQ and the ages of blood Pb level and FSIQ examined is a source of
nondifferential measurement error that could attenuate the association to null.
Pooling data across cohorts improved characterization of the shape of the blood Pb-FSIQ
concentration-response relationship at relatively low blood Pb levels. Most of the low
blood Pb level data were provided by the Boston and Rochester cohorts. Consistent with
the supralinear concentration-response relationship found in several individual cohort
studies, a nonlinear (i.e., log-linear) model was found to fit the pooled data better than a
linear model (Lanphear et al.. 2005; Rothenberg and Rothenberg. 2005). The nonlinear
relationship was indicated further by observations of a greater decrease in FSIQ for a
1 (ig/dL increase in concurrent blood Pb among the 244 (18%) children who had peak
blood Pb levels <10 (ig/dL (-0.80 points [95% CI: -1.7, -0.14]) and the 103 (8%) children
with peak blood Pb levels <7.5 (ig/dL (-2.9 points [95% CI: -5.2, -0.71]). In comparison,
a log-linear analysis that included most of the study population estimated a decrease of
6.9 (95% CI: 4.2, 9.4) FSIQ points for an increase in concurrent blood Pb level from 2.4
to 30 (ig/dL, which suggests a smaller FSIQ decrement per 1 (ig/dL increase in blood Pb
level than results from the piecewise linear models. Among children with peak blood Pb
levels <10 (ig/dL and <7.5 (ig/dL, the mean concurrent blood Pb levels were 4.3 (ig/dL
and 3.2 (ig/dL, respectively (Hornung. 2008).
Similar to individual cohort studies, Lanphear et al. (2005) examined several potential
confounding factors related to SES and parental caregiving. HOME score, birth weight,
maternal IQ, and maternal education were selected for inclusion in the final model
because they were associated (p<0.10) with FSIQ in a model that already contained blood
Pb, or they changed the blood Pb effect estimate by more than 10%. A smaller concurrent
blood Pb-associated decrement in FSIQ was estimated in the adjusted model than in the
unadjusted model (-2.7 points [95% CI: -3.7, -1.7] versus -4.7 points [95% CI: -5.7, -3.6],
respectively, per log increase in concurrent blood Pb level). However, the effect estimate
for blood Pb level remained statistically significant. Other potential confounding factors
such as child sex, maternal prenatal tobacco or alcohol use, maternal age at delivery,
marital status, and birth order were not strongly associated with FSIQ (p>0.10) and did
not alter the effect estimate for concurrent blood Pb level adjusted for the four
aforementioned covariates. Although there may be some residual confounding because
there was some variation among individual studies in how they measured covariates,
these results support an independent effect of Pb.
The few prospective studies published since the 2006 Pb AQCD continued to
demonstrate associations between higher blood Pb level and lower FSIQ, in some cases,
with additional follow-up of previous cohorts. Similar to studies reviewed in the
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2006 Pb AQCD, most recent prospective studies demonstrated associations between
blood Pb level and lower FSIQ in populations or smaller subgroups with mean blood Pb
level 5 to 10 (ig/dL. Jusko et al. (2008) examined 174 Rochester cohort subjects at age
6 years and similar to the overall results of Canfield et al. (2003a) for the cohort at age
5 years, found that the decrease in FSIQ per unit increase in peak blood Pb level was
larger among children with peak blood Pb levels 2.1-10 (ig/dL than among children with
peak blood Pb levels 10-20 (ig/dL (-1.2 points versus -0.32 points per 1 (ig/dL increase in
blood). The age 6 year analysis had similarly extensive consideration for potential
confounding as did Canfield et al. (2003a) (Table 4-3) and also indicated associations
with higher concurrent, infancy average, and lifetime average blood Pb levels (effect
estimates not reported).
Additional evidence was provided for children in Mexico City, albeit in a separate cohort
of children born later with lower blood Pb levels at corresponding ages. Among 150
children born 1987-1992, Schnaas et al. (2006) previously reported larger Pb-associated
decrements in FSIQ for prenatal maternal (28-36 weeks) blood Pb levels than for child
concurrent blood Pb levels between ages 1 and 10 years. In contrast, Kordas et al. (2011)
found a larger decrement in FSIQ at age 4 years per unit increase in blood Pb level for
concurrent blood Pb than cord blood Pb with adjustment for several potential
confounding factors (HOME score not examined) (Table 4-3). The 186 children in the
latter study were born 1994-1995 and at age 4 years had a mean blood Pb level of
8.1 (ig/dL. In Schnaas et al. (2006). the geometric mean blood Pb level at age 4 years was
10.3 (ig/dL. Other than mean age 4 blood Pb levels, the two Mexico City cohorts also
differed in the prenatal blood Pb metric examined and age of FSIQ assessment and could
have differed in the temporal patterns of Pb exposure.
Mazumdar et al. (2011) followed the Boston cohort (Bellinger et al.. 1992) to age
28-30 years, and indicated that the effect of childhood Pb exposures may persist to
adulthood. Only 43 of the original 249 subjects enrolled at birth were examined at age
28-30 years, but they did not differ from the original cohort in childhood blood Pb levels
or demographic characteristics. Higher blood Pb levels measured at age 6 months,
4 years, 10 years, and averaged to age 10 years (means: 3 (ig/dL at age 10 years to
10 (ig/dL at age 1 year) were associated with lower FSIQ in adults with adjustment for
one other factor: sex, birth weight, birth order, gestational age, maternal marital status,
maternal education, maternal IQ, race, maternal prenatal smoking or alcohol use,
childhood average HOME score, subject concussion, or subject current smoking status.
The effect estimates were similar in magnitude for all childhood blood Pb metrics, except
for age 6 month blood Pb level, which was associated with a smaller FSIQ decrement.
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Min et al. (2009) found higher earlier childhood blood Pb levels (age 4 year) to be
associated with decrements in FSIQ in another cohort of children in Cleveland, OH at
ages 4, 9, and 11 years, indicating persistence of effects (Figure 4-2 and Table 4-3 ). But,
similar to the previous Cleveland cohort (Greene et al.. 1992). the recent cohort had high
prevalence of prenatal alcohol and drug exposure. These exposures were weakly
associated with FSIQ or did not influence the blood Pb-FSIQ association, indicating lack
of strong confounding bias. However, the results may be less generalizable.
An important consideration in the evaluation of epidemiologic evidence is the precision
of effect estimates, both within and among studies. There was variability in precision
among studies, which did not appear to be influenced by age of subjects or the extent of
adjustment for potential confounding factors. Analyses of the Port Pirie (n = 375, ages
11-13 years) (Tong etal.. 1996) and Yugoslavia (n = 258, age 7 years) (Wasserman et al..
1997) cohorts and the pooled analysis of 1,333 children (Lanphear et al.. 2005) estimated
more precise effects compared to those of the Boston (n = 148, age 10 years) and
Rochester cohorts (n = 172, age 5 years) (Canfield et al.. 2003a; Bellinger et al.. 1992).
Analyses of the Yugoslavia and Port Pirie cohorts did not necessarily have more or less
extensive adjustment for potential confounding, but the cohorts did include children
living in towns with Pb smelters.
Among prospective studies, a wide range of blood Pb-associated FSIQ decrements was
estimated (Figure 4-2and Table 4-3). This wide range is not unexpected, given
differences among studies in blood Pb level ranges, model specification (linear versus log
linear), lifestage or time period of blood Pb examined, and distribution of potential
confounding factors. The pooled analysis examined study populations of diverse SES,
maternal education, and cultural backgrounds with the same model and indicated
precision of effect (Lanphear et al.. 2005). A series of sensitivity analyses, in which one
cohort was excluded at a time, revealed that no single study was responsible for the
results. Per log increase in blood Pb level, effect estimates excluding one study at a time
fell within a narrow range, -2.4 to -2.9 (blood Pb level ranges for sensitivity analyses not
reported). Precision of effect also was indicated by the similar effect estimates found with
similar model specifications, population blood Pb levels, and sample sizes in the Boston
and Rochester cohorts (Table 4-3). but very different SES and racial distributions of the
cohorts, and different ages of blood Pb and FSIQ examined (Canfield et al.. 2003a;
Bellinger et al.. 1992). These estimates were larger than those found in the Cincinnati,
Port Pirie, and Yugoslavia cohorts but were based on similar extent of adjustment for
potential confounding factors. Several of the smaller blood Pb-associated FSIQ
decrements were based on log-linear models that estimated effects for higher blood Pb
levels (e.g., estimates of 10th percentiles 5.5-12.7 versus 1.4-2.1 (ig/dL). The much
different null associations found in the Cleveland cohort have weaker implications
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because the high prevalence of prenatal alcohol and drug exposure in the cohort may
limit the representativeness of findings (Greene et al.. 1992; Ernhart et al.. 1988).
Evidence from Cross-sectional Studies
The smaller body of cross-sectional studies reviewed in the 2006 Pb AQCD (U.S. EPA.
2006b) found associations of higher concurrent blood (Fulton et al.. 1987) or tooth
(Needleman et al.. 1979) Pb levels with FSIQ decrements in children ages 6-9 years, and
associations also were found in the few recent studies in children ages 3-11 years (Figure
4-2 and Table 4-3). Several cross-sectional studies had larger sample sizes (n = 237-717)
than the prospective studies and produced effect estimates with similar precision. Some
studies had population-based recruitment, high participation rates, and did not indicate
undue selection bias. Previous meta-analyses produced similar combined blood Pb-FSIQ
effect estimates for prospective and cross-sectional studies (Pocock et al.. 1994;
Schwartz. 1994). However, in this ISA, the cross-sectional findings were given less
weight in conclusions regarding Pb-associated effects on cognitive function. The
temporal sequence between Pb exposure and decreases in FSIQ is more difficult to
establish. Among the key potential confounding factors noted at the beginning of Section
4.3.2. cross-sectional studies examined potential confounding by parental education and
SES, parental IQ less so, and parental caregiving quality only in a few studies (Roy et al..
2011; Kim et al.. 2009b: Zailina et al.. 2008; Surkan et al.. 2007; Needleman et al.. 1979).
The meta-analysis by Pocock et al. (1994) noted the lack of adequate control for potential
confounding factors in many of the cross-sectional studies evaluated.
Among the cross-sectional studies, Fulton et al. (1987) had more extensive consideration
O ? v /
for potential confounding. Among 501 children, ages 6-9 years, in Edinburgh, Scotland, a
1 (ig/dL increase in concurrent blood Pb level in the interval between 5.6 and 10 (ig/dL
was associated with a 0.22-point decrease (95% CI: -0.37, -0.06) in FSIQ, after
adjustment for several factors related to SES, parental health and mental health, child
health, and parental caregiving quality (Table 4-3). The effect estimate from this study
was among the smallest produced by cross-sectional studies. The study population was
representative of the source population but had much higher blood Pb levels (geometric
mean: 11.5 (ig/dL) than most current U.S. children.
Most studies that examined populations with mean concurrent blood Pb levels 1.7-
5.4 (ig/dL found associations with FSIQ decrements at lower blood Pb levels. In
exception, among 389 children from urban Boston, MA and rural Farmington, ME with
mean concurrent blood Pb level 2.2 (ig/dL, lower FSIQ was limited to children with
concurrent blood Pb levels 5-10 (ig/dL, with children with blood Pb levels 1-2 (ig/dL
serving as the referent group (Table 4-3) (Surkan et al.. 2007). There was consideration
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for potential confounding by several factors, including age, race, birth weight, SES,
primary caregiver IQ, SES, education and marital status, parenting stress, and maternal
utilization of prenatal or annual health care. Other cross-sectional studies found
associations with relatively low concurrent blood Pb levels; however, higher earlier
childhood Pb exposures could have contributed to the findings. Studies in children age 7
years in Detroit, MI (mean blood Pb level ~5 (ig/dL) found associations with adjustment
for several potential confounding factors (Chiodo et al., 2007; 2004). Although not
necessarily affecting the validity of results, the high prevalence of prenatal alcohol or
drug exposure in these children may limit their generalizability. Among 261 children
ages 8-11 years from four Korean cities with a mean concurrent blood Pb level 1.7 (ig/dL
(detection limit=0.06 (ig/dL), Kim et al. (2009b) found an association between higher
concurrent blood Pb level and lower FSIQ with adjustment for parental education, yearly
income, prenatal and postnatal smoking exposure, birth weight, age, sex, and blood
manganese (Mn) level. The adjusted effect estimate was attenuated but similarly precise
as the unadjusted estimate. The relationship of concurrent blood Pb with FSIQ and verbal
IQ was modified by concurrent blood Mn levels. Blood Pb and Mn levels were not
correlated (r = 0.03, p = 0.64). Higher concurrent blood Pb level was associated with
lower FSIQ in both children in high Mn (above the median of 1.4 (ig/dL) and low Mn
group (below the median of 1.4 (ig/dL) but a larger FSIQ decrement in the 130 children
in the high Mn group (-3.2 points [95% CI: -6.1, -0.23] per 1 (ig/dL increase in the 10th
to 90th percentile interval 0.9-2.8 (ig/dL) compared with the 131 children in the low Mn
group (-2.4 points [95% CI: -6.0, 1.1]). The biological plausibility for the Pb-Mn
interaction is provided by observations that Mn has similar modes of action and cellular
targets as does Pb, i.e., altering Ca2+ metabolism, inducing oxidative damage in neuronal
cells, diminishing dopamine transmission. Among 169 children in Malaysia, ages 6-8
years, concurrent blood Pb (mean ~4 (ig/dL) level was the covariate most strongly
associated with FSIQ decrements (Zailinaetal.. 2008). Only F-statistics were reported.
Family income and parental education were only weakly associated with FSIQ. The
authors reported parental education to be similar between the urban and industrial study
areas and at the lower secondary level. However, the variability in parental education or
family income among subjects was not reported. Thus, it is not clear whether the weak
association of FSIQ with parental education or family income was influenced by the lack
of variability in these factors or whether they were measured inadequately. If measured
inadequately, they could confound blood Pb associations.
Other recent cross-sectional studies found associations in populations of children with
relatively higher concurrent blood Pb levels (means 8.1, 11.5 (ig/dL). However, they did
not consistently find evidence for effect modification by variants in dopamine receptor
genes. Such subgroup analyses are subject to higher probability of finding an association
by chance; however, the modification of Pb-associated effects on cognition by variants in
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dopamine-related genes is biologically plausible. Evidence demonstrates that Pb affects
dopaminergic neurons and dopamine release (Section 4.3.10.8). Further, dopaminergic
activity is a key mediator of cognitive function. The larger of the studies (N = 717
children ages 3-7 years in Chennai, India) with higher concurrent blood Pb levels (mean:
11.5 (ig/dL) found that a 1 (ig/dL higher blood Pb level was associated with a larger
decrease in FSIQ among the 73 children with the Taq A1/A1 dopamine receptor (DRD2)
genotype (-1.2 points [95% CI: -2.3, -0.02] within the blood Pb level interval 6-10 (ig/dL)
than among the 644 children carrying the Taq A2 variant (-0.51 points [-0.92, -0.09])
(RoyetaL 2011). Kordas et al. (2011) did not find effect modification in a smaller study
of 186 children age 4 years in Mexico City with a mean concurrent blood Pb level of
8.1 (ig/dL. Another difference between studies that may have contributed to the
difference in effect modification by the DRD2 variant was a higher mean FSIQ score
observed with Taq A1/A1 genotype in the group in Mexico but no association found in
the group in India.
Summary of Evidence for FSIQ
A large majority of prospective and cross-sectional studies demonstrated associations
between higher blood Pb level and lower FSIQ in children ages 3-17 years (e.g., Figure
4-2 and Table 4-3). While studies performed numerous tests, bias due to increased
probability of finding associations by chance was unlikely because most studies found a
consistent pattern of association across the ages of blood Pb level and FSIQ analyzed.
Across studies, FSIQ was measured with various instruments scored on similar scales
with similar measurement error. The key supporting evidence is provided by the
prospective studies, which better indicated the temporal sequence between Pb exposure
and FSIQ with analysis of blood Pb levels measured earlier in childhood or averaged over
multiple years or tooth Pb levels and FSIQ measured later in childhood. Prospective
studies also had more extensive consideration for potential confounding by maternal IQ
and education, SES, birth weight, smoking exposure, parental caregiving quality, and in a
few cases, other birth outcomes and nutritional factors. Further, the representativeness of
findings is supported by associations found in diverse populations (e.g., Boston, MA;
Cincinnati, OH; Rochester, NY; Cleveland, OH; Mexico City, Mexico; Port Pirie,
Australia; and Kosovo, Yugoslavia) and studies examining populations recruited from
prenatal clinics, hospital maternity departments, or schools with high follow-up
participation and lack of biased follow-up participation by blood Pb level and FSIQ. A
few studies found weak or null associations (i.e., Cleveland, Sydney cohorts) but did not
have major strengths over other studies with respect to methodology or control for
potential confounding and do not weaken the far larger body of supporting evidence.
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The blood Pb-FSIQ association in children also was supported by a pooled analysis of
seven prospective cohorts (Lanphear et al.. 2005) as well as multiple meta-analyses that
combined results across various prospective and cross-sectional studies (Pocock et al.,
1994: Schwartz. 1994: Needleman and Gatsonis. 1990V Schwartz (1994) additionally
demonstrated the robustness of evidence to potential publication bias. While supporting
the robustness of the blood Pb-FSIQ association, these analyses that combine data across
different populations are subject to measurement error. Nondifferential error could result
from the heterogeneity in the tests used to assess FSIQ and the ages of blood Pb and
FSIQ examined. There could be residual confounding due to heterogeneity among studies
in the potential confounding factors examined and the method of assessment.
Across the prospective studies, blood Pb-associated FSIQ decrements at ages 4-17 years
were found with concurrent, prenatal (maternal or cord), early childhood (e.g., age 2 or 4
year), multiple year average, or lifetime average blood Pb levels. Associations also were
found with tooth Pb levels. There is no clear indication of an individual critical lifestage
or time period of blood Pb level associated with FSIQ. Concurrent blood Pb level in
children reflects recent and past Pb exposures. Thus, several observations point to an
effect of cumulative Pb exposure. However, concurrent blood Pb levels tend to be
correlated with those measured during other time periods.
Blood Pb-associated FSIQ decrements were found in populations with mean blood Pb
levels 5-10 (ig/dL. A common finding was a supralinear concentration-response
relationship, i.e., a larger decrement in FSIQ per unit increase in blood Pb level in
children in the lower range of the population blood Pb level distribution. In analyses
restricted to children in the lower range of blood Pb levels (e.g., peak <10 (ig/dL),
associations were found in groups of children with mean blood Pb (concurrent or age 2
year) levels 3-4 (ig/dL (Bellinger. 2008: Canfield. 2008: Hornung. 2008). Results from
the Boston and Rochester cohorts can be considered particularly informative because
within each cohort, there was relative homogeneity in the tests used to assess FSIQ, age
of examination, and methods used to measure potential confounding factors. Further, in
Lanphear et al. (2005). the lower portion of the blood Pb distribution largely comprised
children from the Boston and Rochester cohorts.
In the pooled analysis, Lanphear et al. (2005) found a narrow range of effect estimates,
-2.4 to -2.9 FSIQ points per log increase in concurrent blood Pb level, obtained by
excluding one study at a time. Across individual studies, there was a wide range of effect
estimates reported for blood Pb-associated FSIQ decrements. However, there was
variability among studies in model specification and blood Pb level ranges examined
(Figure 4-2 and Table 4-3). Similarly larger effect estimates were found in the Boston
and Rochester cohorts, which differed in racial and SES distributions. Although these
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studies had smaller sample sizes, they had at least as extensive consideration for potential
confounding as other studies (Canfield et al.. 2003a: Bellinger et al.. 1992). Each study
estimated larger effects for children whose peak blood Pb levels were less than 10 (ig/dL,
-1.8 FSIQ points (95% CI: -3.0, -0.60) per 1 (ig/dL increase in concurrent blood Pb level
in the Rochester cohort (Canfield et al.. 2003a) and -1.6 points (95% CI: -2.9, -0.20) per
1 (ig/dL increase in age 2 year blood Pb level in the Boston cohort (U.S. EPA. 2006b:
Bellinger and Needleman. 2003). These subsets of children had mean blood Pb levels of
3.3 (Rochester) (Canfield. 2008) and 3.8 (ig/dL (Boston) (Bellinger. 2008). lower than
those examined in other prospective studies.
4.3.2.2 Bayley Scales of Infant Development
The Mental Development Index (MDI) of the Bayley Scales of Infant Development is a
widely used test of infant mental development. MDI is a reliable indicator of current
development and cognitive functioning of the infant, integrating cognitive skills such as
sensory/perceptual acuities, discriminations, and response; acquisition of object
constancy; memory learning and problem solving; vocalization and beginning of verbal
communication; and basis of abstract thinking (McCall et al.. 1972). However, the MDI
test is not an intelligence test, and MDI scores, particularly before ages 2-3 years, are not
necessarily strongly correlated with later measurements of FSIQ in children with normal
development. In the review of the MDI evidence, emphasis was placed on results from
examinations at ages 2-3 years, which have test items more similar to those in school-age
IQ tests. Most of the prospective studies reviewed in the 2006 Pb AQCD (U.S. EPA.
2006b) found associations of higher prenatal, earlier infancy, and concurrent blood Pb
level with lower MDI score in children ages 2 to 3 years, and recent studies examined
and found associations with cord blood Pb level (Table 4-4).
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Table 4-4 Associations of blood Pb level with Bayley MDI in children ages 1 to 3 years.
Study
Study Population and Methodological Details
Grouped by strength of study methodology, generalizability3
Blood Pb Metric
Analyzed (ug/dL)
MDI Age and Subgroup
(where analyzed)
Effect Estimate
(95% Cl)b
Bellinger et al. 182 children followed from birth (1979-1981) to age 2
(1987) yr, Boston, MA area.
Prospective. Recruitment from birth hospital. High
follow-up participation. Participants had higher cord
blood Pb level, SES, HOME score, maternal education
and IQ, lower maternal age, were white. Regression
model adjusted for maternal age, race, IQ, education,
years of smoking, and alcohol drinks/week in 3rd
trimester, SES, HOME score, child sex, birth weight,
gestational age, birth order.
Prenatal (cord)
Low: <3
Medium: 6-7
High: >10
Adjusted mean at age 2 yr
in High cord blood Pb
group
vs. Medium group vs. Low
group
-3.8 (-6.3,-1.3)
-4.8 (-7.3, -2.3)
Concurrent reported not
to be associated with
MDI, quantitative data
not reported
Jedrychowski et 384-415 children born 2001-2004 followed prenatally
al. (2009b) to age 3 yr, Krakow, Poland
Prospective cohort examining multiple exposures and
outcomes. Recruitment from prenatal clinic. High
follow-up participation. Log linear regression model
adjusted for maternal education and prenatal smoking,
child sex and birth order. Did not consider potential
confounding by parenting caregiving quality.
Prenatal (cord)
Median: 1.23
Detection limit not reported
Interval analyzed: 0.44-5 =
range < 5
Age 2 yr
Age 3 yr
-1.8 (-3.4,-0.14)
-1.6 (-2.9,-0.21)
Tellez-Rojo et al. 193 children born 1994-1999 followed to age 2 yr,
(2006) Mexico City, Mexico
Tellez-Rojo Cross-sectional. Recruitment from prenatal clinic or
(2008) birth hospital. Participants had older, more educated
mothers, lower cord blood Pb level, and slightly higher
MDI. Linear regression model adjusted for maternal
age and IQ, child sex and birth weight, cohort.
Considered potential confounding by other unspecified
factors.
Concurrent
Median: 2.9
Detection limit not reported
Interval analyzed:
0.8-4.9 = range < 5
Age 2 yr
-1.7 (-3.0,-0.42)
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Table 4-4 (Continued): Associations of blood Pb level with Bayley MDI in children ages 1 to 3 years.
Study
Glaus Henn et al.
(2012)
Study Population and Methodological Details
Grouped by strength of study methodology, generalizability3
455 children born 1997-2000 followed prenatally to
age 3 yr, Mexico City, Mexico
Prospective, subset of cohort above. No selective
participation of subjects. Linear mixed effects
regression adjusted for child sex, hemoglobin, and
gestational age, maternal IQ and education, blood
Pb-Mn interaction. Did not consider potential
confounding by parental caregiving quality.
Blood Pb Metric
Analyzed (ug/dL)
Age 1 yr
Mean(SD): 5.1 (2.6)
Interval analyzed:
2.5-8.4 =10th-90th
percentiles
MDI Age and Subgroup
(where analyzed)
Ages 1 to 3 yr
Blood Mn <2 ug/dL
Blood Mn 2-2.8ug/dL
Blood Mn >2.8 ug/dL
Effect Estimate
(95% Cl)b
-2.5
-0.07 (-0.39, 0.25)
-2.2
Q'SP/fi Pic: nnt availahlp
Hu et al. (2006) 83-146 children born 1997-1999 followed prenatally to
age 2 yr, Mexico City, Mexico
Prospective, subset of Tellez-Rojo et al. (2006).
Recruitment from prenatal clinic. Moderate follow-up
participation. Eligible similar to non-eligible. Log linear
regression model adjusted for maternal age and IQ,
child sex, current weight, height-for-age Z score, and
concurrent blood Pb (in models examining prenatal
blood Pb). Considered potential confounding by other
unspecified factors.
Prenatal maternal 1st
trimester: Mean (SD) 7.1
(5.1) Interval analyzed 2.5
(10th percentile)-10
Prenatal maternal 3rd
trimester: Mean (SD) 6.9
(4.2) Interval analyzed 2.8
(10th percentile)-10
Prenatal cord: Mean (SD)
6.2 (3.9) Interval analyzed
2.5 (10th percentile)-10
Concurrent: Mean (SD)
4.8 (3.7) Interval analyzed
1.6-9.1= 10th-90th
percentiles
Age 2 yr
-0.76 (-1.5,-0.03)
-0.43 (-1.1, 0.27)
-0.06 (-0.87, 0.74)
-0.23 (-0.92, 0.45)
Gomaa et al. 161-197 children followed from birth to age 2 yr,
(2002) Mexico City, Mexico
Prospective. Recruitment from birth hospital.
Moderate participation but high retention, no selective
attrition. Log linear regression model adjusted for
maternal IQ, age, parents in home, breastfeeding
duration, parental education, child hospitalization
status and sex. Did not consider potential confounding
by parental caregiving quality.
Mean (SD)
Prenatal (cord):
6.7(3.4)
Interval analyzed:3.2 (10th
percentile)-10
Concurrent:
8.4(4.6)
Interval analyzed:3.8 (10th
percentile)-10
Age 2 yr
-0.82 (-1.5,-0.15)
-0.01 (-0.09, 0.06)
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Table 4-4 (Continued): Associations of blood Pb level with Bayley MDI in children ages 1 to 3 years.
Study
Study Population and Methodological Details
Grouped by strength of study methodology, generalizability3
Blood Pb Metric
Analyzed (ug/dL)
MDI Age and Subgroup
(where analyzed)
Effect Estimate
(95% Cl)b
Pilsner et al. 221 children born 1994-1995 followed from birth to age
(2010) 2 yr, Mexico City, Mexico
Prospective. Recruitment from birth hospital. Low but
not selective participation. Linear regression model
adjusted for maternal age, IQ, marital status, and
parity, gestational age, lowfolate intake, MTHFR
genotype. Did not consider potential confounding by
parental caregiving quality.
Prenatal (cord)
Mean (SD): 6.7 (3.6)
Interval analyzed: 3.1
(10th percentile)-10
Age 2 yr
Adjust for genotype:
MTHFR 1298
MTHFR677
-0.73 (-1.2,-0.23)
-0.71 (-1.2,-0.21)
Surkan et al. 309 children born 1993-2004 followed from birth to age
(2008) 3 yr, Mexico City, Mexico.
Cross-sectional. Recruited from prenatal or birth
hospital. High participation rate. Linear mixed effects
model adjusted for child sex, maternal age, IQ,
education, self-esteem, parity, grams/day alcohol,
smoking status, cohort. Did not consider potential
confounding by parental caregiving quality.
Age 2 yr
Mean (SD): 6.4 (4.3)
Interval analyzed: 2.0
(10th percentile)-10
Ages 1 to 3 yr
All subjects
High maternal self-esteem
Low maternal self-esteem
-0.18 (-0.45, 0.09)
0.36 (-0.50, 1.2)
-0.31 (-0.60, -0.02)
Vimpani et al. 592 children followed prenatally to age 2 yr, Port Pirie,
(1985) Australia.
Prospective. Residence near Pb smelter. High
baseline participation rate. Linear regression model
adjusted for maternal age, education, IQ, workplace,
and prenatal marital status, paternal education and
workplace, parental relationship, child birth rank,
mouthing activity, oxygen use at birth, apgar score,
neonatal jaundice, size for gestational age. Did not
consider potential confounding by parental caregiving
quality.
Maternal avg prenatal: NR Age 2 yr
Cord: NR
Age 6 mo: NR
Age 2 yr: 20% > 30
Lifetime (to age 2 yr) avg:
NR
-0.06
0.03
-0.40, p <0.05
-0.06
-0.31, p<0.05
Wasserman et al.
(1992)
392 children followed prenatally to age 2 yr, K.
Mitrovica, Pristine, Yugoslavia.
Prospective. 53% live near Pb sources. High follow-
up participation, no selective attrition. Log linear
regression model adjusted for sex, birth order, birth
weight, ethnic group, HOME score, maternal
education, age, and IQ.
Mean (SD) Age 2 yr
Cord: 14.4(10.4)
Age 6 mo: NR
Age 12 mo: NR
Age 18 mo: NR
Concurrent: 35 K.
per 3-fold increase
-1.7, p - 0.12
-1.1, p = 0.34
-1.7, p = 0.17
-1.8, p = 0.16
-2.5, p = 0.03
Mitrovica,8.5 Pristine
Geometric means
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Table 4-4 (Continued): Associations of blood Pb level with Bayley MDI in children ages 1 to 3 years.
Study
Solon et al.
(2008)
Ernhart et al.
(1988: 1987)
Study Population and Methodological Details
Grouped by strength of study methodology, generalizability3
502 children born 1998-2004, Visayas, Philippines.
Cross-sectional. Census based recruitment. No
selective participation of subjects. Two-stage linear
regression model to account for determinants of blood
Pb (sex, roof material, water source, breastfed for
> 4 months) and cognitive function (HOME score,
maternal education and smoking, born premature,
region of residence, years schooling of child).
146-165 children, followed prenatally to age 2 yr,
Cleveland, OH
Prospective. Recruitment at birth hospital. High
follow-up participation. More white, higher IQ,
nonalcohol using mothers not followed. 50% born to
alcoholic mothers. Linear regression adjusted for age,
race, sex, birth order, birth weight, parental education,
maternal IQ, Authoritarian Family Ideology, HOME
score.
Blood Pb Metric
Analyzed (ug/dL)
Concurrent
Mean (SD):
7.1 (7.7)
Interval analyzed: 1.6
(10th percentile)-10
Mean (SD)
Prenatal cord:
6.0(1.8)
Age 6 mo: 10 (3.3)
Concurrent: 17 (6.5)
MDI Age and Subgroup Effect Estimate
(where analyzed) (95% Cl)b
Ages 6 mo to 3 yr -1.07 for population
mean serum folate of
225 ug/dl_, 95% Cl not
available from model
with interaction term
Age 2 yr Variance estimate
.0003, t — 0.21
0.00, p = 0.95
0.00, p = 0.95
MDI = Mental Development Index, MTHFR = methylenetetrahydrofolate reductase
"Studies are grouped according to strength of study design, representativeness of the population characteristics and blood Pb levels examined, and extent of consideration for
potential confounding. All Mexico City studies were kept together. There is not necessarily a continuum of decreasing strength across studies.
bExcept where noted, effect estimates were standardized to a 1 ug/dL increase in blood Pb level for the lowest blood Pb range examined in the study or for blood Pb levels up to
10 ug/dL.
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The prospective studies found blood Pb-associated decrements in MDI in some large
(N = 146-592) populations with mean blood Pb levels 1.3-7.1 (ig/dL. Recruitment of
participants before or at birth without consideration of Pb exposure or maternal IQ, high
to moderate follow-up participation, and nonselective loss-to-follow-up in most studies
increase confidence that the observed associations are not due to selection bias. In
comparisons of blood Pb levels measured at various lifestages, some studies found a
stronger association of MDI with prenatal than child postnatal blood Pb levels (Hu et al.,
2006; Gomaa et al.. 2002; Bellinger et al. 1987). Most prospective studies adjusted for
birth outcomes and maternal IQ and education, and a few found associations with
additional adjustment for other SES indicators, parental smoking, parental alcohol use,
and parental caregiving quality. Cord blood Pb levels were associated with MDI, with
additional adjustment for SES and HOME score in the Boston cohort (Bellinger et al.,
1987) and HOME score in the Yugoslavia cohort (Wasserman et al.. 1992). In the
Cleveland cohort, associations of prenatal maternal and cord, age 6 month, and
concurrent blood Pb levels with MDI at age 2 years became null after adjusting for
covariates including HOME score (Ernhart et al., 1988; Ernhart et al., 1987). However,
about 50% of the original cohort was born to alcoholic mothers. While the high
prevalence of prenatal alcohol exposure may not affect the validity of the results, it may
reduce the generalizability of findings to the general U.S. population of children.
Reporting only correlation coefficients, Cooney et al. (1989b) found that prenatal
maternal and cord blood Pb levels (geometric means: 9.1, 8.1 (ig/dL, respectively) were
not associated with MDI scores in the Sydney cohort at age 2 and 3 years.
A large study of children (n = 384, 415) in Krakow, Poland, found cord blood
Pb-associated decrements in MDI at ages 2 and 3 years with lower cord blood Pb levels
(median 1.23 jig/dL, 5th-95th: 1.24-1.34 jig/dL) (Jedrychowski et al.. 2009b) than
examined in other studies (Table 4-4). However, cord blood Pb levels reflect the
pregnancy blood Pb levels of mothers. Because adult blood Pb levels are influenced by
past Pb exposures, there is uncertainty regarding the Pb exposure magnitude and pattern
of the mothers that contributed to associations between cord blood Pb level and MDI in
children. Jedrychowski et al. (2009a) estimated a larger decrease in age 3-year MDI per
unit increase in cord blood Pb level among the males (51% of subjects) than among the
females (49%). Other observations have indicated increased susceptibility of the
developing male central nervous system (CNS) to environmental insults (Moffitt et al..
2001). Although geometric cord blood Pb levels were similar in males (1.35 (ig/dL) and
females (1.41 (ig/dL), the mean age 3-year MDI score was slightly lower among males
than among females (101 and 105, respectively).
Multiple studies in various Mexico City cohorts reported associations of prenatal
(maternal or cord) or child postnatal blood Pb levels with decrements in MDI in children
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between ages 1 and 3 years (Claus Henn et al., 2012; Pilsner et al.. 2010; Surkan et al..
2008; Hu et al.. 2006; Tellez-Rojo et al.. 2006). In a subset of the cohort previously
examined by Tellez-Rojo et al. (2006). Hu et al. (2006) compared associations among
prenatal maternal blood Pb levels measured at different trimesters among 146 children at
age 2 years. An increase in first trimester maternal blood Pb levels (whole blood or
plasma) was associated with a much larger decrease in MDI score than increases in
maternal third trimester, cord, or child concurrent blood Pb levels (Table 4-4). These
results were adjusted for sex, age 2-year blood Pb level, height-for-age Z score, weight,
maternal age, and maternal IQ. Model covariates did not include SES, maternal
education, or HOME score; however, a larger unspecified list of potential confounding
factors was considered in preliminary analyses.
Consistent with several findings for FSIQ, Tellez-Rojo et al. (2006) found larger effect
estimates in children with lower blood Pb levels. In linear models, a 1 (ig/dL increase in
concurrent blood Pb level was associated with a -1.7 point (95% CI: -3.0, -0.42) change
in age 2-year MDI among 193 children with concurrent blood Pb levels <5 (ig/dL and a
-1.0 point (95% CI: -1.8, -0.26) change among 294 children with concurrent blood Pb
levels <10 (ig/dL. In a follow-up of the full cohort examined in Tellez-Rojo et al. (2006)
to age 3 years, Claus Henn et al. (2012) found inconsistent interactions between blood
Mn and Pb levels. Investigators selected mid-range (2.0-2.8 (ig/dL) blood Mn levels as
the reference group based on previous observations that MDI scores were least affected
by increases in blood Mn level in this group. Larger blood Pb-associated MDI
decrements were found in the 91 children each with blood Mn levels <2.0 (ig/dL and
>2.8 (ig/dL for age 1 year blood Pb level but not age 2 year blood Pb level. Kim et al.
(2009b) also found effect modification by blood Mn levels for the association between
blood Pb level and FSIQ, but in older children ages 8-11 years (Section 4.3.2.1).
Other recent studies in Mexico City examined effect modification by maternal self-
esteem, genetic variants, and nutritional status. Surkan et al. (2008) stratified data by the
level of maternal self-esteem as reported by mothers. Higher age 2-year blood Pb level
was associated with lower MDI score between ages 2 and 3 years among 304 children
with mothers in the lowest three quartiles of self-esteem but not among 75 children with
mothers in the highest quartile of self-esteem (Table 4-4). Model covariates included
cohort, sex, parity, and maternal IQ, age, education, smoking, grams/day alcohol use, and
self-esteem. These findings indicated that maternal psychosocial functioning may
influence the effects of Pb on the mental development of young children.
In another study in Mexico City, higher cord blood Pb level was associated with a lower
MDI score in children at age 2 years (-0.73 points [95% CI: -1.2, -0.23] in MDI per
1 (ig/dL increase in cord blood Pb level in a linear model) (Pilsner et al., 2010).
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Investigators reported a lack of effect modification by genetic variants in the
methylenetetrahydrofolate reductase (MTHFR) enzyme, which is involved in folate
metabolism. The MTHFR 677 TT genotype produces an enzyme with lower metabolic
activity, is associated with lower serum folate levels (Kordas et al.. 2009). and in this
cohort, maternal TT genotype was associated with lower mean child MDI score at age 2
years. Cord blood results from stratified analyses were not reported, thus differences in
the magnitude of association between genetic variants could not be compared.
Consistent with the prospective evidence, a recent cross-sectional analysis indicated an
association between higher concurrent blood Pb level and lower MDI score children in
the Philippines, ages 6 months to 3 years (Solon et al.. 2008). Although HOME score,
premature birth, region of residence, and maternal education and prenatal smoking were
examined as potential confounding factors, adjustment was made in two stages: first,
adjustment for blood Pb determinants and second, adjustment for MDI determinants. This
method may not adequately control for the variance shared by blood Pb level and MDI.
In this cohort, children with lower folate levels had larger Pb-associated decreases in
cognitive function. Among children with folate levels < 230 (ig/mL, 1 (ig/dL increase in
blood Pb level was associated with lower MDI score in the range of-0.80 to -2.4 points.
Among children with folate levels > 305 (ig/mL, blood Pb level was not estimated to
have a negative marginal impact. The results from this study indicated a moderating
effect of folate on blood Pb level associations since folate levels were not independently
associated with MDI. Higher folate level has been associated with lower blood Pb level
(Lee et al.. 2005a). which may be due to the role of folate in increasing Pb excretion by
inhibiting the binding of Pb to blood elements (Tandon et al.. 1987).
In summary, evidence consistently indicates associations of higher blood Pb levels with
lower MDI scores in children ages 2-3 years (Table 4-4). Key evidence was provided by
prospective studies, in particular those that adjusted for maternal IQ and education, SES,
birth outcomes, and HOME score rWasserman et al.. 1992; Bellinger et al.. 1987).
Several large studies contributed to the evidence (N = 146-592), and several studies had
high to moderate follow-up participation and nonselective loss-to-follow-up, which
reduces the likelihood of selection bias. Higher blood Pb level was associated with lower
MDI scores in a few different cohorts in Mexico City and other populations, and while
most adjusted for maternal education and IQ, they did not examine potential confounding
by parental caregiving quality. The lack of association observed in the Cleveland cohort
does not weaken the otherwise compelling evidence. The Cleveland cohort had a high
prevalence of prenatal alcohol exposure, which may not affect the validity of findings but
may limit the generalizability of results. Also, results from the more representative
Boston and Yugoslavia studies provided supporting evidence with adjustment for several
potential confounding factors. MDI at ages 2-3 years was associated with prenatal (cord
4-84
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or maternal), earlier infancy, and concurrent blood Pb levels. In comparisons of blood Pb
levels from different lifestages and time periods, some studies found a stronger
association of MDI with prenatal than child postnatal blood Pb levels (Hu et al.. 2006;
Gomaa et al. 2002; Bellinger et al.. 1987). Pb-associated decrements in MDI were found
mostly in populations with mean blood Pb levels 1.3-7.1 (ig/dL. A cross-sectional
analysis in children in Mexico City found a larger decrement in age 2 year MDI per unit
increase in concurrent blood Pb level among children with blood Pb levels <5 (ig/dL
versus 5-10 (ig/dL or <10 (ig/dL (Tellez-Rojo et al.. 2006). An association was found in
children in Poland with a median cord blood Pb level of 1.2 (ig/dL (Jedrychowski et al..
2009b). However, since cord blood Pb levels reflect blood Pb levels of mothers, and adult
blood Pb levels are influenced by past Pb exposures, the specific maternal Pb exposure
magnitude and pattern contributing to the observed associations are uncertain. Overall,
while evidence indicates associations of higher prenatal and postnatal child blood Pb
levels with lower MDI scores in young children ages 2-3 years, the impact on later
cognitive function is not certain since MDI scores do not necessarily strongly predict
later IQ in children with normal development.
4.3.2.3 Learning and Memory in Children
Epidemiologic Studies of Learning and Memory in Children
The studies reviewed in the 2006 Pb AQCD did not clearly indicate associations between
higher blood Pb level and poorer performance on neuropsychological tests of memory or
learning (i.e., acquisition of new information) in children ages 4-17 years (Table 4-5).
Studies used various tests to assess learning and memory, which may account for some of
the heterogeneity observed in findings. As described in greater detail below, evidence for
both memory and learning from prospective analyses in the Rochester, Boston, and
Cincinnati cohorts was mixed. Previous cross-sectional studies found associations
between higher concurrent blood Pb levels and poorer learning and memory, including
the large (N = 4,853) study of children ages 5-16 years participating in the National
Health and Nutrition Examination Survey (NHANES) III (Lanphear et al.. 2000). Several
recent studies, all cross-sectional, also found associations between higher concurrent
blood Pb level and poorer memory in children ages 5-16 years. Some were variants of
previous studies (Krieg et al.. 2010; Froehlich et al.. 2007): others had limited
implications because of little consideration for potential confounding (Counter et al..
2008: Min et al.. 2007V
4-85
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Table 4-5 Associations between blood Pb levels and performance on tests of learning and memory in children.
Study
Canfield et
al. (2004)
Froehlich
etal.
(2QQ7)
Bellinger
etal.
(1991)
Stiles and
Bellinger
(1993)
Study Population and Methodological Details
Prospective studies presented first, then cross-sectional studies. Within each
category, results are grouped by strength of study methodology, generalizability3
171 children born 1994-1995 followed from age 6 mo to 5 yr, Rochester,
NY
Prospective. Recruitment from study of dust control. Mostly nonwhite.
High follow-up participation, no selective attrition. Linear regression model
adjusted for maternal education, spatial span length. Also considered
potential confounding by prenatal smoking, household income, maternal
IQ, ethnicity, HOME, breastfeeding duration, 1st prenatal visit, spatial
working memory problem, birth weight, marital status, household
crowding, neonatal intensive care unit (NICU) admission, sex, age.
172-174 children born 1994-1995 followed from age 6 mo to 5 yr,
Rochester, NY
Cross-sectional. Same cohort as above. High follow-up participation, no
selective attrition. Linear regression model adjusted for income (spatial
working memory), HOME, maternal IQ, race (spatial span). Also
considered potential confounding by transferrin saturation, prenatal
smoking exposure, maternal education, age, NICU admission, sex.
150-169 children followed from birth (1979-1 981) to age Syr, Boston, MA
area
Prospective. Recruitment at birth hospital. Moderate follow-up
participation. More participants were white, had higher age 2 yr HOME
score, higher postnatal blood Pb levels. Log linear regression model
adjusted for SES, maternal IQ and marital status, preschool attendance,
HOME score, out of home care, # residence changes, medication use in
previous 12 mo, # adults in home, child sex, race, birth weight, and birth
order.
134-145 children followed from birth (1979-1981) to age 10 yr, Boston,
MA area
Prospective. Same cohort as above. Moderate follow-up participation,
participants had higher SES and HOME score. Linear regression model
Blood Pb Metric
Analyzed (ug/dL)
Lifetime (to age 5
yr) avg
Mean (SD): 7.2
(3.6)
Interval analyzed:
3.5 (10th %-ile)-10
Concurrent
Mean (SD): 6.1
(4.9)
Interval analyzed:
1.9 (10th %-ile)-10
Mean (SD)
Earlier childhood
(age2yr): 7.0
(6.6), Interval
analyzed: 1.8 (10th
%-ile)-10
Concurrent: 6.4
(4.1), Interval
analyzed: 2.5 (10th
%-ile)-10
Early childhood
Age 1 yr: NR
Age 2 yr: NR
Memory/Learning Test Effect Estimate
Analyzed (95% Cl)b
Spatial span total errors -0.1 1 (-0.20, -0.02)c
CANTAB
Age 5 yr
Spatial working memory, -0.51 (-1.2, 0.1 6)c
total errors
Spatial span length -0.02 (-0.04, 0)
CANTAB
Age 5 yr
Memory -0.14 (-0.52, 0.25)
Verbal -0.09 (-0.51, 0.34)
Quantitative -0.30 (-0.65, 0.05)
Memory 0.09 (-0.36, 0.54)
MSCA, Age 5 yr
Perseveration score, -0.02 (-0.04, 0)
CVLT 0.03(0.01,0.05)
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Table 4-5 (Continued): Associations between blood Pb levels and performance on tests of learning and memory in children.
Study
Ris et al.
(2QQ4)
Dietrich et
al. (1991)
Lanphear
etal.
(2000)
Study Population and Methodological Details
Prospective studies presented first, then cross-sectional studies. Within each
category, results are grouped by strength of study methodology, generalizability3
adjusted for HOME score, maternal IQ, SES, child stress, sex and birth
order, age 10 yr HOME score, maternal age, marital status, and race, #
residence changes. Also considered potential confounding by family
stress, birth weight, # daycare situations to age 57 mo.
195 children followed prenatally (1979-1985) to age 15-17 yr, Cincinnati,
OH
Prospective. Recruitment at prenatal clinic. High follow-up participation,
no selective attrition. Mostly African-American. Linear regression model
adjusted for SES, maternal IQ, HOME, caregiver education, adolescent
marijuana use, obstetrical complications, child sex and age. Also
considered potential confounding by birth outcomes, maternal age,
maternal prenatal smoking, alcohol, marijuana, and narcotics use, #
previous abortions, stillbirths, gravidity, parity, public assistance, child
health and Fe status.
233-236 children followed prenatally (1979-1984) to age 4 yr, Cincinnati,
OH
Prospective. Recruitment at prenatal clinic. High follow-up participation,
no selective attrition. Mostly African-American. Linear regression adjusted
for SES, birth weight, maternal IQ, prenatal marijuana use, smoking, and
alcohol use, HOME, child race, preschool attendance. Also considered
potential confounding by birth outcomes, maternal age and narcotics use,
# previous abortions, stillbirths, gravidity, parity, caregiver education,
public assistance, child age, sex, health, and Fe status.
4,853 children ages 6-16 yr (born 1972-1988), U.S. NHANES III,
1988-1994
Cross-sectional. Large U.S. representative study of multiple risk factors
and outcomes. Linear regression model adjusted for sex, race/ethnicity,
poverty index ratio, reference adult education, serum ferritin and cotinine
levels. Did not consider potential confounding by parental cognitive
function or caregiving quality.
Blood Pb Metric
Analyzed (ug/dL)
Age 1 yr: NR
All means reported
to be <8
Early childhood
(age 6.5 yr)
Mean (SD): NR
Early childhood
(Age 2 yr): NR
Concurrent: NR
Lifetime avg: NR
Concurrent
Geometric mean
(SD): 1.9(7.0)
63.5% < 2.5
Detection limit =
0.5
Interval analyzed:
1-5
Memory/Learning Test
Analyzed
# trials to 1 st category,
WCST
Age 10 yr
Memory factor score of
CVLT subtests
Learning factor score of
WRAT-3 and WISC-III
subtests
Ages 15-17 yr
Achievement score
KABC
Age 4 yr
Digit Span
WISC-R
Ages 6-16 yr
Effect Estimate
(95% Cl)b
-0.44 (-0.93, 0.05)c
0.01 (-0.02, 0.05)
-0.081 (-0.17, 0.00)
0.06, p >0.05d
0.01, p>0.05d
0.07, p >0.05d
-0.05 (-0.09, -0.01)
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Table 4-5 (Continued): Associations between blood Pb levels and performance on tests of learning and memory in children.
Study
Krieg et al.
(2010)
Surkan et
al. (2007)
Kordas et
al. (2006)
Study Population and Methodological Details
Prospective studies presented first, then cross-sectional studies. Within each
category, results are grouped by strength of study methodology, generalizability3
773 children ages 12-16 yr (born 1975-1982), U.S. NHANES III,
1991-1994
Cross-sectional. Large U.S. representative study of multiple risk factors
and outcomes. Log linear regression model adjusted for sex, caregiver
education, family income, race/ethnicity, test language. Did not consider
potential confounding by parental cognitive function or caregiving quality.
389 children, ages 6-10 years, Boston, MA, Farmington, ME
Cross-sectional. Recruitment from trial of amalgam fillings. High
participation rate. Higher participation of white children in Maine. Analysis
of covariance adjusted for caregiver IQ, child age, SES, race, birth weight.
Also considered potential confounding by site, sex, birth order, caregiver
education and marital status, parenting stress, and maternal utilization of
prenatal and annual health care but not parental caregiving quality.
293 children, ages 6-7 yr, Torreon, Mexico.
Cross-sectional. Recruitment at schools. High participation rate.
Residence near metal foundry. Linear regression model adjusted for child
sex, age, school, birth order, hemoglobin, forgetting homework,
household possessions and crowding, home ownership, maternal
education, family structure, urinary arsenic, tester, school. Did not
consider potential confounding by parental cognitive function or
caregiving quality.
Blood Pb Metric
Analyzed (ug/dL)
Concurrent
Mean (5th-95th):
1.95(1.63-2.27)
Detection limit = 1
Interval analyzed:
1. 7-2.2 = 10th-90th
percentiles
Concurrent
Group 1:1-2
Group 2: 3-4
Group 3: 5-10
Mean (SD): 2.2
(1.6)
Concurrent
Mean (SD): 7.1
(1.7)
Interval analyzed:
2-10.0
Memory/Learning Test
Analyzed
Digit span
WISC-R
Ages 12-16 yr
General memory index,
WRAML
Ages 6-10 yr
Sternberg memory
PPVT
Ages 6-7 yr
Effect Estimate
(95% Cl)b
-0.34 (-0.59, -0.08)
Reference
-0.69 (-4.4, 3.0)e
-6.7 (-12, -1.2)e
-0.16 (-0.37, 0.05)
-0.71 (-1.4, 0.02)
4-8
-------
Table 4-5 (Continued): Associations between blood Pb levels and performance on tests of learning and memory in children.
Study
Study Population and Methodological Details
Prospective studies presented first, then cross-sectional studies. Within each
category, results are grouped by strength of study methodology, generalizability3
Blood Pb Metric
Analyzed (ug/dL)
Memory/Learning Test
Analyzed
Effect Estimate
(95% Cl)b
Chiodo et 208-236 children, age 7.5 yr, Detroit, Ml area
al. (2004) Cross-sectional. Recruitment at prenatal clinic. 100% African-American.
High prevalence of prenatal alcohol and drug exposure. Moderate
participation rate. Log linear regression model adjusted for SES (both
outcomes); caregiver vocabulary, disruption in caregiver (verbal learning);
child age and sex, parity, caregiver education (spatial span). Also
considered potential confounding by HOME score, maternal prenatal
marijuana, smoking, alcohol, or cocaine use, crowding, child life stress,
caregiver age, life stress, and psychopathology, conflict tactics, family
functioning, # children <18 years.
Concurrent
Mean (SD):
5.4(3.3)
Interval analyzed:
2.2-9.5 = 10th-90th
percentiles
Verbal learning, WRAML -0.20, p>0.05d
Corsi Backward Spatial
Span
Age 7.5 yr
-0.22, p>0.05d
CANTAB = Cambridge Neuropsychological Test Automated Battery, MSCA = McCarthy Scale of Children's Abilities, CVLT = California Verbal Learning Test, WCST = Wisconsin Card
Sorting Test, WISC = Wechsler Intelligence Scale for Children, WRAML = Wide Range Assessment of Memory and Learning, PPVT = Peabody Picture Vocabulary Test.
aResults are presented first for prospective studies then for cross-sectional studies. Results from the same cohort are kept together. Within each category, results are grouped by
representativeness of the population characteristics and blood Pb levels examined and extent of consideration for potential confounding. There is not necessarily a continuum of
decreasing strength across studies.
bEffect estimates are standardized to a 1 ug/dL increase in blood Pb level in the lowest range of blood Pb levels examined in the study or the interval from the 10th percentile to
10 ug/dL or the 90th percentile, whichever is lower.
The direction of the effect estimate was changed such that a negative estimate represents poorer performance and a positive estimate represents better performance.
Sufficient data were not available to calculate 95% Cl.
eEffect estimate compares test performance of children in higher blood Pb groups to children in lowest blood Pb group.
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The prospective studies had smaller sample sizes (n = 134-241) than cross-sectional
studies (n = 172-4,853) but greater examination of potential confounding. Further,
recruitment of participants before or at birth, moderate to high follow-up participation,
and in most cases follow-up not biased to higher blood Pb levels and lower cognitive
function reduce the likelihood of selection bias (Table 4-5). Another strength was the
examination of early childhood (e.g., age 1, 2, or 6.5 year) or lifetime average (to age 5
years) blood Pb levels, which better indicated the temporal sequence between Pb
exposure and decrements in learning and memory at ages 5 to 17 years.
There were contrasting associations between blood Pb levels and memory in the
Rochester and Boston cohorts at age 5 years (Canfield et al.. 2004; Bellinger et al.. 1991).
The population means were similar (6.4, 7.1 (ig/dL), but different blood Pb metrics were
examined. In the Rochester cohort, a 1 (ig/dL increase in lifetime average blood Pb level
was associated with 0.11 (95% CI: 0.02, 0.20) more total errors on the spatial span
memory task (i.e., errors in replicating a sequence pattern) with adjustment for maternal
age and spatial span length and consideration for several other factors (Table 4-5)
(Canfield et al.. 2004). Recent evidence in the Rochester cohort extended findings to
associations between poorer performance on spatial working memory tasks and higher
concurrent blood Pb levels (Froehlich et al.. 2007). which also represent past and recent
Pb exposure. In the recent analysis, associations were found with spatial span length
(number of squares recalled correctly) with adjustment for HOME score, maternal IQ,
and race; and with spatial working memory errors with adjustment for family income. In
each analysis of the Rochester cohort, multiple associations were examined; however,
there were consistent patterns of blood Pb-associated decrements in cognitive function
observed across the various indices of memory, learning, and executive function
examined. Coherence for associations with performance on spatial span and spatial
working memory tasks was found with evidence in rodents for Pb-induced impaired
performance on visual-spatial memory tasks in the Morris water maze and working
memory tasks in the radial arm maze, respectively (discussed below). In the Boston
cohort at age 5 years, concurrent blood Pb level was not associated with poorer memory,
as assessed using the McCarthy Scale of Children's Abilities (Bellinger et al.. 1991).
These results were adjusted for more potential confounding factors than results from
other studies, including SES, maternal IQ, and HOME score. Higher age 2 year blood Pb
level was associated with poorer memory at age 5 years, but the association lacked
precision (Table 4-5). Higher age 1 year blood Pb level was associated with better
memory at age 10 years (i.e., fewer errors in recalling word list) as assessed with the
California Verbal Learning Test (Stiles and Bellinger. 1993). In the Cincinnati cohort,
higher blood Pb level was associated with improved learning at age 4 years (with age 2
year or lifetime average blood Pb) (Dietrich et al.. 1991) and not associated with memory
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at ages 15-17 years (with age 6.5 year blood Pb) (Ris et al., 2004) with adjustment for
similar potential confounding factors plus adolescent marijuana use at ages 15-17 years.
With respect to learning, associations in the Boston cohort were inconsistent across the
various tests conducted, ages at which learning was assessed (ages 5, 10 years) and time
periods of blood Pb levels (ages 1, 2, 5, 10 years) examined (Stiles and Bellinger. 1993;
Bellinger et al., 1991). In two different Boston-area cohorts examined at different ages,
poorer learning (number of trials to sort cards properly or number of categories achieved)
as assessed with the Wisconsin Card Sorting Test (WCST) was associated with higher
age 1 year blood Pb level in children ages 10 years (Stiles and Bellinger. 1993) and with
higher childhood tooth Pb levels in young adults ages 19-20 years (-0.6 categories [95%
CI: -1.0, -0.21] per natural log unit increase in tooth Pb level [collected in lst/2nd grade],
with adjustment for parental IQ, maternal age and education, SES, sex, birth order,
current smoking status, drug use, and alcohol use) (Bellinger et al.. 1994a). In the study
of 10 year-old children, poorer performance on the WCST was not consistently found in
association with the various ages of blood Pb level examined (Stiles and Bellinger. 1993).
Components of the WCST fall predominately in the executive function domain; however,
some that measure acquisition of information also fall into the learning domain.
The cross-sectional studies examined potential confounding by parental education and
SES, but a notable omission in many studies was consideration for parental caregiving
quality. Chiodo et al. (2004) found concurrent blood Pb-associated decrements in spatial,
verbal, and working memory with examination of HOME score. While not affecting
internal validity of results, high prevalence of prenatal alcohol exposure in this cohort
may limit their generalizability. HOME score was not always associated with memory.
For example, in the Rochester cohort, only household income remained significantly
associated with total errors in the spatial working memory task and was included in the
final model (Froehlich et al.. 2007). Therefore, the confounding factors may vary among
endpoints. Studies found Pb-associated decrements in learning and memory in
populations with mean lifetime or concurrent blood Pb levels 5.4 to 8 (ig/dL. Despite the
lack of information on HOME score, the cross-sectional analyses of children participating
in NHANES III had several strengths, including large sample sizes (n = 773-4,853), high
participation rates, lower likelihood of selection bias based specifically on Pb exposure
due to the examination of multiple risk factors and outcomes, nationally-representative
results, and examination of the shape of the concentration-response relationship (Krieg et
al.. 2010; Lanphear et al.. 2000). Higher concurrent blood Pb level was associated with
lower digit span score in adolescents ages 12-16 years (Krieg et al.. 2010) and children
6-16 years (Lanphear et al.. 2000). Among children ages 6-16 years, Lanphear et al.
(2000) found the largest decrement in memory score per 1 (ig/dL increase in blood Pb
level in children with concurrent blood Pb levels <2.5 (ig/dL (-0.25 [95% CI: -0.58, 0.08]
4-91
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points versus -0.05 [95% CI: -0.09, -0.01] points among all subjects). A nonlinear
relationship also was found among children ages 7 years living near a metal foundry in
Torreon, Mexico, with a larger Pb-associated decrement in learning and memory found in
children with concurrent blood Pb levels <10 (ig/dL (Kordas et al.. 2006). These findings
of nonlinearity based only on concurrent blood Pb without regard to early or peak
childhood blood Pb levels are less certain because the magnitude and timing of Pb
exposures contributing to the associations are uncertain.
lexicological Studies of Learning and Memory
As described in the preceding sections, blood Pb levels are consistently associated with
decrements in FSIQ in children but show variable associations with performance on tests
of learning and memory. A relationship between Pb exposure and cognitive function
decrements is supported further by evidence for Pb-induced impairments in memory and
learning in animals. In the 2006 Pb AQCD (U.S. EPA. 2006b). several studies reported
Pb-induced impaired memory and learning in animals with Pb exposures that resulted in
blood Pb levels 30-50 (ig/dL; however, some studies observed impairments in rodents
(gestational exposure) and monkeys (lifetime postnatal exposure) with blood Pb levels
15-25 (ig/dL (Altmann et al.. 1993; Rice and Karpinski. 1988; Gilbert and Rice. 1987). A
few recent studies added to the evidence for impaired learning processes, associative
ability, or memory in animals with blood Pb levels in the range relevant to this ISA, 10-
25 (ig/dL (Corv-Slechta et al.. 2010: Niu et al.. 2009: Stangle et al.. 2007). These blood
Pb levels resulted from gestational-lactational, lactational, or lifetime (beginning in
gestation or at birth) Pb exposure in drinking water. Results from animal studies on
learning and other nervous system endpoints that provided concentration-response
information (i.e., those with multiple Pb exposure concentrations) are shown in Figure
4-3 and Table 4-6. These results demonstrate the coherence among inter-related CNS
changes induced by Pb exposure in animals, including deficits in CNS development and
plasticity and alterations in neurotransmitters, which mediate cognitive function.
4-92
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o Highest Concentration
4 Lowest Cone, with Response
A Highest Cone, with No Response
o Lowest Concentration
Outcome Study Study ID Exposure Period
Behavior Beaudin et al. {2007) 1
Grantetal. (1980) 2
Kishietal. (1983) 3
Overmannetal. (1977) A
Cognition Overmannetat (1977) 4
Stangie et al. (2007) 5
Gongand Evans(1997) 6
Li el IS (2009q) 7
Lielal (2010) 8
Motor Function Kishietal. (1983) 3
Overmannelal. (1977) 4
Stangie et al. (2007) 5
Leasureetal. (2008) 14
Physical Kishietal (1983) 3
Development
Gorigand Łvans{1997) 6
Huetal. (200*)) 12
Morphotogy Gongand Evans{1997) 6
Lietal (2009c) 7
List al. (2010) 8
Huetal. (20086) 12
Tavakoli-Nezhadetal (2001) 13
Neurophysiology Huetal. (20086) 12
ProtWDiff/Surv Lielal (2010) 8
Oxidalwe Stress Li et al (2009c) 7
Cortieosterone Virgolin et al. (2005) 10
Virgctini et a). (2008a) 11
Neurotransmstter Virgosmi et al. (200Sa) 1 1
Leasure et al. (2008) 14
Fortune and Lurie(2009) 15
Stress-Cognition Virgoiini et al. (2008t>) 9
Stress - VirgoJini et al (20080) 8
Cocticosterone
VirgolM et al. (2005) 10
Wsoiini et al, (2008a) 11
Stress- Virgeiira et al. (200Gb) 9
Neyrotransmitter
Virgolini et al. (2006a) 11
Neooaial
Ges?atianalMetime
Neonasal
Neonatal
Neonatal
Neonatal
Adult
Neonatal
Neonatal
Neonatal
Neonatal
Neonatal
Neonatal
Neonatal
Adult
Neonatal
Adult
Neonatal
Neonatal
Neonatal
Juvenile/Adult
Neonatal
Neonatal
Neonatal
juvenile/Lifetime
Neonatal
Neonatal
Neonatal
Neonatal
Neonatal
Neonatal
Juvenile/Lifetime
Neonatal
Neonatal
Neonatal
IJfestage/Species
Rat, adult
Rat
Rat, adult
Rat. adult
Rat, adult
Rat. adult
Rat, adult
Mouse, neonate
Mouse, neonate
Rat. adult
Rat, adult
Rat, adult
Mouse, adult
Rat. neonate
Rat, adult
Rat, neonate
Rat, adult
Mouse, adullt
Mouse, neonate
Rat, neonate
Rat, adult
Rat, neonate
Mouse, neonate
Mouse, neonate
Rat, adult
Rat, adult
Rat. adult
Mouse, adule
Mice, neonate
Rat, adult
Rat, adult
Rat. adult
Rat, adult
Rat, adult
Rat, adult
• •
r\ . A*
V-/ OF
• ft
ft
• — •
23— -*»
Q — «•
•„ : JK
™-™..,.™s.~ ,«™™^|p
3 — *
«fi
w
ŁS +
» ••
B — •
* •
Q— «•
* «
a » o
e — m
o — «•
* — *
•—A
«s — *
:Q «.
• A
-------
Table 4-6 Summary of findings from neurotoxicological concentration-
response array presented in Figure 4-3.
Study
ID in
Figure
4-3 Reference
Blood Pb
Level
(ug/dL)
Outcome
1
Beaudin et al.
(2007)
PND52:
13&31
Behavior: Neonatal Pb exposure (birth to 4 weeks of age, drinking water)
during and after lactation, Adult offspring showed over-reactivity to reward
omission and errors in olfactory discrimination task at PND62-64, females.
Grant et al. PND1: Behavior: Gestational, lifetime Pb exposure (drinking water) to dams and
H980) gy pups, Changed locomotor development at PND14, both sexes.
Kishi et al.
(1983)
PND22: Behavior: Neonatal Pb exposure (daily oral gavage of rats PND3-21) during
59 lactation, Changed emotional behavior at PND59-60, males.
Motor function: Neonatal Pb exposure (daily oral gavage of rats PND3-21)
186 during lactation, Motor function (rotarod performance) impaired at PND53-58,
males.
59, 152, &
186
Physical development: Neonatal Pb exposure (daily oral gavage of rats
PND3-21) during lactation, Delayed development of righting reflex at PND10-
17, males.
Overmann
(1977)
PND21: Behavior: Neonatal Pb exposure (oral gavage of rats PND3-21) during
226 lactation, Aversive conditioning affected at PND26-69, both sexes.
33, 174 & Cognition: Neonatal Pb exposure (oral gavage of pups) during lactation,
226 Response inhibition impaired at PND67-89, both sexes.
33, 174,
226
Motor function: Neonatal Pb exposure (oral gavage of pups) during lactation,
Increased nocturnal motor activity, impaired motor coordination (rotarod) at
PND67-89, both sexes.
Stangle et al.
(2007)
PND52:
13&31
Cognition: Neonatal Pb exposure (PND1-PND30, drinking water) during
lactation, Impaired learning with visual discrimination task, heightened
response to errors at PND62, both sexes.
PND52:
31
Motor function: Neonatal Pb exposure (PND1-PND30, drinking water) during
lactation, Alcove latency and response latency significantly affected at
PND62, both sexes.
Gong and Day 21: Cognition: Adult male 21 day Pb exposure (drinking water), Increased
Evans (1997) 33 & 35 hyperactivity with habituation to new cage environment at days 21-42.
85
85
Morphology: Adult male 21 day Pb exposure (drinking water), Increased
hippocampal glial fibrillary acidic protein (GFAP), marker of neuronal injury.
Physical development: Adult male 21 day Pb exposure (drinking water),
Increased brain GFAP.
4-94
-------
Table 4-6 (Continued): Summary of findings from neurotoxicological concentration-
response array presented in Figure 4-3.
Study
ID in
Figure
4-3 Reference
7 Li et al.
(2009c)
8 Li et al.
(201 Ob)
9 Virgolini et al.
(2008b)
10 Virgolini et al.
(2005)
Blood Pb
Level
(ug/dL)
PND21:
80 & 100
40, 80, &
100
40, 80, &
100
80 & 100
PND21:
19
19&30
19&30
Age 3
months:
15
Outcome
Cognition: Gestational & lactational Pb exposure (drinking water), Morris
water maze performance impaired at PND21, both sexes.
Morphology: Gestational & lactational Pb exposure (drinking water),
Increased levels of inflammatory cytokines & exocytosis-related proteins in
brains at PND21 both sexes.
Oxidative stress: Gestational & lactational Pb exposure (drinking water),
Elevated hippocampal levels of pro-inflammatory cytokine TNF-a in brains at
PND21, both sexes.
Cognition: Gestational & lactational Pb exposure (drinking water), Morris
water maze performance impaired at PND21, both sexes
Morphology: Gestational & lactational Pb exposure (drinking water),
Increased hippocampal levels of P-tau and amyloid beta, Alzheimer disease-
associated proteins at PND21, both sexes.
Proliferation/differentation/survival: Gestational & lactational Pb exposure
(drinking water), Increased hippocampal expression of P-tau and amyloid
beta at PND21, both sexes.
Stess-induced Corticosterone: Gestational & lactational Pb exposure
(drinking water, dams) plus stress, Elevated corticosterone with prenatal +
offspring stress was further enhanced with Pb exposure, males.
Stress-induced Cognition: Gestational & lactational Pb exposure (drinking
water, dams) plus stress, Pb often has opposite effect of stress on Fl
performance, both sexes
Stress-induced Neurotransmitter: Gestational & lactational Pb exposure
(drinking water, dams) plus stress, Changes in Fl performance linked to
changes in dopamine and serotonin, males.
Corticosterone: Lifetime Pb exposure (drinking water) from weaning, Pb
exposure alone decreased basal plasma corticosterone levels at 5 months of
age, males.
15&27
Stress-induced Corticosterone: Lfetime Pb exposure (drinking water) from
weaning plus cold stress increased corticosterone levels, males.
11
Virgolini et al.
(2008a)
Measured
in dams
PND21'
31
Corticosterone: Gestational & lactational Pb exposure (maternal, drinking
water), Increase in basal corticosterone levels in non-behavioral tested rats
but not in behaviorally-tested rats, females at age 4-5 months.
Stress-induced Corticosterone: Gestational & lactational Pb exposure
12 (maternal, drinking water) plus maternal stress, Decreases plasma
corticosterone, males behaviorally tested at age 4-5 months.
31
12&31
Neurotransmitter: Gestational & lactational Pb exposure (maternal, drinking
water), Induced NE aberrations in adult rat offspring, both sexes.
Stress-induced neurotransmitter; Gestational & lactational Pb exposure
(maternal, drinking water)+stress: Induced HVA (monoamine
neurotransmitter metabolite) and NE aberrations in adult rat offspring, both
sexes.
4-95
-------
Table 4-6 (Continued): Summary of findings from neurotoxicological concentration-
response array presented in Figure 4-3.
Study
ID in
Fiqure
4-3
12
Reference
Huetal.
(2008b)
Blood Pb
Level
(ug/dl_)
PND1:
4, 10&12
Outcome
Morphology: Gestational Pb exposure (drinking water), Decreased
outgrowth marker PSA-NCAM in rat pups at PND1, both sexes.
neurite
10X19 Neurophysiology: Gestational Pb exposure (drinking water), Decreased
hippocampal sialyltransferase activity in rat pups at PND21, both sexes.
4, 10&12
Physical development: Gestational Pb exposure (drinking water), Decreased
neurite outgrowth marker PSA-NCAM in rat pups at PND1, both sexes.
13 Tavakoli- Timing Morphology: Juvenile/adult Pb exposure (3-6 weeks, starting at PND22,
Nezhad et al. NR: drinking water), Decreased number of spontaneously active midbrain
(2001) 29, & 54 dopamine neurons at end of exposure, males
14
Leasure et al.
(2008)
Peak
during
PNDO-10:
10&42
Motor function: Gestational & lactational Pb exposure (maternal, drinking
water), Decreased rotarod performance in offspring at age 1 year, males.
Neurotransmitter: Gestational & lactational Pb exposure (maternal, drinking
water), Affected dopamine homeostasis in offspring at age 1 year, males.
15 Fortune and Timing Neurotransmitter: Gestational & lactational Pb exposure (maternal, drinking
Lurie (2009) NR: water), Affected offspring superior olivary complex (auditory)
8 & 42, neurotransmitters at PND21, both sexes.
4-96
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Learning and Memory - Morris Water Maze
In the Rochester cohort of children, higher blood Pb level was associated with poorer
performance on tests of spatial memory (Table 4-5). In animals, learning and spatial
memory have been tested using the Morris water maze. These functions are assessed
during training by measuring the time or distance required for a rodent to swim to a
submerged platform using visual cues outside the maze. After animals have learned the
location of the platform, memory for its location is assessed by removing the platform
and measuring the time the animal spends in that area of the maze. The 2006 Pb AQCD
(U.S. EPA. 2006b) reported mixed effects of Pb on behavior in the maze; some studies
found Pb-induced impaired performance in animals whereas others found improved
performance (U.S. EPA. 2006b). Evidence was more consistent for Pb-induced
impairments in long-term memory (relatively permanent storage and use of unchanging
information) which was found in animals with gestational, gestational/lactational, or
gestation/lifetime Pb exposure producing blood Pb levels 23-33 (ig/dL (Yang etal.. 2003;
Jettetal.. 1997; Kuhlmann et al. 1997). Using the Morris water maze, Jett et al. (1997)
found an effect in female juvenile rats of gestational/lactational Pb exposure (Pb acetate
in maternal feed 10 days prior to mating to PND21) on long-term memory but not on
working memory, which reflects temporary storage and manipulation of information
necessary for the completion of complex tasks (Baddelev. 1992). Also using the Morris
water maze, Kuhlmann et al. (1997) compared various lifestages of Pb exposure and
found impaired learning and long-term memory in Long-Evans adult rats exposed to Pb
during gestation and lactation (via maternal diet) or over a lifetime from gestation
through adulthood. Each of these exposure periods produced peak blood Pb levels of
59 (ig/dL, which are higher than those relevant to this ISA. Pb exposure during only the
post-weaning period, which produced more relevant blood Pb levels of 23 and 26 (ig/dL,
did not affect memory.
In contrast with Kuhlmann et al. (1997). recent evidence points to impairments in
memory and learning as assessed with the Morris water maze with postweaning Pb
exposure, albeit with higher blood Pb levels than relevant to this ISA. Impaired learning,
i.e., slower decrease in time to escape from the Morris water maze across training trials,
was found in weanling Sprague-Dawley male rats exposed to 400 or 800 mg/L Pb acetate
for 8 weeks in drinking water (Fan et al.. 2010; Fan et al.. 2009a). Further, various dietary
supplements (by oral gavage) in these weanling males mitigated the effect of Pb on
escape latency to resemble that of control pups at the end of training. In Fan et al.
(2009a). Zn and methionine given before Pb exposure prevented Pb-induced impairments
in spatial memory and learning, whereas glycine, taurine, vitamin C, vitamin Bl, tyrosine
had no effect. In this study, blood Pb levels ranged from 7 to 70 (ig/dL depending on the
type and timing of nutrient supplementation and recovery time. In Fan et al. (2010).
4-97
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Pb-exposed rat pups (mean blood Pb level: 50 (ig/dL) showed deficits in retaining
information about the platform location in probe tests with the platform removed.
Methionine-choline treatment given before or concomitantly with Pb exposure prevented
these impairments in spatial memory. Rats exposed to both Pb and methionine (25, 31
(ig/dL) had lower blood and brain Pb levels than rats exposed only to Pb, but the results
do not identify whether the protective effects of methionine are attributable to its function
as a chelator and/or role as a free-radical scavenger. Choline is important for cell
membranes and neurotransmitter synthesis (Zeisel and Blusztain. 1994). and unrelated to
Pb exposure, choline supplementation of rats PND16-PND30 or GD17-GD12 was shown
to attenuate normal age-related declines in spatial memory (Meek et al.. 2007).
Consistent with Kuhlmann et al. (1997). other recent studies found impairments in
learning and memory in animals as assessed using the Morris water maze with
gestational/lactational Pb exposures, albeit at higher concentrations than those relevant to
this ISA. In Li et al. (2009c). rodents were exposed GD1-PND21 to Pb acetate via the
drinking water of dams (1,000-10,000 ppm) and had corresponding blood Pb levels of
40-100 (ig/dL at PND21. Beginning at weaning, learning and spatial memory were
assessed in pups with a reversal procedure in the Morris water maze. Pb-exposed pups
with blood Pb levels of 80 and 100 (ig/dL had statistically significant increases in escape
latency and number of platform crossings (Li et al.. 2009c). Pups in Li et al. (2009c) were
not separated by sex. Cao et al. (2009) found that long-term postnatal Pb exposure from
birth (2,000 ppm Pb acetate in drinking water) impaired spatial memory in male Wistar
rats as adults (PND81-90), and these effects were exacerbated by long-term
administration of melatonin (3 mg/kg by gastric gavage, 60 days from weaning).
Mechanistic support for effects on learning and memory was provided in this study by
observations in the hippocampal dentate gyrus that Pb exposure also impaired long-term
potentiation (LTP), a major cellular mechanism underlying learning and memory.
Working Memory - Delayed Spatial Alternation
Working memory also can be measured by testing delayed spatial alternation (DSA). In
DSA, an animal receives rewards by alternating responses between two locations or
levers. This test requires working memory because the correct response changes between
trials, and the animal must determine which response is correct based on memory of its
previous response. Impaired working memory is indicated by increased response errors,
decreased percent of correct responses, and increased perseverative errors (i.e., repeatedly
pressing the same lever without moving to the other lever when the reward is moved).
Studies detailed in earlier Pb AQCDs showed that Pb-exposed animals had deficits in
working memory as assessed with DSA (Alberand Strupp. 1996; Rice and Gilbert.
1990a: Rice and Karpinski. 1988; Levin et al.. 1987). Studies in monkeys (ages 6-9
4-98
-------
years) showed that there were multiple lifestages and durations of Pb exposure that
induced poorer performance on DSA tasks, including infancy-only, lifetime from birth, or
lifetime after weaning (Rice and Gilbert. 1990a; Rice and Karpinski. 1988; Levin et al..
1987). Some of these effects were observed in monkeys with blood Pb levels relevant to
this ISA: 15, 26 (ig/dL during Pb exposure. Pb-induced impairments in DSA task
performance have been observed less consistently in rats with juvenile only or juvenile to
adult exposure. In fact, 8.5 month-long Pb exposure beginning at day 21 or 8 months was
shown to increase accuracy of performance on DSA tasks in rats (Cory-Slechta et al..
1991).
Learning and Memory - Y Maze
The three-branch radial Y-maze test evaluates learning as the number of days required to
learn the maze (90% correctly). The Y-maze has a light at the end of one of the branches.
The branch with the illuminated light is the safe area whereas the other two branches are
electrified and deliver a mild electric shock when entered. The spatial memory test
assessed by the Y-maze test is similar to the spatial working memory test used by
Froehlich et al. (2007) in the Rochester cohort, which requires children to search boxes
for a reward and avoid returning to boxes where the reward was previously found. A
recent study using the Y-maze showed impaired learning in Wistar albino rat offspring
exposed to Pb from lactation to adulthood up to 12 weeks of age (300 mg/dL Pb acetate
in dam drinking water and then in offspring drinking water postweaning) (Niu et al..
2009). Pb induced statistically significant impairments in learning at 8, 10 and 12 weeks
of age but not age 6 weeks. These effects on learning were found with blood Pb levels
relevant to this ISA, 17 (ig/dL, as measured at age 6 weeks. Mechanistic support for
Pb-induced learning impairments in this study was provided by observations of
concomitant Pb-induced attenuation in levels of hippocampal glutamate (Section
4.3.10.4). which mediates signaling pathways involved in LTP.
Learning - Schedule-Controlled Behavior Testing
The 2006 Pb AQCD described the effects of Pb exposure on learning in animals as
measured with operant conditioning using Fixed Interval (FI) or Fixed Ratio (FR)
reinforcement schedules (U.S. EPA. 2006b) and indicated differential effects by Pb
exposure concentration, with lower-level (e.g., < 10 mg/kg, blood Pb level: 11 (ig/dL)
and higher-level Pb (e.g., > 10 mg/kg) exposures increasing and decreasing FI response
rate, respectively (Cory-Slechta. 1994). This nonlinear response has been examined
further in recent work by the Cory-Slechta laboratory, much of which also examined the
interaction between stress and Pb exposure. Impaired performance in FI testing with Pb
exposure also supports the effects of Pb on response inhibition, which are discussed in
Section 4.3.3.1.
4-99
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Recent evidence indicated that particular learning impairments induced by certain levels
of Pb exposure were modifiable. Female rats were exposed to 300 ppm Pb acetate via
dam drinking water from birth through lactation PND1-PND17 and then 20 or 30 ppm Pb
via their own drinking water to PND30. Rats exposed to Pb alone (blood Pb levels at
PND52: 13 and 31 (ig/dL) learned the visual discrimination task more slowly than
controls, indicating impaired associative ability (Stangle et al.. 2007). Garavan et al.
(2000) previously had similar observations in rats with gestational plus lacational Pb
exposure; however, mean peak blood Pb levels were >100 (ig/dL. In Stangle et al. (2007).
rats subsequently administered succimer PND31-PND52 (twice daily by oral gavage,
resulting in blood Pb level 2.8 (ig/dL) performed better on visual discrimination tasks
than did Pb-exposed rats. Brain Pb levels also were lower in the Pb+succimer group (196
ng/g) than in the two Pb-only groups (1,040 and 3,690 ng/g). In this study, succimer
alone in the absence of Pb exposure resulted in some impairment in associative ability,
and rats given succimer after a higher Pb concentration (30 ppm, blood Pb level
8.5 (ig/dL) did not have better associative ability than rats exposed to 30 ppm Pb alone
(blood Pb level: 31 (ig/dL). Therefore, succimer administration did not prevent the effects
of all Pb exposure on impairments in associative ability.
Learning Ability with Stress
The paradigm of combined Pb and stress exposure experienced by a laboratory animal
has been examined by the Cory-Slechta laboratory, and they have focused on the
common pathway of altered HPA axis and brain neurotransmitter levels. Effects on
learning varied, depending on the timing of stress, Pb exposure concentration, and sex of
the animal. Pb-stress interactions were found with dietary Pb exposures that resulted in
blood Pb levels relevant to this ISA. Evidence additionally indicates that associations of
Pb exposure and stress with learning deficits (multiple schedule of repeated learning and
performance in females) may be related to aberrations in corticosterone and dopamine.
As indicated in Figure 4-3 and Table 4-6. Pb exposure with stress has been shown to alter
corticosterone levels and exacerbate Pb-induced dopamine release and learning
impairments. For example, learning deficits in female rat offspring at age 2 months were
enhanced following lifetime Pb exposure combined with prenatal stress, i.e., maternal
restraint (Cory-Slechta et al.. 2010). This exposure paradigm involved exposure of dams
to 50 ppm Pb acetate from 2 months prior to mating through lactation and exposure of
their pups from a mixed sex litter to 50 ppm Pb in drinking water through the remainder
of their lifetime (2 months). The peak blood Pb levels measured in pups (age 5-6 days)
ranged from 10 to 13 (ig/dL. Sequence-specific learning deficits were found, but
performance measures (e.g., accuracy, response rate) were not affected. Pb/stress was
found to increase the total number of responses required to learn a response sequence.
Pb/stress exposure also increased dopamine from the frontal cortex and dopamine
4-100
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turnover in the nucleus accumbens, which are involved in mediating cognition. Also,
Pb-exposed offspring with and without maternal stress exposure had statistically
significant decreases in hippocampal nerve growth factor versus controls.
Another study of lifetime Pb exposure (50 or 150 ppm Pb acetate in drinking water of
dams from 2 weeks before pregnancy though lactation and of offspring thereafter)
indicated a potentiation of some effects on learning with Pb and stress co-exposure, with
stress given prenatally via dams or postnatally to offspring. Effects were observed
primarily in female offspring. Compared with no Pb/no stress controls, lifetime 50 ppm
Pb exposure plus prenatal or postnatal stress resulting in mean blood Pb levels
11-16 (ig/dL (measured at different time points) decreased the post-reinforcement pause
(PRP) period in female offspring when examined starting at age 2 months (Rossi-George
et al.. 2011) (Table 4-7. rightmost column). Animals with 150 ppm lifetime Pb exposure
and mean blood Pb levels 25-33 (ig/dL had decreased PRP only with prenatal stress co-
exposure. Within the FI schedule, the PRP represents timing capacity or proper temporal
discrimination and refers to the period during which the animal waits or pauses before
depressing the lever for a reward. In this case, decreased pause or PRP interval in Pb plus
stress-exposed animals indicates that they started responding earlier than did controls.
These results also point to an effect of Pb on increasing impulsivity (Section 4.3.3.1).
Separately, the overall FI response rate, which also indicates impulsivity (i.e., rate of not
withholding responses), was significantly increased with 50 ppm lifetime Pb exposure
alone and with co-exposure to maternal or offspring stress. At 150 ppm, lifetime Pb
increased FI response rate only with co-exposure to stress (maternal or offspring).
Biochemical analysis revealed alterations in frontal cortex norepinephrine, reductions in
dopamine homeostasis in the nucleus accumbens, and enhancement of the striatal
monoamine system as possible mechanistic contributions to Pb-induced impairments in
learning.
4-101
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Table 4-7 Summary of effects of maternal and lifetime Pb exposure on
Fl performance observed by Cory-Slechta laboratory.
Pb (ppm)
Maternal Pbb
Overall rate3
PRPa
Lifetime Pbc
Overall rate3
PRPa
0 ppm
0-PS
o-os
No Significant Effect
No Significant Effect
No Significant Effect
*1 -23%
No Significant Effect
No Significant Effect
No Significant Effect
No Significant Effect
50 ppm
50-NS
50-PS
50-OS
No Significant Effect
No Significant Effect
*f 64.9%
No Significant Effect
No Significant Effect
No Significant Effect
*f 95%
*f 79.2%
*f 74.7%
No Significant Effect
*1 -42%
* 1-39.3%
150 ppm
1 50-NS
1 50-PS
1 50-OS
*f 42.4%
No Significant Effect
*f 59.2%
*i -30.3%
*i -25.7%
No Significant Effect
No Significant Effect
*f 90.7%
*f 78.5%
No Significant Effect
*l -44.7%
No Significant Effect
aBased on calculation of group mean values across session block post-stress challenge for both maternal and lifetime Pb exposure
studies. All calculations represent percent of 0-NS control values; f, represents increase; J,, represents decrease.
"Data from Virgolini et a I. [(2005), as reported by Rossi-George et al. (2011)1.
°Data from current study, Rossi-George et al. (2011)
Notes: PRP = Post-reinforcement pause; PS = Prenatal (maternal stress); OS = Offspring stress.
'Denotes significant effect versus 0 ppm Pb, 0 stress control (p <0.05).
Results demonstrated that lifetime 50 ppm Pb exposure (right columns) alone and with stress induced learning deficits in female rats
at age 2 months as demonstrated by an increase in Fl overall rate. 150 ppm Pb increased overall rate only with stress co-exposure.
50 and 150 ppm Pb decreased PRP only with stress co-exposure. Mechanistically, these authors proposed that associations of Pb
and stress with learning deficits may be related to aberrations in corticosterone and dopamine. Prenatal Pb exposure alone (left
columns) induced learning impairments at 150 ppm but not 50 ppm.
Source: Reprinted with permission of Elsevier Science, Table 1 of Rossi-George et al. (2011).
A separate investigation from the same laboratory similarly indicated a potentiation of
effects in females with Pb and stress co-exposures but with developmental Pb exposure
from two months prior to mating through lactation (50 or 150 ppm Pb acetate in drinking
water) (Virgolini et al.. 2008a). Dams were subject to restraint stress at GD16-GD17.
Prenatal stress or Pb exposure alone did not affect Fl performance in offspring.
Compared with controls, marked increases in response rates on Fl performance were
found in female offspring at age 2-3 months with 50 ppm Pb plus prenatal and offspring
stress. Using the same Pb exposure protocol, Virgolini et al. (2008b) expanded evidence
for Pb-stress interactions with the examination of the effects of additional adult
intermittent stress on Fl performance, corticosterone, and dopamine. Adult stress was
induced by placing rats in a cold 4°C room for 30 minutes or in restraints or by subjecting
rats to previously unexperienced locomotor activity (novelty stress). In female offspring,
adult stress induced by restraint or cold decreased overall Fl response rate and increased
PRP, i.e., decreased impulsivity. These effects were attenuated with 50 and 150 ppm Pb
4-102
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(Figure 4-4. Panel A). In male offspring, restraint stress increased overall FI response
rates and run rate. These effects were attenuated with the lower 50 ppm Pb dose (Figure
4-4. Panel B). In males, 150 ppm Pb increased overall FI response rates and run rates
over those found at baseline and with stress induced by exposure to cold temperature.
A. B.
Restraint Cold Novelty
Restraint Cold Novelty
Qveratf Rate
200-i ISO-, \ 120-1
III ll i||
0 50 150 0 50 150 0 50 150
Run Kate
250 150-, +120-1
hi II II
0 50 150 0 SO 150 0 60 150 0 50 150 0 50 150 0 50 150
Pastreinforcement Pause Time Postreinforameni Pause TVme
500 -I 300-1 125-j 150"| 15Q-I 150- _!_
L. L. ll Ik ll ill
0 -^^™^T^ D SO 150 D M 130 0 50 150
0 50 ISO 0 50 130 D SO 150
Note: * denotes significantly different from 0 ppm Pb. ~denot.es marginally different from 0 ppm Pb, e.g., p = 0.07. + denotes
significantly different from the 50 ppm Pb group. Each column presents results for a particular stressor (placement in restraint, cold
room, novelty) given in adulthood. In females (panel A), gestational/lactational 50 and 150 ppm Pb exposures attenuated the
decreases in overall FI rate and increase in post-reinforcement-pause time induced by restraint or cold stress. In males (panel B),
50 ppm gestational/lactational Pb exposure attenuated the increase in FI overall rate and run rates induced by restraint stress.
Source: Reprinted with permission of Elsevier Science, Virgolini et al. (2008b).
Figure 4-4 Changes in Fixed Interval performance in (A) female and (B) male
offspring with gestational/lactational Pb exposure plus various
stressors given in adulthood.
Pb exposure over various developmental windows in rodents has been shown to affect the
HPA axis, as measured by the levels of corticosterone, the major glucocorticoid involved
in stress responses in rodents. Thus, modulation of corticosterone may provide a
4-103
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mechanistic explanation for learning deficits (FI testing in females) found with Pb and
stress co-exposure in rodents. As examined by the Cory-Slechta laboratory, the effects
exposure to Pb on corticosterone levels varied, depending on sex of the animal, age of the
animal at assessment, and timing and duration of exposure, i.e., developmental
(gestational-lactational), post-weaning, or lifetime (Rossi-George et al., 2011; Cory-
Slechta et al.. 2010: Virgolini et al.. 2008a).
Because animals that are used for FI testing are regularly handled by laboratory personnel
and often participate in other tests of cognition, their baseline level of stress may be
skewed from that of a laboratory animal that constantly remains in a cage without daily
handling. Because effects on the HPA axis are of interest to Pb researchers, the baseline
corticosterone levels of animals that have participated in behavior testing (FI) and those
who have not (NFI) have been compared after gestational/lactational Pb exposure.
Virgolini et al. (2008a) found that baseline corticosterone levels were higher in FI than
NFI animals. Also, the effect of combined gestational/lactational Pb exposure plus
maternal stress on corticosterone was compared in FI and NFI animals. At the baseline
age of 4-5 months, Pb exposure with or without stress did not induce differences in
corticosterone levels in FI females but did in males (Virgolini et al., 2008a). In the FI
males, 50 ppm Pb exposure decreased corticosterone versus control (no Pb exposure),
and 150 ppm Pb exposure elevated corticosterone versus control. In male NFI animals, a
U shaped concentration-response was found, with 50 ppm Pb exposure reducing
corticosterone over than in the controls or with 150 ppm Pb exposure. In the NFI males,
stress did not affect corticosterone levels or interact with the effect of Pb exposure. NFI
females exposed to 150 ppm Pb had significantly elevated corticosterone versus control
(no Pb exposure). These data demonstrate that behaviorally tested animals have altered
HPA axis and altered responses to Pb exposure versus animals that are housed under
conditions without daily handling by caregivers. In animals given stress as adults and not
behaviorally tested, gestational-lactational Pb co-exposure prevented the stress-induced
increase in corticosterone levels in females but not males, which may explain results in
this study for Pb plus stress increases in FI response rates in females (Virgolini et al..
2008b).
Lifetime Pb exposure beginning in gestation (50 or 150 ppm drinking water of dams from
2 months prior to mating through lactation, then continuing in offspring water) induced
increases in basal (age 2 months, before behavioral testing) corticosterone only in female
offspring (Rossi-George et al., 2011) (Figure 4-5). This Pb-related increase in
corticosterone at age 2 months was found in animals with blood Pb levels of 12 and 25-
32 (ig/dL measured at PND21. Pb-stress (either prenatal or prenatal plus offspring)
interactions were observed at age 2 months but not 10 months (final, after behavioral
testing). At age 2 months, Pb plus stress attenuated the Pb-induced elevations in
4-104
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900i
600-
c
300-
corticosterone to baseline levels (Figure 4-5). At 10 months of age, female offspring
exposed to Pb alone or Pb plus stress had lower corticosterone concentrations than
control animals. In males, corticosterone levels were not affected significantly by Pb,
stress, or combined Pb/stress at 2 (basal) or 10 (final) months of age (Figure 4-5) (Rossi-
George etal.. 2011). In another study of lifetime Pb exposure and prenatal stress, stress
alone reduced corticosterone in animals behaviorally tested at age 4 months, and Pb co-
exposure blunted this decrease. In animals at age 11 months not behaviorally tested, only
stress induced a decrease in corticosterone, and no interaction was found with Pb (Cory-
SlechtaetaL 2010). Pb-stress interactions were found in male rats exposed to Pb
postweaning (PND21-age 5 months) and to stress as adults. Males exposed to 50 or 150
ppm Pb plus stress induced by placement in a cold room had higher corticosterone levels
than rats exposed to Pb alone or stress alone (Virgolini et al., 2005).
Female
Male
Basal
Final
PbxS
900
Basal
Final
600
300
G 50 150 D 50 150
Pb Exposure (ppm)
0 50 150 0 SO ISO
Pb Exposure (ppm)
Note: Corticosterone levels are noted on the y-axis. 'denotes 50-NS (50 ppm Pb) or 150-NS (150 ppm Pb) is significantly different
from 0-NS (no Pb, no stress) control; # denotes significantly different from corresponding Pb-NS value; + denotes significantly
different from 50-NS. PS refers to prenatal dam stress, white bar with black dots. OS refers to PS followed by offspring stress, black
bar with white dots. Basal measurements were taken at age 2 months, prior to the initiation of behavioral testing and final
measurements were taken at 10 months, after behavioral testing. Effects of Pb and stress were observed in females not males and
differed between age 2 and 10 months. At age 2 months, Pb exposure alone increased corticosterone but attenuated the prenatal
stress-induced increase in corticosterone. At age 10 months, Pb exposure alone, stress exposure alone (prenatal, or prenatal- plus
postnatal), and combined Pb-stress exposure decreased corticosterone.
Source: Rossi-George et al. (2011).
Figure 4-5 Mean basal and final corticosterone levels of female and male
offspring exposed to lifetime Pb.
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Another study examined the effects of gestational-lactational Pb acetate on the HPA axis
but examined the interaction with maternal restraint stress (GD16-GD17) and offspring
stress induced by injections of saline control or dexamethasone (DEX) (Rossi-George et
al.. 2009). Saline injection induced a spike in corticosterone followed by a time-related
decline. Pb delayed and attenuated the decline in a nonlinear concentration-response in
both sexes with the most profound effects found at the lower 50 ppm Pb dose. Maternal
stress also prolonged the corticosterone response to saline injection and enhanced the
effect of Pb in males. To test the negative feedback of the HPA axis, exogenous
dexamethasone (DEX) was administered to suppress endogenous corticosterone. Pb and
Pb plus maternal stress initially reduced the ability of DEX to suppress corticosterone,
indicating HPA axis hypofunction. In males, with time, Pb and Pb plus stress prolonged
the DEX-induced corticosterone suppression or inhibited the return to the levels in the
control animals (no Pb/no stress). Rossi-George et al. (2011) additionally found that
lifetime Pb significantly impacted the negative feedback by increasing nuclear
glucocorticoid receptor levels. Maternal or offspring stress reduced the Pb effect in
females and increased the effect of Pb in males.
To summarize the results for Pb-stress interactions in animals, Pb exposure when
combined with prenatal maternal or offspring stress was found to exacerbate learning
impairments compared with Pb exposure or stress alone (Rossi-George et al., 2011; Cory-
SlechtaetaL 2010; Virgolini et al.. 2008a). although not uniformly across all tests or Pb
doses or exposure periods. The exacerbation of learning impairments as indicated by
decreases in PRP and increases in the overall response rate in performance on FI tests
was found more consistently with lifetime (beginning in gestation) Pb exposure than with
only gestational-lactational Pb exposure (Table 4-7). The interaction between Pb and
stress may be mediated via effects on corticosterone and dopamine. Pb and stress, each
given alone, have induced changes in corticosterone but not in the same direction across
studies. Some but not all studies have demonstrated Pb-stress interactions on
corticosterone but not always in the same direction. Gestational-lactational and lifetime
Pb exposure were found to blunt the stress-induced increase in corticosterone in female
rats (Cory-Slechta et al.. 2010; Virgolini et al.. 2008b). Maternal prenatal and offspring
stress were found to blunt the increase in corticosterone induced by lifetime Pb exposure
(Rossi-George et al.. 2011). Adult stress was found to enhance the increase in
corticosterone in males induced by postweaning Pb exposure (Virgolini et al., 2005). The
effects of Pb and stress on corticosterone also have differed by sex. A few studies have
shown HPA axis negative feedback hypofunction with gestational-lactational and lifetime
Pb exposure, with the direction of Pb-stress interactions differing between males and
females (Rossi-George et al., 2011; 2009). The combined results show that HPA axis
4-106
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alterations could provide a link for interactions found between Pb and stress in impairing
learning.
4.3.2.4 Executive Function in Children
Epidemiologic evidence presented in the 2006 Pb AQCD (U.S. EPA. 2006b) indicated
associations between higher childhood blood or tooth Pb levels and poorer performance
on tests of executive function in children and young adults. Associations were found with
indices of executive function such as strategic planning, organized search, flexibility of
thought and action to a change in situation, and control of impulses (described in greater
detail in Section 4.3.3.1). Prospective analyses in two Boston area and the Rochester
cohorts provided key evidence with examination of blood Pb levels preceding executive
function testing and adjustment for several potential confounding factors (Canfield et al..
2004: 2003b: Bellinger et al.. 1994a: Stiles and Bellinger. 1993). Further, recruitment of
participants before or at birth, moderate to high follow-up participation, and in most cases
follow-up not biased to those with higher blood Pb levels and lower cognitive function
increase confidence that the observed associations are not due to selection bias (Table
4-8). Among the few recent cross-sectional studies, most found concurrent blood
Pb-associated decrements in executive function, including an analysis of the Rochester
cohort (Froehlich et al., 2007). Evidence from other recent studies had weaker
implications due to the limited consideration of potential confounding (Nelson and Espy.
2009; Vega-Dienstmaier et al., 2006). Several of the prospective and cross-sectional
studies performed multiple tests of cognitive function, including executive function.
However, except for Stiles and Bellinger (1993). a consistent pattern of association was
found across the various tests performed. Thus, the evidence does not appear to be biased
by associations found by chance alone.
Studies in children found Pb-associated decreases in executive function using various
tests including the Intra-Extra Dimensional Set Shift, WCST, and Stroop test (Table 4-8).
As discussed below, studies in animals also demonstrated Pb-induced decrements in
executive function, including rule learning and reversal, which also was associated with
blood Pb levels in children. This coherence between findings in animals and humans for
analogous outcomes further supports a relationship between Pb exposure and decrements
in executive function. In both humans and animals, tests of executive function measure
the ability of subjects to adjust their responses in reaction to changes in reinforcement.
Poorer performance in both children and animals is indicated by more response errors,
fewer correct responses, and perseverative responding (e.g., persistence in making a
previously-rewarded response after anew shift in reinforcement). Biological plausibility
for Pb-associated decrements in executive function also is provided by toxicological
4-107
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evidence for Pb-induced changes in the availability of dopamine (Section 4.3.10.8). a
neurotransmitter that affects executive functions mediated by the prefrontal cortex.
Recent work unrelated to Pb shows that executive function in animals is affected by N-
Methyl-D-aspartic acid or N-Methyl-D-aspartate (NMDA) receptors and dopamine-like
receptors (Herold. 2010). which are two well-characterized targets of Pb.
In the Boston cohort at a relatively older age (10 years), when testing is more reliable, a
1 (ig/dL increase in age 5 year blood Pb level was associated with 0.05 (95% CI: 0.01,
0.09) more perseverative errors on the WCST (errors in sorting cards according to a
change in rule) (Stiles and Bellinger. 1993). In this cohort, results were inconsistent
across the various cognitive tests. However, associations were more consistent for
executive functions assessed by the WCST. In another cohort ages 19-20 years from
towns near Boston, higher tooth Pb levels (from lst/2nd grade) were associated with
more errors on the WCST in sorting by the set rules and poorer performance on the
Stroop Color and Color-word tests, which test the ability of subjects to shift focus to
another dimension of stimulus that defines correct responding (Bellinger et al.. 1994a).
Results from the Rochester cohort at ages 4 and 5 years indicated associations of
concurrent and lifetime average blood Pb level with lower inhibition efficiency in the
Shape School task (i.e., giving correct responses and withholding incorrect responses)
(Canfield et al.. 2003b). poorer problem solving on a spatial planning task (Canfield et
al.. 2004). and poorer rule learning and reversal (Froehlich et al.. 2007). Associations
with Shape School tasks decreased in magnitude and precision with adjustment for
attention ratings, color/shape knowledge, and child IQ. These results suggest that the
effect of Pb exposure on executive function may be mediated through effects on
knowledge or attention. Froehlich et al. (2007) found a larger concurrent blood
Pb-associated decrement in performance on a rule learning and reversal task in the Intra-
Extra Dimensional Set Shift in children without the DRD4 exon III 7-repeat
microsatellite (assessed using a blood Pb-DRD4-7 interaction term, p = 0.042). These
results were unexpected, given evidence that presence of the DRD4-7 repeat is associated
with reduced dopamine-induced signaling in downstream pathways (e.g., cyclic AMP),
and is associated with poorer executive function in this cohort. As the results for effect
modification were based on a smaller subset of subjects (n = 116/174), they could be due
to chance. The association of concurrent blood Pb level with impaired rule learning and
reversal also was greater in boys, who had lower mean scores than girls.
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Table 4-8 Associations between blood or tooth Pb levels and performance on tests of executive function in
children and young adults.
Study
Stiles and
Bellinger
(1993)
Bellinger
etal.
(1994a)
Canfield et
al. (2004)
Study Population and Methodological Details
Presented first for prospective studies then for cross-sectional studies. Within
each category, results are grouped by strength of study methodology,
generalizability3
134 children followed from birth (1979-1981) to age 10 yr, Boston, MA
area
Prospective. Same cohort as above. Moderate follow-up
participation, participants had higher SES and HOME score. Linear
regression model adjusted for HOME score, maternal IQ, SES, child
stress, sex and birth order, age 10 yr HOME score, family stress,
marital status, race (all ages). Plus: birth weight, maternal age,
# daycare situations to age 57 mo (concurrent).
79 young adults followed from 1st-2nd grade (1975-1978) to ages
19-20 yr, Boston, MA area
Prospective. Low follow-up participation. Participation from higher
SES and initial IQ and nonwhite, but no effect on association with
tooth Pb level. Log linear regression model adjusted for parental IQ,
sex, SES, current drug, alcohol and illicit drug use, maternal education
and age, birth order. Also considered potential confounding by other
unspecified factors.
124-151 children born 1994-1995 followed from age 6 mo to 5 yr,
Rochester, NY
Prospective. Recruitment from study of dust control. Mostly nonwhite.
High follow-up participation, no selective attrition. Linear regression
model adjusted for HOME score, maternal prenatal smoking,
household income, child sex (both outcomes). Plus: NICU admission
and average crowding in home (IED); maternal IQ (Stockings of
Cambridge). Also considered potential confounding by breastfeeding
duration, maternal ethnicity, first prenatal visit and education, spatial
working memory problem, age attesting, birth weight, marital status,
spatial span length.
Blood Pb Metric
Analyzed (ug/dL)
Earlier childhood
(age 57 mo): NR
Concurrent: NR
All means < 8
Deciduous tooth
(1st/2nd grade)
Mean (SD): 14
(11)M9/9
10th-90th: 4.3-26
Lifetime (to age 5 yr)
avg
Mean (SD): 7.2(3.6)
10th-90th: 3.5-11.8
Executive Function Effect Estimate
Measure Analyzed (95% Cl)b
Perseverative errors, -0.05 (-0.09, -0.01)c
WCST
AgelOyr -0.05 (-0.11, 0.01 )c
Time to complete color- -0.68 (-1.1, -0.28)c
word test, Stroop test
Perseverative responses, -0.37 (-0.64, -0.1 0)c
WCST
Ages 19-20yr
Stages Completed - IED -0.11 (-0.21, -0.01)
Shift Task
Stockings of Cambridge -0.08 (-0.17, 0)
problems solved in
minimum moves
CANTAB
Age 5 yr
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Table 4-8 (Continued): Associations between blood or tooth Pb levels and performance on tests of executive function in
children and young adults.
Study
Study Population and Methodological Details
Presented first for prospective studies then for cross-sectional studies. Within
each category, results are grouped by strength of study methodology,
generalizability3
Blood Pb Metric
Analyzed (ug/dL)
Executive Function
Measure Analyzed
Effect Estimate
(95% Cl)b
Froehlich 174 children born 1994-1995 followed from age 6 mo to 5 yr,
et al. Rochester, NY
(2007) Cross-sectional. Same cohort as above. Moderate follow-up
participation, no selective attrition. Linear regression model adjusted
for NICU, sex. Also considered potential confounding by income,
HOME score, transferrin saturation, maternal IQ, education, prenatal
smoking status, race, and age
Concurrent
Mean (SD): 6.1 (4.9)
10th-90th: 1.9-11.7
Stages Completed - IED
Set Shift
CANTAB
Age 5 yr
-0.06 (-0.12, 0)
Canfield et 96-126 children born 1994-1995 followed from age 6 mo to 4-4.5 yr,
al. (2003b) Rochester, NY
Cross-sectional. Same cohort as above. Moderate follow-up
participation, no comparison of nonparticipants. Linear mixed effects
model adjusted for child sex and age, maternal IQ, education, marital
status, and prenatal smoking, household income, HOME score. Also
considered potential confounding by birth order, race, gestational age,
attention rating, color/shape knowledge, age 3 FSIQ.
Age 4 yr
Mean: 6.5 Range:
1.7-21
10th-90th: data not
available
Inhibit Efficiency
(# correct-incorrect)/ phase
duration)
Shape School Task
Repeated measures at
ages 4 and 4.5 yr
-0.02 (-0.03, -0.007)
Surkan et 389 children, ages 6-10 years, Boston, MA, Farmington, ME
al. (2007) Cross-sectional. Recruitment from trial of amalgam fillings. High
participation rate. Higher participation of white children in Maine.
Analysis of covariance adjusted for caregiver IQ, child age, SES, race,
birth weight. Also considered potential confounding by site, sex, birth
order, caregiver education and marital status, parenting stress, and
maternal utilization of prenatal and annual health care but not parental
caregiving quality.
Concurrent
Group 1:1-2
Group 2: 3-4
Group 3:5-10
Group 1:1-2
Group 2: 3-4
Group 3:5-10
Perseveration errors,
WCST
Stroop color-word
interference score
Reference
-1.7 (-5.3, 2.0)c'e
-9.2 (-15, -3.8)c'e
Reference
-0.33 (-1.9, 1.2)e
0.75 (-1.6, 3.1)e
Chiodo et 48 children, age 7.5 yr, Detroit, Ml area
al. (2004) Cross-sectional. Recruitment at prenatal clinic. 100% African-
American High prevalence of prenatal alcohol and drug exposure.
Moderate participation rate. Log linear regression model adjusted for
SES, family functioning, # children <18 yr, caregiver vocabulary and
education, child sex, prenatal alcohol exposure. Also considered
potential confounding by HOME score, maternal prenatal marijuana,
smoking, or cocaine use, crowding, child life stress, caregiver age, life
stress, and psychopathology, conflict tactics, disruption in caregiver,
parity, child age.
Concurrent
Mean(SD): 5.4(3.3)
10th-90th: 2.2-9.5
Perseverative errors,
WCST
-0.49, p >0.05C
Percent errors
Age 7.5 yr
-0.74 (-1.5, 0)'
,c,d
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Table 4-8 (Continued): Associations between blood or tooth Pb levels and performance on tests of executive function in
children and young adults.
Study
Cho et al.
(2010)
Study Population and Methodological Details
Presented first for prospective studies then for cross-sectional studies. Within
each category, results are grouped by strength of study methodology,
generalizability3
639 children ages 8-11 yr, born 1997-2000, Seoul, Seongnam, Ulsan,
Incheon, Yeoncheon, Korea
Cross-sectional. School-based recruitment, moderate participation
rate. Log linear regression model adjusted for age, sex, paternal
education, maternal IQ, child IQ, birth weight, urinary cotinine,
residential area. Did not consider potential confounding by parental
caregiving quality.
Blood Pb Metric
Analyzed (ug/dL)
Concurrent
Mean(SD): 1.9
(0.67)
10th-90th: 1.2-2.8
Detection limit = 0.58
Executive Function
Measure Analyzed
Color-word score
Stroop test
Ages 8-1 1 yr
Effect Estimate
(95% Cl)b
0 (-0.02, 0.02)
WCST = Wisconsin Card Sorting Test, IED = Intra-Extra Dimensional, CANTAB = Cambridge Neuropsychological Test Automated Battery.
aResults are presented first for prospective studies then for cross-sectional studies. Within each category, results are grouped by strength of study design, representativeness of the
population characteristics and blood Pb levels examined, and extent of consideration for potential confounding. There is not necessarily a continuum of decreasing strength across
studies.
bEffect estimates are standardized to a 1 ug/dL increase in blood Pb level or 1 ug/g in tooth Pb level in the 10th-90th percentile interval.
°The direction of the effect estimate was changed such that a negative estimate represents poorer performance.
d95% Cl was constructed using a standard error that was estimated from the reported p-value.
eEffect estimates compare test performance of children in higher blood Pb groups to children in lowest blood Pb group.
'Sufficient data were not available to calculate 95% Cl.
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In addition to assessment of early childhood or cumulative Pb biomarkers, a strength of
the prospective studies was the consideration for numerous potential confounding factors.
The potential confounding factors varied among studies based on their association with
executive function and/or influence on the Pb-executive function relationship. Some
prospective studies demonstrated Pb-associated decrements in executive function with
adjustment for SES, maternal IQ, and HOME score (Canfield et al.. 2004; Stiles and
Bellinger. 1993). Others considered and excluded potential confounding by parental
caregiving quality, parental smoking, maternal education, race, or birth outcomes (Table
4i8).
Blood Pb levels were reported for the Rochester cohort, and blood Pb-associated
decrements in executive function were found in the cohort with mean blood Pb levels of
7.2 (ig/dL for lifetime average (to age 5 years) and 6.1 and 6.5 (ig/dL for concurrent (ages
4-5 years). Associations in populations with lower mean blood Pb levels (1.9-5.4 (ig/dL),
as assessed in cross-sectional studies with concurrent blood Pb level, were not as clearly
demonstrated. While these studies adjusted for SES and parental cognitive function, most
did not examine potential confounding by parental caregiving quality, i.e., HOME score.
Chiodo et al. (2004) found a concurrent blood Pb-associated decrease in executive
function in children ages 7 years in Detroit, MI with examination of several potential
confounding factors. While not affecting internal validity, the high prevalence of prenatal
alcohol exposure in the population may limit the generalizability of results. Among
children in New England ages 6-10 years with mean concurrent blood Pb level 2.2 (ig/dL,
poorer performance on the WCST, Trail-making, and Verbal Cancellation tests was
found primarily in the group with blood Pb levels 5-10 (ig/dL (blood Pb group 1-2 (ig/dL
as reference). In this study, higher blood Pb level was not associated with poorer color-
word score in the Stroop test. Cho et al. (2010) did not find a Pb-associated lower color-
word score among children ages 8-11 years in five Korean cities with mean concurrent
blood Pb level 1.9 (ig/dL (detection limit 0.58 (ig/dL). Nelson et al. (2009) found poorer
executive function at age 2-6 years in children with a mean concurrent blood Pb of 6
(ig/dL (compared to 2.5 (ig/dL) but did not consider potential confounding and only
reported F-statistics for group comparisons.
The associations observed in children between blood or tooth Pb levels and poorer
executive function as assessed by the rule learning and reversal components of the Intra-
Extra Dimensional Set Shift, Stroop Test, and WCST are supported by observations in
animals of Pb-induced impairment in analogous measures of cognitive flexibility, tested
with discrimination reversal learning and concurrent random interval (RI-RI) scheduling.
The animal evidence for Pb-induced impairments in executive function was reviewed in
the 2006 Pb AQCD (U.S. EPA. 2006b). and some of these observations were made with
Pb exposures relevant to this ISA. Lifetime dietary Pb exposures beginning at birth or
4-112
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later in infancy that produced blood Pb levels of 19-36 (ig/dL were found to induce
poorer performance on discrimination reversal learning tasks in monkeys ages 5-10 years
(Rice and Gilbert, 1990b; Gilbert and Rice. 1987). Recent work unrelated to Pb has
shown that discrimination reversal learning involves NMDA receptors and dopamine-like
receptors (Herold. 2010). which are two well-characterized targets of Pb. Gestational Pb
exposure (blood Pb of dams >40 (ig/dL) was found to impair cognitive flexibility in
squirrel monkeys, ages 5-6 years, as indicated by a slower shift or lack of a shift to the
lever reinforced more frequently under RI-RI scheduling (Newland et al.. 1994). Rats
also showed Pb-induced impairments in discrimination reversal tasks, but the authors
attributed the changes to poorer associative ability instead of impaired executive function
(Garavan et al.. 2000: Hilson and Strupp. 1997).
4.3.2.5 Academic Performance and Achievement in Children
As described in preceding sections, a large body of evidence demonstrates Pb-associated
decrements in FSIQ in children, with more variable findings for performance on tests of
learning and memory. Lower FSIQ and learning are linked with poorer academic
performance and achievement, which may have important implications for success later
in life. Further, academic performance may better assess the knowledge of an individual
in the actual subject areas studied, whereas aptitude tests are used to predict future
performance. In addition to FSIQ, the 2006 Pb AQCD described associations of higher
blood and tooth Pb levels in children ages 5-18 years with poorer performance on tests of
math, reading, and spelling skills, lower probability of high school completion and lower
class rank, and lower teacher ratings of academic functioning. Associations continued to
be reported in recent studies, including prospective studies examining performance on
academic achievement tests and an additional analysis of adolescents participating in
NHANES (Table 4-9). Multiple testing was common in studies; however, the consistent
pattern of blood Pb-associated decrements in academic performance across the various
tests conducted increases confidence that the evidence is not unduly biased by a higher
probability of associations found by chance alone.
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Table 4-9 Associations between blood or tooth Pb levels and measures of academic performance and
achievement in children and young adults.
Study
Study Population and Methodological Details
Organized by outcome then strength of study methodology,
generalizability3
Blood (ug/dL) or
Tooth Pb Metrics
Analyzed
Measure of
Academic
Performance
or Achievement
Analyzed
Effect Estimate
(95% Cl)b
Studies of tests of academic achievement
Chandramouli 488 children followed from age 4 mo (born 1992) to age 7 yr,
et al. (2009) Avon, U.K.
Prospective. All births in area eligible. Similar characteristics
as U.K. census, high participation at baseline and follow-up.
Participants had better educated mothers, who smoked less,
better home environment. Log linear regression model
adjusted for maternal education and smoking, home
ownership, home facilities score, family adversity index,
paternal SES, parenting attitudes at 6 mo, child sex. Also
considered potential confounding by child IQ.
Earlier childhood
(age 30 mo)
Mean (SD): NR
Standardized
Achievement
Test
Age 7 yr
Per doubling blood Pb
-0.3 (-0.5, -0.1)
Miranda et al. 57,568 children, 4th grade, all counties, NC
(2009) Prospective. Based on data from surveillance databases.
39% state test scores linked to blood Pb data. Regression
model adjusted for sex, age of blood Pb measurement, race,
enrollment in free/reduced lunch program, parental education,
charter school. Did not consider potential confounding by
parental caregiving quality or cognitive function.
Earlier childhood
(9-36 mo)
Median: 4.8
25th-75th: 3-6
Detection limit = 1
Reading
4th grade end-of-
grade test score
Score vs. blood Pb category 1 ug/dL
2 ug/dl_: -0.30 (-0.58, -0.01 )c
3 ug/dl_:-0.46 (-0.73, -0.19)c
4 ug/dl_: -0.52 (-0.79, -0.24)c
5 ug/dl_: -0.80 (-1.08, -0.51 )c
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Table 4-9 (Continued): Associations between blood or tooth Pb levels and measures of academic performance and achievement
in children and young adults.
Study
Min et al.
(2009)
Lanphear et al.
(2000)
Krieg et al.
(2010)
Study Population and Methodological Details
Organized by outcome then strength of study methodology,
generalizability3
267 children followed from birth (1994-1996) to age 11 yr,
Cleveland, OH
Prospective. Recruitment at birth hospital. 86% African-
American with high prevalence of prenatal drug and alcohol
exposure. Moderate follow-up participation to age 4 yr, high
retention to age 1 1 yr. Participants tended to be African-
American and had married mothers. Linear regression model
adjusted for HOME score, maternal birth vocabulary score,
head circumference at birth, prenatal cocaine use (both
outcomes), child sex, maternal prenatal alcohol use, current
caregiver alcohol use. Also considered potential confounding
by maternal education, Fe deficiency, maternal psychological
distress, race.
4,853 children ages 6-16 yr (born 1972-1988), U.S. NHANES
III, 1988-1994
Cross-sectional. Large U.S. representative study of multiple
risk factors and outcomes. Linear regression model adjusted
for sex, race/ethnicity, poverty index ratio, reference adult
education, serum ferritin and cotinine levels. Did not consider
potential confounding by parental cognitive function or
caregiving quality.
766-780 children ages 12-16 yr (born 1975-1982), U.S.
NHANES III, 1991-1994
Cross-sectional. Large U.S. representative study of multiple
risk factors and outcomes. Log linear regression model
adjusted for sex, caregiver education, family income,
race/ethnicity, test language. Did not consider potential
confounding by parental cognitive function or caregiving
quality.
Blood (ug/dL) or
Tooth Pb Metrics
Analyzed
Earlier childhood
(Age 4 yr)
Mean (SD): 7.0
(4.1)
Interval analyzed:
3.0 (10th
percentile)-10
Concurrent
Geometric mean
(SD): 1.9(7.0)
63.5% < 2.5
Ustsction limit —
0.5
Interval analyzed:
1-5
Concurrent
Mean (5th-95th):
1.95(1.63-2.27)
Detection limit = 1
Interval analyzed:
-1700 —
1 . / -Ł. . Z. —
10th-90th
percentiles
Measure of
Academic
Performance
or Achievement
Analyzed
Math
Reading
\ A / i~r A
WJTA
Age 1 1 yr
Math score
Reading score
\ A /I"") A T D
WRAT-R
Ages 6-1 6 yr
Math score
Reading score
\ A /I"") A T" D
WRAT-R
Ages 12-16 yr
Effect Estimate
(95% Cl)b
-0.45 (-0.84, -0.06)
-0.60 (-1.0, -0.19)
-0.70 (-1.0, -0.37)
-0.99 (-1.4, -0.62)
-2.5 (-4.6, -0.38)
-2. 9 (-4. 4, -1.4)
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Table 4-9 (Continued): Associations between blood or tooth Pb levels and measures of academic performance and achievement
in children and young adults.
Study
Surkan et al.
(2007)
Kordas et al.
(2006)
Chiodo et al.
(2007)
Study Population and Methodological Details
Organized by outcome then strength of study methodology,
generalizability3
389 children, ages 6-10 years, Boston, MA, Farmington, ME
Cross-sectional. Recruitment from trial of amalgam fillings.
High participation rate. Higher participation of white children in
Maine. Analysis of covariance adjusted for caregiver IQ, child
age, SES, race, IQ, birth weight. Also considered potential
confounding by site, sex, birth order, caregiver education and
marital status, parenting stress, and maternal utilization of
prenatal and annual health care but not parental caregiving
quality.
293 children, ages 6-7 yr, Torreon, Mexico.
Cross-sectional. Recruitment at schools. High participation
rate. Residence near metal foundry. Linear regression model
adjusted for child sex, age, school, birth order, hemoglobin,
forgetting homework, household possessions and crowding,
home ownership, maternal education, family structure, urinary
arsenic, tester, school. Did not consider potential confounding
by parental cognitive function or caregiving quality.
330-331 children, age 7yr(born 1989-1991), Detroit, Ml area.
Cross-sectional. Recruitment at prenatal clinic. 100%
African-American. High prevalence of prenatal drug exposure.
High participation rate. Linear regression model adjusted for
caregiver education, SES, HOME, maternal IQ and prenatal
marijuana use, child sex (both outcomes). Caregiver
concurrent psychological symptoms (Math), child age,
maternal custody (Reading). Also considered potential
confounding by prenatal cigarettes/day, alcohol use, cocaine
use, # children in home, caretaker marital status, concurrent
alcohol/week, current maternal cigarettes/day, and current
marijuana use.
Blood (ug/dL) or
Tooth Pb Metrics
Analyzed
Concurrent
Group 1:1-2
Group 2: 3-4
Group 3: 5-10
Group 1:1-2
Group 2: 3-4
Group 3: 5-10
Mean (SD): 2.2
(1.6)
Concurrent
Geometric mean
(SD): 7.1 (1.7)
Interval analyzed:
2-10
Concurrent
Mean (SD): 5.0
(3.0)
Interval analyzed:
2.1-8.7 =
10th-90th
percentiles
Measure of
Academic
Performance
or Achievement Effect Estimate
Analyzed (95% Cl)b
Reading score Reference
-0.93 (-3.8, 1.9)c
-5.2 (-9.5, -0.95)c
Math score Reference
WIAT 1.6 (-0.79, 4.0)c
Ages 6-1 0 yr -4.0 (-7.6, -0.43)c
Math -0.42 (-0.92, 0.08)
achievement test
Ages 6-7 yr
Math -0.1 7 (-0.27, -0.07)d
Reading -0.06, p>0.05e
Metropolitan
Aptitude Test
Age 7 yr
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Table 4-9 (Continued): Associations between blood or tooth Pb levels and measures of academic performance and achievement
in children and young adults.
Study
Chiodo et al.
(2004)
Study Population and Methodological Details
Organized by outcome then strength of study methodology,
generalizability3
237 children, age 7.5 yr, Detroit, Ml area
Cross-sectional. Recruitment at prenatal clinic. 100%
African-American. High prevalence of prenatal alcohol and
drug exposure. Moderate participation rate. Log linear
regression model adjusted for SES, HOME, child sex, prenatal
alcohol exposure, caregiver education and vocabulary. Also
considered potential confounding by child age and life stress,
maternal prenatal marijuana, smoking, or cocaine use,
crowding, caregiver life stress, family functioning, # children
0.05e
Studies of School Performance
Fergusson et
al. (1997)
Needleman et
al. (1990)
881 children followed from birth to age 18 yr, Christchurch,
New Zealand
Prospective. Moderate follow-up participation, attrition did not
affect results. Regression model adjusted for maternal age,
punitiveness, standard of living, breastfeeding duration,
parental conflict, grade, residence on busy roads. Also
considered potential confounding by sex, ethnicity, maternal
education, family size, HOME, SES, ethnicity, parental
change, birth order, single parent.
132 young adults followed from 1st/2nd grade (1975-1978) to
age 18 yr, Chelsea, Somerville, MA
Prospective. Recruitment at schools. Low follow-up
participation. Participants had lower tooth Pb, higher parental
education, SES, maternal IQ. Participation status did not
influence results. Logistic regression adjusted for maternal
age at birth, education, and IQ, family size, SES, sex, age at
testing, birth order, alcohol use, mother and child left hospital
together. Did not examine potential confounding by parental
caregiving quality.
Tooth Pb (age 6-8
yr)
Mean (SD): 6.2
(3.7) ug/g
See Ferguson et
al. (1993)
Tooth Pb(1st/2nd
grade) distribution
<10 ppm: 50%
10-19.9 ppm:
22.7%
>20 ppm: 27.3%
Percent leaving
school without
school certificate
Ages 16-18 yr
Failure to
graduate high
school
Highest grade
achieved
0-2 ug/g: 15.6
3-5 ug/g: 16.7
6-8 ug/g: 18.1
9-11 ug/g: 19.7
12+ ug/g: 24.1
p < 0.05
7.4(1.4, 41 )d
OR >20 ppm vs. <10 ppm
-0.03 (-0.05, -0.01)
per natural log increase in tooth Pb
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Table 4-9 (Continued): Associations between blood or tooth Pb levels and measures of academic performance and achievement
in children and young adults.
Study
Study
Leviton
(1993)
Study Population and Methodological Details
Organized by outcome then strength of study methodology,
generalizability3
of teacher ratings of academic performance
et al. 1,923 children followed from birth (1979-1980) to age 8 yr,
Boston, MA area
Prospective. Recruitment from birth hospital. Moderate
participation at baseline and follow-up. Regression model
adjusted for single parent family, gestational age, # children in
home, maternal education, ethnicity, daycare in 1st 3 years.
Also considered potential confounding by other unspecified
factors.
Blood (ug/dL) or
Tooth Pb Metrics
Analyzed
Prenatal (cord)
Mean: 6.8
Tooth Pb (Age 6
yr) Mean: 3.3 ug/g
Measure of
Academic
Performance
or Achievement
Analyzed
Reading difficulty
BTQ, Age 8 yr
Prenatal (cord)
Tooth Pb
Effect Estimate
(95% Cl)b
RR (yes/no) per natural log increase
Girls: 1.7(0.9, 3.3)
Boys: 1.3(0.8,2.2)
Girls: 2.2 (1.1, 4.2)
Boys: 1.2(0.7,2.0)
WRAT-R = Wide Range Achievement Test-Revised, WJTA = Woodcock Johnson-Ill Tests of Achievement, WIAT = Wechsler Individual Achievement Test, WISC = Wechsler
Intelligence Scale for Children, BTQ = Boston Teacher Questionnaire.
"Results are organized by method of outcome assessment. Within each category, results are grouped by strength of study design, representativeness of the population characteristics
and blood Pb levels examined, and extent of consideration for potential confounding. There is not necessarily a continuum of decreasing strength across studies.
bEffect estimates are standardized to a 1 ug/dL increase in blood Pb level in the interval from the 10th percentile of blood Pb level to the 90th percentile or 10 ug/dL, whichever is
lower.
°Effect estimates compare test performance of children in higher blood Pb groups to children in lowest blood Pb group.
d95% Cl was constructed using a standard error that was estimated from the reported p-value.
Sufficient data were not available to calculate 95% Cl.
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Recent prospective studies found associations for early childhood blood Pb levels but did
not have blood Pb measurements available at other time periods for comparison. The
records-based analysis (Miranda et al., 2009; 2007a), recruitment at birth (Min et al.,
2009). multi-factorial nature, or high participation rate of recent studies (Chandramouli et
al., 2009) do not indicate a strong influence of selection bias. Miranda et al. (2009;
2007a) linked higher blood Pb levels measured at ages 0-5 years, as ascertained from a
surveillance database, with lower end-of-grade (EOG) test scores in 8,603 fourth grade
children in seven of the largest counties in North Carolina and then in 57,678 children in
the entire state. A strength of the analyses was the availability of individual-level data on
a large number of children representative of the North Carolina fourth grade population.
The large numbers of children with blood Pb levels 2-5 (ig/dL provided greater power to
examine the effects of Pb in the lower range of blood Pb levels. In each analysis, children
in the blood Pb level category of 2 (ig/dL had lower reading and math fourth grade EOG
scores (p < 0.05) compared with children in the blood Pb level category of 1 (ig/dL
(detection limit = 1 (ig/dL). The reference blood Pb group comprised values below the
detection limit and is subject to greater measurement error. This error likely is non-
differential, which would make it more difficult to find an association if one exists. In
linear models, the decrease in EOG score generally was monotonic across blood Pb
groups. Because these children were born in the early- to mid-1990s and blood Pb levels
were measured earlier in childhood, it is less likely that these children had much higher
past Pb exposures and blood Pb levels that could have influenced the observed
associations.
Due to the records-based study design, investigators had a smaller set of potential
confounding factors available than those considered in the prospective studies described
for other measures of cognitive function. Results were adjusted for sex, race, school-type,
age of blood Pb measurement, parental education, participation in a free or reduced lunch
program as a measure of SES, and in the analysis of seven North Carolina counties,
school district and daily use of a computer as a measure of a stimulating home
environment (Miranda et al., 2007a). While there may be no complete single measure of
SES or parental caregiving quality, the covariates examined in these analyses are not as
well characterized, and the results may be subject to some residual confounding.
In the statewide dataset, compared with children with an earlier (between ages 9 and
36 months) blood Pb level of 1 ng/dL, children with an earlier blood Pb level of 2 (ig/dL
had a 0.30-point lower (95% CI: -0.58, -0.01) fourth grade reading EOG score (Miranda
et al.. 2009). While the linear regression analyses assumed similar effects of Pb across the
distribution, quantile regression indicated differential effects across the EOG distribution
(Figure 4-6). Compared with linear regression, quantile regression is more robust in
response to outliers and can identify whether effects differ at the top and bottom tails of
4-119
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*
I
s
o
o
W
§
u
-8
•12
the outcome distribution rather than at the mean. Per unit increase in blood Pb level, there
was a greater decrease in EOG in lower tail of the EOG distribution than in the middle or
upper portions of the distribution (e.g., leftmost black bar versus rightmost black bar,
Figure 4-6). For example, an increase in blood Pb level from 1 to 10 (ig/dL was
associated with a 2.3-point decrease in EOG score in children in the 5th percentile of
EOG and a 0.8-point decrease in children in the 95th percentile of EOG score. These
findings indicated that children with the lowest EOG performance may be more affected
by Pb exposure. Using quantile regression, Miranda et al. (2009) also showed that while
indicators of lower SES such as lower parental education or enrollment in a school lunch
program accounted for larger decrements in EOG score, an increase in blood Pb level
was independently associated with EOG score decrements that were as large as 1 to 2
points.
Cummulative Deficit: Decrease in EOG scores by multiple risk factors
5% 10% m 20% 25% 30% 35% 40% 45% 50% 55% 60% 65% 70% 75% 80% 85% 90% 95%
t
Baseline (BLL=1 pg/dL, no school
lunch program, parents completed college)
Effect of reduction in parental education
from completed college
to only completed hign school
Income effect as indicated by
enrollment in school lunch program
> Effect of increased BLL from 1 to 5 ug/dL
Quantile
Note: Baseline score calculated for a hypothetical referent individual with a blood Pb level of 1 ug/dL, having parents who completed
college, and not enrolled in the school lunch program (i.e., model covariates = zero). Vertical bars indicate the decrease in EOG
score associated with blood Pb level and covariates in various percentiles (quantiles) of EOG score (lowest to highest, left to right).
An increase in early childhood blood Pb level was associated with a larger decrease in EOG score (larger black bars on left) among
children in the 5th and 10th percentiles of EOG score than children in the 90th and 95th percentiles (smaller black bars on right).
Source: Reprinted with permission of Elsevier Science, Miranda et al. (2009)
Figure 4-6 Greater reduction in End-of-Grade (EOG) scores for an increase in
blood Pb level in lower percentiles of the test score distribution.
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Similar to Miranda et al. (2009). Chandramouli et al. (2009) found associations between
early childhood blood Pb levels (age 30 months) and later academic performance
(Standard Assessment Test [SAT] at age 7 years). In this study of 488 children in the
U.K., who had similar sociodemographic characteristics as those found in the U.K.
census, a doubling of age 30 month blood Pb level was associated with a 0.3-point
decline (95% CI: -0.5, -0.1) in SAT score. Results were adjusted for maternal education
and smoking, home ownership, parental SES and several factors related to caregiving
quality including home facilities score, family adversity index, and parenting attitudes. In
analyses of blood Pb level groups, lower SAT scores were found in children with age
30 month blood Pb levels >5 (ig/dL. Children with blood Pb levels 2-5 (ig/dL did not
have lower SAT scores than children with blood Pb levels 0-2 (ig/dL.
Consistent with prospective studies, cross-sectional studies found associations between
higher concurrent blood Pb level and lower scores on tests of math and reading, including
large studies of children participating in NHANES. Cross-sectional studies considered
potential confounding by SES and caregiver education, but few considered parental
cognitive function or caregiving quality. Higher concurrent blood Pb level was associated
with lower math and reading scores among 4,853 children ages 6-16 years and 766-780
children ages 12-16 years participating in NHANES (Krieg et al.. 2010; Lanphear et al..
2000). The examination of multiple exposures and outcomes in NHANES increases
confidence that associations are not unduly influenced by selection bias related
specifically to Pb exposure. While mean concurrent blood Pb levels of these study
populations were low, ~2 (ig/dL, the influence of higher past Pb exposures on findings
cannot be excluded. Consistent with studies of FSIQ, Lanphear et al. (2000) found a
supralinear concentration-response relationship. A 1 (ig/dL increase in concurrent blood
Pb level was associated with a change in math score of-0.70 points (95% CI: -1.0, -0.37)
among all subjects and -1.1 points (95% CI: -2.0, -0.12) among the 4,043 children ages 6-
16 years with blood Pb levels <5 (ig/dL. A supralinear relationship also was indicated in
children age 7 years in Mexico living near a metal foundry by a larger blood
Pb-associated decrement in math score among children with concurrent blood Pb levels
<10 (ig/dL (Kordas et al.. 2006). These findings of nonlinearity based only on concurrent
blood Pb without regard to early childhood blood Pb are less certain because the
magnitude and timing of Pb exposures contributing to the associations are uncertain. In
contrast with these findings, among children ages 6-10 years in New England, lower
reading and math scores were related to higher concurrent blood Pb levels, i.e., blood Pb
levels 5-10 (ig/dL compared with blood Pb levels 0-2 (ig/dL (Surkan et al.. 2007).
Some studies found blood Pb-associated decreases math or reading test scores after
greater consideration for potential confounding factors, including SES, maternal
education, maternal intelligence, and parental caregiving quality. These associations were
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found in populations with high prevalence of prenatal alcohol or drug exposure (Min et
al.. 2009; Chiodo et al.. 2007; 2004). These exposures, either in initial analyses or by
inclusion in models with blood Pb, were not found to bias blood Pb results. Further, the
results were consistent with those in more representative populations.
Prospective studies in a Boston, MA area cohort and New Zealand cohort found
associations of tooth Pb levels measured at an earlier age (e.g., ages 6-8 years) with
school performance ascertained at ages 18 years from school records (Fergusson et al..
1997. 1993; Needleman et al.. 1990). suggesting the effect of early exposure to Pb may
be persistent. In the New Zealand cohort at ages 12-13 and 18 years, recruitment rate and
follow-up participation were high, and model correction for nonrandom sample attrition
produced robust results, indicating lack of undue selection bias (Fergusson et al.. 1997.
1993). Further, associations observed at ages 12-13 years between higher tooth Pb level
and lower teacher ratings of math, reading, and writing abilities (Fergusson et al.. 1993).
which are subject to greater measurement error, were supported by associations observed
at age 18 years with more objective measures such as lower probability of leaving high
school without a school certificate (i.e., completing high school) and lower scores on
school exams (Fergusson et al.. 1997). In this cohort, Pb-associated decrements in school
performance were found with consideration for potential confounding by several factors
including SES, parental education, HOME score, sex, ethnicity, number of school
changes, perinatal history, breastfeeding, maternal age, and residence in weatherboard
housing and near busy roads.
In one Boston-area cohort, those with age 6-7 year tooth Pb levels >20 ppm had greater
risk (OR: 7.4 [95% CI: 1.4, 40.7]) of dropping out of high school at mean age 18 years
compared with subjects with tooth Pb levels < 10 ppm (Needleman et al.. 1990). The
relatively small sample size (n = 132) and adjustment for several potential confounding
factors, including maternal education, IQ, and age, SES, and subject alcohol use may
have contributed to the imprecision of the effect estimate. Parental caregiving quality was
not considered. Follow-up participation was biased to subjects with lower tooth Pb levels
and higher SES. This selection bias likely did not produce a spurious association with Pb;
however, the results may be less generalizable to the original study population. In another
Boston-area cohort, higher tooth Pb level at age 6 years was associated with higher
probability of having spelling and reading difficulties as rated by teachers in age 8-year
girls but not boys (Leviton et al.. 1993). This study had a large sample size (n = 1,923)
and high follow-up participation but did not adjust for SES or parental caregiving quality.
However, other unspecified potential confounding factors were considered.
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4.3.2.6 Integrated Summary of Cognitive Function in Children
Results from recent epidemiologic and animal studies add to the strong evidence base
reviewed in the 2006 Pb AQCD (U.S. EPA. 2006b) demonstrating that Pb exposure is
associated with decrements in cognitive function in children, based strongly on
associations observed with FSIQ, and also executive function and academic performance.
Associations with performance on tests of learning and memory were less consistently
found (Table 4-5). A large epidemiologic evidence base demonstrates associations of
higher blood Pb level with lower FSIQ in children ages 3-17 years (Figure 4-2 and Table
4-3). with smaller bodies of evidence indicating associations with lower scores on tests of
executive function, and academic performance in children ages 4 to 18 years (Table 4-8
and Table 4-9). There was no clear indication that blood Pb level was more strongly
associated with performance in a particular domain of cognitive function. The
Pb-associated decrements in cognitive function observed in children were strongly
supported by observations in animals of decrements in learning, memory, and executive
function with relevant dietary Pb exposures. In particular, coherence was found between
observations of Pb-associated decrements in performance on spatial span tasks in
children and on tasks in the Morris water maze in animals both of which test visual
spatial memory. Coherence also was found for findings indicating Pb-associated
decrements in performance on spatial working memory tasks in children and in animals
with the radial arm maze (Section 4.3.2.3) and poorer cognitive flexibility in both humans
and animals as assessed in tests of rule learning and reversal (Section 4.3.2.4). Additional
biological plausibility for Pb-associated cognitive function decrements was provided by
toxicological evidence for the effects of Pb on neurophysiological and neurochemical
processes that mediate cognition.
Compelling epidemiologic evidence was described in the 2006 Pb AQCD (U.S. EPA.
2006b) for Pb-associated decrements in FSIQ (also see Section 4.3.2.1 of this ISA).
Across studies, FSIQ was measured with various instruments (i.e., WISC-R, WISC-III,
WPPSI, Stanford-Binet) that were similar in their scoring scale and measurement error.
Associations were found in most of the prospective studies, conducted in the U.S.,
Mexico, Europe, and Australia, in representative populations with moderate to high
follow-up participation and without indication of selective participation among children
with higher blood Pb levels and lower cognitive function (Figure 4-2 and Table 4-3) that
could produce spurious associations. The prospective studies found associations of blood
Pb levels measured concurrently (ages 4-10 years) with FSIQ, measured earlier in life
(i.e., prenatal cord or maternal, age 2 year or 4 year), or averaged over several years or a
lifetime, characterizing the temporal sequence between Pb exposure and cognitive
function decrements better than cross-sectional studies. Multiple testing was common;
however, the consistent pattern of association observed across the ages of blood Pb level
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and/or cognitive test examined in most previous and recent studies increases confidence
that the evidence is not unduly biased by a higher probability of associations found by
chance alone. Another strength of the prospective evidence was the consideration of
several potential confounding factors. As indicated in Table 4-3. results from most
cohorts were adjusted for maternal IQ and education, child sex and birth weight, SES,
and HOME score. Although not considered as frequently, some studies also indicated
lack of confounding by parental smoking, birth order, and nutritional factors. The
robustness of the blood Pb-FSIQ association in children was demonstrated in pooled
analyses of seven prospective cohorts (Lanphear et al.. 2005; Rothenberg and
Rothenberg. 2005) and multiple meta-analyses that combined results across various
prospective and cross-sectional studies (Pocock etal. 1994; Schwartz. 1994; Needleman
and Gatsonis. 1990). with Schwartz (1994) demonstrating the robustness of evidence to
potential publication bias. While supporting the robustness of the blood Pb-FSIQ
association, these analyses that combine data across different populations are subject to
measurement error. Nondifferential error could result from the heterogeneity in the tests
used to assess FSIQ and the ages of blood Pb and FSIQ examined. There also could be
residual confounding due to heterogeneity among studies in the potential confounding
factors examined and the method of assessment.
Among individual studies, a wide range of effect estimates was reported. Comparisons of
effect estimates across studies are difficult because of the variability in population blood
Pb distributions, lifestage or time period of blood Pb examined, type of model examined
(linear versus nonlinear), and tests conducted. The pooled analysis of seven prospective
cohorts demonstrated precision of effect estimates by finding a relatively narrow range of
effect estimates, -2.4 to -2.9 points per log increase in blood Pb level, excluding one
study at a time (Lanphear et al.. 2005). In linear models, a greater decrease in FSIQ
estimated for a 1 (ig/dL increase in concurrent blood Pb for the 244 children who had
peak blood Pb levels <10 (ig/dL (-0.80 points [95% CI: -1.7, -0.14]) and the 103 children
with peak blood Pb levels <7.5 (ig/dL (-2.9 points [95% CI: -5.2, -0.71]). Among
children with peak blood Pb levels <10 (ig/dL and <7.5 (ig/dL, the mean concurrent
blood Pb levels were 4.3 (ig/dL and 3.2 (ig/dL, respectively (Hornung. 2008). Results
from the Boston and Rochester cohorts can be considered particularly informative
because they had lower blood Pb levels compared with other cohorts. Also, compared to
the pooled analyses, there was homogeneity within each cohort in the tests used to assess
FSIQ, the age of examination, and the methods used to measure potential confounding
factors. Similarly large effects were estimated in the Boston and Rochester cohorts,
which differed widely from each other in racial and SES distributions (Canfield et al..
2003a; Bellinger etal.. 1992). While the sample sizes were smaller (Table 4-3). these
studies had at least as extensive consideration for potential confounding as other studies.
Further, each study estimated larger effects for children whose peak blood Pb levels were
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less than 10 (ig/dL, -1.8 points (95% CI: -3.0, -0.60) per 1 (ig/dL increase in concurrent
blood Pb level in the Rochester cohort (n = 101, 59%) (Canfield et al.. 2003a) and -1.6
points (95% CI: -2.9, -0.2) per 1 (ig/dL in age 2-year blood Pb level in the Boston cohort
(n = 48, 32%) (Bellinger and Needleman. 2003). These subsets of children had mean
blood Pb levels of 3.3 (Rochester) and 3.8 (ig/dL (Boston). Some recent cross-sectional
studies estimated smaller effects but with examination of populations with higher
concurrent blood Pb levels (means: 7, 8.1 (ig/dL) using a linear model (Kordas et al..
2011; Min et al.. 2009). Other recent studies estimated similar effects as previous studies
although the log-linear models make comparisons difficult. Among children ages 3-7
years in India, a 1 (ig/dL increase in concurrent blood Pb level within the range from the
10th percentile (6.0) to 10 (ig/dL was associated with a 0.54-point decrease (95% CI:
-0.91, -0.17) in FSIQ (Roy etal.. 2011). Among children ages 8-11 years in Korea, Kim
et al. (2009b) found that a 1 (ig/dL increase in concurrent blood Pb level in the 10th-90th
percentile interval (0.9-2.8 (ig/dL, detection limit = 0.06 (ig/dL) was associated with
decreases in FSIQ of 3.2 points (95% CI: -6.1, -0.24) and 2.4 points (-6.0, 1.1) in children
with blood Mn levels >1.4 (ig/dL and < 1.4 (ig/dL, respectively. In this study, the
potential influence of higher past Pb exposures cannot be excluded. Further, while these
recent studies adjusted for parental education and SES, parental caregiving quality was
not examined. The relatively low blood Pb levels in the Rochester and Boston cohorts,
consideration of peak blood Pb levels, and the adjustment for several potential
confounding factors indicate that their results may be more representative of the effects
of Pb exposure.
Previous prospective studies, several of which contributed to the FSIQ evidence,
provided key evidence for associations of blood or tooth Pb level with decrements in
executive function and academic performance for the reasons described for FSIQ.
Endpoints associated with blood or tooth Pb level included rule learning and reversal,
reading and math skills assessed using neuropsychological tests and school performance
assessed from testing or school records. Higher concurrent blood Pb level was associated
with lower scores on tests of math and reading in large analyses of children participating
in NHANES (Krieg et al.. 2010; Lanphear et al.. 2000). Recent cross-sectional studies
conducted in the U.S., Mexico, Europe, and Asia, also found associations of higher blood
Pb level with poorer executive function or academic performance or achievement. The
few recent prospective studies indicated associations between higher early childhood
blood Pb level, ages 9-36 month and age 30 months, respectively, and poorer academic
performance in fourth-grade children in North Carolina (Miranda et al.. 2009) and in
children ages 7-8 years in the U.K. (Chandramouli et al.. 2009).
In most studies that provided unadjusted and adjusted effect estimates, blood Pb level
was associated with smaller but statistically significant decrements in FSIQ after
4-125
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adjusting for potential confounding factors (Kim et al.. 2009b; Lanphear et al.. 2005;
Canfield et al.. 2003a). The consideration for potential confounding varied among
studies. Most studies adjusted for SES-related factors such as the Hollingshead Index,
household income, and/or parental education. Several, in particular the prospective
studies, adjusted for parental cognitive function and parental caregiving quality (mostly
evaluated as HOME score). Overall, recent studies considered potential confounding by
SES and parental IQ or education; few considered parental caregiving quality. Analyses
of associations between potential confounding factors and blood Pb level and cognitive
function indicated that the confounding factors may vary across populations and
endpoints. In the Cleveland cohort, adjustment for HOME score attenuated the blood or
tooth Pb level-FSIQ relationships (Greene and Ernhart. 1993; Greene et al.. 1992; Ernhart
et al.. 1988). In the Rochester cohort, HOME score met the criteria for adjustment in
models for FSIQ (Canfield et al.. 2003a) but not all measures of memory and executive
function (Froehlich et al.. 2007; Canfield et al.. 2004; 2003b). Controlling for SES is
difficult; it can be highly correlated with Pb exposure and there is no one measure that
fully represents SES. Residual confounding also is likely by factors not considered or
measured inadequately. The combination of evidence from prospective studies that
considered several well-characterized potential confounding factors plus evidence that Pb
exposure induces impairments in cognitive function in animals, in particular, spatial
memory and executive function, which were associated with blood or tooth Pb levels in
children increase confidence that the associations observed between blood and tooth Pb
levels and cognitive function in children represent a relationship with Pb exposure.
With regard to important lifestages and durations of Pb exposure, toxicological evidence
clearly demonstrates impaired learning and memory in animals exposed to Pb
gestationally with or without additional postnatal exposure. The effect of early life Pb
exposures is supported by evidence that processes such as neurogenesis and synaptic
pruning are highly active during the first few years of life (Rice and Barone. 2000; Nolte.
1993). However, evidence in a group of monkeys also indicates impaired learning with
Pb exposure beginning later during the juvenile period, indicating that Pb exposure in
gestation or in infancy is not necessary to induce impairments in cognitive function (Rice.
1992b. 1990; Rice and Gilbert. 1990b). Epidemiologic studies also point to cognitive
function decrements associated with blood Pb levels measured at various lifestages and
time periods. Among studies of young children <3 years, several found stronger
associations of MDI with prenatal (maternal or cord) blood Pb than with postnatal child
blood Pb (Hu et al.. 2006; Gomaa et al.. 2002; Bellinger et al.. 1987; Vimpani et al..
1985). However, in older children, ages 4-17 years, in whom cognitive function is more
stable and reliably measured, decrements in cognitive function were associated with
prenatal and various postnatal (concurrent, early childhood, and multiple year or lifetime
average) blood Pb metrics as well with tooth Pb levels. Evidence did not clearly identify
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an individual critical lifestage, time period, or duration of Pb exposure in terms of risk of
developing cognitive function decrements. Because bone Pb levels contribute to
concurrent blood Pb levels in children and blood Pb levels generally show strong
temporal correlation, associations with concurrent blood Pb levels may reflect an effect
of past and recent Pb exposures.
Previous prospective studies found blood Pb-associated decrements in cognitive function
in populations with mean blood Pb levels 5-10 (ig/dL (Table 4-3. Table 4-5. Table 4-8.
and Table 4-9). In analyses restricted to children in the lower range of the blood Pb
distribution (e.g., peak <10 (ig/dL), associations with FSIQ at ages 6-10 years were found
in groups of children with mean age 2 year or concurrent blood Pb levels 3-4 (ig/dL with
consideration of peak blood Pb levels (Bellinger. 2008; Canfield. 2008; Hornung. 2008).
Additional evidence for effects at lower blood Pb levels was provided recently by
Miranda et al. (2009). who found lower fourth grade EOG scores in children across North
Carolina with age 9-36 month blood Pb level of 2 (ig/dL compared to children with blood
Pb level of 1 (ig/dL. Other recent studies found associations in populations (ages 4-10
years) with mean blood Pb levels (mostly concurrent) 2-8 (ig/dL (Kordas etal.. 2011;
Kim et al.. 2009b; Jusko et al.. 2008; Zailina et al.. 2008; Chiodo et al.. 2007) for FSIQ
but not consistently for other measures of cognitive function (Cho et al.. 2010;
Chandramouli et al.. 2009; Surkan et al.. 2007). Kordas et al. (2011) examined concurrent
blood Pb levels at a young age (e.g., age 4 years) with a mean of 8 (ig/dL; other recent
results for associations at lower blood Pb levels had greater uncertainty regarding
influence of higher past Pb exposures or potential confounding. Recent animal studies
added evidence for impaired learning, associative ability, and memory with blood Pb
levels relevant to this ISA, 10-25 (ig/dL (Corv-Slechta et al.. 2010; Niu et al.. 2009;
Stangle et al.. 2007). Some recent studies from the Cory-Slechta laboratory found greater
learning impairments when Pb exposures were combined with stress, which potentially
may be mediated via effects on corticosterone and dopamine (Rossi-George et al.. 2011;
Corv-Slechta etal.. 2010; Rossi-George et al.. 2009; Virgolini et al.. 2008a; 2008b;
2005).
Biological plausibility for the epidemiologic and animal evidence linking Pb exposure to
decrements in cognitive function is provided by the well-characterized toxicological
evidence for Pb exposure interfering with development of the brain and activity of
neurochemical processes that mediate cognitive function (Section 4.3.10). Pb has been
shown to increase the permeability of the blood-brain barrier and deposit in the CNS. Pb
has been shown to impair neurogenesis, synaptic architecture, and neurite outgrowth.
Cognitive function is mediated by the cortical and subcortical structures of the brain that
integrate function in the hippocampus, prefrontal cortex, and nucleus accumbens using
dopamine and glutamate as primary neurotransmitters. Experimental studies have shown
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that Pb induces changes in dopamine and glutamate release in these regions and
decreases LTP, which is a major cellular mechanism underlying synaptic plasticity and
learning and memory.
4.3.2.7 Epidemiologic Studies of Cognitive Function in Adults
Adults without Occupational Pb Exposures
As described in the preceding section, Pb exposure that begins in gestation and lasts
through the early postnatal period or lifetime exposure that begins in gestation or after
infancy has been shown to induce learning impairments in adult animals. Less well
characterized are learning impairments in adult animals due to adult-only Pb exposures.
As reported in the 2006 Pb AQCD, epidemiologic studies have examined cognitive
performance in adults without occupational Pb exposure primarily in association with
concurrently measured blood and bone Pb levels and have found associations more with
bone Pb levels than blood Pb levels (U.S. EPA. 2006b). Recent studies produced similar
findings and provided new evidence from prospective analyses (Table 4-10).
Evidence was provided by large cohorts examining multiple exposures and outcomes,
reducing the likelihood of selective participation of subjects specifically with higher Pb
exposures and cognitive deficits. Most studies performed multiple tests of cognitive
function. However, associations with bone Pb level were not isolated to a few tests.
Several publications are available; however, many are variant analyses in the same
population (e.g., Normative Aging Study [NAS], NHANES) and all are not considered
independent assessments of the Pb-cognitive function relationship. Further, although
some evidence is available from longitudinal cohorts, most analyses are cross-sectional
examining the association between one measurement of cognitive function and a
concurrent measure of blood or bone Pb level. Because temporality cannot be
determined, causal inference regarding the effects of Pb exposure is limited. In analyses
of bone Pb level, this limitation is mitigated somewhat because bone Pb level reflects
several years of exposure. With blood or bone Pb level, it also is difficult to characterize
the specific timing, duration, frequency, and level of Pb exposure that contributed to
associations observed with cognitive function. This uncertainty may apply particularly to
assessments of blood Pb levels, which in nonoccupationally-exposed adults, reflect both
current exposures and cumulative Pb stores in bone that are mobilized during bone
remodeling (Sections 3.3 and 3.7.3). Data on U.S. blood Pb levels from the 1960s-1980s
indicate that for U.S. adults examined in the studies reviewed in this ISA, past blood Pb
levels likely are within the range considered relevant for this ISA (Sections 1.1 and 4.1).
Because of the greater difficulty in establishing the temporal sequence between Pb
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exposure and cognitive function in cross-sectional studies, in the review of evidence,
emphasis was placed on prospective analyses. Emphasis also was placed on studies that
considered several potential confounding factors such as age, education, SES, smoking,
and alcohol use. Even with the adjustment for age, there is the potential for residual
confounding because of the strong correlation between increasing bone Pb levels and
increasing age. However, the coherence with evidence for cognitive function decrements
in animals induced by long-term Pb exposure provides support for epidemiologic results.
Evidence from Prospective Studies
Key evidence for the effects of Pb exposure on cognitive function of adults has been
provided by recent prospective analyses of the large Baltimore Memory Study (BMS)
and NAS. Strengths of these studies include the repeated assessment of Pb biomarker
levels and cognitive performance, the high follow-up participation of subjects, and lack
of selective attrition by Pb biomarker levels or demographic characteristics. In particular,
the repeated assessments permitted examination of associations of bone Pb levels with
changes in cognitive function over time, which better characterized the temporal
sequence between Pb exposure and subsequent changes in cognitive function. The BMS
and NAS had different strengths, including the sex and race of subjects and potential
confounding factors considered. The BMS included men and women, ages 50-70 years,
residing in Baltimore, MD. A total of 1,140 out of 2,351 (48.5%) subjects participated
from neighborhoods that represented a diversity of race and SES. This study was unique
in that it included a large proportion of African Americans (n = 395). In comparison, the
NAS involved only men (original n = 2,280) residing in the Boston, MA area. Subjects
primarily were white and at enrollment were ages 21 to 80 years and had no current or
past chronic medical conditions. Both studies adjusted for age and education. The BMS
additionally adjusted for household wealth, and the NAS additionally adjusted for intake
of alcohol and smoking. Results from both of these cohorts with different demographics
and analytical methods indicated Pb-associated cognitive function decrements.
In the BMS, longitudinal analyses involved repeat cognitive testing at 14-month
intervals. Most subjects completed follow-up; 91% at the second round of testing and
83% at the third round (Bandeen-Roche et al., 2009). An interquartile range higher
baseline tibia Pb level (12.7 ug/g) was associated with a 0.019 unit per year decrease
(95% CI: -0.031, -0.007) in eye-hand coordination z-score, with adjustment for age, sex,
race, SES, and interviewer, with a larger decrease estimated for African Americans than
for whites (Table 4-10). Results were not homogeneous across the various tests
performed. Tibia Pb levels were more weakly associated with time-related decreases in
language, processing speed, and executive function; however, most effect estimates were
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negative in direction. Further, for language and executive function, tibia Pb level was
associated with greater decreases in scores among whites than African Americans
Similar to the BMS, among NAS men, higher baseline tibia Pb levels were associated
with decreases in cognitive performance overtime in longitudinal analyses with repeated
measures of cognitive function plus a bone Pb-time interaction term in order to estimate
the association between baseline bone Pb level and decline in cognitive test score over
time rWeisskopf et al.. 2007b). This NAS analysis expanded the evidence base by also
finding associations with patella Pb levels. Two measurements of cognitive function,
collected approximately 3.5 years apart were available for 60-70% of participants. Both
tibia and patella Pb levels were associated with decrements in executive function, short-
term memory, and visuospatial skills (as indicated by increased response latency on a
pattern comparison test). The strongest effect was estimated for the latter. Weisskopf et
al. (2007b) also found a nonlinear association with patella Pb, with Pb-associated latency
times becoming worse overtime (i.e., larger values indicating slower response time) up
to approximately 60 ug/g patella Pb then leveling off at higher levels (Figure 4-7). A
20 ug/g difference in patella Pb level was associated with an increase in latency of 0.073
ms (95% CI: 0.04, 0.12) among all men and a 0.15 ms increase among men with patella
Pb level <60 ug/g. Both patella and tibia Pb were associated with fewer errors on the
pattern comparison test. The authors proposed that this association in the unexpected
direction may be related to slowing reaction time to improve accuracy. When the nine
men with the highest bone Pb levels were removed, the association with fewer errors was
no longer statistically significant. However, the authors did not indicate whether the point
estimate changed.
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Table 4-10 Associations of blood and bone Pb levels with cognitive function in adults.
Study Population and Methodological Details
Prospective analyses presented 1st, then cross-
sectional analyses. Within categories, studies are
grouped by strength of methodology, generalizability3
Cognitive Test
Subgroup (where
examined)
Blood Pb Effect
Estimate
(95% Cl)
Bone Pb Effect Estimate
(95% Cl)
Prospective Analyses
Bandeen-Roche et al. (2009)
965, ages 50-70 yr at baseline, BMS, Baltimore,
MD
Large sample of men and women of various
races/ethnicities with repeated measures of
cognitive function and tibia Pb. High follow-up
participation over 28 mo. Marginal linear
regression models adjusted for age, sex,
household wealth, education, race/ethnicity,
interviewer. Did not consider potential confounding
by history of smoking or alcohol use.
Tibia Pb Mean (SD): 19 (12.7) ug/g
NOT EXAMINED
Longitudinal associations
Eye/hand coordination - African-American
Purdue pegboard, trail-
making test White
Cross-sectional associations
Verbal memory/learning African-American
- Rey auditory verbal
learning test White
Language - Boston
naming test
African-American
White
Change in Z-score per
IQR increase:
-0.032 (-0.052, -0.012)/yr
-0.009 (-0.024, 0.006)/yr
0.006 (-0.09, 0.10)
-0.076 (-0.153, 0.001)
0.065 (-0.010, 0.139)
-0.024 (-0.076, 0.028)
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Table 4-10 (Continued): Associations of blood and bone Pb levels with cognitive function in adults.
Study Population and Methodological Details
Prospective analyses presented 1st, then cross-
sectional analyses. Within categories, studies are
grouped by strength of methodology, generalizability3
Cognitive Test
Subgroup (where
examined)
Blood Pb Effect
Estimate
(95% Cl)
Bone Pb Effect Estimate
(95% Cl)
Weisskopf et al. (2007b)
401-750 males, mean age 68.7 yr at baseline,
MAS, Boston, MA area.
Large sample, only men, primarily white. Repeated
measures of cognitive function and tibia Pb.
Moderate follow-up participation over 3.5 yr. Linear
repeated measures analysis adjusted for age,
age2, education, smoking status, current alcohol
intake, yr between bone Pb measurement and first
cognitive test, yr between cognitive tests. Also
evaluated language, computer experience,
physical activity. Did not consider potential
confounding by other SES variables.
Mean (IQR)
Tibia Pb: 20(15) ug/g
Patella Pb: 25 (20) ug/g
NOT EXAMINED
Longitudinal associations
Visuospatial skills - pattern comparison latency
(negative = poorer performance), NES2
Executive function - verbal fluency, WISC-R
Short-term memory-# words recalled, CERAD
Cross-sectional associations
Visuospatial skills - pattern comparison latency
(negative = poorer performance), NES2
Executive function -verbal fluency, WISC-R
Short-term memory - # words recalled,
CERAD
Change in test score over
3.5 yr per IQR increase:
Tibia:-0.079 (-0.12,-0.04)
Patella:-0.073 (-0.12,-0.04)
Tibia:-0.04 (-0.16, 0.08)
Patella: -0.086 (-0.20, 0.03)
Tibia:-0.028 (-0.12, 0.06)
Patella:-0.081 (-0.17, 0.005)
Tibia: 0.03 (-0.11, 0.17)
Patella: 0.02 (-0.11, 0.14)
Tibia:-0.27 (-0.70, 0.16)
Patella:-0.22 (-0.62, 0.17)
Tibia: 0.12 (-0.20, 0.32)
Patella: 0.012 (-0.18, 0.41)
Wang et al. (2007a)
358 males, median age: 67 yr, NAS, Boston, MA
area
Same cohort as above. Subset representative of
full cohort. Linear regression adjusted for age,
years of education, smoking status, pack-years
smoking, nondrinker, grams/day alcohol
consumption, English as first language, computer
experience, diabetes. Did not consider potential
confounding by other SES variables.
Tibia Pb Median (IQR): 19 (15) ug/g
NOT EXAMINED
Mini Mental State Exam
Score
HFE wildtype
1 HFE variant
2 HFE variants
Change in test score per
IQR increase:
-0.02 (-0.10, 0.07)/yr
-0.14 (-0.33, 0.04)/yr
-0.63 (-1.04,-0.21 )/yr
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Table 4-10 (Continued): Associations of blood and bone Pb levels with cognitive function in adults.
Study Population and Methodological Details
Prospective analyses presented 1st, then cross-
sectional analyses. Within categories, studies are
grouped by strength of methodology, generalizability3
Cognitive Test
Subgroup (where
examined)
Blood Pb Eff
Estimate
(95% Cl)
ect
Bone Pb Effect Estimate
(95% Cl)
Cross-sectional Analyses
Shih et al. (2006)
985 adults, mean age: 59 yr, BMS, Baltimore, MD
Large sample of men and women of various
races/ethnicities. Subjects with tibia Pb measured
were more educated and white. Compared
blood/bone associations.
Linear regression adjusted for:
Model A: age, sex, technician, presence of APOE-
Ł4 allele
Model B: Model I, years of education,
race/ethnicity, wealth
Did not consider potential confounding by history
of smoking or alcohol use.
Mean (SD)
Concurrent blood Pb: 3.5 (2.2) ug/dL
Tibia Pb: 18.7 (11.2) ug/g
Glass et al. (2009)
1,001 adults, mean age 59 yr, BMS, Baltimore, MD
Same cohort as above. High participation rate:
91%.
Multilevel hierarchical regression model adjusted
for age, sex, race/ethnicity, education, testing
technician, time of day tested. Investigator
assessed NPH. # of NPHs increase across tertiles.
Did not consider potential confounding by history
of smoking or alcohol use.
Tibia Pb Mean (SD): 18.8 (11.1) ug/g
Language - Boston
naming test
Eye-hand coordination -
Purdue Pegboard, trail
making
Executive functioning -
Purdue Pegboard,
Stroop, trail making test
Visuoconstruction -
Rey complex figure copy
Model A
Model B
Model A
Model B
Model A
Model B
Model A
Model B
Z-score per 1
increase:
-0.006 (-0.03,
-0.002 (-0.02,
-0.011 (-0.03,
-0.008 (-0.03,
-0.014 (-0.03,
-0.010 (-0.03,
-0.019 (-0.05,
-0.014 (-0.04,
ug/dL
0.017)
0.016)
0.01)
0.01)
0.005)
0.007)
0.008)
0.01)
NOT EXAMINED
Language - Boston
naming test
Eye-hand coordination-
Purdue Pegboard, trail
making test
Executive functioning-
Purdue Pegboard,
Stroop, trail making test
Visuoconstruction -
Rey complex figure copy
2nd tertile NPH
3rd tertile NPH
2nd tertile NPH
3rd tertile NPH
2nd tertile NPH
3rd tertile NPH
2nd tertile NPH
3rd tertile NPH
Z-score per 1 ug/g increase:
-0.008 (-0.01, -0.004)
0.0006 (-0.003, 0.004)
-0.008 (-0.01, -0.004)
-0.003 (-0.006, 0.001)
-0.008 (-0.01, -0.004)
-0.001 (-0.005, 0.002)
-0.012 (-0.02, -0.007)
-0.004 (-0.009, 0.0003)
Standardized score per
1 ug/g increase:
0.001 (-0.008, 0.009)b
-0.009 (-0.017, -0.0001 )b
-0.004 (-0.012, 0.005)b
-0.006 (-0.015, 0.002)b
-0.002 (-0.010, 0.006)b
-0.010 (-0.018, -0.002)b
-0.003 (-0.014, 0.008)b
-0.006 (-0.017, 0.005)b
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Table 4-10 (Continued): Associations of blood and bone Pb levels with cognitive function in adults.
Study Population and Methodological Details
Prospective analyses presented 1st, then cross-
sectional analyses. Within categories, studies are
grouped by strength of methodology, generalizability3
Cognitive Test
Blood Pb Effect
Subgroup (where Estimate
examined) (95% Cl)
Bone Pb Effect Estimate
(95% Cl)
Weuve et al. (2006)
720-915 males, ages > 45 yr, MAS, Boston, MA
area
Large sample, only men, primarily white. Moderate
follow-up participation. Compared blood/bone
associations. Linear mixed effects regression
adjusted for age, education, computer experience,
time between Pb and cognitive assessment,
smoking status, grams/day alcohol consumption,
calorie adjusted calcium intake, regular energy
expenditure on leisure time physical activity,
diabetes. Additional adjustment for dietary factors.
n = 766 ALAD1-1, n = 149 ALAD-2
Median (IQR)
Concurrent blood Pb: 5.2 (3) ug/dL
Tibia Pb: 19(15)ug/g
Patella Pb: 27 (21) ug/g
Mini Mental State Exam
Score
ALAD1-1
ALAD-2 carrier
ALAD1-1
ALAD-2 carrier
Test score per IQR
increase:
-0.04 (-0.16, 0.07)
-0.26 (-0.54, 0.01)
Test score per IQR increase:
Tibia:-0.05 (-0.21, 0.12)
Tibia:-0.16 (-0.58, 0.27)
Patella: -0.07 (-0.23, 0.09)
Patella:-0.26 (-0.64, 0.12)
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Table 4-10 (Continued): Associations of blood and bone Pb levels with cognitive function in adults.
Study Population and Methodological Details
Prospective analyses presented 1st, then cross-
sectional analyses. Within categories, studies are
grouped by strength of methodology, generalizability3
Cognitive Test
Blood Pb Effect
Subgroup (where Estimate
examined) (95% Cl)
Bone Pb Effect Estimate
(95% Cl)
Rajan et al. (2008)
486-959 males, ages > 49 yr, MAS, Boston, MA
area
Same cohort as above. Compared blood/bone
associations. Linear regression adjusted for blood
Pb main effect, ALAD genotype, age at cognitive
test, education, grams/day alcohol consumption,
pack-years smoking, English as first language.
Also considered smoking status, income, physical
activity, diabetes, coronary heart disease, n = 818
ALAD1-1, n = 164ALAD-2
Concurrent blood Pb Mean (SD)
ALAD1-1: 5.3 (2.9) ug/dL
ALAD-2 carriers: 4.8 (2.7) ug/dL
Tibia Mean (SD)
ALAD1-1:21.9(13.8)ug/g
ALAD-2 carriers: 21.2 (11.6) ug/g
Patella Mean (SD)
ALAD1-1: 29.3 (19.1) ug/g
ALAD-2 carriers: 27.9 (17.3) ug/g
Z-score*ALAD-2 per IQR
increase:
Visuospatial - constructional Praxis, CERAD -0.05 (-0.23, 0.13)c
Executive function -verbal fluency, CERAD -0.03 (-0.22, 0.16)c
Verbal memory - word list memory, CERAD 0.003 (-0.18, 0.19)c
Perceptual speed - mean response latency
(negative = poorer performance)
continuous performance, NES
-0.18 (-0.42, 0.06)c
Z-score*ALAD-2 per IQR
increase:
Tibia: -0.25 (-0.49, -0.02)c
Patella: 0.02 (-0.19, 0.23)c
Tibia:-0.11 (-0.34, 0.13)c
Patella:-0.03 (-0.24, 0.19)c
Tibia: 0.08 (-0.15, 0.31)c
Patella: 0.14 (-0.07, 0.34)c
Tibia: -0.25 (-0.59, 0.08)c
Patella:-0.16 (-0.44, 0.12)c
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Table 4-10 (Continued): Associations of blood and bone Pb levels with cognitive function in adults.
Study Population and Methodological Details
Prospective analyses presented 1st, then cross-
sectional analyses. Within categories, studies are
grouped by strength of methodology, generalizability3
Cognitive Test
Subgroup (where
examined)
Blood Pb Effect
Estimate
(95% Cl)
Bone Pb Effect Estimate
(95% Cl)
Weuve et al. (2009)
587 females, ages 47-74 yr, Nurses' Health Study,
Boston, MA area
Large sample of only females, primarily white.
Compared blood/bone associations. Generalized
estimating equations adjusted for age and age2 at
Pb assessment, age at cognitive assessment,
education, husband's education, alcoholic
drinks/week, smoking status, physical activity, use
of aspirin, ibuprofen, Vitamin E supplements,
menopausal status, and postmenopausal hormone
use. Additional adjustment for nutritional factors,
medication use, vascular and mental health.
Assessed outcomes over telephone but a mean 5
years after Pb measured.
Mean (SD)
Concurrent blood Pb: 2.9 (1.9) ug/dL
Tibia Pb: 10.5 (9.7) ug/g
Patella Pb: 12.6 (11.6) ug/g
Orientation, registration, immediate verbal
memory with TICS. Immediate and delayed
paragraph recall, category fluency, digit span
backwards (working memory, attention) with
EBMT
Composite Z-score
Composite except letter fluency
Z-score per SD increase: Z-score per SD increase:
-0.015 (-0.069, 0.039)
0.016 (-0.071, 0.039)
Tibia: -0.040 (-0.09, 0.004)
Patella:-0.012 (-0.06, 0.03)
Tibia:-0.05 (-0.10,-0.003)
Patella:-0.033 (-0.08, 0.014)
Krieg and Butler (2009)
2,823 adults, ages 20-59 yr, Large U.S.
representative NHANES III (1991-1994).
Log-linear regression model adjusted for age, sex,
education, family income, race-ethnicity, computer
or video-game familiarity, alcohol use within the
last 3 h, test language, sampling unit and stratum.
Did not consider potential confounding by smoking.
Concurrent Blood mean (SD): 2.88 (6.91) ug/dL
Symbol Digit Substitution
(mean total latency, sec)
Serial digit learning total
score, NES
Ages 20-39 yr
Ages 40-59 yr
Ages 20-39 yr
Ages 40-59 yr
Test score per log
increase:
-0.41 (-1.7, 0.83)d
-1.5 (-3.0, 0.05)d
0.32 (-0.60, 1.3)
-1.1 (-2.6, 0.43)
NOT EXAMINED
4-136
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Table 4-10 (Continued): Associations of blood and bone Pb levels with cognitive function in adults.
Study Population and Methodological Details
Prospective analyses presented 1st, then cross-
sectional analyses. Within categories, studies are
grouped by strength of methodology, generalizability3
Krieg et al. (2009)
1 ,852-1 ,952 adults ages 20-59 yr, 1 ,551-1 ,740
adults ages > 60 yr. Large U.S. representative
NHANES III (1991-1994).
Log linear regression model adjusted for sex, age,
education, family income, race-ethnicity, computer
or video game familiarity, alcohol use in the last 3
h, test language (ages 20-59 yr) and sex, age,
education, family income, race-ethnicity, test
language (ages > 60 yr), sampling unit and
stratum. Did not consider potential confounding by
smoking.
Concurrent Blood Pb Mean (SD)
Age 20-59 yr: 2.85 (7.31 ) ug/dL
Age > 60 yr: 4.02 (3.39) ug/dL
Cognitive Test
Symbol Digit Substitution
(mean total latency)
Serial digit learning total
score
Word recall, number
correct
Story recall, number
correct, Neurobehavioral
Evaluation System
Subgroup (where
examined)
ALADrsI 800435
genotype
Ages 20-59 yr
CC/CG
GG
Ages 20-59 yr
CC/CG
GG
Ages > 60 yr
CC/CG
GG
Ages > 60 yr
CC/CG
GG
Blood Pb Effect
Estimate Bone Pb Effect Estimate
(95% Cl) (95% Cl)
Test score per log
increase NOT EXAMINED
-2.7 (-5.7, 0.30)d
-0.68 (-1.8, 0.48)d
-1.1 (-4.0, 1.8)
0.05 (-1.3, 1.4)
0.02 (-0.33, 0.37)
-0.06 (-0.23, 0.10)
-0.88 (-2.0, 0.26)
0.16 (-0.19, 0.52)
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Table 4-10 (Continued): Associations of blood and bone Pb levels with cognitive function in adults.
Study Population and Methodological Details
Prospective analyses presented 1st, then cross-
sectional analyses. Within categories, studies are
grouped by strength of methodology, generalizability3
Krieg et al. (2010)
1 ,852-1 ,952 adults ages 20-59 yr, 1 ,551-1 ,740
adults, ages > 60 yr. Large U.S. representative
NHANES 111(1991-1994).
Log linear regression model adjusted for sex, age,
education, family income, race-ethnicity, computer
or video game familiarity, alcohol use in the last
3h, test language, sampling unit, stratum (ages
20-59 yr) and sex, age, education, family income,
race-ethnicity, test language, sampling unit, and
stratum (ages > 60 yr). Did not consider potential
confounding by smoking.
Concurrent blood Pb Mean (SD)
Age 20-59 yr: 2.85 (7.31) ug/dL
Age > 60 yr: 4.02 (3.39) ug/dL
Cognitive Test
Symbol Digit Substitution
(mean total latency, sec)
Serial digit learning total
score
Word recall, number
correct
Story recall, number
correct
Neurobehavioral
Evaluation System
Subgroup (where
examined)
By VDR variant
Ages 20-59 yr
CC haplotype
CT haplotype
TC haplotype
TT haplotype
Ages 20-59 yr
CC haplotype
CT haplotype
TC haplotype
TT haplotype
Ages > 60 yr
CC haplotype
CT haplotype
TC haplotype
TT haplotype
Ages > 60 yr
CC haplotype
CT haplotype
TC haplotype
TT haplotype
Blood Pb Effect
Estimate Bone Pb Effect Estimate
(95% Cl) (95% Cl)
Test score per log
increase: NOT EXAMINED
-20 (-44, 4.0)d
0.73 (-1.4, 2.9)d
-2.6 (-5.3, 0.07)d
-3.6 (-7.2, 0.05)d
8.0(0.61, 15.4)
1.0 (-0.89, 2.9)
-1.4 (-3.1, 0.29)
-0.14 (-2.8, 2.5)
-0.65 (-1.5, 0.25)
-0.08 (-0.34, 0.19)
-0.03 (-0.40, 0.33)
-0.08 (-0.76, 0.60)
0.29 (-3.3, 3.9)
0.01 (-0.38, 0.40)
0.07 (-0.64, 0.78)
-0.22 (-0.86, 0.43)
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Table 4-10 (Continued): Associations of blood and bone Pb levels with cognitive function in adults.
Study Population and Methodological Details
Prospective analyses presented 1st, then cross-
sectional analyses. Within categories, studies are
grouped by strength of methodology, generalizability3
Cognitive Test
Subgroup (where
examined)
Blood Pb Effect
Estimate
(95% Cl)
Bone Pb Effect Estimate
(95% Cl)
Van Wijngaarden et al. (2009)
47 adults, ages 55-67 yr, Rochester, NY
Small sample size. Linear regression adjusted for
age, smoking status, educational level, history of
hypertension. Did not consider potential
confounding by other SES variables. Excluded
subjects with BMI >32 kg/m2.
Mean (SD)
Tibia Pb: 2.0 (5.2) ug/g
Calcaneus Pb: 6.1 (8.5) ug/g
Delayed matching to
sample, % correct
CANTAB
NOT EXAMINED
Paired Associate
Learning, total trials
adjusted
(increase = poorer
performance)
CANTAB
Calcaneus
Lowest tertile: 85.408
Medium tertile: 85.34
Highest Tertile: 80.83, p=0.07
Tibia
Lowest tertile: 83.048
Medium tertile: 85.91
Highest tertile: 81.85, p=0.32
Calcaneus
Lowest tertile: 2.618
Medium tertile: 2.68
Highest tertile: 2.74, p = 0.25
Tibia
Lowest tertile: 2.69e
Medium tertile: 2.64
Highest tertile: 2.72, p = 0.70
4-139
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Table 4-10 (Continued): Associations of blood and bone Pb levels with cognitive function in adults.
Study Population and Methodological Details
Prospective analyses presented 1st, then cross-
sectional analyses. Within categories, studies are
grouped by strength of methodology, generalizability3 Cognitive Test
Subgroup (where
examined)
Blood Pb Effect
Estimate
(95% Cl)
Bone Pb Effect Estimate
(95% Cl)
Gao et al. (2008)
188 adults, mean age 69.2 yr, Rural Sichuan and
Shandong Provinces, China.
Small sample size. Subset was younger, had more
education, and higher BMI than full cohort.
Separate ANCOVAs adjusted for age, sex,
education, BMI, orAPOEŁ4. History of smoking
and alcohol consumption not associated with
cognitive score.
Concurrent plasma Pb Mean (SD): 0.39
(0.63) ug/dl_
Composite cognitive Z-
score
Word list learning, word
recall (CERAD), CSID, IU
story recall, Animal
naming fluency test, IU
token test of language
and working memory
Z-score per 1 ug/L
plasma Pb increase:
-0.009 (-0.025, 0.007)
NOT EXAMINED
Note: Effect estimates in bold indicate the stronger association between blood Pb and bone Pb level. IQR = Interquartile range, BMS = Baltimore Memory Study, MAS = Normative
Aging Study, NES2 = Neurobehavioral Evaluation System 2, WISC-R = Wechsler Adult Intelligence Scale-Revised, CERAD = Consortium to Establish Registry for Alzheimer's
disease, HFE = Human Hemochromatosis protein, NPH = Neighborhood Psychosocial Hazard, TICS = Telephone Interview for Cognitive Status, EBMT = East Boston Memory Test,
CANTAB = Cambridge Neuropsychological Test Automated Battery, CSID = Community Screening Instrument for Dementia, IU = Indiana University.
"Studies are presented first for prospective analyses then for cross sectional analyses. Within categories, results are grouped by representativeness of the population characteristics
and blood Pb levels examined and extent of consideration for potential confounding. There is not necessarily a continuum of decreasing strength across studies.
bEffect estimates are for interactions between Pb and category of NPH, with the lowest fertile of NPH (fewer NPH) serving as the reference group.
°Effect estimates are for interactions between Pb and ALAD genotype.
dThe directions of effect estimates were changed to indicate a negative slope as a decrease in cognitive performance.
eResults refer to mean cognitive test scores among tertiles of bone Pb. Tertile concentrations not reported.
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Note: Models are adjusted for age, age squared, education, smoking, alcohol intake, years between bone Pb measurement and first
cognitive test, and years between the cognitive tests. The 9 subjects with the highest patella Pb levels (>89 ug/g bone mineral) were
removed. The estimated concentration-response is indicated by the solid line and the 95% confidence interval by the dashed lines.
The patella Pb level-associated increase in response latency (reference = 0 at mean of patella Pb level) is larger among men with
patella Pb levels <60 ug/g. Patella Pb levels of individual subjects are indicated by short vertical lines on the abscissa.
Source: Reprinted with permission of Williams & Wilkins, Weisskopf et al. (2007b).
Figure 4-7 Nonlinear association between patella Pb level and the relative
change over 3.5 years in response latency on the pattern
comparison test in men from the Normative Aging Study.
Longitudinal analysis of the NAS cohort also indicated that hemochromatosis (HFE)
gene variants modified the blood Pb-cognition association (Wang et al.. 2007a). In
models adjusted for age, years of education, smoking status, pack-years smoking,
nondrinker, grams/day alcohol consumption, English as first language, computer
experience, and diabetes, an interquartile range higher tibia Pb level (15 ug/g) was
associated with a 0.22 point steeper annual decline (95% CI: -0.39, -0.05) in Mini-Mental
State Examination score (MMSE, which assesses cognitive impairment in a number of
domains) among the 130 (36%) men with either the H63D or C282Y variant. The
4-141
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association was found to be nonlinear, with larger Pb-associated declines observed at
higher tibia Pb levels (Figure 4-8. solid curved line). The change in MMSE score
associated with a 15 ug/g higher tibia Pb level was comparable to that found between
NAS men who were 4 years apart in age. Tibia Pb level was not associated with a decline
in MMSE score in men with the HFE wildtype genotype (Figure 4-8. dashed line). Bone
Pb levels did not differ widely by HFE variant. HFE variants, H63D and C282Y, are
associated with hemochromatosis, a disease characterized by higher Fe body burden. Iron
metabolism has been hypothesized to affect neurodegenerative diseases, which may
explain the observed effect modification. However, firm conclusions regarding effect
modification of the bone Pb-cognition relationship by HFE variants are not warranted.
LU _fl9
co az
•= -0.4
0)
o
TO
"5 -0.6
< -0.8
HFE wildtype
HFE variant allele
10 20 30 40
Tibia lead biomarker (ug/g)
50
Note: The lines indicate curvilinear trends estimated from the penalized spline method. Among hemochromatosis (HFE) wild-types,
the association between tibia Pb and annual cognitive decline was nearly null (dashed line). Among variant allele carriers, the
association tended to deviate from linearity (solid line, p = 0.08), with a greater tibia Pb-associated decline in MMSE observed
among men with higher tibia Pb levels. The model was adjusted forage, years of education, smoking status, pack-years smoking,
nondrinker, grams/day alcohol consumption, English as first language, computer experience, and diabetes.
Source: Wang et al. (2007a).
Figure 4-8 Nonlinear association of tibia Pb level with annual rate of
cognitive decline, by hemochromatosis genotype in men from the
Normative Aging Study.
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Evidence from Cross-sectional Studies
Associations between bone Pb levels and decrements in cognitive function in adults also
are supported by evidence from several cross-sectional studies conducted in the BMS and
NAS cohorts and other populations. The cross-sectional studies have contributed
evidence for stronger associations of cognitive function decrements with bone Pb levels
than blood Pb level and for associations with adjustment for additional potential
confounding factors such as diet and medication use. While cross-sectional studies
examined factors that potentially may modify risk of Pb-associated cognitive function
decrements in adults, they each examined different factors and did not produce
conclusive evidence regarding effect modification. These subgroup analyses also are
subject to higher probability of finding an association by chance.
In addition to comparisons of blood and bone Pb levels, cross-sectional analyses of the
BMS included detailed analysis of potential confounding, although smoking and alcohol
use were not examined. Among 985 adults, higher tibia Pb level was associated with
poorer performance in tests of language, processing speed, eye-hand coordination,
executive function, verbal memory and learning, visual memory, and visuoconstruction;
blood Pb level showed associations with executive function and visuoconstruction [Table
4-10 (Shihetal.. 2006)1. Mean (SD) blood and tibia Pb levels were 3.5 (2.2) ug/dL and
18.7 (11.2) ug/g, respectively. Tibia Pb levels were associated with worse performance
on tests in all domains with adjustment for age, sex, testing technician, and presence of
the apolipoprotein (APO)E-e4 allele (potential risk factor for Alzheimer's Disease). The
magnitudes of associations were attenuated with additional adjustment for education,
race, and household wealth; however, in these more fully-adjusted models, effect
estimates for tibia Pb levels remained negative in all domains except language and
processing speed. The strongest association was found for visuoconstruction, which
reflects visuospatial skills and motor skills. A 1 ug/g higher tibia Pb level was associated
with a 0.0044 SD lower (95% CI: -0.0091, 0.0003) visuoconstruction score. Analysis of
tibia Pb as a quadratic term did not indicate a nonlinear relationship with
visuoconstruction.
In contrast with longitudinal results in BMS, race-stratified analyses of persistent effects
in cross-sectional analyses indicated that tibia Pb levels were associated with greater
decreases in performance on tests of eye-hand coordination, executive function, and
verbal memory and learning among whites than among African Americans (Bandeen-
Roche et al., 2009). Among all subjects, tibia Pb-associated decrements in cognitive
performance were modified by neighborhood level psychosocial stress. Specifically,
higher tibia Pb levels were associated with larger decrements, particularly in language,
executive function, and visuoconstruction among subjects living in neighborhoods with a
greater number of psychosocial hazards (e.g., number of violent crimes, emergency calls,
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off-site liquor licenses as assessed by investigators) (Glass et al.. 2009) (Table 4-10).
Results were adjusted for age, sex, race/ethnicity, education, testing technician, and time
of day of testing. Subjects living near more psychosocial hazards had slightly higher tibia
Pb levels. In support of these results, several studies have found Pb-stress interactions in
impaired learning and memory of adult animals with Pb exposures beginning in gestation
and lasting through post-weaning or to the time of testing (Section 4.3.2.3).
The 2006 Pb AQCD (U.S. EPA. 2006b) described cross-sectional associations of both
blood and tibia Pb levels with poorer cognitive performance among 141 NAS men
(Pavton et al.. 1998). Several recent, larger cross-sectional NAS analyses corroborated
previous findings for bone Pb but generally indicated weak associations with concurrent
blood Pb levels and only in groups with specific ALAD genetic variants. In contrast with
the longitudinal analyses, Weisskopf et al. (2007b) found that repeat measures of bone Pb
levels were inconsistently associated with cognitive function (improved and poorer
performance) in cross-sectional analyses. Among 720-915 NAS men 45 years of age and
older, higher concurrent blood and bone Pb levels were associated with lower MMSE
scores among 149 ALAD-2 carriers than among 766 ALAD-1-1 men (Weuve et al..
2006). A 27 (ig/dL higher patella Pb level (interquartile range) was associated with a 0.26
point lower mean MMSE score (95% CI: -0.64, 0.12) among ALAD-2 carriers and a 0.07
point lower score (95% CI: -0.23, 0.09) among noncarriers. Another NAS analysis
(n=486-959) did not find associations of blood or bone Pb levels with cognitive function
in various domains to be modified by ALAD genotype in a consistent direction (Raj an et
al.. 2008). The strongest interaction between higher tibia Pb level and ALAD-2 genotype
was found for visuospatial skills (constructional praxis test), and between patella Pb level
and ALAD-2 genotype for perceptual speed (pattern comparison test). The potential
direction of effect modification by the ALAD-2 genotype is not clear as the greater
affinity of the ALAD-2 enzyme subunit for Pb may increase risk of Pb-associated health
effects by increasing blood Pb levels or may diminish Pb-associated health effects by
sequestering Pb in the bloodstream and decreasing its bioavailability.
Cross-sectional studies examined a larger number of potential confounding factors than
did longitudinal analyses. The NAS found blood and bone Pb-associated decrements in
cognitive function with adjustment for dietary factors, physical activity, medication use,
and comorbid conditions (Rajan et al., 2008; Weuve et al., 2006) (Table 4-10). Additional
potential confounding factors examined in the Nurses' Health Study of 587 healthy
women in the Boston, MA area included mental health and use of aspirin, ibuprofen,
antidepressants, or vitamin E. As in the BMS and NAS, tibia and patella Pb levels were
more consistently associated with cognitive performance than was blood Pb level in the
Nurses' Health Study (Table 4-10) fWeuve et al.. 2009). Blood, patella, and tibia Pb
levels were measured between ages 47 and 74 years and an average of 5 years before
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cognitive testing. Contrary to expectation, higher patella and tibia Pb levels were
associated with higher scores on the "f" naming test (naming words that begin with f). In
separate models, the "f' naming test was omitted from a composite index of all cognitive
tests performed by phone, and a 10 (ig/g higher tibia Pb level was associated with
0.051-point lower (95% CI: -0.10, -0.003) composite cognitive function z-score (Table
4-10). A similar magnitude of decrease was estimated for a 3-year age increase in these
women. The magnitude of association was smaller for an SD increase in patella Pb level
(-0.033 [95% CI: -0.080, 0.014]), and a weak, more imprecise association was found for
an SD unit increase in blood Pb level (-0.016 [95% CI: -0.071, 0.039]).
Several analyses of the large, U.S.-representative NHANES III (1991-1994) population
of men and women investigated effect modification by age and genetic variants. Only
blood Pb levels were available and were measured in samples collected concurrently with
cognitive testing. These analyses adjusted for several of the same potential confounding
factors as other studies, with the exception of smoking. Krieg and Butler (2009) found
blood Pb level to be associated weakly with poorer performance on tests of learning and
visuospatial skills among adults ages 40-59 years and inconsistently in adults ages 20-39
years. Krieg et al. (2009) further found inconsistent associations with word and story
recall in adults ages > 60 years. Because of the different types and numbers of tests
administered, it is difficult to compare findings between adults less than and greater than
age 60 years. In the subset of the population with genetic analysis, blood Pb-cognitive
function associations were not found to be modified by ALAD genetic variants in a
consistent direction (Krieg et al.. 2009). Among adults ages 20-59 years and > 60 years,
higher concurrent blood Pb level was associated with a larger decrement in performance
on some tests in ALAD-2 carriers and other tests in ALAD -1-1 subjects (Table 4-10).
Krieg et al. (2010) found differences in the association between concurrent blood Pb level
and cognitive scores by VDRgene polymorphisms, rs731236 and rs2239185, and by the
VDR haplotype, which have unclear functional relevance. Similar to observations in
adolescent NHANES participants (Section 4.3.2.5). results were not uniform across the
various tests. However, for several tests, blood Pb level was associated with greater
decrements in cognitive performance among adults with the CC genotypes of VDR
polymorphisms.
Other cross-sectional studies with fewer subjects generally produced results consistent
with those from the larger studies described above. A study of 188 rural Chinese men and
women found a weak association between higher plasma Pb levels and a lower composite
cognitive score based on a battery of in-person administered tests (Gao et al.. 2008).
Results were adjusted for age, sex, education, BMI, or APOE-e4 in individual ANCOVA
analyses. Smoking and alcohol use were not associated with cognitive performance in
this group. Pb in plasma is not bound to erythrocytes, as is about 99% of blood Pb, and is
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the fraction delivered directly to soft tissue (Chuang etal.. 2001; Hernandez-Avila et al.,
1998). The results of Gao et al. (2008) may provide information on the cognitive effects
of a more bioavailable fraction of Pb dose; however, because there is little investigation,
firm conclusions for plasma Pb are not warranted. Among 47 men and women in
Rochester, NY (ages 55-67 years), subjects in the highest tertile of calcaneal bone (heel
bone with higher turnover rate than tibia) Pb level performed worse on delayed matching-
to-sample and paired associated learning tasks than subjects in the lowest tertile (exact Pb
levels in tertiles not reported) (van Wijngaarden et al.. 2009). In analyses of tibia Pb
levels, subjects in the highest tertile of tibia Pb level did not consistently perform worse
on the various cognitive tests (Table 4-10).
Adults with Occupational Pb Exposures
The 2006 Pb AQCD (U.S. EPA. 2006b) concluded that in adults, blood Pb levels were
associated with cognitive function more consistently among those with occupational Pb
exposures. These findings were supported by results from a few recent studies of
occupationally-exposed adults. Several of these associations were found with adjustment
for fewer but a similar set of potential confounding factors as in nonoccupational studies;
however, other occupational exposures were not considered. Occupational Mn exposure
also has been linked with decrements in short-term memory and eye-hand coordination.
A prospective analysis was conducted in former male Pb battery workers whose
occupational exposure had ceased 0.02 to 16 years (median: 6) before follow-up testing
in 2001-2004 (Khalil et al.. 2009a). Subjects included 83 of 288 workers (in 2004 mean
age: 54 years, median tibia Pb level: 57 ug/g)and51 of 181 controls (mean age: 55 years,
median tibia Pb level: 12 ug/g) from the 1982 Lead Occupational Study in Pennsylvania.
While the follow-up participation was low, participation was not biased to poor
performers on cognitive tests at baseline. In former Pb-exposed workers, a 10 (ig/g higher
peak tibia Pb levels was associated with a -0.352 change in total cognitive function z-
score (e.g., learning/memory, executive function, general intelligence, spatial function,
psychomotor speed) between 1982 and 2004. In controls, higher tibia Pb levels were
associated with improved performance on several tests. Results were adjusted for age,
education, income, blood pressure, years of employment, years since last worked,
smoking, alcoholic drinks/week, and baseline score. Adjustment for baseline cognitive
function could produce a biased association. Cross-sectional associations indicated
stronger associations of concurrent tibia Pb level than concurrent blood Pb level (median:
12 ug/dL) with poorer cognitive performance in former Pb-exposed workers. Per unit
increase, concurrent blood Pb levels were associated with larger decrements in cognitive
performance in controls than former Pb workers. As in nonoccupationally-exposed
adults, the stronger findings for tibia Pb levels in former Pb-exposed workers indicate
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stronger effects of long-term cumulative Pb exposures than recent exposures on cognitive
function. The associations for concurrent blood Pb levels in controls also may reflect
effects of cumulative exposures.
Blood and tibia Pb levels also were associated with cognitive performance in a follow-up
of 652 Pb-exposed workers (mean age: 43.4 years, mean blood Pb level: 30.9 ug/dL) in
Korea, with recent results showing associations with patella Pb levels (Dorsey et al.,
2006). Higher concurrent patella Pb levels were associated with poorer executive
function and manual dexterity with adjustment for age, sex, education, and job duration.
Associations were not found for verbal or visual memory. The associations for patella Pb
level were not as strong as those previously found for either concurrent blood or baseline
tibia Pb levels in these workers (Schwartz et al., 2005; 2001).
Other occupational studies aimed to characterize factors that either mediate or modify the
association between Pb biomarkers and cognitive function. Among current Pb smelter
workers, associations of a working lifetime time-weighted integrated blood Pb level (an
index of cumulative exposure, mean: 42 ug/dL) (p = 0.09) and tibia Pb level (p = 0.08)
with grooved pegboard test performance were mediated by Pb-associated changes in
cerebral white matter (Bleecker et al.. 2007b). In the same population (n = 112, mean
age: 38 years), higher time-weighted integrated blood Pb level (mean: 34 ug/dL) was
associated with decrements in executive function, learning, and memory among those
with lower cognitive reserve (i.e., < 11th grade reading level by Wide Range
Achievement Test-R) (Bleecker et al.. 2007a). Subjects with lower and higher cognitive
reserve were matched by blood Pb level (mean: 26 ug/dL), and results were adjusted for
age, depression scale, and current alcohol use.
Apolipoprotein E is a transport protein for cholesterol and lipoproteins and has been
found to regulate synapse formation (connections between neurons). A genetic variant,
called the ApoE-e4 allele is a haplotype between 2 exonic SNPs and has been associated
with a two-fold increased risk of developing Alzheimer's disease, although the majority
of such individuals still do not develop the disease. Thus, it is plausible that ApoE-e4
carriers may be biologically susceptible to cognitive dysfunction. A study of 529 U.S.
male, former tetra-ethyl Pb workers found that higher estimated peak tibia Pb levels were
associated with lower scores on tests of executive function, vocabulary, and memory
(Stewart et al.. 2002). and for several tests, larger decrements among the 118 men with at
least one ApoE-e4 allele. Results were adjusted for age, race, education, depression score,
testing technician, and visit number. Men with at least one ApoE-e4 allele had slightly
higher peak tibia Pb levels (mean: 26.2 versus 23.1 ug/g) and a larger percentage of non-
white subjects but were similar in age, education, and time since employment.
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Summary of Cognitive Function in Adults
In summary, consistent with evidence described in the 2006 Pb AQCD, recent studies
found that higher bone Pb levels were associated decrements in cognitive function in
adults without occupational Pb exposure (Table 4-10). Much of this evidence was
provided by analyses of the BMS and NAS, with additional findings reported in the
Nurses' Health Study and smaller populations. Nonetheless, the multiple risk factors and
health outcomes examined in most of these cohorts reduces the likelihood of selection
bias specifically related to Pb exposure and cognitive function. While the NAS and
Nurses' Health Study included primarily white men and white women, respectively, the
BMS examined a more diverse population of men and women of various races and
ethnicities. There was variability in associations across the various domains of cognitive
function tested within studies; however, higher bone Pb levels were associated with
poorer performance in most of the tests conducted. Across populations, higher bone Pb
levels were associated with decrements in executive function, visuospatial skills,
learning, and memory.
Key evidence for bone Pb-associated cognitive decrements was provided by recent
prospective analyses that demonstrated that higher tibia (means: 19, 20 (ig/g) and patella
(mean: 25 (ig/g) bone Pb levels measured at baseline were associated with subsequent
declines in cognitive function over 2- to 4-year periods (Bandeen-Roche et al., 2009;
Weisskopf et al.. 2007b). These findings indicate that long-term Pb exposure may
contribute to ongoing declines in cognitive function in adults. These associations were
found with adjustment for potential confounding by age, education, smoking, and alcohol
use in the NAS and age, sex, race, household wealth, and education in the BMS.
Supporting evidence was provided by most cross-sectional analyses, which adjusted for
several of the potential confounding factors listed above plus dietary factors, physical
activity, medication use, and comorbid conditions (Rajan et al., 2008; Weuve et al..
2006). Cross-sectional studies generally demonstrated larger decrements in cognitive
function in adults in association with tibia or patella Pb levels than with concurrent blood
Pb levels. In comparisons of associations with patella and tibia Pb levels in the NAS and
Nurses' Health Study, tibia Pb levels were not consistently associated with larger
decreases in cognitive performance (Weuve et al.. 2009; Weisskopf et al.. 2007b). In
NHANES analyses, concurrent blood Pb levels were associated with lower cognitive
function in particular age and genetic variant subgroups but not consistently across the
various cognitive tests evaluated (Krieg etal.. 2010; Krieg and Butler. 2009; Krieg et al..
2009). Analyses of NHANES did not have bone Pb measures for comparison.
Because bone Pb is a major contributor to blood Pb levels, blood Pb level also can reflect
longer term exposures, including higher past exposures, especially in adults without
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occupational exposures. Thus, in the NHANES results, it is difficult to characterize the
relative contributions of recent and past Pb exposures to the associations observed
between concurrent blood Pb level and cognitive function. In other cohorts, the
discrepant findings for blood and bone Pb levels indicate that other indicators of
cumulative Pb exposure (that likely included higher past exposures), may be better
predictors of cognitive function in adults than is concurrent blood Pb level. Additional
support for the effects of cumulative or past Pb exposure is provided by the analyses of a
few child cohort as adults, which indicated that higher childhood tooth (from lst/2nd
grade) and blood (e.g., average to age 10 years) Pb levels were associated with
decrements in cognitive function in adults ages 19-30 years (Mazumdar et al., 2011;
Bellinger et al.. 1994a). An uncertainty in the evidence for bone Pb-associated
decrements in cognitive function is the potential residual confounding by age. Although
studies adjusted for age and/or age-squared, the high correlation between increasing age
and bone Pb levels (Section 3.3.5) makes it difficult to distinguish the independent effect
of Pb exposure. However, the coherence with evidence for cognitive function decrements
associated with long-term Pb exposure in animals provides support for associations
observed in human adults.
Cross-sectional analyses provided information on potential effect modification of bone
Pb- and blood Pb-associated decrements in cognitive function in adults by race,
psychosocial stress, and genetic variants. Inconsistencies were found for effect
modification by race in the BMS, ALAD genotype in the NAS and NHANES, and VDR
genotype in NHANES. Larger tibia Pb-associated decrements in cognitive function were
found in NAS men with HFE variants and in BMS subjects living near more psychosocial
hazards. Evidence does not clearly indicate whether the observed effect modification
reflects chance, a change in the toxicokinetics of Pb that alters Pb dose at the biological
site of action, or a direct biological interaction that increases the toxicity of Pb in the
target tissue. However, such effect modification serves to strengthen inferences about
associations between Pb biomarkers and cognitive function since it is unlikely that
potential confounding factors vary by levels of the modifying factor, particularly genetic
variants. However, because there is little available evidence and inconsistent evidence for
some factors, firm conclusions regarding effect modification are not warranted.
In contrast with nonoccupationally-exposed adults, in adults with former or current
occupational Pb exposures, cognitive function decrements were associated with both
blood (means: 31-42 ug/dL) and bone Pb levels. Thus, among Pb-exposed workers, both
current and cumulative Pb exposures may affect cognitive function. Several of these
studies considered confounding by a similar set of factors as studies of
nonoccupationally-exposed adults but did not consider other occupational exposures that
have been linked with cognitive function decrements including Mn. In a prospective
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study of former Pb workers, peak tibia Pb levels were associated more strongly with
cognitive performance than were concurrent blood Pb levels (Khalil et al.. 2009a). These
results also indicate that in the absence of high current Pb exposures, cumulative Pb
exposures may have a greater effect on cognitive function in adults.
4.3.3 Externalizing Behaviors in Children
The effects of Pb exposure on externalizing behaviors such as attention, impulsivity,
hyperactivity, destructive behavior, and truancy have not been examined as extensively
as effects on cognition. Similar to cognitive function, several domains of externalizing
behaviors are inter-related, making it difficult to examine whether Pb exposure has an
effect on a particular domain. Externalizing behaviors are more complex to study than are
cognitive effects, particularly FSIQ. There are fewer objective tests of externalizing
behaviors with as strong psychometric properties or as rigorous validation as IQ tests.
Externalizing behaviors often are assessed using teacher and/or parent ratings which are
subject to greater measurement error. However, domain-specific neuropsychological
assessments are advantageous as they may provide greater insight into whether there is a
particular domain more susceptible to the effects of Pb exposure.
There are three major domains of externalizing behavior disorders: attention deficit
hyperactivity disorder, undersocialized aggressive conduct disorder, and socialized
aggressive conduct disorder (as reviewed inWhitcomb and Merrell 2012). Characteristics
of the attention deficit hyperactivity disorder domain include but are not limited to short
attention span, distractibility, impulsivity, and hyperactivity. The undersocialized
aggressive conduct disorder domain can be characterized by destructive, domineering /
threatening or disruptive behaviors, and irritability and explosiveness. Behaviors included
within the socialized aggressive conduct disorder domain include truancy from home or
school, lying and cheating, and gang membership. In the sections that follow, to the
extent possible, the evaluation of evidence for externalizing behaviors is organized by the
domains described above then by specific behaviors included within a domain. Although
not described explicitly as a part of either domain of conduct disorders, conduct disorders
can be predictors of subsequent delinquency and criminality (Soderstrom et al., 2004;
Babinski etal.. 1999; Pajer. 1998). Therefore, evidence for criminal offenses is reviewed
with conduct disorders.
As with cognitive function, in the evaluation of epidemiologic evidence for externalizing
behaviors, greater emphasis was placed on evidence from more rigorous neuro-
psychological tests, prospective studies with repeated assessments of blood Pb levels and
behavior, studies of older children in whom outcomes are more reliably measured, and
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studies of children whose blood Pb levels are less influenced by higher past Pb
exposures. Similar to cognitive function, associations between Pb biomarker levels and
externalizing behaviors potentially may be confounded by factors such as parental SES,
education, and IQ, as well as quality and stability of the caregiving environment (often
evaluated as HOME scoreTotsika and Sylva. 2004) and nutritional status. Accordingly,
greater weight was given to studies with greater consideration for potential confounding
in the study design or in statistical analyses. Consideration also was given to studies
assessing effects relevant to blood Pb levels in current U.S. children (i.e., <5 (ig/dL).
Externalizing behaviors and related clinically diagnosed disorders such as Attention
Deficit Hyperactivity Disorder (ADHD) are considered to have a strong familial
component (Whitcomb and Merrell. 2012; Dick et al., 2011; Faraone and Doyle. 2001;
Kazdin. 1995; Campbell and Werry. 1986). In recent commentaries to studies reporting
associations between blood Pb level and ADHD in children, Brondum (2011. 2007)
asserted the need for studies to consider confounding by parental history of ADHD.
Parental psychopathology potentially can affect associations between Pb and
externalizing behaviors in the child through various theoretical mechanisms. One, if
related only to the outcome and not to child Pb exposure, parental psychopathology can
be a modifier of the child Pb exposure-externalizing behavior relationship. In this case,
because the familial component of psychopathology explains a large portion of the
variance in child behavior, not removing that variance with statistical adjustment could
mask the smaller magnitude of risk due to other factors, including Pb. Thus, if there is an
association between Pb and externalizing behaviors, it may be detected in an
epidemiologic study by adjusting or stratifying by parental psychopathology.
Characterization of effect modification by parental psychopathology also would depend
on whether parental psychopathology is measured adequately. In the studies of Pb and
externalizing behaviors, few studies have examined parental psychopathology, and none
examined its role as a potential effect modifier. These studies measured parental
psychopathology mostly as parental self report of ever having any among a list of many
psychological, mental health, or psychiatric problems (Cho et al.. 2010; Nicolescu et al..
2010; Silvaet al.. 1988) which is subject to recall error and lacks specificity for the
condition being examined in the study (e.g., attention, hyperactivity, ADHD). Only Wang
et al. (2008d) adjusted child ADHD associations for parent or sibling history of ADHD
diagnosis, as ascertained from clinical records. Thus, in the few available studies,
parental psychopathology and the familial component of externalizing behaviors likely
was measured with error. However, in a potential role in effect modification, failure to
account for parental psychopathology accurately or at all may attenuate or mask an
association with Pb exposure if one exists.
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Another way parental psychopathology potentially can influence a relationship between
child Pb exposure and externalizing behaviors is by producing confounding bias. In this
hypothetical case, parental psychopathology could produce an artifactual association
between child Pb exposure and externalizing behaviors where there is no true association.
However, this confounding bias would require that parental psychopathology not only be
causally related to the outcome but also be correlated with child Pb exposure. One may
speculate that parental psychopathology could affect child Pb exposure indirectly through
parental caregiving quality. Parental caregiving quality is linked with externalizing
behaviors in children (Nuttall etal.. 2012; Campbell 1995; Rothbaum and Weisz. 1994;
Jacobvitz and Sroufe. 1987). Theoretically, parental psychopathology may result in
poorer caregiving quality or neglect which may increase a child's exposure to Pb. There
is not evidence describing this link. Cho et al. (2010) found similar concurrent blood Pb
levels in children with and without parental history of psychopathology, but parental
report of psychopathology is subject to measurement error. Further, if parental
psychopathology is a determinant of Pb exposure, then inclusion as a model covariate
would adjust for Pb exposure itself and underestimate a potential association with blood
Pb. Studies that examined parenting behaviors in parents with current ADHD have
indicated that parents with ADHD show negative parenting control, i.e., over-reactive
disciplining, lack of planning, and disorganization but have not consistently indicated that
parents with ADHD have poorer emotional responsiveness, i.e., involvement with the
child (as reviewed in Johnston et al.. 2012). Thus, it has not been demonstrated that
parents with psychopathology have poorer caregiving quality or are neglectful.
Nonetheless, if caregiving quality is the mechanism by which parental psychopathology
potentially can confound associations between child Pb exposure and externalizing
behaviors (because of a noncausal correlation between Pb exposure and parental
psychopathology), then direct adjustment for parental caregiving quality would help
address potential confounding by parental psychopathology. There still is the possibility
of residual confounding by parental caregiving quality if not measured accurately, but
this applies to any potential confounding factor examined for any outcome in relation to
any exposure. Research has not established a direct relationship between parental
psychopathology and child Pb exposure or one between parental psychopathology and
poorer parental caregiving quality. Thus, based on the currently available evidence, in
this ISA, parental psychopathology itself is not considered to be a potential confounder of
associations between child Pb and externalizing behaviors. Instead, the ISA evaluates
potential confounding by parental caregiving quality in addition to other factors well
documented in the literature to be correlated with both Pb exposure and externalizing
behaviors as described earlier in this section.
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4.3.3.1 Attention, Impulsivity, and Hyperactivity in Children
Within the attention deficit hyperactivity disorder domain of externalizing behaviors,
studies of Pb exposure have focused primarily on attention, impulsivity, and
hyperactivity. Some epidemiologic studies have examined a composite index of multiple
behaviors, and a few studies have examined physician-diagnosed ADHD. The evidence
for each of these outcomes is described below individually to the extent possible. Tests in
animals can measure aspects of both attention and impulsivity; however, in several tests,
a particular behavior that is observed can be identified as predominately reflecting a
decrease in attention or increase in impulsivity. Therefore, the animal evidence is
discussed separately according to the dominant process inferred to characterize a
particular behavior measured by a neuropsychological test.
Epidemiologic Studies of Attention in Children
Among the various externalizing behaviors examined with Pb, most of the evidence for a
relationship with Pb exposure is for measures of attention. Attention is the ability to
maintain a consistent focus on an activity or relevant stimuli and can be assessed by
examining sustained attention, concentration or distractibility. Several epidemiologic
studies reviewed in the 2006 Pb AQCD (U.S. EPA. 2006b) reported associations between
higher blood or tooth Pb levels with attention decrements in children from 1st grade to
age 17 years, including prospective studies described in previous sections for cognitive
function (Ris et al.. 2004; Fergusson et al.. 1993; Leviton et al.. 1993). As described in
this section, most recent studies also found associations of higher blood Pb levels with
attention decrements in children ages 7-17 years. Many previous studies of attention
included children with higher blood Pb levels than most current U.S. children but had
strengths including prospective design and greater consideration for potential
confounding. Some recent studies found blood Pb-associated attention decrements in
populations of children with mean concurrent blood Pb levels 2 to 5 (ig/dL (Cho et al..
2010; NicolescuetaL 2010; Plusquellec et al., 2010; Chiodo et al., 2007) but in contrast
with prospective studies had limitations including cross-sectional design, the potential
influence of higher past Pb exposures, and consideration of fewer potential confounding
factors. In the collective body of literature, most evidence was for attention rated by
teachers, parents, or blinded examiners; however, associations were consistently found
for more objective measures such as performance on the continuous performance test
(CPT) (Figure 4-9 and Table 4-11). Thus, evidence does not indicate undue influence by
biased reporting of attention decrements by parents of children with higher Pb exposures.
In support of the epidemiologic findings for attention decrements in children, some
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animal studies have found Pb exposure to induce less accurate signal detection in tests
with distracting stimuli, which can reflect impaired attention (discussed below).
Most prospective studies that examined attention found associations with blood or tooth
Pb level and had several strengths. With prenatal blood Pb, blood Pb measured 9-11 years
before attention, lifetime average blood Pb, or tooth Pb, prospective studies characterized
the temporal sequence between Pb exposure and attention decrements better than cross-
sectional studies and made reverse causation a less likely explanation for the observed
associations. These studies recruited cohorts from schools or prenatal clinics and had
moderate to high follow-up participation that was not conditional on blood or tooth Pb
levels, which reduces the likelihood of selection bias. Not all results were statistically
significant but overall they showed a pattern of lower attention with higher blood or tooth
Pb level (Figure 4-9 and Table 4-11). Most of these prospective studies were reviewed in
the 2006 Pb AQCD and examined populations with mean blood Pb levels 7-14 (ig/dL.
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Study
Ris etal. (2004)
Cho etal. (2010)
CNodoeial. (2007)"
Chtatoelal. (2004)«
Bumselal (1999)
Canfield etal. (20036)
Nicolescu etal. (2010)
Silva etal. (1988)
Plusquellec etal. (2010)
Kordas etal, (2007)
Roy etal. (2009a)
Chiodoe(al,(2007f
Chiodoetal. (2004)'
Nicolescu etal. (2010)
Nigg etal. (2008)
Nigg etal. (2008)*
Plusquellec etal. (2010)
Mean (SO) blood
Pb(pgfdl) into
8.3
13
8.3
19(0.67)
5.0 (3.0)
5.4(3.3)
5,4(3.3)
GM; 14,3(135-15,1)
GM: 13.9 (13.2-14.6)
6.5
Median 3.7 (2 .5 5.1)
11 (4,9)
5.4 (5.0)
12(6,1)
11 (5,3)
5,0 (3.0)
5.4(3.3)
Median 3,7 (2 .5-5,1)
10(0.49)
10(0.49)
5.4 (5.0)
Blood Pb
irval analyzed
4.3-10
70-10
36-10
1 2-2.8
2.1-B.7
22-9,5
2 2-9.5
137-149
13.4-14.4
No Data
2.6-8.4
5.9-10
14-10.8
5,4-10
5.8-10
21-8,7
22-9.5
26-8,4
0.5-17
0.5-1.7
1.4-11
Outcome
Attention decrement, Continuous Performance Test
Comissions errors. Continuous Performance Test
Response time
Gomissjon errors, Continuous Performance Test
Omission errors
Numberof errors, Continuous Performance Test
Comissions errors
Reaction Time
Attention problems, boys. Child BehaviorChecklist
Attention problems, gins
Attention decrement, confrol phase, Shape Sefsool
Attention decrement, completed phases
Inattention, parent, Gonners
Inattention, teacher
Inattention .parent, Rutter
Inattention, teacher
Off task duration. Infant Behavioral Rating Scale
Off task passive
Inattention. Conners
Attention problems.teacherPROBS-14
Inattenton.Barkley-DuPaul
Go/nogo, Test of Attentions! Performance
Stop task
Hyperactivity/trnpulsivity, Child BehaviorCheckist
Impulsivity
Attention
decrement
Nicolescu etal. (2010) Median 3.7(2.5-5,1) 2.6-8.4
Silva etal. (1988) 11(4.9) 5.9-10
Plusquellecetal.(2010) 5.4(5.0)
Royetal. (2009a) 11(5.3)
Chiodoetal. (2007)« 5.0(3.0)
Chen etal, (2007) 12(5.2)
1.4-11
58-10
2.1-8.7
6.6-10
Cho etal (2010) 19(0.67) 1.2-2.8
Chen etal. (2007) 12(5.2) 6.6-10
Hyperactivity,parent, Conners
Hyperactivity, teacher
Hyperactivity. parent. Rutter
Hyperactivity. teacher
Global activity rate. Infant Benavloral Rating Scale
Hyperactivity, Conners
Hyperactivity, teacher. Achenbach
Hyperaclivity. direct, Conners
Hyperactivity. indirect
Total ADHD rating, teacher, Korean ADHD Rating Scale
Total ADHD rating, parent *
ADHD index, direct, Conners hi
ADHD index, indirect I -
Hyperactivity
Total ADHD
Rating
-0.5 -0.4 -0.3 -02 -0.1 0 0.1 0.2 0.3 0.4 0.5
Standardized change perl (jg/dL increase m
blood Pbtevel in various intervals of blood Pb level
aStandard errors were estimated from p-values or sufficient data were not available to calculate 95% CIs.
Note: Regression coefficients were reported as T- or factor scores or were scaled to their standard deviation to facilitate
comparisons among tests with different scales. Small effect estimates do not necessarily indicate lack of effect or weak effect.
Results are categorized by outcome category. Where available, results from objective tests are presented first. Within categories,
results generally are grouped according to strength of study design, representativeness of the study population characteristics and
blood Pb levels examined, and extent of potential confounding considered. There is not necessarily a continuum of decreasing
strength across studies. Effect estimates are standardized to a 1 ug/dL increase in blood Pb level in the interval from the 10th to the
90th percentile, or to 10 ug/dL, which ever is lower. The percentiles are estimated using various methods and are approximate
values. Effect estimates are assumed to be linear within the interval of blood Pb level examined. Orange, blue, black, and gray
symbols represent associations with prenatal maternal, childhood several years before outcome, concurrent, and lifetime average
blood Pb levels, respectively.
Figure 4-9 Associations of blood Pb levels with attention, impulsivity, and
hyperactivity in children.
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Table 4-11 Additional characteristics and quantitative results for studies presented in Figure 4-9.
Study
Study Population and Methodological Details
Organized by outcome then strength of study
methodology, generalizability3
Blood/Tooth Pb
Metrics Analyzed
(ug/dL or ug/g)
Outcomes Analyzed
Effect Estimate
(95% Cl)b
Bellinger et al. 79 young adults followed from 1st-2nd grade (1975-
Q994a) 1978) to ages 19-20 yr, Chelsea/Somerville, MA
Prospective. Low follow-up participation. Participation
from higher SES and initial IQ and nonwhite but no
effect on association with tooth Pb level. Log linear
regression model adjusted for parental IQ, sex, SES,
current drug, alcohol and illicit drug use, maternal
education and age, birth order. Also considered
potential confounding by other unspecified factors.
Deciduous tooth
(1st/2nd grade) in
ppm
Q1: 2.9-5.9
Q2: 6.0-8.7
Q3: 8.8-19.8
Q4: 19.9-51.8
Mean (SE) Correct Responses
per quartile
Q1: 98.0(1.0)c
Q2: 97.6(1.1)
Q3: 96.9(1.1)
Q4: 94.6(1.1)
Mean (SE) Reaction Time
Errors per quartile
Continuous Performance Test
(CPT), ages 19-20yr
Q1: 361.2(16.5)
Q2: 374.1 (17.3)
Q3: 370.7(17.9)
Q4: 385.0(17.1)
Decrement in attention factor of
CPT outcomes, ages 15-17 yr
Prenatal (maternal)
Age 3-60 mo avg
Age 78 mo
Ris et al. 195 children followed prenatally (1979-1985) to age
(2004) 15-17 yr, Cincinnati, OH
Prospective. Recruitment at prenatal clinic. High
follow-up participation, no selective attrition. Mostly
African-American. Linear regression model adjusted for
SES, maternal IQ, HOME score, adolescent marijuana
use, child sex, birth weight. Also considered potential
confounding by birth outcomes, maternal age, prenatal
smoking, alcohol, marijuana, and narcotics use,
gravidity, # previous abortions, stillbirths, parity,
caregiver education, public assistance, child age,
health, and Fe status.
Mean (SD)
Prenatal maternal:
8.3(3.8)
Age 3-60 mo avg:
13(6.1)
Age 78 mo:
8.3(4.8)
See Wright et al.
(2008)
Intervals analyzed:
4.3, 7.0, 3.6 (10th
percentiles) -10
0.16(0.04, 0.27)
0.11 (0.03, 0.19)
0.12(0.02, 0.22)
Cho et al. 639 children ages 8-11 yr, born 1997-2000, Seoul,
(2010) Seongnam, Ulsan, Incheon, Yeoncheon, Korea
Cross-sectional. School-based recruitment, moderate
participation rate. Log linear regression model adjusted
for age, sex, paternal education, maternal IQ, child IQ,
birth weight, urinary cotinine, residential area. Did not
consider potential confounding by parental caregiving
quality.
Concurrent blood
Mean (SD):
1.9(0.67)
Interval analyzed:
1.2-2.8 =
10th-90th
percentiles
Commission errors, CPT
Response time, CPT
Total ADHD rating, teacher
Total ADHD rating, parent
Korean ADHD Rating Scale IV
Ages 8-11 yr
0.02 (0, 0.04)d
-0.01 (-0.03, 0.02)d
0.04 (0.02, 0.07)d
0.01 (-0.01, 0.03)d
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Table 4-11 (Continued): Additional characteristics and quantitative results for studies presented in Figure 4-9.
Study
Study Population and Methodological Details
Organized by outcome then strength of study
methodology, generalizability3
Blood/Tooth Pb
Metrics Analyzed
(ug/dL or ug/g)
Outcomes Analyzed
Effect Estimate
\b
(95% Cl)°
Nicolescu et al. 83 children ages 8-12 yr (born 1995-1999), Bucharest
(2010) and Pantelimon, Romania
Cross-sectional. Pantelimon near former metal
processing plant. Low correlations blood Pb with blood
Al, Hg. No information on participation rate. Log linear
regression model adjusted for city, sex, age, computer
experience, handedness, eye problems, # siblings,
parental education, prenatal smoking and alcohol use,
parent ever having psychological/psychiatric problem
(for go/no go).
Concurrent blood
Median (25th-
75th): 3.7(2.5-5.1)
Interval analyzed:
2.6-8.4 =
10th-90th
percentiles
Go/no go, KITAP
Inattention parent rating
Inattention teacher rating
Hyperactivity, parent rating
Hyperactivity, teacher rating
German Conners Rating
Ages 8-12 yr
8.2% (-0.02, 21 )e
1.2% (-3.5, 6.7)
4.1% (-0.02, 10)
5.0%(0, 12)
4.1% (-0.02, 11)
Nigg et al. 150 children ages 8-17 yr, Birth yr and location not
(2008) reported
Case-control study of ADHD. Recruitment by
mailings, outreach to clinics, community
advertisements. Could have biased participation by Pb
exposure leading to biased behavioral ratings.
Regression-based path analysis adjusted for sex and
income. Did not consider potential confounding by
parental education or caregiving quality.
Concurrent blood
Mean (SD):
1.03(0.49)
Detection limit=0.3
Interval analyzed:
0.5-1.7 =
10th-90th
percentiles
Stop task
Hyperactivity/impulsivity
Teacher, parent rating, Child
Behavioral Checklist, ADHD
Rating Scale, Conners Rating
Scale
Ages 8-17 yr
0.38(0.16, 0.60)d
0.21 (0, 0.42)d
Needleman et 158 children in 1st/2nd grade (1975-1978), Chelsea,
al. (1979) Somerville, MA
Cross-sectional. Recruitment from schools. Only 15%
selected based on low and high tooth Pb levels. Low
participation rate but no selective participation based
on tooth Pb or teacher ratings. Analysis of covariance
adjusted for paternal SES, parental IQ, and maternal
age, # pregnancies, and education. Did not consider
potential confounding by parental caregiving quality.
Tooth (1st/2nd
grade)
High >20 ppm
(n = 58)
Low <10 ppm
(n = 100)
Reaction time, 12 sec delay
% negative response, teacher
rating
Impulsive
Hyperactive
Mean (SD) seconds0
Low tooth Pb: 0.41 (0.09)
High tooth Pb: 0.47(0.12)
High vs. low tooth Pb
25 vs. 9%, p = 0.01
16 vs. 6%, p = 0.08
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Table 4-11 (Continued): Additional characteristics and quantitative results for studies presented in Figure 4-9.
Study
Study Population and Methodological Details
Organized by outcome then strength of study
methodology, generalizability3
Blood/Tooth Pb
Metrics Analyzed
(ug/dL or ug/g)
Outcomes Analyzed
Effect Estimate
\b
(95% Cl)°
Chiodoetal. 451-466 children, age 7 yr (born 1989-1991), Detroit,
(2007) Ml area.
Cross-sectional. Recruitment at prenatal clinic. 100%
African-American. High prevalence prenatal drug
exposure. High participation rate. Linear regression
model adjusted for child sex, prenatal marijuana use
(commission errors), caregiver education, HOME,
maternal IQ, cocaine use, prenatal alcohol use
(omission errors), child age and sex (hyperactivity),
child age and sex, SES, HOME, caretaker education,
maternal age, prenatal alcohol use, current marijuana
use (attention problems). Also considered potential
confounding by # children in home, caretaker marital
status, concurrent alcohol/week and cigarettes/day,
caregiver concurrent psychological symptoms, and
maternal custody and prenatal cigarettes/day
Concurrent blood
Mean (SD): 5.0
(3.0)
Interval analyzed:
2.1-8.7 =
10th-90th
percentiles
Commission Errors (%), CRT
Omission Errors (%), CRT
Attention problems, teacher
rating, PROBS-14
Hyperactivity, Achenbach
Teacher Report Form
Age 7 yr
-0.08, p >0.05f
0.18(0.04, 0.32)d
0.13(0.03, 0.23)d
0.13(0.03, 0.23)d
Chiodo et al. 164-236 children, age 7.5 yr, Detroit, Ml area
(2004) Cross-sectional. Recruitment at prenatal clinic. 100%
African-American. High prevalence prenatal alcohol
and drug exposure. Moderate participation rate. Log
linear regression model adjusted for SES, # children
0.05f
0.25, p >0.05f
0.44 (0, 0.89)d
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Table 4-11 (Continued): Additional characteristics and quantitative results for studies presented in Figure 4-9.
Study
Study Population and Methodological Details
Organized by outcome then strength of study
methodology, generalizability3
Blood/Tooth Pb
Metrics Analyzed
(ug/dL or ug/g)
Outcomes Analyzed
Effect Estimate
\b
(95% Cl)°
Fergusson et 878 children followed from birth to age 13 yr,
al. (1993) Christchurch, New Zealand
Prospective. Moderate follow-up participation, attrition
did not affect results. Log linear regression model
adjusted for sex, maternal education, family size,
HOME score, # schools attended. Also considered
potential confounding by ethnicity, SES, maternal age,
breastfeeding duration, paternal education, parental
smoking, child birth outcomes, residence on busy
roads, weatherboard housing.
Tooth (age 6-8 yr)
Mean (SD): 6.2
(3.7) ug/g
Inattention/restlessness
and Conners', parent and
teacher ratings
Age 13 yr
0.06(0, 0.12)c
Chandramouli 488 children followed from age 4 mo (born 1992) to
et al. (2009) ages 7-8 yr, Avon, U.K.
Prospective. All births in area eligible. Similar
characteristics as U.K. census, high participation at
baseline and follow-up. Participants had better
educated mothers, who smoked less, better home
environment. Log linear regression model adjusted for
maternal education and smoking, home ownership,
home facilities score, family adversity index, paternal
SES, parenting attitudes at 6 mo, child sex. Also
considered potential confounding by child IQ.
Age 30 mo blood
Not reported
Group 1: 0-<2
Group 2: 2-<5
Group 3: 5-<10
Group 4: >10
OR for increased score
Selective attention decrements
Test of Everyday Attention for
Children Ages 8 yr
Reference
1.03(0.66, 1.61)c
0.99(0.62, 1.57)c
1.14(0.54,2.40)c
Group 1: 0-<2
Group 2: 2-<5
Group 3: 5-<10
Group 4: >10
Hyperactivity, teacher
Strengths and Difficulties
Questionnaire,
Ages 7 yr
Reference
0.84(0.47, 1.52)c
1.25(0.67, 2.33)c
2.82(1.08, 7.35)c
Leviton et al. 1,923 children followed from birth (1979-1980) to age 8
(1993) yr, Boston, MA area
Prospective. Recruited from birth hospital. Moderate
participation at baseline and follow-up. Regression
model adjusted for single parent family, gestational
age, # children in home, maternal education,
ethnicity, daycare in 1st 3 years. Also considered
potential confounding by other unspecified factors
Prenatal (cord)
blood
Mean: 6.8
Tooth Pb (Age 6
yr)
Mean: 3.3 ug/g
Daydreaming problems, Boston
Teacher Questionnaire, age 8 yr
RR (yes/no) per natural log
increase
Prenatal (cord)
Girls: 1.3(0.8, 2.2)c
Boys: 1.0(0.6, 1.5)c
Tooth Pb
Girls: 1.5(0.9, 2.6)c
Boys: 1.1 (0.7, 1.7)c
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Table 4-11 (Continued): Additional characteristics and quantitative results for studies presented in Figure 4-9.
Study
Study Population and Methodological Details
Organized by outcome then strength of study
methodology, generalizability3
Blood/Tooth Pb
Metrics Analyzed
(ug/dL or ug/g)
Outcomes Analyzed
Effect Estimate
\b
(95% Cl)°
Burns et al. 322 children followed from birth (1979-1982) to ages
(1999) 11-13 yr, Port Pirie, Australia.
Prospective. Moderate follow-up participation.
Participants had higher birth weight, older mothers,
less educated fathers. Log linear regression model
adjusted for maternal age, prenatal smoking status, IQ,
concurrent psychopathology, and education, birth
weight, type of feeding, length of breastfeeding,
paternal education and occupation, birth order, family
functioning, parental smoking, marital status, HOME
score, child IQ.
Lifetime avg (to
age 11-1 Syr)
blood
GM (5th-95th)
Boys: 14.3
(13.5-15.1)
10th-90th:
13.7-14.9
Girls: 13.9
(13.2-14.6)
10th-90th:
13.4-14.5
Attention problems, boys
Attention problems, girls
Maternal rating by Child
Behavior Checklist
Ages 11-1 Syr
0.02 (-0.04, 0.08)
0.07(0.02, 0.12)
Wasserman et 191 children followed prenatally (1984-1985) to age
al. (2001) 5 yr, Pristina, Yugoslavia.
Prospective. High follow-up participation. Participants
had lower maternal education, were Albanian, had
higher average blood Pb at age 4 yr. Generalized
estimating equations with log blood Pb adjusted for
child sex, birth weight, and age, ethnicity, HOME score,
maternal education and smoking. Did not consider
potential confounding by maternal IQ.
Lifetime avg (to
age 5 yr) blood
Mean (SD): 7.2
Only log blood Pb
reported
Attention problems
Maternal rating, Child Behavior
Checklist
Repeated measures
ages 4-5 yr.
Change in raw score per log
increase in blood Pb:
0.06 (-0.21, 0.33)c
Chen et al. 721 children in TLC trial followed from ages 1-3 to 5 yr,
(2007) Baltimore, MD; Cincinnati, OH; Newark, NJ;
Philadelphia, PA
Prospective. Mostly African-American. Multi-city, high
participation rate. Eligible for chelation trial because of
high age 1-3 yr blood Pb levels (20-44 ug/dL). Linear
regression-based path analysis adjusted for city, race,
sex, language, single parent, parental education and
employment, age at blood Pb measurement, caregiver
IQ. Considered potential confounding by chelation
treatment but not parental caregiving quality. Direct =
independent of IQ. Indirect = mediated through IQ.
Age 2 year blood
Mean (SD): 26
(5.1)
Interval analyzed:
20-33 =10th-90th
percentiles
Hyperactivity Index
ADHD index
Direct: 0.07 (-0.07, 0.20)
Indirect: 0.03 (0, 0.05)
Direct: 0.04 (-0.09, 0.18)
Indirect: 0.05(0.03, 0.13)
Concurrent blood
Mean (SD): 12
(5.2)
Interval analyzed:
6.6 (10th
percentile)-10
Hyperactivity Index
ADHD index
Parent ratings, Conners Scale-
Revised, Age 5 yr
Direct: 0.08 (-0.06, 0.21)
Indirect: 0.04 (-0.06, 0.13)
Direct: 0.04 (-0.10, 0.18)
Indirect: 0.07(0.03, 0.11)
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Table 4-11 (Continued): Additional characteristics and quantitative results for studies presented in Figure 4-9.
Study
Study Population and Methodological Details
Organized by outcome then strength of study
methodology, generalizability3
Blood/Tooth Pb
Metrics Analyzed
(ug/dL or ug/g)
Outcomes Analyzed
Effect Estimate
\b
(95% Cl)°
Canfield et al. 96-127children born 1994-1995 followed ages 6 mo-4-
(2003b) 4.5 yr, Rochester, NY
Cross-sectional. Recruitment from study of dust
control. Mostly nonwhite. Moderate follow-up
participation, no comparison of nonparticipants. Linear
mixed effects model adjusted for child age, gestational
age, maternal IQ and education, HOME, race,
color/shape knowledge, child IQ (both outcomes). Plus:
race (control phase), child sex (inhibit phase). Also
considered potential confounding by birth weight,
household income, prenatal smoking exposure.
Age 4 yr blood
Mean: 6.5
Range: 1.7-21
10th-90th: data not
available
Attention decrement, control
phase
Attention decrement, inhibit
phase
Examiner rating during Shape
School Task
Repeated measures at ages 4
and 4.5 yr
0.01 (-0.01, 0.04)
0.008 (-0.02, 0.04)
Silva et al. 535 children age 11 yr (born 1972-1973), Dunedin,
(1988) New Zealand
Cross-sectional. Moderate participation rate.
Participants were of higher SES and non-Maori. Log
linear regression adjusted for SES, maternal verbal
skills, age at 1st birth, and mental health symptoms,
change in residence and school, solo parenting,
child/parent separation, family relations, marriage
guidance sought, child sex, birth order, and IQ. Did not
consider potential confounding by parental caregiving
quality.
Concurrent blood
Mean (SD): 11
(4.9)
Interval analyzed:
5.9 (10th
percentile)-10
Inattention, parent
Inattention, teacher
Hyperactivity, parent
Hyperactivity, teacher
Rutter Behavior Questionnaire
Age 11 yr
0.06 (-0.03, 0.16)
0.15(0.06, 0.25)
0.13(0.03, 0.23)
0.12(0.01, 0.22)
Plusquellec et 84-98 children, ages 5-6 years (born 1994-1997), Inuit
al. (2010) communities, Quebec, Canada
Cross-sectional. Study of multiple exposures. Low but
no selective participation by blood Pb, PCBs, or Hg.
Log linear regression model adjusted for birth weight,
sex, parity, caregiver education (impulsivity); birth
weight, SES, child blood Hg (off task duration),
maternal prenatal binge drinking, # children in
home,cord and child serum fatty acids (global activity
rate). Also examined potential confounding by,#
residents per room, caretaker psychological distress,
nonverbal reasoning, and linguistic acculturation,
HOME, maternal prenatal illicit drug use and
cigarettes/day, serum Se. p > 0.20 for associations with
blood Hg or PCBs.
Concurrent blood
Mean (SD): 5.4
(5.0)
Interval analyzed:
1.4-11 = 10th-90th
percentiles
Off task duration
Impulsivity
Global activity rate
Examiner ratings, modified
Infant Behavior Rating Scale
Ages 5-6 yr
0.02 (0, 0.04)d
0.02 (0, 0.04)d
0.014 (-0.01, 0.03)d
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Table 4-11 (Continued): Additional characteristics and quantitative results for studies presented in Figure 4-9.
Study
Study Population and Methodological Details
Organized by outcome then strength of study
methodology, generalizability3
Blood/Tooth Pb
Metrics Analyzed
(ug/dL or ug/g)
Outcomes Analyzed
Effect Estimate
\b
(95% Cl)°
Kordas et al. 157 children ages 6-8 yr (born 1993-1995), Torreon,
(2007) Mexico
Cross-sectional. 26% selected from larger study for
classroom observation. Residence near metal foundry.
Linear regression model adjusted for age, sex, SES,
home ownership, crowding in home, maternal
education, child living with 1 or both parents, forgetting
homework. Also considered potential confounding by
micronutrients but not parental caregiving quality.
Concurrent blood
Mean (SD): 12
(6.1)
Interval analyzed:
5.4 (10th
percentile)-10
Off task passive behavior
Examiner rating, instrument
developed by investigator
Ages 6-8 yr
0.034(0.01, 0.06)
Roy et al. 756 children ages 3-7 yr (born 1998-2003), Chennai,
(2009) India
Cross-sectional. Recruitment at schools. No
information provided on participation. Log linear
regression model adjusted for age, sex, average
monthly income, hemoglobin, parental education, #
other children, cluster of school and classroom. Did not
consider potential confounding by parental caregiving
quality.
Concurrent blood
Mean (SD): 11
(5.3)
Interval analyzed:
5.8 (10th
percentile)-10
Inattentive z-score
Hyperactive z-score
ADHD index z-score
Teacher ratings, Conners'
ADHD/DSM-IV Scales
Ages 3-7 yr
0.03(0.01, 0.06)
0.02 (-0.01, 0.04)
0.02(0, 0.05)
Rabinowitz et 493 children, grades 1-3, Taiwan
al. (1992) Cross-sectional. Some reside near smelter.
Recruitment from schools. High participation rate.
Logistic regression model adjusted for sex, # adults at
home. Also considered potential confounding by grade,
child longest hospital stay parental education, SES,
birth outcomes, handedness, language at home,
prenatal maternal medicine, alcohol, and smoking but
not parental caregiving quality.
Tooth (grades 1-3)
Mean (SD): 4.6
(3.5)
Hyperactivity Syndrome
Boston Teacher Questionnaire
Grades 1-3
OR vs. <2.3 ug/g as
reference
2.3-7 ug/g:
1.9(0.53, 6.5)c
>7
2.8(0.68, 12)c
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Table 4-11 (Continued): Additional characteristics and quantitative results for studies presented in Figure 4-9.
Study
Froehlich et al.
(2009)
Study Population and Methodological Details
Organized by outcome then strength of study
methodology, generalizability3
2,588 children, ages 8-15 yr (born 1986-1996), U.S.
NHANES 2001 -2004
Cross-sectional. U.S. representative results, study of
multiple risk factors and outcomes. Logistic regression
adjusted for child serum cotinine, sex, age, preschool
attendance, birth weight, income/poverty ratio,
maternal age and prenatal smoking, race/ethnicity. Did
not consider potential confounding by parental
education or caregiving quality.
Blood/Tooth Pb
Metrics Analyzed
(ug/dL or ug/g)
Concurrent blood
Tertilel: 0.2-0.8
Tertile2: 0.9-1.3
TertileS: >1.3
1% below
detection limit of
0.3
Outcomes Analyzed
ADHD DSM-IV criteria met
(yes/no)
Parental rating, ADHD module
of Diagnostic Interview
Schedule for Children
Age 8-15 yr
Effect Estimate
(95% Cl)b
OR vs. 0.2-0.8 ug/dL as
reference
0.9-1.3ug/dL:
1.7(0.97, 2.9)g
>1.3ug/dL:
2.3(1.5, 3.8)g
"Results are organized by outcome examined, then grouped by strength of study design, representativeness of the study population characteristics and blood Pb levels examined, and
extent of potential confounding considered. There is not necessarily a continuum of decreasing strength across studies.
bEffect estimates are standardized to a 1 ug/dL increase in blood Pb level in the interval from the 10th percentile of blood Pb level to 10 ug/dL or the 90th percentile, whichever is lower
and scaled to the standard deviation of the test score to facilitate comparisons among tests that are scored on different scales. For studies with 10th percentiles of blood Pb level > 10
ug/dL, effect estimates are standardized to a 1 ug/dL increase in blood Pb level in the interval from the 10th to 90th percentile of blood Pb level.
°Results not presented in Figure 4-9 because tooth Pb analyzed or data reported only for log blood Pb.
dStandard error was estimated from the reported p-value.
eResults represent the change in false alarm rate.
'Sufficient data were not available to calculate 95% CIs.
9Results not presented in Figure 4-9 because OR or RR reported in papers.
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Prospective studies in Cincinnati and in Chelsea/Somerville, MA found Pb-associated
attention decrements in adolescents and young adults as indicated by increases in
commission and omission errors or reaction time on the objective CPT (Ris et al.. 2004;
Bellinger et al.. 1994a). In the CPT, subjects are assessed for their ability to maintain
focus during a repetitive task and respond to targets or inhibit responses. Stronger, more
consistent results were found in the larger Cincinnati cohort with the analysis of blood Pb
levels as a continuous variable. In 195 mostly African-American, lower SES adolescents
ages 15-17 years, Ris et al. (2004) found attention decrements (composite of CPT
outcomes) in association with higher prenatal maternal, age 3-60 month average, and age
78 month blood Pb levels, particularly among males. The mean (SD) blood Pb levels as
reported in Wright et al. (2008) were 8.3 (3.8) (ig/dL for maternal prenatal, 13.4 (6.1) for
age 3-60 month average, and 8.3 (4.8) for age 78 month. Among 79 primarily white,
higher SES young adults ages 19-20 years, Bellinger et al. (1994a) found that compared
with the group with lst/2nd grade tooth Pb levels 2.9-5.9 ppm, the group with tooth Pb
levels >19.9 ppm had fewer correct responses on the CPT and had longer reaction times
for correct responses but did not commit more commission errors (responding to a
nontarget) (p = 0.25 for overall sustain factor). Although blood Pb levels at older ages
were not examined and results do not exclude an effect of more recent Pb exposures, the
combined evidence from these prospective CPT studies of points to an effect on attention
of Pb exposures from earlier in development.
In addition to objective tests of attention, most prospective studies found Pb-associated
attention decrements in school-aged children as rated by teachers, parents, or independent
examiners (Figure 4-9 and Table 4-11). In children ages 7-13 years in Australia, New
Zealand, and Boston, MA, lower ratings of attention were associated with higher lifetime
(to age 11-13 years) average (Burns etal.. 1999). prenatal cord (Leviton et al., 1993). and
tooth (collected at ages 6-8 years) Pb levels (Fergusson et al.. 1993; Leviton et al.. 1993).
The mean blood Pb levels in these populations were 13.9 (ig/dL for lifetime average
(Burns etal.. 1999) and 6.8 (ig/dL for prenatal cord (Leviton etal. 1993). In separate
cohorts, associations of lifetime average blood Pb level (Burns etal.. 1999) and tooth Pb
level (Leviton et al.. 1993) with attention ratings were stronger among girls than boys.
Other prospective studies found lack of association with early childhood or lifetime
average blood Pb levels. Age 30 month blood Pb levels was not associated with teacher
rating of selective inattention in U.K. children ages 7-8 years (Chandramouli et al.. 2009).
Lifetime average blood Pb level was not associated with maternal rating of attention
problems in the Yugoslavia cohort at age 4-5 years (Wasserman et al.. 2001). However,
ratings may be less reliable in younger children, in whom patterns of behavior are less
well established.
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Another strength of prospective studies, both those finding and not finding associations,
was their examination of many potential confounding factors, either by inclusion in
statistical models with Pb or by examination in preliminary analyses. Higher blood or
tooth Pb levels were associated with attention decrements with adjustment for parental
education, IQ, and caregiving quality as well as SES in most studies (Ris et al.. 2004;
Burns et al.. 1999; Fergusson et al.. 1993) and in fewer studies, birth outcomes and
exposures to smoking, alcohol, or drugs. These results increase confidence that the
associations observed with Pb biomarkers reflect a relationship with Pb exposure.
Most cross-sectional studies found associations between higher concurrent blood, tooth,
or hair Pb level and attention decrements among children ages 3-16 years. With analysis
limited to concurrently measured Pb biomarkers, the directionality of effects is difficult
to characterize. While these studies considered potential confounding by SES, child age,
sex, family structure, or parental education, most did not examine potential confounding
by parental caregiving quality. Cross-sectional studies comprised a mix of previously
reviewed and recently published studies and a range of concurrent blood Pb levels, means
1-13.2 (ig/dL (Table 4-11). Among studies of children with mean concurrent blood Pb
levels 11.1-13.2 ng/dL, most (Roy et al.. 2009; Kordas et al.. 2007; Silvaetal.. 1988) but
not all (Liu et al.. 20lib) found associations with lower attention ratings. Associations
with blood or hair Pb levels were found in most of the study populations that lived near
metal sources in Mexico or China (Liuetal. 20lib: Bao et al.. 2009; Kordas et al..
2007). Needleman et al. (1979) found longer reaction time at varying intervals of delay
and higher ratings of distractibility in children in grades 1-2 in Chelsea and Somerville,
MA with tooth Pb levels > 24 ppm than those with tooth Pb <6 ppm (Table 4-11). Some
studies included young children (e.g., ages 3, 4 years), in whom behavioral ratings may
be less reliable (Liuetal.. 20 lib: Roy et al.. 2009).
Evidence of Pb-associated attention decrements in populations with mean concurrent
blood Pb levels 1-6.5 (ig/dL (ages 4-17 years) was more variable. A large group of these
studies considered several potential confounding factors including SES, parental
education, parental IQ, and parental caregiving quality (Plusquellec et al.. 2010; Chiodo
et al.. 2007; 2004; Canfield et al.. 2003b). These studies also had population-based
recruitment and moderate to high follow-up, without indication of biased participation by
children with higher blood Pb levels and poorer attention. However, in these higher-
quality studies of children with lower blood Pb levels, there was a difference in
consistency of evidence between studies of younger children ages 4-5 years and older
children ages 7 years. Chiodo et al. (2007; 2004) examined children ages 7 years in
Detroit, MI and found concurrent blood Pb-associated poorer performance on several
tests of attention on the CPT and lower teacher or study examiner ratings of attention. In
these populations with high prevalence of prenatal alcohol or drug exposure, results did
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not indicate confounding by these other exposures. However, the presence of these other
exposures may limit the generalizability of results.
Among the lower concurrent blood Pb level studies that included younger children, ages
4-5 years, evidence was mixed. An association between higher concurrent blood Pb level
and attention decrements was found among Inuit children ages 5-6 years in Quebec,
Canada (Plusquellec et al., 2010) but not the Rochester cohort (Canfield et al., 2003b).
These results were based on maternal or study examiner ratings. Ratings may be less
reliable in younger children, in whom patterns of behavior are less well established.
Canfield et al. (2003b) initially found associations between higher concurrent blood Pb
level and lower ratings of attention, but they were attenuated with adjustment for child
color and shape knowledge and age 3-year FSIQ, which suggested that poorer knowledge
of the task parameters may increase distraction. However, Cho et al. (2010) found an
association between concurrent blood Pb level and lower attention ratings with
adjustment for child IQ, indicating the relationship between attention and cognitive
function may vary across populations. In the study of Inuit children, concurrent blood Pb
level but not cord blood Pb level was associated with duration of off-task behavior as
rated by study examiners (Plusquellec et al., 2010). In this population with potential co-
exposures to polychlorinated biphenyls (PCBs) and mercury (Hg) from high consumption
offish, blood levels of PCBs and Hg were reported not to be associated with off task
duration, reducing the likelihood of confounding by these co-exposures.
Other studies examining lower concurrent blood Pb levels, means 1.9 and 3.7 (ig/dL,
found associations with attention decrements in children ages 8-12 years as assessed by
CPT or parent/teacher ratings (Cho etal.. 2010: Nicolescuet al.. 2010). Nigg et al. (2008)
did not provide quantitative results but only reported that the association between
concurrent blood Pb level (mean 1 (ig/dL, detection limit=0.3 (ig/dL) and inattention-
disorganization in children ages 8-17 years lost statistical significance with adjustment
for sex and family income. Despite some evidence for attention decrements at relatively
low blood Pb levels, these studies had several limitations in addition to their cross-
sectional design. In these older children, higher earlier childhood Pb exposures may have
contributed to observed associations. These studies adjusted for parental education but
not consistently for other SES factors. They did not consider potential confounding by
parental caregiving quality. Further, sufficient information was not reported to assess
whether participation was biased to those with higher Pb exposure and attention
decrements. Among children ages 8-11 years in Korea, higher concurrent blood Pb level
(mean 1.9 (ig/dL, detection limit=0.58 (ig/dL) was associated with a higher teacher rating
of inattention and more commission errors on the CPT but not other parameters of the
CPT (Figure 4-9 and Table 4-11) (Cho etal.. 2010). Urinary cotinine was more strongly
associated with CPT performance than blood Pb level and the primary cause of
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attenuation of blood Pb effect estimates. Past Pb exposures especially may have had an
influence in the study of children ages 8-12 years in Romania, 55% (n = 46/83) of whom
lived near a former metal processing plant and had higher concurrent blood Pb levels
(mean 5.1 versus 3.2 (ig/dL) (Nicolescu et al.. 2010). Among all subjects, a 1 (ig/dL
increase in concurrent blood Pb level in the 10th to 90th percentile interval (1.8-7.1
(ig/dL) was associated with a 5.2% higher (95% CI: -0.02, 13%) teacher rating of
inattention and a weaker, imprecise 1.5% (95% CI: -4.4, 8.5) higher parent rating. Blood
Pb levels were poorly correlated (r = -0.06, 0.05) with blood levels of aluminum (Al) or
Hg, and neither of these other metals was associated with attention ratings. Results did
not change substantially in an analysis that excluded the 5 children with blood Pb levels
> 10 (ig/dL.
lexicological Studies of Attention
In support of the associations described in the preceding sections for blood Pb level with
attention decrements in children, studies have found Pb-induced decreases in attention in
animals, although results have not been consistent across studies. Although tests in
animals can measure aspects of both attention and impulsivity, behaviors measured with
signal detection tests with distraction can be inferred as predominately assessing
sustained attention. In this test, animals earn food rewards by responding to a target
stimulus and not responding to a distracter light. Poorer sustained attention and greater
distractibility are indicated by lack of response to the target and increased response to the
distracter light, respectively. The 2006 Pb AQCD (U.S. EPA. 2006b) reported
inconsistent effects of Pb exposure in animals on performance in this test. For example,
postweaning Pb exposure that produced blood Pb levels of 16 and 28 ug/dL induced
small decreases in attention in adult rats as indicated by small increases in omission and
commission errors but only during sessions with long intervals between stimuli (Brockel
and Cory-Slechta. 1999a). Lifetime Pb exposure from birth (blood Pb levels 15,
25 (ig/dL) was found to induce distractibility in monkeys at age 9-10 years as indicated
by increased responding to irrelevant cues, i.e., distracting stimuli, in a spatial
discrimination reversal task. Repeated reversal testing revealed that these deficits likely
were not due to sensory or motor impairment (Gilbert and Rice. 1987).
In animals, Pb-induced decrements in attention have been inferred from tests that were
designed to assess attention but that have elicited behaviors that suggest deficits in
attention. For example, a study reported that impaired performance on auditory threshold
tasks in Pb-exposed monkeys was likely due to lack of attention (Laughlin et al.. 2009).
Rhesus monkeys were exposed to Pb acetate from gestation (drinking water of mothers,
3 months prior to mating) to birth or postnatally from birth to age 5.5 months at weaning
and had resultant bone Pb levels at 11 years of 7 and 13 (ig/g for prenatal and postnatal
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groups, respectively, and average blood Pb levels during Pb exposure of 35 and
46 (ig/dL, respectively. Animals were tested at age 13 years when blood Pb levels had
returned to baseline levels. The inability of some of the monkeys to engage or focus
attention on the task at hand yielded fewer available measurements in Pb-exposed
animals versus controls. These observations were made in monkeys with higher peak
blood Pb levels than those relevant to this ISA.
Epidemiologic Studies of Impulsivity in Children
Compared to attention, measures specific to impulsivity have been examined in few
epidemiologic studies of children and mostly in recent cross-sectional studies. However,
evidence in children indicates Pb-associated poorer performance on tests of response
inhibition which finds coherence with similar evidence in animals. Response inhibition is
a measure of impulsivity and in children has been assessed with stop signal tasks, which
measures the execution of action in response to stimuli and the inhibition of that action
when given a stop signal. Recent cross-sectional studies found that children with higher
concurrent blood Pb levels had greater responses with stop signals. Among children ages
8-12 years in Romania, a 1 (ig/dL increase in concurrent blood Pb level in the 10th-90th
percentile interval 1.8-7.1 (ig/dL was associated with a 10% (95% CI: 0, 16) increased
false-alarm rate in responses to stop signals with adjustment for city, sex, age, computer
experience, handedness, eye problems, number of siblings, parental education, prenatal
alcohol and smoking exposure, and parental self report of ever having a psychopathology
(Nicolescu et al.. 2010). Children in one town lived near a metal processing plant;
however, associations were not found with blood Pb levels of Al and Hg. In a case-
control study of children with ADHD, Nigg et al. (2008) found that higher concurrent
blood Pb level was associated with poorer response inhibition on a stop task
A few studies found associations between higher concurrent tooth (Needleman et al..
1979) or blood (Tlusquellec et al.. 2010: Nigg et al.. 2008) Pb level with higher ratings of
impulsivity. Nigg et al. (2008) found associations of concurrent blood Pb levels with
parent and teacher ratings of a composite hyperactivity/impulsivity index in a group of
children (location not reported) with and without ADHD (ages 8-17 years) with a mean
blood Pb level of ~1 (ig/dL (detection limit = 0.3 (ig/dL). Regression-based path analysis
showed that the association between blood Pb level and hyperactivity-impulsivity ratings
was mediated by (i.e., occurred through) poorer performance on the more objective stop
task. Path analysis produces partial regression coefficients that measure the effect of one
variable on another controlling for previous variables. Because hyperactivity and
impulsivity are characteristics of ADHD, ratings likely are higher among ADHD cases
than controls. If the differential reporting of hyperactivity/impulsivity by ADHD status is
unrelated to Pb exposure or blood Pb level, it is less likely to bias associations with Pb.
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However, differential rating between cases and controls could produce an artifactual
association with blood Pb level if participation by children with ADHD is biased to those
with higher Pb exposures and blood Pb levels. Nigg et al. (2008) also used path analysis
to show that the Pb-associated increase in hyperactivity/impulsivity was independent of
an association with IQ. With adjustment for sex and income, concurrent blood Pb level
was directly associated with hyperactivity/impulsivity, and the association was not
completely mediated by the blood Pb-IQ association. Instead, the association between
blood Pb level and IQ was found to be mediated by the association of Pb with
hyperactivity/impulsivity. In a population of Inuit children at ages 5-6 years, the
relationship between concurrent blood Pb level and motor function was found to be
mediated by (i.e., occur through) the association of blood Pb level with impulsivity
(Fraser et al.. 2006).
These studies of impulsivity were similar in their adjustment for potential confounding by
SES, sex, and parental education (Table 4-11). A few adjusted for smoking exposure.
However, parental IQ or caregiving quality were not examined in most studies.
Plusquellec et al. (2010) has the most extensive consideration for potential confounding,
excluding parental IQ, HOME score, micronutrient levels, Hg, and PCBs (potential co-
exposures from high consumption offish) from final models with blood Pb based on their
lack of strong association with impulsivity. There also was no biased participation
according to the exposures examined. A limitation of this study is the young age of some
children at examination (study range: age 5-6 years). The variable nature of behaviors in
younger children may limit the implications of findings.
lexicological Studies of Impulsivity
The associations described between higher blood Pb level and greater impulsivity in
children are supported by findings in animals for Pb-induced increased perseveration and
impaired ability to inhibit inappropriate responses. In animals, these effects are supported
by studies reviewed in the 1986 and 2006 Pb AQCDs (U.S. EPA. 2006b. 1986b) and
recent studies. Animal studies provide more consistent evidence for the effects of Pb
exposure on impulsivity in animals than on sustained attention. As mentioned earlier,
behaviors displayed by animals in a variety of tests can be identified as predominately
reflecting impulsivity. These include tests of Differential Reinforcement of Low Rates of
Responding (DRL), FI schedule performance, FI with Extinction, or Fixed Ratio
(FR)/waiting-for-reward. Greater impulsivity is indicated by premature responses,
decreased pause time between two scheduled events, and increased perseveration.
Behaviors observed in tests of operant conditioning with FI reinforcement schedules also
have been used to indicate impaired learning in animals (Section 4.3.2.3), and the
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interactions observed between Pb exposure and maternal or offspring stress also may
apply to effects on impulsivity. Maternal exposure to 150 ppm Pb with and without stress
co-exposure was found to increase overall FI rate and decrease PRP in rats (Table 4-7).
Lifetime (from gestation) Pb exposure resulting in blood Pb levels 11-16 (ig/dL increased
overall FI rate without stress co-exposure and decreased PRP with stress co-exposure
(Table 4-7) (Rossi-George et al.. 2011). Discrimination reversal learning has been shown
to be affected by Pb exposure. In these tasks, an animal is trained to choose between two
alternative responses and then is required to reverse that association. Perseveration or
lack of inhibition of the original response can be interpreted to involve impulsivity.
Spatial and non-spatial discrimination reversal was significantly affected in monkeys
after Pb exposure during infancy, after infancy, or continuously from birth, and was
exacerbated with distracting stimuli (Rice. 1990; Rice and Gilbert. 1990b; Gilbert and
Rice. 1987). These monkeys had blood Pb levels in the range of 15-36 (ig/dL, which
includes those relevant to this ISA. Contrary to the findings in monkeys, Hilson and
Strupp (1997). found Pb exposure (Pb acetate in drinking water GD1-PND28, blood Pb
levels 26, 51 (ig/dL) in rats to slow reversal learning in an olfactory discrimination task.
But, analysis of the response patterns showed that Pb exposure shortened the
perseverative responding phase of reversal learning and lengthened the post-perseverative
phase of chance responding, indicating impairments in associative ability not response
inhibition.
The effects of Pb exposure on impulsivity also have been demonstrated as shorter time
Pb-exposed animals will wait for reward in FR/waiting for reward testing. In this test,
animals can obtain food by pressing a lever a fixed number of times (FR component).
Then, free food is delivered at increasingly long time intervals, so long as the animal
inhibits additional lever presses. Animals can reset the schedule to return to the FR
component at any time. Brockel and Cory-Slechta (1998) exposed male Long-Evans rats
to 0, 50, or 150 ppm Pb acetate in drinking water from weaning, which produced
respective blood Pb levels of <5, 11, and 29 ug/dL after 3 months of exposure. After
40 days of exposure, the 150 ppm Pb-exposed rats responded more quickly in the FR
component and reset the schedule (thus shortening the waiting period) more often than
did the 50 ppm Pb-exposed rats and controls. In the waiting component, average wait
time was significantly lower in both Pb exposure groups compared to controls. The rats
exposed to 150 ppm Pb also had higher response rates, earned more reinforcers per
session, but had a higher response-to reinforcement-ratio than did the 50 ppm Pb group
and controls, which indicated less efficient responding. Mechanistic understanding of the
Pb-induced effects on responding in the FR/waiting for reward test was provided by a
study with similar postweaning exposure to 0, 50, and 150 ppm Pb that yielded respective
blood Pb levels of <5, 10, and 26 ug/dL after 3 and 7 months of exposure. Administration
of a dopamine (D2) receptor agonist reversed the Pb-induced effects on increasing FR
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response rates and decreasing mean waiting time, suggesting a role for dopamine-like
receptors in Pb-induced effects on waiting behavior (Brockel and Cory-Slechta. 1999b).
Pb-induced impulsivity appears to be related to emotionality, which was found in rats
trained to perform an olfactory discrimination task, albeit at higher Pb exposures than
those relevant to this ISA. Rats were given early postnatal Pb exposure (300 ppm
Pb acetate via dam drinking water PND1-PND17 then either 20 or 300 ppm in their own
drinking water PND18-PND30), which produced blood Pb levels of 40-60 and
100-140 (ig/dL. The rats were tested as young adults on a food-motivated olfactory
discrimination task in which rewards for correct responses were occasionally and
unpredictably omitted. Pb-exposed animals were more sensitive both to their own errors
and to reward omission than controls, suggesting a lowered capacity for regulating
arousal and emotion. Administration of the chelator succimer after the Pb exposure
period (PND31-PND52) normalized reactivity to reward omission and errors in the
Pb-treated rats, but increased reactivity in the control animals (Beaudin et al.. 2007).
The same laboratory using the same Pb exposure protocol also found heightened
reactivity to errors in tests of visual discrimination and visual sustained attention (Stangle
et al.. 2007). Pb-induced increases in impulsivity were found in rats at PND62 after
exposure to 300 ppm Pb (blood Pb level 31 (ig/dL at PND52) but not 20 ppm Pb (blood
Pb level 13 (ig/dL). Impulsivity was inferred from a higher number of premature
responses after the basic rules of the task had been learned. Succimer reduced blood Pb
level to 3-9 (ig/dL at PND52 but did not reduce premature responding at PND62,
indicating persistence of the effect. Finally, monkeys exposed to Pb from birth to time of
testing at age 3-4 years with blood Pb levels 15 and 25 (ig/dL showed slower acquisition
of a DRL schedule, which requires an animal to respond at a low rate for rewards.
However, the effect was reversible. Pb-exposed monkeys eventually acquired
reinforcement rates equal to those of controls (Rice and Gilbert. 1985).
In summary, several studies in animals indicate that Pb exposure of rodents and non-
human primates from birth or after weaning changes behavior in ways consistent with
increased impulsivity, primarily as indicated by impaired response inhibition. Some
observations of Pb-induced impulsivity in animals were made with blood Pb levels
considered relevant for this ISA. The observations for Pb-induced increases in
impulsivity in animals provide support for associations found in children of higher blood
and tooth Pb levels with lower response inhibition and higher ratings of impulsivity.
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Hyperactivity
Studies reviewed in the 2006 Pb AQCD (U.S. EPA. 2006b) indicated associations
between higher concurrent blood Pb level or tooth Pb level and higher parent or teacher
ratings of hyperactivity in children ages 6-11 years in the U.S., Asia, and New Zealand
(Rabinowitz et al.. 1992; Silvaetal.. 1988; Gittleman and Eskenazi. 1983; Needleman et
al.. 1979; David et al.. 1976). The case-control or cross-sectional design of studies limited
understanding of the temporal sequence between Pb exposure and hyperactivity. Several
recent studies, including a prospective study (Chandramouli et al.. 2009). also found
associations between blood Pb level and hyperactivity as rated by teachers and parents
(Figure 4-9 and Table 4-11). Overall, studies indicated associations in children ages 3-12
years with mean concurrent blood Pb levels 3.7-12 (ig/dL.
The recent prospective study of children in the U.K. helps address some of the limitations
of cross-sectional studies by demonstrating an association between higher early
childhood (age 30 months) blood Pb level and higher teacher ratings of hyperactivity
later in childhood at age 7-8 years and by adjusting for several potential confounding
factors, including SES, home facilities score, and family adversity index (Chandramouli
et al.. 2009). In addition to the prospective design, the study had high participation rates
at baseline and follow-up from a population with similar characteristics as reported in the
U.K. census. Increases in hyperactivity were found primarily in the group of children
with blood Pb levels >10 (ig/dL and were independent of associations with IQ.
The cross-sectional evidence is provided by previous and recent studies conducted in
populations with a range of blood Pb levels and with variation in the extent of
consideration for potential confounding. A few studies found concurrent blood Pb-
associated higher ratings of hyperactivity in populations with mean concurrent blood Pb
levels 5-6 (ig/dL and examined potential confounding by a large number of factors.
Among children ages 5-6 years in Quebec, Canada, Plusquellec et al. (2010) found that
hyperactivity (assessed as observer rating and computer program) was related to
concurrent blood Pb level but not HOME score, caretaker IQ or education, or other
prevalent exposures in the community including Hg and PCBs. Among children age 7
years in Detroit, MI, hyperactivity was related to concurrent blood Pb levels but not SES,
parental education or IQ, or HOME score (Chiodo et al.. 2007). While there is confidence
in the internal validity of results from these studies, the applicability to other populations
may be limited. The former study included younger children in whom behavioral patterns
are less fixed, and the latter study examined a population with high prevalence of prenatal
drug exposure.
Other cross-sectional studies found Pb-associated increases in hyperactivity but examined
fewer potential confounding factors. Adjustment for SES, maternal education and IQ was
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common; however, few adjusted for parental caregiving quality. Associations were found
with tooth Pb in children in grades 1-3 (Rabinowitz et al.. 1992; Needleman et al.. 1979).
Several (Roy et al.. 2009; Chen et al.. 2007; Silvaetal.. 1988) but not all (Liu et al..
20 lib) studies found associations in populations of children ages 3-11 years in the U.S.,
New Zealand, and Asia with relatively high concurrent blood Pb levels (means 10-12
(ig/dL). Nicolescu et al. (2010) found an association in children ages 8-12 years in
Romania with relatively lower concurrent blood Pb levels, median 3.7 (ig/dL. Results
were adjusted for several potential confounding factors including age, sex, parental
education, and prenatal smoking or alcohol exposure. Hyperactivity was not related to Al
or Hg, other prevalent exposures in the community. However, potential confounding by
other SES variables or parental caregiving quality was not examined.
A recent study did not clearly indicate an effect of Pb on inducing hyperactivity in mice.
Pb exposure during gestation and the early postnatal period (PND10) (27 and 109 ppm Pb
acetate: < 10 and 42 (ig/dL peak blood Pb levels, respectively) increased motor activity of
male mice at age 1 year with co-treatment with amphetamines but not female mice
(Leasure et al.. 2008). Without amphetamines, Pb induced less activity of mice, and the
lower Pb dose inhibited activity more than the higher Pb dose did.
Ratings of Attention Deficit Hyperactivity Disorder-related Behaviors
In addition to finding associations with attention, impulsivity, and hyperactivity, some of
the recent epidemiologic studies described in the preceding sections found associations
between higher concurrent blood Pb level and higher parental and teacher ratings of
ADHD-related behaviors (Cho etal. 2010: Nicolescu etal.. 2010: Froehlich et al.. 2009:
Roy et al.. 2009). calculated as a composite of the various behaviors that are evaluated in
the diagnosis of ADHD. The strengths and limitations of these studies have been
described in the preceding sections. Main limitations were the cross-sectional design,
lack of validation of ADHD ratings with a clinical diagnosis, and lack of consideration
for some potentially important confounding factors. Studies considered age, sex, SES or
parental education but usually not both, and none considered parental caregiving quality.
Thus, the evidence specifically for these total ADHD index ratings were emphasized less
than evidence for individual behaviors in drawing conclusions about the effects of Pb
exposure on attention, impulsivity, and hyperactivity.
The large, U.S. representative analysis of children participating in NHANES 2001-2004
found an association between concurrent blood Pb level in children ages 8-15 years and
parental assessment of child ADHD-related behaviors using the Diagnostic Interview
Schedule for Children which uses DSM-IV criteria to identify children at increased risk
of meeting diagnostic criteria for ADHD (Froehlich et al.. 2009). Compared with children
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with concurrent blood Pb levels <0.8 (ig/dL (4% below detection limit of 0.3 (ig/dL),
children with concurrent blood Pb levels >1.3 (ig/dL had an elevated odds of parentally-
rated ADHD-related behaviors with an OR of 2.3 (95% CI: 1.5, 3.8). These results were
adjusted for prenatal and current household smoking exposure, sex, age, race/ethnicity,
income to poverty ratio, preschool attendance, maternal age, and birth weight. A similar
OR was estimated in analyses restricting the third tertile to children with concurrent
blood Pb levels 1.3-5.0 (ig/dL. The strongest association was observed in children with
both high blood Pb level and prenatal tobacco smoke exposure. Compared to children
with blood Pb levels <0.8 (ig/dL and no prenatal exposure to tobacco smoke, children
with blood Pb levels >1.3 (ig/dL and prenatal exposure to tobacco smoking had the
highest odds of parentally-rated ADHD-related behavior (OR: 8.1 [95% CI: 3.5, 18.7]).
Although ADHD-related behavior was associated with low concurrent blood Pb levels
(1.3-5 (ig/dL), the contribution of higher past Pb exposures of adolescents born in the late
1980s cannot be excluded. Roy et al. (2009) also found an association with teacher
ratings of ADHD-related behaviors using DSM criteria in children in Chennai, India;
however, the study population included some very young children (i.e., age 3 years) and
had relatively high concurrent blood Pb levels (mean: 1 1.4 (ig/dL).
Other recent cross-sectional studies found concurrent blood Pb-associated higher ratings
of ADHD-related behaviors using instruments that do not follow DSM criteria. Among
children ages 8-1 1 years in Korea, Cho et al. (2010) found a stronger relationship with a
total ADHD-related behaviors index as rated by teachers than parents. Mean ADHD
ratings by teacher and parents were similar (both 9. 1); however, parental ratings had
greater variability (SD: 1 1 .5 for parents and 8.6 for teachers), which could have
contributed to differences in association. Among children in Romania, concurrent blood
Pb level was associated similarly with parent and teacher ratings of ADHD-related
behaviors (Nicolescu et al.. 2010). As with individual behaviors described in preceding
sections, blood Al and Hg levels were not associated with ratings of ADHD-related
behaviors. Based on a log-linear model, a 1 (ig/dL increase in concurrent blood Pb level
within the 10th-90th percentile interval (1.8-7.1 (ig/dL) was associated with a 5.9%
increase (95% CI: 0, 13.7) in teacher rating of ADHD-related behavior. The association
did not change substantially in an analysis that excluded the 5 children with blood Pb
levels > lO
Attention Deficit Hyperactivity Disorder in Children
The 2006 Pb AQCD (U.S. EPA. 2006b) did not review studies of prevalence or incidence
of ADHD diagnosis but noted lack of conclusive evidence for the effect of Pb exposure
on ADHD based on a few small studies comparing blood or urine Pb levels between
children with and without hyperactivity as identified by parents, teachers, or schools
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(Gittleman and Eskenazi. 1983; David etal.. 1972). As described in the preceding
section, several recent cross-sectional studies found associations of concurrent blood Pb
level with parent and teacher ratings of a total ADHD index, a composite index of
attention, impulsivity, and hyperactivity. Results from a small body of recent studies also
indicate associations of higher concurrent blood Pb level with prevalence of diagnosed
ADHD in children ages 4-17 years (Nigg etal.. 2010: 2008: Wang et al.. 2008d: Braun et
al.. 2006). All of the studies were cross-sectional; thus, the temporal sequence between
Pb exposure and ADHD incidence cannot be established. Further, while there are
diagnostic guidelines for ADHD, the exact criteria or specific behaviors required can
vary. Thus, within studies, there may be variations among subjects in the types of
behaviors they displayed that led to a diagnosis of ADHD. While there is coherence with
evidence from prospective studies in other populations for associations of Pb biomarker
levels with characteristic behaviors, evidence specifically for ADHD prevalence was
emphasized less than evidence for individual behaviors in drawing conclusions about the
effects of Pb exposure on attention, hyperactivity, and impulsivity.
Associations between concurrent blood Pb level and ADHD prevalence were found in
case-control studies conducted in different populations of children. While a limitation of
these studies is potential selection bias arising from the nonrandom population sample, a
common strength is their independent diagnosis of ADHD in a structured manner using
parental and teacher ratings of behavior followed by independent assessment by multiple
clinicians using DSM-IV criteria (Nigg etal.. 2010; 2008; Wang et al.. 2008d). Nigg et
al. (2010; 2008) found an association between concurrent blood Pb level and ADHD
diagnosis in relatively small (n = 150, 236) groups of children ages 6-17 years from the
same community, with controls selected from healthy children who responded to the
same mailings, community advertisements, or outreach to local clinics as cases. Wang et
al. (2008d) found an association in a larger (n = 1,260) group of children in China, with
controls selected from children admitted as outpatients to the same two pediatric clinics
for respiratory infections.
Braun et al. (2006) found an association in children ages 4-15 years participating in
NHANES 1999-2002. ADHD was ascertained by parent-report of ADHD diagnosis or
use of stimulant medication, which is subject to reporting bias; however, the examination
of multiple risk factors and outcomes in NHANES reduces the likelihood of biased
participation and reporting of ADHD by parents of children specifically with higher Pb
exposure. NHANES is not a random sample, but a strength over other studies that
examined the prevalence of ADHD diagnosis is the large (n = 4,704) sample size and the
nationally-representative results produced with adjustment for sampling weights in
models. Surveillance data indicate that states with a higher percentage of children with
blood Pb levels > 10 (ig/dL have lower prevalence of diagnosed ADHD (CDC. 2012.
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201 Ic). These data do not support the potential for positive confounding of associations
observed in the NHANES population by regional differences in blood Pb levels and
prevalence of ADHD diagnosis. However, there is the potential for attenuation of
associations with blood Pb levels if adjustment is not made for region.
With respect to blood Pb levels associated with ADHD diagnosis, analyses of the
concentration-response relationship indicated monotonic increases in ORs across blood
Pb level groups (Wang et al.. 2008d: Braun et al.. 2006). In the analysis of children in
NHANES ages 4-15 years, compared to children with concurrent blood Pb level
<0.8 (ig/dL (7% below detection limit of 0.3 (ig/dL), children with concurrent blood Pb
level >2.0 (ig/dL (maximum not reported) had higher prevalence of ADHD with an OR
of 4.1 (95% CI: 1.2, 14.0). A similar OR was estimated for children with blood Pb levels
2.0-5.0 (ig/dL (Braun et al.. 2006). In the study of children in China, the highest OR was
found in children ages 4-12 years with concurrent blood Pb levels > 10 (ig/dL but also
was elevated in the group with blood Pb levels 5-10 (ig/dL (OR: 4.92 [95% CI: 3.47,
6.98] compared with children with blood Pb level <5 (ig/dL) (Wang et al.. 2008d).
Similar to findings in NHANES, Nigg et al. (2010; 2008) found associations at relatively
low concurrent blood Pb levels, i.e., population means ~1 (ig/dL (1.3%, 0% below
detection limit of 0.3 (ig/dL). However, because adolescents were examined, higher past
Pb exposures could have contributed to the observed associations. Blood Pb levels are
higher in early childhood, and among children participating in NHANES who were born
1984-1998, some likely had higher early-life Pb exposures as a results of the use of
leaded gasoline in the U.S. (Braun et al.. 2006).
Consideration for potential confounding varied among studies. In three-way analyses of
covariance, Nigg et al. (2008) adjusted for sex and household income, and Nigg et al.
(2010) adjusted for maternal IQ and prenatal smoking exposure. However, in preliminary
analyses, Nigg et al. (2010) considered blood hemoglobin, household income, age, sex,
and maternal smoking as potential confounding factors. The analysis of children
participating in NHANES adjusted for age, sex, race, prenatal smoking exposure,
postnatal smoker in the home, preschool/child care attendance, health insurance
coverage, and ferritin levels but initially considered poverty to income ratio, birth weight,
and admission to the neonatal intensive care unit (Braun et al.. 2006). The results for
children in China were adjusted for similar covariates and also family (parent and sibling)
history of ADHD diagnosis, ascertained from clinical records (Wang et al.. 2008d).
Family history of ADHD was selected as a covariate based on its association with child
ADHD; no information was provided on its association with child blood Pb level. None
of the studies of ADHD prevalence considered potential confounding by current parental
caregiving quality.
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4.3.3.2 Conduct Disorders
Epidemiologic Studies of Conduct Disorders in Children
As described in the introduction to externalizing behaviors, there are two domains of
conduct disorders: undersocialized aggressive conduct disorder and socialized aggressive
conduct disorder (as reviewed inWhitcomb and MerrelL 2012). In this evaluation of a
relationship with Pb exposure, the two domains are combined because they cannot be
easily distinguished in the available epidemiologic studies that combine various
behaviors into a composite index, find associations with behaviors within each domain,
or examine outcomes characterized by behaviors in each domain. This section also
evaluates the evidence for criminal offenses which can be predicted by earlier conduct
disorders (Soderstrom et al.. 2004: Babinski et al.. 1999: Paier. 1998).
The 2006 Pb AQCD (U.S. EPA. 2006b) described several prospective studies that found
associations of higher blood, tooth, or bone Pb levels with behaviors related to conduct
disorders in children as rated by parents or teachers or criminal offenses in adolescents
(also see Table 4-12). Supporting evidence from recent prospective studies included
follow-up of previous cohorts to older ages. Recent cross-sectional studies found
associations between concurrent blood Pb level and ratings of behaviors related to
conduct disorders, but aside from difficulty in establishing temporality, these recent
cross-sectional studies were limited by their less extensive consideration for potential
confounding.
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Table 4-12 Associations between Pb biomarker levels and behaviors related to conduct disorders in children
and young adults.
Study
Study Population and Methodological Details
Organized by outcome category. Within each category, results are
grouped by strength of study methodology, generalizability3
Blood (ug/dL),
Tooth, or Bone (ug/g)
Pb Metrics Analyzed
Outcomes Analyzed
Effect Estimate
(95% Cl)b
Prospective Studies of Behavioral Ratings
Dietrich et al.
(2001)
Burns et al.
(1999)
186 children followed prenatally (1979-1985) to ages 15-17 yr,
Cincinnati, OH
Recruitment at prenatal clinic. Moderate follow-up participation
but no selective attrition. Mostly African-American. Linear
regression model adjusted for HOME, parental IQ, current
SES, birth weight. Also considered potential confounding by
maternal age, prenatal smoking, alcohol use, and marijuana
use, other birth outcomes, ear infections, Fe status, sex, age,
caregiver education, public assistance, preschool attendance,
# children and adults in home.
322 children followed from birth (1979-1982) to ages 11-1 Syr,
Port Pirie, Australia.
Moderate follow-up participation. Participants had higher birth
weight, older mothers, less educated fathers. Log linear
regression model adjusted for maternal age, prenatal smoking
status, IQ, concurrent psychopathology, and education, birth
weight, type of feeding, length of breastfeeding, paternal
education and occupation, birth order, family functioning,
parental smoking, marital status, HOME score, child IQ.
0-6 yr avg blood
Median (5th-95th):
12(6.0-26)
See Wright et al.
(2008)
Lifetime (age 11-13 yr)
avg blood
GM (5th-95th)
Boys 14.3(13.5-15.1)
Girls: 13.9(13.2-14.6)
Intervals analyzed
(10th-90th percentiles)
Boys: 13.7-14.9
Girls: 13.4-14.5
Self-Report of Delinquent
Behavior Score
Parental Report of
Predelinquent and
Delinquent Behavior
Score
A AC A ~7 \ iv
Ages 15-17 yr
Aggressive Score, boys
Aggressive Score, girls
Delinquent score, boys
Delinquent score, girls
Maternal rating by Child
Behavior Checklist at
ages 11-13 yr
0.10(0.01, 0.19)
0.09 (-0.02, 0.20)
0.17(0.08, 0.26)
0.10(0, 0.21)
0.06(0.02, 0.09)
0.01 (-0.01, 0.04)
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Table 4-12 (Continued): Associations between Pb biomarker levels and behaviors related to conduct disorders in children and
young adults.
Study
Chandramouli
et al. (2009)
Needleman et
al. (1996)
Wasserman et
al. (2001)
Study Population and Methodological Details
Organized by outcome category. Within each category, results are
grouped by strength of study methodology, generalizability3
488 children followed from age 4 mo (born 1992) to ages 8 yr,
Avon, U.K.
All births in area eligible. Similar characteristics as U.K.
census, high participation at baseline and follow-up.
Participants had better educated mothers, who smoked less,
better home environment. Log linear regression model
adjusted for maternal education and smoking, home
ownership, home facilities score, family adversity index,
paternal SES, parenting attitudes at 6 mo, child sex. Also
considered potential confounding by child IQ.
212 boys selected from prospective cohort followed from 1st
grade to ages 11-14 yr, Pittsburgh, PA
Nested case-control. Low follow-up participation. Participants
had higher SES, lower maternal IQ, smaller family size, higher
IQ, were nonwhite. ANCOVA with behavior scores adjusted for
maternal age, IQ, occupation, and education, presence of both
parents in home, # children in family, race, history of medical
problems, age, score at age 7 yr. Did not consider potential
confounding by parental caregiving quality.
191 children followed prenatally (1984-1985) to ages 4-5 yr,
Pristina, Yugoslavia.
Recruitment from prenatal clinics. High follow-up participation,
participants had less educated mothers, higher average blood
Pb, were Albanian. Log linear regression model adjusted for
sex, ethnicity, age, maternal education and smoking history,
HOME score, birth weight.
Blood (ug/dL),
Tooth, or Bone (ug/g)
Pb Metrics Analyzed
Age 30 mo blood
Mean (SD): NR
Group 1: 0-<2
Group 2: 2-<5
Group 3: 5-<10
Group 4: >10
Bone at ages 9-14 yr:
NR
High bone Pb: upper
bound of quintile 3 and
higher
Low bone Pb: rest of
quintile 3 and lower
Exact levels NR
Lifetime (to age 4-5 yr)
avg blood
Mean of log= 0.86,
Mean -7.2
Outcomes Analyzed
Antisocial activities
Parent or teacher rating
by Antisocial Behaviour
Interview at age 8 yr
Delinquent score (square
root)
Aggressive score (square
root)
Parent rating by Child
Behavior Checklist at age
11-14yr
Aggressive Score
Delinquent Score
Maternal rating by Child
Behavior Checklist at
Effect Estimate
(95% Cl)b
ORs for increased
score
Reference
0.93(0.47, 1.83)c
1.44(0.73, 2.84)c
2.90(1.05, 8.03)c
LowPb: 1.18
High Pb. 1.45, p-0.04
Low Pb: 2.43
High Pb: 2.98, p=0.009
Per log increase:
0.08 (-0.19, 0.35)
0.32(0.03, 0.61)
ages 4-5 yr
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Table 4-12 (Continued): Associations between Pb biomarker levels and behaviors related to conduct disorders in children and
young adults.
Study
Study Population and Methodological Details
Organized by outcome category. Within each category, results are
grouped by strength of study methodology, generalizability3
Blood (ug/dL),
Tooth, or Bone (ug/g)
Pb Metrics Analyzed
Outcomes Analyzed
Effect Estimate
\b
(95% Cl)°
Bellinger et al. 1,782 children followed from birth (1979-1980) to age 8 yr,
(1994b) Boston, MA area
Recruitment at birth hospital. Moderate follow-up participation.
More participants were white, had lower cord blood Pb levels,
better birth outcomes. Log linear regression model adjusted for
prepregnant weight, race, Cesarean section, maternal marital
status, prenatal care, paternal education, colic, child current
medication use, sibship size, sex, birth weight. Also
considered potential confounding by public assistance,
prenatal smoking, maternal education but not parental
caregiving quality.
Tooth (age 6 yr)
Mean (SD):
3.4(2.4)
Interval analyzed:
1.2-6.3 = 10th-90th
percentiles
Total externalizing
behavior T score,
(inattentive, nervous-
overactive, aggressive)
Teacher rating by Child
Behavior Profile at age 8
yr
0.51 (0.21, 0.81)
Chen et al. 721 children participating in TLC trial, followed from age 1-3 to
(2007) 5 yr, Baltimore, MD; Cincinnati, OH; Newark, NJ; Philadelphia,
PA.
Multi-city, high participation rate. Eligible for chelation trial
because of high age 1-3 yr blood Pb levels (20-44 ug/dL).
Linear regression-based path analysis adjusted for city, race,
sex, language, single parent, parental education and
employment, age at blood Pb measurement, caregiver IQ.
Considered potential confounding by chelation treatment but
not parental caregiving quality. Direct = independent of IQ.
Indirect = mediated through IQ.
Age 2 year
Mean(SD): 26(5.1)
Interval analyzed: 20-
33=10th-90th
percentiles
Oppositional Index Score
Direct
Indirect
0.12 (-0.08, 0.31)
0.04(0.002, 0.07)
Concurrent
Mean(SD): 12(5.2)
Interval analyzed: 6.6
(10th percentile)-10
Direct
Indirect
0.12 (-0.08, 0.32)
0.05 (-0.03, 0.13)
Cross-sectional studies of Behavioral Ratings
Braun et al. 2,867 children ages 8-15 yr (born 1986-1996), U.S. NHANES
(2008) 2001-2004
Large multi-location study of multiple risk factors and
outcomes. Subjects with available data were older, white,
higher SES, with lower blood Pb levels, higher birth weight,
and fewer had household smokers. Logistic regression model
adjusted for child age, sex, and race, poverty income ratio,
maternal age, prenatal smoke exposure, cotinine levels. Did
not consider potential confounding by parental caregiving
quality.
Concurrent blood
Q1: 0.2-0.7
Q2: 0.8-1.0
Q3: 1.1-1.4
Q4: 1.5-10
Detection limit = 0.3
Conduct disorder
Parental report using
Diagnostic Interview
Schedule for Children-
Caregiver Module at ages
8-1 Syr
ORs (yes/no)
Q1: Reference
Q2: 7.2(1.1,49)c
Q3: 12(2.4, 65)c
Q4: 8.6(1.9, 40)c
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Table 4-12 (Continued): Associations between Pb biomarker levels and behaviors related to conduct disorders in children and
young adults.
Study
Chiodo et
(2QQ7)
Study Population and Methodological Details
Organized by outcome category. Within each category, results are
grouped by strength of study methodology, generalizability3
al. 452-460 African-American children, age 7 yr (born
1989-1991), Detroit, Ml area
Recruitment at prenatal clinic. High prevalence prenatal drug
exposure. High participation rate. Linear regression model
adjusted for sex (both outcomes), caretaker education, HOME,
maternal prenatal alcohol use, current marijuana use
(delinquent behavior), maternal age, # children in home (social
problems). Also considered potential confounding by SES,
child age, maternal prenatal and current drug and alcohol use,
and IQ, caretaker current psychopathology.
Sciarillo et al. 201 children (born 1984-1987) ages 2-5 yr, Baltimore, MD.
(1992)
Nigg et al
(2008)
Studies
Wright et
(2008)
Convenience sample. No information on participation rate.
100% African-American. Linear regression adjusted for
maternal education, employment status, marital status, current
depressive symptom score, # preschool children in the home,
child age, sex, and Fe deficiency.
150 children, ages 8-17 yr, birth yr and location NR.
Case-control study of ADHD. Recruitment by community
advertisements, mailings, outreach to clinics. Moderate
participation rate from eligibles. Could have biased
participation by Pb exposure leading to biased behavioral
ratings. Did not consider potential confounding factors.
of Criminal Offenses
al. 250 adults followed from birth (1979-1984) to age 19-24 yr,
Cincinnati, OH
Prospective. Recruitment at prenatal clinic. 90% African-
American. Moderate follow-up participation. No selective
attrition. Negative binomial regression model adjusted for
maternal IQ and education, sex, SES. Also considered
potential confounding by maternal prenatal smoking,
marijuana use, narcotic use, and prior arrests, HOME score,
birth weight, # children in the home, public assistance in
childhood.
Blood (ug/dL),
Tooth, or Bone (ug/g)
Pb Metrics Analyzed
Concurrent blood
Mean (SD):
5.0(3.0)
Interval analyzed:
2.1-8.7 = 10th-90th
percentiles
Concurrent blood
Mean (SD)
Low: 9.2 (2.9)
High: 28 (10)
Interval analyzed: 5.9
(10th percentile of low
group)-10
Concurrent blood
Mean (SD):
1.03(0.49)
Detection limit=0.3
Age 6 yr blood Pb
Median (5th-95th):
6.8(3.4-18)
Age 0-6 yr avg
Median (5th-95th):
A f~\ ic n I~)C\
IZ (D.U-ZD)
Outcomes Analyzed
Delinquent behavior
Social problems
Teacher rating by
Achenbach Teacher
Report Form at age 7 yr
Total Behavioral Problem
score (aggressive,
destructive, somatic
problems, sleep problems,
depressed, social
withdrawal, etc)
Maternal rating by Child
Behavior Checklist at
ages 2-5 yr
Oppositional defiant
disorder Index
Parent, teacher rating
Conners Rating Scale-
Revised
Criminal arrests From
county records at ages
1 9-24 yr
Age 6 yr
Age 0-6 avg
Effect Estimate
(95% Cl)b
0.09(0, 0.1 8)d
0.10(0, 0.20)d
0.18(0.04, 0.32)
Pearson r =
0.18, p<0.05
RRs (yes/no):
1.05(1.01, 1.09)
1.01 (0.98, 1.05)
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Table 4-12 (Continued): Associations between Pb biomarker levels and behaviors related to conduct disorders in children and
young adults.
Study
Fergusson et
al. (2008)
Needleman et
al. (2002)
Study Population and Methodological Details
Organized by outcome category. Within each category, results are
grouped by strength of study methodology, generalizability3
853 children followed from birth (1977) to age 21 yr,
Christchurch, New Zealand
Prospective. High follow-up participation. Participants had
lower SES. Negative binomial regression model adjusted for
maternal education and prenatal cigarettes/ day, ethnicity,
family conflict, physical abuse in childhood. Also considered
potential confounding by traffic density in childhood, maternal
age, paternal education, average family income, maternal use
of punishment, parental drug use, parental bonding, child
marijuana use, parent alcohol problems.
340 adolescents, ages 12-18 yr (born 1978-1986), Pittsburgh,
PA area.
Case-control. 194 cases (county program) and 146 controls
(Pittsburgh high schools) from different sources. Low
participation from cases. Logistic regression adjusted for race,
parental education and occupation, both parents in home,
neighborhood crime rate, #children in home. Did not consider
potential confounding by parental caregiving quality.
Blood (ug/dL),
Tooth, or Bone (ug/g)
Pb Metrics Analyzed
Tooth
(age 6-8 yr)
Mean: 6.2
Concurrent Bone
Mean (SD) in ppm
Cases:
11 (33)
f^ r\n4-rr\\e-* •
Controls.
*1 C /QO\
1 .0 (oZ)
Outcomes Analyzed
Number convictions for
property or violent
offenses
From police records at
ages 14-21 yr
Delinquent status
From Juvenile Court
records at ages 12-18 yr
Effect Estimate
(95% Cl)b
0.20 (0, 0.40)
ORs (yes/no) for bone
Pb level >25 ppm
White:
3.8(1.1, 13)c
African-American:
2.2(0.5, 10.0)c
"Results are organized by outcome category, behavior ratings then documented criminal offenses. Within each category, studies are grouped by strength of study design,
representativeness of the study population characteristics and Pb biomarker levels examined, and extent of consideration for potential confounding. There is not necessarily a
continuum of decreasing strength across studies.
bUnless otherwise specified, effect estimates are standardized to a 1 ug/dL increase in blood Pb level or 1 ug/g increase in bone or tooth Pb level in the interval from the 10th
percentile to 10 ug/dL blood Pb or 10 ug/g tooth or bone Pb or the 90th percentile, whichever is lower.
°Odds in higher quantile of blood or bone Pb level compared to that in lowest quantile of blood or bone Pb level (reference).
d95% Cl was estimated from a reported p-value of 0.05.
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Collectively, the evidence from prospective studies indicated associations of teacher and
parental ratings of aggressive, antisocial, and delinquent behavior with biomarkers of
cumulative Pb exposure, i.e., age 0-6 year average blood, lifetime (to ages 4-5 or 11-13
years) average blood, tooth, and bone Pb level. Associations with tooth (Bellinger etal..
1994b) and bone (Needleman et al., 1996) Pb level, collected prior to or at the same time
as behavioral assessments, respectively, were found with adjustment for several potential
confounding factors as noted above, with the exception of parental caregiving quality.
Collective evidence from prospective studies indicated associations between blood Pb
level and ratings of behaviors related to conduct problems with population mean blood
Pb levels 6.8-14.3 (ig/dL (prenatal maternal, age 6 year, lifetime average to 4-5 or 11-13
years) (Dietrich et al.. 2001; Wasserman et al.. 2001; Burns etal.. 1999). In a study of
children ages 7-8 years in the U.K. born in the 1990s, Chandramouli et al. (2009) added
to the evidence for associations at relatively high blood Pb levels by finding that
compared with children ages 8 years with age 30 month blood Pb levels 0-2 ug/dL,
children with blood Pb levels >10 ug/dL had increased odds of greater antisocial
activities as rated by parents or teachers with an OR of 2.9 (95% CI: 1.1, 8.0). The
Boston cohort had lower childhood blood Pb levels (seeBellinger et al.. 1992): however,
tooth Pb level was mostly strongly associated with a higher total externalizing behavior
score, which also included attention, overactive/nervous, and aggression (Bellinger et al..
1994b). Weaker associations were found specifically with aggressive or destructive
behavior.
As described above, cross-sectional studies also indicated blood Pb-associated higher
ratings of conduct disorders. Associations were found with lower blood Pb levels than in
prospective studies, concurrent means 1 and 5 (ig/dL (ages 7-17 years). However, for
these children, the influence of higher Pb exposures earlier in childhood cannot be
excluded. Further, because of their various limitations discussed below, cross-sectional
results were less of a consideration in drawing conclusions about the effects of Pb
exposure on conduct disorders. In the recent analysis of 2,867 children ages 8-15 years
participating in NHANES 2001-2004, Braun et al. (2008) analyzed blood Pb level as a
categorical variable and found higher prevalence of conduct disorder as ascertained by
parents (Diagnostic Interview Schedule for Children which follows DSM-IV criteria)
with concurrent blood Pb levels in the range of 0.8 to 1.0 ug/dL. Compared with children
with blood Pb levels 0.2-0.7 ug/dL (6% below detection limit of 0.3 ug/dL), the OR in
children with blood Pb levels 0.8-1.0 ug/dL was 7.2 (95% CI: 1.1, 49). The wide 95%
CIs could have resulted from the small numbers of cases of conduct disorder. For
example, there were 22 children rated as having conduct disorder in the group with blood
Pb levels 0.8-1.0 ug/dL. Nigg et al. (2008) found a blood Pb-associated higher rating
(parent or teacher) of oppositional defiant disorder in a population with similarly low
concurrent blood Pb levels, means ~1 ug/dL. However, the implications are limited
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because of the ADHD case-control design (Section 4.3.3.1) and lack of consideration of
potential confounding factors. Further, because both studies examined adolescents who
likely had higher earlier childhood Pb exposures, there is uncertainty regarding the level,
timing, frequency, and duration of Pb exposure that contributed to the observed
associations.
Consideration for potential confounding varied among the cross-sectional studies of
behaviors related to conduct disorders. With the exception of Nigg et al. (2008). many
considered sex, race, SES, and parental education. However, only a few considered
parental caregiving quality (i.e., HOME score). An analysis of children in Baltimore, MD
adjusted results for child Fe deficiency but examined children ages 2-5 years and
analyzed a total behavior problem score, which included various externalizing and
internalizing behaviors (Sciarillo et al.. 1992). Chiodo et al. (2007) found that higher
concurrent blood Pb level was associated with higher teacher ratings of social problems
and delinquent behavior in children ages 7 years in Detroit, MI with consideration for
HOME score. The study population had high prevalence of prenatal drug exposure. Drug
exposure did not meet the criteria for inclusion in the model, indicating lack of
confounding by this factor. While prenatal drug exposure may not affect internal validity
of the results, it may limit the generalizability of the results to the general U.S. population
of children. Lack of representativeness also may pertain to the results of Chen et al.
(2007). who examined children participating in a chelation trial because of high blood Pb
levels (20-44 ug/dL) at ages 1-3 years.
While few in number, evidence from prospective studies also indicated associations of
biomarkers of earlier childhood Pb exposure with delinquent and criminal acts as
objectively assessed from government records. These studies of delinquent and criminal
acts examined Pb levels in blood or tooth samples collected in the 1980s when Pb
exposures were much higher than those of the current U.S. population (Fergusson et al..
2008; Wright et al.. 2008). However, the prospective study design and consideration for
several potential confounding factors increase confidence that the observed associations
represent a relationship with Pb exposure.
In the Cincinnati cohort, prenatal maternal 1st trimester, age 6 year, and age 0-6 year
average blood Pb levels were associated with self- and parent-reported delinquent and
antisocial acts at ages 15-17 years (Dietrich et al.. 2001). Wright et al. (2008) recently
extended these findings to include associations with criminal and violent criminal arrests
at ages 19-24 years. In models that adjusted for maternal IQ and education, sex, and SES,
the relative risks (RRs) for total arrests per 1 ug/dL increment in blood Pb level were 1.07
(95% CI: 1.01, 1.13) for prenatal blood Pb level, 1.01 (95% CI: 0.97, 1.05) forage 0-6
year average blood Pb level, and 1.05 (95% CI: 1.01, 1.09) for blood Pb level at age
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6 years. Interactions terms for blood Pb by sex were not statistically significant; however,
the attributable risk was considerably higher for males (0.85 arrests/year [95% CI: 0.48,
1.47] for age 6 year blood Pb) than females (0.18 [95% CI: 0.09, 0.33]). A strength of
Wright et al. (2008) was the detailed examination of potential confounding by a large
number of variables (Table 4-12). All of the examined covariates were weakly correlated
with blood Pb levels (r = 0.24-0.35), thereby reducing the potential for confounding by
the examined factors. Nonetheless, variables such as maternal IQ and education, sex, and
SES were included in the model because they were associated with p < 0.05 with arrests
in the full multivariate model or changed the blood Pb level estimate by more than 10%.
HOME score was similar between subjects with and without criminal arrest records and
did not meet the criteria for inclusion in final models.
The study of the New Zealand cohort also considered several potential confounding
factors including several related to SES, family functioning, and parental bonding (Table
4-12) (Fergusson et al.. 2008). Per log increase in Pb in teeth obtained between ages 6
and 8 years, there was a 0.49 (95% CI: 0.16, 0.82) increase in the number of documented
violent or property convictions at ages 14-21 years. Results were adjusted for SES,
ethnicity, parental criminal offenses, and parental alcoholism. The effect estimate for
tooth Pb level decreased in adjusted models and was found to account for <1% in the
variance of criminal convictions; however, the association remained statistically
significant.
The epidemiologic studies described above varied in study design, examination of
specific behaviors related to conduct disorders, and method of assessment. The
consistency of association of Pb biomarker levels with conduct disorders was
corroborated in a recent meta-analysis (Marcus et al., 2010) that included 19 studies
(several of which are described above) with atotal of 8,561 children and adolescents
(range of mean ages: 3.5-18.4 years). Effect estimates were converted to Pearson
correlation coefficients, and the combined effect estimate was r = 0.19 (95% CI: 0.14,
0.23). The key finding of this study was the robustness of association to many between-
study sources of heterogeneity. In the meta-analysis, effect sizes did not significantly
differ by population mean blood Pb level, by prospective or cross-sectional design, by the
various measures of conduct disorders examined (i.e., oppositional behavior,
delinquency, externalizing problem composite measure), or by outcomes assessed using
self-report, parent report, teacher report, or criminal records. Adjustment for covariates
such as SES, birth weight, parental IQ, and home environment did not attenuate the
relationship between Pb biomarker level and conduct disorders. In addition to
strengthening the evidence for the independent associations of Pb biomarker levels with
conduct disorders, the results indicated that the lack of adjustment for any particular
covariate, including HOME score, does not warrant limiting inferences from a particular
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study. The major source of heterogeneity in effect estimates was the biomarker of Pb
examined. A larger magnitude of effect was estimated for hair Pb levels compared with
bone, tooth, or blood Pb levels, which had similar effect sizes. Hair Pb has been
purported to be a better indicator of cumulative Pb exposure compared to bone Pb levels
due to the high turnover of bone throughout childhood and into adolescence (Marcus et
al.. 2010); however, an empirical basis for interpreting hair Pb measurements in terms of
Pb body burden or exposure has not been firmly established (Section 3.3.4.2).
Several studies of conduct disorders aimed to characterize whether associations with Pb
biomarkers were independent of effects on cognitive function. These studies found that
associations of Pb biomarkers with conduct disorders remained statistically significant in
a model that additionally adjusted for child IQ or academic achievement, indicating that
Pb exposure may have a direct effect on conduct disorders independent of its effect on
cognitive function (Chandramouli et al., 2009; Fergusson et al., 2008; Burns etal.. 1999).
However, simple statistical adjustment for cognitive function indices may underestimate
an effect of Pb on conduct disorders because a decrement in cognitive function may lie
on the causal pathway to behavioral problems. Chen et al. (2007) used path analysis to
characterize the direct effects and indirect effects (mediated/occurring through child IQ)
of blood Pb level on an oppositional index at age 5 years and total externalizing behavior
ratings at age 7 years. There was evidence of a direct effect of Pb exposure, particularly
for total externalizing behavior ratings at age 7 years (Table 4-12). These findings may
have limited applicability to the general U.S. population given that some children in the
study had been given chelators at ages 1-3 years because of high blood levels, and it is
possible that the observed associations were due to the residual effect of high earlier
childhood blood Pb levels (20-44 (ig/dL).
lexicological Studies of Aggression
While recent studies were not identified, evidence available in the 2006 Pb AQCD (U.S.
EPA. 2006b) pointed to the effects of Pb on changes in social behavior of rodents and
nonhuman primates. Most observations comprised Pb-induced increases in social and
sexual investigation, as indicated by sniffing, grooming, following, mounting, and
lordosis behavior. In animals, the social behavior most comparable to findings for
conduct disorders in children is aggression; however, the effects of Pb on aggression in
animals were inconsistent. In animals, aggression was assessed as threats, attacks, bites,
chases, and offensive posture in encounters with other animals, and Pb exposure was
found to not affect, decrease, and increase aggression. Pb exposure generally was not
found to affect aggression in juvenile animals; however, increased aggression was found
in adult animals with high concentration (>2,500 ppm) gestational plus postnatal dietary
Pb exposure.
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Delville (1999) found Pb-induced increases in aggression with the lowest concentration
Pb exposure examined among all animal studies of aggression. Golden hamsters exposed
to 100 ppm Pb acetate GD8-PND42 in drinking water had blood Pb levels of 10 to
15 ug/dL at PND42. As adults at PND45, Pb-exposed animals displayed more aggression
as measured by attacking and biting an intruder put in the cage. In mice, higher Pb
exposure produced mixed findings. BK:W mice exposed to 1,300 ppm Pb acetate in
drinking water from gestation through age 18 weeks displayed increased social and
sexual investigation but not aggression (femur Pb level at 34 weeks: 5,364 uM Pb/g ash
in males, 4,026 uM Pb/g ash in females) (Donald et al., 1986). In another study from the
same laboratory, BK:W mice were exposed to 2,500 ppm Pb acetate in drinking water
from gestation through age 17-18 weeks, and found shorter latencies to aggression in
Pb-exposed male mice than in controls (Donald etal. 1987). In juvenile Long-Evans
hooded rats, lactation-only (PND1-PND21) exposure to 670 ppm Pb chloride in drinking
water increased "rough and tumble" play behavior at PND26 which was not characterized
as aggression because of the lack of injury, submissive posturing, or escape attempts in
encounters with other animals (Holloway and Thor. 1987).
4.3.3.3 Integrated Summary of Externalizing Behaviors
Attention, Impulsivity, and Hyperactivity
Attention, impulsivity, and hyperactivity are included within the attention deficit
hyperactivity disorder domain of externalizing behaviors. Although not as extensively as
cognitive function, several epidemiologic studies have examined the relationship between
Pb exposure in children and attention, impulsivity, and hyperactivity in children and
young adults. A majority has examined attention, and some of these studies also
examined impulsivity or hyperactivity. Few studies have examined ratings of a composite
index of ADHD-related behaviors or ADHD diagnosis and are not the focus of this
evaluation. The focus of the evaluation is on evidence for attention but draws on
coherence with evidence for impulsivity and hyperactivity, evidence in animals, and that
describing potential modes of action. The collective epidemiologic evidence base for
attention comprises many prospective studies, which also were reviewed in the 2006 Pb
AQCD (U.S. EPA, 2006b). as well as many cross-sectional studies, some reviewed
previously and some recently published. Whether prospective or cross-sectional or
previously reviewed or recently published, most of the epidemiologic studies found
associations between childhood blood or tooth Pb levels and attention decrements,
impulsivity, and hyperactivity (Figure 4-9 and Table 4-11). Not all results were
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statistically significant, but results mostly showed a pattern of attention decrements,
impulsivity, and hyperactivity with higher blood or tooth Pb level.
Whether prospective or cross-sectional, most studies had population-based recruitment
from prenatal clinics, hospitals at birth, or schools and had moderate to high participation.
A few prospective studies had increased loss-to-follow-up in certain groups, for example,
lower SES, lower earlier FSIQ, lower HOME score. This potential selection bias can
reduce the generalizability of findings to the original study population, but there was not
strong indication that participation was biased to those with higher blood Pb levels and
greater attention decrements, impulsivity, or hyperactivity. Multiple testing was common;
however, in most studies, the consistent pattern of association observed across the ages of
blood Pb level and/or behavior examined increases confidence that the evidence is not
unduly biased by the increased probability of finding associations by chance alone.
With analysis of prenatal, earlier childhood (e.g., age 6 year) or lifetime average (e.g., to
ages 4-5 years or 11-13 years) blood or tooth Pb levels and measures of attention from
later in childhood between ages 4 and 20 years, the prospective studies better
characterized the temporal sequence between exposure and outcome. In prospective
studies, associations of blood or tooth Pb levels with attention were found in diverse
populations ages 8-20 years in North America, Australia, and New Zealand. These
prospective studies examined populations with higher blood Pb levels (means 7-14
(ig/dL) than did cross-sectional studies. Prospective studies found associations of blood
and tooth Pb levels with attention as assessed using neuropsychological tests or ratings by
parents or teachers. Observations of associations across the various methods of
assessment increase confidence that the collective evidence is not unduly influenced by
lower attention ratings biased to parents of children with higher blood Pb levels or Pb
exposures. Prospective studies that examined attention with the CPT found associations
with blood or tooth Pb level. Increases in commission and omission errors or reaction
time were associated with higher prenatal (maternal) and earlier childhood (age
3-60 month average, age 78 month) blood Pb levels in the Cincinnati cohort at ages 15-17
years (Ris et al.. 2004) and with higher tooth Pb (from lst/2nd grade) levels in Boston-
area young adults at ages 19-20 years (Bellinger et al.. 1994a). Results from prospective
studies also indicated lower parent and teacher ratings of attention in association with
higher lifetime average blood Pb levels in children ages 11-13 years in Port Pirie,
Australia (Burns etal. 1999) and with tooth Pb levels (from ages 6-8 years) in children
ages 8-13 years in New Zealand, and Boston, MA (Fergusson et al., 1993; Leviton et al..
1993). The mean blood Pb levels across the populations examined were 6.8 and
8.3 (ig/dL for prenatal (maternal or cord) blood, 13.4 for age 3-60 month average, and 14
(ig/dL for lifetime average blood. A recent prospective study in the U.K. found that
children ages 7-8 years with age 30 month blood Pb levels > 10 (ig/dL had higher teacher
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ratings of hyperactivity but not selective attention compared with children with blood Pb
levels 0-2 (ig/dL (Chandramouli et al.. 2009). A null association with attention was found
in the Yugoslavia cohort with a mean lifetime (to ages 4-5 years) blood Pb level of 7.2
(ig/dL (Wasserman et al.. 2001). However, behaviors may be less reliably measured in
the young children ages 4-5 years.
In children, attention was associated with biomarkers of Pb exposure representing several
different lifestages and time periods. Prospective studies did not examine a detailed Pb
biomarker history, and results do not identify an individual critical lifestage, time period,
or duration of Pb exposure associated with attention decrements in children. Associations
in prospective studies for attention decrements with tooth Pb level, early childhood
average and lifetime average blood Pb levels point to an effect of cumulative Pb
exposure. Prospective studies did not examine indicators of more recent Pb exposure. A
strength of the prospective studies was their more extensive consideration for potential
confounding. Although the specific factors varied by study, prospective studies of
attention adjusted for factors such as SES, parental IQ, maternal education, parental
caregiving quality, self drug use, prenatal drug and alcohol exposure, and birth outcomes.
As described in Section 4.3.3. a few studies have tried to assess potential confounding by
parental psychopathology (Cho et al.. 2010; Nicolescu et al.. 2010; Silvaet al.. 1988) but
with the use of parental self-report, which may be subject to measurement error. At
present, there is not evidence to support parental psychopathology as a direct
confounding factor, i.e., contributing to a spurious association with Pb biomarkers when
there is none. However, parental psychopathology has been speculated to influence child
Pb exposure via a potential relationship with parenting caregiving quality. A clear
relationship between parental psychopathology and poorer parenting behavior has not
been established (as reviewed in Johnston et al.. 2012). and adjustment for parental
psychopathology may control for Pb exposure itself. However, if there is anoncausal
correlation between parental psychopathology and Pb exposure and it is mediated by (i.e.,
occurs through) parental caregiving quality, then direct adjustment for parental
caregiving quality may help address potential confounding by parental psychopathology.
Most cross-sectional studies found attention decrements in children ages 3-16 years in
association with concurrently measured blood or tooth Pb level. Among studies that
examined multiple behaviors, most also found associations with impulsivity and/or
hyperactivity. A limitation of cross-sectional associations with concurrent blood Pb levels
is the greater uncertainty about the potential for reverse causation. The cross-sectional
studies included a mix of previous and recent studies and a wide range of population
mean blood Pb levels (1-11.5(ig/dL). There also was a mix of preschool-aged and
schoolaged children. Among studies of populations with relatively high concurrent blood
Pb levels (means 11.1-13.2 (ig/dL), most found that higher blood Pb levels were
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associated with higher parent or teacher ratings of attention decrements and hyperactivity
(Rov et al.. 2009: Kordas et al. 2007: Silvaetal.. 1988). Higher tooth Pb levels also were
associated with higher teacher ratings of hyperactivity in children in grades 1-3
(Rabinowitz et al.. 1992: Needleman et al.. 1979). Needleman et al. (1979) also found an
association with impulsivity ratings and reaction time. These cross-sectional studies of
higher blood Pb levels found associations with adjustment for SES, child age, sex, family
structure, and parental education, but most did not examine parental caregiving quality.
Some cross-sectional studies examined lower blood Pb levels (means 1-6.5 (ig/dL for
ages 4-17 years) closer to those measured in most current U.S. children. There was
heterogeneity in the extent of potential confounding and ages of children examined,
which makes drawing conclusions difficult. Among the lower blood Pb level studies that
had more extensive consideration for potential confounding, most found higher
concurrent blood Pb levels to be associated with attention decrements and also with
impulsivity or hyperactivity. All of these studies adjusted for or considered potential
confounding by SES as well as parental IQ, education, and caregiving quality. In the
Rochester cohort, concurrent blood Pb level (mean 6.5 (ig/dL) was not associated
strongly with attention as rated by study examiners (Canfield et al.. 2003b). However, in
the young children ages 4 years, behaviors may be less reliably measured. Also, in the
Rochester cohort, the association with attention was attenuated only after adjustment for
color/shape knowledge and child IQ. These inter-related outcomes may share some
variance. In another study of younger children ages 5-6 years in Canada, higher
concurrent blood Pb level was associated with higher examiner ratings of attention
decrements, impulsivity, and hyperactivity (Plusquellec et al.. 2010). Among older
children (age 7 years) in Detroit, MI, higher concurrent blood Pb level (means: 5.0, 5.4
(ig/dL) was associated with higher examiner or teacher ratings of attention decrements
and hyperactivity (Chiodo et al.. 2007: 2004). These studies also found associations with
attention decrements as assessed by the CPT, although not strongly with all examined
indices of the CPT. In these populations with high prevalence of prenatal alcohol or drug
exposure, results did not indicate confounding by these other exposures. However, the
presence of these other exposures may limit the generalizability of results. Further, in
older children, the contribution of higher Pb exposures earlier in childhood cannot be
excluded. While higher-quality cross-sectional evidence in populations with mean blood
Pb levels 5-6.5 (ig/dL show associations of blood Pb with attention decrements,
impulsivity, and hyperactivity, there are limitations related to young age of subjects or
generalizability of the study population.
Other cross-sectional studies examined even lower concurrent blood Pb levels, 1-3.7
(ig/dL and found associations with attention, impulsivity, and hyperactivity in children
ages 8-17 years (Cho et al.. 2010: Nicolescu et al.. 2010: Nigg et al.. 2008). These
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associations were found with parental or teacher ratings, with attention measured by the
CPT, and impulsivity measured with response inhibition tests. Associations between
blood Pb level and attention or impulsivity also were found with adjustment for child IQ
(Cho etal.. 2010; Nigg et al.. 2008). supporting an effect of Pb exposure on these
behaviors independent of effects on cognitive function. Despite finding associations with
more objective measures of attention or impulsivity, a limitation of these studies was
their less consistent consideration for potential confounding factors. Studies varied in
adjusting results for SES and parental education, and none considered parental caregiving
quality. A limitation specific to Nigg et al. (2008) is the case-control study design of
children with and without ADHD could result in biased ratings of hyperactivity by blood
Pb level if participation by children with ADHD is biased to those with higher Pb
exposure.
Recent cross-sectional studies also found associations between higher concurrent blood
Pb levels and higher parental ratings of a composite index of ADHD-related behaviors,
including a large NHANES analysis that used DSM-IV criteria (Froehlich et al.. 2009). In
the few available studies, concurrent blood Pb levels were associated with prevalence of
diagnosed ADHD in children (Section 4.3.3.1). There is coherence with evidence from
prospective studies for associations of blood and tooth Pb levels with attention,
hyperactivity, and impulsivity, which are behaviors evaluated in the diagnosis of ADHD.
However, the small number of studies, their cross-sectional or case-control design,
inconsistent adjustment for SES and parental education, and lack of consideration for
potential confounding by parental caregiving quality preclude conclusions regarding the
relationship between Pb exposure and ADHD specifically.
As described above, consideration of potential confounding factors varied particularly
among cross-sectional studies. Regardless of study design, adjustment for SES is difficult
as it is highly correlated with Pb exposure and there is no single measure that fully
represents SES. Residual confounding also is likely by factors not considered or factors
measured with error. However, in addition to evidence from prospective studies that
considered several well-characterized potential confounding factors, the coherence with
evidence that Pb exposure induces greater distractibility and impulsivity in animals
increases confidence that the associations of blood and tooth Pb levels with attention
decrements and impulsivity observed in children represent a relationship with Pb
exposure. Evidence in rodents and monkeys demonstrates Pb-induced impaired response
inhibition by decreased interresponse times and increased overall response rates in FI
reinforcement schedules (Table 4-7) (Rossi-George et al.. 2011). slower acquisition of a
DRL schedule (Rice and Gilbert. 1985). and response perseveration on tests of
discrimination reversal learning and FR/waiting for reward (Brockel and Cory-Slechta.
1999b. 1998; Gilbert and Rice. 1987). Coherence is found particularly with evidence in
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children for decrements in response inhibition as assessed using the stop signal task.
Impulsivity in animals was found with relevant gestational, post-weaning, and lifetime
(from gestation or birth) Pb exposures, i.e., producing blood Pb levels 11-29 (ig/dL. Pb
exposure has not consistently induced attention decrements or distractibility in rats and
monkeys (Brockel and Cory-Slechta. 1999a; Gilbert and Rice. 1987). The findings in
children and animals for Pb-associated impulsivity are supported by evidence that Pb
affects dopaminergic neurons of the frontal cortex and striatum of the brain by altering
dopamine release and receptor density. The circuitry in these regions is thought to
mediate response inhibition. Attention decrements, impulsivity, and hyperactivity also
have been linked with changes in the hippocampus, and evidence describing the effects of
Pb on hippocampal functions further supports the mode of action for the effects of Pb on
these behaviors.
Conduct Disorders
As described at the beginning of Section 4.3.3. there are two domains of conduct
disorder: undersocialized aggressive conduct disorder and socialized aggressive conduct
disorder. Although not examined as extensively as cognitive function or other
externalizing behaviors, previous and recent prospective studies consistently demonstrate
Pb-associated increases in delinquent behavior, aggression, antisocial activities, and
destructive behavior in children as rated by parents and teachers (Table 4-12). The
available evidence does not point to an effect of Pb exposure in one particular domain of
conduct disorders. Supporting evidence also is provided by studies showing associations
with criminal offenses in adolescents and young adults as assessed with government
records. Most studies examined multiple measures of conduct disorders; however, the
consistent pattern of association observed across the ages of blood Pb level and/or
behaviors examined increases confidence that the evidence is not unduly biased by an
increased probability of finding associations by chance alone. Recent cross-sectional
studies found associations between concurrent blood Pb level and ratings of conduct
disorders, but several had limitations aside from study design (Nigg et al.. 2008; Chen et
al.. 2007; Chiodo et al.. 2007). including limited consideration for potential confounding.
A particularly informative cross-sectional study was that finding a 7.2 (95% CI: 1.1, 49)
higher odds of parentally-assessed conduct disorder in 2,867 children ages 8-15 years
participating in NHANES with concurrent blood Pb levels 0.8-1.0 (ig/dL compared with
blood Pb levels 0.2-0.7 (ig/dL (Braun et al.. 2008). There were 22 children rated as
having conduct disorder in the group with blood Pb levels 0.8-1.0 (ig/dL, which may
have contributed to the wide 95% CI. The association also could have been influenced by
higher past Pb exposures of the children. Further, while adjustment was made for age,
sex, race, poverty to income ratio, and smoking exposure, potential confounding by
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parental caregiving quality was not examined. Evidence of Pb-induced aggression in
animals was mixed in adult animals with lifetime Pb exposure beginning in gestation and
not indicated in juvenile animals.
The evidence from prospective studies for ratings of behaviors related to conduct
disorders in children is substantiated by analyses of school-aged children and young
adults (ages 7-17 years) in populations from various locations and SES (i.e., U.K.,
Cincinnati, Australia) with high participation rates, lack of indication of substantial
selection bias, and consideration of several potential confounding factors including
multiple SES-related factors, parental caregiving quality, smoking exposure, and birth
outcomes (Chandramouli et al.. 2009: Dietrich etal.. 2001: Burns etal.. 1999). Pb
biomarker levels were associated with both parent and teacher ratings of behaviors
related to conduct disorders, reducing the likelihood of higher behavioral ratings biased
to parents of children with higher Pb biomarker levels. Recent prospective studies of
criminal offenses in adolescents and young adults strengthened previous evidence with
consideration for potential confounding by factors such as SES, smoking, drug, and
alcohol exposure, and parental caregiving quality (Fergusson et al.. 2008: Wright et al..
2008). In the Cincinnati cohort, a 1 ug/dL increase in age 6 year blood Pb level was
associated with an increased risk of criminal arrests with an RR of 1.05 (95% CI: 1.01,
1.09) with adjustment for maternal IQ and education, sex, and SES (Wright et al.. 2008).
In the New Zealand cohort, a log unit increase in tooth Pb (from ages 6-8 years) was
associated with a 0.49 (95% CI: 0.16, 0.82) increase in the number of documented
property or violent convictions at ages 14-21 years (Fergusson et al.. 2008) with
adjustment for SES, ethnicity, parental criminal offenses, and parental alcoholism.
Further support for Pb-associated increases in conduct disorders was provided by a recent
meta-analysis that found that the evidence was robust to heterogeneity in study design,
definition and assessment method of conduct disorders, potential confounding factors
examined, and population mean blood Pb levels (Marcus et al., 2010). As described in
Section 4.3.3. there is not evidence to support parental psychopathology as a direct
confounding factor. There hypothetically could be potential confounding by parental
psychopathology via a relationship with parental caregiving quality. However, Pb-
associated increases in behaviors related to conduct disorders have been found with
adjustment for parental caregiving quality.
Associations of measures of conduct disorders at ages 7-24 years (parent/teacher
behavior ratings and criminal offenses) with earlier childhood blood (e.g., age 30 month,
age 6 year), early childhood average blood (e.g., age 0-6 year), lifetime average blood (to
age 11-13 years), tooth, and bone Pb levels pointed to the effects of early childhood or
cumulative Pb exposures. Associations were found with a mean lifetime average (to age
11-13 years) blood Pb level of 14 (ig/dL, mean age 6 year blood Pb level of 6.8 (ig/dL,
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mean prenatal maternal blood Pb level of 7.8 (ig/dL, and age 30 month blood Pb levels
>10 (ig/dL. Recent cross-sectional studies found associations with concurrent blood Pb
level and lower blood Pb levels, means 1 and 5 (ig/dL (ages 7-17 years), but the study
limitations detailed above limit strong inferences regarding the effects of Pb exposure on
conduct disorders in these populations. Most prospective studies did not analyze Pb
biomarker levels at multiple lifestages and time periods, including later childhood and
more recent adulthood, or examine differences in association between Pb biomarker
levels measured at various lifestages and time periods. The evidence does not identify an
individual critical lifestage, time period, or duration of Pb exposure associated with
conduct disorders in children or exclude an effect of more recent Pb exposures.
4.3.4 Internalizing Behaviors in Children
4.3.4.1 Epidemiologic Studies of Internalizing Behaviors in
Children
Most studies of the effects of Pb on behavior in children have focused on externalizing
behaviors such as attention, impulsivity, and conduct disorders. However, several studies
also have linked biomarkers of Pb exposure in children with internalizing behaviors
characterized by directing feelings and emotions inward, i.e., withdrawn behavior,
symptoms of depression, fearfulness, and anxiety. As with externalizing behaviors,
internalizing behaviors can be classified into various domains: withdrawn behavior,
somatic symptoms, anxiety/depression. Studies examining relationships with Pb exposure
examined multiple behaviors and did not clearly indicate that Pb exposure affected a
particular domain of internalizing behaviors. Therefore, the evidence for internalizing
behaviors is examined as a single group. Some studies found stronger associations for
externalizing behaviors than for internalizing behaviors (Plusquellec et al., 2010;
Wassermanetal.. 2001; Bellinger et al.. 1994a; Sciarillo et al.. 1992). Others did not find
a clear difference in the strength of association (Roy et al., 2009; Chiodo et al., 2004;
Bellinger et al.. 1994b). which is not unexpected since externalizing and internalizing
behaviors are positively correlated. Internalizing behaviors frequently were assessed
using the Child Behavior Checklist, and as with externalizing behaviors, were rated by
parents and/or teachers. Associations with both parent and teacher ratings increase
confidence that biased reporting of internalizing behaviors by parents of children with
higher blood Pb levels did not unduly influence the collective body of evidence. Most
studies had moderate to high follow-up participation. With the exception of the
Yugoslavia cohort, participation was not biased to those with higher ratings of
internalizing behaviors and higher blood Pb levels. Additionally, in most studies, a
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consistent pattern of association was observed across the ages of blood Pb level and/or
multiple behaviors examined, which increases confidence that the evidence is not
strongly biased by an increased probability of finding associations by chance alone.
The potential for parental psychopathology to influence associations between child Pb
and externalizing behaviors was described in Section 4.3.3. Parental psychopathology has
the potential to influence associations between child Pb and internalizing behaviors in the
same ways since internalizing behaviors also have a familial component. As with
externalizing behaviors, available evidence does not support consideration of parental
psychopathology as a direct potential confounder of child Pb-internalizing behavior
associations. Instead, the ISA evaluates potential confounding by parental caregiving
quality (speculated to mediate a potential correlation between parental psychopathology
and Pb exposure) and other factors well documented in the literature to be correlated with
both Pb exposure and internalizing behaviors such as SES and parental education.
Key evidence was provided by prospective studies in various populations, i.e., Boston,
Port Pirie, Australia, and Yugoslavia. Collectively, these studies found higher ratings for
internalizing behaviors in children (n = 322-1,273, ages 3-13 years) in association with
cord blood, concurrent blood (age 3 year), lifetime average blood, and tooth Pb levels
(Wasserman et al., 2001; Burns et al., 1999; Wasserman et al., 1998; Bellinger et al..
1994b). In the Port Pirie cohort, Burns et al. (1999) found that higher lifetime average
blood Pb levels (mean: ~14 (ig/dL) were associated with parental ratings of externalizing
behaviors more strongly in boys and with internalizing behaviors (i.e., withdrawn,
anxious/depressed, composite) more strongly in girls ages 11-13 years, which may
indicate sex differences in the effect of Pb or differences in the types of behaviors that are
observed and reported in girls versus boys. Based on a log-linear model, a 1 (ig/dL
increase in lifetime average blood Pb level was associated with increased odds of an
anxious/depressed rating above the median of 1.04 (95% CI: 1.0, 1.09) among 159 boys
ages 11-13 years (in the 10th-90th percentile interval of blood Pb level 13.7-14.9 (ig/dL)
and 1.07 (95% CI: 1.01, 1.14) among 163 girls ages 11-13 years (in the 10th-90th
percentile interval of blood Pb level 13.3-14.9 (ig/dL). These associations were found
with the adjustment for SES factors, HOME score, family functioning score, and current
maternal psychopathology (General Health Questionnaire).
Differences between externalizing and internalizing behaviors also were found in the
Yugoslavia cohort but by age of assessment and blood Pb levels. This cohort was
examined between ages 3 and 5 years, ages at which behaviors may be less reliably
assessed. Among 379 children ages 3 years from the higher and lower Pb exposure
towns, higher cord and concurrent blood Pb levels (respective means: 14, 25.8 (ig/dL)
were associated with higher maternal ratings of anxious-depressed, withdrawn, and
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externalizing behaviors, with stronger associations found for concurrent blood Pb level
rWasserman et al.. 1998). Among 191 children ages 4-5 years from the lower Pb
exposure town, higher lifetime average blood Pb level (mean: 7.2 (ig/dL) was associated
with higher ratings of delinquent behavior and internalizing behaviors, with stronger
associations found for delinquent behavior (Wasserman et al.. 2001). A log increase in
higher lifetime average blood Pb level was associated with a 0.22 log higher (95% CI:
-0.03, 0.47) rating of withdrawn behavior and 0.19 log higher (95% CI: -0.05, 0.43)
rating of anxious-depressed behavior. Results at each age were adjusted for HOME score,
sex, ethnicity, and maternal education. Additional covariates included residence type at
age 3 years and age, birthweight, and maternal smoking history at ages 4-5 years.
With regard to important lifestages or durations of Pb exposure, results from prospective
studies did not clearly demonstrate differences in association among Pb biomarkers
measured at various lifestages or time periods. The importance of cumulative Pb
exposure was indicated by associations found with lifetime average blood Pb levels in the
Port Pirie cohort (to age 11-13 years) and in the Yugoslavia cohort (to age 5 years) and
with tooth Pb (from age 6 years) levels in Boston children at age 8 years. In Boston
children, Pb levels measured in teeth (mean: 3.4 (ig/g) but not cord blood were associated
with a teacher rating of internalizing behavior score > 60 at age 8 years (Bellinger et al..
1994b). In another Boston-area cohort, tooth (collected at first or second grade) Pb levels
were not associated with self-rated symptoms of depression (Profile of Mood States
questionnaire) at ages 19-20 years (Bellinger etal., 1994a). Prospective studies did not
analyze a detailed history of Pb biomarker levels to evaluate persistence of effects of
early exposure or to identify an individual critical lifestage or time period of Pb exposure
associated with increases in internalizing behaviors. The available evidence does not
preclude an effect of later childhood or more recent Pb exposure.
In the Cincinnati cohort, Dietrich et al. (1987) with structural equations showed that
associations of prenatal maternal and age 10 day blood Pb level (respective means: 8.3,
4.9 (ig/dL) with poorer mood in infants ages 6 months (n = 185) were indirect, meaning
they were mediated through associations of Pb with shorter gestation or lower birth
weight. These results suggest that Pb may exert its effects by impairing nervous system
development. The fetal period is an active period of neuronal differentiation, dendritic
branching, and synaptogenesis, which if impaired by Pb exposure, could have broad
implications on subsequent neurodevelopment. There are few such analyses, and the
implications are limited by the lower reliability of mood assessed in infancy.
Cross-sectional studies found associations between concurrent blood (Liuetal.. 20 lib;
Roy et al.. 2009) or hair (Bao et al.. 2009) Pb levels and teacher and parent ratings of
internalizing behaviors in children in China and India (n = 303-756, ages 3-16 years).
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Mean concurrent blood Pb levels in children ages 3-7 years were 11 and 14 (ig/dL.
Results were adjusted for family income and parental education but not caregiving
quality. In the few studies of populations with mean blood Pb levels ~5 (ig/dL, results
were inconsistent. Chiodo et al. (2004) found an association with internalizing behaviors
with adjustment for SES, maternal hard drug use, and marital status in 175 children age
7.5 years in Detroit, MI who had high prevalence of prenatal alcohol exposure. HOME
score, maternal education, and prenatal alcohol and drug exposure were not found to
influence associations with blood Pb level. While prenatal alcohol and drug exposure
may not affect the internal validity of results, they may limit generalizability of the
results to the general population of U.S. children. A study that examined 79-91 Inuit
children (age 5 years) in Quebec, Canada with Hg and PCB co-exposures from a high
fish diet, did not find statistically significant associations between concurrent blood Pb
level and internalizing behaviors with consideration of potential confounding by parental
caregiving quality, caregiver education and IQ, blood Hg or PCB levels, and prenatal
smoking and alcohol exposure (Plusquellec et al.. 2010).
Associations of Pb biomarkers with internalizing behaviors in children were observed
with consideration for a wide range of potential confounding factors, most commonly,
age, birth outcomes, parental education, and other SES-related factors. Parental
caregiving quality was evaluated in few studies. Blood Pb-associated higher ratings of
internalizing behaviors were found with adjustment for HOME score in the Yugoslavia
cohort (Wasserman et al., 2001; 1998). and HOME, family functioning, and current
maternal psychopathology (General Health Questionnaire) in the Port Pirie cohort (Burns
et al., 1999). Several studies, each of which adjusted for a different set of covariates,
found slightly attenuated or larger effect estimates in multivariate models than unadjusted
models (Wasserman et al., 2001; Burns etal., 1999; Bellinger et al.. 1994b). Collectively,
these observations increase confidence that the associations observed with Pb biomarkers
reflect a relationship with Pb exposure.
4.3.4.2 lexicological Studies of Internalizing Behaviors
As in epidemiologic studies, toxicological studies have focused more on cognitive
function and externalizing behaviors and less on internalizing behaviors. The associations
described in the preceding section for Pb biomarker levels with ratings of withdrawn
behavior and symptoms of depression and anxiety in children are supported by findings
of Pb-induced anxiety, emotionality and depression-like behaviors in animals.
Emotionality has been indicated by increased disruption in performance and frustration in
response to errors and reward omission in visual or olfactory discrimination task trials in
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Pb-exposed rats (Beaudin et al.. 2007; Stangle et al.. 2007). In each study, Long-Evans
rats were exposed to Pb during and after lactation (300 ppm Pb acetate via dam drinking
water PND1-PND17 then either 20 or 300 ppm PND17-PND30 in own water, with
respective blood Pb levels of 13 and 31 (ig/dL on PND52). However, this exposure
regimen in mice was reported to produce higher peak blood Pb levels at the end of
exposure, i.e., 40-60 or 100-140 (ig/dL. In Beaudin et al. (2007). greater disruption in
performance (i.e., failure to enter testing alcove) after committing errors and having
rewards omitted was found in rats with blood Pb levels 13 and 31 (ig/dL tested at age
9-15 weeks. Pb-exposed rats also had greater response latency with reward omission as
indicated by the entrance into the testing alcove but failure to respond within a set period
of time. In Stangle et al. (2007). increased reactivity to errors and reward omission was
found in rats with blood Pb levels 31 (ig/dL. In rhesus monkeys, emotional dysregulation
was indicated by tactile defensiveness after exposure to Pb acetate/50% glucose in 4 cc
daily milk formula from PND8 to ages 1-2 years to produce target blood Pb levels
35-40 (ig/dL (Moore et al.. 2008).
In other studies, rats showed emotionality and depression-like behavior in the open field
test or forced swim test (i.e., Porsolt's Test) with Pb exposure. The open field test
monitors activity levels and movements of animals in three dimensions. Depression-like
behavior is indicated by freezing behavior and low levels of activity. Emotionality is
indicated by grooming or freezing. In the forced swim test, animals are placed in a
vertical cylinder of water from which there is no escape and monitored for duration of
struggling or attempt to escape. Animals that stop quickly are ascribed a depression-like
phenotype. As for many other neurobehavioral responses, sex-specific differences were
found. Pb-exposed male Wistar rats showed increased emotionality in the open field test
as indicated by fewer counts of rearing and longer time freezing. Pb-exposed females
showed a depression-like phenotype in the forced swim test as indicated by longer time
of immobility (de Souza Lisboa et al.. 2005). While measured blood Pb levels of rats
were low, 5-7 (ig/dL, they were measured after a lag in exposure (PND70) and produced
by oral gavage (10 mg/day to dams), a route that may be less relevant to human routes of
Pb exposure. Pb-induced immobility in the forced swim test also was found with 6-week
postnatal (PND30-PND72) Pb acetate exposure via drinking water but producing a mean
blood Pb level of 40 (ig/dL 6-7 weeks after exposure ceased. Reducing internal Pb dose
with the chelator succimer reversed Pb-induced immobility (Stewart et al.. 1996).
Depression initially may seem like an unexpected effect of immune modulation, but it has
been linked to an interaction between the CNS and the immune system via alterations in
cytokines such as IL-6 (Section 4.6.6.1). Dyatlov and Lawrence (2002) found that dietary
Pb exposure through lactation and a brief period after weaning (500 (JVI, PND1-PND22,
resultant blood Pb level: 17 (ig/dL) potentiated sickness behavior in mice in response to
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bacterial infection. Sickness behavior was evidenced by an increase in serum IL-6 levels
with an accompanying decrease in food and water intake and decrease in body weight
gain. This phenotype was correlated with decreases in the populations of several T cell
subtypes. Pb exposure also potentiated release of IL-1(3, which plays an important role in
inflammatory responses to infection and has been shown to inhibit hippocampal
glutamate release in young but not aged animals. Sickness behavior also was induced in
Pb-exposed animals with IL-6 and IL-1 administration without infection, further
supporting a role for immunomodulation in mediating sickness behavior.
Pb exposure had mixed effects on anxiety-related responses as measured by the elevated
plus maze, which assesses behavior of rodents in an unfamiliar environment. The maze is
elevated above the floor and consists of two arms that are enclosed with walls intersected
with two arms that have no walls. The animal is placed in the center of the maze, and
longer latency to enter an open arm, and lower frequency and duration of entries into an
open arm are indicative of anxiety. Postnatal (PND1-PND30) exposure of female rats to
2,000 ppm Pb acetate in drinking water, yielding blood Pb levels 34 (ig/dL, induced an
increase in anxiety-related behavior at PND60, as indicated by a lower percentage of
open arm entries and less time spent in the open arms (Foxetal. 2010). In a study of
gestational/lactational (GD1-PND24) exposure to 2.84 mg/mL Pb acetate trihydrate in
drinking water, Pb-exposed Sprague-Dawley rats did not differ in anxiety-related
responses from controls. Blood Pb levels of these Pb-exposed rats were higher than those
relevant to this ISA, 69.8 (ig/dL at PND25 (Molina etal.. 2011).
4.3.4.3 Summary of Internalizing Behaviors
Internalizing behaviors, i.e., withdrawn behavior, symptoms of depression and anxiety,
have been examined less than externalizing behaviors (e.g., attention, conduct disorders).
However, several prospective studies found associations of higher parental and teacher
ratings of internalizing behaviors with higher tooth or lifetime average blood Pb levels in
children ages 8-13 years (Burns etal.. 1999; Bellinger etal.. 1994b) and higher cord,
concurrent, and lifetime average blood Pb levels in children ages 3-5 years (Wasserman
et al., 2001; 1998). Collectively, the lack of biased participation by subjects with higher
blood Pb levels and associations found with both parent and teacher ratings increase
confidence that the evidence is not unduly influenced by biased reporting of behaviors by
parents of children with higher blood Pb levels. The prospective studies found
associations with adjustment for several potential confounding factors, including age,
birth outcomes, SES, parental education, and parental caregiving quality. As described in
Section 4.3.3. there is not evidence to support consideration of parental psychopathology
as a direct confounding factor. There hypothetically could be potential confounding by
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parental psychopathology via a relationship with parental caregiving quality. However,
Pb-associated increases in internalizing behaviors have been found with adjustment for
parental caregiving quality. In prospective studies, associations were found with tooth Pb
levels and lifetime average, cord, and concurrent blood Pb levels. While, there is not a
clear indication of an individual critical lifestage or time period of Pb exposure associated
with internalizing behaviors, several observations point to an effect of cumulative Pb
exposure. These results do not preclude an effect of later childhood or more recent Pb
exposures. In prospective studies, associations with internalizing behaviors were found
with mean lifetime average blood Pb levels of 14 (ig/dL (to age 11-13 years) (Burns et
al.. 1999) and 7.2 (ig/dL (to age 4.5-5 years) (Wasserman et al.. 2001). In the Yugoslavia
cohort at ages 4-5 years, in whom behaviors may be less reliably rated, lifetime average
blood Pb level was associated more strongly with the rating of delinquent behavior than
ratings of internalizing behaviors (Wasserman et al.. 2001). Cross-sectional studies found
concurrent blood Pb-associated increases in internalizing behaviors in children ages 3-16
years. Studies that considered potential confounding by a similar set of factors as the
prospective studies examined children ages 5-7 years with mean concurrent blood Pb
levels ~5 (ig/dL and found inconsistent associations (Plusquellec et al.. 2010; Chiodo et
al.. 2004).
Evidence in children is supported by observations that dietary Pb exposure resulted in
depression-like behavior and emotionality in rodents (gestational-lactational, PND1
to PND 22 or 30, post-lactation) and rhesus monkeys (neonatal exposure to age 1-2
years) (Moore et al.. 2008). with some evidence in rodents at blood Pb levels relevant to
this ISA (13, 17 (ig/dL) (Beaudin et al., 2007; Dyatlov and Lawrence. 2002). Evidence
for Pb-induced anxiety in animals is mixed. Mode of action support is provided by well-
documented evidence for Pb-induced changes in the HPA axis (Section 4.3.2.3) and
dopaminergic and GABAergic systems (Sections 4.2.2.2. 4.3.10.4. and 4.3.10.8). which
are involved in regulating mood and emotional state.
4.3.5 Psychopathological Effects in Adults
4.3.5.1 Epidemiologic Studies of Psychopathological Effects in
Adults
The potential effects of Pb exposure on psychopathological effects (e.g., anxiety,
depression, schizophrenia) have been examined less so in adults than in children and less
than cognitive function in adults. Most studies of Pb and psychopathologies in adults
have examined mood states, which are an integral part of the neurocognitive test battery
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of the World Health Organization (WHO). There is evidence that indices of the Profile of
Mood States may be particularly sensitive to toxicant exposures (Johnson et al.. 1987).
As with other nervous system endpoints in adults, evidence reviewed in the
2006 Pb AQCD (U.S. EPA. 2006b) pointed to effects in Pb-exposed workers. Higher
prevalence of self-reported symptoms of depression, anxiety, and tension was found
among Pb-exposed workers (n = 43-576, mean concurrent or peak blood Pb levels: 31-
79 (ig/dL) in association with higher blood Pb levels or compared with unexposed
controls (n = 24-181, mean blood Pb levels: 15-38 ng/dL) (Schwartz et al.. 2005:
Maizlish et al.. 1995; Parkinson et al.. 1986; Baker etal., 1985; Baker etal. 1984; Lilis et
al., 1977). Several studies considered potential confounding by age, sex, and education.
Medical conditions, smoking, and alcohol use were examined in one or two studies. Only
Maizlish et al. (1995) examined other occupational exposures (previous occupational
exposure to solvents). Occupational Mn exposure has been linked with anxiety and
depression but was not examined as a potential confounder of Pb associations. Most
studies were cross-sectional, which makes less certain the temporal sequence between Pb
exposure and development of psychopathological effects.
The few studies of adults without occupational Pb exposures participating in NAS and
NHANES demonstrated associations of concurrent blood and bone Pb level with
symptoms of depression and anxiety. As bone Pb is a major contributor to blood Pb
levels in adults without current occupational Pb exposure, cross-sectional associations for
both blood and bone Pb may indicate effects of cumulative Pb exposure. Although
specific covariates varied among studies, these previous and recent cross-sectional
studies found associations with adjustment for several potential confounding factors,
including age, education, smoking, alcohol use, employment status, and SES. Further, the
examination of multiple exposures and outcomes in these studies reduces the likelihood
of participation conditional specifically on Pb exposure and psychopathological effects.
Analyses of 526 men ages 48-70 years in the NAS indicated associations of both higher
concurrent blood (mean: 6.3 [SD: 4.2] (ig/dL), patella (mean: 32.1 [SD: 19.8] (ig/g), and
tibia (mean: 21.9 [SD: 13.5] (ig/g) Pb levels with higher prevalence of a combined index
of self-reported depression and anxiety (Rhodes et al., 2003). In a recent analysis of 744
NAS men ages 48-94 years, Rajan et al. (2007) found associations of symptoms assessed
using the Brief Symptom Inventory with patella and tibia Pb levels (measured within 1.3
years of outcomes). A 14 (ig/g increase in tibia Pb level was associated with an increased
odds of an anxiety score one standard deviation above the mean of 1.18 (95% CI: 0.98,
1.42) and of depression score above the median of 1.16 (95% CI: 0.97, 1.38). Similar
effect estimates were found for patella Pb level. Effect modification by ALAD genotype
was not in a consistent direction. For most of the evaluated mood symptoms, ORs for
tibia bone Pb levels were larger among the 587 men with the ALAD 1-1 genotype. In
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contrast, ORs for associations between patella Pb levels and symptoms such as
depression and positive symptom distress index were larger among the 119 ALAD-2
carriers. In the NAS, effect modification by ALAD genotype also was inconsistent for
associations between tibia Pb levels and cognitive performance (Rajan et al.. 2008)
(Section 4.3.2.7). The relationship between ALAD-2 genotype and Pb bioavailability is
not clear (Section 4.2.3.3).
A recent analysis of 1,987 adults ages 20-39 years participating in NHANES 1999-2004
was the largest study of psychopathological effects in adults and included both men and
women of multiple races and ethnicities (Bouchard et al.. 2009). However, only
concurrent blood Pb levels were available for analysis. Major depressive disorder, panic
disorder, and generalized anxiety disorder were examined using the WHO Composite
International Diagnostic Interview, which follows DSM criteria. Adults with concurrent
blood Pb levels >0.7 (ig/dL (detection limit = 0.2, 0.3 (ig/dL) had higher prevalence of all
three self-reported disorders. Adults in the highest quintile of concurrent blood Pb level
(> 2.11 (ig/dL) had the highest OR for major depressive disorder (OR: 2.3 [95% CI: 1.1,
4.8]) and panic disorder (OR: 4.9 [95% CI: 1.3, 18]) compared with nonsmoking adults
with blood Pb levels 0.3-0.7 (ig/dL with adjustment for age, sex, race, education, and
poverty to income ratio. A monotonic increase in ORs was not found across the quintiles
of blood Pb levels. For all endpoints, ORs were larger in analyses excluding current
smokers. While associations were found with relatively low concurrent blood Pb levels,
there is uncertainty regarding the magnitude, timing, frequency, and duration of Pb
exposure that contributed to the observed associations.
In analyses of small cohorts in California and New England born 1959-1966, Opler et al.
(2008; 2004) reported associations between higher levels of cord plasma 5-ALA and
subsequent diagnosis of schizophrenia spectrum disorder (ascertained from records and
structured interview) in adolescents and adults. Because of the lack of direct
measurements of Pb biomarker levels, post hoc analysis, and limited consideration for
potential confounding, conclusions regarding a relationship between schizophrenia and
Pb and are not warranted. Investigators measured 5-ALA levels in stored serum samples
as surrogates for Pb exposure only citing previous observations of a high agreement
(90%) between samples with 5-ALA levels > 9.05 ng/mL and blood Pb levels >
15 (ig/dL. In the California cohort, 5-ALA level > 9.05 ng/mL was associated with
schizophrenia spectrum disorder with an OR of 2.43 (95% CI: 0.99, 5.96), adjusted for
maternal age at delivery. In pooled analyses of the California and New England cohorts
(n = 200), 5-ALA level > 9.05 ng/mL was associated with schizophrenia spectrum
disorder with an OR of 1.92 (95% CI: 1.05, 3.52), with adjustment for maternal age and
education. An adjusted OR was not presented for the New England cohort alone, and it
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appeared that the association in the pooled dataset was influenced more by that found in
the California cohort.
4.3.5.2 Toxicological Studies of Mechanisms for Schizophrenia
An environmental origin of schizophrenia was proposed years ago (Tsuang. 2000). and
while epidemiologic evidence is inconclusive, toxicological studies have found Pb
exposure to induce changes that have been associated with schizophrenia-related effects
(Figure 4-10). Pb exposure has been shown to reduce function in the NMDA receptor
(NMDAR) and decrease hippocampal neurogenesis. Pb may bind a divalent cation site in
the NMDAR and allosterically inhibit glycine binding (Hashemzadeh-Gargari and
Guilarte. 1999). Antagonists of the NMDAR glycine site have been shown to exacerbate
schizophrenia symptoms in affected individuals and induce a schizophrenic phenotype in
unaffected subjects (Coyle and Tsai. 2004). Developmental Pb exposure inhibits
neurogenesis in animal models (Section 4.3.10.9). Evidence indirectly suggests that a
decrease in hippocampal degenerate gyrus (DG) neurogenesis could be a mode of action
for Pb-associated schizophrenia induction. Decreased neurogenesis is seen in patients
with schizophrenia (Kempermann et al.. 2008; Reif et al.. 2006) and animal models of
schizophrenia (Maedaetal.. 2007). and clozapine, a treatment for schizophrenia, restores
hippocampal DG neurogenesis in animal models of schizophrenia (Maeda et al.. 2007)
(Figure 4-10). These DG pathways are also NMDAR-dependent.
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Human Populations or Cohorts
ALAD and Schizophrenia
Associated in Human Cohort
(Epidemiologic)
NMDAR Pharmacologically-
Pb exposure > Antagonism < induced Schizophrenia in
Animal models
Figure 4-10 Hypothetical representation of the contribution of Pb exposure to
the development of a phenotype consistent with schizophrenia.
4.3.5.3 Summary of Psychopathological Effects in Adults
Studies of Pb exposure and behavior in adults have focused on mental disorders rather
than the externalizing behaviors that have been examined in children. Evidence links
occupational Pb exposure with self-reported increases in symptoms of depression and
anxiety, although the cross-sectional design and potential confounding by other
occupational exposures such as Mn limit the implications of findings from these
occupational studies. However, supporting evidence is provided by a few but large
(n = 744 and 1,787) cross-sectional studies in nonoccupationally-exposed adults that
found associations of concurrent blood (Bouchard et al., 2009) and bone (Rajan et al.,
2007) Pb levels with symptoms of depression and anxiety as assessed with widely-used
structured questionnaires such as the Brief Symptom Inventory and the WHO Composite
International Diagnostic Interview. Evidence was provided by the NAS study of men
(primarily white) and study of men and women (various races/ethnicities) participating in
NHANES, both of which involve the examination of multiple exposures and outcomes.
Although specific factors varied among studies, studies in adults with and without
occupational Pb exposure found associations with adjustment for several confounding
factors, including age, sex, and education, and in a few studies, smoking and alcohol use.
The cross-sectional nature of these studies makes uncertain the temporal sequence
between Pb exposure and development of psychopathological effects and the critical
level, timing, frequency, and duration of Pb exposure. Both blood and bone Pb levels in
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adults reflect cumulative exposure, and it is uncertain what are the relative contributions
of past versus recent Pb exposures to the observed associations. The epidemiologic
evidence is supported by observations that Pb exposure (gestational-lactational,
lactational, postlactational) induces depression-like and emotional phenotypes in animals
(Section 4.3.4.2). The mode of action is supported by evidence for Pb-induced changes in
the HP A axis and dopaminergic and GABAergic CNS processes, which are involved in
regulating mood and emotional state.
An epidemiologic study found a cord plasma 5-ALA-schizophrenia association but was
insufficiently informative because of the use of an indirect biomarker of Pb exposure.
However, toxicological studies have shown that Pb exposure decreases NMDAR function
and hippocampal DG neurogenesis, which are found in animal models of schizophrenia
(agitation, trouble finding food, reduced swimming behavior).
Epidemiologic evidence also indicates associations of Pb biomarker levels with
symptoms of depression and anxiety in children and adults as rated by self, parents, or
teachers (Section 4.3.4.1). Differences between children and adults for a relationship
between Pb and other behaviors may be related to differences in what endpoints are
examined. Several studies in children and young adults have examined attention and
conduct disorders; studies of older adults did not examine such behaviors. Differential
effects in children and adults also may be expected given the predominance of different
neurophysiological processes operating at different ages, for example, neurogenesis and
brain development in children and neurodegeneration in adults. Differences in the effects
of Pb exposure between children and adults also may be related to differences in Pb
exposure profiles by age.
4.3.6 Sensory Organ Function
4.3.6.1 Auditory Function
Epidemiologic Studies of Auditory Function in Children
Although not as widely examined as cognitive and behavioral outcomes, several studies
found associations of higher blood Pb level with higher hearing thresholds or poorer
auditory processing in children, with evidence limited largely to that described in the
2006 Pb AQCD (U.S. EPA. 2006b). The prospective Cincinnati study with repeat
measures of blood Pb prenatally to age 5 years provided information on the temporal
sequence between Pb exposure and auditory effects and potential critical time periods of
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Pb exposure and had extensive consideration for potential confounding. In 215 Cincinnati
subjects at age 57 months, poorer auditory processing was associated with higher prenatal
maternal, neonatal (age 10-day), yearly age 1 to 5 year (means: 10.6-17.2 (ig/dL), and
lifetime average blood Pb levels, with the strongest associations found for neonatal blood
Pb level (mean: 4.8 [SD: 3.3] (ig/dL) (Dietrich et al., 1992). A 1 (ig/dL higher neonatal
blood Pb level was associated with a 0.26-point (p < 0.10) and 0.20-point (p < 0.01)
lower score on the total and left ear Filtered Word test (indicating incorrectly identified,
filtered, or muffled words), respectively, with adjustment for hearing screen, social class,
HOME score, birth weight, gestational age, obstetrical complications, and maternal
alcohol consumption. Overall, the findings pointed to a stronger effect of Pb exposure in
infancy.
Additional support was provided by the large U.S. cross-sectional NHANES II
(n = 4,519) (Schwartz and Otto. 1987) and Hispanic Health and Nutrition Examination
Survey (HHANES, n = 3,262) (Schwartz and Otto. 1991) studies. The examination of
multiple exposures and outcomes in these studies reduces the likelihood of participation
conditional specifically on Pb exposure and hearing function. Each study found an
association between higher concurrent blood Pb level and higher hearing thresholds in
children (ages 4-19 years). In HHNANES, an increase in concurrent blood Pb level
(median: 8 (ig/dL) from 6 to 18 (ig/dL also was associated with a 15% increase in the
percentage of children with a substandard hearing threshold at 2,000 Hz. Higher
concurrent blood Pb level also was associated with higher hearing thresholds across
several frequencies in a smaller (n = 155) study of similarly aged (4-14 years) children in
Poland with similar blood Pb levels (median: 7.2 (ig/dL [range: 1.9-28 (ig/dLJ) (Osman et
al.. 1999). In each of the studies in children, associations persisted with concurrent blood
Pb levels <10 (ig/dL. Each of these studies adjusted for different potential confounding
factors, but in stepwise regression analyses, most considered parental education, SES,
colds, and nutritional factors. Across studies, associations were found with adjustment for
factors such as age, sex, ethnicity, family income, concurrent or past colds, antibiotic use,
degree of urbanization, and Apgar score (assessment of newborn overall well-being).
Mechanistic support for associations with higher hearing thresholds in children was
provided by a few studies that found associations of blood Pb level with lower brainstem
auditory evoked potentials in children. Brainstem auditory evoked potentials measure
nerve electrical activity and are indicators of neurological auditory function. In
prospective analyses of the Mexico City cohort, decreased brainstem auditory evoked
potentials were associated with prenatal blood Pb levels at age 9 days (n = 30)
(Rothenberg etal., 1994b) and ages 5-7 years and with postnatal blood Pb levels at ages
5-7 years (n = 100-113) (Rothenberg et al.. 2000). At age 5-7 years, the shape of the
concentration-response relationship differed between prenatal maternal and postnatal
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(ages 1 and 4 years) blood Pb level. Linear associations were found for age 1 year and
age 4 year blood Pb level (means reported for age 18 and 60 months: 10.8 and 8.0 ug/dL,
respectively). Prenatal maternal blood Pb levels showed a biphasic relationship, with
lower evoked potentials found with increasing blood Pb levels 1-8 ug/dL and higher
evoked potentials found with increasing blood Pb levels 8-30 ug/dL. Results were
adjusted for age, sex, and head circumference. Associations with increased latencies of
auditory evoked potentials were found in small studies (n = 29, 49) of children and adults
with higher blood Pb levels (i.e., range 6-84 ug/dL) than most of the current U.S.
population of children (Holdstein et al. 1986; Robinson et al., 1985).
Recent cross-sectional studies aimed to identify the locus in the auditory system affected
by Pb exposure in the examination of a population of children (n = 53, 117, ages 2-18
years) living in Pb glazing communities in Ecuador with higher blood Pb levels than
those relevant to this ISA (means 34 and 38 ug/dL) (Buchanan et al., 2011; Counter et
al.. 2011). Concurrent blood Pb level was not correlated with the acoustic stapedius reflex
(Counter et al., 2011) or distortion product otoacoustic emissions (Buchanan et al., 2011),
indicating lack of effect on the auditory brainstem or inner ear, respectively. Other loci
were not examined, and potential confounding was not considered.
Epidemiologic Studies of Auditory Function in Adults
Studies of auditory function reviewed in the 2006 Pb AQCD consistently indicated
associations between blood Pb levels and prolonged latencies in auditory evoked
brainstem potentials in occupationally-exposed adults (U.S. EPA. 2006b). A few recent
studies found increases in hearing thresholds in Pb-exposed workers. A recent analysis of
the NAS provided evidence in nonoccupationally-exposed men for associations of tibia
Pb levels with hearing loss.
Among 448 men in the NAS, higher tibia Pb level (mean: 22.5 ug/g) at mean age 64.9
years, measured up to 20 years after initial hearing testing, was associated with a faster
rate of increase in hearing threshold for frequencies of 1, 2, and 8 kHz and a pure tone
average. Men were free of hearing loss at baseline and had hearing tested repeatedly
(median 5 times per subject) over a median of 23 years (Park etal.. 2010). Blood Pb was
not examined. In cross-sectional analyses adjusted for age, race, education, body mass
index, pack-years of cigarettes, diabetes, hypertension, occupational noise (based on a
job-exposure estimate), and presence of a noise notch (indicative of noise-induced
hearing loss), higher patella Pb level (mean 32.5 ug/g, measured within 5 years of hearing
test) was associated with a higher hearing threshold for frequencies greater than 1 kHz. A
21 ug/g (interquartile range) increase in patella Pb level was associated with pure tone
average hearing loss with an OR of 1.48 (95% CI: 1.14, 1.91) in adjusted analyses.
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Similar, but slightly weaker associations were found for tibia bone Pb levels. In the NAS,
bone Pb levels were measured after the initial hearing measurement but reverse causation
is unlikely since bone Pb is an indicator of cumulative Pb exposure, and tibia Pb has a
half-life on the order of decades (Section 3.3). Bone Pb levels increase with age, and
although age was included as a model covariate, residual confounding by age is possible.
Recent cross-sectional studies added evidence for associations between higher concurrent
blood Pb levels and higher hearing thresholds in adults with potential or actual
occupational Pb exposures. A hospital-based case-control study examined workers from
diverse occupations and examined potential confounding by other occupational
exposures. Cases included workers referred for hearing testing (average hearing
thresholds above 25 dB), and controls comprised workers with normal hearing thresholds
who were having routine occupational health examinations (Chuang et al. 2007).
Geometric mean blood Pb levels were 10.7 ug/dL forthe 121 cases and 3.9 ug/dL for the
173 controls. In models that adjusted for age, smoking status, alcohol consumption, years
of noise exposure, as well as Mn, arsenic (As), and selenium (Se) levels in blood, higher
blood Pb levels were associated with higher hearing threshold (0.5-6 kHz). The potential
selection bias arising from the nonrandom population sample may limit implications of
these findings. Other studies found associations of higher concurrent blood Pb level with
increased hearing thresholds or hearing loss in Pb-exposed workers (n = 183-259) but had
limited or no consideration for potential confounding (Hwang et al.. 2009; Forst et al..
1997) and/or examined workers with a mean blood Pb level of 57 ug/dL (Wu et al.,
2000).
lexicological Studies of Auditory Function
The 2006 Pb AQCD (U.S. EPA. 2006b) described increases in hearing thresholds or
latencies in brainstem auditory evoked potentials in nonhuman primates (n = 5-7)
exposed to lifetime Pb from gestation or birth to ages 8-13 years (resulting in blood Pb
levels 33-150 (ig/dL at peak or during testing) (Rice. 1997; Lilienthal and Winneke.
1996). Rice (1997) found that half of the pure tone detection thresholds were above the
control range at certain frequencies. In Lilienthal and Winneke (1996). the Pb-related
increased latency in auditory evoked potentials persisted after Pb exposure was
terminated, and blood Pb levels had decreased to < 5 (ig/dL. The observations in
monkeys are consistent with the epidemiologic associations described above but were
related to higher Pb exposures than those relevant to this ISA. In addition to indicating
hearing loss, brainstem auditory evoked potentials can indicate impaired synaptic
maturation and incomplete neuron axon myelination leading to impaired neuronal
conduction (Gozdzik-Zolnierkiewicz and Moszynski. 1969). Thus, the findings from Rice
(1997) and those described in the preceding sections for children may indicate that Pb
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exposure impairs auditory nerve conduction. Studies in animals with blood Pb levels
>300 (ig/dL found that the cochlear nerve was especially sensitive to Pb exposure
(Gozdzik-Zolnierkiewicz and Moszynski. 1969).
A recent study found similar effects in monkeys with similar blood Pb levels as Lilienthal
and Winneke (1996). but with only early-life Pb exposure. Laughlin et al. (2009) studied
rhesus monkeys exposed to Pb acetate prenatally to birth or postnatally from birth
through weaning at age 5.5 months (maternal drinking water, 3 months prior to mating
until weaning, resulting in bone Pb levels at age 11 years of 7 and 13 (ig/g for prenatal
and postnatal groups, respectively, and blood Pb levels averaged over Pb exposure of 35
and 46 (ig/dL, respectively). Auditory threshold testing and threshold task testing was
conducted at 13 years of age after blood Pb levels had returned to those found in controls.
At birth, animals were cross fostered, creating a control group, a prenatal Pb group, and a
postnatal Pb group; however, Pb-exposed animals were analyzed as a single group. Pb
exposure induced small, statistically nonsignificant elevations in auditory thresholds in
animals. Behavioral testing during the auditory threshold task indicated less engagement
and responsiveness in Pb-exposed animals. This study has multiple limitations that could
have contributed to the lack of statistically significant aberrations, including limited
power with the examination of 5 animals per group, the inability of some of the monkeys
to engage or focus on the task at hand which resulted in fewer available measurements,
differences between the sexes in attention, and mixing of the postnatal Pb and prenatal
Pb-exposed animals into one group.
Summary of Auditory Function
Children
Several studies indicated that higher blood Pb levels are associated with decrements in
auditory function in children, as evidenced by increases in hearing thresholds. Results
from the prospective Cincinnati cohort study (n = 215) provide key evidence for
associations of prenatal maternal, neonatal (age 10 day), yearly age 1 to 4 year, and
lifetime average blood Pb levels with increased hearing thresholds at age 57 months
(Dietrich et al., 1992), and large (n = 3,000-4,000) cross-sectional analyses of children
participating in NHANES and HHANES provide supporting evidence for concurrent
blood Pb levels (Schwartz and Otto. 1991. 1987). The examination of multiple exposures
and outcomes in NHANES and HHANES reduces the likelihood of participation
conditional specifically on Pb exposure and hearing function. In the Cincinnati cohort,
mean blood Pb levels were 4.8 (ig/dL for neonatal and 17.4 (ig/dL for lifetime average. In
HHANES, the median concurrent blood Pb level was ~8 (ig/dL. Across studies,
associations were found with adjustment for factors such as age, sex, ethnicity, family
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income, concurrent or past colds, antibiotic use, degree of urbanization, and Apgar score.
Potential confounding by parental education, nutritional factors, environmental noise, and
maternal smoking also was considered.
Mechanistic evidence was provided by observations of associations between blood Pb
level with lower auditory evoked potentials in children, particularly associations found in
the prospective analysis of children ages 5-7 years in Mexico City with prenatal maternal,
age 1 year, and age 4 year blood Pb levels (Rothenberg et al.. 2000). Pb-related increases
in hearing thresholds and latencies in brainstem auditory evoked potentials also were
found in nonhuman primates with lifetime Pb exposure beginning in gestation or birth
(Rice. 1997; Lilienthal and Winneke. 1996). although auditory assessments were made in
adult animals ages 8-13 years. Shorter-duration Pb exposure from gestation to birth or
from birth to age 5.5 months was found to induce small, statistically nonsignificant
elevations in auditory thresholds and decreased responsiveness in behavioral testing
performed with the threshold task (Laughlin et al.. 2009). In animals, auditory effects
were found with higher blood Pb levels (i.e., 33-150 (ig/dL at peak or testing) than those
relevant to this ISA; thus, it is difficult to ascertain biological plausibility for observations
in children.
Adults
In adults, increased hearing thresholds and hearing loss were associated with bone Pb
levels in 448 NAS men who were unlikely to have had occupational Pb exposures (Park
et al., 2010) and with concurrent blood Pb levels in adults with potential occupational Pb
exposure. In the NAS, the examination of multiple exposures and outcomes reduces the
likelihood of participation conditional specifically on Pb exposure and auditory function.
Among NAS men, higher tibia Pb levels were associated with a faster rate of increase in
hearing thresholds over a 23 year follow-up with adjustment for age, race, education,
body mass index, pack-years of cigarettes, diabetes, hypertension, occupational noise,
and presence of a noise notch. Tibia Pb levels were measured up to 20 years after initial
hearing testing, and while there is uncertainty regarding the temporal sequence between
Pb exposure and changes in hearing thresholds, Pb in tibia has a half-life on the order of
decades. Temporality also is difficult to establish in the cross-sectional occupational
studies. A hospital-based case-control study found an association between higher
concurrent blood Pb levels and higher hearing thresholds among workers with relevant
blood Pb levels (means 10.7 and 3.9 ug/dL in workers with and without hearing
problems, respectively) (Chuang et al., 2007). Among other factors, results were adjusted
for other metal exposures. Similar auditory effects were found in adult animals ages 8-13
years with lifetime Pb exposures but with higher blood Pb levels (i.e., 33-150 (ig/dL) than
those relevant to this ISA.
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4.3.6.2 Visual Function
The sections that follow review studies examining relationships of Pb exposure with
visual function as related to effects on the central nervous system. Studies examining
other effects on the visual system such as cataracts are described in Section 4.9.5.
Epidemiologic Studies of Visual Function
The little available epidemiologic evidence does not clearly describe an effect of relevant
Pb exposures with effects on the retina in humans. Among children (n = 100-113, ages 7-
10 years) in Mexico City examined prospectively, Rothenberg et al. (2002b) reported an
association between maternal first trimester blood Pb levels 10.5-32.5 ug/dL and
supernormal electroretinograms (ERGs), the impact of which on visual impairment is not
clear. In a case-control study of macular degeneration (condition due to retinal damage)
in adults, Pb concentrations were elevated in retinal tissue of 36 cases (50 eye donors at
time of autopsy; median [IQR]: 12.0 [8-18] ng/g Pb) versus 25 controls (72 eyes; median
[IQR]: 8.0 [0-11] ng/g Pb) (Erie et al.. 2009V The implications of results in children and
adults are limited by the lack of rigorous statistical analysis and consideration for
potential confounding. The macular degeneration results are limited further by the
inability to establish directionality of effects and uncertainty in representativeness of the
control population.
lexicological Studies of Visual Function
As described in the 1986 and 2006 Pb AQCDs (U.S. EPA. 2006b. 1986bX a majority of
the evidence for the effects of Pb on the visual system has been provided by toxicological
studies. Pb exposure of animals during perinatal development and adulthood has been
found to reduce visual acuity and produce changes describing potential mechanisms of
action such as alterations in the retina (Foxetal.. 1997; Fox and Sillman. 1979). CNS
visual processing areas (Costa and Fox. 1983). and subcortical neurons involved in vision
(Cline et al.. 1996). For example, environmentally-relevant doses of Pb (10~3 (iM) in
tadpoles inhibited the growth of neurons in the subcortical retinotectal pathway, the main
efferent from the retina (Cline etal. 1996). Pb-related aberrations in electrical responses
in retinal cells, as measured by ERGs, have been found in rodents and nonhuman
primates. Recent research on ERGs expands upon the extant evidence by examining
animals with lower Pb exposures or blood Pb levels.
Work in nonhuman primates with Pb exposure during development or over a lifetime
(peak blood Pb levels 50-115 (ig/dL) showed dysfunction in temporal visual function
(responses to different frequencies of light flicker) under high luminance but no change
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in spatial function (Rice. 1998). A recent study found no effects of infancy Pb exposure
(PND8-age 26 weeks via commercial milk formula, producing blood Pb levels of
35-40 ug/dL) on spatial acuity of rhesus monkeys as assessed with the modified Teller
preferential looking paradigm (Laughlin et al.. 2008). In monkeys, effects on vision were
tested with higher Pb exposures than those relevant to this ISA. Low-level developmental
Pb exposure was found to result in sensorimotor deficits in adult zebrafish (Rice et al..
2011). Fish that were exposed as embryos (2 to 24 hours post-fertilization) to water
containing 0.03 uM PbCl2 had impaired response to visual stimulation (a rotating bar)
under low light conditions. These zebrafish also failed to respond normally to
mechanosensory stimulation (0.01 and 0.03 uM PbCl2), showing a significantly impaired
startle response.
Animal toxicological evidence also shows that the lifestage of exposure and the dose of
Pb contribute to the complex and variable effects of Pb on ERGs (summarized in Table
4-13). The biological relevance of these variable findings is uncertain. Female rats
exposed postnatally to 200 or 2,000 ppm Pb acetate exposure via dam drinking water
from birth through lactation, resulting in blood Pb levels of 19 and 59 ug/dL at weaning,
respectively, had subnormal scotopic ERGs (decreased A- and B-wave amplitudes) with
decreased sensitivity and temporal resolution when assessed at 90 days of age (Fox et al..
1991). Similar results were obtained in multiple studies conducted in in vitro models
(Otto and Fox. 1993: Fox and Farber. 1988: FoxandChu. 1988). Monkeys exposed to
relatively high levels of Pb continuously from the prenatal period to age 7 years (350 or
600 ppm Pb acetate, resulting in blood Pb levels of 40 and 50 ug/dL, respectively) had
persistently increased maximal ERG amplitude (B-wave only, supernormality) and
increased mean ERG latency when assessed 2 years after Pb exposure was terminated
when blood Pb levels were <10 ug/dL (Lilienthal et al.. 1988).
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Table 4-13 Summary of Pb-related effects observed on the visual system.
Study,
Species,
Sex
Fox et al.
(2008)
Long-
Evans Rat,
F
Lilienthal et
al. (1988)
Rhesus
Monkey,
M&F
Fox et al.
(1997)
Long-
Evans Rat,
F
Rothenberg
etal.
(2002b)
Human
children, M
&F
Pb
Exposure
Protocol
and Dose
DW,
Prenatal-
PND10
Low, 27 ppm
Moderate,
55 ppm
High,
109 ppm
DW,
Prenatal-
lifetime
350 ppm
600 ppm
DW,
PND1-21
200 ppm
2,000 ppm
Prenatal
maternal
1st trimester
blood
Maximal
Blood Pb
Level
(ug/dL)
12
24
46
-50
-115
19
59
10.5-32.5
Progenitor Retinal Retinal
Cell Pro- Cell Dopamine
ERG Effect liferation Apoptosis Levels
Supernormal Yes „ °. , , °^.e" .
Supernormal Yes "oi. . . Dos/- .,
r affected dependent J,
Subnormal No Yes , °^.e" .
Supernormal — — —
Supernormal — — —
Subnormal — Yes —
Subnormal — Yes —
Supernormal — — —
Retinal
Cell Layer
Thickness
T
T
I
I
I
—
F = females, M = males, DW = Drinking water exposure, - denotes not examined in the study.
Pb exposure beginning in the gestational period (Pb acetate in drinking water from
2 weeks before mating to PND10) also induced supernormal ERGs in adult Long-Evans
rats, with low (27 ppm) and moderate (55 ppm) doses that produced blood Pb levels
10-12 (ig/dL and 21-24 (ig/dL (Fox et al.. 2008) (Figure 4-11 and Table 4-13). This
exposure window represents the developmental period for the retina of the rat and is
analogous to gestational human retinal development. Subnormal ERGs were induced by
the high 109 ppm dose (Figure 4-11). which produced blood Pb levels 40-46 (ig/dL.
Results of this rodent study demonstrated persistent supernormal scotopic rod
photoreceptor-mediated ERGs in animals with blood Pb levels relevant to this ISA. These
findings were consistent with the association observed between supernormal ERGs and
prenatal maternal blood Pb levels 10.5-32.5 ug/dL in male and female children
(Rothenberg et al.. 2002b). The functional relevance of findings is uncertain as
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supernormal scotopic ERGs may be recorded without other overt ophthalmologic
changes and are rarely seen in the clinical setting (Terziivanov et al.. 1982).
I5t
r±
Hiflh iruansitv
(test stimulus 21
H il- -it.,1.11-,
ItfEtSliCWllUtJI
II •]!• 'il iti '..ly
IIKfl stimulus II
Note: *p <0.05. Low Pb = 27 ppm Pb acetate, blood Pb level 10-12 ug/dL, Moderate Pb = 55 ppm, blood Pb level 21-24 ug/dL,
High Pb = 109 ppm, blood Pb level 40-46 |jg/dL. Relative to controls (gray bars), low (white bars) or moderate (blue bars)
Pb exposure from gestation through postnatal day (PND)10 induced supernormal ERGs whereas high Pb exposure (black bars)
induced subnormal ERGs.
Source: Fox et al. (2008)
Figure 4-11 Retinal a-wave and b-wave electroretinogram (ERG) amplitude in
adult rats after prenatal plus early postnatal Pb exposure.
Animal studies indicate that the dose of Pb and the exposure lifestage not only
differentially affect functional tests, i.e., ERG, but also differentially affect retinal cell
numbers and morphology. Concomitant with Pb-induced supernormal ERGs, Fox et al.
(2008) found that 27 and 55 ppm gestational plus early postnatal Pb exposure increased
neurogenesis of rod photoreceptors and rod bipolar cells without affecting apoptosis of
Miiller glial cells and increased the number of rods in central and peripheral retina (Table
4-13). Concomitant with subnormal ERGs, higher-level gestational plus early postnatal
Pb exposure (109 ppm, blood Pb level 40-46 (ig/dL) decreased the number of rods in the
central and peripheral retina (Fox et al.. 2008). Early postnatal (PND1-PND21) Pb
exposure (200 or 2,000 ppm, producing blood Pb levels 19 and 59 (ig/dL) induced
scotopic ERG subnormality in adult rats, decreased the number of rods in the central and
peripheral retina, and decreased the retinal Zn concentration (Foxet al.. 1997) (Table
4-13). Similar observations were made in separate work in mice. Low and moderate
doses of Pb from gestation to PND10 (27 or 55 ppm Pb acetate in dam drinking water,
producing blood Pb levels of 12 and 25 ug/dL, respectively) induced greater and
prolonged rod bipolar cell neurogenesis and greater thickness and cell number of the
outer and inner neuroblastic layers of the retina (Giddabasappa et al.. 2011). As in rats,
with a higher dose of Pb (109 ppm Pb acetate, resulting in blood Pb levels of 56 ug/dL),
there was no increased rod neurogenesis in mice. Nitric oxide has been shown to regulate
retinal progenitor cell proliferation in chick embryos (Magalhaes et al.. 2006). Thus,
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these authors postulated that impaired NO production may contribute to aberrant retinal
cell proliferation (Giddabasappa et al.. 2011).
Mechanistic understanding of the effect of Pb on the visual system includes the capability
of Pb to displace divalent cations, inhibit physiological enzymes, alter cell proliferation
and apoptosis, perturb normal neuroanatomical formation (cytoarchitecture in the brain),
and affect neurotransmitters. The effects of Pb on the retina were shown to be mediated
by inhibition of cGMP phosphodiesterase (PDE) by Pb (Srivastava et al.. 1995; Fox and
Farber. 1988). Without regard to Pb exposure, pharmacological inhibition of cGMP PDE
has been linked with visual problems including alterations in scotopic ERGs (Laties and
Zrenner. 2002). Postnatal Pb exposure of animals (peak blood Pb levels: 19, 59 (ig/dL) or
in vitro Pb exposure of rods isolated from these animals elevated cGMP which
contributed to elevated rod calcium concentration (Fox and Katz. 1992) and subsequently
induced apoptotic cell death in a concentration-dependent manner.
Pb has been shown to affect a plethora of neurotransmitters in the brain, and it has
recently been shown to affect neurotransmitters in the retina. With the aforementioned
gestational to PND10 exposure, Pb induced a concentration-dependent decrease in adult
rat retinal synthesis of dopamine, which has functions in retinal growth and development
and signal transduction in rods and cones (Fox et al., 2008) (Figure 4-12). As discussed in
the 2006 Pb AQCD (U.S. EPA. 2006b). other mode of action support for the effects of Pb
on the visual system is provided by observations of Pb-induced decreased Na+/K+ATPase
activity which have been reported in vitro and in vivo. Also, structural changes from
chronic Pb exposure (birth to age 6 years) of monkeys included cytoarchitectural changes
in visual projection areas of the brain. Maximum blood Pb levels in the low and high
dose groups reached 20 (ig/dL and 220 (ig/dL, respectively (Reuhl et al.. 1989).
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Control Low Moderate High
GLE
Note: *p <0.05. GLE = Gestational Pb exposure to postnatal day 10. Low Pb = 27 ppm Pb acetate, blood Pb levels 10-12 ug/dL,
Moderate Pb = 52 ppm, blood Pb levels 21-24 ug/dL, High Pb = 109 ppm, blood Pb levels 40-46 ug/dL. A. DA = dopamine,
B. DOPAC = dopamine metabolite, C. HVA = dopamine metabolite, D. DOPAC/DA and HVA/DA = ratio of dopamine metabolite to
dopamine. Pb exposure decreased dopamine, DOPAC, and HVA in a concentration-dependent manner in light-adapted animals
(blue bars). In dark adapted animals (black bars), Pb exposure decreased dopamine, DOPAC, and HVA but not always in
concentration-dependent manner. Pb exposure decreased DOPAC/DA and HVA/DA.
Source: Fox et al. (2008)
Figure 4-12 Retinal dopamine metabolism in adult control and gestationally
Pb-exposed (GLE) rats.
Summary of Visual Function
Increasing maternal first trimester blood Pb levels (10.5-32 Lig/dL) were associated with
supernormal ERGs in children in Mexico City at ages 5-7 years (Rothenberg et al..
2002b). The animal evidence showed ERGs in different directions depending on lifestage
of Pb exposure and blood Pb level. Supernormal ERGs were found in adult rats with
prenatal plus early postnatal (PND10) Pb exposure that produced blood Pb levels of 12
and 24 ug/dL (Fox et al., 2008). The implication of supernormal ERGs on visual
impairment is not clear, and the biological relevance of the nonlinear concentration-
response is not clear. In these animals, Pb exposure also increased rod cell neurogenesis
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and decreased dopamine. Toxicological studies also found other retinal effects including
alterations in morphology and cell architecture, signaling, enzyme inhibition,
neurotransmitter levels, neuroanatomical development, cell proliferation, and retinal cell
apoptosis. Adult monkeys were found to have supernormal ERGs and decreased temporal
visual function with developmental or lifetime Pb exposure (Rice. 1998). but infant
monkeys had no change in visual acuity with infancy Pb exposure (Laughlin et al.. 2008).
In each case, Pb exposures were higher (blood Pb levels > 35 (ig/dL) than those relevant
to this ISA. Pb-associated visual system effects in human adults are not well
characterized and are limited to a case-control study finding higher Pb concentrations in
retinal tissue from macular degeneration cases (Erie et al., 2009) that did not perform
rigorous statistical analysis or examine potential confounding factors.
4.3.7 Motor Function
Some studies in children have assessed fine motor function, i.e., response speed,
dexterity, as part of a battery of neurodevelopmental testing, and most have found
associations with blood Pb level. Fewer studies have examined gross motor function,
i.e., postural balance, action tremor, agility, but also have found associations with blood
Pb level in children. Poorer motor function also was found in Pb-exposed workers.
Key evidence from prospective analyses of the Cincinnati and Yugoslavia cohorts
demonstrated associations of blood Pb levels with poorer motor function with either
adjustment for or consideration for several potential confounding factors related to SES,
parental caregiving quality and education, smoking exposure, birth outcomes, sex, and
child health. In the Cincinnati cohort, higher earlier childhood blood Pb levels (age
0-5 year average [median: 11.7 (ig/dL] or age 78 month) were associated with poorer fine
(n = 195) (Ris et al.. 2004) and gross motor function (n = 91) (Bhattacharya et al.. 2006)
assessed in adolescence (ages 12, 15-17 years). In this cohort recruited from birth,
follow-up participation was high and not conditional on blood Pb levels. The fine motor
function results were adjusted for maternal IQ, SES, HOME score, sex, and adolescent
marijuana consumption. Collectively, these findings suggest the persistence of effects of
earlier childhood Pb exposure to later childhood; however, later childhood or concurrent
blood Pb levels were not examined. Assessments in Cincinnati cohort children at age 6
years indicated associations of concurrent (mean: 10.1 (ig/dL), lifetime average (mean:
12.3 (ig/dL), and neonatal (mean: 4.8 (ig/dL) but not prenatal maternal (mean: 8.4 (ig/dL)
blood Pb levels with poorer upper limb dexterity, fine motor composite score (n = 245)
(Dietrich et al.. 1993a). and poorer postural balance (n = 162) (Bhattacharya et al.. 1995).
These results were adjusted for HOME score and race. Additional covariates included
maternal IQ, SES, and sex for fine motor functions (Dietrich et al.. 1993a) and height,
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BMI, birth weight, bilateral ear infection, nutritional factors, and foot area for postural
balance (Bhattacharya et al.. 1995). Blood Pb levels were associated with fine and gross
motor function in unadjusted and adjusted analyses, increasing confidence that
confounding by the examined covariates did not unduly bias the observed associations.
Prospective analysis of the Yugoslavian cohort indicated an association of lifetime
average blood Pb level (exact levels not reported) with decrements in fine but not gross
motor function at age 4.5 years in 283 children (Wasserman et al., 2000a). Although only
50% of the cohort was examined and participation was greater among children with
lower SES and HOME score, participation was not conditional on higher blood Pb levels.
Evidence from cross-sectional studies for associations between motor function and
concurrent blood Pb level was mixed, with studies examining mean concurrent blood Pb
levels 2-3 (ig/dL not finding decrements in motor function. Surkan et al. (2007) found
that higher concurrent blood Pb level was associated with better fine motor function as
indicated by faster finger tapping speed among 534 children (ages 6-10 years) in
New England. Results were adjusted for several potential confounding factors such as
age, sex, caregiver IQ, SES, and race. Min et al. (2007) found weak, imprecise
associations in 61 similarly aged children (7-16 years) in Korea. Concurrent blood Pb
level (mean: 5.0 (ig/dL) was associated with greater sway oscillation, action tremor, and
alternating arm movements in 110 Inuit children (ages 4-6 years) in Quebec, Canada
(Despres et al.. 2005) with consideration for potential confounding by factors such as
HOME score, maternal education and several nutrient levels. The population of Inuit
children was selected from subsistence fishing communities, who also have higher
exposure to Hg and PCBs. Several indices of fine and gross motor function were
associated with blood Pb level, with adjustment for these other exposures. Poorer fine
motor function also was associated with higher concurrent blood Pb level in 814 children
ages 3-7 years in India with mean concurrent blood Pb level 11.4 (ig/dL with adjustment
for similar factors as the aforementioned studies (Palaniappan et al., 2011).
An association of Pb exposure with poorer motor function in adults was found in
Pb-exposed workers (Iwata et al.. 2005). although implications are limited by the cross-
sectional design, high concurrent blood Pb levels (mean: 40 ug/dL), and lack of
consideration for potential confounding by other occupational exposures. Among 121 Pb
smelter workers in Japan, higher blood Pb level was associated with greater sagittal sway
with eyes open (p <0.05) and eyes closed (p <0.01) and transversal sway with eyes closed
(p <0.05) with adjustment for age, height, smoking status, and drinking status. A mean
benchmark dose level (Budtz-Jorgensen et al.. 2001; NRC. 2000) of 14.3 ug/dL was
calculated for postural sway from a linear model. A supralinear concentration-response
function fit the data slightly better than a linear function did.
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Pb exposure has shown mixed effects on endurance, balance, and coordination in animals
as measured by rotarod performance, treadmill testing, and visual discrimination task.
Lower concentration gestational plus early postnatal (to PND10) Pb exposure (27 ppm,
producing peak blood Pb level < 10 (ig/dL at PNDO-PND10) resulted in significantly
poorer rotarod performance (i.e., falling off more quickly) than did higher exposure
(109 ppm, blood Pb level: 33-42 (ig/dL) in male (but not female) adult mice, indicative of
a nonlinear concentration-response relationship (Leasure et al., 2008). Other rotarod
experiments examining various speeds of rotarod rotation, higher than relevant Pb
exposures (blood Pb levels >59 (ig/dL), and/or routes of exposure with uncertain
relevance to humans, yielded mixed results (Kishi et al., 1983; Grant etal., 1980;
Overmann. 1977). An effect on motor function was indicated by increased alcove latency
of rats performing a visual discrimination task but with Pb exposure (PND1-30) blood Pb
levels of 31 (ig/dL (Stangle et al.. 2007). Herring gull chicks injected with a single i.p.
bolus dose of Pb (100 mg/kg Pb acetate, selected to represent exposure in the wild) had
slower development of motor skills versus control birds, as assessed by the treadmill test
(Burger and Gochfeld. 2005).
In summary, epidemiologic evidence demonstrates associations of higher blood Pb levels
with poorer fine and gross motor function in children ages 3-17 years. Little evidence is
available in adults. Prospective analyses of the Cincinnati and Yugoslavia cohorts
(n = 91-283) that considered several potential confounding factors such as SES and child
health found associations with earlier childhood blood Pb levels (i.e., age 78 month, 0-5
year average) in adolescents (Bhattacharya et al.. 2006; Ris et al.. 2004) and neonatal,
lifetime average, and concurrent blood Pb levels in children ages 4-6 years (Wasserman
et al.. 2000b; Bhattacharya et al.. 1995; Dietrich et al.. 1993a). In the Cincinnati cohort,
neonatal blood Pb levels were lower than concurrent or lifetime average blood Pb levels
at age 6 years (means 4.8, 10.1, 12 (ig/dL, respectively). In cross-sectional studies that
examined similar potential confounding factors, results were inconsistent in populations
with mean blood Pb levels 2-5 (ig/dL (Min et al.. 2007; Surkan et al.. 2007; Despres et
al.. 2005). While Pb exposure had mixed effects on motor function in animals, mice with
relevant blood Pb levels (peak < 10 (ig/dL) after gestational-early postnatal Pb exposure
showed poorer balance (Leasure et al., 2008).
4.3.8 Seizures in Animals
Previous studies did not consistently show that Pb exposure induced seizures in animals.
Pb-induced seizures were found in male Wistar rats exposed to Pb acetate postnatally
(250-1,000 ppm in drinking water PND60-PND90, resulting in blood Pb levels of
20-43 (ig/dL), as indicated by a decrease in the elapsed time required to develop the first
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myoclonic jerk and tonic-clonic seizure (Arrieta et al.. 2005). Also, the dose of the
seizure-inducing agent pentylenetetrazol (PTZ) required to induce seizures significantly
decreased in all Pb dose groups. Other studies showed no effect of Pb exposure on
seizures (Schwark et al.. 1985; Alfano and Petit 1981). In a study of early postnatal
Pb acetate exposure (2,000 ppm in drinking water PND1-PND25), Pb had variable effects
on induction of seizures in Sprague Dawley rats examined at PND25 or PND50,
depending on the convulsant-inducing agent administered (Chen and Chan. 2002). These
variable effects were hypothesized to be due to the selective effects on inhibitory and
excitatory neurotransmission by age and blood Pb level, which were 47 (ig/dL and
11 (ig/dL at PND25 and PND50, respectively.
Recent investigation expanded on the work by Arrieta et al. (2005) by showing that Pb
exposure may induce seizure activity in another rodent species, BALB/c mice. Adult
(ages 2-3 months) male BALB/c mice were exposed to Pb acetate for 30 days via
drinking water (range of blood Pb levels 50-400 ppm Pb groups: 6.4-18 (ig/dL)
(Mesdaghinia et al.. 2010). Except for 50 ppm Pb exposure, all other Pb concentrations
significantly reduced the thresholds efface and forelimb clonus, myoclonic twitch,
running and bouncing clonus, and tonic hindlimb extension. High-level Pb administered
by bolus injection (200 mg/kg Pb acetate or 50 mg/kg Pb nitrate, single injection, 2 days,
blood Pb levels >90 (ig/dL) also induced epileptic form activity or seizures in adult male
Wistar rats (Krishnamoorthy et al.. 1993).
4.3.9 Neurodegenerative Diseases
4.3.9.1 Alzheimer's Disease
Lower scores on the MMSE, which is widely used as a screening tool for dementia, have
been associated with higher bone Pb level in NAS men (Wang et al.. 2007a: Weisskopf et
al., 2004; Wright et al., 2003) but inconsistently with blood Pb level in adults (Weuve et
al.. 2006; Nordberg et al.. 2000). Direct evidence regarding the effects of Pb exposure on
Alzheimer's disease is limited to studies reviewed in the 2006 Pb AQCD (U.S. EPA.
2006b). which did not find higher occupational exposure to Pb (Graves et al.. 1991) or Pb
level in the brains (Haraguchi et al.. 2001) in Alzheimer's disease cases than unaffected
controls. Overall, the latter studies have sufficient limitations (e.g., case-control design
that may be subject to reverse causation, limited consideration for potential confounding)
such that evidence is inconclusive regarding the effect of Pb exposure on Alzheimer's
disease.
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Despite inconclusive epidemiologic evidence, toxicological evidence indicates that Pb
exposure in early life promotes Alzheimer's disease-like pathologies in the brains of aged
adult animals. Alzheimer's disease is characterized by amyloid-beta peptide (Ab)
plaques, hyper-phosphorylation of the tau protein, neuronal death and synaptic loss. In
the last decade, the developmental origins of adult health and disease paradigm and the
similar Barker hypothesis have indicated that early life exposures can produce aberrant
effects in adults. Bolin et al. (2006) connected oxidative DNA damage in adult rats with
indicators of plaque formation after lactational Pb exposure that produced blood Pb levels
of 46 (ig/dL. Wu and colleagues (2008a) made similar observations with infantile Pb
exposure of monkeys. These results suggest the need to directly examine the long-term
effects of developmental exposure to toxicants rather than relying on adult exposure
alone to predict potential health risks in adults (Dietert and Piepenbrink. 2006).
The fetal basis of amyloidogenesis has been examined extensively by the Zawia
laboratory in both rodents and nonhuman primates. Mechanistically, Ab plaques originate
from the cleavage of the amyloid precursor protein (APP). In male rats exposed to Pb as
infants (200 ppm Pb acetate PND1-PND20 in dam drinking water, resulting in pup
PND20 blood Pb level of 46 (ig/dL and cortex 0.41 (ig/g wet weight of tissue) or as
adults, infancy Pb exposure induced APP gene expression in the aged animal brains
(Basha et al., 2005). A bimodal response was observed, with a significant increase in
APP expression above that in control animals first manifesting in infancy and again in old
age (20 months). A concomitant bimodal response was observed in specificity protein 1
(Spl), a transcription factor involved in gene expression in the early development of an
organism and known to be related to APP expression. Ab was also significantly elevated
in the aged animals developmentally exposed to Pb. Adult-only (age 18-20 months)
exposure to Pb did not alter APP or Spl expression or Ab production.
Consistent with findings in rodents, Wu et al. (2008a) found that Pb exposure in infancy
(PND1-PND400, 1.5 mg/kg per day in infant formula) resulted in significantly higher
gene expression of APP and Spl and significantly higher protein expression of APP and
Ab in aged female monkey cortex tissue (23 year-old Macaca fascicularis) from a cohort
of animals established in the 1980s by Rice (1992a. 1990). At PND400, the monkeys had
blood Pb levels of 19-26 (ig/dL. In old age when amyloid plaques had manifested, blood
Pb levels and brain cortex Pb levels had returned to control levels. Together, the rodent
and nonhuman primate evidence concurs, and indicates that developmental Pb exposure
and not adult-only exposure induces elevations in neuronal Alzheimer's Disease-related
plaque proteins in aged animals.
Mechanistic understanding of Ab production and elimination after Pb exposure was
examined in human SH-SY5Y neuroblastoma cells exposed to Pb concentrations of 0, 5,
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10, 20, and 50 (iM for 48 hours. Pb was found to affect two separate pathways to increase
Ab. Pb exposure induced both the overexpression of APP and repression of neprilysin, a
rate-limiting enzyme involved in Ab metabolism or removal (Huang et al., 201 la).
Further mechanistic understanding of how Ab peptide formation is affected by Pb
exposure was provided by Behl et al. (2009). The choroid plexus is capable of removing
beta-amyloid peptides from the brain extracellular matrix. Pb was shown to impair this
function, possibly by inhibiting the metalloendopeptidase, insulin-degrading enzyme,
which metabolizes Ab (Behl et al.. 2009). In another study, lactational Pb exposure of
Long-Evans hooded rat pups induced perturbations in DNA binding of SP1 via its Zn
finger protein motif (Basha et al., 2003). This effect of Pb was ameliorated by exogenous
Zn supplementation.
An additional study with gestational plus lactational Pb exposure (1,000-10,000 ppm,
dam drinking water, resultant offspring blood Pb levels: 40-100 (ig/dL) showed that the
rodent hippocampus as early as PND21 contained neurofibrillary changes, commonly
used a marker for Alzheimer's disease. These changes manifested as hyper-
phosphorylated Tau, which comprises neurofibrillary tangles, and increased tau and beta
amyloid hippocampal protein levels (Li et al.. 2010b). Neurofibrillary changes were
accompanied by poorer performance in tests of learning and memory.
In summary, recent studies showed that Pb exposure of rats and monkeys during infancy
or during gestation/lactation induced significant increases in neuronal plaque-associated
proteins such as Ab-peptide, activation of Ab-supporting transcription factors, and
hyperphosphorylation of tau, all of which are pathologies found in humans with
Alzheimer's disease. These pathologies were not found with adult-only Pb exposure of
animals, further demonstrating that early life Pb is a critical window for Pb-induced
Alzheimer's-like pathologies in animals. The few epidemiologic studies have not linked
higher Pb exposure with Alzheimer's disease. These case-control studies lacked
consideration for potential confounding. The animal evidence indicates that
epidemiologic studies assessing concurrent brain Pb levels or occupational Pb exposure
may not have examined the etiologically relevant exposure period. However, the
observations that were made in experimental animals with high Pb exposure and blood
Pb levels (40-100 (ig/dL) may have uncertain relevance to this ISA. Further, animals
were not behaviorally assessed for dementia.
4.3.9.2 Amyotrophic Lateral Sclerosis
The 2006 Pb AQCD (U.S. EPA. 2006b) reported mixed epidemiologic findings for an
association between Pb and ALS based on case-control studies, several of which relied on
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indirect methods of assessing Pb exposure. Case-control studies that measured blood Pb
levels produced contrasting results. A study of 16 ALS cases (mean blood Pb level:
12.7 (ig/dL) and 39 controls (mean blood Pb level: 10.8 (ig/dL) found a small difference
in the mean concurrent blood Pb level (Vinceti etal.. 1997). A larger study of 109 cases
and 256 controls that examined concurrent blood and bone Pb levels in a New England-
area population found higher odds of ALS among subjects with concurrent blood Pb
levels 3-14 (ig/dL (e.g., OR: 14.3 [95% CI: 3.0, 69.3] for n = 55 with blood Pb levels
3-4 (ig/dL compared with blood Pb levels 10 (ig/g patella Pb and > 8 (ig/g tibia Pb), but lacked precision. For example, compared
with subjects with tibia Pb level <7 (ig/g, the OR for tibia Pb levels 8-14 (ig/g was 1.6
(95% CI: 0.5, 5.6). Results were adjusted for age, education, and hours/day inactive.
Potential confounding by smoking was not considered. Also in this population, an
estimate of cumulative Pb exposure based on occupational history was found to be
associated with ALS (Kamel et al.. 2002). The stronger findings for blood Pb level were
surprising given that bone Pb level is a better biomarker of cumulative Pb exposure. One
explanation for these findings is reverse causation. Blood was collected from people who
already had ALS, and reduced physical activity among those with ALS could lead to
more bone turnover and greater release of Pb from bones into circulation in ALS cases
than controls.
Since the 2006 Pb AQCD, a few additional studies of ALS have been conducted with the
same New England-area case-control group. Kamel et al. (2005) reported that the
association between blood Pb level and ALS was not modified by the ALAD genotype,
and Kamel et al. (2008) found that higher tibia and patella Pb levels were associated with
longer survival time among 100 of the original 110 ALS cases with adjustment for age,
sex, and ever smoking. Results were not altered by the additional adjustment for
education, BMI, family history of ALS, or concurrent physical activity. Higher blood Pb
levels were not associated with survival time. These paradoxical findings that point to a
protective effect of Pb are not easily explainable but find coherence with results for
Pb-induced increased survival time in an ALS mouse model (see below). On one hand,
the cases with longer survival time may have higher bone Pb levels because they reflect a
longer period of cumulative exposure. On the other hand, with longer survival time, there
could be greater progression of disease and less mobility. Decreased mobility would tend
to increase bone resorption, lower bone Pb levels, and increase blood Pb levels overtime.
The latter hypothesis is a less likely explanation for findings in this New England cohort
because higher bone Pb levels were more strongly associated with longer survival time
than was blood Pb level.
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Another case-control study examined concurrent blood Pb levels and ALS among 184
cases (33 were either progressive muscular atrophy or primary lateral sclerosis, mean
blood Pb level: 2.41 (ig/dL) and 194 controls (mean blood Pb level: 1.76 (ig/dL) (Fang et
al.. 2010). The cases were recruited from the National Registry of U.S. Veterans with
ALS, and controls were recruited from among U.S. Veterans without ALS and frequency
matched by age, gender, race, and past use of the Veterans Administration system for
health care. A doubling of concurrent blood Pb level was associated with ALS with an
OR of 1.9 (95% CI: 1.3, 2.7) with adjustment for age and a collagen protein as an
indicator of bone formation. The association did not differ substantially by of the degree
of bone turnover but was slightly higher among ALAD 1-1 carriers. The association with
blood Pb level was similar in analyses that excluded the progressive muscular atrophy
and primary lateral sclerosis cases. The similar results by degree of bone turnover suggest
that reverse causation is not likely explaining the association between blood Pb level and
ALS. However, as in other ALS case-control studies, the directionality of effects is
difficult to establish. This study did not have measures of bone Pb to assess the
association with biomarkers of longer-term Pb exposure.
Although epidemiologic studies have provided inconsistent evidence for associations of
Pb biomarker levels with ALS in adults, toxicological studies have found that Pb
exposure affects neurophysiologic changes associated with ALS. For example, chronic
postnatal Pb exposure from weaning onward (200 ppm Pb acetate in drinking water,
resultant blood Pb level: 27 (ig/dL) reduced astrocyte reactivity and induced increased
survival time in the superoxide dismutase transgenic (SOD1 Tg) mouse, which has SOD
mutations found in humans with familial ALS (Barbeito et al.. 2010). In this model, Pb
exposure did not significantly increase the onset of the ALS disease. These findings
provide coherence with the association observed between bone Pb level and longer
survival time in patients diagnosed with ALS (Kamel et al.. 2008).
Research outside of the Pb field has suggested different mechanisms for ALS initiation
versus ALS progression, i.e., motor neuron function versus astrocyte and microglia
function (Yamanaka et al.. 2008; Boillee et al.. 2006). Astrocyte vascular endothelial
growth factor (VEGF) was examined for its involvement in the effects of Pb on
increasing survival time in the ALS mouse model. Lower VEGF expression has been
linked with risk of ALS in humans and ALS-like symptoms in animals. Baseline VEGF
levels were elevated in astrocytes from the ventral spinal cord of untreated SOD1
Tg mice versus untreated nontransgenic mice. VEGF was not induced in the astrocytes of
Pb-treated nontransgenic mice. In comparison, Pb-exposed SOD1 Tg mice, which had
longer survival time, also had significant elevations in astrocyte VEGF (Barbeito et al.,
2010). These findings for Pb-induced effects on astrocytes in a mouse model for ALS
may provide a mechanistic explanation for Pb effects on increasing survival time in ALS.
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Others reported that VEGF administration to the SOD1 Tg mice significantly reduced
glial reactivity, a marker or neuroinflammation (Zheng et al.. 2007). Using a cell-based
co-culture system of neurons and astrocytes isolated from Pb-exposed SOD1 Tg mice,
Barbeito et al. (2010) found that an up-regulation of VEGF production by astrocytes was
protective against motor neuron death in the SOD1 Tg mouse cells. Thus, in vivo and in
vitro results indicate that chronic Pb exposure resulted in increased survival time in an
ALS mouse model and was correlated with higher spinal cord VEGF levels, which made
astrocytes less cytotoxic to surrounding motor neurons (Barbeito et al.. 2010).
In summary, there is inconsistent evidence of association between indicators of Pb
exposure (history of occupational exposure, Pb biomarker levels) and ALS prevalence
and survival time in humans. Other limitations include the potential for reverse causation
and bias due to survival time in the case-control studies and the lack of objective
assessment of occupational exposure. While several studies have considered potential
confounding by age, education, and physical activity, few have considered smoking.
Toxicological evidence also points to Pb exposure increasing survival in a mouse model
of ALS and has suggested explanations including Pb-induced increases in VEGF
expression and subsequent reduction in glial activity and protection of motor neurons
against inflammation.
4.3.9.3 Parkinson's Disease
Previous Pb AQCDs reviewed a few studies, some ecological (Rybicki etal.. 1993;
Aquilonius and Hartvig. 1986) and some case-control relying on questionnaire data or
occupational history (Gulson et al.. 1999; GorelletaL 1997; Tanner etal.. 1989) that
indicated associations between exposure to heavy metals, particularly Pb, and risk of
Parkinson's disease. The limited number of previous studies, weak study designs, and
weak Pb exposure assessment did not permit firm conclusions. Recent studies maintain
several of these limitations but have indicated associations with bone Pb levels.
A recent large case-control study (330 cases, 308 controls) examined a population in the
Boston, MA area with virtually no occupational exposures to Pb (Weisskopf et al.. 2010).
Subjects in the highest quartile of tibia Pb level (>16.0 (ig/g) had higher odds of
Parkinson's disease compared to those in the lowest quartile (< 3.1 (ig/g) (OR: 1.91 [95%
CI: 1.01, 3.60]) with adjustment for age, race, pack-years smoking, education, and
recruitment site. Cases and controls were recruited from several different sources
including movement disorder clinics and the NAS, which could have introduced biased
participation by Pb exposure or reduced representativeness to the target population. In the
NAS, cases were ascertained from self-report, which may introduce measurement error.
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However, when analyses were restricted to cases recruited from movement disorder
clinics and their spouse, in-law, or friend as controls, the results were even stronger (OR:
2.57 [95% CI: 1.11, 5.93] for tibia Pb >13.9 (ig/g). Although the use of spouse, in-law,
and friend controls can introduce bias, this is expected to be toward the null as these
groups are likely to share many exposures. Weisskopf et al. (2010) did not adjust results
for Mn exposure, which has been associated with Parkinsonian symptoms and could
potentially confound associations between Pb and Parkinson's disease. However, unlike
occupational exposure to Pb, environmental Pb exposure is less likely to be correlated
with environmental Mn exposure. Thus, it is less likely that the associations observed
with Pb were confounded by co-occurring Mn exposure.
Coon et al. (2006) conducted a smaller case-control study of 121 Parkinson's disease
patients and 414 controls frequency-matched by age, sex, and race, all receiving health
care services from the Henry Ford Health System in Michigan. Subjects in the highest
quartile of both tibia (OR: 1.62 [95% CI: 0.83, 3.17] for levels > 15 (ig/g) and calcaneus
(OR: 1.50 [95% CI: 0.75, 3.00] for levels > 25 (ig/g) bone Pb levels had higher but
imprecise odds of Parkinson's disease compared to those in the lowest quartiles (0-6 (ig/g
for tibia and 0-12 (ig/g for calcaneus). Subjects in the highest quartile of whole-body
lifetime Pb level (> 81 (ig/g, estimated using PBPK modeling) had the highest OR: 2.27
(95% CI: 1.13, 4.55) versus the lowest quartile, 0-40 (ig/g. These results were adjusted
for age, race, sex, pack-years smoking, regular coffee consumption, and regular alcohol
use, but Mn exposure was not considered. It was not clear what the extent of occupational
exposure to Pb was among the participants; however, a previous Henry Ford Health
System study had linked occupational Pb exposure to Parkinson's disease (Gorell et al..
1997). Thus, it is uncertain whether the observed associations were confounded by
co-occurring Mn exposure.
In summary, a few recent case-control studies expand on previous evidence by finding
associations of tibia and calcaneus bone Pb levels, biomarkers of cumulative Pb
exposure, with Parkinson's disease in adults. The associations observed with biomarkers
of cumulative Pb exposure increase confidence that associations are not explained by
reverse causation. However, firm conclusions are not warranted. While associations were
adjusted for potential confounding by age, sex, race, and education, Mn co-exposure was
not considered.
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4.3.9.4 Essential Tremor
The few available case-control studies of essential tremor have found associations with
concurrent blood Pb levels. The 2006 Pb AQCD (U.S. EPA. 2006b) described case-
control studies that found associations between concurrent blood Pb levels and essential
tremor in New York City metropolitan area populations (Louis et al., 2005; Louis et al.,
2003). In the larger study, mean (SD) blood Pb levels were 3.3 (2.4) (ig/dL in the 100
essential tremor cases and 2.6 (1.6) (ig/dL in the 143 controls (Louis et al., 2003). In the
other study, mean (SD) blood Pb levels were 3.5 (2.2) (ig/dL in the 63 essential tremor
cases and 2.6 (1.5) (ig/dL in the 101 controls (Louis et al., 2005). In Louis et al. (2005).
the magnitude of association was larger among the 35 ALAD-2 carriers than among 129
adults homozygous for ALAD-1. History of occupational Pb exposure was similarly rare
in cases and controls (2%).
Recently, Dogu et al. (2007) reported on 105 essential tremor cases selected from a
movement disorder clinic in Turkey and 105 controls (69 spouses and 36 other relatives)
living in the same district. With adjustment for age, sex, education, smoking status,
cigarette pack-years, and alcohol use, a 1 ug/dL higher blood Pb level (measured at the
time of study recruitment) was associated with essential tremor with an OR of 4.19 (95%
CI: 2.59, 6.78). This OR was much larger than that obtained in the New York area study
(OR: 1.19 [95% CI: 1.03, 1.37]) (Louis et al.. 2003). The magnitude of association in
Dogu et al. (2007) is even more striking because so many of the controls were spouses
who are expected to share many environmental exposures as cases. Most of the essential
tremor cases were retired at the time of the study; however past occupational history was
not examined.
In summary, a small body of studies indicates associations between blood Pb level
measured at the time of the study and prevalence of essential tremor in adults. However,
because of the case-control design, reverse causation cannot be excluded as a potential
explanation for the observed associations since loss of physical activity and subsequent
bone resorption may lead to an increase in blood Pb level. Further, the level, timing,
frequency, and duration of Pb exposure associated with essential tremor are uncertain.
History of occupational Pb exposure was not consistently examined, and potential
confounding by Mn exposure was not examined.
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4.3.9.5 Toxicological Studies of Cell Death Pathways
A common element of the neurodegenerative diseases described above is neuronal cell
death. Studies reviewed in the 2006 Pb AQCD (U.S. EPA. 2006b) documented that Pb
exposure induced cell death or apoptosis in various models including rat brain (Tavakoli-
Nezhadetal.. 2001). retinal rod cells (He etal.. 2003; He et al. 2000). and cerebellar
neurons (Oberto etal.. 1996). Recent studies produced similar findings, in most cases, in
animals with higher blood Pb levels than those relevant to this ISA. Long-term (40 days)
exposure to 500 ppm Pb in drinking water was found to increase pro-apoptotic Bax
protein levels and the number of apoptotic cells in the hippocampus in young (exposure
starting at 2-4 weeks of age) and adult (exposure starting at 12-14 weeks of age) male rats
with blood Pb levels 91 and 98 (ig/dL (Sharifi et al.. 2010). Apoptosis was verified by
light and electron microscopy. These investigators found Pb-induced apoptosis of PC 12
neuronal cells in vitro (Sharifi and Mousavi. 2008). Another study followed the
developmental profile of changes in various apoptotic factors in specific brain regions of
rats exposed to 2,000 ppm Pb acetate during lactation (to PND20) via drinking water of
dams (Chao et al.. 2007). At PND20, male offspring blood Pb level was 80 (ig/dL. Pb
exposure significantly upregulated hippocampal mRNA for various apoptotic factors
including caspase-3, Bcl-x, and Brain-derived neurotrophic factor (BDNF) on PND12,
PND15 and PND20. The cortex of these male pups showed upregulation of Bcl-x and
BDNF on PND20 and PND25. The cerebellum did not have elevated apoptotic mRNA
levels in this model. Thus, in this study, Pb-induced apoptosis varied by age and brain
region in male offspring.
Pb exposure also has been shown to induce apoptosis of spinal cord cells during spinal
cord development in chicks treated with 150 or 450 (ig Pb acetate in ovo at
embryonic day 3 or 5 and visualized six days later (Miiller et al.. 2012). TUNEL positive
cells, indicating DNA fragmentation induced by apoptosis, were at significantly higher
levels in Pb-exposed animals and were visualized in all layers of the developing spinal
cords. Also, levels of glial fibrillary acidic protein (GFAP), a factor important in neuronal
migration and cellular differentiation during nervous system development, was
significantly attenuated in spinal cords of Pb-exposed chicks. Liu et al. (201 Ob) examined
apoptotic effects in 30 day-old male rats that were treated with Pb acetate once daily for
6 weeks via intragastric infusion. Four groups: control, low (2 mg/kg BW), medium (20
mg/kg BW), and high (200 mg/kg BW) had blood Pb levels of 1-7.5 ug/dL;
4.5-11.4 ug/dL; 8.9-42 ug/dL; and 48-73 ug/dL, respectively. Pb induced hippocampal
neuronal apoptosis (TUNEL positive staining, statistically significant at all Pb doses) and
downregulation of hippocampal XIAP (significant at high dose only) and Smac
(statistically nonsignificant trend) at the termination of the 6 week treatment. In another
study, Pb exposure (500 ppm Pb acetate in drinking water for 8 weeks) of adult male rats
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induced regional-specific changes in brain apoptotic proteins poly(ADP-ribose)
polymerase, Bcl-2, and caspase-3 with a greater effect observed in the hippocampus and
cerebellum than the brainstem or frontal cortex (Kiran Kumar et al., 2009).
In summary, a small body of epidemiologic studies found Pb-associated increases in
essential tremor and Parkinson's Disease in adults. However, the implications of these
findings are limited because of the potential for reverse causation to explain cross-
sectional associations observed with blood Pb level, and the potential for confounding by
Mn exposure. However, toxicological evidence supports an effect of Pb on
neurodegeneration by demonstrating that Pb exposure during various lifestages, early
postnatal or adulthood, induces neuronal apoptosis in animals. Several of these
observations were made with routes (i.e., i.p.) or concentrations of Pb exposure that may
not be relevant to this ISA.
4.3.10 Modes of Action for Pb Nervous System Effects
4.3.10.1 Effects on Brain Physiology and Activity
The 2006 Pb AQCD (U.S. EPA. 2006b) reviewed a small body of available
epidemiologic studies demonstrating associations of Pb biomarkers with
electrophysiologic and physical changes in the brains of young adults (Yuan et al., 2006;
Cecil et al.. 2005) and children (TVIeng etal.. 2005; Trope etal.. 2001) as assessed by
magnetic resonance imaging (MRI) or spectroscopy (MRS). The implications of previous
findings were limited by the small sample sizes (n = 12-45) and limited consideration for
potential confounding. Recent studies examining MRI data were limited largely to the
Cincinnati cohort as adults (ages 19-24 years). In addition to supporting associations of
childhood blood Pb levels with physiological changes in the brain of adults, these recent
analyses expanded on previous studies by including larger sample sizes, aiming to
characterize potentially important lifestages of Pb exposures, and evaluating potential
links between changes in brain activity and functional neurodevelopmental effects. While
overall there are few studies in few populations, by showing physical and physiologic
changes in areas of the brain associated with neurodevelopmental function, the evidence
provides biological plausibility for the associations observed between Pb biomarker
levels and cognitive function and behavior.
In prospective analyses of the Cincinnati cohort as adults (ages 20-23 years, n = 35, 42),
Cecil et al. (2005) and Yuan et al. (2006) conducted functional MRI during a verb
generation language task and found that higher age 3-78 month average blood Pb level
(mean 14.2 (ig/dL) was associated with decreased activation in the left frontal gyrus and
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left middle temporal gyrus, regions implicated in semantic language function. Yuan et al.
(2006) considered birth weight, marijuana consumption, sex, SES, gestational age, and
IQ as potential confounding factors. Whereas previous analyses of the Cincinnati cohort
focused on activity in specific regions of the brain, Cecil et al. (2011) examined brain
metabolites. Higher age 3-78 month average blood Pb levels (mean: 13.3 (ig/dL) were
associated with lower levels of N-acetylaspartate (NAA) and creatine (Cr) in the basal
ganglia and lower levels of choline in white matter in 159 adults, ages 19-23 years. These
results were adjusted for age and FSIQ; however, several other unspecified factors were
considered. Lower levels of NAA, Cr, and choline are linked to decreased neuronal
density and alteration in myelin. A recent prospective analysis of 31 men in the NAS
cohort similarly reported an association between biomarkers of cumulative, long-term Pb
exposure and changes in brain metabolites in older adults. Weisskopf et al. (2007a) found
higher tibia and patella Pb levels to be associated with a higher myoinositol/Cr ratio in
the hippocampus measured more than 10 years after bone Pb was measured and adjusted
for age. Myoinositol/Cr ratio may be indicative of glial activation and is a signal
reportedly found in the early stages of HIV-related dementia and Alzheimer's disease.
Other studies in the Cincinnati cohort as young adults found that childhood average blood
Pb levels were associated with altered brain architecture. Among 91 adults ages 20-26
years, Brubaker et al. (2009) found associations of age 3-78 month average blood Pb
levels (mean: 12.9 (ig/dL) with diffusion parameters that were indicative of less
organization of fibers throughout white matter. Results were adjusted for maternal IQ,
prenatal alcohol and tobacco exposure, and adult marijuana use. In regions of the corona
radiata, higher blood Pb levels were associated with less myelination axonal integrity. In
regions of the corpus callosum, higher blood Pb levels were associated with greater
myelination and axonal integrity. The differential impact among neural elements may be
related to the stage of myelination development present at various time periods.
Another study of 157 Cincinnati cohort adults ages 19-24 years provided evidence of
region-specific reductions in gray matter volume in association with age 3-78 month
average blood Pb levels (mean: 13.3 (ig/dL) with adjustment for sex (Cecil et al.. 2008).
The most affected regions included frontal gray matter, specifically the anterior cingulate
cortex, and the ventrolateral prefrontal cortex (i.e., areas related to executive functions,
mood regulation, and decision-making). Further, fine motor factor scores were positively
correlated with gray matter volume in the cerebellar hemispheres; adding blood Pb level
as a variable to the model attenuated this correlation. These findings suggested that
changes observed with MRI may mediate the association between blood Pb levels and
decrements in motor function. The functional relevance of these structural changes in the
brain also is supported by observations from other studies that link changes in brain
architecture and activity to changes in cognitive function (e.g., visuoconstruction, visual
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memory, eye-hand coordination) (Schwartz et al., 2007) and behavioral problems
(impulsivity, aggression, violence) (Yang et al.. 2005; Raine et al.. 2000). In a subsequent
comparison of blood Pb levels measured at various lifestages in 157 Cincinnati cohort
adults ages 19-24 years, Brubaker et al. (2010) found that higher blood Pb levels at older
ages (annual means from ages 3-6 years, means: 9.6-16.3 (ig/dL) were associated with
greater losses in gray matter volume than were age 1 or 2 year (means: 10.6, 17.2 (ig/dL),
3-78 month average or maximum blood Pb levels (mean: 23.1 (ig/dL). Both Cecil et al.
(2008) and Brubaker et al. (2010) found that Pb-associated reductions in gray matter were
more pronounced in males than females in the Cincinnati cohort.
Studies of Pb-exposed workers (n = 15-532) also found associations of concurrent blood
(means: 17-63.5 (ig/dL) and tibia (mean 14.5 (ig/dL) Pb levels with changes in brain
structure and physiology, supporting the effects of chronic Pb exposure. Pb-associated
changes included white matter lesions, smaller brain volumes, less total gray matter, and
lower levels of brain metabolites such as NAA and Cr (Hsieh et al.. 2009b: Jiang et al..
2008; Bleecker et al.. 2007b; Stewart et al.. 2006) with adjustment for similar factors as
associations for cognitive function. Other occupational exposures were not examined. In
a few of these occupational groups, Pb-associated brain changes were linked to poorer
performance in cognitive function tests (Caffo et al.. 2008; Bleecker et al.. 2007b).
Higher concurrent blood Pb level also was associated with lower NAA/Cr ratio in small
cross-sectional studies that included children (n = 6, 16, ages 4-21 years), although
neither study considered potential confounding (Meng et al.. 2005; Trope etal. 2001).
All subjects had normal MRIs with no evidence of structural abnormalities. Thus, the
biological relevance of the observed physiological changes is unclear. Additionally, the
representativeness of findings is uncertain because results were based on comparisons of
subjects with relatively high blood levels (means: 38, 40 (ig/dL) to those with blood Pb
levels <10 (ig/dL.
In summary, results in a few populations indicate associations of childhood blood or adult
tibia Pb levels with changes in brain structure and physiology in adults assessed by MRI
or MRS. Associations were found in children, but implications are more limited because
of small samples sizes, lack of consideration of potential confounding, and high blood Pb
levels of the children examined. Evidence from the prospective Cincinnati cohort studies
improves characterization of the temporal sequence between Pb exposure and changes in
brain structure and physiology. Several studies linked these changes to functional
changes in cognitive performance or motor skills. Because of the small samples sizes of
several studies and limited consideration for potential confounding, implications of the
effects of Pb exposure specifically on changes in brain structure and physiology are not
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certain. However, the evidence provides biological plausibility for the associations
observed between Pb biomarker levels and cognitive function and behavioral problems.
4.3.10.2 Oxidative Stress
Because the brain has the highest energy demand and metabolism of any organ, energy
homeostasis is of utmost importance. Energy imbalance can increase the susceptibility of
the highly energetic brain tissue to stressors and cell death. Pb has been shown to induce
energy imbalance by inhibiting various enzymes involved in energy production or
glucose metabolism including glyceraldehyde-3-phosphate dehydrogenase, hexokinase,
pyruvate kinase, and succinate dehydrogenase (Verma et al., 2005; Yun and Hover, 2000;
Regunathan and Sundaresan. 1984; Sterling et al.. 1982). Mitochondria produce ATP or
energy through oxidative phosphorylation. Aberrant mitochondrial function can decrease
the energy pool and contribute to ROS formation via electron transport chain disruption.
ATP depletion can also affect synaptic and extracellular neurotransmission. The
mitochondrial Na+/K+ATPase is important in maintaining the inner mitochondrial
membrane potential A*Pm (delta psifsub m]) and the functioning of the mitochondria.
Gestational Pb exposure of mice was found to impair mitochondrial function and energy
production in neuronal cells and produce concomitant increases in mitochondrial and
cellular ROS production. In these mice, 1,000 ppm Pb acetate in dam drinking water
resulted in offspring blood Pb levels of 4 (ig/dL and cerebella Pb levels of 7.2 (ig/g dry
weight) (Baranowska-Bosiacka et al.. 201 Ib). Cerebellar granular cells were harvested
from control and Pb-exposed animals at PND8. These neuronal cells were cultured for
5 days in vitro, at which point various mitochondrial parameters were measured. With Pb
exposure, ROS were significantly increased in both the cortical granule cells and in the
mitochondria. Intracellular ATP concentration and adenylate energy charge values were
significantly decreased in cells of Pb-exposed mice versus controls. Neuronal
Na+/K+ATPase activity was significantly lower in cortical granule cells from Pb-exposed
mice versus cells from controls. Mitochondrial mass was unaffected with Pb treatment,
but mitochondrial membrane potential was significantly decreased with Pb exposure.
Energy imbalances also were found in Wistar rats (PND15) of each sex injected daily for
2 weeks with Pb acetate (15mg/kg BW, i.p., resulting in a mean blood Pb level of
31 (ig/dL; control blood Pb level 3 (ig/dL) (Baranowska-Bosiacka et al.. 201 la). ATP and
ADP were significantly decreased in various brain regions with Pb exposure, with the
cerebellum and hippocampus more strongly affected than the forebrain cortex. Also,
expression of the pro-inflammatory P2XR receptor was enhanced in the glial fraction,
indicating the astrocyte pool may be involved in the pathological changes found in
Pb-exposed immature rat brains. Mitochondrial energy imbalances also were found in
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Pb-exposed crayfish that were placed under hypoxic conditions which induced a decrease
in metabolism (Morris et al.. 2005).
In rats, Pb exposure has been shown to induce oxidative stress, in some cases, with
concomitant functional CNS changes. Exposure of adult male rats to 4,000 ppm
Pb acetate in drinking water for 6 weeks increased brain levels of lipid peroxides (LPO)
and lowered levels of antioxidants including nitric oxide (NO), total antioxidant capacity
(TAC), glutathione (GSH), glutathione-S-transferase (GST), and superoxide dismutase
(SOD). In red blood cells, Pb levels were positively correlated with LPO levels and
negatively correlated with NO levels (Hamed etal. 2010). These effects of Pb on
oxidative stress parameters were attenuated with co-exposure to green tea extract (1.5%),
which reduced brain (1.9 to 1.2 ppm) and blood Pb levels of rats (0.773 to 0.654 ppm). In
a study of adult male Wistar albino rats, Pb acetate treatment by i.p. (20 mg/kg daily,
5 days) elevated lipid peroxidation, neuronal damage, and brain tissue DNA
fragmentation and decreased antioxidant GSH levels and antioxidant enzyme activity,
(Abdel Moneim et al.. 2011 a). These effects were attenuated with co-administration of
the polyunsaturated fatty acid flaxseed oil (oral gavage l,000mg/kg body weight for
5 days, 1 hour prior to Pb dosing). Flaxseed oil co-treatment also was associated with
significantly lower blood Pb level in Pb-exposed animals and control animals, indicating
that flaxseed oil may alter Pb toxicokinetics in animals. Another study provided indirect
evidence of Pb-induced oxidative stress with observations that Pb-induced (2,000 ppm
Pb acetate in drinking water PND1-PND67) impairments in long-term potentiation
(LTP), paired-pulse reactions, and input/output functions in the DG of male and female
Wistar rats were significantly attenuated with co-treatment with the antioxidant quercetin
(i.p. 30 mg/kg BW, PND60-PND67) at PND67 (Hu et al.. 2008a). Quercetin-treated
animals had significantly less hippocampal Pb than did the animals exposed only to Pb.
Oxidative stress may be involved in neurodegenerative pathologies including Alzheimer's
disease. Hydrogen peroxide-induced oxidative stress has been shown to induce
intracellular accumulation of Ab in human neuroblastoma cells (Misonou et al.. 2000).
Oxidative stress-induced DNA damage can be measured as the ratio of the adduct
8-hydroxy-2'-deoxyguanosine to 2-deoxyguanosine (8-oxo-dG/2-dG). 2-dG is a product
of oxidative cleavage and is oxidized to form 8-oxo-dG. Pb-induced changes in the
8-oxo-dG to 2-dG ratio were examined recently as a mechanism underlying
neurodegeneration. Similar to Ab levels, changes in the 8-oxo-dG to 2-dG ratio showed a
biphasic relationship in the brains of rats exposed to 2,000 ppm Pb acetate via drinking
water of dams from PND1-PND20 (blood Pb level 46 (ig/dL) (Bolin et al.. 2006). The
8-oxo-dG to 2-dG ratio decreased early in exposure (PND5) but increased at age
20 months. No increase was found in animals exposed to Pb from age 18 to 20 months
(blood Pb level: 60 (ig/dL). Activity of the base-excision DNA repair enzyme 8-oxo-
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guanine DNA glycosylase was unaffected by Pb exposure. Similar findings were reported
in a monkey study (Wu et al.. 2008a). The ratio of 8-oxo-dG to 2-dG in the brains of aged
monkeys (23 years) was significantly elevated above that in controls with Pb exposure
only in infancy (PND1-PND400, infant formula, blood Pb levels: 19-26 (ig/dL at
PND400) (Wu et al., 2008a). Thus, evidence in rats and monkeys suggests a possible role
for oxidative stress in Pb-induced neurodegenerative effects and indicates that early life
Pb exposure induces oxidative DNA damage and amyloidogenesis in aged animals.
4.3.10.3 Nitrosative Signaling and Nitrosative Stress
The NO system is increasingly being recognized as a signaling system in addition to its
more classical role as a marker of cellular stress. Pb exposures during the gestational-
early postnatal period (GD6-PND21) (Chetty et al.. 2001) and after lactation only (Fan et
al.. 2009a) were found to reduce hippocampal levels of NO or neuronal NO synthase. In
the hippocampus, NO mediates LTP, which is considered to be a major cellular
mechanism underlying learning and memory. Thus, observations of Pb-induced changes
in hippocampal NO may provide a mechanistic explanation for the effects of Pb on
cognitive function decrements. Fan et al. (2009a) found reduced hippocampal NOS and
NO in weanling male rats after either 4 or 8 weeks of Pb exposure resulting in blood Pb
levels of 48 and 70 (ig/dL, respectively. In the same study, dietary supplementation with
taurine or glycine concomitant with 8 weeks of Pb exposure induced significant increases
in hippocampal NOS, whereas Pb plus dietary supplementation with vitamin C,
methionine, tyrosine, or vitamin Bl decreased hippocampal NOS. In this study,
co-exposure of specific nutrients also prevented Pb-induced impairments in learning as
evidenced by lack of increased escape latency in the Morris water maze. Dietary
supplementation with tyrosine, methionine, or ascorbic acid after 4 weeks of Pb exposure
(blood Pb levels upon cessation of exposure and after 4-week lag: 48 and 8.1 (ig/dL,
respectively) reversed Pb-induced decrements in NO/NOS. Zn supplementation given
after Pb exposure had no effect on the NO system.
4.3.10.4 Synaptic Changes
Previous toxicological studies pointed to an effect of developmental Pb exposure on
synapse development, which mechanistically may contribute to multiple Pb-related
aberrant effects, including changes in LTP and facilitation. Facilitation of a neuronal
terminal is defined as the increased capability to transmit an impulse down a nerve due to
prior excitation of the nerve. Earlier work showed that Pb exposure resulted in altered
density of dendritic hippocampal spines (Kiraly and Jones. 1982; Petit and Leboutillier.
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1979). reduced synapse formation (Chen et al., 2005b). and abnormal long-term plasticity
(Nihei et al.. 2000). In a recent study, Li et al. (2009c) focused on inflammatory
endpoints and synaptic changes after gestational plus lactational Pb exposure
(1,000-10,000 ppm Pb acetate via drinking water of dams, producing offspring blood Pb
levels 40-100 (ig/dL, respectively at PND21). Hippocampal TNF-a was significantly
elevated with Pb exposure, and proteins that comprise the SNARE complex were all
changed with Pb exposure. The SNARE complex of synaptic proteins includes SNAP-25,
VAMP-2 and Syntaxin la and is essential in exocytotic neurotransmitter release at the
synapse. Thus, Li et al. (2009c) found significant differences in hippocampal synaptic
protein composition and increased pro-inflammatory cytokine levels in the brains of
Pb-exposed offspring.
Recent research using the Drosophila larval neuromuscular junction model showed that
compared with unexposed controls, Pb-exposed larvae had significant increases in
intracellular calcium and significant delays in calcium decays back to baseline levels at
the pre-synaptic neuronal bouton (as stimulated with multiple action potentials, also
called AP trains). Pb-exposed larvae had reduced activity of the plasma membrane
Ca2+ATPase, which is responsible for extravasations of calcium from the synaptic
terminal (He et al.. 2009). Intracellular calcium in Pb-exposed larvae was no different
from that in controls under resting conditions or in neurons with stimulation by a single
action potential. Pb media concentrations in these experiments were 100 or 250 (iM with
the body burden of Pb from the lower dose estimated to be 13-48 (iM per larvae. After
stimulation of the axon, facilitation of the excitatory post-synaptic potential, which is
dependent on residual terminal calcium, was significantly elevated in Pb-exposed larvae
versus control (He et al.. 2009). The data from this synapse study demonstrate that
developmental Pb exposure affected the plasma membrane Ca2+ATPase, induced changes
in the intracellular calcium levels during impulse activation, and produced changes in
facilitation of the neuronal networks of Drosophila. Thus, the neuromuscular junction is a
potential site of Pb interaction.
Neurotransmission is an energy-dependent process as indicated by the presence of
calcium-dependent ATP releases at the synaptic cleft. At the synapse, ATP is
metabolized by ectonucleotidases. Acute exposure (96 hours) of male and female
zebrafish to 2 (ig/dL Pb acetate in their water induced significant decreases in ATP
hydrolysis in brain tissue (Senger et al.. 2006). This dose is deemed to be
environmentally relevant. With chronic exposure (30 days), Pb acetate promoted the
inhibition of ATP, ADP and AMP hydrolysis. These findings were consistent with
findings in rodents (Baranowska-Bosiacka et al.. 201 la). The authors hypothesized that at
30 days, this Pb-induced change in nucleotide hydrolysis was likely due to post-
translational modification because expression of enzymes responsible for the hydrolysis,
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NTPDasel and 5'-nucleotidase, were unchanged (Senger et al., 2006). Thus, Pb was
shown to affect nucleotidase activity in the CNS of zebrafish, possibly contributing to
aberrant neurotransmission.
Another enzyme important in synaptic transmission at cholinergic junctions in the CNS
and at neuromuscular junctions is acetylcholinesterase (AChE). After 24 hours of
exposure to 2 (ig/dL Pb acetate in water, AChE activity was significantly inhibited in
zebrafish brain tissue (Richetti et al.. 2010). AChE activity returned to baseline by 96
hours and maintained baseline activity after 30 days of exposure. Thus, Pb was shown to
affect synaptic homeostasis of AChE in the brains of zebrafish only transiently.
Pb has been shown to act as an antagonist of the NMDA receptor (NMDAR). The
NMDAR is essential for proper presynaptic neuronal activity and function. Primary
cultures of mouse hippocampal cells exposed to 10 or 100 (iM Pb during the period of
synaptogenesis had loss of two proteins necessary for presynaptic vesicular release,
synaptophysin (Syn) and synaptobrevin (Syb) but no change in a similar protein
synaptotagmin (Syt) (Neal et al.. 2010a). This deficit was found in both GABAergic and
glutamatergic neurons. Pb also induced an increase in the number of presynaptic contact
sites. But, these sites may have been nonfunctional as they lacked the protein receptor
complexes necessary for proper vesicular exocytosis. Another factor involved in
maturation and signaling of presynaptic neurons is BDNF, which is synthesized and
released by postsynaptic neurons and regulated by the NMDAR. In mouse hippocampal
cells, both pro-BDNF and BDNF release were significantly attenuated with Pb exposure
(Neal et al.. 2010a). Further, exogenous BDNF administration rescued the
aforementioned Pb-related effects on presynaptic proteins. Thus, results from this cell
culture model showed that Pb-related presynaptic aberrations are controlled by NMDAR-
dependent BDNF effects on synaptic transmission.
Glutamate is another neurotransmitter that is released from presynaptic neurons and via
interactions with the NMDAR causes postsynaptic neuron depolarization. A recent study
of Wistar albino rats exposed to Pb postnatally from birth to age 12 weeks (drinking
water 3 x 104 (ig/dL Pb acetate, resulting in blood Pb levels of 17 (ig/dL at age 6 weeks)
showed decreased learning ability, decreased hippocampal glutamate at 6, 8, 10 and
12 weeks of age, as well as significant decrements in the hippocampal glutamate
synthesis-related enzymes aspartate aminotransferase and alanine aminotransferase (Niu
et al.. 2009).
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4.3.10.5 Blood Brain Barrier
Two barrier systems exist in the body to separate the brain or the CNS from the blood:
the blood brain barrier (BBB) and the blood cerebrospinal fluid barrier (BCB). The BBB,
formed by tight junctions at endothelial capillaries forming the zonulae occludens
(occludins, claudins, and cytoplasmic proteins), separates the brain from the blood and its
oncotic and osmotic forces, allowing for selective transport of materials across the BBB.
Pb exposure during various developmental windows has been shown to increase the
permeability of the BBB of animals (Dvatlov et al.. 1998; Struzvnska et al.. 1997b;
Moorhouse et al.. 1988; Sundstrom et al.. 1985). Possibly due the underdevelopment of
the BBB early in life, prenatal and perinatal Pb exposure has been found to result in
higher brain Pb accumulation than have similar exposures later in life (Moorhouse et al..
1988). The choroid plexus and cerebral endothelial cells that form the BBB and BCB
tight junctions have been shown to accumulate Pb more than other cell types and regions
of the CNS. Studies reviewed in the 2006 Pb AQCD showed that the chemical form of Pb
and its capability to interact with proteins and other blood components affect its
capability to penetrate the BBB (U.S. EPA. 2006b). Pb also has been shown to
compromise the BCB and decrease the cerebrospinal fluid level of transthyretin, which
binds thyroid hormone in the cerebrospinal fluid. Low thyroid hormone levels in
pregnant women have been linked with IQ deficits in their children (Lazarus. 2005).
Recent research with male weanling rats exposed to Pb acetate via drinking water showed
leaky cerebral vasculature, an indication of a compromised BBB, as detected
histologically with lanthanum nitrate staining of the brain parenchyma (Wang et al..
2007b). Cerebral vasculature leakiness was ameliorated or resembled that of controls
after Fe supplementation. The cerebral vasculature leakiness may by explained by
observations of significant Pb-induced decreases in the BBB tight junction protein
occludin in the hippocampus, brain cortex, and cerebellum of these weanling animals.
Occludin levels were rescued to control levels with Fe supplementation. This loss of
integrity at the junctional protein level was affirmed with additional experiments using
the rat brain vascular endothelial cell line RBE4, in which 10 (iM Pb acetate exposure for
2, 4, 8, 16 and 24 hours resulted in decreases in junctional proteins occludin and claudin
5 as well as scaffold proteins ZO1 and ZO2 (Balbuena et al.. 2011). Because gene
expression for these junctional and scaffold proteins did not show decrements, it was
determined that these protein decrements were due to post-translational modifications.
Pb exposure also was found to contribute to leakiness of the BBB by decreasing the
resistance across the junction (Balbuena et al.. 2010). An in vitro co-culture system
employing endothelial cells (RBE4 or bovine brain microvascular endothelial cells) and
astrocytes (primary Sprague-Dawley neonatal pup astrocytes, GD21) served as the barrier
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between Pb-containing media and neurons. Pb acetate exposure (1 and 10 (iM) for 14
hours significantly impaired transendothelial electrical resistance, a marker of BBB
integrity, in a concentration-dependent manner.
Long-term Pb exposure of adult mice was found to increase regional edema and BBB
permeability (Lopez-Larrubia and Cauli. 2011). Adult male rats exposed to Pb acetate in
drinking water for 4 or 12 weeks (50 or 500 ppm, resulting in blood Pb levels of 12 and
55 (ig/dL, respectively) were assessed by diffusion weighted imaging for changes in
apparent diffusion coefficent (ADC), a measure of tissue water diffusivity that changes
under pathological conditions like cerebral edema. With 12 weeks of exposure, 50 ppm
Pb increased the ADC values in the cerebellum and mesencephalic reticular formation,
and 500 ppm Pb exposure significantly increased ADC in the corpus callosum and
caudate putamen. With 4 weeks of exposure, 500 ppm Pb significantly increased the
water ADC in the hippocampus, mesencephalic reticular formation, and cerebellum but
not in other brain areas. The brain areas with elevated ADC also showed increased BBB
permeability as measured with evans blue dye.
4.3.10.6 Cell Adhesion Molecules
Classic cell adhesion molecules including neural cell adhesion molecule (NCAM) and the
cadherins are junctional or cell surface proteins that are critical for cell recognition and
adhesion. While direct effects of Pb on cell adhesion molecules have not been described,
the calcium-dependency of these molecules suggests possible interactions with
competing cations like Pb and a possible contribution to Pb effects on CNS barrier
function disruption, neurite outgrowth, synaptic plasticity, learning, and memory
(Prozialeck et al.. 2002).
4.3.10.7 Effects on Glial Cells
Astroglia and oligodendroglia are supporting cells in the nervous system that maintain the
extracellular space in the brain and provide structural support to neurons, deliver
nutrients to neurons, and promote myelination. Glial cells provide immune surveillance
in the brain and contribute to inflammation-mediated pathologies. In Wistar rats, Pb
treatment (15 mg/kg of Pb acetate, i.p.) during early postnatal maturation was observed to
produce chronic glial activation with inflammation and neurodegeneration (Struzynska et
al., 2007). Among the cytokines detected in the brains of these Pb-treated rats were
IL-1(3, TNF-a and IL-6. Glial cells have been shown to serve as Pb sinks in the
developing and mature brain by sequestering Pb (Tiffany-Castiglioni etal.. 1989). This
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glial sequestration of Pb has been accompanied by a decrease in brain glutamine and a
reduction in glutamine synthetase activity in the astroglia; astroglia take up released
glutamate and convert it to glutamine. Pb has been shown to induce hypomyelination and
demyelination (CoriaetaL 1984) mediated through the oligodendrocytes with younger
animals found to be more susceptible to the effects of Pb (Tiffany-Castiglioni et al.,
1989). Pb accumulation in young glial cells may contribute to a lifelong exposure of
neurons to Pb as Pb is released from the sink over time. Thus, Pb accumulation in glial
cells can contribute to continual damage of surrounding neurons (Holtzman et al.. 1987).
Glial transmitters
Evidence indicates that glial transmission is affected with Pb exposure and that the
NMDAR may be involved in this aberrant glial transmission. To determine the
contribution of the gliotransmitter serine to Pb-mediated changes in LTP, Sun et al.
(2007) exposed rats to Pb acetate from gestation through lactation to PND28 via maternal
drinking water and collected hippocampal sections. CA1 section LTPs were examined
using in vitro patch clamp monitoring. Chronic Pb exposure impaired the magnitude of
hippocampal NMDAR-dependent LTPs, but the magnitude of the LTPs was restored with
supplementation with D-serine (Sun et al., 2007). which is an NMDAR agonist (Bear and
Malenka. 1994). The use of 7-chlorokynurenic acid, an antagonist of the glycine binding
site of the NMDAR, which also is the binding site of D-serine, effectively abolished the
rescue of LTP by D-serine. NMDAR-independent LTP hippocampal neurotransmission
was inhibited in slices of Pb-exposed mossy-CA3 synapses and was not rescued by
exogenous D-serine supplementation.
4.3.10.8 Neurotransmitters
Pb has been shown to compete with calcium for common binding sites and second
messenger activation. When Pb activates a calcium-dependent system in the nervous
system, it can contribute to aberrant neurotransmitter regulation and release because this
system intimately relies on calcium signaling for its homeostasis. Pb also has been shown
to interfere with other physiological divalent cations. Pb-related alterations in
neurotransmission are discussed in further detail below.
Monoamine Neurotransmitters and Stress
The monoamine neurotransmitters include dopamine (DA), serotonin (5HT), and
norepinephrine (NE). Combined exposures of maternal stress and Pb exposure can
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synergistically enhance neurobehavioral impairments in offspring and can sometimes
potentiate an effect that would otherwise be sub-threshold. Virgolini et al. (2008a) found
enhanced DA and NE release in male rats and enhanced NE release in female rats after
developmental Pb exposure (50 or 150 ppm in drinking water, 2 months prior to mating
through lactation, resulting in blood Pb levels of 11 (ig/dL and 35 (ig/dL, respectively)
and combined maternal and offspring stress. In most cases, stress potentiated the effects
of Pb exposure on offspring NE and DA concentrations. Regional 5HT levels were
unaffected in offspring with Pb exposure alone. Pb (50 and 150 ppm) combined with
stress (maternal and/or offspring) significantly potentiated 5HT levels in the frontal
cortex in females and in the nucleus accumbens (NAC) and striatum in male offspring.
The concentration of 5-Hydroxyindoleacetic acid (5HIAA), the main metabolite of 5HT,
was significantly increased in the striatum of male offspring with 150 ppm Pb exposure
alone. With 50 ppm Pb, stress potentiated striatal and frontal cortex 5HIAA in males.
Potentiation of 5HIAA levels in females was significant in the NAC with 50 ppm Pb
exposure; stress alone also significantly increased 5HIAA levels in the NAC of females
with no Pb exposure. Pb-induced changes in brain neurochemistry with or without
concomitant stress exposure are complex with differences varying by brain region,
neurotransmitter type, and sex of the animal.
Monoamine Neurotransmitters and Auditory Function
Earlier work showed that perinatal Pb exposure of rats induced increased tyrosine
hydroxylase, DA, and cerebral cortex catecholamine neurotransmission (Devi et al..
2005; Leret et al.. 2002; Bielarczvk et al.. 1996). Earlier publications examining various
time windows, durations, and doses of Pb exposure indicated varying effects on
monoamine neurotransmitters. In recent work, these neurotransmitters, among others,
have been implicated in Pb effects on auditory function in various integration centers of
the brainstem including the lateral superior olive (LSO) and the superior olivary complex
(SOC). Among various functions, the SOC is vital for sound detection in noisy settings.
A recent study in mice found significant decreases in immunostaining of LSO and SOC
brainstem sections for monoamine vesicular transporter VMAT2, 5HT, and dopamine
beta-hydroxylase (DbH, a marker for NE) after gestational-lactational Pb exposure (10 or
100 (iM Pb acetate from the formation of breeding pairs to PND21) (Fortune and Lurie.
2009). This exposure period corresponds to the period of auditory development in the
mouse. Statistically significant decreases in VMAT2 and DbH were found in mice with
blood Pb levels of 8.0 and 42 (ig/dL; however, decrements in 5HT were statistically
significant only in mice with 8.0 (ig/dL blood Pb level. Immunostaining for tyrosine
hydroxylase and transporters including VGLUT1, VGAT, VAChAT indicated that they
were unaffected by developmental Pb exposure. These data provide evidence that specific
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regions of the brainstem involved in auditory integration are affected by developmental
Pb exposure via effects on monoamine neurotransmitters. The Pb-induced effects on the
monoamine system of the portion of the brainstem involved in auditory function provide
possible mechanistic explanation for the epidemiologic and toxicological evidence for
Pb-associated decrements in auditory processing (Section 4.3.6.1).
Dopamine
The 2006 Pb AQCD (U.S. EPA. 2006b) detailed evidence for Pb-related decreased
dopaminergic cell activity in the substantia nigra and ventral segmental areas. Earlier
studies with postnatal or adult Pb exposure reported changes in DA metabolism, as
indicated by changes in DA and DOPAC, a DA metabolite. Expanding upon these
findings, a recent study assessed DA metabolism in various brain regions of year-old
male C57BL/6 mice after gestational plus lactational Pb exposure (Leasure et al.. 2008).
Both 27 and 109 ppm Pb acetate induced significant elevations in DOPAC concentration
and the DOPAC to DA ratio in the forebrain. In the forebrain, DA was significantly
decreased with 27 ppm Pb and significantly elevated with 109 ppm Pb compared to
controls. In the striatum, DOPAC was significantly elevated with both doses, but DA was
significantly elevated only with the higher dose. The striatum ratio of DOPAC to DA was
not significantly different from that in controls. These recent data expand upon the
monoamine literature base which indicates that Pb exposure of rats during the
gestational/lactational, lactational, or postweaning period producing blood Pb levels
9-34 (ig/dL increases sensitivity of the dopamine receptors (D2 and D3) (Gedeon et al.,
2001; Cory-Slechta et al., 1992), produces higher DA levels (Devi etal., 2005; Leretet
al.. 2002). and enhances catecholamine neurotransmission in the cerebral cortex,
cerebellum, and hippocampus (Devi et al., 2005).
The interaction of DA and the NO system in the striatum was studied after prenatal Pb
exposure (Nowak et al.. 2008). Blood Pb levels were not reported in this study, but
similarly treated Wistar rat pups had blood Pb levels at parturition in range of
50-100 (ig/dL (Grant etal.. 1980). 7-nitroinidazole (7-NI), a selective inhibitor of nNOS,
enhanced amphetamine-evoked DA release in the rat striatum (Nowak et al., 2008).
Prenatal Pb exposure attenuated the facilitatory effect of 7-NI on DA release in the
striatum. This interaction was ROS-independent; using spin trap measurements,
investigators found no significant concentration changes in hydroxyl radical with Pb
exposure (Nowak et al., 2008). Thus, the neuronal NO system appears to be involved in
certain aspects of Pb-related dopaminergic changes.
In various animal models, the loss of retinal DA, dopamine turnover (DOPAC:DA ratio),
or Zn was associated with Pb-induced abnormal rod-mediated scotopic ERGs. These
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effects may explain observations of Pb-associated subnormal or supernormal ERGs
observed in animals and children (Rothenberg et al.. 2002b: Lilienthal et al.. 1994;
Lilienthal et al.. 1988; Alexander and Fishman. 1984) (Section 4.3.6.2), although the
biological relevance of the variable effects of Pb exposure on subnormal versus
supernormal ERGs is not clear.
NMDA Receptors
The glutamate receptor, NMDAR, has been shown to contribute to synaptic plasticity,
and Pb exposure at various developmental stages has been shown to contribute to
aberrations in LTP or LTD in the hippocampus via reduced NMDA current, among other
mechanisms (Liu et al.. 2004). The 2006 Pb AQCD (U.S. EPA. 2006b) indicated that Pb
attenuated the stimulation of glutamate release, which in turn, affected LTP. Further, the
effects of Pb exposure on decreasing the magnitude of LTP and increasing the threshold
of the LTP in the hippocampus were found to be biphasic or nonlinear. NMDAR
subtypes have been shown to be significantly decreased with developmental Pb exposure
(Guilarte and McGlothan. 1998). Recent evidence indicated Pb-related decreases in the
gene expression and protein level of NMDAR subunits NR1, NR2A, and NR2B in
weanling male rats exposed to 4* 104 (ig/dL Pb acetate in drinking water for 8 weeks.
Several of these responses were attenuated with methionine-choline co-exposure (Fan et
al., 2010). Other recent mechanistic studies found that pretreatment of primary fetal brain
neuronal rat cultures with glutamic acid, a NMDAR agonist, reversed Pb-induced
reductions in NMDAR subunits (Xu and Raj anna. 2006) whereas pretreatment with the
NMDA antagonist MK-801 exacerbated Pb-induced NMDAR deficits (Xu and Raj anna.
2006). Further strengthening the link among Pb exposure, NMDAR function, and
learning, Guilarte et al. (2003) demonstrated that rats exposed to 1,500 ppm Pb during
gestation and lactation then reared in isolation, had reduced expression of hippocampal
NR1, reduced induction of BDNF mRNA, and learning impairment. These effects were
attenuated in Pb-exposed rats reared in an enriched environment with toys.
Other Glutamate Receptors
The metabotropic glutamate receptor (mGluR) is another well-recognized target of Pb
toxicity. In vitro (GD18 fetal rat hippocampal neurons, 0.01, 100, 1 (iM Pb chloride in
culture media) and in vivo experiments (gestational-lactational Pb acetate exposure; 500,
2,000, 5,000 ppm in dam drinking water, with respective weanling rat blood Pb levels of
18, 57, 186 (ig/dL) showed that Pb exposure decreased mGluR5 mRNA and protein in a
concentration-dependent manner (Xu et al., 2009c). Recent evidence indicates a role for
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mGluRS in synaptic plasticity, LTP, and LTD; thus, the Pb-related attenuation of mGluS
expression may represent a mechanism by which Pb impairs learning and memory.
4.3.10.9 Neurogenesis
Studies continued to show that Pb exposure decreases neurogenesis (i.e., proliferation of
neuronal cells) in the hippocampus, which is important in LTP, spatial learning, neuronal
outgrowth, and possibly effects related to schizophrenia pathophysiology. Coherence for
these findings is provided by evidence for Pb-induced decreases in NMDAR, which
mediates the integration of new neurons into existing neuronal pathways in the adult
hippocampal DG. Earlier work by Schneider et al. (2005) showed that postnatal Pb
exposure (PND25 to PND50 or PND55, 1,500 ppm Pb acetate in chow, resulting in blood
Pb level of 20 (ig/dL) of male Lewis rats induced significant decrements in BrdU
incorporation (an indicator of DNA replication) at PND50-PND55.
Recent publications affirm this previous finding with different sex of animals, dosing and
exposure time windows. Postnatal Pb exposure of Wistar rat pups from PND1 to PND30
(2,000 ppm Pb acetate, resulting in blood Pb levels of 34 and 6.5 (ig/dL at PND21 and
PND80, respectively) induced a statistically significant decrement in the number of new
cells (BrdU positive cells) in the DG at PND80 (Fox etal.. 2010) (Figure 4-13).
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„ 600°
1 5000
1
Ł 4000
-o «
m g 3000
0
JS 2000
Z 1000
n
-
— , —
-
•
-
-
Con
1.5r
•
0.5
O.O1-
Con
Note: Light micrograph pictures of BrdU positive cells (proliferating cells undergoing DMA replication) shown as black dots, show
more proliferating cells in control (A) than Pb-exposed rats (B). Counts of BrdU positive cells (C) but not Volume of hippocampus
dentate gyrus (D), are lower in Pb-exposed rats (black bars) than control (white bars) rats. *p <0.05 vs. control. Rats were exposed
to 2,000 ppm Pb acetate postnatal day 1-30 (blood Pb levels 34 ug/dL) and were examined at postnatal day 80.
Source: Reprinted with permission of Elsevier Science, Fox et al. (201 0)
Figure 4-13 Neurogenesis (production of new cells) in the rat hippocampal
dentate gyrus after early postnatal Pb exposure.
In another study, lifetime Pb exposure beginning in gestation (1,500 ppm Pb acetate in
chow from 10 days before mating to PND50 or PND78, resulting in blood Pb levels
26 (ig/dL) induced significant decrements in hippocampal granule cell neurogenesis in
adult Long-Evans rats (Verina et al., 2007). Also, Pb-exposed animals had significant
decreases in brain volume in the stratum oriens (SO) region of the hippocampus,
specifically in the mossy fiber terminals of the SO. Pb-exposed animals also showed a
significant decrease in the length-density of immature or newly-formed neurons in the
outer portion of the DG. These findings show that Pb exposure at doses relevant to this
ISA induced significant decreases in adult hippocampus granule cell neurogenesis and
morphology, potentially providing mechanistic explanations for Pb-induced neuronal
aberrations and impairments in downstream effects such as learning and memory.
Exposure of zebrafish embryos to Pb (50-700 (iM Pb acetate in embryo medium from 0
to 6 days post hatch) caused significant apoptosis of brain cells (increased TUNEL
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positive brain cells) and decreased brain levels of some (gfap and huC) but not all (crestin
and neurogeninl) genes involved in neurogenesis (Dou and Zhang. 2011).
4.3.10.10 Neurite Outgrowth
As described in the 2006 Pb AQCD (U.S. EPA. 2006b). Pb was shown to decrease
neurite outgrowth in vitro and mediate such effects via protein kinase mediated pathways
(MAPK/ERK); earlier work had documented decreased primary DA neuron outgrowth
with 0.001 (iM Pb exposure in vitro (Schneider et al.. 2003). A recent study showed that
gestational exposure of female Wistar rats to 500-4,000 (iM Pb chloride (resulting in
offspring blood Pb levels up to 12 (ig/dL) significantly decreased offspring hippocampal
neurite outgrowth and reduced the expression of hippocampal polysialylated neural cell
adhesion molecule (PSA-NCAM), NCAM, and sialyltransferase (Hu et al.. 2008b). PSA-
NCAM is transiently expressed in newly formed neurons during the period of neurite
outgrowth from embryogenesis until the early postnatal period and is down-regulated in
the brains of adults except in areas known to exhibit synaptic plasticity (Seki and Arai.
1993). NCAM is important for memory formation, plasticity, and synapse formation, and
its suppression by gestational Pb exposure may represent a mechanism that mediates
Pb-associated impairments in cognitive function.
4.3.10.11 Epigenetics
Many investigators are beginning to show that environmental chemical exposures and air
pollution exposure are associated with epigenetic changes in humans (Baccarelli and
Bollati. 2009; Pavanello et al.. 2009; Tarantini et al.. 2009; Bollati et al.. 2007).
Epigenetic changes involve changes in DNA expression without changes in the DNA
sequence, and these changes may be heritable. Epigenetic changes include changes in
histone modification, DNA methylation, miRNA, or pathways that affect these processes.
Differential epigenetic modification has the potential to contribute to disease by silencing
or activating genes in an aberrant manner. For example, a recent study identified
differential methylation of a specific locus in monozygotic twins discordant for
schizophrenia (Dempster et al., 2011); Pb was not examined in this study.
DNA methyltransferases catalyze the transfer of a methyl group to DNA and are
important in epigenetics (i.e., silencing of genes like tumor suppressors) and imprinting.
DNA methyltransferase activity was significantly decreased in cortical neurons from
aged monkeys at ages 20-23 years after infancy exposure (1.5 mg/kg Pb acetate via diet
per day PND1-PND400, blood Pb level 19-26 (ig/dL at PND400) and in fetal mouse
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brain cells 1 week after 24-hour exposure to 0.1 (iM Pb in culture (Wu et al., 2008a).
Cerebral cortex tissue was obtained from female monkeys examined in this study for
further analysis. Decreases in DNA methyltransferases (Dnmtl, DnmtSa) were noted in
control monkey brains as they aged, and these decreases were exacerbated by infancy Pb
exposure (Bihaqi et al., 2011). Another enzyme involved in DNA methylation, methyl
CpG binding protein 2, showed a similar trend as the Dnmts. Profiles of the histone
modifying proteins increased with age in control animals. These age-related increases
were significantly attenuated in Pb-exposed animals. Methyltransferases catalyze
biological methylation reactions using cofactor S-adenosyl methionine (SAM) as the
methyl donor. In rats, SAM exposure after gestational-lactational Pb exposure
(1,500 ppm Pb acetate via drinking water of dams followed by 20-22 days of daily
20 mg/kg BW SAM exposure of offspring) improved hippocampal LTP and Morris water
maze performance at PND44-PND54 (Cao et al.. 2008). Thus, the improved cognition
and synaptic plasticity observed with co-exposure to Pb and the methyl donor SAM
suggest that methylation reactions may be involved in Pb-associated effects on cognition.
4.3.10.12 Cholesterol and Lipid Homeostasis
Various pathological conditions are associated with elevated plasma free fatty acids or
elevated cholesterol. Adult male rats exposed to Pb acetate (200, 300, or 400 ppm) in
their drinking water for 12 weeks had increased cholesterogenesis and phospholipidosis
in brain tissue (Ademuyiwa et al.. 2009). Pb-induced changes in brain cholesterol showed
an inverse U concentration-response relationship, with the largest increase in brain
cholesterol observed with 200 ppm Pb followed by 300 ppm Pb. Animals exposed to
400 ppm Pb did not have significant changes in brain cholesterol. In a separate study, Pb
treatment (single dose 100 (imol/kg, i.v.) was shown to depress the activity of
cholesterol-7-a-hydroxylase, an enzyme involved in biosynthesis of bile acid, which
mediates elimination of cholesterol from the body (Kojima et al.. 2005). In Ademuyiwa
et al. (2009). Pb exposure significantly increased brain triglycerides by 83% at 300 ppm
and by 108% at 400 ppm. At 200 ppm, Pb exposure induced a statistically nonsignificant
decrease in brain triglycerides. All three Pb concentrations induced significantly
increased brain phospholipids. Interestingly, plasma free fatty acids were significantly
elevated in a concentration-dependent manner; plasma triglycerides and cholesterol were
unaffected by Pb exposure. The molar ratio of brain cholesterol to phospholipids, an
indicator of membrane fluidity, was significantly increased at 200 and 300 ppm Pb
exposure, indicating increased membrane fluidity. Brain Pb in all dose groups was below
the limit of detection (0.1 ppm). Blood Pb levels for the 0; 200; 300; and 400 ppm Pb
exposures were 7; 41; 61; and 39 (ig/dL, respectively; higher than those relevant to this
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ISA. In summary, a recent study found that adult 12-week Pb exposure significantly
increased brain cholesterol, triglycerides, and phospholipids as well as significantly
increased plasma free fatty acids in rats. The increase in cholesterol was more prominent
at the lower 200 ppm Pb dose. The impacts of these Pb-related changes in
phospholipidosis and cholesterogenesis in the brain on downstream nervous system
effects are not well characterized.
4.3.11 Lifestages and Time Periods of Pb Exposure and Neurodevelopmental
Effects
Environmental exposures during critical lifestages can affect key physiological systems
that orchestrate plasticity (Teinberg. 2007). Exposures during the prenatal and/or early
postnatal period may be especially detrimental for neurodevelopmental effects because of
active neuronal growth and/or synaptogenesis/pruning structure that occur during these
periods (Rice and Barone. 2000; Nolte, 1993). However, brain development has been
shown to continue throughout adolescence. MRI studies of the Child Psychiatry Branch
cohort of children and adults ages 3-30 years have shown peak total cerebral volume at
age 10.5 and 14.5 years in females and males, respectively (Giedd et al.. 2009; Lenroot
and Giedd. 2006). The volume of the cerebellum was found to peak two years later.
Lateral ventricular volume and white and gray matter volume also were found to increase
throughout adolescence. Gray matter volume peaked 1 to 3 years earlier in females than
males. These observations that brain development is active throughout childhood and in
adolescence indicate the potential for neurodevelopment to be altered later in childhood.
Epidemiologic studies consistently show that blood Pb levels measured during various
lifestages and time periods, including the prenatal period, early childhood, later
childhood, and averaged over multiple years, are associated with cognitive function
decrements and increases in externalizing and internalizing behaviors. These observations
of Pb-associated elevated risk of neurodevelopmental decrements in children are well
supported by findings in animals that prenatal and/or early postnatal or lifetime Pb
exposure alters brain development via changes in synaptic architecture (Section 4.3.10.4)
and neuronal outgrowth (Section 4.3.10.10) and leads to impairments in memory and
learning (Section 4.3.2.3) and increases in impulsivity (Section 4.3.3.1). In monkeys, Pb
exposures during multiple lifestages and time periods, including lifetime, infancy, or
juvenile to adulthood, resulted in impaired cognitive function, although not on all tests
(Rice. 1992b: Rice and Gilbert. 1990a: Rice. 1990; Rice and Karpinski. 1988). On one
test of executive function in the same monkeys at ages 5-6 years, impairments were
found with lifetime Pb exposure starting from birth or starting at age 400 days but not
infancy-only exposure (Rice and Gilbert. 1990b). The observations indicate that
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gestational or infancy Pb exposures are not necessary to induce cognitive function
decrements in juvenile animals.
Unlike other organ systems, the unidirectional nature of CNS development limits the
capability of the developing brain to compensate for cell loss, and environmentally-
induced cell death can result in a permanent reduction in cell numbers (Bayer. 1989).
Hence, when normal development is altered, the early effects have the potential to persist
into adult life even in the absence of concurrent exposure, magnifying the potential public
health impact. Some epidemiologic evidence indicates associations of earlier childhood
blood or tooth Pb levels with cognitive function decrements, attention decrements, and
increases in behaviors related to conduct disorders in adolescents or adults (Mazumdar et
al..2011; Fergusson et al.. 2008; Wright et al.. 2008; Ris et al. 2004; Bellinger et al.
1994a: Stiles and Bellinger. 1993). These epidemiologic studies did not examine adult
blood Pb levels, thus the relative influence of adult Pb exposure cannot be ascertained. In
the Boston cohort, stronger associations observed for age 2 year blood Pb level than
concurrent blood Pb level with FSIQ decrements at ages 57 months and 10 years
indicated an effect of earlier rather than later childhood Pb exposures (Bellinger et al..
1992; 1991). The persistence of effects of early exposures is supported by findings of
impaired learning in adults monkeys that had infancy-only Pb exposure (Rice. 1992b:
Rice and Gilbert. 1990a; Rice, 1990). A few available toxicological studies also found
that infancy Pb exposure (but not adult exposure in rats) led to neurodegenerative
amyloid plaque formation in the brains of aged rodents and monkeys (Section 4.3.9.1).
With repeated assessments of children (prenatally to later childhood and early adulthood),
the prospective cohort studies have provided data to compare the neurodevelopmental
effects associated with blood Pb levels measured at different lifestages and time periods.
In the collective body of evidence, cognitive function decrements in children have been
associated with prenatal, early childhood, childhood average, and concurrent blood Pb
levels, without clear indication of a single critical lifestage or duration of Pb exposure
related to risk of neurodevelopmental effects in children. The identification of critical
developmental periods with regard to risk of neurodevelopmental decrements from Pb
exposure has been complicated by the high degree of correlation of the blood Pb levels of
children over time and the confounding of age and peak blood Pb levels (Hornung et al..
2009; Lanphearetal.. 2005: Dietrich et al.. 1993a).
As described in detail in the 2006 Pb AQCD (U.S. EPA. 2006b). several studies with
varying lengths of follow-up demonstrated associations of prenatal blood Pb levels
(maternal and umbilical cord) with decrements in cognitive function throughout
childhood and into early adulthood (Section 4.3.2). These findings are consistent with the
observations of active CNS development occurring during prenatal development as
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described above. Substantial fetal Pb exposure may occur from mobilization of maternal
skeletal Pb stores, which may be related to past maternal Pb exposures (Gulson et al..
2003; Hu and Hernandez-Avila. 2002), and the transfer of Pb across the placenta
(Section 3.2.2.4). Among 94-211 mother-child pairs in Albany County, NY, maternal
pregnancy-cord blood Pb level correlations of 0.53-0.81 were reported, depending on the
stage of pregnancy, indicating the influence of maternal blood Pb levels on newborn
blood Pb levels (Schell et al.. 2003). Depending on the magnitude of child exposure, the
contribution of maternal blood Pb on child blood Pb levels appears to diminish rapidly
over a period of a few months following birth, after which child blood Pb levels may be
influenced mainly by postnatal Pb exposures (Section 3.4.1). Thus, associations of
neurodevelopmental outcomes assessed after infancy with postnatal blood Pb levels may
reflect effects of postnatal Pb exposures.
In most studies of very young children, ages <2 years, decrements in MDI score were
associated with higher prenatal (maternal or cord) and concurrent blood Pb levels (Table
4-14). Among studies that examined blood Pb levels at multiple time periods, some found
a larger magnitude decrement in MDI per unit increase in prenatal blood Pb than
concurrent blood Pb (Hu et al.. 2006; Gomaaetal.. 2002) (Table 4-14). The mean was
higher for prenatal blood Pb level than concurrent in one study (Hu et al.. 2006). whereas
another study reported similar means for prenatal and concurrent blood Pb levels (Gomaa
et al.. 2002). Bellinger et al. (1987) did not report effect estimates for concurrent blood
Pb level and only indicated lack of statistically significant association with MDI. In the
Yugoslavia cohort, per log increase in blood Pb level, the MDI decrement at age 2 years
was larger for concurrent blood Pb than for prenatal cord blood Pb (Wasserman et al.,
1992). Concurrent blood Pb levels were higher than prenatal cord blood Pb levels. The
collective evidence indicates that both prenatal and postnatal child Pb exposures may
contribute to neurodevelopmental effects in children at to age 2 years, with some
indication that prenatal Pb exposure has a stronger effect.
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Table 4-14 Associations of cognitive function with blood Pb levels measured at various lifestages and time
periods in prospective studies.
Study3
Cognitive
Bellinger et
(1987)
Study Population and Methodological Details
Within studies, effect estimates are presented in order of
lifestage or time period of blood Pb measurement, with
largest effect estimate in bold3
Assessment at Age 2 Years and Younger
al. 182 children followed from birth (1979-1981) to age 2
yr, Boston, MA area.
Recruitment from birth hospital. High follow-up
participation. Participants had higher cord blood Pb
level, SES, HOME score, maternal education and IQ,
lower maternal age, were white. Regression model
adjusted for maternal age, education, race, IQ, years
of smoking, and alcohol drinks/week in 3rd trimester,
SES, HOME score, child sex, birth weight, gestational
age, birth order.
Dietrich et al. 245-280 children followed prenatally to age 6 mo,
(1986) Cincinnati, OH
No information on participation rate. Log linear
regression model adjusted for birth weight, gestation,
sex. Also considered potential confounding by SES,
HOME score, prenatal smoking and alcohol use,
maternal Fe binding capacity, child sex, race.
Huetal.
(2006)
83-146 children born 1997-1999 followed prenatally to
age 2 yr, Mexico City, Mexico
Recruitment from prenatal clinic. Moderate follow-up
participation. Eligible similar to non-eligible. Log linear
regression model adjusted for maternal age and IQ,
child sex, current weight, height-for-age Z score, and
concurrent blood Pb (in models examining prenatal
blood Pb). Considered potential confounding by other
unspecified factors.
Blood Pb Metrics
Analyzed (ug/dL)
Prenatal (cord)
Mean (SD): 6.6 (3.2)
Low: < 3
Medium: 6-7
High: >10
Mean (SD) Prenatal
(maternal): 8.0 (3.8)
Concurrent: 4.5 (2.9)
Mean (SD)
Prenatal (maternal 1st
trimester): 7.1 (5.1)
Prenatal avg NR
Earlier childhood at
12 mo: 5.2(3.4)
Concurrent: 4.8 (3.7)
Outcome
Analyzed
Overall Bayley
MDI among
Ages 6, 12, 18,
and 24 mo
Bayley MDI
Age 6 mo
Bayley MDI
Age 24 mo
Effect Estimate (95% Cl)b
Score in high cord blood Pb group
vs. Low cord blood: -4.8 (-7.3, -2.3)c
vs. Medium cord blood: -3.8 (-6.3, -1.3)c
Concurrent reported to not to be
significantly associated with overall MDI,
no quantitative data reported.
Per log increase in blood Pb:
Prenatal: -15 (-27, -3)
Concurrent: -15 (-25, -5)
Per log increase in blood Pb:
Prenatal 1st trimester: -4.1 (-8.1, -0.17)
Prenatal avg: -3.5 (-7.7, 0.63)
Age 12 month: -2.4 (-6.2, 1.5)
Concurrent: -1.0 (-3.9, 1.9)
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Table 4-14 (Continued): Associations of cognitive function with blood Pb levels measured at various lifestages and time
periods in prospective studies.
Study3
Gomaa et al.
(2002)
Wasserman
et al. (1992)
Ernhart et al.
(1988: 1987)
Study Population and Methodological Details
Within studies, effect estimates are presented in order of
lifestage or time period of blood Pb measurement, with
largest effect estimate in bold3
161-197 children followed from birth to age 2 yr,
Mexico City, Mexico
Recruitment from birth hospital. Moderate participation
but high retention, no selective attrition. Log linear
regression model adjusted for maternal IQ, age,
parents in home, breastfeeding duration, parental
education, child hospitalization status, and sex. Did not
consider potential confounding by parental caregiving
quality.
392 children followed prenatally to age 2 yr, K.
Mitrovica, Pristina, Yugoslavia.
53% live near Pb sources. High follow-up participation,
no selective attrition. Log linear regression model
adjusted for sex, birth order, birth weight, ethnic group,
HOME score, maternal education, age, and IQ.
146-165 children, followed prenatally to age 2 yr,
Cleveland, OH
Prospective. Recruitment at birth hospital. High
follow-up participation. More white, higher IQ,
nonalcohol using mothers not followed. 50% born to
alcoholic mothers. Linear regression adjusted for age,
race, sex, birth order, birth weight, parental education,
maternal IQ, Authoritarian Family Ideology, HOME
score.
Blood Pb Metrics
Analyzed (ug/dL)
Mean (SD)
Prenatal (cord):
6.7(3.4)
Concurrent:
8.4(4.6)
Mean (SD)
Prenatal (cord): 14(10)
Concurrent:
35 K. Mitrovica, 8.5
Pristina
Mean (SD)
Prenatal cord:
6.0(1.8)
6 mo:
10(3.3)
Concurrent:
17(6.5)
Outcome
Analyzed
Bayley MDI
Age 24 mo
Bayley MDI
Age 24 mo
Bayley MDI
Age 2 yr
Effect Estimate (95% Cl)b
Per log increase in blood Pb:
Prenatal: -4.9 (-9.0, -0.89)
Concurrent: -0.09 (-0.58, 0.40)
Per log increase in blood Pb:
Prenatal: -3.5 (-8.6, 1.7)
Concurrent: -5.3 (-10, -0.53)
Variance estimates:
Prpnatal' 0 0003 t - -0 91 d
Age 6 mo: 0.00, p = 0.95d
Concurrent: 0.00, p = 0.95d
Cognitive Assessment at Age 3 Years and Older
Canfield et al.
(2003a)
171-172 children born 1994-1995 followed from age 6
mo to 5 yr, Rochester, NY
Recruitment from study of dust control. 73% nonwhite.
Moderate follow-up participation but no selective
attrition. Linear regression model adjusted for child
sex, Fe status, birth weight, maternal race, education,
IQ, income, and prenatal smoking status, HOME
score.
Mean (SD)
Infancy avg (6-24 mo):
7.0(3.8)
Peak:
11(7.1)
Concurrent:
5.8(4.1)
Lifetime (to age 5 yr)
avg: 7.4 (4.3)
FSIQ
Stanford-Binet
Age 5 yr
Per 1 ug/dL increase in blood Pb:
Infancy avg: -0.53 (-0.93, -0.13)
Peak: -0.26, (-0.47, -0.05)
Concurrent: -0.61 (-0.99, -0.24)
Lifetime avg: -0.57 (-0.93, -0.20)
4-251
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Table 4-14 (Continued): Associations of cognitive function with blood Pb levels measured at various lifestages and time
periods in prospective studies.
Study3
Study Population and Methodological Details
Within studies, effect estimates are presented in order of
lifestage or time period of blood Pb measurement, with
largest effect estimate in bold3
Blood Pb Metrics
Analyzed (ug/dL)
Outcome
Analyzed
Effect Estimate (95% CI)D
Wasserman 332 children followed prenatallyto ages 3-4 yr(1991),
et al. (1994) K. Mitrovica, Pristine, Yugoslavia
Moderate follow-up participation. More participants
were male, from K. Mitrovica, Albanian, and had lower
maternal IQ and education and HOME score. Log
linear regression model adjusted for HOME score,
maternal age, IQ, and education, language, sibship
size, birth weight, child sex.
Mean (SD)
Prenatal (cord):
14(11)
Geometric means
Concurrent
K. Mitrovica: 40
Pristine: 9.6
FSIQ
McCarthy GCI
Age 4 yr
Per log increase in blood Pb:
Prenatal cord:-7.1 (-12, -3.1)
Age 2 yr:-10.4 (-15,-5.7)
Concurrent: -9.4 (-14, -4.6)
Wasserman 390 children followed prenatally (1985-1896) to age 7
et al. (2000a) yr, K. Mitrovica, Pristine, Yugoslavia
Moderate follow-up participation. More participants
were from K. Mitrovica, Albanian, with lower maternal
IQ and education and HOME score. Regression model
adjusted for HOME score, maternal age, education,
IQ, and # live births, child sex, birthweight, ethnicity,
age.
Mean of average during
period
Prenatal: 10
Age 0-2: 13
Age 2-3: 19
Age 2-4: 19
Age 2-5: 19
Age 2-7: 17
FSIQ
McCarthy GCI,
ages 3-4 yr
WPPSI, age 5
yr
WISC-III, age 7
yr
Prenatal: -6.1 (-8.7, -3.4) per log increase
in blood Pb
Postnatal blood Pb analyzed as
categorical variable
No increase prenatal to postnatal:
Reference
>50% increase in postnatal only from birth
to age 2 yr:-1.8 (-3.5, -0.05)c
>50% increase in postnatal from birth to
age 2 yr and from age 2 yr onward:
-2.7 (-4.9, -0.52)c
Bellinger et al. 148 children followed from birth (1979-1981) to age 10
(1992) yr, Boston, MA area
Moderate follow-up participation. Participants had
higher SES and HOME scores. Linear regression
model adjusted for HOME score (age 10 and 5 yr),
maternal race, age, IQ, and marital status, SES, child
sex, birth order, and stress. Plus: # residence changes
(2 yr), family stress, birth weight, # daycare situations
to 5 yr (10 yr). Also considered potential confounding
by psychiatric factors, child serum ferritin levels.
Mean (SD)
Age 6 mo:
6.7(7.0)
Earlier childhood Age 2
yr:
6.5(4.9)
Concurrent:
2.9(2.4)
FSIQ Per 1 ug/dL increase in blood Pb:
WISC-Revised Age 6 mo: -0.13 (-0.42, 0.16)
Age 10 yr Age 2 yr: -0.58 (-0.99, -0.17)
Concurrent -0.46 (-1.5, 0.56)
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Table 4-14 (Continued): Associations of cognitive function with blood Pb levels measured at various lifestages and time
periods in prospective studies.
Study3
Study Population and Methodological Details
Within studies, effect estimates are presented in order of
lifestage or time period of blood Pb measurement, with
largest effect estimate in bold3
Blood Pb Metrics
Analyzed (ug/dL)
Outcome
Analyzed
Effect Estimate (95% CI)D
Dietrich et al. 199-231 children followed from birth (1979-1985) to
(1993b) age 6.5 yr, Cincinnati, OH
Recruitment at prenatal clinic. High follow-up
participation. Participants had slightly higher age 1 yr
blood Pb levels. Primarily African-American. Linear
regression model adjusted for HOME score, child birth
weight, birth length, and sex, maternal IQ and prenatal
smoking. Also considered potential confounding by
perinatal complications, prenatal maternal substance
abuse, nutritional status.
Mean (SD)
Prenatal (maternal)
8.3(3.7)
Age 4-5 yr avg
11.8(6.3)
Concurrent:
NR
Lifetime (to age 6.5 yr)
avg: NR
FSIQ Per 1 ug/dL increase in blood Pb:
WISC-Revised Prenatal: 0.15 (-0.26, 0.56)
Age 6.5 yr Concurrent: -0.33 (-0.60, -0.06)
Lifetime avg: -0.13 (-0.35, 0.09)
Baghurst 494 children followed from birth (1979-1982) to ages 7-
et al. (1992) 8 yr, Port Pirie, Australia
Moderate follow-up participation. Participants had
higher SES, more breastfeeding, less maternal
smoking. Log linear regression model adjusted for sex,
birth weight, birth order, feeding method, breastfeeding
duration, parental education and smoking status,
maternal age and IQ, SES, HOME score, parents
living together.
Means of 2nd quartile
Prenatal (cord): 7.4
Age 0-2 yr: 17
Lifetime (to age 7 yr)
avg: 16
FSIQ
WISC-Revised
Age 7-8 yr
Per log increase in blood Pb:
Prenatal: 0.6 (-2.1, 3.3)
Age 0-2 yr: -4.6 (-8.7, -0.48)
Lifetime avg: -3.7 (-8.6, 1.2)
Schnaas et al. 136-150 children followed from birth (1987-1992) to
(2006) age 10 yr, Mexico City, Mexico.
Recruitment at prenatal clinic. Low follow-up
participation. Participants had higher SES, FSIQ,
higher blood Pb level before age 5 yr, lower at older
ages. Log linear mixed effects model adjusted for SES,
maternal IQ, child sex, birth weight, & postnatal blood
Pb, age of 1st FSIQ test, random slope for subject.
Also considered potential confounding by HOME
score. Covariates assessed in pregnancy or within age
6 mo.
Geometric Mean (5th-
95th)
Prenatal (maternal
28-36 week): 7.8(2.5-
25)
Age Syr: 9.3(3.8-18)
Age 6-10 yr avg: 6.2
(range = 2.2-19)
FSIQ Per log increase in blood Pb:
WISC-Revised Prenatal: -4.0 (-6.4, -1.7)
Ages 6-10 yr Age 5 yr: -0.32 (-4.3, 3.4)
Age 6-10 yr avg: -2.5 (-4.1, -0.81)
4-253
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Table 4-14 (Continued): Associations of cognitive function with blood Pb levels measured at various lifestages and time
periods in prospective studies.
Study3
Study Population and Methodological Details
Within studies, effect estimates are presented in order of
lifestage or time period of blood Pb measurement, with
largest effect estimate in bold3
Blood Pb Metrics
Analyzed (ug/dL)
Outcome
Analyzed
Effect Estimate (95% CI)D
Ris et al. 195 children followed prenatally (1979-1985) to age
(2004) 15-17 yr, Cincinnati, OH
Prospective. Recruitment at prenatal clinic. High
follow-up participation, no selective attrition. Mostly
African-American. Linear regression model adjusted
for SES, maternal IQ, HOME, adolescent marijuana
use, obstetrical complications. Also considered
potential confounding by birth outcomes, maternal age,
maternal prenatal smoking, alcohol, marijuana, and
narcotics use, # previous abortions, stillbirths,
gravidity, parity, public assistance, child sex, age,
health and Fe status, caregiver education.
Prenatal, Earlier
childhood (age 6.5 yr),
Earlier childhood avg
(age 3-78 mo):
NR
Learning/IQ
factor score
WISC-III,
WRAT-3
indices
Ages 15-17 yr
Per 1 ug/dL increase in blood Pb:
Prenatal: -0.08 (-0.18, 0.03)
Age 6.5 yr: -0.08 (-0.17, 0.00)
Age 3-78 mo avg: -0.03 (-0.09, 0.02)
Lanphear 1,333 children pooled from Boston, Cincinnati,
et al. (2005) Cleveland, Mexico City, Port Pirie, Rochester, and
Yugoslavia cohorts.
Uniform analytic method applied to cohorts from
diverse locations and demographic characteristics.
Blood Pb levels and FSIQ measured at different ages.
Several sensitivity analyses to examine heterogeneity
of results by cohort, confounding, and model
specification. Log linear regression model adjusted for
HOME score, birth weight, maternal IQ and education.
Also considered potential confounding by child sex,
birth order, marital status, maternal age, prenatal
smoking status and alcohol use.
Median (5th-95th)
Early childhood (mean
ages 6-24 mo):
12.7(4.0-35)
Peak:
18(6.2-47)
Lifetime avg (to ages
4.8-1 Oyr):
12(4.1-35)
Concurrent:
9.7 (2.5-33)
FSIQ Per log increase in blood Pb:
Various tests Early childhood: -2.0 (-3.3, -0.81)
Ages 4.8-10 yr Peak: -2.9 (-4.1, -1.6)
Concurrent: -3.0 (-4.3, -1.8)
Lifetime avg:-2.7 (-3.7,-1.7)
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Table 4-14 (Continued): Associations of cognitive function with blood Pb levels measured at various lifestages and time
periods in prospective studies.
Study3
Pocock et al.
(1994)
Study Population and Methodological Details
Within studies, effect estimates are presented in order of
lifestage or time period of blood Pb measurement, with
largest effect estimate in bold3
Meta-analysis of 5 prospective studies (1,260 children)
from Port Pirie and Sydney, Australia, Cincinnati,
Cleveland, Boston
Meta-analysis of combining covariate-adjusted effect
estimates from individual studies.
Blood Pb Metrics
Analyzed (ug/dL)
Earlier childhood (2 yr)
range in means:
6.8-21
Around birth, Postnatal:
NR
Outcome
Analyzed
FSIQ
Various tests
Ages 5-10 yr
Effect Estimate (95% Cl)b
Per doubling of blood Pb from 10-20
ug/dL:
Around birth: 0.18 (-1.0, 1.4)
Age 2 yr: -1.9 (-2.8, -0.85)
Postnatal mean: -0.88 (-2.0, 0.26)
MDI = Mental Development Index, FSIQ = full-scale IQ, NR = Not reported, McCarthy GCI = McCarthy General Cognitive Index, WPPSI = Wechsler Preschool and Primary Scale of
Intelligence, WISC = Wechsler Intelligence Scale for Children
"Results are presented first for MDI in children up to age 2 years, then for FSIQ in children ages 3 years and older. Within studies, effect estimates are presented in order of increasing
lifestage or time period of blood Pb measurement, with the largest effect estimate in bold. After age of assessment, studies are grouped by strength of study design,
representativeness of the study population characteristics and blood Pb levels examined, and extent of consideration for potential confounding. There is not necessarily a continuum of
decreasing strength across studies.
bEffect estimates are standardized to a 1 ug/dL or log increase in blood Pb level in analyses of blood Pb as a continuous variable.
°Effect estimates represent comparisons between children in different categories of blood Pb level, with children in the lower blood Pb category serving as the reference group.
dSufficient data were not available to calculate 95% Cl.
4-255
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Prenatal and early postnatal (age 6 month) blood Pb levels also were associated with
cognitive function in studies that included school-aged children (ages 5-17 years) (Table
4-14). However, most of these studies also found cognitive function decrements in
association with postnatal blood Pb levels, and results did not identify an individual
critical postnatal time period of blood Pb measurement associated with cognitive function
decrements. In the Yugoslavia cohort, the FSIQ decrement per unit increase in blood Pb
level was larger for prenatal maternal than concurrent blood Pb when children were
examined at ages 3-7 years (Wasserman et al.. 2000a). but larger for concurrent than
prenatal maternal or cord blood Pb when children were examined at age 4 years
(Wasserman et al., 1994). In the Cincinnati cohort, the association with FSIQ was
stronger for concurrent than for prenatal blood Pb levels at age 6.5 years (Dietrich et al..
I993b), but in adolescents ages 15-17 years, increases in both prenatal and earlier
childhood (age 6.5 years) blood Pb levels were associated with decrements in a
learning/memory composite score (Ris et al., 2004). In the Boston and Port Pirie cohorts,
increases in age 2 year or 0-2 year average blood Pb levels were associated with larger
FSIQ decrements at ages 10 and 7 years than were concurrent or lifetime average blood
Pb levels, respectively (Baghurst et al.. 1992; Bellinger etal.. 1992). However, in the Port
Pirie cohort, were was an association found for lifetime average (to age 7 years) blood Pb
level. Among children ages 6-10 years in Mexico City, per unit increase, prenatal blood
Pb level was associated with a larger FSIQ decrement than were blood Pb levels
measured at each age from 1 to 5 years or averaged from age 6 to 10 years (Schnaas et
al., 2006). In contrast, results from the Rochester cohort at age 5 years indicated that the
FSIQ decrement per unit increase in blood Pb level was larger for lifetime average and
concurrent blood Pb than peak blood Pb (Canfield et al.. 2003a). Collectively, the
epidemiologic findings indicate that blood Pb levels measured at various postnatal time
periods, early childhood (e.g., age 2 year), lifetime average (e.g., to 7 years), later
childhood (e.g., age 6-10 year average), and concurrent blood Pb levels are associated
with decrements in cognitive function when assessed in populations that include school-
aged children ages 5-17 years.
Consistent with individual studies, analyses combining studies pointed to associations of
FSIQ in school-aged children with blood Pb levels measured at various lifestages and
time periods. The analysis pooling data from seven prospective studies found that
increases in early childhood (age 6-24 month average), peak, concurrent, and lifetime
average blood Pb levels were associated with decreases in FSIQ in children ages
4.8-10 years (Table 4-14). Investigators reported that the model with concurrent blood Pb
level explained the largest proportion of variance in FSIQ (R2) (Lanphear et al.. 2005). In
a meta-analysis of results from five prospective studies (Pocock etal.. 1994). the decrease
in FSIQ per unit increase in blood Pb level was larger for peak (around age 2 years)
blood Pb level than blood Pb level measured around birth or after age 2 years. Deciduous
4-256
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tooth Pb levels have been associated with decrements in cognitive function, decrements
in attention, as well as increases in impulsivity and hyperactivity in children and young
adults (Table 4-8. Table 4-9. and Table 4-11). These results that indicate cumulative Pb
exposure over several years may contribute to neurodevelopmental effects in children. In
school-aged children, concurrent blood Pb levels reflect recent exposures and past Pb
exposures that are mobilized from bone remodeling to blood. Thus, associations with
concurrent blood Pb levels also may reflect an effect of cumulative past and more recent
Pb exposures.
Studies conducted in the Cincinnati cohort examined diverse nervous system effects and
found that prenatal or neonatal blood Pb levels were associated with poorer auditory
processing in children ages 5 years (Dietrich et al.. 1992) and higher parental ratings of
delinquent behavior in adolescents ages 15-17 years (Dietrich et al.. 2001) but
inconsistently with motor function decrements at ages 4-10 years (Bhattacharya et al..
2006; Bhattacharya et al.. 1995; Dietrich et al.. 1993a) and not with cognitive function in
children at age 6 years (Dietrich et al.. 1993b). These findings suggest that the critical
timing of Pb exposure may vary among nervous system endpoints.
Some studies have aimed to improve the characterization of important lifestages and time
periods of Pb exposure by examining children in whom blood Pb levels are not strongly
correlated overtime (i.e., children whose blood Pb level ranking changed overtime)
(Hornung et al.. 2009; Schnaas et al.. 2006; Chen et al.. 2005a; Wasserman et al.. 2000a;
Tong etal.. 1998; Bellinger et al.. 1990). Some calculated specific blood Pb metrics that
reflected the change over time (i.e., ratio of blood Pb measured at different times, amount
of change over time), and collectively most results indicated FSIQ decrements in
association with higher concurrent blood Pb levels but did not conclusively demonstrate
stronger findings for early or concurrent blood Pb levels (Table 4-15).
Schnaas et al. (2006) did not calculate a blood Pb level metric to reflect the change over
time but found a weak correlation between maternal blood Pb levels at 28-36 weeks of
pregnancy and repeated measures of child blood Pb level between ages 1 and 10 years
(Pearson r < 0.23). In this study, children in Mexico City were followed prenatally to age
10 years. In models analyzing blood Pb levels at different ages individually or together in
one mixed effects model that included prenatal and multiple postnatal blood Pb measures,
the decrement in FSIQ per unit increase in blood Pb level was largest for maternal
28-36 week blood Pb level (Table 4-14). In the model that included blood Pb level at
multiple ages, analysis of variance inflation factors indicated a lack of collinearity among
the serial blood Pb measures
4-257
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Table 4-15 Comparisons of blood Pb-cognitive function associations in groups of children with different
temporal trends in blood Pb levels.
Study
Study Population and
Methodological Details3
Blood Pb Analyzed (ug/dL) Outcome Analyzed Effect Estimate (95% Cl)
Hornung et al. 397 children (born 1979-1995) followed
(2009) from age 2 to 6 yr, Rochester, NY and
Cincinnati, OH.
Moderate follow-up participation,
representative of full cohorts. Regression
model adjusted for city, HOME score, birth
weight, maternal IQ and education, log
childhood average blood Pb level.
Geometric mean (5th-95th):
Early childhood age
2 yr: 8.9 (3.0-24)
Concurrent: 6.0 (1.9-18)
6.3% with higher age 6 yr
than age 2 yr blood Pb level
FSIQ
WPPSIandWISC-R
Age 6 yr
Difference in FSIQ at age 6 yr per 1 unit
increase in age 6:2 year blood Pb ratio:
-7.0 (-10,-4.0)
Estimates a 7-point lower FSIQ in children
with 50% higher blood Pb level at age 6
than 2 years (ratio = 1.5) compared to
children with a 50% lower blood Pb level
at age 6 than 2 years (ratio = 0.5)
Bellinger et al. 170 children followed prenatally to age 57
(1990) mo, Boston, MA area
High follow-up participation, no comparison
with nonparticipants. Log regression
adjusted for HOME score, SES, maternal
IQ and age, sex, ethnicity.
High prenatal (cord): >10
Concurrent
Low: <3
Medium: 3-10
High: >10
Change in z-score
McCarthy GCI at age
57 mo
MDI at 24 mo
High prenatal/Low concurrent
0.42 (-0.15, 0.99)
High prenatal/Medium concurrent
0.15 (-0.14, 0.44)
High prenatal/High concurrent
-0.15 (-0.56, 0.26)
Wasserman 390 children followed prenatally (1985-
et al. (2000a) 1986) to age 7 yr, K. Mitrovica, Pristine,
Yugoslavia
Moderate follow-up participation. More
participants were from K. Mitrovica,
Albanian, with lower maternal IQ and
education and HOME score. Regression
model adjusted for HOME score, maternal
age, education, IQ, and # live births, child
sex, birthweight, ethnicity, age.
Mean of average during
period
Prenatal: 10
Age 0-2: 13
Age 2-3: 19
Age 2-4: 19
Age 2-5: 19
Age 2-7: 17
FSIQ
McCarthy GCI, ages
3-4 yr
WPPSI, age Syr
WISC-III, age7yr
No change from prenatal to postnatal:
Reference
>50% increase prenatal to postnatal only
birth to age 2 yr: -1.8 (-3.5, -0.05)
>50% increase prenatal to postnatal from
birth to age 2 yr and from age 2 yr
onward: -2.7 (-4.9, -0.52)
4-258
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Table 4-15 (Continued): Comparisons of blood Pb-cognitive function associations in groups of children with different
temporal trends in blood Pb levels.
Study
Study Population and
Methodological Details3
Blood Pb Analyzed (ug/dL) Outcome Analyzed Effect Estimate (95% Cl)
long et al. 326 children followed from birth
(1998) (1979-1982) to age 11-13 yr, Port Pirie,
Australia
Moderate follow-up participation.
Participants had weaker early blood
Pb-cognitive function association. ANOVA
to compare mean change in FSIQ among
children with different degrees of change in
blood Pb across time.
Means
Early childhood age 2 yr: 21
Age 11-13 yr: 7.9
Mean change in
cognitive function z-
scores
between ages 2 and
11-13yr
Bayley MDI, age 2 yr
WISC-R, ages 11-13
yr
<10.2 ug/dL decline: 0.03 (-0.15, 0.21)
10.2-16.2 ug/dl_ decline: 0.04 (-0.15, 0.23)
>16.2 ug/dl_ decline: -0.01 (-0.20, 0.18)
p = 0.74 for difference among groups
Difference in age 7 FSIQ
Low age 2, Low age 7:
Reference
Low age 2, High age 7:
-3.6 (-6.4, -0.7)
High age 2, Low age 7:
0 (-2.8, 2.7)
High age 2, High age 7:
-3.7 (-6.2,-1.3)
Chen et al. 617 children participating in the TLC trial
(2005a) from age 12-33 mo to age 7 yr, Baltimore,
MD; Cincinnati, OH; Newark, NJ
77% African-American. 50% given
chelation at ages 12-33 mo, blood Pb
levels 20-44 ug/dL. High follow-up
participation. Regression model adjusted
for city, race, sex, age at blood Pb
measurement, language, single parent,
caregiver IQ, parental education and
employment. Considered potential
confounding by chelation but not parental
caregiving quality.
Mean (SD):
Age 2 yr: 26 (5.1)
Age Syr: 12(5.2)
Age 7 yr: 8.0 (4.0)
Low age 2 yr: <24.9
Low age 7 yr: <7.2
WISC-III
Age 7 yr
WPPSI = Wechsler Preschool and Primary Scale of Intelligence, WISC = Wechsler Intelligence Scale for Children, GCI = General Cognitive Index
aStudies are grouped by strength of study design, representativeness of the study population characteristics and blood Pb levels examined, and extent of consideration for potential
confounding. There is not necessarily a continuum of decreasing strength across studies.
4-259
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Combining the Cincinnati and Rochester cohorts (n = 397), Hornung et al. (2009)
analyzed the ratio of age 6 blood Pb level to age 2 year blood Pb level to assess the
association of FSIQ at age 6 years with a change in blood Pb level over time. Blood Pb
levels measured at various ages were highly correlated (r = 0.64-0.96), but investigators
found a range of temporal trends in blood Pb levels (range in age 6:2 year blood Pb level:
0.11-4.1, mean: 0.58). This range was analyzed in the model as the predictor. The model
estimated that for children with the same average (ages 6 months-6 years) blood Pb level,
those with a higher age 6:2 year blood Pb ratio had a lower FSIQ at age 6 years (Table
4-15). Although 6.3% of children had a ratio > 1, representing higher age 6 than age 2
blood Pb level, model results were not dependent on these data points. An example of
model results is illustrated in Figure 4-14. The three temporal blood Pb patterns have
similar areas under the curve, indicating similar childhood average blood Pb levels. Thus,
differences in FSIQ more likely would be attributable to differences in temporal exposure
pattern. Among the three blood Pb patterns illustrated, the model estimated the lowest
FSIQ for children with an age 6:2 year blood Pb ratio of 1.25 or children who had a 25%
increase in blood Pb level from 2 to 6 years of age (Figure 4-14).
18
16
14
12
6-year:2-year
ratio = 0.5
IQ = 89.0
6-year:2-year
ratio = 1.25
IQ = 83.7
3 4
Age (years)
Note: FSIQ was estimated for different patterns of blood Pb level changes overtime: 1) peak at age 2 years, i.e., age 6:2 year
ratio = 0.5 (blue diamonds); 2) peak at 5 years, i.e., age 6:2 year ratio = 1.25 (black triangles); and 3) constant blood Pb level, i.e.,
age 6:2 year ratio = 1 (white squares). The 3 patterns have a similar cumulative blood Pb level (10 ug/dL) as indicated the areas
under the curve. Among the 3 patterns, the lowest FSIQ at age 6 years was estimated for blood Pb level peaking at age 5 years.
Source: Hornung et al. (2009)
Figure 4-14 Estimated FSIQ for three different temporal patterns in blood Pb
level from ages 2 to 6 years in Rochester and Cincinnati cohorts.
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Results from both the Boston and Yugoslavia cohorts demonstrated cognitive function
decrements in children with increases in blood Pb level from the prenatal to postnatal
period (Wasserman et al., 2000a; Bellinger etal., 1990) (Table 4-15). In the Boston
cohort, children with high prenatal cord (> 10 (ig/dL) and high concurrent (> 10 (ig/dL)
blood Pb level had a decrease in cognitive function from age 24 to 57 months (Bellinger
et al.. 1990). In contrast, children with high prenatal blood Pb but low concurrent
(<3 (ig/dL) blood Pb level had an increase in cognitive function from age 24 to
57 months (Table 4-15). Findings from this study indicated that by age 5 years, children
with higher prenatal blood Pb levels appeared to recover the Pb-associated decrements in
cognitive function unless concurrent blood Pb levels remained high. The investigators
also demonstrated that optimal sociodemographic characteristics (e.g., higher HOME
score, SES, maternal IQ and age, female) also protected against decrements in cognitive
function associated with higher postnatal blood Pb levels. Similarly, environmental
enrichment with items such as toys, tunnels, and running wheels in cages given after
gestational-lactational Pb exposure was found to protect rats from deficits in spatial
learning in adulthood (Guilarte et al.. 2003). The Yugoslavia study examined in detail
during what periods of postnatal development were increases in blood Pb level associated
with FSIQ decrements. Children with at least a 50% increase in blood Pb level from the
prenatal period only up to age 2 year (early postnatal increase) and children with
increases to age 2 years and to 7 years (early and later postnatal increase) had lower FSIQ
compared with children with no increase in blood Pb level from the prenatal or postnatal
period (Wasserman et al., 2000a) (Table 4-15). The decrease in FSIQ was greater among
children with both early and later postnatal increases in blood Pb level. Collectively,
these results from the Boston and Yugoslavia cohorts suggest that cognitive development
is not fixed early in childhood and can be affected negatively or positively by postnatal
Pb exposures or other factors that influence cognitive function.
Chen et al. (2005a) found a stronger influence of higher concurrent than age 2 year blood
Pb levels on FSIQ at age 7 years among children participating in a multi-city chelation
trial. Children with higher concurrent blood Pb levels (> median 7.2 (ig/dL) had lower
FSIQ at age 7 years, regardless of whether blood Pb level at age 2 years was low or high
(less than or greater than the median of 24.9 (ig/dL, respectively). Blood Pb levels at ages
2 and 7 years were weakly correlated (r = 0.27). In another analysis of the same study
population (Liu et al.. 2002) plus analysis of another chelation trial, results directly
demonstrated increases in cognitive function in children with decreases in blood Pb levels
over follow-up of 6 to 36 months (Ruff et al., 1993). Liu et al. (2002) found such an
association in the group not receiving chelation. Because these children had initial high
blood Pb levels (20-55 (ig/dL) at ages 1 to 7 years, the findings may have limited
generalizability to the general population of children currently living in the U.S.
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Examining an older cohort (Port Pirie, Australia) at ages 11-13 years, Tong et al. (1998)
found some evidence of persistence of effects of higher early-life blood Pb level (Table
4-15). This conclusion was based on the analysis of groups of children with different
degrees of decline in blood Pb levels between ages 2 and 11-13 years. Although the mean
blood Pb level in the study population declined overall from 21.2 (ig/dL at age 2 years to
7.9 (ig/dL at age 11-13 years, the magnitude of decline varied among children. The
change in cognitive function between ages 2 and 11-13 years did not significantly differ
between children with the largest decline (>16 (ig/dL) in blood Pb level over that time
period and children with the smaller decline (<10 (ig/dL) (Table 4-15). Children with the
largest decline in blood Pb level between ages 7 and 11-13 years had a slightly smaller
decline in FSIQ over that time period. Thus, the findings indicated that cognitive function
assessed in adolescents may be influenced by higher early-life blood Pb levels but also
suggest some impact of more recent declines in blood Pb level. The results do not
preclude an independent association with concurrent blood Pb level.
To conclude, the collective body of epidemiologic evidence does not strongly identify an
individual critical lifestage or timing of Pb exposure with regard to neurodevelopmental
effects in children. Cognitive function decrements and increases in externalizing and
internalizing behaviors assessed at ages 3-24 years have been associated with prenatal,
early childhood (e.g., age 1 or 2 year), lifetime average (to age 4-5, 7, or 11-13 years),
and concurrent (at ages 3-17 years) blood Pb levels as well as with childhood tooth Pb
levels. The identification of critical lifestages and time periods of Pb exposure is
complicated further by the fact that blood Pb levels in older children, although affected
by recent exposure, are also influenced by Pb stored in bone due to rapid growth-related
bone turnover in children relative to adults. Thus, associations of neurodevelopmental
effects with concurrent blood Pb level in children may reflect the effects of past and/or
recent Pb exposures (Section 3.3.5.1). Evidence indicates that prenatal blood Pb levels
are associated with mental development in very young children
-------
(ig/dL) or larger declines overtime (i.e., 8, 14 (ig/dL) than those expected for most of the
current population of U.S. children. Nonetheless, these findings are consistent with the
understanding that the nervous system continues to develop throughout childhood. The
epidemiologic evidence for associations of neurodevelopmental effects with multiple
lifestages or time periods of Pb exposure, including more recent exposures, is supported
by evidence in monkeys for Pb exposures during multiple lifestages and time periods,
including infancy only, lifetime starting from birth, or lifetime starting during the juvenile
period, inducing impairments in cognitive function when assessed between ages 6 and 10
years (Rice. 1992b: Rice and Gilbert. 1990a: Rice. 1990; Rice and Gilbert. 1990b: Rice
and Karpinski. 1988).
4.3.12 Examination of the Pb Concentration-Response Relationship
With successive Pb AQCD and supplements, epidemiologic and toxicological studies
find that progressively lower blood Pb levels are associated with effects on cognitive
function and behavior. For example, among children, such effects were observed in
association with blood Pb levels in the range of 10-15 (ig/dL (for prenatal maternal or
cord) in the 1986 Addendum (U.S. EPA. 1986c) and 1990 Supplement to the 1986
Addendum (U.S. EPA. 1990a) and 10 (ig/dL and lower in the 2006 Pb AQCD (U.S. EPA.
2006b). Further, in the 2006 Pb AQCD, several individual studies and pooled analyses
estimated a supralinear concentration-response relationship in children, i.e., greater
decrements in cognitive function per unit increase in blood Pb level among children in
lower strata of blood Pb levels compared with children in higher strata of blood Pb level
(Figure 4-15 and Table 4-16). Some evidence in animals also indicates nonlinear
concentration-response relationships for various endpoints, including those related to
learning impairments. Some toxicological evidence points to larger absolute effects in
lower Pb exposure groups (relative to control groups) than in higher exposure groups.
However, these toxicological findings of U- or inverted U-shaped relationships are
distinct from epidemiologic findings of supralinear relationships in that some
toxicological relationships do not indicate Pb-induced impairments at higher exposure
concentrations.
Several prospective epidemiologic studies found supralinear concentration-response
relationships for concurrent, early childhood (age 2 year), and lifetime average (to age 5
years) blood Pb levels and cognitive function decrements in children ages 2-16 years.
Most studies used a blood Pb level of 10 (ig/dL to define lower and higher strata and
reported mean or median blood Pb levels in the lower strata of 3-7 (ig/dL (Bellinger.
2008; Canfield. 2008; Hornung. 2008; Tellez-Roio. 2008). Except for the pooled
analysis, the lower strata of blood Pb levels comprised >30% of the study population,
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indicating that the blood Pb-cognitive function relationships calculated for the lower
strata likely are not outliers or unrepresentative of the overall study population. Studies in
the Boston and Rochester cohorts each examined different ages of blood Pb and FSIQ but
lower blood Pb levels compared to other studies (means ~6 (ig/dL), and a greater FSIQ
decrement per unit increase in concurrent blood Pb level was found for children whose
peak blood Pb levels < 10 (ig/dL compared with all children examined (Bellinger and
Needleman. 2003; Canfield et al. 2003a: Bellinger et al.. 1992). A nonlinear blood Pb-
FSIQ relationship also was indicated in pooled analyses of seven prospective studies
involving a wider range of blood Pb levels (5th-95th percentiles 2.5-33.2 (ig/dL) by
statistical significance of blood Pb level modeled as quadratic and cubic functions
(Lanphear et al.. 2005) or better fit for a log-linear than linear model (Rothenberg and
Rothenberg. 2005).
.. . Blood Pb Metric
Study Analyzed
Prospective
Bellmgerand Needleman A 0
(2003) Age^yr
Canfield etal. (2003) Concurrent
Lanphearetal. (2005) Concurrent
Cross-sectional
Tellez-Rojoetal. (2006) Concurrent
Lanphearetal. (2000) Concurrent
Kordas etal. (2006) Concurrent
Outcome t . ,..
FSIQ All subjects
Peak<10a*
FSIQ All subjects
Peak< 10=
FSIQ Peak>10°
Peak<10"
Peak>7.5a
Peak<7.5°
BayleyMDI >10b
<10b
<5b
Math score All subjects
<7.5
<5
<2.5
Math score All subjects
Sample size
1,203 -»-
4,853 *
532 -«-
Change in cognitive function test score perl ug/dL
increase in blood Pb level (95% Cl)
aStrata were defined by the peak blood Pb level measured in child at any point during follow up. b95% Cl estimated from reported
p-value.
Note: Results are presented first for prospective analyses then for cross-sectional analyses. FSIQ = full-scale IQ, MDI = mental
development index. Effect estimates (concentration-response) are presented for a 1 ug/dL increase in blood Pb level. Black symbols
represent effect estimates among all subjects or subjects in the higher blood Pb stratum. Blue symbols represent effect estimates in
lower blood Pb strata.
Figure 4-15 Comparison of associations between blood Pb level and cognitive
function among various blood Pb strata.
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Table 4-1 6
Study
Bellinger et al.
(1992)
Bellinger and
Needleman
(2003)
Bellinger (2008)
1 1 G PPA
U.O. CrM
(2006b)
Canfield et al.
(2003a)
Canfield (2008)
Jusko et al.
(2008)
Additional characteristics and quantitative results for studies presented in Figure 4-15.
Study Population and
Methodological Details3
148 children followed from birth
(1979-1 981) to age 10 yr, Boston area,
MA.
Prospective. Recruitment at birth hospital.
Participation by 59% of original cohort but
88% from age 5 yr. Participants had higher
SES, HOME score. Linear regression
model adjusted for HOME score (age 10
and 5), child stress events, race, maternal
\Q, age, marital status, SES, sex, birth
order, # residence changes before age 5 yr
171 children born 1994-1995 followed from
age 6 mo to 5 yr, Rochester, NY.
Prospective. Recruitment from study of
dust control. 73% nonwhite. Moderate
follow-up participation but no selective
attrition. Mixed effects models adjusted for
child sex, Fe status, and birth weight,
maternal race, education, \Q, income and
prenatal smoking, HOME score
174 children born 1994-1995 followed from
age 6 mo to 6 yr, Rochester, NY.
Prospective. Same cohort as above. High
Blood Pb Metric
Analyzed (ug/dL)
Early childhood
(age 2 yr)
Mean (SD)
All subjects:
6.5(4.9)
Peak <10:
•3 K
o.o
(range: 1-9.3)
Detection limit not
reported
Concurrent
Mean (SD)
All subjects:
5.8(4.1)
Da'alf
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Table 4-16 (Continued): Additional characteristics and quantitative results for studies presented in Figure 4-15.
Study
Lanphear et al.
(2005)
Hornung (2008)
Tellez-Rojo et
al. (2006)
Tellez-Rojo
(2008)
Lanphear et al.
(2000)
Study Population and
Methodological Details3
1,333 children pooled from Boston,
Cincinnati, Cleveland, Mexico City,
Port Pirie, Rochester, and Yugoslavia
cohorts.
Prospective. Large, uniform analysis
pooling diverse cohorts. Included 84% of
eligible children. Linear regression model
adjusted for HOME score, birth weight,
maternal IQ and education. Considered
child sex, marital status, birth order,
prenatal alcohol consumption and smoking
status, maternal age.
384 children followed from birth
(1994-1995, 1997-1 999) to age 2 yr,
Mexico City, Mexico.
Cross-sectional. Recruitment from
prenatal clinic or birth hospital. Participants
had higher maternal education, lower cord
blood Pb. Linear regression model
adjusted for sex, birth weight, maternal IQ.
Considered maternal age and other
unspecified factors.
4,853 children ages 6-16 yr, NHANES III
1988-1994.
Cross-sectional. Large study of multiple
exposures and outcomes. Linear
regression model adjusted for U.S. region,
sex, race/ethnicity, reference adult
education, poverty index ratio, marital
status, serum ferritin level, serum cotinine
level. Did not consider parental IQ or
caregiving quality.
Blood Pb Metric
Analyzed (ug/dL) Outcome
Concurrent FSIQ
Mean (range) Various tests
Peak > 10: Ages4.8-10yr
13.9(0.1-71.7)
Doalf 7.5:
12.9(0.1-71.7)
Peak <7.5:
3.2 (0.9-7.4)
Concurrent Bayley MDI
Mean (SD) Age 2 yr
>10: NR
<10:
4.3(2.3)
c in-
*J 1 \J m
6n / *i A \
.» (1.4)
<5:
2.9(1.1)
Concurrent Math Score
Geometric mean (SD) WRAT
All subjects: Ages 6-16 yr
1.9(7.0)
Subgroups: NR
Blood Pb stratum
(ug/dL)
1,089 subjects peak > 10
244 subjects peak <10
1,203 subjects peak > 7.5
103 subjects peak<7.5
90 subjects > 10
294 subjects <10
101 subjects 5-10
193 subjects <5
All 4,853 subjects
4,681 subjects <10
4,526 subjects <7.5
4,043 subjects <5
2,467 subjects <2.5
Effect Estimate
(95% Cl)b
-0.13 (-2.3, -0.03)
-0.80 (-1.7, 0.14)
-0.16 (-0.24, -0.08)
-2.9 (-5.2, -0.71)
0.07 (-0.61, 0.75)c
-1.0 (-1.8, -0.25)c
-0.94 (-2.1, 0.23)c
-1.7 (-3.0, -0.42)c
-0.70 (-1.0, -0.37)
-0.89 (-1.5, -0.26)
-1.1 (-1.8, -0.30)
-1.1 (-2.0, -0.12)
-1.3 (-3.2, 0.64)
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Table 4-16 (Continued): Additional characteristics and quantitative results for studies presented in Figure 4-15.
Study
Kordas et al.
(2006)
Study Population and
Methodological Details3
532 children ages 6-8 yr, Torreon, Mexico
Cross-sectional. Recruitment at schools.
Residence near metal foundry. High
participation. Linear regression model
adjusted for sex, age, hemoglobin, family
possessions, forgetting homework, house
ownership, crowding, maternal education,
birth order, family structure, urine arsenic,
tester, school. Did not consider parental IQ
or caregiving quality.
Blood Pb Metric
Analyzed (ug/dL)
Concurrent
Mean (SD)
All subjects: 11.4
(6.1)
<10: NR
Outcome
Math score
Achievement
test
Ages 6-8 yr
Blood Pb stratum
(ug/dL)
All 532 subjects
293 subjects <10
Effect Estimate
(95% Cl)b
-0.1 7 (-0.28,
-0.42 (-0.92,
-0.06)
0.08)
FSIQ = Full-scale IQ, WISC-R = Wechsler Intelligence Scale for Children-Revised, WPPSI-R = Wechsler Preschool and Primary Scale of Intelligence-Revised, MDI = Mental
Developmental Index, NR = Not reported, WRAT = Wide Range Achievement Test.
aStudies are grouped by strength of study design, representativeness of the study population characteristics and blood Pb levels examined, and extent of consideration for potential
confounding. There is not necessarily a continuum of decreasing strength across studies.
bMost effect estimates are derived from linear models and are presented for a 1 ug/dL increase in blood Pb level.
°Results not included in Figure 4-15 because nonparametric analysis did not produce 95% CIs for various strata of blood Pb levels.
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A few cross-sectional studies demonstrated larger Pb-associated decreases in cognitive
function in groups with concurrent blood Pb levels <5 (ig/dL than groups with higher
levels. Tellez-Rojo et al. (2006) estimated a larger decrement in age 2 year Bayley MDI
per unit increase in blood Pb level for children with concurrent blood Pb levels <5 (ig/dL
compared with children with blood Pb levels 5-10 (ig/dL, and >10 (ig/dL (Figure 4-15
and Table 4-16). However, it is not clear what the implications of age 2 year MDI results
may be on cognitive function at later ages. Among children ages 6-16 years participating
in NHANES 1988-1994, Lanphear et al. (2000) found larger decrements in reading and
math skills and memory per unit increase in blood Pb level in children with concurrent
blood Pb levels <2.5 (ig/dL than children with levels <5 (ig/dL, <7.5 (ig/dL, <10 (ig/dL,
and all subjects. However, these children born 1972-1989 likely had higher blood Pb
levels (Figure 3-18) and Pb exposures earlier in childhood, which may have contributed
to the observed associations.
Several (Min et al.. 2009: Jusko et al.. 2008: Schnaas et al.. 2006) but not all
(Palaniappan et al., 2011) recent studies found evidence for a nonlinear blood
Pb-cognitive function relationship in nonparametric regression analyses using splines or
LOWES S with smoothing parameters that did not produce quantitative results for
individual blood Pb groups. Similar to the pooled analyses of the seven prospective
cohorts, these relationship were evaluated for a wide range of blood Pb levels. In the
Rochester and Mexico City cohorts, the blood Pb-FSIQ relationship was more negative
for children with lower blood Pb levels, specifically for peak blood Pb levels 2.1-
10 (ig/dL (overall range: 2.1-46 (ig/dL) in the Rochester cohort at age 6 years (Jusko et
al., 2008) and for prenatal maternal week 28-36 blood Pb levels <6 (ig/dL (overall
5th-95th percentile 2.5-25 ng/dL) in the Mexico City cohort at ages 6-10 years (Schnaas
et al., 2006). In a formal test of nonlinearity, Schnaas et al. (2006) found the nonlinear
blood Pb term to fit the data better than a linear term. Among 278 children of age 4 years
(blood Pb range: 1.3-24 (ig/dL) who had high prenatal alcohol and drug exposure, Min et
al. (2009) reported a p-value of 0.19 for a restricted cubic spline term for blood Pb level
and described the covariate-adjusted concurrent blood Pb level-FSIQ curve to be more
negative at blood Pb levels 1.3-7 (ig/dL. Among 814 children in India ages 3-7 years,
Palaniappan et al. (2011) mostly found linear associations between concurrent blood Pb
level (range: 2.6-41 (ig/dL) and indices of cognitive function. The exception was visual-
motor skills, for which a greater blood Pb-associated decline was found with blood Pb
levels >30 (ig/dL. The linearity versus nonlinearity of the blood Pb-FSIQ concentration-
response relationship within a lower, more narrow range of blood Pb levels has not been
examined in detail.
Few studies of adults have examined whether the relationship between blood or bone Pb
level and cognitive function is described better with a linear or nonlinear function. In
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analyses of adults in NHANES, only log-linear models were used to fit the data (Krieg et
al..201Q: Krieg and Butler. 2009: Krieg et al.. 2009). Nonlinearity in the BMS and NAS
cohorts was examined with the use of quadratic terms, penalized splines, or visual
inspection of bivariate plots (Bandeen-Roche et al.. 2009; Wang et al.. 2007a: Weisskopf
et al., 2007b; Shih et al.. 2006). There was some evidence for nonlinearity in prospective
analyses of the NAS cohort (Figure 4-7 and Figure 4-8). but not all results indicated
greater declines in cognitive function per unit increase in bone Pb level in the lower bone
Pb groups. Wang et al. (2007a) found that among NAS men with an HFE gene variant,
there was a larger decline in MMSE score per unit increase in tibia Pb level at higher
tibia Pb levels (Figure 4-8). In the BMS cohort, linear relationships were indicated for
various tests of cognitive function by a statistically nonsignificant quadratic term (Shih et
al., 2006) or spline (Bandeen-Roche et al., 2009) for tibia Pb level.
Although not specific to Pb exposure, attenuation of concentration-response relationships
at higher exposure or dose levels has been reported in the occupational literature, and
explanations have included greater exposure measurement error and saturation of
biological mechanisms at higher levels and larger proportions of at-risk populations at
lower exposure levels (Stavner et al.. 2003). Hypotheses for nonlinearity in the
relationship between Pb and cognitive function include a lower incremental effect of Pb
due to covarying risk factors such as low SES, poor caregiving environment, higher
exposure to other environmental risk factors (Schwartz. 1994). different mechanisms
operating at different exposure levels, and confounding by omitted or misspecified
variables. The contribution of these factors to the supralinear relationship observed
between blood Pb levels and cognitive function in children has not been examined in
many epidemiologic studies to date. Some studies found that risk factors such as SES,
parental education, and parental caregiving quality explain a greater proportion of
variance in cognitive function than does blood Pb level (Wasserman et al.. 1997; Greene
et al.. 1992). Recently, among 57,678 fourth grade children across North Carolina,
Miranda et al. (2009) found that lower parental education and enrollment in a
free/reduced fee lunch program accounted for larger decrements in EOG scores than did
blood Pb level across the various quantiles of EOG score distribution (Figure 4-6). In a
meta-analysis, the blood Pb-FSIQ relationship was more negative in studies of lower SES
populations than in studies of higher SES populations (Schwartz. 1994).
Few studies have examined effect modification of the blood Pb level-cognitive function
relationship by covarying risk factors such as sociodemographic factors, and the limited
evidence is inconclusive. None of these studies examined effect modification within
specific strata of blood Pb levels. Among children in the Boston cohort in the higher cord
blood Pb level group (>10 (ig/dL), cognitive function between ages 24 and 57 months
increased in the group that was female and had higher HOME score, SES, and maternal
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IQ (Bellinger et al., 1990). In the Port Pirie, Australia cohort, larger blood Pb-associated
FSIQ decrements were found in groups with lower SES at age 11-13 years (Tong et al..
2000) but not lower HOME score at age 4 years (McMichael et al., 1992). Similar effects
of blood Pb level on FSIQ were estimated in the higher SES Boston cohort and the lower
SES Rochester cohort (Table 4-3). Overall, evidence does not clearly indicate whether
the blood Pb-cognitive function relationship is modified by factors such as SES or other
sociodemographic characteristics or whether these differences can explain the observed
nonlinear concentration-response relationship.
Results from the pooled analysis by Rothenberg and Rothenberg (2005) do not indicate
that residual confounding by covariates explains the nonlinear blood Pb-FSIQ
relationship. Modeling maternal IQ, HOME score, maternal education, and birthweight as
spline functions (df = 2) did not significantly improve model fit either with a linear blood
Pb term or log blood Pb term, which indicated that the improved model fit with log-
specification of blood Pb level was not influenced by the modeling of covariates as linear
or nonlinear functions.
Bowers and Beck (2006) postulated that a supralinear concentration-response function
necessarily will be found in a model with a log-normally distributed independent variable
and a normally distributed outcome variable. However, as discussed in the
2006 Pb AQCD, this modeling strategy was not employed in the epidemiologic analyses
showing a supralinear concentration-response function. FSIQ scores generally were not
forced into a normal distribution. Normalized FSIQ scores were not the basis for
individual findings from four of the seven studies included in the pooled analysis by
Lanphear et al. (2005) or the results pooling the seven cohorts (Hornung et al.. 2006).
Further, a log-linear model (a linear relationship between IQ and the log of blood Pb)
provided a better fit of the pooled data.
Results from prospective analyses in the Boston and Rochester cohorts for associations
between blood Pb level as a continuous variable and FSIQ in groups of children in the
lower segment of the population blood Pb distribution have not identified a threshold in
the range of blood Pb levels examined. In the Boston cohort, higher age 2 year blood Pb
levels were associated with lower FSIQ at age 10 years in children with blood Pb levels
1-9.3 (ig/dL whose peak blood Pb levels were less than 10 (ig/dL (Bellinger. 2008;
Bellinger and Needleman. 2003). Schwartz (1994) explicitly assessed evidence for a
threshold in the Boston cohort data by regressing FSIQ and blood Pb level on covariates
such as age, race, maternal IQ, SES, and HOME score and fitting a nonparametric
smoothed curve to the residuals of each regression model (variation in FSIQ or blood Pb
level not explained by covariates). A 7-point decrease in FSIQ was found over the range
of blood Pb residuals below 0 (corresponding to the mean blood Pb level of 6.5 (ig/dL),
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indicating an association between blood Pb level and FSIQ well below the mean blood Pb
level. Results from the Rochester cohort did not identify a threshold either. Higher
concurrent blood Pb levels (range 0.5-8.4 (ig/dL, detection limit = 1 (ig/dL) were
associated with lower FSIQ at ages 3 and 5 years in children with peak blood Pb levels
<10 (ig/dL (Canfield et al.. 2003a). A threshold also was not identified for the association
between concurrent blood Pb level and MDI score at age 2 years among children in
Mexico City (Tellez-Rojo et al.. 2006). Higher blood Pb level was associated with a
lower MDI score in the group of children with blood Pb levels <5 (ig/dL (range
0.8-4.9 (ig/dL) (Tellez-Rojo. 2008). Other results based on nonparametric regression
analyses did not identify a threshold for Pb-associated cognitive function decrements in
children, but results have weaker implications because the blood Pb levels in the
examined populations were higher (e.g., minimum 2.1 (ig/dL, 5th percentiles 2.5 (ig/dL)
(Jusko et al.. 2008; Schnaas et al.. 2006; Lanphear et al.. 2005) than those in most current
U.S. children.
In conjunction with downward trends in population blood Pb distributions (Figure 3-16).
more sensitive quantification methods have improved measurement of blood Pb levels in
the lower portion of the population distribution (e.g., in NHANES, from 0.6 (ig/dL in
1999-2002 to 0.025 (ig/dL in 2003-2004). Consequently, the examination of groups of
children (ages 4-15 years) with lower blood Pb levels, in the range of <1 to 16 (ig/dL, has
indicated Pb-associated cognitive function decrements or increases in externalizing
behaviors at lower blood Pb levels (Cho et al.. 2010; Kim et al.. 2009b; Miranda et al..
2009: Braun et al.. 2008: Miranda et al.. 2007a: Braun et al.. 2006). In the studies
examining concurrent blood Pb levels, the potential contribution of higher past Pb
exposures obscures assessment of a threshold. For example, Braun et al. (2008: 2006)
found higher odds of parental reports of conduct disorder and ADHD among children
ages 4-15 years with concurrent blood Pb levels ~1.0 (ig/dL compared with children with
blood Pb levels <0.8 (ig/dL (6 or 7% below detection limit of 0.3 (ig/dL). However,
children likely had higher early childhood blood Pb levels, and adolescents in the study
were born in the 1970s during the use of leaded gasoline.
Results from Miranda et al. (2009) did not identify a threshold with the examination of
blood Pb levels measured between ages 6 and 36 months during 1995-1999. Among
57,678 children in North Carolina, lower 4th grade EOG scores were found in children in
the blood Pb level category of 2 (ig/dL (exact values not reported) compared with
children in the blood Pb level category of 1 (ig/dL, which included children with blood
Pb levels at or below the detection limit. Other analyses of blood Pb level as a categorical
variable did not clearly address the identification of a threshold for the blood
Pb-cognitive function relationship. However, these analyses of categories that encompass
a wider range of blood Pb levels may be less sensitive. Cognitive function decrements
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were limited to children ages 7-8 years in the U.K. with age 30-month blood Pb levels 5-
10 (ig/dL (compared with blood Pb levels 0-2 (ig/dL) (Chandramouli et al.. 2009) and
children ages 6-10 years in New England with concurrent blood Pb levels 5-10 (ig/dL
(compared with blood Pb levels 1-2 (ig/dL) (Surkan et al.. 2007).
Some toxicological studies found nonlinear relationships between Pb exposure and
effects related to impaired learning and memory in animals, although these differ from
the nonlinear relationships found in epidemiologic studies of children in that nonlinearity
in animal studies has been demonstrated as lower and higher Pb exposures having effects
in opposite directions (U- or inverse U-shaped curves). Epidemiologic results for
nonlinearity indicate negative effects over the entire blood Pb distribution examined.
Results summarized across multiple studies in multiple species demonstrated that lower
Pb exposures increased FI response rates relative to controls, and higher Pb exposures
decreased FI response rates (Cory-Slechta. 1994). Increased FI response rates indicate
impaired learning by reflecting the impaired ability of animals to respond according to a
fixed schedule of reinforcement (Section 4.3.2.3). Consistent with previous findings,
Rossi-George et al. (2011) found that 50 ppm lifetime Pb exposure increased FI overall
rate in rats tested at age 2 months whereas 150 ppm Pb exposure did not affect FI
responses. Nonlinear effects of Pb on learning were not observed with longer duration
exposures (e.g., 8-11 months) (Cory-Slechta. 1990). The nonlinear effects of Pb on
impaired learning were supported by evidence in animals indicating that lower and higher
Pb exposures differentially activate underlying mechanisms. Gilbert et al. (1999) found
reduced LTP magnitude in adult rats exposed to 1,000-5,000 ppm but not 10,000 ppm
Pb acetate in drinking water from GDI6. LTP is one indication of synaptic plasticity
(Section 4.3.10.4). which is considered to be a major cellular mechanism underlying
learning and memory. Synaptic plasticity also is affected by glutamatergic
neurotransmission via its NMDA receptor (Section 4.3.10.8). and reduced glutamate
release in the hippocampus was found in adult rats exposed to Pb acetate from
GDI5-GDI6 with blood Pb levels 27-40 (ig/dL but not with blood Pb levels of
62-117 (ig/dL (Laslev and Gilbert. 2002).
Dopaminergic neurotransmission is involved in many CNS processes including
cognition, behavior, and motor function. The shape of the Pb-DA concentration-response
relationship varied among rodent studies. Some studies found that lower Pb exposures
(e.g., 27, 50 ppm) did not affect or increased DA activity relative to controls and higher
Pb exposure (109, 150, 250 ppm) (Leasure et al.. 2008; Virgolini et al.. 2005; Lewis and
Pitts. 2004). However, sometimes in the same study, in other regions of the brain or
under some conditions, higher Pb exposures (109, 150 ppm) increased or impaired DA
activity (Leasure et al.. 2008; Virgolini et al.. 2005). These differential responses of DA
may be related to the diverse CNS effects of DA in different regions of the brain. For
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example, the increased forebrain dopamine turnover with 27 ppm gestational/lactational
Pb acetate exposure was accompanied by less spontaneous activity in male mice
compared with male mice exposed to 109 ppm Pb (Leasure et al., 2008).
In vitro results indicated differential effects on calcineurin enzyme activity, with inhibited
activity resulting from higher Pb exposure (>2* 10"4 (iM) and stimulated activity from
lower Pb exposure (Kern and Audesirk. 2000). While calcineurin activity has been found
to modulate learning, LTP, and behavior in animals, lower calcineurin activity has been
associated with both impaired and improved performance on learning and memory tasks
(Zeng et al.. 2001). Thus, it is uncertain whether altered calcineurin activity can explain
the nonlinear relationships observed between Pb exposure and impaired learning in
animals. At lower concentrations, Pb may displace calcium at its binding sites on
calmodulin and by acting as a calmodulin agonist at the catalytic A subunit of calcineurin
stimulate calcineurin activity. At higher Pb exposure, Pb may bind directly to a separate
calcium-binding B subunit, override the calmodulin-dependent effect and turn off the
activity of calcineurin. Lasley and Gilbert (2002) found that 2,000 ppm but not 5,000 or
10,000 ppm Pb acetate exposure of rats (in drinking water starting at GD15-GD16)
inhibited glutamate release by acting as a calcium mimetic.
Some toxicological studies have found nonlinear relationships for non-cognitive
outcomes in animals. U-shaped Pb concentration-response relationships were found for
spontaneous motor activity level and latency to fall from rotarod (Leasure et al., 2008).
Inverted U-shaped relationships were found for hippocampal neurogenesis related to Pb
exposure duration (24 hours not 28 days after BrdU injection) or concentration (Fox et
al.. 2008; Gilbert et al.. 2005). Evidence also points to differences in hormone production
by Pb exposure concentration. In male mice with long-term Pb exposure
(PND21-9 months of age), basal corticosterone levels were significantly lower with
50 ppm Pb than with 150 ppm Pb or controls (Virgolini et al., 2005). The visual system in
animals also has shown to be affected differentially by lower versus higher Pb exposure
(GD1-PND10, pup blood Pb levels 10-12, 24-27, and 42-46 jig/dL). Inverted U-shaped
concentration-response curves were observed for rod photoreceptor numbers in the retina
(Giddabasappa et al.. 2011) and retinal layer thickness (Foxetal.. 2010). These
dichotomous histological findings may give insight to the complex Pb-associated changes
in ERG wave amplitudes that vary by exposure timing and dose (Section 4.3.6.2).
To conclude, several studies found a supralinear blood Pb-cognitive function
concentration-response relationship in children but not adults based on comparisons of
effect estimates in lower and higher strata of blood Pb level and results from
nonparametric regression modeling. Explanations for this supralinear relationship have
not been well characterized by epidemiologic studies. Evidence from the prospective
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studies in the Boston and Rochester cohorts has not identified a threshold for
Pb-associated cognitive function decrements in the range of blood Pb levels examined.
Increases in childhood blood Pb levels ranging from those at or below the detection limit
of 1.0 ug/dL to 9.3 ug/dL (means: 3.3 and 3.8 ug/dL) were associated with cognitive
function decrements at ages 3 to 10 years in children whose peak blood Pb levels were
less than 10 ug/dL (Bellinger. 2008: Canfield. 2008: Bellinger and Needleman. 2003:
Canfield et al., 2003a). Further, a recent study found an association between higher blood
Pb levels measured at ages 6-36 months (1995-1999) and lower 4th grade EOG scores in
57,678 children in North Carolina with blood Pb levels at or below the detection limit of
1 ug/dL to 16 ug/dL (Miranda et al., 2009). Concurrent blood Pb levels in the range of
0.8-9.8 ug/dL (detection limit not reported) were associated with MDI decrements in
children age 2 years in Mexico City (Tellez-Rojo. 2008: Tellez-Rojo et al., 2006). The
lack of a reference population with blood Pb levels reflecting pre-industrial Pb exposures
limits the ability to identify a threshold. Analysis of ancient bones in pre-industrialized
societies suggests that "background" blood Pb levels in pre-industrial humans may have
been more than 50-fold lower than those of the current U.S. population (Flegal and
Smith. 1992) and lower than the levels at which neurodevelopmental effects have been
examined. Thus, the current evidence does not preclude the possibility of a threshold for
neurodevelopmental effects in children existing with lower blood levels than those
currently examined. While distinct from supralinear relationships observed in
epidemiologic studies, some toxicological studies showed that lower Pb exposures (e.g.,
50 ppm in drinking water) induced learning and memory impairments in animals
compared to control exposures or higher Pb exposures (e.g., 150 ppm). Additional
toxicological evidence suggests that differentially activated mechanisms at lower and
higher Pb exposures such as reduced LTP and hippocampal glutamate release with lower
Pb exposures may provide explanation for the impaired learning and memory observed
with lower Pb exposures in animals.
4.3.13 Confounding in Epidemiologic Studies of Nervous System Effects
In addition to Pb exposure, many factors influence cognitive function and behavior in
children, including parental IQ and education, SES of the family, quality of the
caregiving environment, and other environmental exposures (Wasserman and Factor-
Litvak, 2001). These other risk factors often are correlated with blood, tooth, and bone Pb
levels, thus, a major challenge to observational studies examining associations of Pb
biomarker levels with cognitive function and behaviors in children has been the
assessment and control for potential confounding factors. By definition, a confounder is
associated with both the independent variable and the outcome and consequently has the
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potential to bias the association between the independent variable of interest and the
outcome. Most epidemiologic studies of Pb biomarkers in children have examined
potential confounding by parental IQ and SES-related variables such as parental
education, household income, and the Hollingshead Four-Factor Index of Social Position,
which incorporates education and income of both parents. Fewer but still several studies
have examined confounding by quality of the caregiving environment (i.e., HOME
score), birth weight, and smoking exposure. A relatively smaller number of studies have
considered nutritional status, other environmental exposures, parental substance abuse, or
parental psychopathology. Studies have varied with respect to the number of potential
confounding factors examined, with some studies considering multiple SES-related
factors and other studies focusing on a smaller set. The extent of confounding by a
particular factor likely varies across studies, depending on the population examined.
Thus, the impact of adjustment for specific covariates on the Pb effect estimate also
likely varies across studies.
Various methods have been used to control for potential confounding: examining a
population relatively homogeneous in SES, examining populations in which factors are
not correlated, conducting multivariate regression, characterizing the change in the Pb
biomarker effect estimate with adjustment for a covariate, and examining associations in
different strata of a covariate. The evidence derived from each of these control strategies
is discussed below. No single method is without limitation, and adjustment for SES is
difficult as it is highly correlated with Pb exposure and there is no single variable that
fully represents SES. Residual confounding also is likely by factors not considered. The
combination of evidence from prospective studies that considered several well-
characterized potential confounding factors plus evidence that Pb exposure impairs
cognitive function and behavior in animals, in particular, visual-spatial memory,
executive function, and response inhibition, which are also found to be affected in
children, increase confidence that the associations observed between Pb biomarker levels
and neurodevelopmental effects in children represent a relationship with Pb exposure.
In the Boston prospective study, potential confounding by SES was reduced by study
design and statistical adjustment for SES. The study subjects were from middle- to upper-
middle-class families, a majority with married, college-educated parents. Hence, the
potential for confounding by SES in this study was considerably less compared to other
studies examining similar outcomes. In this cohort, higher prenatal cord and concurrent
blood Pb levels were associated with FSIQ decrements at age 57 months, and higher age
2 year and age 57 month blood Pb levels were associated decrements in FSIQ and
executive function at age 10 years (Stiles and Bellinger. 1993; Bellinger et al., 1992;
Bellinger et al.. 1990). In contrast, blood Pb levels measured at various ages were weakly
associated with cognitive function decrements at age 7 years in the Sydney, Australia
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cohort of middle-SES children (i.e., 20% mothers with > high school education) (Cooney
et al.. 1991). However, a relationship between Pb exposure and cognitive function
decrements is supported by a similar magnitude of blood Pb-associated FSIQ decrement
found in the Boston and the low-SES Rochester cohort (majority of mothers with <
college education and annual income <$15,000), with adjustment for similar covariates
(Figure 4-2 and Table 4-3) (Canfield et al.. 2003a). For some outcomes, larger effects
were estimated in the Boston cohort than in other cohorts.
Blood Pb levels also were associated with cognitive function decrements in populations
in which higher blood Pb levels were not correlated with lower SES (Factor-Litvak et al..
1999; Bellinger et al.. 1987). In the Yugoslavia cohort, blood Pb levels at age 4 years
were higher in groups with higher maternal education, maternal IQ, and HOME score in
the city near Pb sources and were lower in the distant city. Among all children, higher
blood Pb level was associated with lower FSIQ and learning and memory scores and with
higher ratings of internalizing behaviors (Factor-Litvak et al.. 1999). In the Boston
cohort, parental education, social class, HOME score, and birth outcomes were similar
among low (<3 (ig/dL), medium (6-7 (ig/dL), and high (> 10 (ig/dL) cord blood Pb level
groups. Further, adjusting for these and other demographic factors, Bellinger et al. (1987)
found that children in the high cord blood Pb group had a 4.8-point lower Bayley MDI
score at age 2 years than did children in the low cord blood Pb group.
The primary method used by epidemiologic studies to control for potential confounding,
in particular recent studies of children with blood Pb levels more similar to current U.S.
levels, has been multivariate regression. Some studies modeled a set of covariates based
on a priori evidence, whereas others selected specific covariates based on their
association with the outcome in a model with all potential covariates and/or a greater than
10% change in the blood Pb level effect estimate. Studies also varied in the number of
potential confounding factors included in models. Some included multiple SES-related
variables, whereas others analyzed one or two factors. Regardless of the method used to
select model covariates or the number of covariates included, studies consistently found
associations of higher blood Pb level with cognitive function decrements and poorer
behavior. The evidence suggests that confounding by particular factors may vary across
populations and increases confidence that the associations observed with
neurodevelopmental effects in children represent a relationship with Pb exposure.
The consistency in finding associations across populations with different SES and
co-exposures and across studies examining different covariates was reinforced in pooled
and meta-analyses (Marcus etal. 2010; Lanphear et al.. 2005; Schwartz. 1994). Pooling
data from seven international prospective cohorts, Lanphear et al. (2005) found similar
FSIQ decrements per log increase in blood Pb level (-2.6% to +8.6% difference) by
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excluding one study at a time. These results indicated a relatively robust pooled estimate
despite between-study differences in population characteristics, including SES. In a meta-
analysis, Schwartz (1994) found a relatively narrow range of blood Pb-FSIQ effect
estimates among studies despite large between-study differences in the correlation
between blood Pb level and SES. A wider range of effect estimates would be expected if
omitted SES factors confounded the association. A recent meta-analysis of the
association between blood Pb level and conduct problems from earlier and recent studies
of children found that adjustment for SES and home environment did little to attenuate
the association (Marcus etal. 2010).
Among the several studies that provided both unadjusted and adjusted effect estimates,
most indicated that blood Pb level was a statistically significant predictor of cognitive
function (e.g., FSIQ, executive function, learning, memory) in children ages 5-10 years
before and after adjusting for potential confounders. Although most effect estimates
changed by 20-50% in multivariate models, most remained within the 95% CI of the
unadjusted estimate (Min et al., 2009; Kordas et al., 2006; Schnaas et al., 2006; Canfield
et al.. 2003a; Dietrich et al.. 1993b; Bellinger et al. 1992). Such observations were made
in previous analyses of the Boston and Rochester cohort with mean blood Pb levels 6.5
and 5.8 (ig/dL, respectively, with adjustment for SES, maternal IQ and education, and
HOME score (Canfield et al., 2003a; Bellinger et al.. 1992). These analyses also adjusted
for or considered potential confounding by indicators of nutritional status. Recent studies
of children of a similar age range and mean blood Pb levels also found statistically
significant associations (as indicated by correlation and/or regression coefficients)
between blood Pb level and cognitive function before and after adjustment for similar
covariates (Min et al.. 2009; Chiodo et al.. 2007); however, because of the high
prevalence of prenatal alcohol or drug use in these populations, their results may be less
generalizable than results from the Rochester and Boston cohorts.
Blood Pb level also was a statistically significant predictor of cognitive function after
adjustment for covariates such as maternal education and IQ, SES, and HOME score in
with children with higher mean blood Pb levels, 8-14 (ig/dL (Kordas et al.. 2006; Schnaas
et al., 2006; Tong and Lu. 2001; Dietrich et al., 1993b). Exceptions include multiple
analyses of the Cleveland cohort, in which blood Pb level was estimated to have a weak
and imprecise or null effect after adjustment for potential confounding factors (Greene et
al.. 1992; Ernhart et al.. 1989; Ernhart et al.. 1988). Analyses of the Cleveland cohort
considered similar potential confounding factors as other studies. HOME score was the
major factor accounting for the attenuation of the Pb effect estimate in the Cleveland
cohort. An analysis of the Yugoslavia cohort, which adjusted for most of the same
covariates as the Cleveland analyses reported larger magnitude blood Pb-cognitive
function associations in covariate-adjusted models (Factor-Litvak et al.. 1999). The
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collective findings in children indicate potential confounding by the SES-related and
demographic factors examined in the literature base but also demonstrate that blood Pb
level is an independent predictor of cognitive function decrements in children with
adjustment for these factors.
A challenge to separating the effects of Pb exposure from those related to SES and
parental caregiving quality is their frequently high correlation with blood Pb levels. In
such cases, it is difficult to know how much variation in the outcome to attribute to each
of the risk factors (Needleman and Bellinger. 2001). For example, due to the high
correlation between blood Pb level and SES, a model that includes SES may
underestimate the Pb effect because some of the variance in outcome due to Pb is
mistakenly attributed to the variance due to SES. This misattribution may be exacerbated
when multiple correlated variables are included in the same model (i.e., overcontrol). The
relationships observed for Pb biomarker levels with SES and parental caregiving quality
may indicate that they are proxies or determinants of Pb exposure rather than a
confounder of the association of interest. Lower SES in urban children is closely linked
to residence in older, poorer condition housing that, in turn, may increase exposure of
children to environmental Pb and risk of cognitive function decrements (Clark et al.,
1985). In such cases where SES is a determinant of Pb exposure, statistical adjustment for
SES will result in overcontrol of the Pb effect (Bellinger. 2004a). This type of
overcontrol could apply to results from the New Zealand cohort, which were adjusted for
residence in older wooden housing, which is associated with higher exposure to Pb paint
and accumulated dust and soil and higher child tooth Pb levels (Tergusson et al.. 1988a.
b). However, even in models with older wooden housing, Pb remained a statistically
significant predictor of poorer reading skills and teacher ratings of school performance.
SES has been shown to be an effect modifier of the Pb-child cognitive function
relationship in some populations. Larger blood Pb-associated decreases in cognitive
function were found with lower SES in some studies (Ris et al., 2004; Tong et al., 2000;
Bellinger et al.. 1990) and higher SES in a meta-analysis (Schwartz. 1994). In cases of
effect modification, potential confounding by SES is less likely.
In summary, the collective epidemiologic evidence consistently demonstrates
associations of higher blood and tooth Pb levels with cognitive function decrements and
poorer behavior in children. These associations have been observed in diverse
populations in the U.S., Mexico, Europe, Asia, and Australia. Associations have been
observed across studies that use different methods to control for confounding and adjust
for different potential confounding factors but commonly, maternal IQ and education,
SES, and HOME score. Several studies have found associations with additional
adjustment for smoking exposure, child birth outcomes, and nutritional factors. No single
method to control for potential confounding is without limitation, and there is potential
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for residual confounding by unmeasured factors. However, the consistency of findings
among different populations and study methods with consideration of several well-
characterized potential confounding factors as described above increases confidence that
the associations observed between Pb biomarker levels and neurodevelopmental effects
in children represent a relationship with Pb exposure. Biological plausibility is provided
by the coherence with extensive evidence in animals with Pb exposures that produce
blood Pb levels relevant to this ISA and that is not subject to confounding by factors such
as social class and correlated environmental factors. Further, Pb exposure has been shown
to induce decrements in visual-spatial memory, rule learning and reversal, and response
inhibition, which also have been associated with blood or tooth Pb levels in children.
Additional support for the epidemiologic evidence is provided by extensive toxicological
evidence describing modes of action for Pb-induced cognition and behavioral problems,
including changes in neurogenesis, synaptic pruning, and neurotransmitter function in the
hippocampus, prefrontal cortex, and nucleus accumbens of the brain (Section 4.3.10).
4.3.14 Public Health Significance of Associations between Pb Biomarkers and
Neurodevelopmental Effects
As described in Section 4.3.2.1. most studies found that a 1 (ig/dL increase in blood Pb
level was associated with decrements in FSIQ in school-aged children in the range of <1
to 2 points, depending on the model and blood Pb level range examined (Figure 4-2 and
Table 4-3). Similarly, a 1 (ig/dL increase in blood Pb level typically was associated with
lower scores on tests of executive function (Table 4-8) and academic performance (Table
4-9), and higher ratings of behavioral problems (Figure 4-9. Table 4-11. Table 4-12) on
the order of less than 1 standard deviation. Such findings prompt consideration of the
public significance of blood Pb level-associated effects on decrements in cognitive
function and behavior in children, specifically, whether the magnitudes of change have
consequences on the health and life-success of individuals. According to the WHO,
"Health is a state of complete physical, mental and social well-being and not merely the
absence of disease or infirmity" (WHO. 1948). By this definition, even decrements in
health status that are not severe enough to meet diagnostic criteria might be undesirable if
they reflect a decrement in the well-being of an individual. Also, decreases in health
indices or life-success may not be observable except at the population level. The
American Thoracic Society discussed the need to consider the prevalence of exposures in
the population and exposure to other risk factors in evaluating whether shifts in the
population-level risk are adverse (ATS. 2000). Neurodevelopmental decrements
measured in childhood may set affected children on trajectories more prone toward lower
educational attainment and financial well-being. Thus, early effects in children may have
lifetime consequences. There also may be groups in the population at increased risk of
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neurodevelopmental effects from Pb exposure. For example, some evidence points to
larger blood Pb-associated decrements in cognitive function in children with lower SES
(Ris et al.. 2004; Tong et al.. 2000; Bellinger et al.. 1990). whereas a meta-analysis found
a larger effect estimate for studies with higher SES (Schwartz. 1994).
It has been argued that blood Pb-associated decrements in IQ points less than 3 or 4
points are meaningless given that such changes are within the standard error of a single
test (i.e., the statistic that defines the range within which the true value of an individual is
likely to lie) (Kaufman. 2001). However, this argument incorrectly assumes that
conclusions drawn from individual-level data apply to populations. Evidence does not
indicate that the standard error is nonrandom, i.e., biased in one direction. Hence, there is
no reason to expect that children with higher blood Pb levels systematically test lower
than their true IQ value or that children with lower blood Pb levels test higher than their
true IQ value. Thus, in a population of children, on a given assessment, some children
will test lower than their true value and others will test higher than their true value. In
such cases, between-group differences will be measureable on a population basis. Error in
the measurement of IQ in an individual will contribute nondifferential error on a
population-level and make it more difficult to detect an association if one exists.
The issue of individual-level versus population-level risk also pertains to the implications
of the magnitude of decrease in cognitive function or increase in behavioral problems per
unit increase in blood Pb level. Although fractional changes in FSIQ, memory, or
attention may not be consequential for an individual, they may be consequential on a
population level, especially in the two tails of the distribution (Bellinger. 2007. 2004b).
For example, 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. 2007. 2004b). In statistical exercises not specific to Pb or
analysis of data collected from individuals, Weiss (1990. 1988) predicted 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. In this hypothetical model,
with a reduction in population mean IQ from 100 to 95, the percentage of individuals
predicted to score above 130 (two standard deviations above the mean) decreases from
2.3% to 0.99%. Weiss (1988) stated that the implication of such a loss transcends the
current circumscribed definitions of risk. In addition to implications on the loss of
intellectual ability, the loss of a few IQ points potentially could result in the loss of
academic opportunities in children. For example, schools or programs for the gifted have
used IQ cut-offs (e.g., score of 130) to screen or accept applicants. In another
hypothetical analysis presented in the 2006 Pb AQCD (U.S. EPA. 2006b). based on a
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blood Pb-FSIQ effect estimate of-0.9 points/(ig/dL (the median effect estimated for
blood Pb levels <10 (ig/dL), the fraction of the population with an FSIQ <80 is predicted
to more than double from 9% with a blood Pb level of 0 (ig/dL to 23% with a blood Pb
level of 10 (ig/dL (Figure 4-16). The proportion with an FSIQ <70, which often requires
community support to live (WHO. 1992). is predicted to increase from a little over 2%
with a blood Pb level of 0 (ig/dL to about 8% with a blood Pb level of 10 (ig/dL [(Figure
4-16) and (U.S. EPA. 2006b)1. These theoretical exercises estimate that for an individual
in the low range of the IQ distribution, a Pb-associated decline of a few points might be
sufficient to drop that individual into the range associated with increased risk of
educational, vocational, and social failure.
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10
Note: The results presented are based on a theoretical analysis of changes in population FSIQ using a concentration-response
estimate of-0.9 FSIQ points/ug/dL, which was the median estimate from studies reviewed in the 2006 Pb AQCD for blood Pb levels
<10 ug/dL. The analysis estimates increasing proportions of individuals with FSIQ <80 and <70 with increasing blood Pb level.
Source: 2006 Pb AQCD (U.S. EPA, 2006b).
Figure 4-16 Hypothetical effect of increasing blood Pb level on the proportion
of the population with FSIQ <80 and <70 points.
The hypothetical predictions presented in Weiss (1990. 1988) and the 2006 Pb AQCD
(U.S. EPA. 2006b) have been supported by findings from analyses using data from
studies of blood Pb levels and FSIQ decrements in children. Among children in 1st and
2nd grades from towns around Boston, MA, Needleman et al. (1982) found that a 4-point
downward shift in the study population mean verbal IQ estimated for tooth Pb levels
>24 ppm was associated with a 3-fold increase in the percentage of children with a verbal
IQ of <80 and a decrease in the percentage achieving a verbal IQ >125 from 5% to 0%.
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The aforementioned hypothetical analyses and those using data collected from
individuals that estimate Pb-associated changes in the population IQ distribution assume
that the magnitude of change is equal across segments of the IQ distribution. Few studies
of Pb and cognitive function have examined whether the effect of Pb varies across the
distribution of cognitive function. However, in a recent study of fourth graders across the
entire state of North Carolina, Miranda et al. (2009) found that higher blood Pb level
measured once in each child between age 9 months and 3 years was associated with
larger decreases in fourth grade EOG scores in the lower segment of the EOG
distribution. An increase in blood Pb level from 1 to 10 (ig/dL was estimated to decrease
EOG score by 0.8 points in the 95th percentile of EOG scores but by 2.3 points in the 5th
percentile of EOG score (Figure 4-6). These findings by Miranda et al. (2009) based on
analysis of a large database representative of fourth graders in North Carolina indicate
that a shift in the population mean from increased Pb exposure may increase the
proportions of children at the lower end of the cognitive function distribution over that
estimated by theoretical analyses.
In summary, the public health significance of evidence demonstrating associations
between increases in blood Pb levels and decrements in IQ of children in the range of a
few points is supported by hypothetical predictions that a shift in the population mean
increases the proportion of individuals in the lower range of cognitive function and
decreases the proportion of individuals in the upper range of cognitive function. These
changes in the population distribution also were found in children in 1st and 2nd grade in
Massachusetts in whom higher tooth Pb level was associated with decrements in IQ
(Needleman et al.. 1982). Results from a few studies suggested that the public health
significance of decrements in cognitive function associated with higher blood Pb levels
may be greater in certain groups of children, for example, those in the lower range of
cognitive function (Miranda et al.. 2009) and those in lower SES groups (Ris et al.. 2004;
Tong et al.. 2000; Bellinger et al.. 1990). On a population-level, small Pb-associated
decreases in cognitive function could increase the number of individuals at increased risk
of educational, vocational, and social failure and decrease the number of individuals with
opportunities for academic and later-life success.
4.3.15 Summary and Causal Determination
The collective body of epidemiologic and toxicological evidence integrated across that
reviewed in the 2006 Pb AQCD (U.S. EPA. 2006b) and recent studies indicates
relationships between Pb exposure and a range of nervous system effects. In children,
effects on cognitive function include FSIQ, learning, memory, executive function, and
academic performance. Externalizing behaviors in children were evaluated in two groups:
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1) attention, impulsivity, hyperactivity, and ADHD; and 2) conduct disorders which
included aggression, delinquency, and criminal offenses. Effects on internalizing
behaviors include withdrawn behavior and symptoms of depression, fearfulness, and
anxiety. Other nervous system effects evaluated in children are auditory function, visual
function, and motor function. Relationships for Pb exposure with cognitive function,
auditory function, and visual function also were evaluated in adults. Other nervous
system effects in adults examined in relation to Pb exposure include those related to
psychopathological effects such as symptoms of depression, anxiety, and panic disorder.
Additionally, effects on neurodegenerative diseases include Alzheimer's disease, ALS,
Parkinson's disease, and essential tremor. The subsequent sections describe the
evaluation of evidence for each of these outcome groups with respect to causal
relationships with Pb exposure using the framework described in Table II of the
Preamble. The application of key evidence, supporting or contradictory, to the causal
framework is summarized in Table 4-17.
4.3.15.1 Evidence for Cognitive Function Decrements in Children
A causal relationship between Pb exposure and cognitive function decrements in children
is supported by several lines of evidence: 1) findings from prospective studies in diverse
populations consistently demonstrating associations of higher blood and tooth Pb levels
with lower FSIQ and performance on tests of executive function and academic
performance in children ages 4-17 years and young adults ages 18-20 years (Section
4.3.2); 2) coherence with evidence in animals for impairments in learning, memory, and
executive function with relevant Pb exposures; and 3) evidence describing modes of
action (Table 4-17).
Clear support for Pb-associated cognitive function decrements in children, as described in
the 2006 Pb AQCD (U.S. EPA, 2006b). was provided by prospective epidemiologic
studies indicating associations of higher earlier childhood (e.g., prenatal maternal or cord,
age 2 or 4 year), concurrent (ages 4-10 year), and childhood average blood and tooth Pb
levels with lower FSIQ in children ages 4-17 years (Table 4-17 and Section 4.3.2.1).
Evidence for such Pb biomarker metrics better characterized the temporal sequence
between Pb exposure and decrements in cognitive function than did cross-sectional
analyses. Across studies, FSIQ was measured with various instruments (i.e., WISC-R,
WISC-III, WPPSI, Stanford-Binet) that were similar in scoring scale and measurement
error. Associations were found in most of the prospective studies, conducted in the U.S.,
Mexico, Europe, and Australia in representative populations, most of which had moderate
to high follow-up participation without indication of selective participation among
children with higher blood Pb levels and lower cognitive function (Table 4-3). Another
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strength of the evidence from prospective studies was the consideration for several
potential confounding factors. As indicated in Table 4-3. results from most cohort studies
were adjusted for maternal IQ and education, child sex and birth weight, SES, and
HOME score. Although not considered as frequently, some studies also found
associations with adjustment for parental smoking, birth order, and nutritional factors.
The consistency and reproducibility of the blood Pb-FSIQ association in children were
demonstrated further in pooled analyses of seven prospective studies (Lanphear et al..
2005; Rothenberg and Rothenberg. 2005) as well as multiple meta-analyses that
combined results across various prospective and cross-sectional studies (Pocock et al..
1994; Schwartz. 1994; Needleman and Gatsonis. 1990). with Schwartz (1994)
demonstrating the robustness of evidence to potential publication bias. These pooled or
meta-analyses combine data that are somewhat heterogeneous in the tests used to assess
FSIQ and the ages of blood Pb and FSIQ examined which could be sources of non-
differential measurement error. Heterogeneity in the potential confounding factors
examined and the method of assessment could produce residual confounding.
Among individual studies, a wide range of blood Pb-FSIQ effect estimates was obtained,
which is not unexpected given the wide range of blood Pb levels examined and modeling
methods used (i.e., linear, log-linear). The pooled analysis of seven prospective cohorts
demonstrated precision of effect estimates by applying a uniform method across
populations (Lanphear et al.. 2005). A narrow range of estimates was obtained by
excluding one study at a time, -2.4 to -2.9 points per log increase in concurrent (ages 4.8-
10 years) blood Pb level. Results from several individual studies, in particular prospective
studies, indicated a supralinear concentration-response relationship which estimated a
greater decrement in cognitive function per unit increase in blood Pb level among
children in lower strata of blood Pb levels than children in higher strata of blood Pb levels
(Figure 4-15 and Table 4-16).
Among the largest effect estimates were those found in the Boston and Rochester cohorts
(Canfield et al.. 2003a; Bellinger et al.. 1992). which had relatively smaller sample sizes
but considered several potential confounding factors as listed above and in Table 4-3.
examined lower blood Pb levels than did other prospective studies, and examined cohorts
that differed in SES. Thus, their results may be more representative of the effects of Pb
exposure on FSIQ. In the Boston cohort, a 1 (ig/dL increase in age 2 year blood Pb level
was associated with a -1.6 (95% CI: -2.9, -0.2) point change in FSIQ at age 10 years in 48
children with age 2-year blood Pb levels 1-9.3 (ig/dL (detection limit not reported) and
peak blood Pb levels <10 (ig/dL (Bellinger. 2008: Bellinger and Needleman. 2003). In
the Rochester cohort, a 1 (ig/dL increase in concurrent blood Pb level was associated a
-1.8 (95% CI: -3.0, -0.60) change in FSIQ at age 5 years in 101 children with concurrent
blood Pb levels 0.5-8.4 (ig/dL (detection limit = 1 (ig/dL) and peak blood Pb levels
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<10 (ig/dL (Canfield. 2008; Canfield et al. 2003a). In the pooled analysis, a 1 (ig/dL
increase in concurrent blood Pb level was associated a -2.9 (95% CI: -5.2, -0.71) change
in FSIQ at ages 4.8-10 years in 103 children with a mean concurrent blood Pb level of
3.2 (ig/dL and peak blood Pb levels <7.5 (ig/dL (Hornung. 2008; Lanphear et al.. 2005).
Results from the Boston and Rochester cohorts can be considered particularly
informative because within each cohort, there was greater homogeneity in the tests used
to assess FSIQ, the age of examination, and the methods used to measure potential
confounding factors. Further, in Lanphear et al. (2005). 80% of the data for the lower
portion of the blood Pb distribution was provided by the Boston and Rochester cohorts.
These observations do not identify a threshold in the ranges of blood Pb level examined.
Null or weak associations were limited to a few cohorts, namely, the Cleveland and
Sydney cohorts (Greene et al.. 1992: Coonevetal.. 1991: 1989a. b; Ernhart et al.. 1988).
The Cleveland and Sydney studies were not outliers with respect to population mean
blood Pb levels or the specific confounding factors considered (Table 4-3). and the
Cleveland cohort had high prevalence of maternal prenatal substance abuse which may
limit the representativeness of results.
Other previous studies estimated smaller magnitude of effects for the blood Pb-FSIQ
association in children ages 4-13 years with either linear or log-linear models but
examined higher blood Pb levels (means: 7.8-16 (ig/dL) and did not analyze the
concentration response at the lower range of the study population blood Pb distribution
(Table 4-3). Most recent cross-sectional studies supported associations between higher
concurrent blood Pb levels and decrements in FSIQ. Some found associations in
populations of children ages 6-11 years with lower blood Pb levels than other studies,
i.e., means 1.7-5.0 (ig/dL (Table 4-17). In exception, among children ages 6-10 years in
New England with a mean concurrent blood Pb level 2.2 (ig/dL, lower FSIQ was found in
the group with blood Pb levels 5-10 (ig/dL (Surkan et al.. 2007). In many of these studies,
the potential influence of higher past Pb exposures cannot be excluded. Although Zailina
et al. (2008) had limited consideration for potential confounding, results from most of
these studies were adjusted for SES and parental IQ or education. Chiodo et al. (2007)
found an association with extensive consideration for potential confounding, including
parental caregiving quality and prenatal drug exposure, which was prevalent in the study
population.
A causal relationship between Pb exposure and cognitive function decrements in children
also is supported by previous prospective studies (several of which contributed to the
FSIQ evidence) that found associations of blood or tooth Pb level with decrements in
executive function and academic performance in populations ages 4-20 years (Sections
4.3.2.4. 4.3.2.5. and Table 4-17). The bodies of evidence for executive function, and
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academic performance are smaller than that for FSIQ but consistently indicate
associations with blood or tooth Pb level. Associations with performance on tests of
learning and memory were less consistently found across populations (Section 4.3.2.3).
In most studies, previous and recent, multiple testing was common; however, the
consistent pattern of association observed across the ages of blood Pb level and/or
cognitive test examined increases confidence that the evidence is not unduly biased by an
increased probability of associations found by chance alone. While recent studies of
executive function and academic performance adjusted for SES and parental education
and/or cognitive function, few examined potential confounding by parental caregiving
quality.
Adding to the evidence for Pb-associated cognitive function decrements in children were
recent prospective studies that indicated associations between higher earlier childhood
blood Pb level and poorer academic performance in school-aged children (Chandramouli
et al.. 2009; Miranda et al.. 2009). Among 57,678 children in North Carolina, lower
fourth grade EOG scores were found in children with age 9-36 month blood Pb levels
categorized as 2 (ig/dL compared with children with blood Pb levels categorized as
1 (ig/dL (detection limit = 1 (ig/dL), with adjustment for sex, race, school type, age of
blood Pb measurement, parental education, and enrollment in a free lunch program as an
indicator of SES (Miranda et al.. 2009). Results from this study also indicated a greater
incremental association of blood Pb level with decrement in EOG score among children
in the lower end of the EOG distribution. Chandramouli et al. (2009) found lower school
achievement tests scores in 488 children age 7 years in U.K. in the children with age
30 month blood Pb levels >5 (ig/dL. These results were adjusted for several potential
confounding factors, including SES, parental education, SES, home facilities score, and
family adversity. Recent cross-sectional studies conducted in the U.S. (Krieg etal. 2010;
Surkan et al.. 2007) found associations of concurrent blood Pb level with decrements in
executive function and academic performance in children, including the large analysis of
>700 children ages 12-16 years participating in NHANES (Krieg etal.. 2010). Cho et al.
(2010) did not find an association between concurrent blood Pb level and executive
function among children in Korea ages 8-11 years with a mean blood Pb level 1.9 (ig/dL.
Several studies found associations of higher prenatal, earlier infancy (e.g., age 1 year),
and concurrent blood Pb levels with lower Bayley MDI scores in children ages 2 and 3
years (Table 4-4). Similar to studies of FSIQ, Tellez-Rojo et al. (2006) estimated a larger
decrement in age 2 year MDI per unit increase in concurrent blood Pb level for children
in Mexico City with blood Pb levels <5 (ig/dL compared with children with blood Pb
levels 5-10 (ig/dL, and >10 (ig/dL (Figure 4-15 and Table 4-16). MDI is a well-
standardized measure of current infant mental development. However, the test of MDI is
not an intelligence test, and MDI scores, particularly before ages 2-3 years, are not
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necessarily strongly correlated with later measurements of FSIQ in children with normal
development.
A causal relationship between Pb exposure and cognitive function decrements in children
is further supported by consistent observations in rodents and monkeys of decrements in
learning, associative ability, memory, and executive function with relevant dietary Pb
exposures. In particular, coherence was found between evidence in children and animals
of Pb-associated decrements in visual-spatial memory, working memory (Section 4.3.2.3)
and rule learning and reversal (Section 4.3.2.4). Previous studies in monkeys ages 5-10
years demonstrated impairments in learning, memory, and executive function with
dietary Pb exposures during infancy only, lifetime after early infancy, and lifetime from
birth that produced blood Pb levels of 15-36 (ig/dL, some relevant to this ISA (Rice.
1992b: Rice and Gilbert. 1990a: Rice. 1990: Rice and Gilbert. 1990b: Rice and
Karpinski. 1988; Gilbert and Rice. 1987).
Several recent toxicological studies added to the evidence for impaired learning,
associative ability, and memory in animals with blood Pb levels, 10-25 (ig/dL after
gestational-lactational, lactational, or lifetime (with and without gestational) Pb exposure
(Corv-Slechtaetal.. 2010: Niu et al. 2009: Stangle et al.. 2007). Together, the
prospective epidemiologic and toxicological studies provide evidence for the temporal
sequence between Pb exposure and decrements in cognitive function. Additional
biological plausibility for Pb-associated cognitive function decrements was provided by
toxicological evidence for the effects of Pb on modes of action for cognitive function
(Section 4.3.10). Pb has been shown to increase the permeability of the blood-brain
barrier and deposit in the target CNS. Pb has been shown to impair neurogenesis,
synaptic architecture, and neurite outgrowth. The high activity of these processes during
fetal and infant development provides biological plausibility for the effects of childhood
Pb exposure on decrements in cognitive function. Cognitive function is mediated by the
cortical and subcortical structures of the brain that integrate function in the hippocampus,
prefrontal cortex, and nucleus accumbens using dopamine and glutamate as primary
neurotransmitters. Experimental studies have shown that Pb induces changes in dopamine
and glutamate release in these regions and decreases the magnitude of LTP, which is a
major cellular mechanism underlying synaptic plasticity and learning and memory.
With regard to critical lifestages and time periods of Pb exposure, toxicological evidence
clearly demonstrates impaired learning and memory in animals exposed to Pb
gestationally with or without lactational exposure. This evidence is well supported by
knowledge that processes such as neurogenesis, synaptogenesis, and synaptic pruning are
very active during this developmental period. However, evidence in monkeys also
indicates impaired cognitive function at ages 5-8 years with Pb exposure starting after
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early infancy (Rice. 1992b; Rice and Gilbert. 1990a; Rice. 1990; Rice and Gilbert.
1990b). Epidemiologic studies also found cognitive function decrements associated with
blood Pb levels measured during various lifestages and time periods. Distinguishing
among the effects of Pb exposures at different time periods is difficult in epidemiologic
studies due to the high correlations commonly found among blood Pb levels within
children overtime. Studies of young children ages 6 months to 3 years found lower MDI
in association with higher prenatal maternal or cord and postnatal child blood Pb levels;
some found larger magnitude decrements for prenatal blood Pb (Hu et al.. 2006; Gomaa
et al.. 2002). However, in older children, ages 4-17 years, in whom cognitive function is
more stable and reliably measured, larger decrements in cognitive function were found in
association with increases in postnatal blood Pb levels (concurrent, earlier childhood, and
cumulative average) than increases in prenatal blood Pb levels. There was no clear
indication of an individual critical lifestage, timing, or duration of Pb exposure associated
with cognitive function decrements in children. Because of the contribution of bone Pb
levels to concurrent blood Pb levels in children, associations with concurrent blood Pb
levels may reflect an effect of past and/or recent Pb exposures.
The consideration for potential confounding varied among studies. Most studies adjusted
for SES-related factors such as the Hollingshead Index, household income, and/or
parental education. Several, in particular the prospective studies, additionally adjusted for
parental cognitive function and caregiving quality mostly evaluated as HOME score. A
few studies considered nutritional factors. Not many recent studies considered potential
confounding by parental caregiving quality or nutritional factors. Controlling for SES is
difficult as it can be highly correlated with Pb exposure and there is no single measure
that fully represents SES. Residual confounding also is possible by factors not considered
or measured inadequately. The combination of evidence from prospective studies that
considered several well-characterized potential confounding factors plus evidence that Pb
exposure induces impairments in cognitive function in animals, in particular, for similar
constructs as those associated with Pb biomarkers in children, increase confidence that
the associations observed between blood Pb levels and cognitive function in children
represent a relationship with Pb exposure.
In conclusion, multiple prospective studies conducted in diverse populations consistently
demonstrate associations of higher blood and tooth Pb levels with lower FSIQ, executive
function, and academic performance and achievement in children. Most studies examined
representative populations and had moderate to high follow-up participation without
strong indication of selection bias. Associations between blood Pb level and cognitive
function decrements were found with adjustment for several potential confounding
factors, most commonly, SES, parental IQ, parental education, and parental caregiving
quality. Among studies that examined early childhood blood Pb levels (i.e., age < 3
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years), considered peak blood Pb levels in their analysis (i.e., peak < 10 (ig/dL), or
examined concurrent blood Pb levels in young children (i.e., age 4 years), blood Pb-
associated decrements in cognitive function were found in populations of children ages 4-
11 years with mean or group blood Pb levels in the range of 2 to 8 (ig/dL. Neither
epidemiologic nor toxicological evidence has identified an individual critical lifestage or
time period of Pb exposure within childhood that is associated with cognitive function
decrements. Several epidemiologic studies found a supralinear concentration-response
relationship. Examination of children with blood Pb levels in the range <1 (at or below
detection limits) to 10 (ig/dL, with consideration of early or peak childhood blood Pb
levels, has not identified a threshold for cognitive function decrements. Evidence in
children was clearly supported by observations of Pb-induced impairments in learning,
memory, and executive function in rodents and monkeys. Several studies in animals
indicated cognitive impairments with prenatal, lactational, and lifetime (with or without
prenatal) Pb exposures that resulted in blood Pb levels of 10-25 (ig/dL. The mode of
action for Pb-associated cognitive function decrements is supported by observations of
Pb-induced impairments in neurogenesis, synaptogenesis and synaptic pruning, LTP, and
neurotransmitter function in the hippocampus, prefrontal cortex, and nucleus accumbens.
The associations consistently found for FSIQ and other measures of cognitive function in
prospective studies of children with adjustment for SES, parental education, and parental
caregiving quality and the biological plausibility provided by evidence in animals for
impairments in learning, memory, and executive function with relevant Pb exposures and
evidence describing modes of action are sufficient to conclude that there is a causal
relationship between Pb exposure and decrements in cognitive function in children.
4.3.15.2 Evidence for Externalizing Behaviors: Attention,
Impulsivity, and Hyperactivity in Children
Attention, impulsivity, and hyperactivity are evaluated together because they are included
within the attention deficit hyperactivity disorder domain of externalizing behaviors (as
reviewed in Whitcomb and Merrell. 2012). Although examined less extensively than
cognitive function, a causal relationship between Pb exposure and attention decrements,
impulsivity, and hyperactivity in children is supported by multiple lines of evidence: 1)
findings from prospective studies in diverse populations for associations with blood or
tooth Pb levels; 2) coherence with evidence in animals with relevant Pb exposures and; 3)
evidence for plausible modes of action (Table 4-17). As described below, this conclusion
is based on evidence primary for attention decrements assessed with neuropsychological
testing or rated by parents or teachers and supporting evidence for impulsivity and
hyperactivity. The few studies showing blood Pb-associated increases in ADHD
prevalence are not a major consideration because of the cross-sectional or case-control
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design of studies and inconsistent consideration for potential confounding by factors such
as SES, parental education, or parental caregiving quality (Section 4.3.3.1).
Prospective studies provided key evidence for associations of childhood blood and tooth
Pb levels with attention decrements or hyperactivity in children ages 7-17 years and
young adults ages 19-20 years. Not all results were statistically significant, but results
within and across studies mostly showed a pattern of attention decrements or
hyperactivity with higher blood or tooth Pb level. The pattern of association observed
within studies increases confidence that the evidence is not unduly biased by the
increased probability of finding associations by chance alone. The results from
prospective studies better characterized the temporal sequence between Pb exposure and
measured behaviors than did cross-sectional studies. Associations were found in
populations in the U.S., U.K., Australia, and New Zealand (Table 4-17). Most studies had
population-based recruitment from prenatal clinics, hospitals at birth, or schools and had
moderate to high participation. A few prospective studies had greater loss-to-follow-up in
certain groups, for example, those with lower SES, earlier FSIQ, or HOME score. This
potential selection bias may have reduced the generalizability of findings to the original
study population, but there was not a strong indication that participation was biased to
those with higher blood Pb levels and greater behavioral decrements.
Key evidence provided by prospective studies supported associations of blood and tooth
Pb levels with attention decrements as assessed using neuropsychological tests or parent
or teacher ratings using widely-used structured questionnaires (Section 4.3.3.1). Thus, the
collective evidence does not appear to be unduly influenced by biased reporting of such
behaviors by parents of children with higher blood or tooth Pb levels. Prospective studies
that examined attention with the CPT found associations of higher prenatal or earlier
childhood (age 6 year, age 3-60 month) blood Pb levels or tooth (from lst/2nd grade) Pb
levels with increases in commission, omission, or reaction time errors in adolescents ages
15-17 years in Cincinnati (Ris et al.. 2004) and young adults 19-20 years in
Chelsea/Somerville, MA (Bellinger etal., 1994a). Results from prospective studies also
indicated associations of lifetime (to ages 11-13 years) average blood Pb levels and tooth
(from ages 6-8 years) Pb levels with lower parental and teacher ratings of attention in
children ages 8-13 years in Australia, New Zealand, and Boston, MA (Burns et al.. 1999;
Fergusson et al., 1993; Leviton et al., 1993). A recent prospective study in the U.K. found
that children ages 7-8 years with age 30 month blood Pb levels > 10 (ig/dL had higher
teacher ratings of hyperactivity but not selective attention compared with children with
blood Pb levels 0-2 (ig/dL (Chandramouli et al.. 2009). Results from children
participating in a chelation trial suggested a blood Pb-hyperactivity relationship that was
mediated through an association with IQ (Chen et al.. 2007). Previous findings for
hyperactivity were limited to cross-sectional and case-control studies (Section 4.3.3.1).
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Lifetime average blood Pb levels were not associated with ratings of attention in younger
children ages 4-5 years in Yugoslavia (Wasserman et al.. 2001). However, attention may
be less reliably rated in such young children.
Attention decrements were associated with biomarkers of Pb exposure representing
various lifestages and time periods. Prospective studies did not examine a detailed Pb
biomarker history, and results do not identify an individual critical lifestage, time period,
or duration of Pb exposure associated with attention decrements. Associations in
prospective studies for attention decrements with tooth Pb level, early childhood average
and lifetime average blood Pb levels point to an effect of cumulative Pb exposure.
Prospective studies did not examine biomarkers Pb exposure measured closer to the time
of behavioral assessment. Across prospective studies, attention decrements were
associated with blood Pb levels that had population means of 6.8 and 8.3 (ig/dL for
prenatal (maternal or cord) blood, 13.4 for age 3-60 month average blood, and 14 (ig/dL
for lifetime average blood. In Yugoslavia study with null results, the mean lifetime
average blood Pb level was 7.2 (ig/dL, similar to those associated with attention
decrements in some of the other prospective studies.
A causal relationship between Pb exposure and attention, hyperactivity, and impulsivity
is supported further by the consideration for several potential confounding factors in
prospective studies. Although the specific factors varied by study, prospective studies
adjusted for factors such as SES, parental IQ, maternal education, self drug use, prenatal
drug and alcohol exposure, birth outcomes, and parental caregiving quality. Parental
caregiving quality has been speculated as a means by which parental psychopathology
may influence child Pb exposure. However, a clear relationship between parental
psychopathology and poorer parenting behavior has not been established (as reviewed in
Johnston et al.. 2012). and adjustment for parental psychopathology may adjust for Pb
exposure itself. At present, there is not evidence to support consideration of parental
psychopathology as a direct confounding factor, i.e., contributing to a spurious
association with Pb biomarkers when there is none. However, if there is a noncausal
correlation between parental psychopathology and Pb exposure and it is mediated by
parental caregiving quality, then adjustment for parental caregiving quality also may help
address potential confounding by parental psychopathology.
Consistent with previous prospective studies, cross-sectional studies found associations
of higher concurrent blood or tooth Pb level with attention decrements, impulsivity, and
hyperactivity in children ages 3-16 years (Table 4-17). Because of the greater uncertainty
about the potential for reverse causation, cross-sectional results are a lesser consideration
in conclusions. The cross-sectional studies included a mix of previous and recent studies,
a mix of preschool-aged and school-aged children, and a wide range of population mean
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blood Pb levels (1-12 (ig/dL). Among the studies examining populations in the lower part
of this blood Pb distribution (means: 5-6.5 (ig/dL for ages 4-7.5 years) and with more
extensive consideration for potential confounding, most found associations with attention
decrements, hyperactivity, and/or impulsivity (Plusquellec et al. 2010; Chiodo et al..
2007; Chiodo et al.. 2004). All of these studies adjusted for or considered potential
confounding by SES as well as parental IQ, education, and caregiving quality. Among
children ages 4-6 years in whom behavior may be less reliably measured, concurrent
blood Pb level was associated with higher examiner ratings of attention decrements,
impulsivity, and hyperactivity in Inuit children ages 5-6 years in Canada (Plusquellec et
al., 2010) but not in the Rochester cohort at age 4.5 years (Canfield et al., 2003b). Also,
in the Rochester cohort, the association with attention was attenuated only after
adjustment for color/shape knowledge and child IQ. These inter-related outcomes may
share some variance. Among older children (age 7-7.5 years) in Detroit, MI, higher
concurrent blood Pb level (means: 5; 5.4 (ig/dL) was associated with higher examiner or
teacher ratings of attention decrements and hyperactivity (Chiodo et al.. 2007; Chiodo et
al.. 2004). These studies also found associations with attention decrements as assessed by
the CPT, although not strongly with all examined indices of the CPT. These populations
had high prevalence of prenatal alcohol or drug exposure, but results did not indicate
confounding by these other exposures. Nonetheless, the presence of these other exposures
may limit the generalizability of results. While higher-quality cross-sectional evidence in
populations with mean blood Pb levels 5-6.5 (ig/dL show associations of blood Pb level
with attention decrements, impulsivity, and hyperactivity, there are limitations related to
young age of subjects, generalizability of the study population, or potential contribution
of higher blood Pb levels earlier in childhood.
Other cross-sectional studies examined even lower concurrent blood Pb levels (means:
1-3.7 (ig/dL) and found associations with attention decrements, impulsivity,
hyperactivity, or a composite index of ADHD-related behaviors in children ages 8-17
years (Cho etal.. 2010: Nicolescu etal. 2010: Froehlich et al.. 2009: Nigg et al.. 2008).
These associations were found with parental or teacher ratings, with attention measured
by the CPT, and impulsivity measured with response inhibition tests. However, results
have weaker implications because of less extensive consideration for potential
confounding, case-control study design, and uncertainty regarding the influence of higher
blood Pb levels earlier in childhood. Studies varied in adjusting results for SES and
parental education, and none considered parental caregiving quality. Most studies of
populations with relatively high concurrent blood Pb levels (means 11-12 (ig/dL for ages
3-11 years) also found that higher blood Pb levels were associated with higher parent or
teacher ratings of attention decrements and hyperactivity (Roy et al.. 2009: Kordas et al..
2007: Silvaet al.. 1988). Most cross-sectional studies of higher blood Pb levels found
associations with adjustment for SES, child age, sex, family structure, and parental
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education, but most did not examine parental caregiving quality. With the exception of
findings from the Rochester cohort (Canfield et al.. 2003b). studies generally found
associations between blood Pb level and attention decrements, impulsivity, and
hyperactivity with adjustment for child IQ or other measure of cognitive function (Cho et
al..201Q; Chandramouli et al.. 2009; Nigg et al.. 2008; Silvaetal.. 1988). These findings
add support for higher Pb exposures having effects on attention, impulsivity, and
hyperactivity, independent of effects on cognitive function.
Further support for a causal relationship between Pb exposure and attention, impulsivity,
and hyperactivity is provided by observations of greater impulsivity in animals with
relevant dietary Pb exposures. Evidence in rodents and monkeys demonstrates Pb-
induced impaired response inhibition by decreased interresponse times and increased
overall response rates in FI reinforcement schedules (Table 4-7) (Rossi-George et al..
2011), slower acquisition of a DRL schedule (Rice and Gilbert. 1985). and response
perseveration on tests of discrimination reversal learning and FR/waiting for reward
(Brockel and Cory-Slechta. 1999b. 1998; Gilbert and Rice. 1987). In particular,
coherence is found with observations in children of associations between blood Pb levels
and poorer performance on the stop signal task, which also measures response inhibition.
Impulsivity in rodents and monkeys was found with gestational, post-weaning and
lifetime (from gestation or birth) dietary Pb exposures that resulted in blood Pb levels 11-
29 (ig/dL (Table 4-17 and Section 4.3.3.1). In animals, the effects of Pb exposure on
sustained attention were inconsistent as assessed with signal detection or discrimination
reversal tests with distracting stimuli (Table 4-17 and Section 4.3.3.1). A recent study did
not clearly indicate an effect of Pb on activity in male mice as an increase was found only
with amphetamine co-treatment (Leasure et al.. 2008): thus, the findings may not be
directly comparable to observations of Pb-associated increases in hyperactivity in
children.
Additional support for a causal relationship between Pb exposure and attention
decrements, impulsivity, and hyperactivity is provided by evidence describing modes of
action. These behaviors have been linked with changes in the prefrontal cerebral cortex,
cerebellum, and hippocampus, and Pb exposure has been found to affect development
and neuronal processes in these regions. For example, Pb has been found to affect
dopaminergic neurons of the frontal cortex and striatum of the brain by altering dopamine
release and receptor density. Other lines of evidence supporting the mode of action for
the effects of Pb exposure on attention, impulsivity, and hyperactivity include Pb-induced
changes in neurogenesis, synapse formation, and synaptic plasticity.
In conclusion, although examined less extensively than cognitive function, several
prospective studies demonstrated associations of blood or tooth Pb levels measured years
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before outcomes with attention decrements and hyperactivity in children 7-17 years and
young adults ages 19-20 years as assessed using objective neuropsychological tests and
rated by parents and teachers. Most of these prospective studies examined representative
populations without indication of participation conditional on blood Pb levels and
behavior. The results from prospective studies were adjusted for potential confounding by
SES as well as parental education and caregiving quality, with some studies also
considering parental cognitive function, birth outcomes, substance abuse, and nutritional
factors. In prospective studies, blood Pb-associated attention decrements and
hyperactivity were found in populations with prenatal (maternal or cord), age 3-60 month
average, age 6 year, or lifetime average (to age 11-13 years) mean blood Pb levels of 7 to
14 (ig/dL and in groups with age 30 month blood Pb levels >10 (ig/dL. Most well-
conducted cross-sectional studies that examined several potential confounding factors
found associations of attention decrements, impulsivity, and hyperactivity in children
ages 5-7.5 years with concurrent blood Pb levels with means of 5-5.4 (ig/dL but cannot
establish temporality or exclude the influence of higher blood Pb levels earlier in
childhood. Biological plausibility for observations in children is provided by several
findings in animals for increases in impulsivity or impaired response inhibition with
relevant post-weaning and lifetime Pb exposures that resulted in blood Pb levels of 11 to
29 (ig/dL. The mode of action for Pb-associated attention decrements, impulsivity, and
hyperactivity is supported by observations of Pb-induced impairments in neurogenesis,
synaptic pruning, and dopamine transmission in the prefrontal cerebral cortex,
cerebellum, and hippocampus. The consistency of epidemiologic evidence for attention
decrements and hyperactivity from prospective studies and the biological plausibility
provided by evidence for Pb-induced impulsivity in animals and for underlying modes of
action is sufficient to conclude that there is a causal relationship between Pb exposure
and effects on attention, impulsivity, and hyperactivity in children.
4.3.15.3 Evidence for Externalizing Behaviors: Conduct Disorders
in Children and Young Adults
There are two domains of conduct disorders: undersocialized aggressive conduct disorder
and socialized aggressive conduct disorder (as reviewed in Whitcomb and Merrell. 2012).
In this evaluation of a relationship with Pb exposure, the two domains are combined
because it is difficult to differentiate between these two domains in the available
epidemiologic studies. Criminal offenses are included in the evaluation because they can
be predicted by earlier conduct disorders (Soderstrom et al., 2004; Babinski et al., 1999;
Pajer. 1998). A causal relationship is likely to exist between Pb exposure and conduct
disorders in children and young adults based primarily on epidemiologic evidence
(Section 4.3.3.2). Key evidence is provided by most previous and recent prospective
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studies that found associations of blood Pb levels with higher parent and teacher ratings
of delinquent, aggression, and antisocial behavior in children ages 8-17 years from
diverse locations and SES (i.e., U.K., Cincinnati, Port Pirie, Australia) (Table 4-12 and
Table 4-17). Associations were found with lifetime average blood Pb levels in boys ages
11-13 years in Port Pirie, Australia (Burns et al.. 1999). with age 30 month blood Pb
levels in children ages 8 years in the U.K. (Chandramouli et al.. 2009). and age 0-6 year
average blood Pb levels in the Cincinnati cohort at age 15-17 years (Dietrich et al.. 2001).
Additional support was provided recent prospective studies finding associations of higher
earlier childhood blood (age 6 years) or tooth (from ages 6-8 years) Pb levels with
criminal offenses in young adults in Cincinnati and Christchurch, New Zealand, ages
19-24 years, as assessed through government records (Fergusson et al.. 2008; Wright et
al.. 2008). In the Cincinnati cohort, a 1 (ig/dL increase in age 6 year blood Pb level was
associated with an increased risk of criminal arrests at age 19-24 years with an RR of
1.05 (95% CI: 1.01, 1.09). The moderate to high follow-up participation and associations
found with parent and teacher ratings of behaviors related to conduct disorders do not
provide strong evidence for biased participation or reporting of behaviors for children
with higher blood Pb levels. Studies of criminal offenses and ratings of behaviors related
to conduct disorders found associations with adjustment for several potential confounding
factors such as SES, exposure to smoking, drugs, or alcohol, and parental caregiving
quality. Adjustment for parental caregiving quality may also address potential
confounding by parental psychopathology. Parental caregiving quality has been
speculated as a means by which parental psychopathology may influence child Pb
exposure. However, a clear relationship between parental psychopathology and poorer
parenting behavior has not been established (as reviewed in Johnston et al.. 2012). There
is not evidence to support consideration of parental psychopathology as a direct
confounding factor.
Supporting evidence was provided by cross-sectional studies, including the large analysis
of 2,867 children ages 8-15 years participating in NHANES, which found that compared
with children with concurrent blood Pb levels 0.2-0.7 (ig/dL (detection limit=0.3 (ig/dL),
children with concurrent blood Pb levels 0.8-1.0 (ig/dL had higher odds of conduct
disorder as assessed by parents with adjustment for age, sex, race, poverty income ratio,
and smoking exposure (Braun et al.. 2008). Chiodo et al. (2007) found an association
between concurrent blood Pb level and teacher ratings of social problems and delinquent
behavior in children ages 7 years in Detroit, MI after extensive consideration for potential
confounding factors including parental caregiving quality as well as prenatal drug
exposure, which was prevalent in the population. Further supporting the consistency of
association between blood Pb levels and behaviors related to conduct disorders in
children, a recent meta-analysis found that evidence was robust to heterogeneity in study
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design, definition and assessment method of conduct disorders, potential confounding
variables examined, and range of blood Pb levels (Marcus et al.. 2010).
Associations of measures of conduct disorders in children and young adults with earlier
childhood, earlier childhood average, and lifetime average blood Pb levels, tooth Pb
levels, and concurrent bone Pb levels point to the effects of early childhood or cumulative
Pb exposures. Most prospective studies did not analyze Pb biomarker levels at multiple
lifestages or time periods and thus did not provide information on potential associations
with more recent blood Pb levels or differences in association among Pb biomarkers at
various time periods. With respect to blood Pb levels, an association with criminal
offenses was found in young adults ages 19-24 years with a mean age 6 year blood Pb
level of 6.8 (ig/dL, and associations with behaviors related to conduct disorders were
found in children age 8 years with age 30 month blood Pb levels >10 (ig/dL, boys ages
11-13 years with a mean lifetime average blood Pb level of 14 (ig/dL, and children ages
15-17 years with a median age 6 year blood Pb level of 12 (ig/dL (Table 4-17). Cross-
sectional studies found associations with concurrent blood Pb level that were lower: 0-8-
1.0 (ig/dL for ages 8-15 years and a mean of 5 (ig/dL for children age 7 years. However,
the influence by higher past Pb exposures cannot be excluded.
Despite evidence from several high-quality epidemiologic studies that have examined
several potentially important confounding factors, an important uncertainty remains.
Evidence of Pb-induced aggression in animals was inconsistent, with increases in
aggression found in some studies of adult animals with gestational plus lifetime Pb
exposure but not juvenile animals (Section 4.3.3.2). The lack of biological plausibility
from animal evidence and the relatively small body of epidemiologic evidence produces
uncertainty regarding the effects of Pb exposure independent from other factors.
In conclusion, the few prospective studies consistently indicate that earlier childhood (age
30 months, 6 years) or lifetime average (to age 11-13 years) blood Pb levels or tooth
(from ages 6-8 years) Pb levels are associated with criminal offenses in young adults ages
19-24 years and with higher parent and teacher ratings of behaviors related to conduct
disorders in children ages 8-17 years. Pb-associated increases in conduct disorders were
indicated in populations with mean blood Pb levels 7-14 (ig/dL. Supporting evidence is
provided by a few cross-sectional studies and a meta-analysis of prospective and cross-
sectional studies. Associations with lower blood Pb levels, as found in the cross-sectional
studies, may have been influenced by higher earlier Pb exposures. Overall, associations
were found without indication of strong selection bias and with adjustment for SES,
family functioning, parental education, IQ, and caregiving quality, and maternal smoking
and substance abuse. Evidence for Pb-induced aggression in animals is mixed. The
consistent epidemiologic evidence from prospective and cross-sectional studies for
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criminal offenses and measures of conduct disorders, but uncertainty due to lack of clear
evidence for aggression in animals, is sufficient to conclude that a causal relationship is
likely to exist between Pb exposure and conduct disorders in children and young adults.
4.3.15.4 Evidence for Internalizing Behaviors in Children
A causal relationship is likely to exist between Pb exposure in children and internalizing
behaviors, including withdrawn behavior and symptoms of depression, fearfulness, and
anxiety based on evidence from a few epidemiologic studies and a few animal
toxicological studies. Internalizing behaviors have been examined to a lesser extent than
cognitive function or externalizing behaviors. However, supporting evidence is provided
by a few previous prospective studies that found associations of higher lifetime average
blood or tooth Pb levels with higher parent and teacher ratings of withdrawn behavior or
symptoms of anxiety and depression in school-aged children, 8-13 years (Burns et al..
1999; Bellinger et al.. 1994b) (Section 4.3.4.1. Table 4-17). These prospective studies
followed children from birth and had moderate follow-up participation to later childhood.
Participation was not conditional on early childhood blood Pb levels, and associations
were found with both parent and teacher ratings of internalizing behaviors, reducing the
likelihood of undue influence of biased participation and behavioral ratings by parents of
children with higher Pb exposures. Internalizing behaviors were assessed with widely-
used structured questionnaires such as the Child Behavior Checklist but not assessed as
clinically-diagnosed conditions such as depression.
The analysis of the Port Pirie, Australia cohort had the most extensive consideration of
potential confounding. Among only the 163 girls, ages 11-13 years, Burns et al. (1999)
found that a 1 (ig/dL increase in lifetime average blood Pb level in the interval 13.3-14.9
(ig/dL (10th-90th percentiles) was associated with an increased odds of an
anxious/depressed parental rating above the median of 1.07 (95% CI: 1.01, 1.14) with the
adjustment for several SES-related variables and factors related to parental caregiving
including HOME score, family functioning score, and current maternal psychopathology.
In a Boston-area cohort, Pb level in deciduous teeth (from age 6 years) was associated
with a higher teacher rating of a composite of anxious and social withdrawn behaviors in
children ages 8 years with adjustment for receiving public assistance at birth and
maternal education (Bellinger et al.. 1994b). Tooth Pb level was not associated with
internalizing behaviors in another Boston-area cohort at age 19-20 years (Bellinger et al..
1994a). In the Yugoslavia cohort, higher lifetime average blood Pb levels (mean: 7.2
(ig/dL) were associated with higher maternal ratings of anxious-depressed and withdrawn
behaviors in 191 children ages 4-5 years, with stronger associations found with
delinquent behaviors (Wasserman et al.. 2001). Behavior ratings may be less reliable in
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these younger children. With respect to critical lifestages and durations of Pb exposure,
evidence from prospective studies for associations with tooth Pb levels and lifetime
average blood Pb levels indicates an effect of cumulative Pb exposure on increasing
internalizing behaviors in children. While these studies adjusted for SES, birth outcomes,
and parental education, few considered parental caregiving quality. Thus, there is
uncertainty regarding potential confounding by parental caregiving quality.
Cross-sectional studies, including several recent studies, indicated associations between
concurrent blood Pb levels and internalizing behaviors in children ages 3-16 years.
Among the cross-sectional studies with more extensive consideration for potential
confounding and examination of relatively low concurrent blood Pb levels, i.e., means
~5 (ig/dL, results were inconsistent (Table 4-17). Previously, Chiodo et al. (2004) found
that among children age 7 years in Detroit, MI, HOME score, SES, maternal education,
and prenatal alcohol and drug exposure did not influence associations between blood Pb
level (mean: 5 (ig/dL) and internalizing behaviors. While the high prevalence of prenatal
alcohol and drug exposure may not affect internal validity of results, they may limit the
generalizability of findings. A lack of association between concurrent blood Pb level and
internalizing behaviors was reported in Inuit children age 5 years in Quebec, Canada with
a mean blood Pb level of 5.4 (ig/dL with consideration of potential confounding by
parental caregiving quality, caregiver education and IQ, blood Hg and PCB levels, and
prenatal smoking and alcohol exposure (Plusquellec et al.. 2010).
Some of the uncertainty in the small epidemiologic evidence base regarding a
relationship between Pb exposure and internalizing behaviors is addressed by the
biological plausibility provided by the coherence with evidence from a few studies that
relevant Pb exposures resulted in depression-like behavior or emotionality in rodents, i.e.,
dietary lactational Pb exposure producing blood Pb levels of 13 and 17 (ig/dL (Beaudin et
al.. 2007; Dvatlov and Lawrence. 2002). Other studies found Pb-induced increases in
emotionality, depression-like behavior, and tactile defensiveness in animals with blood
Pb levels >30 (ig/dL after gestational/lactational, lactational, or other early postnatal Pb
exposure (Section 4.3.4.2). Biological plausibility for Pb-associated increases in
internalizing behaviors also is provided by evidence that describes mode of action,
including Pb-induced changes in the HPA axis (Section 4.3.2.3) and dopaminergic and
GABAergic systems ((Sections 4.2.2.2. 4.3.10.4. and 4.3.10.8). which are found to affect
mood and emotional state.
In conclusion, prospective studies in a few populations demonstrate associations of
higher lifetime average blood (mean: -14 (ig/dL, age 11-13 years) or childhood tooth
(from ages 6-8 years) Pb levels with higher parent and teacher ratings of internalizing
behaviors such as withdrawn behavior and symptoms of depression, fearfulness, and
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anxiety in children ages 8-13 years. The lack of selective participation by blood Pb level
and associations found with parental and teacher behavioral ratings do not provide strong
indication of selection bias. The few cross-sectional studies in populations ages 5 and 7
years with mean concurrent blood Pb levels of 5 (ig/dL produced inconsistent findings.
Pb-associated increases in internalizing behaviors were found with adjustment for
maternal education and SES-related variables, and in a few studies, consideration for
parental caregiving quality. Some uncertainty in the epidemiologic evidence for the
effects of Pb on internalizing behaviors is addressed by the biological plausibility
provided by a few findings in animals with dietary lactational Pb exposure producing
blood Pb levels relevant to this ISA. Additional toxicological evidence supports modes of
action, including Pb-induced changes in the HPA axis and dopaminergic and GABAergic
systems. The evidence from prospective studies in a few populations of children and the
coherence with evidence from a few animal studies with relevant Pb exposures and with
evidence for mode of action but some uncertainty related to potential confounding by
parental caregiving quality in studies of children is sufficient to conclude that a causal
relationship is likely to exist between Pb exposure and internalizing behaviors in
children.
4.3.15.5 Evidence for Auditory Function Decrements in Children
A causal relationship is likely to exist between Pb exposure and auditory function
decrements in children based on primarily on epidemiologic evidence provided by a
previous prospective study indicating associations of poorer auditory processing with
infancy, childhood, and lifetime average blood Pb levels in 215 children age 5 years
(Dietrich et al.. 1992) and previous large (n = >3,000) cross-sectional NHANES and
HHANES studies showing associations of increasing hearing thresholds with concurrent
blood Pb levels in children ages 4-19 years (Schwartz and Otto. 1991. 1987). The high
follow-up participation from birth in the Cincinnati cohort and the examination of
multiple exposures and outcomes in NHANES and HHANES reduce the likelihood of
biased participation by children specifically with high blood Pb levels.
The epidemiologic evidence in children is strengthened by the consideration for several
potential confounding factors. In the Cincinnati cohort at age 5 years, higher prenatal
maternal, neonatal (age 10 day, mean: 4.8 (ig/dL), and various postnatal blood Pb levels
were associated with poorer auditory processing with adjustment for SES, HOME score,
several birth outcomes, and maternal alcohol consumption and with consideration for
factors such as maternal smoking and child health (Dietrich et al.. 1992). In NHANES
and HHANES, higher concurrent blood Pb levels (median: 8 (ig/dL) were associated with
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increased hearing thresholds with adjustment for age, sex, race, family income, parental
education, and nutritional factors (Schwartz and Otto. 1991. 1987).
Additional support for a relationship between Pb exposure and auditory function
decrements in children is provided by evidence supporting modes of action. A previous
prospective study in children in Mexico City (n = 100-113) found associations of prenatal
maternal (range: 1-8 (ig/dL) and age 1 and 4 year blood Pb levels (age 2 year mean:
10.8 (ig/dL) with lower auditory evoked potentials (Rothenberg et al.. 2000). Increased
hearing thresholds or latencies in brainstem auditory evoked potentials were found in
adult monkeys (ages 8-13 years) with lifetime Pb exposure beginning at gestation or birth
(Rice. 1997; Lilienthal and Winneke. 1996). Smaller effects were induced by Pb
exposure during gestation only or lactation only (to age 5 months) (Laughlin et al., 2009).
In animals, auditory effects were examined with higher Pb exposures than those relevant
to this ISA (i.e., blood Pb levels 33-150 (ig/dL at peak or testing); thus, it is difficult to
assess coherence with observations in children.
In conclusion, evidence from a prospective study and cross-sectional studies in a few
U.S. populations of children indicates associations of higher blood Pb level with
increases in hearing thresholds and decreases auditory processing or auditory evoked
potentials with adjustment for potential confounding by SES in most studies and by child
health and nutritional factors in some studies. The high participation rates, particularly in
the prospective study with follow-up from birth, reduce the likelihood of biased
participation by children with higher blood Pb levels. Across studies, associations were
found with blood Pb levels measured at various time periods, including prenatal maternal,
neonatal (10 day, mean 4.8 (ig/dL), lifetime average (to age 5 years), and concurrent
(ages 4-19 years) blood Pb levels (median 8 (ig/dL). Evidence for Pb-associated increases
in hearing thresholds or latencies of auditory evoked potentials was found in adult
monkeys with lifetime dietary Pb exposure. However, these effects in adult animals were
found with higher peak or concurrent blood Pb levels (i.e., 33-150 (ig/dL) than those
relevant to this ISA; thus, the biological plausibility for epidemiologic observations is
unclear. The evidence in children, particularly that from a prospective study, but
uncertainties related to effects on auditory function in juvenile animals with relevant Pb
exposures, is sufficient to conclude that a causal relationship is likely to exist between Pb
exposure and decrements in auditory function in children.
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4.3.15.6 Evidence for Visual Function Decrements in Children
The effects of relevant Pb exposures on visual function related to CNS changes
specifically in children have been examined in few toxicological and epidemiologic
studies. An epidemiologic study in children (Rothenberg et al.. 2002b) and toxicological
studies in adult rats found Pb-associated supernormal ERGs (Fox et al.. 2008). However,
the relevance of supernormal ERGs to visual function is unclear. Adult animal studies
also showed subnormal ERGs depending on the timing of Pb exposure and blood Pb
level. Toxicological studies also have characterized potential mechanisms leading to
retinal changes including alterations in morphology and cell architecture, signaling,
ATPase activity, neurotransmitter levels, cell proliferation, and retinal cell apoptosis. A
study of infant monkeys did not find infancy Pb exposure to affect on visual acuity
(Laughlin et al.. 2008) and used higher exposures than those relevant to this ISA (blood
Pb levels 35-40 (ig/dL). Because the available epidemiologic and toxicological evidence
is of insufficient quantity, quality, and consistency, the evidence is inadequate to
determine that a causal relationship exists between Pb exposure and visual function in
children.
4.3.15.7 Evidence for Motor Function Decrements in Children
A causal relationship is likely to exist between Pb exposure and motor function
decrements in children based mostly on evidence from previous prospective
epidemiologic studies with some supporting toxicological evidence. In the Cincinnati
cohort, higher neonatal, concurrent, and lifetime average blood Pb levels (means: 4.8, 10,
and 12 (ig/dL, respectively) were associated with poorer fine and gross motor function in
children at age 6 years (Bhattacharya et al.. 1995; Dietrich et al.. 1993a). and higher age
0-5 year average (mean: 12 (ig/dL) and 78 month blood Pb levels were associated with
poorer fine and gross motor function in children at ages 15-17 years (Bhattacharya et al..
2006; Ris et al.. 2004). In the Yugoslavian cohort, higher lifetime average blood Pb level
was associated with decrements in fine but not gross motor function at age 4.5 years
(Wasserman et al.. 2000b). In these studies, follow-up participation was not conditional
on blood Pb levels and motor function, reducing the likelihood of selection bias affecting
results. Motor function was assessed using varied but widely-used, structured tests. The
evidence from the Cincinnati and Yugoslavia cohorts is substantiated by the
consideration of several potential confounding factors such as SES, parental caregiving
quality, child health, and in adolescents, marijuana use (Ris et al.. 2004). In the
prospective studies, the population mean blood Pb levels were higher than those of most
of the current U.S. population. Recent cross-sectional studies examining concurrent blood
Pb levels that were lower, means 2-5 (ig/dL, produced mixed results. Higher concurrent
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blood Pb level was associated with poorer motor function with adjustment for several
potential confounding factors in 110 Inuit children ages 4-6 years living in subsistence
fishing communities (Despres et al., 2005) but not in a more representative population of
534 children ages 6-10 years in New England (Surkan et al.. 2007) or in children ages 7-
16 years in Korea (Min et al., 2007).
Epidemiologic evidence is supported by observations of poorer balance as assessed with
the rotarod test in male (not female) mice with relevant blood Pb levels, i.e., 10 (ig/dL
after dietary Pb exposure from gestation to PND10 (Leasure et al., 2008). Other
toxicological studies produced mixed results for effects on endurance, coordination, and
balance (Section 4.3.7) but are less relevant to this ISA because of the higher Pb exposure
concentrations examined, i.e., those producing blood Pb levels >30 (ig/dL.
In conclusion, evidence from prospective studies in a few populations indicates
associations of decrements in fine and gross motor function with higher neonatal,
concurrent, and lifetime average blood Pb levels in children ages 4.5-6 years and with
higher earlier childhood (ages 0-5 year average, age 78 months) blood Pb levels in
children ages 15-17 years. The means for these blood Pb metrics ranged from 4.8 to 12
(ig/dL. These associations were found with adjustment for several potential confounding
factors, including SES, parental caregiving quality, and child health and without
indication of substantial selection bias. Evidence from cross-sectional studies was less
consistent. The biological plausibility for associations observed in children is provided by
a study that found poorer balance in male mice with relevant gestational to early postnatal
(PND10) Pb exposures. The evidence in children, particularly from the few prospective
studies, and the coherence with limited available findings in mice is sufficient to
conclude that a causal relationship is likely to exist between Pb exposure and decrements
in motor function in children.
4.3.15.8 Evidence for Cognitive Function Decrements in Adults
A causal relationship is likely to exist between Pb exposure and cognitive function
decrements in adults based on prospective and cross-sectional epidemiologic studies that
indicate associations with bone Pb level and coherence with evidence in animals (Table
4-17). Key evidence for bone Pb levels was provided by recent prospective analyses of
the BMS and NAS, with support provided by the cross-sectional Nurses' Health Study.
The multiple risk factors and health outcomes examined in these studies reduces the
likelihood of biased participation and/or follow-up by adults specifically with higher Pb
exposure and lower cognitive function. While the NAS and Nurses' Health Study
examined primarily white men and white women, respectively, the BMS examined a
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more ethnically diverse population of men and women, increasing the generalizability of
findings. There was variability in associations across the various domains of cognitive
function tested within studies; however, bone Pb levels were associated with decrements
in most of the neuropsychological tests performed. In many studies, bone Pb levels were
associated with poorer executive function, visuospatial skills, learning, and memory.
Recent evidence from prospective analyses of the NAS and BMS cohorts expanded upon
previous cross-sectional evidence by improving characterization of the temporal
sequence between Pb exposure and cognitive function declines in adults (n = 358-965,
mean ages 59-69 years) by demonstrating that higher tibia (means: 19, 20 (ig/g) or patella
(mean: 25 (ig/g) bone Pb levels measured at baseline were associated with subsequent
declines in cognitive function over 2- to 4-year periods (Bandeen-Roche et al., 2009;
Weisskopf et al.. 2007b). The specific potential confounding factors considered differed
between studies; both studies adjusted for age and education. Additional adjustment was
made for household wealth in the BMS and current alcohol use and current smoking in
the NAS.
Evidence from most cross-sectional analyses supported associations between higher bone
Pb level and decrements in cognitive function in adults. A strength of cross-sectional
studies overall was the adjustment for many of the potential confounding factors
described above and also dietary factors, physical activity, medication use, and comorbid
conditions (Rajan et al., 2008; Weuve et al., 2006). Cross-sectional studies generally
demonstrated larger decrements in cognitive function in adults per unit increase in tibia
or patella Pb levels than concurrent blood Pb levels. Results from the NAS and Nurses'
Health Study did not clearly indicate a difference in association with cognitive
performance between tibia and patella Pb levels (Weuve et al., 2009; Weisskopf et al..
2007b). In NHANES analyses, higher concurrent blood Pb levels were associated with
lower cognitive function in particular age and genetic variant subgroups but not
consistently across the various cognitive tests conducted (Krieg et al.. 2010; Krieg and
Butler. 2009; Krieg et al.. 2009). NHANES did not have bone Pb measures for
comparison.
Because bone Pb is a major contributor to blood Pb levels, blood Pb level also can reflect
longer term exposures, including higher past exposures, especially in adults without
occupational exposures (Sections 3.3 and 3.7.3). Thus, in the NHANES results, it is
difficult to characterize the relative contributions of recent and past Pb exposures to the
associations observed between concurrent blood Pb level and cognitive function. The
discrepant findings for blood and bone Pb levels in other studies indicate that cumulative
Pb exposure that likely included higher past exposures may be a better predictor of
cognitive function in adults than is concurrent blood Pb level. Additional support for the
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effects of cumulative or past Pb exposure is provided by analyses of a few child cohorts
as adults, which indicate that childhood tooth (from lst/2nd grade) and blood (average
to age 10 years) Pb levels are associated with decrements in cognitive function in adults
ages 19-30 years (Mazumdar et al.. 2011; Bellinger et al.. 1994a). An uncertainty in the
evidence for bone Pb levels is potential residual confounding by age. Increasing age is
highly correlated with increasing bone Pb level (Section 3.3.5.2). and distinguishing
Pb-related declines in cognitive function from age-related declines with model
adjustment is difficult. One explanation for the more variable findings in adults than in
children may be that cognitive reserve may compensate for the effects of Pb exposure on
learning new information. Among current smelter workers, blood Pb-associated
decrements in cognitive function were larger among workers with lower cognitive
reserve (Bleecker et al., 2007a). Compensatory mechanisms may be overwhelmed with
age and with higher long-term or cumulative Pb exposure represented by higher bone Pb
levels.
Higher blood and bone Pb levels were associated with cognitive function decrements in
adults with current or former occupational Pb exposures (Table 4-17). Some studies
examined current workers with a mean concurrent blood Pb level of 31 ug/dL (Dorsey et
al.. 2006; Stewart et al.. 2002). Among adults with current occupational Pb exposures,
both concurrent and cumulative exposures may affect cognitive function. Several of these
studies considered potential confounding by a similar set of factors as did studies of
adults without occupational Pb exposures but did not examine other occupational
exposures such as manganese. In the prospective study of adults with former
occupational Pb exposure, peak tibia Pb levels were associated more strongly with
cognitive performance than were concurrent blood Pb levels (Khalil et al.. 2009a). Thus,
in the absence of higher current Pb exposures, cumulative Pb exposures may have a
greater effect on cognitive function in adults.
Additional support for a relationship between Pb exposure and cognitive function
decrements in adults is provided by the coherence with evidence in adult animals that
lifetime Pb exposure starting from gestation, birth, or after weaning induces learning
impairments (Table 4-17 and Section 4.3.2.3). Biological plausibility also is provided by
evidence describing the effects of Pb on modes of action underlying cognitive function.
Cognitive function is mediated by actions of the neurotransmitters dopamine and
glutamate in the hippocampus, prefrontal cortex, and nucleus accumbens. Experimental
studies have shown that Pb induces changes in neurotransmitter release in these regions.
Studies also have shown Pb-induced decreases in the magnitude of LTP, which is a major
cellular mechanism underlying learning and memory.
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In conclusion, in adults without occupational exposure, recent prospective studies in the
NAS and BMS cohorts indicate associations of higher baseline tibia (means 19, 20 (ig/g)
or patella (mean 25 (ig/g) Pb levels with declines in cognitive function in adults (>age 50
years) over 2- to 4-year periods. While the specific covariates differed between studies,
these bone Pb-associated cognitive function decrements were found with adjustment for
potential confounding factors such as age, education, SES, current alcohol use, and
current smoking. Supporting evidence is provided by cross-sectional analyses of the
NAS, BMS, and the Nurses' Health Study, which found stronger associations with bone
Pb level than concurrent blood Pb level. Cross-sectional studies also considered more
potential confounding factors, including dietary factors, physical activity, medication use,
and comorbid conditions. The multiple exposures and health outcomes examined in many
studies reduces the likelihood of biased participation by adults specifically with higher Pb
exposure and lower cognitive function. The collective evidence indicates associations in
cohorts of white men and women and a cohort of more ethnically diverse men and
women. The specific timing, frequency, duration, and magnitude of Pb exposures
contributing to the associations observed with bone Pb levels are uncertain. Also
uncertain is the potential for residual confounding by age. The effects of recent Pb
exposures on cognitive function decrements were indicated in Pb-exposed workers by
associations found with blood Pb levels, although these studies did not consider potential
confounding by other workplace exposures. The biological plausibility for the effects of
Pb exposure on cognitive function decrements in adults is provided by findings that
lifetime Pb exposures from gestation, birth, or after weaning induce learning impairments
in adult animals and by evidence for the effects of Pb on altering neurotransmitter
function in the hippocampus, prefrontal cortex, and nucleus accumbens. The associations
between bone Pb level and cognitive function decrements consistently found in the few
prospective and cross-sectional studies of adults without occupational Pb exposure, the
coherence with animal findings, and toxicological evidence supporting modes of action,
but uncertainties related to potential residual confounding by age in epidemiologic
studies, are sufficient to conclude that a causal relationship is likely to exist between
long-term cumulative Pb exposure and cognitive function decrements in adults.
4.3.15.9 Evidence for Psychopathological Effects in Adults
Evidence indicates that a causal relationship is likely to exist between Pb exposure and
psychopathological effects in adults, based on the associations found between concurrent
blood or bone Pb level and self-reported symptoms of depression, anxiety, and panic
disorder in large cross-sectional studies of adults (i.e., NHANES, NAS) and some
supporting toxicological evidence. Higher cord blood 5-ALA level was associated with
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schizophrenia in adults in California (Opler et al.. 2008; 2004). but because of the lack of
assessment of blood or bone Pb levels, results were not a major consideration in
conclusions. In both NHANES and NAS, the high participation rates and examination of
multiple exposures and outcomes reduces the likelihood that findings are influenced by
biased participation or reporting of symptoms by subjects specifically with higher Pb
exposures. Symptoms of depression and anxiety were assessed with widely-used
structured questionnaires such as the Profile of Mood States, but there is uncertainty
regarding the effects of Pb exposure on clinically-diagnosed disorders.
Epidemiologic evidence for associations between blood or bone Pb level and
psychopathological effects is strengthened by the adjustment for several potential
confounding factors including age, education, smoking, alcohol use, employment status,
and SES. Among adults ages 20-39 years participating in NHANES, 369 adults with
concurrent blood Pb level > 2.1 (ig/dL had the highest OR for self-reported major
depressive disorder (OR: 2.3 [95% CI: 1.1, 4.8]) and panic disorder (OR: 4.9 [95% CI:
1.3, 18]) compared with the 449 adults with blood Pb levels 0.3-0.7 (ig/dL (Bouchard et
al.. 2009). Among NAS men ages 48-94 years, higher concurrent blood, patella, and tibia
Pb levels were associated with self-reported anxiety (nervousness, tension), depression,
and phobic anxiety (a subcategory of anxiety but defined specifically as persistent fear,
irrational and disproportionate to a stimulus) (Rajan et al.. 2007; Rhodes et al.. 2003). A
14 (ig/g increase in tibia Pb level was associated with a self-reported anxiety score one
standard deviation above the mean with an OR of 1.2 (95% CI: 0.98, 1.4) (Rajan et al..
2007). Because of the cross-sectional design of studies, the temporal sequence between
Pb exposure and psychopathological symptoms in adults is uncertain. This uncertainty is
somewhat reduced with results for tibia Pb, since it is an indicator of cumulative Pb
exposure. For results with blood and bone Pb level, there is uncertainty regarding the
specific level, timing, frequency, and duration of Pb exposure associated with
psychopathological effects.
The epidemiologic evidence for Pb-associated psychopathological effects is supported by
the coherence with findings in rodents that dietary lactational Pb exposure with or
without additional post-lactational exposure resulted in depression-like behavior and
emotionality in rodents, with some evidence at blood Pb levels relevant to this ISA (13,
17 (ig/dL) (Beaudin et al.. 2007; Dyatlov and Lawrence. 2002). Other studies found
Pb-induced increases in depression-like behavior in animals with higher than relevant
blood Pb levels (Section 4.3.4.2). Further support for Pb-associated increases in
psychopathological effects in adults is provided by evidence that describes modes of
action, including Pb-induced changes in the HPA axis (Section 4.3.2.3) and dopaminergic
and GABAergic systems (Sections 4.2.2.2. 4.3.10.4. and 4.3.10.8). which are found to
affect mood and emotional state.
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In conclusion, cross-sectional studies in a few populations demonstrate associations of
higher concurrent blood, patella, or tibia Pb levels with self-reported symptoms of
depression and anxiety in adults. The examination of multiple exposures and outcomes in
the available studies does not provide strong indication of biased reporting of
psychopathological effects by adults specifically with higher Pb exposures. In adults,
Pb-associated increases in symptoms of depression and anxiety were found with
adjustment for age, SES, education, and in the NAS, daily alcohol intake and pack-years
smoking. The biological plausibility for epidemiologic evidence is provided by
observations of depression-like behavior or emotionality in animals with dietary
lactational Pb exposure, with some evidence at blood Pb levels relevant to this ISA and
by toxicological evidence supporting modes of action, including Pb-induced changes in
the HPA axis and dopaminergic and GABAergic systems. The associations of blood and
bone Pb level with symptoms of psychopathological effects found in the few studies of
adults without occupational Pb exposure as well as the biological plausibility provided by
the coherence with findings from a few rodent studies and evidence for underlying modes
of action, but uncertainties related to residual confounding of bone Pb results by age in
epidemiologic studies are sufficient to conclude that a causal relationship is likely to exist
between Pb exposure and psychopathological effects in adults.
4.3.15.10 Evidence for Auditory Function Decrements in Adults
The evidence is suggestive of a causal relationship between Pb exposure and auditory
function decrements in adults based primarily on findings from a few epidemiologic
studies. Key evidence in humans is provided by the recent analysis of NAS males in
which a 15 (ig/g higher tibia Pb level at mean age 64.9 years was associated with a 0.05
dB/year (95% CI: 0.017, 0.083) increase in hearing threshold for a pure tone average
frequency (Parketal.. 2010). Results were adjusted for baseline age, race, education,
occupational noise, BMI, pack-years smoking, noise notch, diabetes, and hypertension.
Although the generalizability of results in this primarily white population of men is
limited, high follow-up participation and the examination of multiple exposures and
outcomes in this cohort reduces the likelihood that findings are biased by selective
participation of men specifically with higher Pb exposures. Bone Pb levels were
measured up to 20 years after the initial hearing measurement; however, tibia Pb level is
considered an indicator of cumulative Pb exposure since the half-life of Pb in bone is on
the order of decades (Section 3.3). Bone Pb levels increase with age, and although age
was included as a model covariate, residual confounding by age is possible. Supporting
evidence was provided by a recent case-control study of adults attending a hospital for
occupational health exams. Despite limitations of a nonrandom population and uncertain
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comparability of controls, the examination of multiple metals reduces the likelihood of
biased participation by higher Pb exposure. Higher concurrent blood Pb level was
associated with hearing loss with adjustment for several factors including blood levels of
Mn, As, and Se (Chuang et al. 2007).
Pb-associated increases in hearing thresholds or latencies in brainstem auditory evoked
potentials were found in adult monkeys ages 8-13 years with lifetime Pb exposure in
drinking water from gestation or birth (Rice. 1997; Lilienthal and Winneke. 1996). A
recent study found small increases in hearing thresholds in monkeys at age 13 years with
Pb exposure only during gestation or lactation (to age 5.5 months) (Laughlin et al.. 2009).
However, these effects were found with higher blood Pb levels (i.e., at peak or testing
33-150 (ig/dL) than those relevant to this ISA; thus, the biological plausibility for
epidemiologic observations is unclear.
In conclusion, the evidence provided by the analysis of NAS men for associations of
higher tibia Pb level with a higher rate of elevations in hearing threshold over 20 years
but uncertainties related to effects on auditory function in adult animals with relevant Pb
exposures, is suggestive of a causal relationship between Pb exposure and auditory
function decrements in adults.
4.3.15.11 Evidence for Visual Function Decrements in Adults
The effects of relevant Pb exposures on visual function related to CNS changes,
specifically in adults have been examined in some toxicological studies, but
epidemiologic evidence is sparse. Studies in adult animals show differential effects on
ERGs, depending on the timing and concentration of Pb exposure. Although studies have
linked Pb-induced ERG changes to changes in rod cell proliferation, retinal cell
apoptosis, and dopamine, the implications of findings are unclear because the functional
relevance of the observed ERG changes is unclear. Decreased visual acuity was found in
adult monkeys with lifetime (to age 7-9 years) Pb exposure but at higher concentrations
than those relevant to this ISA, 50-115 (ig/dL (Rice. 1998). A case-control study found
higher Pb concentrations in retinal tissue of macular degeneration adult cases (Table
4-17) but was limited by the inability to establish directionality of effects, the uncertain
representativeness of the control population, and the lack of rigorous statistical analysis
or examination of potential confounding factors. Because the available epidemiologic and
toxicological evidence is of insufficient quantity, quality, and consistency, the evidence is
inadequate to determine that a causal relationship exists between Pb exposure and visual
function decrements in adults.
4-308
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4.3.15.12 Evidence for Neurodegenerative Diseases in Adults
Epidemiologic and toxicological studies combined produced inconclusive evidence for
Alzheimer's disease and ALS but found associations between indicators of Pb exposure
and neurodegenerative diseases such as Parkinson's disease and essential tremor.
However, study limitations as described below limit the implications of findings.
The few case-control studies of essential tremor found higher concurrent blood Pb levels
in cases than controls (Section 4.3.9.4). A common limitation of these studies was the
potential for reverse causation. Reduced physical activity among cases could result in
greater bone turnover and greater release of Pb from bones into blood in cases than
controls. Some case-control studies found adults with Parkinson's disease to have higher
bone Pb levels, which are not likely to increase with decreases in physical activity
(Weisskopf et al., 2010; Coon et al., 2006). While some of these studies of Parkinson's
disease and essential tremor considered potential confounding by factors such as age, sex,
race, education, and alcohol consumption, they did not consider Mn co-exposure.
Epidemiologic findings for Parkinson's disease are supported by limited available mode
of action evidence for Pb-induced decreased dopaminergic cell activity in the substantia
nigra, which contributes to the primary symptoms of Parkinson's disease.
The few case-control studies of Alzheimer's disease did not find higher prevalence of
occupational Pb exposure or higher brain Pb levels in cases but did not measure Pb in
blood or bone (Section 4.3.9.1). Toxicological studies indicated that gestational-
lactational or infancy Pb exposure of monkeys and rats (but not adult-only exposure in
rats) induced pathologies that underlie Alzheimer's disease, including the formation of
amyloid plaques and neurofibrillary tangles in the brains of aged animals (Section
4.3.9.1). While these results suggest the need to consider early-life Pb exposure in
epidemiologic studies, some indicate that effects may be attributable to the high Pb
exposure concentrations tested, i.e., producing blood Pb levels >40 (ig/dL in rats (Li et
al.. 2010b: Basha et al.. 2005). Studies of ALS have not consistently found higher Pb
biomarker levels or history of occupational Pb exposure among ALS cases and controls
(Section 4.3.9.2). and a recent study found that higher tibia and patella Pb levels were
associated with longer survival time among ALS cases (Kamel et al., 2008).
4-309
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In conclusion, while evidence is inconclusive for ALS and Alzheimer's disease, a few
case-control studies each found higher blood Pb levels in adults with essential tremor and
higher bone Pb levels in adults with Parkinson's disease. Because of the inconclusive
evidence for some diseases and limitations such as reverse causation for essential tremor
and the lack of consideration for potential confounding by Mn exposure for both essential
tremor and Parkinson's disease, the evidence is inadequate to determine that there is a
causal relationship between Pb exposure and neurodegenerative diseases.
4-310
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Table 4-17 Summary of Evidence Supporting Nervous System Causal Determinations.
Attribute in Causal
Framework3
Key Evidence
References
Pb Biomarker Levels
Associated with Effects0
Cognitive Function Decrements in Children - Causal
Consistent
associations from
multiple, high quality
epidemiologic
studies with relevant
blood Pb levels
Evidence from prospective studies for decrements in
FSIQ in association with prenatal, earlier childhood,
peak, concurrent, lifetime average blood Pb levels
and tooth Pb levels in children ages 4-17 yr in
multiple U.S. locations, Mexico, Europe, Australia
Evidence from prospective studies for lower scores
on tests of executive function and academic
performance in association with earlier childhood or
lifetime average blood Pb levels or tooth Pb levels in
children ages 5-20 yr in multiple U.S. locations, U.K,
New Zealand. Associations less consistent for
learning and memory.
Supporting evidence from cross-sectional studies of
children ages 3-16 yr, but most did not consider
potential confounding by parental caregiving quality.
Includes large NHANES III analysis.
Outcomes assessed using widely-used, structured
questionnaires.
Several studies indicate supralinear C-R
relationship, with larger decrements in cognitive
function per unit increase in blood Pb at lower blood
Pb levels in children ages 5-10 yr
Canfield et al. (2003a),
Bellinger et al. (1992),
Jusko et al. (2008),
Dietrich et al. (1993b).
Schnaas et al. (2006).
Wasserman et al. (1997).
long et al. (1996).
Lanphearet al. (2005)
Plus Table 4-3. Section 4.3.2.1
Canfield et al. (2004).
Stiles and Bellinger (1993).
Miranda et al. (2009: 2007a),
Fergusson et al. (1997, 1993),
Leviton et al. (1993).
Chandramouli et al. (2009)
Sections 4.3.2.3. 4.3.2.4. 4.3.2.5
Surkan et al. (2007).
Kim et al. (2009b).
Roy et al. (2011),
Lanphearet al. (2000),
Froehlich et al. (2007),
Chiodo et al. (2007: 2004)
Canfield et al. (2003a).
Bellinger et al. (1992).
Jusko et al. (2008),
Kordas et al. (2006),
Lanphearet al. (2005)
Plus Table 4-16
Blood Pb (various time periods & lifestages):
Means 3-16 ug/dL
With consideration of peak or early childhood
blood Pb levels: Means 3-8 ug/dL for
concurrent (age 4, 5 yr), age 2 yr
Blood Pb (various time periods & lifestages):
Means 4.8-7.2 ug/dL, Groups with early
childhood blood Pb 2-16 ug/dL and 5-10 ug/dL
Tooth Pb (ages 6-8 yr): means 3.3, 6.2 ug/g
Concurrent (ages 3-16 yr) blood Pb : Means
1.7-12 ug/dL, Group (ages 6-10 yr) with blood
Pb5-10ug/dL
Groups with peak blood Pb <10 ug/dL:
concurrent mean 3.3 ug/dL, age 2 year mean
3.8 ug/dL
4-311
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Table 4-17 (Continued): Summary of Evidence Supporting Nervous System Causal Determinations.
Attribute in Causal
Framework3
Key Evidence
References
Pb Biomarker Levels
Associated with Effects0
Additional
epidemiologic
evidence to help
rule out chance,
bias, and
confounding with
reasonable
confidence
Several epidemiologic studies found associations
with adjustment for SES, maternal IQ and education,
HOME score. Several adjust for birth weight,
smoking. A few, nutritional factors.
Epidemiologic studies had population-based
recruitment, most with moderate to high follow-up
participation not conditional on blood or tooth Pb
level or cognitive function.
Pooled and meta-analyses demonstrate the
consistency of association
Table 4-3. Table 4-5:
Table 4-8. Table 4-9.
Sections 4.3.2.1, 4.3.2.3
4.3.2.4. and 4.3.2.5
Lanphearet al. (2005),
Pocock et al. (1994).
Schwartz (1994)
Consistent evidence
in animals with
relevant exposures
to help rule out
chance, bias, and
confounding with
reasonable
confidence
Impaired learning and associative ability in juvenile
and adult animals as indicated by performance in
tasks of visual discrimination, water maze, y maze,
and operant conditioning with schedules of
reinforcement with relevant dietary Pb exposure.
Impaired learning, memory, executive function in
adult monkeys as indicated by poorer performance
on delayed spatial alternation and spatial
discrimination reversal learning tasks with dietary Pb
exposures.
Stangle et al. (2007),
Niu et al. (2009),
Cory-Slechta et al. (2010),
Altmann et al. (1993).
Section 4.3.2.3
Gilbert and Rice (1987).
Rice and Karpinski (1988).
Sections 4.3.2.3 and 4.3.2.4
Blood Pb (after prenatal/ lactation, lactation
only, prenatal/lifetime Pb exposure): 10-
25 ug/dl_
Blood Pb (after lifetime Pb exposure from
birth): 15, 25 ug/dl_
4-312
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Table 4-17 (Continued): Summary of Evidence Supporting Nervous System Causal Determinations.
Attribute in Causal
Framework3
Key Evidence
References
Pb Biomarker Levels
Associated with Effects0
Evidence describes
mode of action:
Impaired neuron
development
Synaptic changes
LTP
Neurotransmitter
changes
Decreased neurogenesis in hippocampus DG
(involved in LTP and learning). Decreased NMDAR
(involved in integration of new neurons into existing
neuronal pathways). Decreased neurite outgrowth.
Found in animals with dietary gestational-lactational,
lactational, post-lactational (3-8 weeks), lifetime from
gestation Pb exposures.
Decreased synaptic development. Changes in
synaptic protein composition. Decreased ATP and
AchE, which both mediate neurotransmission.
Found in animals with dietary gestational with or
without additional lactational Pb exposures.
Decreased magnitude, increased threshold of LTP
with gestational-lactational or lifetime Pb exposure.
Changes in dopamine metabolism. Increased
sensitivity of dopamine receptor. Increased
catecholamine transmission in cerebral cortex,
cerebellum, hippocampus. Decreased glutamate
and expression of glutamate receptor, NMDAR.
Found in animals with dietary gestational-lactational,
lactational, or post-lactational Pb exposure.
Sections 4.3.10.9 and 4.3.10.10
Section 4.3.10.4
Sections 4.3.12, 4.3.10.7,
4.3.10.8
Section 4.3.10.8
4-313
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Table 4-17 (Continued): Summary of Evidence Supporting Nervous System Causal Determinations.
Attribute in Causal
Framework3
Key Evidence
References
Pb Biomarker Levels
Associated with Effects0
Externalizing Behaviors: Attention, Impulsivity, and Hyperactivity in Children - Causal
Consistent
associations from
multiple, high quality
epidemiologic
studies with relevant
blood Pb levels
Evidence from prospective studies for attention
decrements and hyperactivity in association with
prenatal (maternal or cord), early childhood, and
lifetime avg blood Pb and tooth Pb levels in children
ages 7-17 yr and young adults 19-20 yr in U.S.,
U.K., Australia, New Zealand.
Association not found with ratings of attention
problems in children ages 4-5 yr in whom ratings
may be measured less reliably
Associations found with neuropsychological tests
(CPT) and teacher and parent ratings using widely-
used, structured questionnaires.
Burns et al. (1999),
Ris et al. (2004).
Fergusson et al. (1993).
Bellinger et al. (1994a).
Chandramouli et al. (2009),
Leviton et al. (1993)
Section 4.3.3.1
Wasserman et al. (2001)
Section 4.3.3.1
Blood Pb:
Means 6.8, 8.3 ug/dL (prenatal maternal or
cord), 8.3 ug/dL age 6 yr, 13.4 age 3-60
month, 14 ug/dL (lifetime avg to age 11-13 yr)
Group with age 30 mo >10 ug/dL
Tooth Pb (ages 6-8 yr): Means: 3.3, 6.2 ug/g
Blood Pb: Mean 7.2 ug/dL for lifetime (to age
4-5 yr) avg
Consistent
associations from
multiple, high quality
epidemiologic
studies with relevant
blood Pb levels
(Continued)
Supporting evidence from higher quality cross-
sectional studies (consideration for several potential
confounding facors) for associations of concurrent
blood Pb level with attention decrements,
impulsivity, and hyperactivity in children ages 5-7.5
yr. Some populations had high prenatal drug or
alcohol exposure.
No association in Rochester cohort, age 4.5 yr.
Associations found in most other cross-sectional
studies with less extensive consideration for
potential confounding
Chiodo et al. (2007: 2004)
Plusquellec et al. (2010)
Section 4.3.3.1
Canfield et al. (2003b)
Section 4.3.3.1, Table 4-11
Concurrent (ages 5-7.5 yr) blood Pb: Means
5.0-5.4 ug/dl_
Concurrent (age 4.5 yr) blood Pb: Mean 6.5
Concurrent (ages 3-17 yr) blood Pb:
Means 1-12 ug/dL
Additional
epidemiologic
evidence to help
rule out chance,
bias, and
confounding with
reasonable
confidence
Most prospective and some cross-sectional studies
found associations with adjustment for SES,
maternal education, and parental caregiving quality.
Some also considered parental IQ, smoking, birth
outcomes. A few considered substance abuse,
nutritional factors.
Studies had population-based recruitment with
moderate to high follow-up participation not
conditional on blood or tooth Pb level.
Section 4.3.3.1. Table 4-11
4-314
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Table 4-17 (Continued): Summary of Evidence Supporting Nervous System Causal Determinations.
Attribute in Causal
Framework3
Key Evidence
References
Pb Biomarker Levels
Associated with Effects0
Consistent evidence
in animals with
relevant exposures to
help rule out chance,
bias, and confounding
with reasonable
confidence
Impulsivity indicated by premature responses,
increased response perseveration, decreased pause
time between events on tests with Fl, FR/waiting for
reward, or Differential reinforcement of low respose
rate in rodents and monkeys with relevant dietary
postnatal Pb exposures.
Inconsistent evidence for sustained attention in rats
and monkeys with relevant postnatal dietary Pb
exposures.
Increased activity in male mice found only with Pb
and amphetamine co-exposure.
Rossi-George et al. (2011).
Brockel and Cory-Slechta
Q999b, 1998),
Rice and Gilbert (1985).
Gilbert and Rice (1987)
Section 4.3.3.1
Sustained attention:
Gilbert and Rice (1987),
Brockel and Cory-Slechta
(1999a)
Hyperactivity:
Leasure et al. (2008)
Section 4.3.3.1
Blood Pb: 15, 25 ug/dL in monkeys with
lifetime exposure, 11-16 ug/dL in rats with
lifetime exposure from gestation; 11-29 ug/dL
in rats with 3, 7 mo post-weaning exposure
Blood Pb: 15, 25 ug/dL in monkeys after
lifetime (after birth) exposure; 16, 28 ug/dL in
rats after 34-day post-weaning exposure
Evidence describes
mode of action
Same as above for cognitive function in children
4-315
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Table 4-17 (Continued): Summary of Evidence Supporting Nervous System Causal Determinations.
Attribute in Causal
Framework3
Key Evidence
References
Pb Biomarker Levels
Associated with Effects0
Externalizing Behaviors: Conduct Disorders in Children and Young Adults (e.g., delinquent, aggressive, antisocial behavior, and criminal offenses) -
Likely Causal
Consistent results
from a few high-
quality
epidemiologic
studies with relevant
blood or tooth Pb
levels
Evidence from prospective studies for higher parent
and teacher ratings of aggressive, antisocial,
delinquent behavior in children ages 8-17 yr in U.S.,
U.K., Australia in association with earlier childhood
or lifetime average blood Pb levels.
Evidence from prospective studies for criminal
offenses in young adults ages 19-24 yr in Cincinnati
and New Zealand in association with earlier
childhood blood or tooth Pb levels.
Association found in Yugoslavia cohort at ages 4-5
years, when behavior ratings may be less reliable.
Studies had moderate to high follow-up participation,
not conditional on blood Pb level.
Most studies considered potential confounding by
SES, parental IQ and education, other SES factors,
parental caregiving quality, family functioning,
parental smoking or substance abuse.
Supporting evidence from cross-sectional studies of
children ages 7-15 years, including NHANES
analysis and study of children in Detroit, Ml that had
extensive consideration for potential confounding.
Dietrich et al. (2001),
Burns et al. (1999),
Chandramouli et al. (2009),
Section 4.3.3.2
Wright et al. (2008).
Fergusson et al. (2008),
Section 4.3.3.2
Wasserman et al. (2001)
Table 4-12
Blood Pb: lifetime (to age 11-13 yr) avg mean
14 ug/dL, age 6 year mean 12 ug/dL, group
with age 30 month blood Pb >10 ug/dL
Age 6 yr blood Pb mean 6.8 ug/dL
Tooth Pb (age 6-8 yr) mean: 6.2 ug/g
Mean lifetime (to ages 4-5 yr) avg: 7.2
Braun et al. (2008)
Chiodo et al. (2007)
Section 4.3.3.2
Groups ages 8-17 years with concurrent blood
Pb 0.8-1 Oug/dL
Concurrent (age 7 year) blood Pb mean: 5.0
Consistency supported by meta-analysis indicating
similar effect estimates by study design, potential
confounding factors considered
Teacher and parental ratings derived from widely-
used, structured questionnaires.
Marcus et al. (2010)
Section 4.3.3.2
Blood Pb range of study means (concurrent,
lifetime avg): 1.0-26 ug/dL
4-316
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Table 4-17 (Continued): Summary of Evidence Supporting Nervous System Causal Determinations.
Attribute in Causal
Framework3
Key Evidence
References
Pb Biomarker Levels
Associated with Effects0
But, uncertainty
because of
inconsistent animal
evidence with
relevant exposures
and small number of
epidemiologic
studies
Aggression observed in adult hamsters with
gestational-lifetime dietary Pb exposure. Other
evidence in adult animals with similar duration
exposure inconsistent.
Aggression generally not found in juvenile animals
with lactational Pb exposure.
No specific mode of action examined with Pb
exposure.
Delville (1999)
Section 4.3.3.2
Blood Pb after gestational-lifetime exposure:
10-15ug/dL
Internalizing Behaviors in Children (e.g., withdrawn behavior, symptoms of depression, anxiety, tearfulness) - Likely Causal
Consistent results
from a few high-
quality
epidemiologic
studies with relevant
exposures
Evidence from prospective studies for higher ratings
of internalizing behaviors in children ages 8-13 yr in
Boston and Port Pirie cohorts in association with
tooth or lifetime average blood Pb levels.
Associations also found in children age 4-5 yr in
Yugoslavia in association with lifetime average
blood Pb level; stronger association found for
delinquent behavior.
Studies had population-based recruitment with
moderate follow-up participation. Participation not
conditional on tooth/blood Pb levels and behavior.
Associations found with teacher and parent ratings
on widely used, structured questionnaires.
Inconsistent associations in cross-sectional studies
that had extensive consideration for potential
confounding and examined relatively low concurrent
blood Pb levels.
Burns et al. (1999),
Bellinger et al. (1994b)
Section 4.3.4.1
Wasserman et al. (2001)
Blood Pb lifetime (to age 11-13 yr) average
mean: -14 ug/dL
Tooth Pb (age 6 yr) mean: 3.4 ug/g
Blood Pb lifetime (to age 4-5 yr) average
mean: 7.2 ug/dL
Chiodo et al. (2004)
Plusquellec et al. (2010)
Section 4.3.4.1
Concurrent blood Pb means (ages 5, 7 yr):
5, 5.4 ug/dl_
And, supporting
animal evidence
with relevant
exposures from a
few studies
Dietary lactational Pb exposure increased
emotionality, depression-like behavior in juvenile
rodents
Beaudin et al. (2007)
Dyatlov and Lawrence (2002)
Section 4.3.4.2
Blood Pb at PND52 after PND1-PND30
exposure: 13 ug/dL
Blood Pb at PND22 after PND1-PND22
exposure: 17 ug/dL
Some evidence
describing mode of
action
Pb-induced changes in HPA axis and dopaminergic
and GABAergic systems
Sections 4.3.2.3, 4.2.2.2,
4.3.10.4,4.3.10.8
4-317
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Table 4-17 (Continued): Summary of Evidence Supporting Nervous System Causal Determinations.
Attribute in Causal
Framework3
Key Evidence
References
Pb Biomarker Levels
Associated with Effects0
But, uncertainty
regarding potential
confounding
Epidemiologic associations found with adjustment
for SES, birth outcomes, parental education. Few
studies adjusted for or considered HOME score.
Uncertainty regarding potential confounding by
parental caregiving quality.
Burns et al. (1999).
Bellinger et al. (1994b).
Wasserman et al. (2001)
Chiodo et al. (2004)
Plusquellec et al. (2010)
Section 4.3.4.1
Auditory Function Decrements in Children - Likely Causal
Consistent findings
from a few high-
quality
epidemiologic
studies with relevant
blood Pb levels
Prospective study found associations of prenatal
(maternal), neonatal, yearly age 1 to 5 yr, lifetime
avg blood Pb levels with poorer auditory processing
in children at age 5 yr in Cincinnati.
Information was not provided on participation rates.
Results were adjusted for SES, HOME score, birth
outcomes, obstetrical complications, maternal
smoking. Several other factors considered.
Supporting evidence from cross-sectional studies for
increased hearing thresholds in children ages 4-19
yr participating in NHANES and HHANES in
association with higher concurrent blood Pb levels.
Examination of multiple exposures and outcomes in
NHANES and HHANES reduces likelihood of
selection bias specifically by blood Pb levels.
Studies considered potential confounding by age,
sex, race, income, parental education, nutritional
factors.
Dietrich et al. (1992)
Section 4.3.6.1
Blood Pb means: neonatal (10 day) 4.8 ug/dL,
yearly age 1 to 5 year 10.6-17.2 ug/dL, lifetime
(to age 5 yr) avg NR
Schwartz and Otto (1991. 1987)
Section 4.3.6.1
Blood Pb median:
HHANES: 8 ug/dl_
NHANES: NR
But, uncertainty
because of lack of
animal evidence in
juveniles and at
relevant exposures
Increased hearing thresholds, decreased auditory
evoked potentials in adult monkeys (ages 8-13
years) with lifetime dietary Pb exposure.
Changes in neurotransmitters in regions of brain
involved in auditory integration in mice with
gestational-lactational Pb exposure
Rice (1997)
Lilienthal and Winneke (1996)
Section 4.3.6.1
Fortune et al. (2009)
Section 4.3.10.8
Blood Pb after lifetime exposure: 33-150 ug/dL
Blood Pb after gestational-lactational
exposure: 8.2, 42 ug/dL
4-318
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Table 4-17 (Continued): Summary of Evidence Supporting Nervous System Causal Determinations.
Attribute in Causal
Framework3
Key Evidence
References
Pb Biomarker Levels
Associated with Effects0
Visual Function Decrements in Children - Inadequate
The available
evidence is of
insufficient quantity,
quality, or
consistency
High maternal 1s trimester blood Pb associated with
supernormal ERG in a study of children. No
potential confounding considered. Uncertain
functional relevance of ERG findings.
Rothenberg et al. (2002b)
Section 4.3.6.2
Group with prenatal maternal blood Pb 10.5-
32.5 ug/dl_
Higher than relevant Pb exposures did not affect
visual acuity in infant monkeys but decreased visual
acuity in adult monkeys.
Examination of retinal ERGs limited to adult rats
(see below for Visual Function Decrements in
Adults)
Laughlin et al. (2008)
Section 4.3.6.2
Section 4.3.6.2
Blood Pb after postnatal day 8-week 26
exposure: 35-40 ug/dL
Motor Function Decrements in Children - Likely Causal
Consistent findings
from a few high-
quality
epidemiologic
studies with relevant
blood Pb levels
Evidence from prospective studies for fine and gross
motor function decrements in children ages 4.5-17
yr in Cincinnati, Yugoslavia in association with
neonatal, earlier childhood, concurrent, lifetime avg
blood Pb levels.
High follow-up participation, no selective attrition in
Cincinnati cohort, higher loss-to-follow-up in
Yugoslavia cohort with lower maternal IQ, HOME.
Both studies adjusted for maternal IQ, parental
education, SES, HOME score
Studies used various, widely-used tests to assess
outcomes.
Mixed evidence for lower (concurrent) blood Pb
levels from cross-sectional studies that considered
several potential confounding factors.
Ris et al. (2004),
Dietrich et al. (1993a),
Bhattacharya et al. (2006:
1995),
Wasserman et al. (2000b)
Section 4.3.7
Blood Pb means
Cincinnati: neonatal (10 day) 4.8 ug/dL,
concurrent (age 6 yr) 11.6 ug/dL, lifetime (to
age 15-17 yr) avg 12.3 ug/dL, age 0-5 yr avg
11.7ug/dL
Yugoslavia: NR
Section 4.3.7
No decrease in motor function found in children
ages 6-10 yr in New England.
Decrease found in Inuit children ages 4-6 years in
Canada.
Surkan et al. (2007)
Despres et al. (2005).
Concurrent blood Pb mean: 2.2 ug/dL
Concurrent blood Pb mean: 4.1 ug/dL
4-319
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Table 4-17 (Continued): Summary of Evidence Supporting Nervous System Causal Determinations.
Attribute in Causal
Framework3
But, uncertainty
because of limited
animal evidence at
relevant exposures
Key Evidence13
Poorer balance (fell off rotarod more quickly) in adult
mice with gestational-lactation dietary Pb exposure
Inconsistent results with higher than relevant
postnatal exposures
References'3
Leasure et al. (2008)
Section 4.3.7
Section 4.3.7
Pb Biomarker Levels
Associated with Effects0
Peak blood Pb after gestational -lactational
exposure -10 ug/dL
Blood Pb after 2-7 week postnatal Pb
exposure >59 ug/dL
Cognitive Function Decrements in Adults - Likely Causal
Consistent findings
from high-quality
epidemiologic
studies with relevant
bone Pb levels
Prospective analyses in MAS cohort of white men
and BMS cohort of men and women of diverse
ethnicities found cognitive function decrements over
2 to 4 years in association with bone Pb levels.
Baseline participation rates differed but high follow-
up participation, not conditional on bone Pb levels.
Different potential confounding factors considered.
Adjustment for age and education in both cohorts,
SES in BMS, smoking and alcohol use in MAS.
Supporting cross-sectional evidence from MAS,
BMS, and also women in Nurses' Health Study with
adjustment for additional potential confounding
factors, including dietary factors, medications,
physical activity, comorbid conditions.
Associations with concurrent blood Pb level found in
men and women participating in NHANES in certain
age and genetic variant groups but not consistently.
Weisskopf et al. (2007b),
Bandeen-Roche et al. (2009)
Table 4-10 and
Section 4.3.2.7
Baseline means:
19, 20 ug/g tibia
25 ug/g patella
Nurses Healthy Study:
Weuve et al. (2009)
Section 4.3.2.7
Table 4-10
Krieg et al. (2009),
Krieg et al. (2010),
Krieg and Butler (2009)
Section 4.3.2.7
Concurrent means:
11 ug/g tibia
13 ug/g patella
Concurrent blood Pb means: 3-4 ug/dL
Some studies found associations in former and
current Pb-exposed workers with relevant blood Pb
levels and bone Pb levels. Most studies adjusted for
age and education. Some also adjusted for
depression and/or alcohol use; none considered
other occupational exposures such as Mn.
Outcomes assessed using various but widely used,
structured instruments.
Dorsey et al. (2006),
Schwartz et al. (2002)
Stewart et al. (2002)
Section 4.3.2.7
Concurrent blood Pb means: 31 ug/dL
Peak tibia Pb mean: 26 ug/g.
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Table 4-17 (Continued): Summary of Evidence Supporting Nervous System Causal Determinations.
Attribute in Causal
Framework3
Key Evidence
References
Pb Biomarker Levels
Associated with Effects0
Consistent evidence
in animals with
relevant exposures
Impaired learning, memory, and executive function
in monkeys ages 7-10 years with lifetime dietary Pb
exposures from birth.
Impaired learning in animals with lifetime dietary Pb
exposures starting in gestation.
Rice and Karpinski (1988).
Gilbert and Rice (1987)
Section 4.3.2.3
See above for
cognitive function in children.
Blood Pb at age 200 days: 15, 25 ug/dL
Evidence clearly
describes mode of
action
Same as above for cognitive function in children
But, important
uncertainty remains
Few prospective studies available to address
reverse causation
Uncertainty regarding potential residual confounding
of bone Pb results by age
Psychopathological Effects in Adults (e.g., symptoms of depression, anxiety, panic disorder) - Likely Causal
Consistent findings
from a few high-
quality
epidemiologic
studies with relevant
blood and bone Pb
levels
A few population-based cross-sectional studies
indicate associations of higher concurrent blood,
patella, or tibia Pb level with increased reporting of
symptoms of depression, anxiety, panic disorder in
adults.
Studies had high participation rates and examined
multiple exposures and outcomes and include
NHANES analysis of 1,987 adults.
Studies used widely used, structured instruments to
assess outcomes but not diagnosed conditions.
Bouchard et al. (2009),
Rhodes et al. (2003).
Rajan et al. (2008)
Section 4.3.5.1
Bouchard et al. (2009)
Section 4.3.5.1
Concurrent blood Pb mean: 6 ug/dL
Tibia Pb mean: 22 ug/g
Patella Pb mean: 31, 32 ug/g
Group with concurrent blood Pb > 2.11 ug/dL
And, some
supporting animal
evidence with
relevant exposures
Same as above for internalizing behaviors in
children.
But, uncertainty
regarding potential
confounding
Associations found with adjustment for education,
age, and SES. Mn not considered.
Also alcohol use and pack-years smoking in MAS.
However, uncertainty regarding generalizability
because of limited number of studies and potential
for residual confounding of bone Pb results by age.
Section 4.3.5.1
Rhodes et al. (2003).
Rajan et al. (2007)
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Table 4-17 (Continued): Summary of Evidence Supporting Nervous System Causal Determinations.
Attribute in Causal
Framework3
Key Evidence
References
Pb Biomarker Levels
Associated with Effects0
Auditory Function Decrements in Adults - Suggestive
Limited
epidemiologic
evidence with
relevant bone or
blood Pb levels, one
high-quality study
Prospective study found association between tibia
Pb level and higher rate of increase in hearing
threshold over 23 yr in MAS.
Population comprises only males, primarily white,
but study examines multiple exposures and
outcomes and has high follow-up participation.
Results adjusted forage, race, BMI, education,
diabetes, hypertension, smoking, occupational
noise.
Supporting evidence from case-control study finding
higher blood Pb levels in workers from various
occupations with hearing loss with adjustment for
age, smoking, alcohol consumption, years of noise
exposure, blood Mn, As, Se
Park et al. (2010)
Section 4.3.6.1
Tibia Pb mean: 22.5 ug/g, measured near end
of follow-up
Chuang et al. (2007)
Section 4.3.6.1
Concurrent blood Pb geometric mean in
cases: 10.7 ug/dL
Lack of animal
evidence at relevant
exposures
Increased hearing thresholds, decreased auditory
evoked potentials in adult monkeys (ages 8-13
years) with gestational, lactational, lifetime dietary
Pb exposure.
Rice (1997)
Lilienthal and Winneke (1996)
Laughlin et al. (2009)
Section 4.3.6.1
Blood Pb after gestational, lactational, or
lifetime exposure: 33-150 ug/dL
Visual Function Decrements in Adults - Inadequate
The available
evidence is of
insufficient quantity,
quality, or
consistency
Epidemiologic evidence limited to case-control study
of higher Pb in retinal tissue of macular
degeneration adult cases. Potential confounding not
considered.
Erie et al. (2009)
Median Pb in retina tissue
Cases: 12.0 ng/g
Controls: 8.0 ng/g
Higher than relevant Pb exposures decreases visual
acuity in adult monkeys but not in infant monkeys.
Rice et al. (1998)
Blood Pb: 50-115 ug/dL with lifetime exposure
to age 7-9 yr
Supernormal or subnormal ERGs in adult rats
depending on timing and dose of Pb exposure that
were linked to changes in rod cell proliferation,
retinal cell apoptosis, dopamine. Uncertain
functional relevance.
Fox et al. (2008: 1997),
Section 4.3.6.2
Blood Pb: 12, 24 ug/dL with gestational-
lactational exposure, 19 ug/dL with lactational
exposure
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Table 4-17 (Continued): Summary of Evidence Supporting Nervous System Causal Determinations.
Attribute in Causal
Framework3
Key Evidence
References
Pb Biomarker Levels
Associated with Effects0
Neurodegenerative Diseases (e.g., Alzheimer's disease, ALS, Parkinson's disease, Essential tremor) - Inadequate
The available
evidence is of
insufficient quantity,
quality, or
consistency
The few available case-control studies found higher
blood or bone Pb levels or whole-body lifetime
burden (PBPK modeling) in cases with Parkinson's
disease and essential tremor. Results for ALS are
mixed, association found with increased survival
time.
Studies subject to selection and recall bias and
reverse causation (for blood Pb studies) that could
produce artifactual associations.
Some studies consider potential confounding by
age, sex, smoking, alcohol use, education, and
activity levels, but not Mn exposures.
Case-control studies did not show associations
between occupational history of Pb exposure or
brain Pb levels and Alzheimer's disease.
Parkinson's disease:
Section 4.3.9.3
Weisskopf et al. (2010).
Coon et al. (2006)
Essential tremor:
Section 4.3.9.4
Louis et al. (2005: 2003),
Dogu et al. (2007)
ALS: Section 4.3.9.2
Kamel et al. (2008: 2005: 2002).
Vinceti et al. (1997).
Fang et al. (2010)
Graves et al. (1991),
Hariguchi et al. (2001),
Section 4.3.9.1
Parkinson's disease: group with tibia Pb levels
>16 ug/g, whole-body lifetime burden > 81
M9/9
Essential tremor: concurrent blood Pb means
2.7-3.5 ug/dL
ALS: Groups with concurrent blood Pb 3-
14 ug/dL, concurrent blood Pb means 2.4,
12.7ug/dL
Some evidence
describes mode of
action
Amyloid plaques found in brains of adult monkeys
and rodents with infancy-only Pb exposures.
Effects in monkeys at ages 20-23 yr when blood Pb
levels had returned to baseline.
In rodents, no effect with adult-only Pb exposure.
Increases in neuronal cell apoptosis found animals
with Pb exposure
Section 4.3.9.1
Wu et al. (2008a)
Basha et al. (2005)
Section 4.3.9.5
Blood Pb with exposure to age 400 days: peak
32-36 ug/dL
Blood Pb with lactational exposure: 46 ug/dL
"Described in detail in Table II of the Preamble.
bDescribes the key evidence and references contributing most heavily to causal determination and, where applicable, to uncertainties or inconsistencies. References to earlier sections
indicate where full body of evidence is described.
""Describes the Pb biomarker levels with which the evidence is substantiated and blood Pb levels in animals most relevant to this ISA.
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4.4 Cardiovascular Effects
4.4.1 Introduction
The 2006 Pb AQCD (U.S. EPA. 2006b) concluded that both epidemiologic and animal
toxicological studies support the relationship between increased Pb exposure and
increased cardiovascular effects, in particular, increased blood pressure (BP) and
increased incidence of arterial hypertension. Although fewer in number, epidemiologic
studies demonstrated associations of blood and bone Pb levels with other cardiovascular
diseases (CVDs) in adults, such as ischemic heart disease, cerebrovascular disease,
peripheral vascular disease, and CVD-related mortality. As the cardiovascular and renal
systems are intimately linked, cardiovascular effects can arise secondarily to Pb-induced
renal injury (Section 4.5). Toxicological studies also provided compelling evidence
supporting the biological plausibility for Pb-associated cardiovascular effects by
characterizing a number of the underlying mechanisms by which Pb exposure can lead to
human cardiovascular health effects. Such studies demonstrated that the Pb content in
heart tissue of animals reflects the increases in blood Pb levels (Lai etal.. 1991).
indicating that the cardiovascular morbidity associated with blood Pb levels may
represent the effects of the bioavailable Pb in the target tissue. The strongest evidence
supported the role of oxidative stress in the pathogenesis of Pb-induced hypertension.
Additionally, several toxicological studies characterized other pathways or cellular,
molecular, and tissue events promoting the Pb-induced increase in BP. These
mechanisms included inflammation, adrenergic and sympathetic activation, renin-
angiotensin-aldosterone system (RAAS) activation, vasomodulator imbalance, and
vascular cell dysfunction.
With regard to the concentration-response relationship, a meta-analysis of human studies
found that each doubling of blood Pb level (within a range extending from 1 at the lower
end to above 40 ug/dL at the upper end, measured concurrently in most studies) was
associated with a 1 mmHg increase in systolic BP and a 0.6 mmHg increase in diastolic
BP (Nawrot et al., 2002). On a population-wide basis, the estimated effect size could
translate into a clinically significant increase in the segment of the population with the
highest BP. In a moderately-sized population, a relatively small effect size thus has
important health consequences for the risk of sequelae of increased BP, such as stroke,
myocardial infarction, and sudden death. It was also noted that most of the reviewed
studies examining bone Pb levels, biomarkers of cumulative Pb exposure, also showed
increased BP (Cheng et al., 2001; Hu et al., 1996a) or increased hypertension with
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increasing bone Pb level (Lee et al.. 200 la). Across studies, over a range of bone Pb
concentrations (<1.0 to 96 ug/g), every 10 ug/g increase in bone Pb was associated with
increased odds ratios of hypertension between 1.28 and 1.86. Studies observed an
average increase in systolic BP of-0.75 mmHg for every 10 ug/g increase in bone Pb
concentration over a range of <1 to 52 ug/g.
With regard to etiologically-relevant timing of Pb exposure, toxicological evidence
demonstrated increases in BP after long-term (>4 weeks) Pb exposure. In epidemiologic
studies, cardiovascular outcomes were most often examined in cross-sectional studies
with one or a limited number of Pb biomarker measurements, so uncertainty exists as to
the specific Pb exposure level, timing, frequency, and duration that contributed to the
observed associations. While associations of adult bone Pb (particularly tibia Pb) with
health outcomes in adults are indicative of effects related to past or cumulative exposures,
interpretation of similar associations involving adult blood Pb levels, especially those
measured concurrently with outcomes, is complicated by the higher past exposures
generally observed in U.S. adults populations. Detailed interpretation of Pb in blood and
bone are provided in Sections 3.3 and 3.7.3. Briefly, higher past Pb exposures in adults
increased their bone Pb stores which contribute to current blood Pb levels through the
normal process of bone remodeling, as well as periods of increased bone remodeling and
loss (e.g., osteoporosis and pregnancy). Due to the long latency period for the
development of increased BP and CVD, associations of cardiovascular effects with low
concurrent blood Pb levels (e.g., population means 1.6-4 ug/dL) in adults may be
influenced by higher past Pb exposures (Section 3.4.1).
Past air Pb concentration and blood Pb data provide context for the cardiovascular
studies. Section 2.2 notes that the peak U.S. use of Pb anti-knock additives in automobile
gasoline and peak industrial emissions occurred between 1968 and 1972 and Pb in
gasoline was finally banned from use in 1996. Section 2.5 shows that the annual average
of the maximum 3-month average air Pb concentration for trends monitors across the
U.S. decreased from 1.3 ug/m3 in 1980 to 0.14 ug/m3 in 2010. The mean 2010 3-month
rolling average for non-source monitors (i.e., monitors not in close proximity to specific
sources) was an order of magnitude lower than the 2010 trends site average. Studies of
blood Pb during this time period of peak air Pb concentrations in the U.S. (1968-1980)
show higher blood Pb levels than 2009-2010 levels presented in Section 3.4. Among
subjects ages 1-74 years examined in the U.S. NHANES II study (1976-1980) by Pirkle
et al. (1994). the geometric mean blood Pb level was 12.8 ug/dL. When stratified by age,
studies report geometric mean blood Pb levels in adults (age 20-79 years in 1968-1980)
of 13.1 - 16.2 ug/dL (Pirkle et al.. 1994; Tepper and Levin. 1975) and in children (age
1 mo - 6 years in 1970-1980) of 15.0-30.4 ug/dL (Pirkle et al.. 1994: Billick et al.. 1979:
Fineetal.. 1972).
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This section reviews the published studies pertaining to the cardiovascular effects of Pb
exposure in humans, experimental animals, isolated vascular tissues, and cultured
vascular cells. With the large and strong existing body of evidence serving as the
foundation, emphasis was placed on studies published since the 2006 Pb AQCD.
Epidemiologic and toxicological studies continued to augment the evidence for increases
in BP and hypertension development associated with long-term Pb exposure and
expanded the evidence for the biological pathways of these effects. Epidemiologic studies
strengthened the evidence for associations between Pb biomarkers and cardiovascular
effects after adjusting for potential confounding factors such as age, SES, diet, alcohol
use, BMI, comorbidities, and smoking. Emphasis was placed on studies that had
extensive consideration for confounding and prospective study designs. The
epidemiologic evidence was substantiated with results from several available prospective
studies demonstrating the directionality of effects by indicating associations between Pb
biomarkers and the subsequent incidence of cardiovascular health effects.
4.4.2 Blood Pressure and Hypertension
4.4.2.1 Epidemiology
The most commonly used indicator of cardiovascular morbidity was increased BP and its
derived index, hypertension. Hypertension in these studies was defined as diastolic and/or
systolic BP above certain cut-points or use of anti-hypertensive medicines. The BP cut-
points were established by reference to informed medical opinion, but BP cut-points
defining hypertension have been lowered over time, as medical knowledge has improved.
Consequently, different studies using "hypertension" as a cardiovascular outcome may
have assigned different cut-points, depending on the year and location of the study and
the individual investigator. All of the recent studies in the current review used similar
criteria for hypertension (e.g., systolic BP at or above 140, diastolic BP at or above 90, or
use of anti-hypertensive medications). Studies in the medical literature show that elevated
BP is associated with increased risk of CVD including coronary disease, stroke,
peripheral artery disease, and cardiac failure. Coronary disease (i.e., myocardial
infarction, angina pectoris, and sudden death) is the most lethal sequelae of hypertension
(Ingelsson et al., 2008; Chobanian et al., 2003; Pastor-Barriuso et al., 2003; Prospective
Studies Collaboration. 2002; Kannel. 2000a. b; Neatonetal. 1995).
Earlier, U.S. EPA (1990a) reviewed the then available studies examining Pb exposure
and BP and hypertension outcomes which included evaluation of several studies
conducting analysis of the data in NHANES II (1976-80). They noted that across a range
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of 7 to 34 (ig/dL, no evident threshold was found below which concurrent blood Pb was
not significantly related to BP. U.S. EPA (1990a) concluded that a small but positive
association exists between blood Pb levels and increases in BP. Quantitatively, the
relationship appeared to hold across a wide range of blood Pb values and, furthermore, an
estimated mean increase of 1.5-3.0 mmHg in systolic BP appeared to occur for every
doubling of blood Pb concentration in adult males and less than 1.0-2.0 mmHg for adult
females. U.S. EPA (1990a) further concluded that the plausibility of these relationships
observed in epidemiologic studies of human populations being of a causal nature is
supported by controlled experimental animal studies demonstrating increased BP clearly
attributable to Pb. Subsequently, the 2006 Pb AQCD (U.S. EPA. 2006b) reviewed the
literature examining Pb exposure and effects of BP and hypertension, published after the
1990 document as discussed in Section 4.4.1.
Several recent general population and occupational cohort and cross-sectional studies
strengthened the evidence for associations of blood and bone Pb levels with measures of
BP (Figure 4-17 and Table 4-18) and with the prevalence and incidence of hypertension
(Figure 4-18 and Table 4-19). Further, recent studies expanded evidence, finding
differences in association among racial/ethnic groups, perceived stress, diet, and genetic
variants, and thus, identified populations potentially at increased risk of Pb-associated
cardiovascular effects.
In a cross-sectional analysis, Martin et al. (2006) examined the associations of concurrent
blood and tibia Pb levels with BP and hypertension in a large, community-based study of
older adults (N = 964, age ranging from 50 to 70 years) in Baltimore, MD. Although
cross-sectional in design, a key strength of this study was the extensive consideration of
potential confounding variables. Four models evaluated associations for BP and
hypertension. The base model included age, sex, BMI, sodium intake, potassium intake,
total cholesterol, time of day, testing technician, and hypertensive medication use. Other
models added SES, race/ethnicity, or both as covariates. Blood Pb but not tibia Pb level
was a strong predictor of BP in all models; a 1 (ig/dL increase in concurrent blood Pb
level was associated with an approximately 1 mmHg increase in systolic BP and an
approximately 0.5 mmHg increase in diastolic BP. Tibia Pb but not blood Pb was
associated with hypertension in logistic regression models. The authors applied
propensity analysis to their models to better account for the effect of other risk factors for
hypertension such as race/ethnicity, age, and SES that were strongly associated with tibia
Pb level. The propensity score analysis and model adjustment did not substantially
change the numerical findings and conclusions (e.g., tibia Pb and hypertension were
positively associated independently of race/ethnicity and SES), indicating that neither
SES nor race/ethnicity confounded the association between tibia Pb level and
hypertension. No evidence for effect modification by race/ethnicity was found either.
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Martin et al. (2006) concluded that Pb in blood has a short term effect on BP and that Pb
contributes to hypertension risk as a function of cumulative, chronic exposure (as
represented as bone Pb in this population). While different aspects of Pb exposure may
contribute differentially to increases in BP and hypertension, it is important to note that
concurrent blood Pb levels in adults also reflect cumulative Pb exposure. Thus, its
association with BP may not reflect a short term effect but may instead reflect an effect of
cumulative Pb exposure.
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Reference
Martin etal. (2006)
Glenn etal. (2006)
Weaveretal. (2008)
Scinicarielloetal. (2010)
Martin etal. (2006)
Glenn etal. (2006)
Peters etal. (2007)
Elmarsafawy et al. (2006)
Weaveretal. (2008)
Petersetal.(2007)
Martin etal. (2006)
Scinicarielloetal. (2010)
Martin etal. (2006)
Perlstein etal. (2007)
Zhang etal. (2010)
Population
Adults, Baltimore, MD
Korean Pb Workers
Korean Pb Workers
NHANES III - Blacks
NHANES III - Mexicans
Adults, Baltimore, MD
Korean Pb Workers
NASmen, High Stress
NASmen, Low Stress
NASmen, High calcium
NASmen, Low calcium
Korean Pb Workers
NASmen, High Stress
Adults, Baltimore, MD
NHANES III -Whites
NHANES III - Blacks
NHANES III - Mexicans
Adults, Baltimore, MD
NASmen
NASmen
NASmen
NASmen
NASmen
NASmen
NASmen
NASmen
NASmen, HFE Wild-type
NASmen, HFE H63D
NASmen, HFE C282Y
NASmen, Any HFE variant
NASmen, HFE Wild-type
NASmen, HFE H63D
NASmen, HFE C282Y
NASmen, Any HFE variant
Pb Distribution a
2.9(2.0,4.4)
27.2(19.3,28.3)
1.4(0.6,3.6)
2.0(1.0,3.9)
15.7(10.5,23.5)
18.1(12.2,26.9)
18.1(12.2,26.9)
21.6(12.0)
21.6(12.0)
74.3(67.3,82.0)
26.9(18.4,39.3)
2.9(2.0,4.4)
1.6(0.8,3.3)
1.4(0.6,3.6)
2.0(1.0,3.9)
15.7(10.5,23.5)
12.4(4.4)
7.4(0.6)
5.4 (0.5)
3.9(0.3)
40.9 (14)
29.4(2.2)
18.9(1.4)
14.1(1.4)
18(12,27)
19(14,26)
20(14,27)
19(14,27)
26(17,34)
27(19,37)
25(17,37)
26(18,37)
PbBiomarker
Blood Pb
Blood Pb (concurrent) — I
Blood Pb (longitudinal) — 1
Blood Pb —
In Blood Pb
In Blood Pb
Tibia Pb — V
Tibia Pb (historical) — C
Tibia Pb
Tibia Pb — C
Tibia Pb
Tibia Pb
Patella Pb — <
Patella Pb
Blood Pb
In Blood Pb
In Blood Pb
In Blood Pb
Tibia Pb
m
Blood Pb
Blood Pb *
Blood Pb 9
Tibia Pb
Tibia Pb
Tibia Pb ^ O
Tibia Pb
Tibia Pb —
Tibia Pb
Tibia Pb
Patella Pb -<
Patella Pb
Patella Pb
Patella Pb
SBP
— •—
t—
~
— • —
r»-
> —
—
— o —
— o —
> —
~°~
DBP
• —
-•—
D-
PP
^^^^^^
o —
f\
>-
— o —
o-
-4
Change in BP (mmHg 95% Cl) per
lu.g/dL increase in blood Pb or 10 u.g/g bone Pb
aPb distributions present the median (IQR), which were estimated from the mean and SD assuming a normal distribution.
bEffect estimates were standardized to 1 ug/dL blood Pb or 10 ug/g bone Pb.
Note: In general, results are categorized by specific BP parameter, then by Pb biomarker. For categories with multiple studies, the
order of the studies follows the order of discussion in the text. Results display associations (95% Cl) of a 1 ug/dL increase in blood
Pb level (closed circles) or 10 ug/g increase in bone Pb (open circles) with systolic BP (SBP; blue), diastolic BP (DBP; red), and
pulse pressure (PP; purple) in adults.
Figure 4-17 Associations of blood and bone Pb levels with systolic BP,
diastolic BP, and pulse pressure in adults.
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Table 4-18 Additional characteristics and quantitative data for associations of blood and bone Pb with BP
measures for studies presented in Figure 4-17.
Study
Study Population /
Methodology
Parameter Pb Data
Statistical Analysis
Effect Estimate
P (95% Cl)
Martin et al.
(2006)
Cross-sectional
964 men and women,
ages 50-70 yr, 40%
African American, 55%
White, 5% other, in
Baltimore, MD
BP Concurrent Mean
Blood Pb:
Mean (SD): 3.5 (2.3) ug/dl_
African American: 3.4 (2.3)
White: 3.5 (2.4)
Tibia Pb:
Mean (SD): 18.8
(12.4) ug/g
African American: 21.5
(12.6)
White: 16.7(11.9)
Extensive analysis of potential
confounding factors. Multiple
linear regression base models
adjusted for age, sex, BMI,
antihypertensive medication use,
dietary sodium intake, dietary
potassium intake, time of day,
testing technician, serum total
cholesterol. SES, race/ ethnicity
also included in models that are
presented in Figure 4-17, and
tabulated here.)
Blood Pb:
SBP: 1.05(0.53,1.58)
DBP: 0.53(0.25, 0.81)
mmHg per ug/dL blood Pb
Tibia Pb:
SBP: 0.07 (-0.05, 0.14)
DBP: 0.05 (-0.02, 0.08)
mmHg per ug/g bone Pb
Glenn et al.
(2006)
Longitudinal
575 Pb exposed
workers, ages 18-65 yr,
in South Korea
(10/1997-6/2001)
BP Blood Pb mean (SD):
Visit 1: 20.3 (9.6), Women
Visit 2: 20.8 (10.8), Women
Visits: 19.8(10.7), Women
Visit 1: 35.0(13.5), Men
Visit 2: 36.5(14.2), Men
Visits: 35.4(15.9), Men
Multivariable models using GEE
were used in longitudinal
analyses. Models were adjusted
for visit number, baseline age,
baseline age squared, baseline
lifetime alcohol consumption,
baseline BMI, sex, baseline BP
lowering medication use, alcohol
consumption.
Model 1 (short-term)
Blood Pb (longitudinal):
0.09(0.01, 0.16)
Blood Pb (concurrent):
0.08 (-0.01, 0.16)
Model 4 (short and longer-
term)
Blood Pb (longitudinal):
0.09(0.01, 0.16)
Blood Pb (concurrent):
0.10(0.01, 0.19)
mmHg per 10 ug/dL blood Pb
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Table 4-18 (Continued): Additional characteristics and quantitative data for associations of blood and bone Pb with BP
measures for studies presented in Figure 4-17.
Study
Weaver et al.
(2008)
Study Population /
Methodology
Cross-sectional
652 current and former
Pb workers in
South Korea
(12/1999-6/2001)
Same cohort as Glenn
et al. (2006)
Parameter Pb Data
BP Concurrent Blood Pb:
Mean (SD):
30.9(16.7) ug/dL
Concurrent Patella Pb:
Mean (SD):
75.1 (101.1) ug/g
Statistical Analysis
Linear regression model adjusted
forage, sex, BMI, diabetes,
antihypertensive and analgesic
medication use, Pbjob duration,
work status, tobacco and alcohol
use
Effect Estimate
P (95% Cl)
SBP
Patella Pb:
0.0059 (-0.008, 0.02)a
Blood Pb:
0.1007(0.02, 0.18)a
mmHg per 1 ug/dL blood Pb
or 1 ug/g patella Pb
Interaction between blood
Pb/patella Pb with ALAD and
vitamin D receptor
polymorphisms not
significant.
Perlstein et al.
(2007)
Cross-sectional
593 elderly men (mostly
white) enrolled in MAS
in Greater Boston, MA
area (1991-1997)
PP Blood Pb:
Overall mean (SD): 6.12
(4.03) ug/dL
Quintile means (SD):
Q1:2.3(0.8)ug/dL
Q2: 3.9 (0.3) ug/dL
Q3: 5.4 (0.5) ug/dL
Q4: 7.4 (0.6) ug/dL
Q5: 12.4 (4.4) ug/dL
Tibia Pb:
Overall median: 19 ug/g
Quintile means (SD):
01:7.4(3.2) ug/g
Q2: 14.1 (1.4) ug/g
Q3: 18.9(1.4) ug/g
Q4: 24.9 (2.2) ug/g
Q5: 40.9 (14) ug/g
BP association assessed using
Spearman correlation
coefficients.
PP association (adjusted mean
difference) assessed using
multiple linear regression model
adjusted for age, height, race,
heart rate, waist circumference,
diabetes, family history of
hypertension, education level
achieved, smoking, alcohol
intake, fasting plasma glucose,
and ratio of total cholesterol to
HDL cholesterol
PP
4.2 (1.9, 6.5) mmHg higher in
men with tibia Pb >19 ug/g
(median) compared with men
with tibia Pb below the
median
Blood Pb (mean difference):
Q5:-1.49 (-4.93, 1.94)
Q4:-1.39 (-4.94, 2.15),
Q3: -2.56 (-5.78, 0.67)
Q2: -4.37 (-7.88-0.86)
Q1: Referent group
Tibia Pb (mean difference):
Q5: 2.58 (-1.15, 6.33)
Q4: 2.64 (-0.93, 6.21)
Q3: -0.73 (-4.27, 2.82)
Q2: -3.02 (-6.48, 0.44)
Q1: Referent group
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Table 4-18 (Continued): Additional characteristics and quantitative data for associations of blood and bone Pb with BP
measures for studies presented in Figure 4-17.
Study
Peters et al.
(2007)
Elmarsafawy et al.
(2006)
Study Population /
Methodology
Longitudinal and
Cross-sectional
513 elderly men (mostly
white, mean age 67 yr)
enrolled in MAS in
Greater Boston, MA
area
Cross-sectional
471 elderly men (mostly
white, mean age 67 yr)
enrolled in MAS in
Greater Boston, MA
area
Parameter Pb Data
BP Tibia Pb:
mean (SD):
21.5 (13.4) ug/g
Patella Pb:
Mean (SD):
31.5(19.3) |jg/g
BP Blood Pb:
Mean (SD):
6.6 (4.3) |jg/dL
Tibia Pb:
Mean (SD):
21.6 (12.0) ug/g
Patella Pb:
Mean (SD):
31 7 (18 3) uq/q
\s i.i y i \j .\s j |_i\^f\^
Statistical Analysis
Logistic and linear regression
models adjusted for age, age
squared, sodium, potassium, and
Ca2+ intake, family history of
hypertension, BMI, educational
level, pack-years of smoking,
alcohol consumption, and
physical activity
Linear regression models
adjusted for age, BMI, family
history of hypertension, history of
smoking, dietary sodium intake,
and cumulative alcohol ingestion
Lack of consideration for potential
confounding by SES-related
variables.
Effect Estimate
P (95% Cl)
SBP
Tibia Pb/ High Stress:
3.57(0.39,6.75)
Tibia Pb/ Low Stress:
0.21 (-1.70, 1.29)
per SD increase in tibia Pb
Patella Pb/ High Stress:
2.98 (-0.12, 6.08)
per SD increase in patella Pb
Patella Pb/ Low Stress:
NR
SBP
Tibia Pb
High Ca2+ group
(>800 mg/day):
0.40.(0.11, 0.70)
Low Ca group
(<800 mg/day):
0.19(0.01, 0.37)
mmHg per ug/g tibia Pb
4-332
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Table 4-18 (Continued): Additional characteristics and quantitative data for associations of blood and bone Pb with BP
measures for studies presented in Figure 4-17.
Study Population /
Study Methodology Parameter Pb Data
Zhang et al. Cross-sectional PP Wild type HFE
(2010a) 61 9 elderly men (mostly Tibia Pb:
white, mean age 67 yr) Median (IQR):
enrolled in the MAS in 8 (12-27) pg/g
Greater Boston, MA Patella Pb-
area Median (IQR):
26(1 7-37) |jg/g
C282Y HFE
Tibia Pb:
Median (IQR):
20(1 4-27) |jg/g
Patella Pb:
Median (IQR):
25(1 7-37) |jg/g
H63D HFE
Tibia Pb:
Median (IQR):
19(1 4-26) |jg/g
Patella Pb:
Median (IQR):
27(1 9-37) |jg/g
Statistical Analysis
Linear mixed effects regression
models with repeated
measurements adjusted for age;
education; alcohol intake;
smoking; daily intakes of Ca2+,
sodium, and potassium; total
calories; family history of
hypertension; diabetes; height;
heart rate; high-density
lipoprotein (HDL); total
cholesterol: HDL ratio; and waist
circumference
Effect Estimate
P (95% Cl)
PP
mmHg per 13 ug/g Tibia Pb:
Wild Type HFE:
0.38(0,1.96)
H63D HFE:
3.30(0.16,6.46)
C282Y HFE:
0.89 (0, 5.24)
Any HFE variant:
2.90(0.31, 5.51)
mmHg per 19 ug/g Patella
Pb:
Wild Type HFE:
0.26(0, 1.78)
H63D HFE:
2.95 (0, 5.92)
C282Y HFE:
0.55(0, 1.66)
Any HFE variant:
2.83 (0.32,5.37)
4-333
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Table 4-18 (Continued): Additional characteristics and quantitative data for associations of blood and bone Pb with BP
measures for studies presented in Figure 4-17.
Study
Study Population /
Methodology
Parameter Pb Data
Statistical Analysis
Effect Estimate
P (95% Cl)
Scinicariello et al.
(2010)
Cross-sectional
6,016 NHANES III
(1988-1994) participants
age > 17 yr
BP Concurrent Blood Pb:
Mean (SE):
Overall:
2.99 (0.09) |jg/dL
Non-Hispanic Whites:
2.87(0.09)
Non-Hispanic Blacks
3.59(0.20)
Mexican American
3.33(0.11)
Multivariable linear regression of
log-transformed blood Pb level
adjusted for age, sex, education,
smoking status, alcohol intake,
BMI, serum creatinine levels,
serum Ca2+, glycosylated
hemoglobin, and hematocrit
SBP
Non-Hispanic whites:
1.05(0.32, 1.78)
Non-Hispanic blacks:
2.55(1.59, 3.51)
Mexican Americans:
0.84 (-0.06, 1.74)
DBP
Non-Hispanic whites:
-0.14 (-1.1, 0.82)
Non-Hispanic blacks:
1.99(1.13,2.85)
Mexican Americans:
0.74 (-0.005, 1.48)
mmHg per unit increase in
logarithmic (In)-transformed
blood Pb
Significant interactions with
blood Pb and ALAD genotype
observed in relation to SBP
for non-Hispanic whites and
non-Hispanic blacks
4-334
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Table 4-18 (Continued): Additional characteristics and quantitative data for associations of blood and bone Pb with BP
measures for studies presented in Figure 4-17.
Study
Study Population /
Methodology
Parameter Pb Data
Statistical Analysis
Effect Estimate
P (95% Cl)
Navas-Acien et al.
(2008)b
Longitudinal and
Cross-sectional
Meta-analysis of studies
using bone Pb as an
exposure metric and BP
as the outcome
(8 studies)
BP
Inverse variance weighted
random-effects meta-analyses
BP Pooled Estimates
mmHg per 10 ug/g Tibia Pb
Prospective SBP
0.33 (-0.44, 1.11)
Cross-sectional SBP
0.26 (0.02, 0.50)
Cross-sectional DBP
0.02 (-0.15, 0.19)
Hypertension
per 10 ug/g patella Pb
Cross-sectional hypertension
OR: 1.04(1.01, 1.07)
Meta-analysis estimate
hypertension
OR: 1.04(0.96, 1.12)
Yazbeck et al.
(2009)°
Cross-sectional
971 pregnant women,
ages 18-45 yr,
in France
BP Midpregnancy Blood Pb:
PIH group mean (SD): 2.2
(1.4)
No PIH group mean (SD):
1.9(1.2)
Multivariable logistic regression
models adjusted for maternal
age; Cd, Mn, and Se blood levels;
hematocrit; parity; BMI;
pregnancy weight gain;
gestational diabetes; educational
level; SES; geographic residence;
and smoking status and alcohol
consumption before and during
pregnancy
Log-transformed blood Pb at
mid-pregnancy
SBP:
r= 0.08; p = 0.03
DBP:
r = 0.07; p = 0.03
Significant correlations also
observed after 24 weeks of
gestation and after 36 weeks
of gestation.
a95% CIs estimated from given p-value.
bReference not included in Figure 4-17. because it is a meta-analysis.
""Reference not included in Figure 4-17. because only correlations were reported
4-335
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Reference
Study Location Population Blood Pbc (ug/dL) Comparison
Martin etal. (2006) Baltimore, MD
Elmarsafawy etal. (2006) NAS
Yazbeck etal. (2009) France
Muntneretal. (2005)
Scinicarielloetal. (2010) NHANES III
1998-1994
Park etal. (2009)
Elmarsafawy etal. (2006) NAS
Martin etal. (2006) Baltimore, MD
Peters etal. (2007) Boston, MA
Petersetal.(2007) Boston, MA
Elmarsafawy etal. (2006) NAS
LowCalcium
High Calcium
PregnantWomen
Non-Hispanic Whites
Non-Hispanic Blacks
Mexican Americans
Non-Hispanic Whites
Non-Hispanic Blacks
Mexican Americans
Overall
3.5(2.3)
6.6(4.3)
6.6(4.3)
<1.20
1.2-1.7
1 71-2 30
>2.30
<1.06
1.06-1.63
1.63-2.47
>2.47
<1.06
1.06-1.63
1.63-2.47
>2.47
<1.06
1.06-1.63
1.63-2.47
>2.47
0.7-1.4
1.5-2.3
2.4-3.7
3.8-52.9
2.4-3.7
0.7-1.4
1.5-2.3
2.4-3.7
3.8-52.9
2.4-3.7
0.7-1.4
1.5-2.3
2.4-3.7
3.8-52.9
2.4-3.7
3.52(0.10)
perl ug/dL
perl ug/dL
perl ug/dL ;,
Reference
Reference
Q2vQl - —
Q3 v Ql -.' /-
Q4 v Ql -i'^-
Reference ;
Q2vQl
Q3 v Ql -1 ,
Q4 v Ql J — ;
Reference
Q2vQl — .,
Q4vQl 1 1
Reference '
Q2 v Ql — H ,
Reference ;
Q2 v QI ; — e
ALAD^vl11 ' ^
Reference ;
Q3 v Ql i "
perl ug/dL Ł">
White Men
Black Men
White Women
BlackWomen
Men <50
Men>50
Women <50
Women >50
LowCalcium
High Calcium
High Stress
High Stress
LowCalcium
High Calcium
Tibia Pb ' (ug/g)
21.6(12.0)
21.6(12.0)
18.8(12.4)
21.5(13.4)
Patella Pb ' (ug/g)
31.5(19.3)
31.7(18.3)
31.7(18.3)
per 10 ug/g
per 10 ug/g
per 10 ug/g
per 10 ug/g
per 10 ug/g
per 10 ug/g
per 10 ug/g
23456
Odds Ratio (95% Cl)
Note: Studies are categorized by Pb biomarker (blood Pb [closed circles] or bone Pb [open circles]) for hypertension prevalence and
incidence3. Within each category, studies generally are presented in order of discussion in the text.
"The outcomes plotted are hypertension prevalence with the exception of Yazbeck et al. (2009) which measured pregnancy induced
hypertension and Peters et al. (2007) which measured hypertension incidence.
bb: ALAD2 vs. 1 indicates comparison between ALAD 2 carriers (e.g., ALAD1-2 and ALAD2-2) and ALAD 1 homozygotes
(e.g., ALAD1-1).
°c: Effect estimates were standardized to a 1 |jg/dL increase in blood Pb (closed circles).
dd: Effect estimates were standardized to a 10 |jg/g increase in bone Pb (open circles).
Figure 4-18 Odds ratios (95% Cl) for associations of blood Pb and bone Pb
with hypertension prevalence and incidence.
4-336
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Table 4-1 9
Study
(Ordered as they
appear in the text)
Martin et al.
(2006)
Weaver et al.
(2008f
Peters et al.
(2QQ7)
Additional characteristics and quantitative data for results presented in Figure 4-18 for associations
of blood and bone Pb with hypertension measures.
Study Population
and Methodology
Cross-sectional
964 men and women,
ages 50-70 yr,
40% African
American, 55%
White,
5% other,
in Baltimore, MD
Cross-sectional
652 current and
former Pb workers in
South Korea
(12/1999-6/2001)
Longitudinal
513 elderly men
(mostly white, mean
age 67 yr) enrolled in
MAS in Greater
Boston, MA area
Parameter
Hypertension
(current use of
antihypertensive
medication, mean
SBP > 140 mmHg or
DBP >90 mmHg)
Hypertension
(mean SBP > 140
mmHg, DBP > 90
mmHg; and/or use of
antihypertensive
medications; or
physician diagnosis)
Hypertension
(mean SBP >140
mmHg, DBP >90
mmHg; or physician
diagnosis)
Pb Data
Blood Pb:
Mean (SD):
3.5 (2.3) ug/dL
Tibia Pb:
Mean (SD):
18.8 (12.4) ug/g
Blood Pb:
Mean (SD):
31.9 (14.8) |jg/dL
Patella Pb:
Mean (SD):
37.5 (41 .8) ug/g
Tibia Pb:
Mean (SD):
21 .5 (13.4) ug/g
Patella Pb:
Mean (SD):
Q*1 CMOQXiirW^
oi .0 (iy.o) ug/g
Statistical Analysis
Logistic regression models
adjusted for age, sex, BMI,
antihypertensive medication
use, dietary sodium intake,
dietary potassium intake, time
of day, testing technician, and
serum homocysteine
Lack of consideration for
potential confounding by SES-
related variables.
Logistic regression models
adjusted for age, sex, BMI,
diabetes, antihypertensive and
analgesic medication use, Pb
job duration, work status,
tobacco and alcohol use
Cox proportional hazards
models adjusted for age, age
squared, sodium, potassium,
and Ca2+ intake, family history
of hypertension, BMI,
educational level, smoking,
alcohol consumption, baseline
SBP and DBP, and physical
activity
Effect Estimate (95% Cl)
Blood Pb level:
OR=1.02(0.87, 1.19)
Tihia Ph-
I lUld r U.
OR=1.24(1.05, 1.47)
Quantitative results not reported.
None of the examined Pb
exposure metrics (blood, patella,
and logarithmic (In)-transformed
patella) were significantly
associated with hypertension
Risk of Hypertension Incidence
High Stress
RR=2.66 (1.43, 4.95) per SD
increase in tibia Pb
RR=2.64 (1.42, 4.92) per SD
increase in patella Pb
4-337
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Table 4-19 (Continued): Additional characteristics and quantitative data for results presented in Figure 4-18 for associations
of blood and bone Pb with hypertension measures.
Study
(Ordered as they
appear in the text)
Elmarsafawy et
al. (2006)
Study Population
and Methodology
Cross-sectional
471 elderly men
(mostly white, mean
age 67 yr) enrolled in
MAS in Greater
Boston, MA area
Parameter
Hypertension
(mean SBP > 160
mmHg, DBP > 95
mmHg; and/or
physician diagnosis
with current use of
antihypertensive
medications)
Pb Data
Blood Pb:
Mean (SD):
6.6 (4.3) ug/dl_
Tibia Pb:
Mean (SD):
21 .6 (12.0) ug/g
Patella Pb:
Mean (SD):
31.7 (18.3) ug/g
Statistical Analysis
Logistic regression models
adjusted for age, BMI, family
history of hypertension, history
of smoking, dietary sodium
intake, and cumulative alcohol
ingestion
Effect Estimate (95% Cl)
Low Ca2+ group
(<800 mg/day):
Blood Pb OR:
1.07(1.00, 1.15)
Tibia Pb OR:
1.02(1.00, 1.04)
Patella Pb OR:
1I"M / *1 nn *1 MO\
.01 (1.00, 1.03)
High Ca2+ group
(>800 mg/day):
Blood Pb OR:
1.03(0.97, 1.11)
Tibia Pb OR:
1.01 (0.97, 1.04)
Patella Pb OR:
1.01 (0.99, 1.03)
Per ug/dL blood Pb
or ug/g tibia or patella Pb
Yazbeck et al.
(2009f
Cross-sectional
971 pregnant women,
ages 18-45 yr, in
France
PIH
(SBP > 140 mmHg or
DBP > 90 mmHg
after the 22nd week
of gestation)
Blood Pb (ug/dL)
PIH group mean
(SD): 2.2 (1.4)
No PIH group mean
(SD):
1.9(1.2)
Q1: <1.20
Q2: 1.20-1.70
Q3: 1.71-2.30 Q4:
>2.30
Multivariable logistic regression
models adjusted for maternal
age, Cd, Mn, and Se blood
levels, parity, hematocrit, BMI,
gestational diabetes,
educational levels, SES,
geographic residence, and
smoking status during
pregnancy
PIH
Blood Pb
OR=3.29(1.11, 9.74)
per 1 unit increase in log maternal
blood Pb level
Q1: Reference group
Q2: OR=1.84 (0.77, 4.41)
Q3: OR=2.07 (0.83, 5.13)
Q4: OR=2.56 (1.05, 6.22)
4-338
-------
Table 4-19 (Continued): Additional characteristics and quantitative data for results presented in Figure 4-18 for associations
of blood and bone Pb with hypertension measures.
Study
(Ordered as they
appear in the text)
Study Population
and Methodology
Parameter
Pb Data
Statistical Analysis
Effect Estimate (95% Cl)
Muntner et al.
(2005)
Cross-sectional
9,961 NHANES
(1999-2002)
participants, ages >18
yr
Hypertension
(current use of
antihypertensive
medication, SBP
> 140 mmHg, or DBP
>90 mmHg)
Concurrent Blood Pb
(ug/dL)
Overall Mean (95%
Cl):
1.64(1.59-1.68)
Q1: <1.06
Q2: 1.06-1.63
Q3: 1.63-2.47
Q4:>2.47
Multivariable logistic regression
models adjusted for age, sex,
diabetes mellitus, BMI,
cigarette smoking, alcohol
consumption, high school
education, and health
insurance status
Monotonic increase in OR
across blood Pb level groups.
Non-Hispanic white:
Q1: Reference group
Q2: OR=1.12(0.83, 1.50)
Q3: OR=1.03(0.78, 1.37)
Q4: OR=1.10(0.87, 1.41)
Non-Hispanic black
Q1: Reference group
Q2: OR=1.03(0.63, 1.67)
Q3: OR=1.12(0.77, 1.64)
Q4: OR=1.44 (0.89, 2.32)
Mexican American
Q1: Reference group
Q2: OR=1.42 (0.75, 2.71)
Q2: OR=1.48 (0.89, 2.48)
Q3: OR=1.54 (0.99, 2.39)
p fortrend=0.04
4-339
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Table 4-19 (Continued): Additional characteristics and quantitative data for results presented in Figure 4-18 for associations
of blood and bone Pb with hypertension measures.
Study
(Ordered as they
appear in the text)
Scinicariello et
al. (2010)
Study Population
and Methodology
Cross-sectional
6,016 NHANES III
(1988-1994)
participants, ages
>17yr
Parameter
Hypertension
(current use of
antihypertensive
medication, SBP
> 140 mmHg, or DBP
>90 mmHg)
Pb Data
Concurrent Blood Pb
(ug/dL)
Mean (SE):
2.99 (0.09)
Q1: 0.7-1.4
Q2: 1.5-2.3
Q3: 2.4-3.7
Q4: 3.8-52.9
Statistical Analysis
Multivariable logistic regression
model adjusted for
race/ethnicity, age, sex,
education, smoking status,
alcohol intake, BMI, serum
..... _ 9+
creatmme levels, serum Ca ,
glycosylated hemoglobin, and
hematocrit
Effect Estimate (95% Cl)
Non-Hispanic whites:
Q1: Reference group
Q2: POR=1.21 (0.66,2.24)
Q3: POR=1. 57 (0.88, 2.80)
Q4: POR=1. 52 (0.80, 2.88)
ALAD1 -2/2-2:
POR= 0.76 (0.17, 3.50)
ALAD-1: Reference group
Mean (SE):
Non-Hispanic Whites:
2.87(0.09)
Non-Hispanic Blacks:
3.59(0.20)
Mexican American:
3.33(0.11)
Non-Hispanic blacks:
Q1: Reference
Q2: POR=1.83(1.08, 3.09)
Q3: POR=2.38 (1.40, 4.06)
Q4: POR=2.92(1.58, 5.41)
ALAD1-2/2-2:
POR= 3.40 (0.05, 219.03)
ALAD-1: Reference group
Mexican Americans:
Q1: Reference
Q2: POR=0.74 (0.24, 2.23)
Q3: POR=1.43(0.61, 3.38)
Q4: POR=1.27 (0.59, 2.75)
ALAD1-2/2-2:
POR= 0.49 (0.08, 3.20)
ALAD-1: Reference group
ALAD2 carriers in non-Hispanic
white population in highest blood
Pb quartile with blood Pb level <10
ug/dL associated with
hypertension compared to ALAD1
homozygous individuals
(POR=1.86[1.00, 3.49]).
4-340
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Table 4-19 (Continued): Additional characteristics and quantitative data for results presented in Figure 4-18 for associations
of blood and bone Pb with hypertension measures.
Study
(Ordered as they Study Population
appear in the text) and Methodology Parameter
Park et al. Cross-sectional Hypertension
(2009b) 12,500 NHANES III
(1988-1994)
participants, mean
age 49.9 yr
Pb Data
Concurrent Blood Pb
Mean and SE (ug/dl_)
Overall:
3.52(0.10)
White men
<50 yr:
4.02(0.16)
> 50 yr:
4.92(0.18)
Black men
<50 yr:
4.55(0.15)
> 50 yr:
7.57 (0.22)
White women
<50 yr:
2.09(0.07)
> 50 yr:
3.53(0.12)
Black women
<50 yr:
2.52 (0.09)
> 50 yr:
4.49(0.16)
Statistical Analysis
Logistic regression models
adjusted for age, education,
smoking status, cigarette
smoking, BMI, hematocrit,
alcohol consumption, physical
activity, antihypertensive
medication use, and diagnosis
of type-2 diabetes
Effect Estimate (95% Cl)
OR perSD
(0.75 ug/dl_) in log blood Pb:
Overall: 1.12(1.03, 1.23).
White men:
1.06(0.92, 1.22)
Black men:
1.17(0.98, 1.38)
White women:
1.16(1.04, 1.29)
Black women:
1.19(1.04, 1.38)
Men <50 yr:
0.98(0.80, 1.22)
Men >50 yr:
1.20(1.02, 1.41)
Women <50 yr:
1.23(1.04, 1.46)
Women >50 yr:
1.09(0.94, 1.26)
aNot included in Figure 4-18 because OR data were not reported.
4-341
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In an occupational cohort in South Korea, Glenn et al. (2006) simultaneously modeled
multiple Pb dose measures of individuals collected repeatedly over four years of follow
up. Thus, through the assessment of cross-sectional and longitudinal relationships with
BP, this study provided key insight on potentially important time periods of Pb exposure
and also informed the directionality of association. The initial blood Pb level was used as
a baseline covariate and the difference in blood Pb level between visits was computed for
each subsequent visit. The bone Pb measures (tibia Pb at visits 1 and 2, patella Pb at visit
3) were used to indicate historical exposure and cumulative dose. Four models were
specified: Model 1 was conceptualized to reflect short-term changes in BP associated
with recent dose; Model 2 to reflect longer-term changes associated with cumulative dose
controlling for the association of baseline BP with recent dose; Model 3 to reflect longer-
term changes associated with cumulative dose controlling for cross-sectional influence of
cumulative dose on baseline BP; and Model 4 to reflect both short-term change with
recent dose and longer-term change with cumulative dose. Concurrent blood Pb and
increases in blood Pb between visits were associated with increases in systolic BP in
Model 1 (short-term dose) and Model 4 (short- and longer-term dose). No association
was observed between BP and tibia Pb at baseline while higher tibia Pb was associated
with a decrease in systolic BP in each of the models.
Glenn et al. (2006) was strengthened by the analysis of associations between changes in
blood Pb and changes in BP overtime within individual subjects. These results indicate
that circulating Pb (e.g., blood Pb) may act continuously on systolic BP and reductions in
blood Pb may contribute to reductions in BP, while cumulative Pb exposure (represented
by bone Pb in this study) may contribute to hypertension incidence by different
mechanisms over longer time periods and in older subjects. This analysis in subjects
(mean [SD] age at baseline 41.4 [9.5] years) with a low prevalence of hypertension
suggests at least one of the biological pathways that influences how systolic BP responds
to Pb operates over a relatively rapid timeframe. This may reflect an immediate response
to Pb at a biochemical site of action as a consequence of the biologically available Pb
circulating in blood. A persistent effect of cumulative doses over a lifetime may occur via
other mechanisms. Bone Pb level may exert influence on blood Pb levels and
consequently on BP in an aging population with prolonged Pb exposure. Thus, the
findings contribute important information regarding the various short and long-term
exposure relationships with increases in BP and hypertension. It is important to
acknowledge the uncertainty regarding the applicability of these findings regarding short-
term and long-term effects in Pb workers with relatively high current Pb exposures
contributing to blood Pb levels (mean blood Pb levels overtime: 20-37 (ig/dL) to adults
in the U.S. general population whose concurrent blood Pb levels are influenced more by
Pb mobilized from bone stores. Further, for bone Pb analysis, the potential for bone Pb,
BP, and hypertension findings to be impacted by residual confounding by age may be a
4-342
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factor to consider in older populations since exposure studies of older cohorts (NAS/
mean age >60 years; (Wilker et al.. 2011; Kim et al.. 1997)) indicate that bone Pb is
correlated with age.
In a separate cross-sectional analysis of the same occupationally-exposed group in year
three of follow-up, Weaver et al. (2008) examined associations of concurrent patella Pb
and blood Pb level with systolic BP, diastolic BP, and hypertension and effect
modification by ALAD (the genotypes have different affinities for Pb) and vitamin D
receptor (VDR) polymorphisms. None of the Pb biomarkers were associated with
diastolic BP. Patella Pb alone was not significantly associated with systolic BP. However,
blood Pb, either alone or with patella Pb, was significantly associated with higher systolic
BP. The patella Pb-age and blood Pb-age interactions were not statistically significant.
There were no significant associations of blood Pb or patella Pb with hypertension status
or effect modification by age or sex. Further, interactions between polymorphisms of the
VDR and of ALAD variants with blood Pb and patella Pb on systolic BP were not
statistically significant. Mean blood Pb level was high (30.9 (ig/dL) compared to
non-occupational groups.
Weaver et al. (2010) provided the results of further analysis of this Korean worker cohort,
with a focus on determining the functional form of the concentration-response
relationships. In a log linear model, the coefficient indicated that every doubling of blood
Pb level was associated with a systolic BP increase of 1.76 mmHg. The J test, a statistical
test for determining which, if either, of two functional forms of the same variable
provides a superior fit to data in non-nested models (Davidson and MacKinnon. 1981).
returned ap-value of 0.013 in favor of the natural log blood Pb level over the linear blood
Pb level specification. This analysis indicates that the systolic BP increase in this cohort
is better described as a logarithmic function of blood Pb level within the range of the
study than by a linear function.
Several analyses in the NAS cohort of predominantly white older men in the greater
Boston area found associations of blood and bone Pb level with BP and hypertension, and
they indicated effect modification by calcium intake, perceived stress, and HFE gene
variants. In a cross-sectional analysis, Perlstein et al. (2007) found a statistically
significant association between blood Pb and diastolic BP in adjusted models. The
subjects in this study had at least one bone Pb measurement during the years 1991-1997
and were not on antihypertensive medication at the time of the measurement. While tibia
Pb was not significantly associated with BP, it was associated with pulse pressure (PP).
Men with tibia Pb above the median (19 (ig/g) had a higher mean PP (4.2 mmHg [95%
CI: 1.9, 6.5]) compared to men with tibia Pb below the median. The trend toward
increasing PP with increasing quintile of tibia Pb was statistically significant although
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none of the confidence intervals for PP referenced to the lowest quintile of tibia Pb
(<7.4 (ig/g) excluded the null value.
Peters et al. (2007) examined cross-sectionally the modification of the associations of
tibia and patella Pb with BP and hypertension by self-reported stress (assessed by
questionnaire) in NAS men. High stress also has been linked with higher BP, potentially
via mechanisms also affected by Pb exposure including activation of sympathetic
pathways, ROS, and the HPA axis. Among all subjects, higher bone Pb level was
associated (statistically nonsignificant) with greater odds of hypertension status and
higher systolic BP. As indicated in Figure 4-19. the association between systolic BP and
tibia Pb differed between those with high and low self-reported stress ((3 for tibia Pb x
stress interaction = 3.77 mmHg [95% CI: 0.46, 7.09]) per SD increase in tibia Pb. Stress
also was found to modify the patella Pb-BP association ((3 for patella Pb x stress
interaction = 2.60 mmHg [95% CI: -0.95, 6.15] per SD increase in patella Pb). Neither
bone, self-reported stress, nor their interaction was associated significantly with diastolic
BP. Peters et al. (2007) also used Cox proportional hazards models to assess the
interaction of stress and bone Pb level in the development of hypertension among those
free of hypertension at baseline. The results of this analysis showed that increasing tibia
and patella Pb were associated with greater risk of developing hypertension among those
with high stress compared to those with lower perceived stress (RR of developing
hypertension among those with high stress: 2.66 [95% CI: 1.43, 4.95] per SD increase in
tibia Pb and 2.64 [95% CI: 1.42, 4.92] per SD increase in patella Pb). These results
provide evidence supporting adults with higher stress as a population potentially at
increased risk of Pb-associated cardiovascular effects. Earlier, Cheng et al. (2001)
examined the NAS cohort in 474 subjects without hypertension (mean [SD] blood Pb
level: 5.87 [4.01]) at baseline measurement and analyzed linear models with patella Pb
and reported that only patella Pb level was associated with a significant increase in the
rate ratio for hypertension using a Cox's proportional hazards model.
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140
135-
Ji 130-
1
125~
g 120-
115-
110
O High perceived stress
• Low perceived stress
Trend (high stress)
Trend (low stress)
-10 0 10
30
50
70
90
Tibia lead (ug/g)
Source: Peters et al. (2007)
Figure 4-19 The relationship between tibia Pb and estimated systolic BP
(SBP) for those with high self-reported stress versus those with
low self-reported stress.
Elmarsafawy et al. (2006) examined the modification of the relationship between Pb and
hypertension by dietary calcium, with 467 subjects from the NAS. Responses on a semi-
quantitative dietary frequency questionnaire with one-year recall were used to estimate
calcium intake. Effect modification by calcium intake (dichotomized at 800 mg/day) was
examined using interaction terms in logistic regression models and by conducting
analyses stratified on the calcium variable. Increasing bone and blood Pb increased the
odds of hypertension, particularly among subjects with low dietary calcium (i.e., <800
mg/day).
Zhang et al. (2010a) examined the effect of polymorphisms of the hemochromatosis gene
(HFE) on the relationship of bone Pb with PP in NAS men. HFE polymorphisms promote
Fe absorption and have been shown to modify the impact of adult cardiac function. The
H63D variant is associated with hemochromatosis, a disease characterized by higher
Fe body burden. Subjects had up to three PP measurements during the 10 year study
period. The overall results demonstrated a strong relationship between bone Pb and PP in
this study, similar to an earlier cross-sectional PP study of many of the same subjects
(Perlstein et al.. 2007). Zhang et al. (2010a) extended these findings by demonstrating
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larger increases in PP per unit increase in tibia and patella Pb level among those with the
H63D variant compared to those with the wild-type or the C282Y variant.
A small number of cross-sectional studies examined and found that blood Pb level was
associated with hypertension in pregnancy. Yazbeck et al. (2009) examined a
community-based group of pregnant women in France and unlike most other studies,
adjusted for potential confounding by blood concentrations of Cd, Mn, and Se. Pregnancy
induced hypertension (PIH) was defined as systolic BP >140 mmHg and/or diastolic BP
>90 mmHg during at least two clinic visits after week 22 of gestation. Patients with pre-
existing chronic hypertension were excluded. The mean (SD) blood Pb levels measured
during pregnancy were 2.2 (1.4) (ig/dL in PIH cases and 1.9 (1.2) (ig/dL in normotensive
women. An association between blood Pb and PIH was observed (OR 3.29 [95% CI:
1.11, 9.74] per unit increase in log-transformed blood Pb level). Cd and Se concentrations
were comparable between PIH and no PIH groups. Adjustment for the metals slightly
attenuated but did not eliminate the association between blood Pb levels and the risk of
PIH. Investigators observed no significant interactions among blood Pb level, any of the
other elements, and maternal characteristics in predicting the risk of PIH. Interaction
between blood Se and Pb concentrations was not significant, and the putative protection
effects of Se through antioxidative properties were not found in this study.
Wells et al. (20 lib) measured the relationship of cord blood Pb with BP in 285 women at
admission to the Johns Hopkins Hospital in Baltimore, MD, during labor and delivery.
Women with cord blood Pb levels in the highest quartile for the study group
(>0.96 (ig/dL) had significantly higher systolic and diastolic BP (upon admission and for
maximum BP) compared to women in the first quartile (<0.46 (ig/dL). The level of
uncertainty at these levels of exposure is difficult to estimate. The authors used
Benchmark Dose Software V2.1, developed by the EPA, to estimate the blood Pb level
(benchmark dose or BMD) and the associated lower confidence limit (BMDL) that was
associated with one SD increase in BP. In this study group, one SD is approximately
equivalent to a 10% increase above the mean for the first quartile blood Pb reference
group. The BMD approach was used only as a means of quantifying the relationship of
blood Pb with BP in this population. This analysis indicated that the BMDL for the
maternal blood Pb level (estimated from cord blood Pb levels) that was associated with a
1 SD increase in all BP outcomes was about 1.4 ug/dL. These reported results are similar
to those reported in the 2006 Pb AQCD as well as those found 25 years ago but with
blood Pb levels an order of magnitude lower in the more recent study. However,
uncertainty exists as to the specific Pb exposure level, timing, frequency, and duration
that contributed to the observed associations.
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Recent analyses using NHANES data continued to indicate associations of Pb biomarkers
with BP and hypertension. Muntner et al. (2005) previously used the NHANES
1999-2002 data to indicate that concurrent blood Pb levels were associated with
hypertension, peripheral artery disease (PAD), and chronic kidney disease. The PAD
results are discussed later in Section 4.4.3.5. and chronic kidney disease results are
discussed in Section 4.5.2.1. Blood Pb increased regularly with age (geometric means
[95% CIs]: 1.28 (ig/dL [1.23, 1.33] in the 18-39 year age group to 2.32 (ig/dL [2.20, 2.44]
in the 75 year and older age group). Associations were observed between concurrent
blood Pb level and hypertension across race/ethnicity groups with significant trends
observed for non-Hispanic blacks and Mexican Americans.
In the NHANES III 1988-1994 population, Scinicariello et al. (2010) found a gene-
environment interaction between blood Pb level and ALAD genotype (the genotypes
have different affinities for Pb) in relation to systolic BP and diastolic BP in a cross-
sectional analysis. These interactions varied across race/ethnicity strata. The strongest
associations were observed among non-Hispanic blacks (Figure 4-17. Table 4-18). A
statistically significant interaction was observed between concurrent blood Pb level and
ALAD 1-2/2-2b (i.e., ALAD2 carriers) among non-Hispanic whites and non-Hispanic
blacks. Scinicariello et al. (2010) also found an interaction between ALAD genotype and
blood Pb level in the association with hypertension. Statistically significant associations
between concurrent blood Pb level and hypertension were observed among non-Hispanic
blacks and nonsignificant increases were observed among non-Hispanic whites and
Mexican Americans (with the exception of Mexican Americans in the second quartile of
blood Pb level) (Figure 4-18. Table 4-19). In addition, non-Hispanic white ALAD2
carriers in the highest blood Pb level quartile 3.8-52.9 (ig/dL) had a significantly higher
association with hypertension compared with ALAD1 homozygous individuals in the
highest quartile of blood Pb. In the same NHANES population, Park et al. (2009b)
predicted bone Pb levels using a model developed with NAS data. Concurrent blood Pb
was associated with hypertension overall in the NHANES population, with larger
associations observed among black men and women as well as older adults (Figure 4-18.
Table 4-19). Associations also were observed with estimated bone Pb.
4.4.2.2 Toxicology
Studies on the effect of Pb (as blood Pb level) on systolic BP in unanesthetized adult rats
consistently reported an increase in BP with increasing blood Pb level as shown in Figure
4-20 (results summarized in Table 4-20). An array of studies has provided evidence that
long-term Pb exposure (>4 weeks), resulting in blood Pb levels less than 10 (ig/dL, which
is relevant to humans in the U.S., can result in the onset of hypertension (after a latency
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period) in experimental animals that persists long after the cessation of Pb exposure (U.S.
EPA. 2006b). Tsao et al. (2000) presented evidence for increased systolic and diastolic
BP in rats with blood Pb levels somewhat similar to the current U.S. adult population
(mean [SD]: 2.15 [0.92] Lig/dL blood Pb; 140 [7] mmHg systolic BP, 98 [7] mmHg
diastolic BP) compared to untreated controls (mean [SD]: 0.05 [0.05] Lig/dL blood Pb;
127 [7] mmHg systolic BP, 88 [7] mmHg diastolic BP). As this was the lowest Pb level
tested, no evidence of a threshold was evident. Further, a test for linear trend revealed a
statistically significant, positive trend for increasing BP with increasing blood Pb levels
up to 56 Lig/dL (e.g., mean [SD]: 5.47 [2.1] Lig/dL blood Pb; 143 [6] mmHg systolic BP,
97 [8] mmHg diastolic BP), with the effect leveling off at higher blood Pb levels.
200
180
160
8
a
120
100
10
15 20 25
Blood Pb Level (|ig/dL)
JO
40
-Bravo etal. 2007
-Rizzietal. 2009
-Chang etal. 1997 Chang etal. 2005 Heydari etal. 2006 Nakhoul etal. 1992
Rizzietal. 2009 Tsao etal. 2000 Zhangetal. 2009 Fiorimet al. 2011
Note: Crosses represent standard error for blood Pb and BP measurements. If no crossbar is present, error results were not
reported. Arrows represent higher doses tested.
Figure 4-20 Changes in BP after Pb exposure (represented as blood Pb level)
in unanesthetized adult rats across studies.
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Table 4-20 Characteristics of studies of blood Pb with BP measures in animals
presented in Figure 4-20.
Reference3
Fiorim et al.
(2011)
Nakhoul et
al. (1992)
Chang et al.
(2005)
Tsao et al.
(2000)
Rizzi et al.
(2009)
Chang et al.
(1997)
Heydari et al.
(2006)
Bravo et al.
(2QQ7)
Zhang et al.
(2009a)
Lifestage;
Sex;
Strain
Adult; M,
Wistar Rat
Adult; M
SH Rat
Adult; M
Wistar Rat
Adult, M
Wistar Rat
Adult; M
Wistar Rat
Adult; M
Wistar Rat
Adult; M
Sprague-
Dawley
Rat
Adult; M
Sprague-
Dawley
Rat
Adult; M
Wistar Rat
Exposure
Duration
7 days
8 weeks
8 weeks
8 weeks
8 weeks
8 weeks
12 weeks
14 weeks
40 weeks
Exposure Level;
Route
4ug/100g
followed by
0.05 ug/100 g
daily;
intramuscular
100 ppm;
drinking water
20,000 ppm then
removal and
measurements
1-7 mo after;
drinking water
100 -20,000 ppm;
drinking water
30, 90 ppm;
drinking water
500 ppm; drinking
water
100 ppm; drinking
water
100 ppm; drinking
water
100 ppm; drinking
water
Mean
[SEM]b
Blood Pb
Level
(ug/dL) n
9.98 [1.7] 12
5.3 [3] 7
Range: 5
4.5 to 83
Range of 10
means:
2.15 [0.29]
to
85.76 [1.29]b
7.6 [1.3], 11
19.3 [3.4]
29.1 [0.6]b 10
26.8 [2.2] 6
23.7[1.9]b 12
28.4[1.1]b 8-10
ASBP
(mmHg;
lowest
blood Pb
level
compared
with
control)0
16
28
13.8
13
13.3
58
25.8
30
15.3
Comments
Spontaneously
hypertensive
rat model
Studies are presented in order of increasing duration of exposure.
bStandard deviation converted to SEM.
""Difference in systolic BP (SBP) between group means not within one exposure group.
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Experimental animal studies continued to provide evidence that long-term Pb exposure
results in sustained arterial hypertension after a latency period. Systolic BP increased in
rats after exposure to 90-10,000 ppm Pb (as Pb acetate in drinking water) for various time
periods that resulted in blood Pb levels between 19.3-240 (ig/dL (Mohammad et al.
2010; Zhang et al.. 2009a: Badavi et al.. 2008; Grizzo and Cordellini. 2008; Reza et al..
2008: Bravo et al.. 2007: Vargas-Robles et al.. 2007: Hevdari et al.. 2006: Bagchi and
Preuss. 2005). Past studies have shown statistically significant elevations in BP in rats
with lower blood Pb levels. For example, long-term Pb exposure to spontaneously
hypertensive rats (resulting in mean [SEM] blood Pb level: 5.3 [3] (ig/dL) led to
increased BP (Nakhoul et al., 1992). Consistent with measurements of systolic BP by tail-
cuff plethysmography, Pb exposure (100 ppm for 14 weeks; mean blood Pb level:
24 (ig/dL) also caused an increase in intra-aortic mean arterial pressure (Bravo et al..
2007). In a study that tested low levels of Pb exposure (30 ppm; mean blood Pb level:
7.6 (ig/dL), a statistically significant increase in systolic BP was not observed despite
elevated blood Pb level after 8 weeks of treatment. Nonetheless, there was a trend of
higher BP with increased blood Pb levels (Rizzi et al.. 2009).
Studies found that Pb-induced increases in BP persisted long after cessation of Pb
exposure. Bagchi and Preuss (2005) found that elevated systolic BP was maintained for
210 days after cessation of Pb exposure (10,000 ppm Pb acetate in water, 40 days,
monitored for one year). However, chelation therapy using Na2CaEDTA returned
systolic BP to levels comparable to those in rats not treated with Pb (Bagchi and Preuss.
2005). Chang et al. (2005) reported a partial reversibility of effect after cessation of Pb
exposure, where Pb-induced elevated BP decreased but did not return to control levels
7 months post Pb exposure. After Pb exposure was removed, blood, heart, aorta, and
kidney Pb levels decreased quickly within the first three months (Chang et al.. 2005).
Pb-induced elevated systolic BP persisted for one month following Pb exposure
cessation, followed by obvious decreases in BP until 4 months after Pb exposure
cessation. Between 4 and 7 months after Pb exposure cessation, the still-elevated BP did
not decrease further, thus never returning to control BP levels. Decreases in BP were
strongly correlated with decreases in blood Pb level after exposure cessation.
The aforementioned studies all assessed the relationship between long-term exposure
(>4 weeks) of rats to Pb and measures of BP. However, recent research also investigated
BP elevation occurring after short-term treatment with Pb (<4 weeks). Studies found
increased systolic BP after 7 days of Pb treatment (daily injections resulting in mean
[SEM] blood Pb levels of 9.98 [1.7] (ig/dL) (Fiorim et al.. 2011) and after 2 weeks of Pb
exposure (100 ppm via drinking water) (Sharifi et al.. 2004). A study utilizing intra-
arterial pressure measurements found that a single high-dose Pb injection in rats
(resulting in mean [SEM] blood Pb levels of 37 [1.7] (ig/dL) increased systolic arterial
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pressure after only 60 minutes (Simdes et al.. 2011). The injection of Pb into the rat may
not allow for extrapolation of these results to humans since this is not a comparable Pb
exposure route. These studies suggest that there is the potential for an increase in BP
following short-term Pb treatment. It is possible that the increases in BP following short-
and long-term Pb exposures are occurring through separate mechanisms; however,
studies using both short- and longer-term Pb exposure have correlated increased BP with
an activation of the renin-angiotensin system (i.e., increase in angiotensin converting
enzyme (ACE) activity) (Section 4.4.2.3). Several of these aforementioned studies used
the injection route of Pb administration, and the relevance of these bolus doses over short
periods of time to human routes of short-term exposure is uncertain. However, it is
important to acknowledge that the results were similar to those from the study that
examined short-term exposure to Pb via drinking water.
4.4.2.3 Hypertension Modes of Action
The 2006 Pb AQCD (U.S. EPA. 2006b) examined a number of mechanisms leading to
Pb-induced hypertension, including oxidative stress, hormonal and blood pressure
regulatory system dysfunction, vasomodulation, and cellular alterations. As described
below, recent studies in experimental animals and cells further supported roles for these
potential mechanisms in mediating hypertension from Pb exposure.
Oxidative Stress Response - Reactive Oxygen Species and Nitric Oxide
Several studies discussed in the 2006 Pb AQCD demonstrated a role for oxidative stress
in the pathogenesis of Pb-induced hypertension, mediated by the inactivation of nitric
oxide (NO) and downregulation of soluble guanylate cyclase (sGC) (Dursun et al. 2005;
Attrietal.. 2003: Gonick et al.. 1997: Vazirietal.. 1997: Khalil-Manesh et al.. 1994:
Khalil-Manesh et al.. 1993b). Pb-induced reduction of biologically active NO was found
not to be due to a reduction in NO-production capacity (Vaziri and Ding. 2001: Vaziri et
al.. 1999a): instead, it was found to result from inactivation and sequestration of NO by
ROS (Malvezzi et al., 2001: Vaziri etal.. 1999b). Oxidative stress from Pb exposure in
animals may be due to upregulation of NAD(P)H oxidase (Ni et al., 2004: Vaziri et al..
2003). induction of Fenton and Haber-Weiss reactions (Ding etal. 2001: Ding et al..
2000). and failure of the antioxidant enzymes, CAT and GPx, to compensate for the
increased ROS (Farmand et al.. 2005: Vaziri et al.. 2003). Many biological actions of
NO, such as vasorelaxation, are mediated by cGMP, which is produced by sGC from the
substrate GTP. Oxidative stress also has been found to play a role in Pb-induced
downregulation of sGC (Farmand et al.. 2005: Courtois et al.. 2003: Marques et al..
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2001). Thus, the reduction of the vasodilator NO from inactivation and sequestration by
Pb-induced ROS leads to increased vasoconstriction and BP.
Pb-induced oxidative stress also has been found to induce renal tubulointerstitial
inflammation which plays a crucial role in models of hypertension (Rodriguez-Iturbe et
al.. 2005; Rodriguez-Iturbe et al.. 2004). Tubulointerstitial inflammation from treatment
with Pb has been coupled with activation of the redox sensitive NF-KB (Ramesh et al..
2001). Pb-induced hypertension, inflammation, and NF-KB activation can be ameliorated
by antioxidant therapy (Rodriguez-Iturbe et al., 2004). There is mixed evidence to
suggest that Pb-induced hypertension may also be promoted by activation of PKC leading
to enhanced vascular contractility (Valencia et al.. 2001; Watts et al.. 1995).
Recent studies continued to provide evidence for the role of ROS and NO metabolism in
Pb-induced hypertension and vascular disease. Increased systolic BP after Pb exposure
was accompanied by increased superoxide (O2~) and O2"-positive cells (Bravo et al..
2007; Vargas-Robles et al.. 2007). elevated urinary malondialdehyde (MDA, a measure
of lipid peroxidation) (Bravo et al.. 2007). and increased 3-nitrotyrosine (Vargas-Robles
et al.. 2007). Inhibition of NAD(P)H oxidase, an enzyme that generates O2~ and hydrogen
peroxide, was able to block Pb-induced (1 ppm) aortic contraction
to 5-hydroxytryptamine (5-HT) (Zhang et al.. 2005). Increases in systolic BP, intra-aortic
mean arterial pressure, and MDA after Pb exposure (100 ppm; mean blood Pb level:
23.7 (ig/dL) were also prevented by treatment with the immunosuppressant,
mycophenolate mofetil (MMF) (mean blood Pb level in MMF-treated animals: 27 (ig/dL)
(Bravo et al.. 2007). MMF has been shown to inhibit endothelial NAD(P)H oxidase,
which could explain how it decreases Pb-induced increases in oxidative stress and BP.
MMF was not found to alter blood Pb levels of animals. Red grape seed extract and
ascorbic acid supplementation were also able to protect rats from Pb-induced (100 ppm)
increased BP and heart rate, perhaps through the antioxidant properties of both the extract
(Badavi et al.. 2008) and vitamin C (Mohammad et al.. 2010). Red grape seed extract did
not alter the accumulation of Pb in blood, indicating that its protective effect was not
mediated through altered Pb toxicokinetics; however, internal doses of Pb were not
measured in the vitamin C study to clarify the mechanism of action of vitamin C. Another
study found that the antioxidant, anti-inflammatory chemical, curcumin, as well as
physical exercise training reversed Pb-induced increases in serum creatinine kinase-MB
(CK-MB), low density lipoprotein (LDL), heart high-sensitivity C-reactive protein
(hs-CRP), and MDA. Pb-induced decreases in serum total antioxidant capacity, high
density lipoprotein (HDL), and heart glutathione peroxidase (GPx) were also reversed by
curcumin and exercise. However, internal doses of Pb were not measured to clarify the
mechanism of action in this study (Roshan et al.. 2011).
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Exposure to Pb can also affect the activity and levels of antioxidant enzymes. Male ($)
and female (9) rats exposed to Pb for 18 weeks (100-1,000 ppm) had altered responses in
antioxidant enzymes in heart tissue (Sobekova et al.. 2009; Alghazal et al.. 2008a). Pb
exposure in female rats increased the activity of cardiac SOD, GST, GR, and GPx
(MOO ppm) and increased cardiac thiobarbituric acid reactive substances (TEARS, a
measure of lipid peroxidation) (1,000 ppm). Pb exposure in male rats did not affect the
activity of SOD or production of TEARS, however decreased the activity of GST and GR
(>100 ppm). Male and female rats also accumulated different amounts of Pb in the
cardiac tissue after similar Pb exposure (<$ 100 ppm: 205% of control, 1,000 ppm: 379%;
9 100 ppm: 246%, 1,000 ppm: 775%), which could explain the sex differences observed
in antioxidant enzyme responses.
Oxidative stress can trigger a cascade of events that promote cellular stress, renal
inflammation, and hypertension. As was shown previously (Rodriguez-Iturbe et al..
2005). Pb exposure can increase renal NF-KB, which was associated with
tubulointerstitial damage and infiltration of lymphocytes and macrophages (Bravo et al..
2007). These events could also be ablated by MMF treatment, likely due to its anti-
inflammatory and antioxidant properties. Pb also was found to induce inflammation in
human endothelial cells as a model for vessel intima hyperplasia (Zeller et al.. 2010). The
pro-inflammatory cytokine, interleukin (IL)-8 protein and mRNA were increased,
concentration- and time-dependently, after in vitro Pb exposure (5-50 (iM). Enhanced
IL-8 production was mediated through activation of the transcription factor Nrf2 (but not
NF-KB, hypoxia inducible factor-1, or aryl hydrocarbon receptor), as shown through
increased nuclear translocation and Nrf2 cellular knockdown experiments. Additionally,
measures of endothelial stress, NQO1 and HO-1 protein, were induced by Pb exposure
(Zeller et al.. 2010). Pb treatment (20 ppm, i.p., 3 days/week, 8 weeks) increased the
inflammatory markers hs-CRP and CK-MB in rat hearts (Roshan etal.. 2011).
Oxidative stress affects vascular reactivity and tone through inactivation and
sequestration of NO, causing a reduction in biologically active NO. Recent studies
affirmed past conclusions on the interplay of ROS and NO metabolism in the
cardiovascular effects of Pb. Elevated systolic BP and altered vasorelaxation after Pb
exposure was accompanied by a decrease in total nitrates and nitrites (NOX) (Mohammad
etal.. 2010; Zhang et al.. 2007a; Hevdari et al.. 2006). Serum NOX levels in Pb-treated
rats remained depressed for 8 weeks and then reversed after 12 weeks, despite continued
elevation in systolic BP (Hevdari et al.. 2006). This return of serum NOX levels to levels
similar in controls could be a result of compensatory increases in endothelial NOS
(eNOS) attempting to replenish an over-sequestered NO supply. With this in mind,
studies showed increased eNOS protein expression after long-term Pb exposure in kidney
(Zhang et al.. 2007a) and isolated cultured aorta (Vargas-Robles et al.. 2007). No change
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in inducible NOS was observed in isolated cultured aorta after 1 ppm Pb exposure (Zhang
et al.. 2007a). In contrast to long-term exposure, Pb treatment over a short time period
(daily injections resulting in mean [SEM] blood Pb levels of 9.98 [1.7] (ig/dL) was found
to increase iNOS and phosphorylated eNOS protein (Fiorim et al.. 2011) which may
cause an increase in NO production and a short-term increase in NO bioavailability. This
increase in NO bioavailability early after Pb exposure could be the immediate
compensatory mechanism against the elevation in BP.
NO, also known as endothelium-derived relaxing factor, is a potent endogenous
vasodilator. Toxicological studies continued to investigate the effects of Pb on
NO-dependent vascular reactivity by using NO stimulating vasodilators, such as
acetylcholine (ACh) and sodium nitroprusside (SNP), and NO inhibiting
vasoconstrictors, such as L-NAME. Studies provided mixed evidence; however, results
suggested that Pb disrupts the vasorelaxant response to NO in the aorta due to damage to
the endothelium. Pb exposure (1 ppm and 100 (JVI, 1 hour) decreased ACh-induced
vasorelaxation, which triggers the release of NO from the endothelial cell, in isolated rat
' OO ?
tail artery, suggesting damage to the endothelium (Silveira et al.. 2010; Zhang et al..
2007a). In aortic rings of perinatally exposed rats (1,000 ppm through pregnancy and
lactation, mean blood Pb level: 58.7 (ig/dL), blocking NOS with L-NAME abolished the
relaxant response evoked by ACh (Grizzo and Cordellini. 2008). However, there was no
change observed in the relaxation response to ACh by Pb alone (Fiorim et al.. 2011; Rizzi
et al.. 2009; Grizzo and Cordellini. 2008). Conversely, Skoczynska and Stojek (2005)
found that Pb exposure (50 ppm; blood Pb level 11.2 (ig/dL) enhanced NO-mediated
vasodilation by ACh in rat mesenteric arteries, and NOS inhibition enhanced the ACh
relaxant response. A number of studies found that Pb exposure did not affect smooth
muscle integrity since SNP-induced vasorelaxation, which is endothelium independent,
was unchanged (Fiorim et al.. 2011; Silveira et al.. 2010; Rizzi et al.. 2009; Grizzo and
Cordellini. 2008).
NO also was found to play a role in the interaction between Pb and the vasoconstrictor
response. Blocking NOS with L-NAME or inhibiting iNOS specifically, which decreases
NO production, increased the contraction of aortic rings in response to the
vasoconstrictor phenylephrine (PHE), and Pb exposure potentiated this response (Fiorim
et al., 2011). Also, L-NAME increased the Pb pressor response to PHE after perinatal Pb
exposure (1,000 ppm through pregnancy and lactation, blood Pb level 58.7 (ig/dL)
(Grizzo and Cordellini. 2008). Conversely, in rat renal interlobar arteries, Pb exposure
blunted the increase in renal angiotensin II (Angll)-mediated contraction from NOS
inhibition by L-NAME (Vargas-Robles et al., 2007). Treatment with the SOD mimetic
tempol, which would increase NO bioavailability, decreased, but did not eliminate, the Pb
pressor response (Silveira et al.. 2010).
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In summary, recent studies continued to provide evidence for the role of ROS in
Pb-induced hypertension and vascular disease by indicating Pb-induced increases in ROS
and modulation of cardiovascular responses by antioxidant substances. Additionally,
recent studies continued to show that Pb-induced hypertension and vascular responses are
mediated primarily via inactivation of NO not via inhibition of NO production.
Vascular Reactivity
Pb exposure-induced alterations of the adrenergic system, which can increase peripheral
vascular resistance, and thereby arterial pressure, may be one mediator of Pb-induced
hypertension. Pb exposure in animals can increase stimulation of the sympathetic nervous
system (SNS), as shown by increased plasma levels of norepinephrine (NE) and other
catecholamines (Carmignani et al., 2000; Chang et al., 1997) and decreased (3 adrenergic
receptor density and (3 agonist-stimulated cAMP production in the aorta and heart (Tsao
et al., 2000; Chang etal., 1997). These stimulatory effects on the SNS paralleled the
effects of Pb on BP, cardiac contractility, and carotid blood flow. Pb-induced elevations
in arterial pressure and heart rate were abrogated by ganglionic blockade (Simdes et al.,
2011; Lai et al.. 2002). Arterial pressure and heart rate gradually decreased 7 months
after Pb exposure cessation as did the Pb-induced SNS alterations (Chang et al.. 2005).
Increases in BP can be caused by activation of the SNS, which can lead to vascular
narrowing, in turn, resulting in increased total peripheral resistance. In this neural
mechanism, activation of the SNS leads to vasoconstriction, whereas inhibition leads to
vasodilation. It has been suggested that Pb leads to increased vascular reactivity to
catecholamines (i.e., epinephrine, NE, and dopamine), hormones of the SNS. Indeed, the
isolated mesenteric vessel bed from Pb-treated rats (50 ppm with blood Pb level:
11.2 (ig/dL, but not 100 ppm with blood Pb level: 17.3 (ig/dL) exhibited increased
reactivity to NE (Skoczynska and Stojek. 2005). However, in another study, 100 ppm Pb
did not affect the NE-induced contractile response after 10 months of exposure (Zhang et
al.. 2009a). suggesting a small range of Pb doses affects pressor response to NE.
Catecholamines act primarily through the adrenergic and dopaminergic receptors.
Antagonists of a 1-adrenergic, a2-adrenergic, (3-adrenergic, and dopamine Dl receptors
were found to abolish Pb-induced aortic contraction (Fazli-Tabaei et al.. 2006; Heydari et
al.. 2006). However, the a 1-adrenergic receptor agonist, PHE, induced aortic contractions
and these were enhanced by treatment with Pb (100 ppm; blood Pb level: 26.8 (ig/dL),
indicating a specific role for the a 1-adrenergic receptor (Silveira et al.. 2010; Grizzo and
Cordellini. 2008; Heydari et al.. 2006). Removal of the endothelium blunted the PHE-
induced contraction. Conversely, short-term Pb treatment (7 days, i.p.) decreased the
contractile response induced by PHE in rat aortas resulting in a decreased vascular
reactivity (Fiorim et al.. 2011). This decrease may represent a compensatory response
4-355
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attempting to correct the Pb-induced BP elevation. Additionally, Pb blunted the
isoproterenol-induced relaxation, supporting a role for the (3-adrenoceptors in mediating
the effects of Pb on vascular reactivity (Vassallo et al., 2008; Hevdari et al., 2006).
Recently, there was mixed evidence for Pb disrupting vascular reactivity to other pressor
agents. Pb (1 ppm) treatment of isolated rat thoracic aorta increased 5-HT induced
contraction, which was endothelium dependent, but not due to 5-HT2B receptor
expression (Zhang et al.. 2005). Follow-up of this study in whole animals found, on the
contrary, that Pb (100 ppm; blood Pb level: 28.4 (ig/dL) decreased the maximum
contractile response to 5-HT, but did not affect 5-HT plasma levels or 5-HT2B receptor
expression (Zhang et al.. 2009a). In addition, Pb exposure (100 ppm, 12 weeks) increased
the renal vascular response to Angll in isolated perfused kidneys from Pb-exposed rats
(Vargas-Robles et al.. 2007).
Studies continued to investigate the effects of Pb on NO-dependent vascular reactivity by
using NO stimulating vasodilators, such as ACh and SNP, and NO inhibiting
vasoconstrictors, such as L-NAME. These studies were discussed in a preceding
subsection (Oxidative Stress Response).
Renin-Angiotensin-Aldosterone and Kininergic Systems
The adrenergic system also affects the renin-angiotensin-aldosterone system (RAAS),
which is responsible for fluid homeostasis and BP regulation, and has been shown to be
affected by Pb exposure. A meta-analysis found that Pb exposure (resulting in blood Pb
levels: 30-40 (ig/dL) increased plasma renin activity and renal tissue renin in young but
not old rats (Vander. 1988). Exposure of experimental animals to Pb also induced
increases in plasma, aorta, heart, and kidney angiotensin converting enzyme (ACE)
activity; plasma kininase II, kininase I, and kallikrein activities; and renal Angll positive
cells (Rodriguez-Iturbe et al.. 2005; Sharifi et al.. 2004; Carmignani et al.. 1999). ACE
activity declined over time while arterial pressure stayed elevated, suggesting that the
RAAS may be involved in the induction, but not the maintenance of Pb-induced
hypertension in rats.
Recent studies continued to implicate the RAAS in the development of Pb-induced
hypertension, especially during early exposure in young animals. Angll, a main player in
the RAAS, induces arteriolar vasoconstriction leading to increased BP. Pb exposure
increased the vascular reactivity to Angll (Vargas-Robles et al.. 2007). Acute
(60 minutes) or short-term (7 days) treatment of rats to Pb increased the plasma ACE
activity (Fiorim et al.. 2011; Simdes etal.. 2011). and Fiorim et al. (2011) additionally
found this increase to be correlated with the Pb-induced increase in systolic BP.
4-356
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However, after these short time points there were no changes in the Angll receptors 1 or
2 protein levels or expression. Treatment with the Angll receptor (ATiR) blocker,
Losartan, or the ACE inhibitor, Enalapril, blocked the Pb-induced systolic BP increase
(Simdes etal.. 2011) and decreased the PHE-induced vasoconstrictor response in
Pb-treated aortas (Fiorim et al., 2011). Similarly, treatment with Losartan resulted in a
greater decrease in systolic BP in highly Pb-exposed rats (10,000 ppm Pb, 40 days; blood
Pb level >240 (ig/dL after exposure, 12-13 (ig/dL after chelation after 1 year) compared
to control rats, and this decrease continued into later periods of follow-up (day 283)
(Bagchi and Preuss. 2005). Increased systolic BP after early exposure to Pb corresponded
with increased water intake, urine output, potassium excretion, and decreased urinary
sodium and urine osmolality. These functional changes in renal behavior are consistent
with the actions of a stimulated RAAS. Lower level Pb (100 ppm, 14 weeks; range of
blood Pb levels: 23.7-27 (ig/dL) exposure increased renal cortical Angll content and the
number of tubulointerstitial Angll-positive cells (Bravo et al., 2007). This heightened
intrarenal angiotensin corresponded with sodium retention and increased systolic BP and
was ablated by the anti-inflammatory antioxidant, MMF. Sodium reabsorption is
important for the maintenance of BP, and Na+ transporters play a key role in this process.
In other studies, Pb exposure increased activity and levels of the a-1 subunit protein of
Na+/K+ATPase, which plays a major role in Na+ reabsorption and is regulated by the
RAAS (Fiorim et al., 2011; Simdes et al., 2011). These studies point to the activation of
the RAAS in the course of Pb-induced hypertension, particularly in the early stages of
elevated BP.
Vasomodulators
The balance between production of vasodilators and vasoconstrictors is important in the
regulation of BP and cardiovascular function. The 2006 Pb AQCD reported that Pb did
not affect all vasomodulators in the same way. Urinary excretion of the vasoconstrictor,
thromboxane (TXB2), and the vasodilatory prostaglandin, 6-keto-PGFla, was unchanged
in rats with Pb-induced hypertension (Gonick et al., 1998). However, in vitro Pb
exposure promoted the release of the prostaglandin precursor, arachidonic acid, in
vascular smooth muscle cells (VSMCs) via activation of phospholipase A2 (Dorman and
Freeman. 2002). Plasma concentration and urinary excretion of the vasoconstrictive
peptide, endothelin (ET)-3 was increased after low- (100 ppm), but not high-level
(5,000 ppm) Pb exposure in rats (Gonick et al.. 1997: Khalil-Manesh et al.. 1994: Khalil-
Manesh et al., 1993b). Antagonism of the ET receptor A blunted the downregulation of
sGC and cGMP production by Pb in isolated rat artery segments, suggesting that some of
the hypertensive effects of Pb exposure may be mediated through ET (Courtois et al.,
2003). Additionally, Pb-exposed animals exhibited fluid retention and a
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concentration-dependent decline in the vasodilator, atrial natriuretic factor (ANF)
(Giridhar and Isom. 1990). Results from these studies suggest that Pb may interfere with
the balance between vasodilators and vasoconstrictors that contribute to the complex
hormonal regulation of vascular contraction and BP.
The imbalance in vasomodulators is one explanation for the concentration-dependent
vasoconstriction observed in some animals after Pb exposure (Valencia et al.. 2001;
Watts et al.. 1995; Piccinini et al.. 1977). However, vasoconstriction after Pb exposure
was not reported in all studies (Shelkovnikov and Gonick, 2001) and is likely varied
depending on the type of vessel used, the Pb concentration employed, and the animal
species being studied. Studies have reported Pb-induced attenuation of ACh-mediated
and NO-mediated vasodilation (Marques et al., 2001; Oishi et al.. 1996) in some, but not
all vascular tissues and in some, but not all studies (Purdy etal. 1997). These effects
have been variably attributed to Pb-mediated activation of PKC and direct action on the
VSMCs through the Ca2+ mimetic properties of Pb among other possibilities (Valencia et
al..2Q01; Watts etal. 1995; Piccinini et al.. 1977).
A recent study investigated the role of the endothelial-derived vasoconstrictor, ET-1, in
Pb-induced hypertension. ET-1 from the endothelium acts on the ETA-type receptors
located on the vascular smooth muscle layer and may be involved in vascular reactivity
by NO and cyclooxygenase (COX) derivatives. Pb exposure (1 ppm, 24 hours) to rat
aortic segments decreased expression of sGC-(31 subunit, an enzyme involved in
NO-induced vasodilation, and increased expression of COX-2 in an
endothelium-dependent manner (Molero et al., 2006). Even though Pb treatment did not
alter ET-1 or ETA-type receptor protein expression in this system, blocking the ETA-type
receptors partially reversed Pb-induced changes in sGC and COX-2 in vascular tissue.
These results suggest that the endothelium and ET-1 may contribute to Pb-induced
hypertension through activation of ETA-type receptors that alter expression of COX-2
and sGC-(31 subunit, which affects NO signaling.
COX-2 blockade has been shown to prevent Pb-induced downregulation of sGC
expression (Courtois et al., 2003). Inhibition of COX-2 also decreased the Pb-induced
pressor response to ACh (Grizzo and Cordellini. 2008) and PHE (Silveira et al.. 2010) in
experimental animals. These results suggest that Pb-induced vascular reactivity may
depend on the participation of a COX-derived vasoconstrictor, such as prostaglandins,
prostacyclins, orthromboxanes.
In summary, recent studies continued to show that Pb exposure affects vasomodulatory
pathways that are important for the maintenance of vascular tone; however, results
indicated that not all vascular cell types are similarly affected by Pb exposure. Further,
effects appeared to vary according to the concentration of Pb exposure. Pb exposure has
4-358
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been shown to interrupt baseline or endogenous NO-mediated vasodilation of vessels via
alterations in PKC, sGC, VSMC, endothelial cells, NADPH oxidase, and Ca2+ levels.
Recent studies indicated that Pb exposure may affect vascular reactivity by increasing
COX-2 and COX-2-dependent vasoconstrictors. Also, the vasoconstrictor endothelin may
contribute to Pb-induced vasomodulation via similar pathways as NO including effects
on sGC and COX-2.
4.4.2.4 Summary of Blood Pressure and Hypertension
The 2006 Pb AQCD (U.S. EPA. 2006b) reported a clear association between higher
blood Pb levels and higher BP. The effect was modest, but robust, as determined by a
meta-analysis (Nawrot et al.. 2002) of over 30 cross-sectional and prospective studies
comprising over 58,000 adults (Figure 4-21). In the meta-analysis, each doubling of
concurrent blood Pb was associated with a 1 mmHg increase in systolic BP and a
0.6 mmHg increase in diastolic BP. In addition, U.S. EPA (1990a) reviewed studies
available prior to 1990, a period in the U.S. when Pb exposures from air were probably at
the highest level, that examined Pb exposure and BP outcomes, which included several
studies of the population represented in NHANES II (1976-80). They noted that across a
blood Pb range of 7 to 34 (ig/dL, no evident threshold was found below which the blood
Pb level was not significantly related to BP. U.S. EPA (1990a) concluded that a small but
positive association exists between blood Pb levels and increases in BP.
Recent epidemiologic studies supported this association at lower concurrent blood Pb
levels (in populations with mean blood Pb levels < 2 (ig/dL) and added to the evidence
base regarding populations potentially at increased risk (i.e., high stress, genetic variants)
and regarding associations of bone Pb levels with BP and hypertension in populations
with mean bone Pb levels less than 20 (ig/g. As these studies were mostly cross-sectional
in design and were conducted in adults whose concurrent blood Pb levels are influenced
both by current Pb exposures and past Pb exposures mobilized from bone, uncertainty
exists over the Pb exposure conditions that contributed to the associations observed
between concurrent blood Pb level and increased BP and hypertension (Sections 3.3 and
3.7.3).
A recent prospective study in Pb workers found independent associations of both baseline
blood Pb level and subsequent changes in blood Pb over follow-up with changes in BP
over follow-up and associations of bone Pb level with hypertension (Glenn et al.. 2006).
Although these Pb workers had higher current Pb exposure compared with
nonoccupationally-exposed adults, the results indicated that different mechanisms may
4-359
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mediate shorter-term Pb-associated increases in BP and longer-term Pb-associated
development of hypertension.
Key evidence was further provided by a recent cross-sectional study in an ethnically
diverse community-based cohort of women and men aged 50-70 years that reported
associations between both blood and tibia Pb levels and BP with extensive consideration
of potential confounding factors (Table 4-18) (Martin et al., 2006). Additionally, a recent
epidemiologic study provided evidence for associations in an adult cohort between blood
Pb level and BP and hypertension with relatively low blood Pb levels; a positive
relationship was found in the NHANES adult data (1999-2002) with a geometric mean
concurrent blood Pb level of 1.64 (ig/dL (Muntner et al.. 2005). However, as noted
above, in adults, uncertainty exists regarding the magnitude, timing, frequency, and
duration of Pb exposure that contribute to the associations observed with concurrent
blood Pb levels.
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POCOCK 84
KROMHOUT 85
ORSSAUD 85
WEISS 86
DE KORT 87
i nrkPTT HI \ *
PARKINSON 87 i-»
RABINOWITZ 87
ELWOOD (0 88 (-
ELWOOD (HP) 88 i-
ELWOOD (HP) 88 H
GARTSIDE (W) 88
nABT^inF (Rl ftfi
GARTSIDE (W) 88 H
F, ADTQIDF fRl fifl
ucpl (PW) RR
NERI 88 <
GRANDJEAN 89 i—
fiCJANn IFAN RQ i—
RFIMFR RQ
Aon^Tni i on
APOSTOLI 90
MORRIS 90
MORRIS 90 <
SHARP 181 90
CTAfClQP'M 9
n*
*»- i /-,»
• i O
• 0*
-V-^ 9
• I .-.»
-•^ — i o*
-i 0
n*
• i O
• 0*
• 9
—i 9
h»H 9
»-H 0*
-n o*
H 9
h»H 9
•i 0*9
III 1 1 1 1 I 1
-15 -10 -5 0 5 10 15 20 25 30
SBP
7379
152
431
89
105
116
428
3851
1136
865
856
2705
350
2827
407
288
2193
504
546
58
254
271
145
106
117
132
398
133
439
1703
1661
809
1319
798
339
345
186
2563
2763
1471
1329
439
1188
543
4685
1761
5138
2197
58518
Study Key: C - Caerphilly Study; HP - Welsh Heart Program; W - Whites; B - Blacks; Nl - Non-immigrants; I - Immigrants;
FW- Foundry Workers; CS - Civil Servants; P - PheeCad (Public Health and Environmental Exposure to Cadmium) Study.
Note: Individual study results are presented in each row. The rightmost columns indicate the sex of subjects and study sample size.
Circles represent individual groups and squares represent the combined association sizes. Open circles denote a nonsignificant
association size that was assumed to be zero.
Source: Reprinted with permission of MacMillan Press, Nawrot et al. (2002)
Figure 4-21 Meta-analysis of change in systolic BP (SBP), in mmHg, with 95%
Cl, associated with a doubling in the blood Pb concentration.
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In concordance with epidemiologic evidence, collectively, the animal toxicological
studies providing blood Pb level and BP measurements reported higher BP with higher
blood Pb levels in adult rodents (Figure 4-20). While the contribution of low concurrent
blood Pb levels to the findings is difficult to ascertain in adult humans, animal
toxicological studies provide support for low blood Pb level effects with increases in BP
observed in groups of animals with long-term dietary Pb exposure resulting in blood Pb
levels as low as 2 (ig/dL (Rizzi et al.. 2009; Tsao et al.. 2000; Nakhoul et al.. 1992).
However, the majority of animal toxicological studies showing Pb-induced hypertension
were conducted at higher Pb exposure levels that result in blood Pb levels >10 (ig/dL.
Evidence in animal studies also suggests the potential for increased BP following short-
term (4 weeks) Pb treatment that included injected bolus doses that may have uncertain
relevance to human routes of Pb exposure (Fiorim et al.. 2011; Simdes et al.. 2011;
Sharifi et al.. 2004). These effects may be partially reversible; a study demonstrated
partial reversibility (not to levels in controls) of Pb-induced elevations in BP following
Pb exposure cessation or chelation (Chang et al.. 2005).
Epidemiologic studies continued to investigate the relationship between bone Pb and BP.
A recently published meta-analysis (Figure 4-22) (Navas-Acien et al.. 2008) included
several studies (three prospective, five cross-sectional) that individually showed that bone
Pb level was associated with systolic BP but not diastolic BP. In the cross-sectional
studies, a pooled estimate indicated an increase in systolic BP of 0.26 mmHg (95% CI:
0.02, 0.50) per 10 (ig/g tibia Pb. In the longitudinal studies, a 0.33 mmHg (95% CI: -0.44,
1.11) increase was estimated per 10 (ig/g bone Pb. Most studies also reported associations
of bone Pb with hypertension. Pooled odds ratios for hypertension of 1.04 (95% CI: 1.01,
1.07) per 10 (ig/g increase in tibia Pb and 1.04 (95% CI: 0.96, 1.12) per 10 (ig/g increase
in patella Pb were reported.
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First author, year
Tibia lead
Prospective
Glenn 2006"
Cheng 2001 1!
Glenn 2003"
Cross-sectional
Lee 2001"
Hu 96'VCher>g 01
Martin18 2006
Schwartz" 2000
Korrick15 1999
Patella lead
Cheng" 2001
Hu 96'5/Cheng 01
Komck'«1999
Mean Increase in SBP (95%CI) Increase In DBP (95%CI) Hypertension RR or OR (95%CI)
Lead
(no/a)
38.4 -0.02 (-0.03 to 0.004) 1
21 .9 --
14.7 0.78(0.24101.31)
Overall: 0.33 (-0.44 to 1.11)
.',
Increase i
37.2 0.20 (-0.05 to 0.45)
'32.1 1.01(0.01102.02)
18.8 0.20 (-0.80 to 1.10)
14.4 0.74 (-0.73 to 2.21)
13.3 -
Overall: 0.26 (0.02 to 0.50)
31.4 -
3 32.1 0.29 (-0.36 to 0.95}
17.3 -
i _ :
|
• 0.07 (-0.30 to 0.45) — fc
"
|_
\
11 -1012 BBS 1 1 1 1 S
SBP (mmHg / year) Increase in DBP (mmHg /year) Hypertension RR
-•- -0.02 (-0.20 to 0.17) -1
• 0 "0 ( 0 30 to 0 70)
• - 0 35 ( 0 75 to 1 *l 5)
0.02 (-O.15toO.19)
— 1
-
1.05(1.00101.11)
1.15(0.97101.35)
"* 1.13(0.98101.29)
— • 0.90 (0.70 to 1.17) ^ — •
1.03(1.00101.05)
1.04 (1.01 to 1.07)
1.14(1.01101.28)
1.09 (0.98 to 1.22)
1.00(0.98101.03) |
1.04(0.96101.12)
-•-
•
•
•
•
1
-1012 -1012 0.85 1 1.2 1.5
Increase in SBP (mmHg) Increase in DBP (mmHg) Hypertension OR
In the Normative Aging Study, Hu et al. (1996a) reported the cross-sectional association between bone Pb levels and the
prevalence of hypertension and Cheng et al. (2001) reported the cross-sectional association between bone Pb levels and systolic
BP in study participants free of hypertension at baseline.
Note: The studies are ordered by increasing mean bone Pb levels. The area of each square is proportional to the inverse of the
variance of the estimated change or log relative risk. Horizontal lines represent 95% confidence intervals. Diamonds represent
summary estimates from inverse-variance weighted random effects models. Because of the small number of studies, summary
estimates are presented primarily for descriptive purposes. RR indicates risk ratio.
Source: Reprinted with permission of Lippincott Williams & Wilkins, Inc., Navas-Acien et al. (2008)
Figure 4-22 Meta-analysis of an increase in systolic BP (SBP) and diastolic BP
(DBP) and relative risk of hypertension per 10 ug/g increase in
bone Pb levels.
A few recent epidemiologic studies also emphasized the potential interaction between
measures of long-term Pb exposure, i.e., bone Pb levels, and factors such as chronic
stress and HFE genetic variants to moderate or modify the relationship of BP and
hypertension with Pb. For example, among NAS men, tibia Pb level was associated with
a larger risk of developing hypertension in an originally nonhypertensive group among
men with higher self-reported stress (Peters et al.. 2007).
In addition to stress, recent epidemiologic studies investigated effect modification by
race/ethnicity and genetic variants. In the NHANES 1988-1994 population of adults, the
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association of concurrent blood Pb with systolic BP was higher among Mexican
Americans. In the same NHANES population, the association between blood Pb level and
hypertension was higher among non-Hispanic Blacks with the ALAD2 allele (see Figure
4-17 and Figure 4-18 for results) (Scinicariello et al.. 2010). Additionally, the association
between blood Pb and PP was larger among NAS men with the HFE H63D variant
(Figure 4-17) (Zhang et al.. 2010a). PP represents a good predictor of cardiovascular
morbidity and mortality and an indicator of arterial stiffness. The aforementioned genes
are related to Fe metabolism and have been linked with differences in Pb distribution in
blood and bone. Park et al. (2009a) provided further evidence of variants in
Fe metabolism genes impacting the association of bone Pb levels with QT interval
changes (see Table 4-21 for results).
Animal toxicological evidence continued to build on the evidence characterizing the
mechanisms leading to these Pb-induced cardiovascular alterations. Biological
plausibility for the consistent associations observed between blood and bone Pb and
cardiovascular effects is provided by enhanced understanding of Pb-induced oxidative
stress including NO inactivation, endothelial dysfunction leading to altered vascular
reactivity, activation of the RAAS, and vasomodulator imbalance.
4.4.3 Vascular Effects and Cardiotoxicity
Not only has Pb been shown to increase BP and alter vascular reactivity, but Pb can alter
cardiac function, initiate atherosclerosis, and increase cardiovascular mortality. Past
toxicological studies have reported that Pb can increase atheromatous plaque formation in
pigeons, increase arterial pressure, decrease heart rate and blood flow, and alter cardiac
energy metabolism and conduction (Prentice and Kopp. 1985; Revis etal., 1981). A
limited number of available epidemiologic studies discussed in the 2006 Pb AQCD (U.S.
EPA, 2006b) provided evidence of associations of blood Pb level with ischemic heart
disease (IHD) and peripheral artery disease (PAD).
4.4.3.1 Effects on Vascular Cell Types
The endothelial layer is an important constituent of the blood vessel wall, which regulates
macromolecular permeability, VSMC tone, tissue perfusion, and blood fluidity. Damage
to the endothelium is an initiating step in development of atherosclerosis, thrombosis, and
tissue injury. Given that epidemiologic and toxicological evidence suggests that long-
term Pb exposure is associated with a number of these conditions, numerous
toxicological studies have investigated and found an effect of Pb on endothelial
4-364
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dysfunction. A recent occupational study found that endothelial function assessed by
flow-mediated dilatation was impaired in highly Pb-exposed workers (mean blood Pb
levels: 24.1 (ig/dL in workers versus 7.8 (ig/dL in unexposed controls) (Por^baet al.,
2010).
The endothelial layer makes up only a small part of the vascular anatomy; the majority of
the vessel wall is composed of VSMCs, which work in concert with the endothelial cells
(EC) in contraction and relaxation of the vessel, local BP regulation, and atherosclerotic
plaque development. Since Pb has been shown repeatedly to result in hypertension and
vascular disease in experimental animals, studies continued to investigate and find an
effect of Pb on VSMCs.
In in vitro assays, Pb (50 (iM, 2 weeks) stimulated VSMC invasiveness in isolated human
arteries leading to the invasion of medial VSMC into the vessel intima and development
of intimal hyperplasia, a key step in atherosclerotic progression (Zeller et al.. 2010). In
addition, treatment with Pb (50 (JVI, 12 hours) promoted VSMC elastin expression and
increased arterial extracellular matrix in isolated human arteries. VSMC invasiveness was
also increased in culture by treatment with supernatant of Pb-treated human EC (50 (iM),
suggesting that Pb-exposed ECs secrete an activating compound. This compound was
confirmed to be IL-8. Pb exposure (5-50 (iM) was able to, in a concentration-dependent
manner, increase IL-8 synthesis and secretion in human umbilical vein EC cultures
through activation of the transcription factor Nrf2. Neutralization of IL-8 could block
VSMC invasion and arterial intima thickening (Zeller et al.. 2010). This study provides
evidence that Pb exposure stimulates ECs to secrete IL-8 in an Nrf2-dependent manner
which stimulates VSMC invasion from the vessel media to intima leading to a vascular
thickening and possibly atherogenesis.
A number of CVDs, including atherosclerosis, are characterized by increased
inflammatory processes. Numerous studies have shown that Pb exposure is associated
with an inflammatory environment in vascular tissues of humans and animals as indicated
by higher levels of inflammatory mediators like prostaglandin E2 (PGE2). Human aortic
VSMCs treated with Pb (1 (iM, 1-12 hours) exhibited increased secretion of PGE2 time-
dependently through enhanced gene transcription (Chang etal.. 2011). This was preceded
by a Pb-induced increase in the gene expression of cytosolic phospholipase A2 (cPLA2)
and COX-2, two rate limiting enzymes in the regulation of prostaglandins. The induction
of these enzymes was mediated by activation of ERK1/2, MEK1, and MEK2. Further
investigation of the entrance of Pb into the cell revealed that inhibition of the store-
operated calcium channels (SOC) could only partially suppress cPLA2 and COX
activation by Pb; however, inhibition of epidermal growth factor receptor (EGFR)
attenuated Pb-induced PGE2 secretion and activation of cPLA2 and COX. A follow-up to
4-365
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this study found that Pb treatment (1(JVI) of a human epithelial cell line increased COX-2
gene expression, promoter activity, and protein (Chou et al.. 2011). Inhibition of NF-KB
decreased the Pb-induced COX-2 activation; whereas EGFR inhibition blocked COX-2
upregulation and NF-KB nuclear translocation. Overall these results suggest that Pb can
induce pro-inflammatory events in VSMC in the form of increased PGE2 secretion and
expression of cPLA2 and COX-2 through activation of EGFR via ERK1/2 and NF-KB
pathways.
Damage to the endothelium is a hallmark event in the development of atherosclerosis.
Past studies have shown that Pb exposure results in de-endothelialization, impaired
proliferation, and inhibition of endothelium repair processes after injury (Fujiwara et al..
1997; Uedaetal.. 1997; Kaiietal.. 1995; Kishimoto et al.. 1995). However, Pb exposure
was not found to lead to nonspecific cytotoxicity at low exposure levels (2-25 (iM) as
shown by the lack of release of lactate dehydrogenase (LDH) from Pb-treated bovine
aortic EC (Shinkai et al.. 2010). Instead, Pb induced specific apoptosis of EC (caspase3/7
activation) through endoplasmic reticulum (ER) stress that was protected against by the
ER chaperones glucose-regulated protein 78 (GRP78) and glucose-regulated protein 94
(GRP94). GRP78 and GRP94 play key roles in the adaptive unfolded protein response
that serves as a marker of and acts to alleviate ER stress. Exposure of ECs to Pb induced
GRP78 and GRP94 gene (2-25 \M) and protein (GRP78 [5-25 \M\ and GRP94
[10-25 (JVI]) expression through activation of the IREl-JNK-AP-1 pathways (Shinkai et
al.. 2010). This finding suggests that the functional damage in ECs caused by Pb
exposure may be partly attributed to induction of ER stress.
4.4.3.2 Cholesterol
As blood cholesterol rises so does the risk of coronary heart disease. Previous
occupational studies (Ademuviwa et al.. 2005a; Bener et al.. 200Ib; Kristal-Boneh et al..
1999) examining higher than most current adult blood Pb levels (>40 (ig/dL) reported
higher total cholesterol levels related to Pb exposure, but mixed results for HDL, LDL,
and triglycerides. More recently, Poreba et al. (2010). in an occupational study, reported
no significant differences in parameters of lipid metabolism between Pb exposed workers
(mean blood Pb level: 25 (ig/dL) and unexposed individuals. Conversely, Kamal et al.
(2011) reported that occupational Pb exposure (mean blood Pb level: >40 (ig/dL) was
associated with higher levels of triglycerides, total cholesterol, and LDL, and decreased
HDL-C. Other Pb studies adjusted models for total cholesterol to control for this coronary
heart disease risk factor. Higher mean total cholesterol with higher blood Pb levels has
been reported in an NHANES study (Menke et al.. 2006). In developing models to
predict bone Pb levels, Park et al. (2009b) noted in an NAS study that total and HDL
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cholesterol were selected as 2 of 18 predictors for the bone Pb level model. Their findings
suggested that higher Pb exposure in nonoccupationally-exposed men may be associated
with higher total and HDL cholesterols. In support of epidemiologic evidence, a recent
toxicological study reported increased LDL and decreased HDL in rats treated with Pb
(20 ppm, i.p., 3 days/week, 8 weeks) (Roshan et al., 2011). The major risk factor that
lipids represent for heart disease make relating lipid levels to Pb exposures an interesting
but challenging hypothesis to test.
4.4.3.3 Atherosclerosis
A small number of toxicological and cross-sectional epidemiologic studies provide
evidence for increased atherosclerosis and intimal medial thickening (IMT) due to Pb
exposure. The association of stroke subtypes and severity of cerebral atherosclerosis was
examined in relation to a single concurrent blood Pb level and total 72-hour urinary Pb
level (body Pb store-EDTA mobilization test) in a cross-sectional study of 153 patients
(mean age 63.7 years) receiving digital subtraction angiography in Chang Gung
Memorial Hospital in Taiwan from 2002 to 2005 (Lee et al., 2009). In an analysis
adjusted for age, sex, hypertension, diabetes, triglyceride, uric acid, smoking, and alcohol
consumption, a 1 (ig increase in urine Pb was associated with > 50% stenosis in the
intracranial carotid system with an OR of 1.02 (95% CI: 1.00, 1.03). Urine Pb was not
associated with greater stenosis in the extracranial or vertebrobasilar systems. Blood Pb
level was not associated with greater stenosis in any region. As the development of
atherosclerosis is a lifelong process, body Pb stores, analyzed by total 72-hour urine Pb
amount, may more strongly be associated with atherosclerosis than are single blood Pb
measurements.
A recent study correlated greater carotid artery IMT with higher concurrent serum Pb
levels (mean [SD] 0.41 [0.38] (ig/dL) in hemodialysis patients (Ari etal.. 2011). A few
available recent occupational studies also presented evidence for increased measures of
atherosclerosis in highly Pb-exposed adult populations with mean blood Pb levels around
25 (ig/dL. Por^ba et al. (2011 a) reported increased local arterial stiffness and more
frequent left ventricular diastolic dysfunction in Pb-exposed workers with hypertension
compared to non-exposed controls with hypertension. Occupational exposure to Pb
(mean blood Pb levels: 24 (ig/dL in workers, 8.3 (ig/dL in non-exposed group) was also
associated with greater IMT and atherosclerotic plaque presentation, analyzed by Doppler
ultrasound (Poreba et al., 2011).
Zeller et al. (2010) examined human radial and internal mammary arteries exposed to Pb
in culture and reported a concentration-dependent increase in arterial intimal thickness
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(statistically nonsignificant at 5 (iM Pb, significant at 50 (iM Pb, 2 week treatment) and
intimal extracellular matrix accumulation (50 (iM). Also, Pb promoted EC proliferation
(5 and 50 (JVI, 72 hours) and VSMC elastin expression (50 (JVI, 12 hours), as discussed
above (Section 4.4.3.1) (Zeller et al.. 2010). Another study showed that Pb exposure
(100 ppm in drinking water for 10 months; mean blood Pb level 28.4 (ig/dL) of rats also
increased the aortic media thickness, media-lumen ratio, and medial collagen content
(Zhang et al., 2009a). These morphological changes to the vessel due to Pb exposure
indicate initiation of arteriosclerosis and could be the cause of decreased contractile
response of the vessel due to altered visco-dynamic vessel properties. Alternatively, these
vascular changes could be an effect of Pb-induced hypertension.
4.4.3.4 Heart Rate Variability
HRV and BP are regulated, in part, by the sympathetic and parasympathetic nervous
systems. Changes in either may increase the risk of cardiovascular events. HRV is
defined as the oscillation in the interval between consecutive heart beats and between
consecutive instantaneous heart rates. Decreases in HRV have been associated with
cardiovascular mortality/morbidity in older adults and those with significant heart disease
(Task Force of the European Society of Cardiology and the North American Society of
Pacing and Electrophysiology. 1996). In addition, decreased HRV may precede some
clinically important arrhythmias, such as atrial fibrillation, as well as sudden cardiac
death, in high risk populations (Chen and Tan. 2007; Sandercock and Brodie. 2006).
Pb has been shown not only to affect vascular contractility in animals, but also is
associated with cardiac contractility. The 2006 Pb AQCD (U.S. EPA. 2006b) described
one study that investigated Pb-induced alterations in HRV (Cheng etal. 1998). Cheng et
al. (1998) found increasing duration of corrected QT interval (QTc) with increasing bone
Pb levels in men <65 years, but not in men > 65 years. Bum et al. (2011) and Park et al.
(2009a) followed up on this previous NAS cohort (Cheng etal.. 1998) (details found in
Table 4-21). Bum et al. (2011) prospectively examined the association between blood and
bone Pb levels and the development of electrocardiographic (ECG) conduction
abnormalities among 600 men who were free of ECG abnormalities at the baseline
assessment. A second ECG was obtained for 496 men 8.1 (SD: 3.1) years later on
average. Baseline Pb concentrations in blood (mean [SD]: 5.8 [3.6] (ig/dL), patella bone
(mean [SD]: 30.3 [17.7] (ig/g), and tibia bone (mean [SD]: 21.6 [12.0] (ig/g) were similar
to those found in other samples from the general U.S. adult population and much lower
than those reported in occupationally exposed groups. Higher tibia Pb was associated
with increases in QTc interval and QRSc duration. Compared with those in the lowest
tertile of baseline tibia Pb (<16 (ig/g), participants in the highest tertile (>23 (ig/g) had a
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7.94 msec (95% CI: 1.42, 14.45) greater increase in QTc interval and a 5.94 msec (95%
CI: 1.66, 10.22) greater increase in QRSc duration over 8 years after adjusting for
covariates: age, education, smoking, BMI, albumin-adjusted serum calcium, and diabetes
status at baseline, and years between ECG tests and QT-prolongation drugs at the time of
ECG measurement. There were no statistically significant associations with patella or
blood Pb levels. These associations with tibia bone Pb levels were observed in men with
relatively low blood and bone Pb concentrations who were free of cardiac conduction
abnormalities at baseline and were examined prospectively. Thus, they indicate that long-
term cumulative Pb exposure may increase the risk of developing cardiac abnormalities.
Uncertainty exists as to the specific Pb exposure level, timing, frequency, and duration
contributing these associations observed for tibia Pb levels. A recent occupational study
reported lower HRV and abnormal parameters of heart rate turbulence in Pb-exposed
workers (mean blood Pb levels: ~25 (ig/dL) compared to control subjects (Poreba et al..
201 Ib).
Park et al. (2009a) cross-sectionally examined whether polymorphisms in genes known to
alter Fe metabolism (HFE, transferrin [TF] C2, heme oxygenase-1 [HMOX-1]) modify
the association between Pb biomarker levels and the QT interval. Investigators examined
associations in data stratified on polymorphisms in the three genes. They also analyzed
interaction models with cross-product terms for genotype and the Pb biomarker. The
distributions of all genotypes but the HFE variant, H63D, were in Hardy-Weinberg
equilibrium. Subjects homozygous for the other HFE variant, C282Y, had higher bone Pb
levels and those homozygous for H63D and heterozygous with both C282Y and H63D
had lower bone Pb levels. The antioxidant HMOX-1 L variant (longer repeats of GT,
associated with lower enzyme inducibility) alone, compared to the wild type, showed a
statistically significant interaction with tibia Pb (11.35 msec longer QTc interval for each
13 (ig/g increase in bone Pb in L-allele variants). No other gene variant alone showed
different Pb-associated QTc intervals from those in wild types, either for tibia and patella
Pb or for (linear) concurrent blood Pb. Lengthening of QTc with higher tibia and blood
Pb was more pronounced with an increase in the total number of gene variants, driven by
a joint effect between HFE variants and HMOX-1 L allele. There was a trend observed
with blood and tibia Pb-associated QTc interval increasing with increasing number of
gene variants from 0 to 3. This study provided further evidence of gene variants
modifying associations of Pb biomarkers with cardiovascular effects.
The interaction of key markers of the metabolic syndrome with bone Pb levels in
affecting HRV was cross-sectionally investigated in a group of 413 older adults with
patella Pb measurements in the NAS (Park et al.. 2006). Metabolic syndrome was defined
to include three or more of the following: waist circumference >102 cm,
hypertriglyceridemia (>150 mg/dL), low HDL cholesterol (<40 mg/dL in men), high BP
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>130/85 mmHg, and high fasting glucose (>110 mg/dL). Men using antihypertensive
medication or diabetes medications were counted as high BP or high fasting glucose,
respectively. The strongest relationships between patella Pb levels and lower HRV were
observed among those with three or more metabolic abnormalities. A trend was observed
for larger patella Pb-associated decreases in HRV with increasing number of metabolic
abnormalities. These results suggest multiplicative effects of cumulative Pb exposure and
metabolic abnormalities on key predictors of CVD. Park et al. (2006) also reported the
penalized spline fits to bone Pb in models assessing only main effects of bone Pb. The
optimal degree of smoothing determined by the generalized cross-validation criterion for
all HRV measures was 1, which indicated that the associations were nearly linear. The
spline fits and associated statistics showed that the bone Pb main effects on HRV
measures were linear. However, the relationship with the LF/HF HRV measure was
linear with log(LF/HF).
Increased incidence of arrhythmia and atrioventricular conduction block was found in
rats after 12 weeks of Pb exposure (100 ppm; mean blood Pb level 26.8 (ig/dL) (Reza et
al.. 2008). Also, Pb exposure for 8 weeks increased heart rate and systolic BP. These
increases corresponded with increased cardiac contractile force and prolonged ST
interval, without alteration in QRS duration or coronary flow. In contrast, another study
using rat right ventricular strips found that Pb (100 (iM) exposure, in a concentration-
dependent manner, reduced myocardial contraction by reducing sarcolemmal Ca2+ influx
and myosin ATPase activity (Vassallo et al.. 2008). This study also found that Pb
exposure changed the response to inotropic agents and blunted the force produced during
contraction. Conversely, past studies have found that Pb exposure increases intracellular
Ca2+ content (Laletal.. 1991; Favalli et al.. 1977; Piccinini et al.. 1977). which could
result in increased cardiac output and hypertension.
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Table 4-21 Characteristics and quantitative data for associations of blood and bone Pb with other CVD
measures, i.e., HRV, PAD, and IHD in recent epidemiologic studies.
Study
(Ordered as they
appear in the text)
Study Population/
Methodology
Parameter
Pb Data
Statistical Analysis
Effect Estimate
(95% Cl)a
Heart rate variability
Eum et al. (2011)
Longitudinal
600 elderly men
(mostly white, mean
age 67 yr) free of
electrographic
abnormalities at the
time of baseline
ECG enrolled in
MAS in Greater
Boston, MA area
(496 with follow-up
ECG 8 years later)
ECG conduction
(QTc,
QRSc,
JTc,
QT prolongation,
JT prolongation,
IVCDC,
AVCD,
Arrhythmia)
Baseline Blood Pb:
Mean (SD): 5.8 (3.6) ug/dL
Baseline Patella Pb:
Mean(SD): 30.3 (17.7) ug/g
Baseline Tibia Pb:
Mean (SD): 21.6 (12.0) ug/g
Q1: <16 ug/g (n = 191)
Q2: 16.0-23 ug/g (n = 208)
Q3: >23ug/g(n = 195)
Repeated measures linear
regression adjusted for
age, education, smoking,
BMI, albumin-adjusted
serum Ca2+, and diabetes
status at baseline, years
between ECG tests, and
QT-prolongation drugs at
the time of ECG
measurement.
Tibia Pb:
Adjusted 8-year change:
QTc:
Q2 vs. Q1 (reference):
7.49(1.22, 13.75) msec,
Q3vs. Q1:
7.94(1.42, 14.45) msec
p for trend = 0.03
QRSc:
Q2vs. Q1:
0.52 (-3.60, 4.65) msec
Q3vs. Q1:
5.94(1.66, 10.22) msec
p for trend = 0.005
No associations with patella or
blood Pb
Park et al.
(2009a)
Cross-sectional QTcb interval
613 elderly men
(mostly white, mean
age 67 yr) enrolled
in MAS in Greater
Boston, MA area
(8/1991 -12/1995)
Blood Pb:
Median (IQR):
5 (4-7) ug/dL
Patella Pb:
Median (IQR):
26(1 8-37) ug/g
Tibia Pb:
Median (IQR):
19 (14-27) ug/g
Linear regression models
adjusted forage, BMI,
smoking status, serum
Ca2+, and diabetes. No
SES indicator was
considered.
Per IQR (3 ug/dL)
increase in blood Pb:
1.3 (-0.76, 3.36) msec
after 8-year follow up
Per IQR (19 ug/g)
increase in patella Pb:
2.64(0.13, 5. 15) msec
Per IQR (13 ug/g)
increase in tibia Pb:
2.85 (0.29, 5.40) msec
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Table 4-21 (Continued): Characteristics and quantitative data for associations of blood and bone Pb with other CVD
measures, i.e., HRV, PAD, and IHD in recent epidemiologic studies.
Study
(Ordered as they
appear in the text)
Study Population/
Methodology Parameter Pb Data
Statistical Analysis
Effect Estimate
(95% Cl)a
Park et al. (2006)
Cross-sectional
413 elderly men
(mostly white, mean
age 67 yr) enrolled
in MAS in Greater
Boston, MA area
(11/14/2000-
12/22/2004)
HRV
(SDNN, HF, HFnorm,
LF, LFnorm, LF/HF)
Tibia Pb:
Median (IQR):
19.0(11-28) |jg/g
Patella Pb (measured within
6 mo of HRV:
Median (IQR):
23.0(15-34) |jg/g
Estimated3: Median (IQR):
16.3(10.4-25.8)|jg/g
Log linear regression
models adjusted for age,
cigarette smoking, alcohol
consumption, room
temperature, season
(model 2) BMI, fasting
blood glucose, HDL
cholesterol, triglyceride,
use of (3-blockers,
Ca2+ channel blockers,
and/or ACE inhibitors. No
SES indicator was
considered.
PerlQR(17ug/g)tibiaPb,
Model 2
HF:-0.9 (-3.8, 2.1)
normalized units (nu)
LF: 0.9 (-2.0, 3.9) nu
Log LF/HF:
3.3 (-10.7, 19.5)(%)
Per IQR (15.4) ug/g estimated
patella Pb, Model 2
HF:-0.6 (-3.1, 1.9)nu
LF: 0.6 (-1.9, 3.1) nu
Log LF/HF: 3.0 (-8.7, 16.2)
Effect estimates were more
pronounced among those with
greater* metabolic
abnormalities.
Peripheral artery disease
Muntner et al.
(2005)
Cross-sectional
9,961 NHANES
(1999-2002)
participants
PAD
Range Concurrent Blood Pb: Logistic regression models OR
Q1: <1.06ug/dL,
Q2: 1.06-1.63 ug/dL
Q3: 1.63-2.47 ug/dL
Q4: >2.47 ug/dL
adjusted for age, Q-|:
race/ethnicity, sex, diabetes Q2:
mellitus, BMI, cigarette Q3:
smoking, alcohol Q4:
consumption, high school
education, health insurance
status
1.00 (Reference),
1.00(0.45, 2.22),
1.21 (0.66, 2.23),
1.92(1.02, 3.61)
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Table 4-21 (Continued): Characteristics and quantitative data for associations of blood and bone Pb with other CVD
measures, i.e., HRV, PAD, and IHD in recent epidemiologic studies.
Study
(Ordered as they Study Population/
appear in the text) Methodology Parameter Pb Data
Statistical Analysis
Effect Estimate
(95% Cl)a
Navas-Acien et
al. (2005)
Cross-sectional
790 participants,
age > 40 yr, from
NHANES
(1999-2000)
PAD
Concurrent urinary Pb:
Mean (10th-90th percentile):
0.79 (0.2-2.3) ug/L
Logistic regression
adjusted for the following:
Model 1: age, sex, race,
and education
Model 2: covariates above
plus smoking status
Model 3:covariates above
plus urinary creatinine
Per IQR increase in urinary Pb
Model 1:
OR: 1.17(0.81, 1.69)
Model 2:
OR: 1.17(0.78, 1.76)
ModelS:
OR: 0.89(0.45, 1.78)
Array of metals in urine also
evaluated.
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Table 4-21 (Continued): Characteristics and quantitative data for associations of blood and bone Pb with other CVD
measures, i.e., HRV, PAD, and IHD in recent epidemiologic studies.
Study
(Ordered as they
appear in the text)
Study Population/
Methodology Parameter Pb Data
Statistical Analysis
Effect Estimate
(95% Cl)a
Ischemic Heart Disease
Jain et al. (2007)
Longitudinal
837 elderly men
(mostly white, mean
age 67 yr) enrolled
in MAS in Greater
Boston, MA area
(1991-2001)
IHD
(Ml or angina
pectoris)
Baseline Blood Pb Mean
(SD):
Non-cases
6.2 (4.3) |jg/dL;
Cases
7.0 (3.8) |jg/dL
Baseline Patella Pb Mean
(SD):
Non-cases
30.6(19.7) |jg/dL;
Cases
36.8 (20.8) |jg/dL
Cox proportional hazards
models adjusted for age,
BMI, education, race,
smoking status, pack-years
smoked, alcohol intake,
history of diabetes mellitus
and hypertension, family
history of hypertension,
DBP, SBP, serum
triglycerides, serum HDL,
and total serum cholesterol
Blood Pb level > 5 ug/dL
Per 1 SD increase in Pb
biomarker
OR over 10-year follow-up:
1.73(1.05,2.87)
Ln [blood Pb]
OR: 1.45(1.01,2.06)
Ln [patella Pb level]
OR: 2.64 (1.09, 6.37)
Ln [tibia Pb level ]
OR: 1.84(0.57, 5.90)
Baseline Tibia Pb
Mean (SD):
Non-Cases
21.4(13.6) ug/g;
Cases
24.2 (15.9) ug/g
Cases:
Blood Pb range:
1.0to20.0ug/dL
Patella Pb range:
5.0 to 101 ug/g
Tibia Pb range:
-5 to 75 ug/g
Estimated patella Pb accounts for declining trend in patella Pb levels between analysis of bone Pb and HRV.
Heart-rate-corrected QT interval calculated by Bazett's formula
°IVCD, intraventricular conduction defect; AVCD, atrioventricular conduction defect
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4.4.3.5 Peripheral Artery Disease
Peripheral artery disease (PAD) is an indicator of atherosclerosis and measured by the
ankle brachial index, which is the ratio of BP between the posterior tibia artery and the
brachial artery. PAD is typically defined as an ankle brachial index of less than 0.9.
Muntner et al. (2005). whose results describing the association of blood Pb and
hypertension in the NHANES 1999-2002 data set for adults were discussed previously,
also examined the association of blood Pb with PAD (details found in Table 4-21). The
authors observed an increasing trend in the odds of PAD with increasing concurrent
blood Pb level. The OR for PAD comparing the fourth quartile of blood Pb (>2.47 (ig/dL)
to the first quartile of blood Pb (<1.06 jig/dL) was 1.92 (95% CI: 1.02, 3.61). Key
potential confounding factors Table 4-21) were adjusted for in the analysis. These results
are consistent with those from a previous NHANES analysis by Navas-Acien et al.
(2004) reviewed in the 2006 Pb AQCD.
Navas-Acien et al. (2004) reported a trend of increasing OR for PAD with increasing
quartile of concurrent blood Pb or Cd in adults who were 40 years of age in the
1999-2000 NHANES population. These authors tested both Pb and Cd in separate
models, tested the metals simultaneously, and tested the interaction between the metals.
The correlation coefficient between natural log Pb and natural log Cd was 0.32
(p <0.001). Although the interaction was not statistically significant, when blood Pb and
blood Cd were in the same model, the ORs were diminished slightly. Both showed
statistically significant trends of increasing OR with increasing quartile of the metal.
These results indicate that blood Cd levels did not confound the association between
blood Pb level and PAD. In a subsequent analysis, Navas-Acien et al. (2005) used the
same 1999-2000 NHANES dataset, but constructed PAD models using a suite of urine
metal concentrations. Power was reduced in this study because only 659-736 subjects
(compared to 2,125) had spot urine metal tests in the data set. Urinary Cd, but not urinary
Pb, was consistently associated with PAD in all models. Associations also were observed
with urinary antimony and tungsten. Spot urine Pb measurements are less reliable
compared to blood Pb measurements (Section 3.3.3). In Navas-Acien et al. (2005). the
urinary Pb level association with PAD was sensitive to adjustment for urinary creatinine,
indicating that spot urine Pb measurements are affected by differences in urine dilution.
This finding illustrates the limited reliability of spot urine Pb measurements compared to
blood Pb measurements.
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4.4.3.6 Ischemic Heart Disease
A few cross-sectional studies discussed in the 2006 Pb AQCD (U.S. EPA. 2006b)
indicated associations between Pb biomarker levels and increased risk of cardiovascular
outcomes associated with IHD, including left ventricular hypertrophy (Schwartz. 1991)
and myocardial infarction (Gustavsson et al., 2001). Recently, Jain et al. (2007) provided
prospective evidence for the incidence of IHD (physician confirmed MI, angina pectoris)
among older adult males enrolled in the NAS who were followed during the period of
September, 1991 to December, 2001 (details found in Table 4-21). All subjects had blood
Pb and bone Pb measurements with no IHD at enrollment. Fatal and nonfatal cases were
combined for analysis. Baseline blood, tibia, and patella Pb levels were log-transformed.
Blood Pb level and patella Pb level were associated with increased risk of IHD over the
10-year follow-up period. When blood Pb and patella Pb were included simultaneously in
the model, each of their hazard ratios (HRs) was only moderately attenuated (HR: 1.24
[95% CI: 0.80, 1.93] per SD increase in blood Pb and HR: 2.62 [95% CI: 0.99, 6.93] per
SD increase in patella Pb). When blood Pb and tibia Pb were included simultaneously in
the model, their risk estimates were only moderately attenuated (HR: 1.38 [95% CI: 0.89,
2.13] per SD increase in blood Pb and HR: 1.55 [95% CI: 0.44, 5.53] per SD increase in
tibia Pb). These findings indicate that both blood and bone Pb levels are independently
associated with IHD incidence.
IHD, characterized by reduced blood supply to the heart, may result from increased
thrombosis. In support of the epidemiologic evidence, a recent animal study suggested
that Pb exposure promotes a procoagulant state that could contribute to thrombus
formation (Shin et al., 2007). In a rat model of venous thrombosis, Pb treatment (i.v.
25 mg/kg) resulted in increased thrombus formation, although i.v. Pb treatment may have
uncertain relevance to human routes of Pb exposure. Additionally, Pb treatment to human
erythrocytes (red blood cells, RBCs) increased coagulation at a dose of 5 (iM and
thrombin generation in a concentration-dependent manner at doses from 2-5(iM. This
enhanced procoagulant activity in Pb-treated RBCs was the result of increased outer cell
membrane phosphatidylserine (PS) surfacing (human RBCs: 2-5 (iM Pb; rat RBCs: 5 (iM
Pb). Similar to these in vitro results, PS externalization on erythrocytes was increased in
Pb-treated rats (i.v. 50-100 mg/kg, not 25 mg/kg). Increased PS externalization was likely
the result of increased intracellular calcium (5 (iM Pb), enhanced scramblase activity
(5-10 (iM Pb), inhibited flippase activity (5-10 (iM Pb), and ATP depletion (1-5 (iM Pb)
after Pb exposure (Shin et al.. 2007).
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4.4.3.7 Summary of Vascular Effects and Cardiotoxicity
There are a limited number of studies in a limited number of populations that investigate
the associations between Pb biomarkers and cardiotoxicity and cardiovascular effects
other than BP or hypertension (Table 4-21). As presented in Table 4-21. these studies
demonstrated associations between various biomarkers of Pb exposure and clinical
cardiovascular outcomes such as atherosclerosis, IHD, PAD, and HRV occurrence in
adult populations after adjusting for potential confounding by variables such as age, sex,
education, BMI, smoking, alcohol consumption, and diabetes. In a limited body of
studies, mixed evidence of association between occupational exposure to Pb and altered
cholesterol was reported.
A limited number of studies have evaluated markers of subclinical atherosclerosis such as
PAD (i.e., ankle-brachial index) and generalized atherosclerosis (i.e., IMT) following Pb
exposure in humans or animals. Concurrent blood Pb levels (population means
>2.5 (ig/dL) were associated with greater odds of PAD in adults in NHANES analyses
(Muntner et al.. 2005; Navas-Acien et al.. 2004). Since these effects were observed in
adults that may have had higher past exposure to Pb, there is uncertainty as to the specific
Pb exposure level, timing, frequency, and duration that contributed to the observed
associations. A recent study involving both human and toxicological studies observed
Pb-mediated arterial IMT, an early event in Pb-induced atherogenesis (Zeller et al..
2010). A second study in rats reports increased aortic media thickness following Pb
exposure (Zhang et al.. 2009a). Toxicological studies of Pb-induced endothelial
dysfunction, VMSC invasiveness, and inflammation in isolated vascular tissues and cells
provide mechanistic evidence to support the biological plausibility of these vascular
effects and cardiotoxicity. Studies in isolated tissues and cells found that Pb stimulated
the synthesis and secretion of IL-8 in ECs, which was responsible for stimulating VSMC
invasion into the vessel intimal layer. Pb treatment also increased extracellular matrix and
elastin, primary sites for lipid deposition in the vessel wall.
Several studies of the NAS cohort report associations between biomarkers of Pb exposure
and diseases associated with coronary heart disease (CHD), such as HRV, IHD, and MI.
A prospective NAS study reported that higher baseline tibia Pb was associated with
increases in QTc interval and QRSc duration over an 8-year follow-up period (Bum et al..
2011). In addition, in the NAS cohort of older adult men, blood Pb (> 5 (ig/dL) and
patella Pb levels were associated with increased incidence of IHD (Jain et al.. 2007). A
recent study provided evidence for the interaction between biomarkers of Pb exposure
and the HFE C282Y and HMOX-1 L variants on the prolonged QT interval in
nonoccupationally-exposed older men (Park et al.. 2009a). Also, in the NAS population,
bone Pb levels were associated with larger decreases in HRV parameters among subjects
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identified as having metabolic abnormalities (Park et al., 2006). These metabolic
abnormalities, abdominal obesity, hypertriglyceridemia, low HDL cholesterol, high
BP/medication use, or high fasting glucose, have been shown to be associated with
increased risk of cardiovascular events.
Overall, the relatively few available studies provide support for associations between Pb
biomarkers and other cardiovascular conditions including subclinical atherosclerosis and
CHD. A number of these are quality studies from two cohorts, NAS and NHANES with
adequate sample size that account for potential confounding, with some being conducted
prospectively.
4.4.4 Cardiovascular Function and Blood Pressure in Children
4.4.4.1 Introduction
The study of cardiovascular function effects in relation to blood Pb levels in children
potentially offers unique information on several topics. First, this research examines
endpoints that may be predictive of future cardiovascular pathology, and consequently
may offer information on the potential cardiovascular effects of Pb exposure in an
understudied population, children, in whom cardiovascular events are not likely. Second,
examination of cardiovascular changes that may be antecedent to later-in-life effects,
such as increased BP and changes in other CVD-related endpoints, may inform the
understanding of the time course of cardiovascular changes associated with Pb exposure.
Finally, these studies address gaps in knowledge regarding Pb exposure effects in
populations of children (ages 5-15 years and 19-29 years) with mostly concurrent mean
blood Pb levels in the range of 1-5 (ig/dL.
An important aspect to the literature about the association between cardiovascular effects
and blood Pb levels in children is that the blood Pb levels of children may better reflect
relatively recent Pb exposure and its effect on CVD than blood Pb levels do in adults
because of the much longer exposure history of adults during which Pb exposures were
commonly much higher than they are today. However, in older children there is still
uncertainty regarding the frequency, duration, timing, and magnitude of exposure
contributing to the blood Pb levels measured. The much lower prevalence of
cardiovascular effects in children, however, poses a challenge to investigations of
potential relationships with Pb exposures. For example, the prevalence of hypertension in
children (9 to 10 years old) ranges from to 2 to 5 percent (Daniels. 2011; Steinthorsdottir
et al.. 2011). while more than half of people aged 60 to 69 years have hypertension
(Chobanian et al.. 2003). Accordingly, much larger study populations are required to
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provide similar statistical power for such studies in children as compared to adult studies.
Further, in drawing interpretations from such studies with regard to potential effects of Pb
exposures at later ages, it is additionally important to recognize that compensatory
mechanisms in children may be more active than in adults, and the cardiovascular tissue
of the young may be less susceptible to damage than that of adults.
The limited numbers of cardiovascular studies published on children have examined
endpoints such as total peripheral resistance (TPR), BP, and autonomic nervous system
activation. These recent and earlier studies are presented in Table 4-22. Multiple single
pollutant studies in New York State evaluated two child cohorts born in the 1990s after
Pb was removed from gasoline in the U.S. with mean blood Pb levels of 4.62 and
1.01 (ig/dL (Gump etal.. 2011; Gump et al.. 2009; Gump et al. 2007; Gump et al.. 2005).
The cohort examined in Gump et al. (2009; 2007; 2005) was a subset of the Oswego
Children's Study designed as a prospective, prenatal polychlorinated biphenyls study.
Gump et al. (2005) reported cardiovascular outcomes associated (p<0.2) with other
prenatal toxicants (i.e., mercury, dichlorodiphenyldichlorethylene, hexachlorobenzene);
however, the effects of postnatal and prenatal Pb were found even after covariate control
for all non-Pb toxicants that were related to the outcome. Zhang et al. (2011 a) examined
children in Mexico City born from 1994 to 2003, when Pb was being taken out of
gasoline in Mexico as indicated by Martinez et al. (2007). The geometric means for cord
and concurrent blood Pb levels of the children (ages 7-15 years) in the Mexico City
cohort were 4.67 and 2.56 (ig/dL, respectively.
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Table 4-22 Studies of child cardiovascular endpoints and Pb biomarkers.
Study Study
(Ordered as they Population/
appear in the text) Methodology
Parameters
Blood Pb Data3
Statistical Analysis
Effect Estimates/Results
Gump et al.
(2005)
Prospective
122 children age
9.5 yr in Oswego,
NY (born at a
single hospital in
New York from
1991-94)
SBP, DBP, SV
(stroke volume),
CO (cardiac
output), TPR
(total peripheral
vascular
resistance)
Cord blood Pb:
GM (GSD):
2.56 ug/dl_ (1.16)
Childhood (mean age of
measurement: 2.6 yr)
blood Pb: GM (GSD):
4.06 ug/dl_ (1.14)
MDL: 1 ug/dl_
Multivariate linear regression
models examined the
relationship of blood Pb with
change in z-score for
outcome (post- and pre-
stress). Potential
confounders considered:
HOME score, SES, birth
weight, child BMI, child sex.
Per 1 ug/dL increase in childhood blood
Pb level:
TPR: 0.088 (95% Cl: 0.023, 0.153)
dyne-s/cm5
SV:-0.069 (95% Cl: -0.124, -0.015) ml_
CO:-0.056 (95% Cl:-0.113, 0.001)
L/min
DBP: 0.069 (95% Cl: -0.001, 0.138)
mmHg
Per 1 ug/dL increase in cord blood Pb
level:
SBP: 12.16 (95% Cl: 2.44, 21.88)
mmHg
Blood Pb levels were not associated
with baseline cardiovascular levels.
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Table 4-22 (Continued): Studies of child cardiovascular endpoints and Pb biomarkers.
Study
(Ordered as they
appear in the text)
Study
Population/
Methodology
Parameters
Blood Pb Data3
Statistical Analysis
Effect Estimates/Results
Gump et al.
(2007)
Prospective
122 children age
9.5 yr in
Oswego, NY
SBP, TPR Childhood (mean age of
measurement: 2.6 yr)
blood Pb:
GM (GSD):
4.06 ug/dl_ (1.14)
Linear regression models
adjusting for the same
covariates as in Gump et al.
(2005). Separate models
testing whether Pb is a
mediator of SES
associations (Sobel test) and
whether Pb moderates SES
associations (Pb-SES
interaction).
Blood Pb was a mediator of the SES-
TPR relationship
SES alone: -0.62 dyne-s/cm5 (p <0.05)
SES with Blood Pb: -0.40 dyne-s/cm5
(p >0.10), change in R2 attributable to
SES: -55.3%
Blood Pb was a potential moderator of
the SES-TPR relationship.
Blood Pb x SES interaction: p = 0.07
Blood Pb was a moderator of SES-SBP
relationship
Pb x SES interaction: p = 0.007
At blood Pb levels >4 ug/dL, SES not
significantly associated with SBP
Gump et al.
(2009)
Prospective
122 children age
9.5 yr in
Oswego, NY
Salivary cortisol Cord blood Pb:
GM (GSD):
2.56 ug/dl_ (1.16)
Childhood (mean age of
measurement: 2.6 yr)
blood Pb: GM (GSD):
4.06 ug/dl_ (1.14)
Linear regression to examine
whether blood Pb level
mediates or moderates the
relationship between SES
and salivary cortisol as in
Gump et al. (2007)
Blood Pb was a mediator of the SES-
cortisol association. SES was no longer
significantly associated with cortisol
after adjusting for blood Pb level. R2 for
SES decreased by 40, 33, and 50% for
cortisol measured at 21, 40, and 60 min.
Blood Pb was not a significant
moderator of SES-cortisol association.
Blood Pb x SES interaction term was
not statistically significant
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Table 4-22 (Continued): Studies of child cardiovascular endpoints and Pb biomarkers.
Study Study
(Ordered as they Population/
appear in the text) Methodology
Parameters
Blood Pb Data3
Statistical Analysis
Effect Estimates/Results
Gump et al.
(2011)
Cross-sectional
140 children
ages 9-11 yr
Oswego, NY
SBP, TPR, HRV
(heart rate
variability) in
response to
acute stress
(mirror tracing
task)
Concurrent blood Pb:
GM: 1.01 ug/dl_
Quartiles:
Q1: 0.14-0.68 ug/dl_
Q2: 0.69-0.93 ug/dl_
Q3: 0.94-1.20 ug/dl_
Q4: 1.21-3.76 ug/dl_
MDL: 0.28-0.34 ug/dl_
Outcomes were analyzed as
continuous variables for the
pre-stress values or the
change post- and pre-stress.
Regression models were
adjusted for sex, SES, BMI,
and age.
Blood Pb levels associated with
autonomic and cardiovascular
dysregulation in response to stress -
greater vascular resistance, reduced
stroke volume, and cardiac output
Change in SBP (mmHg) across
quartiles: Q1: 5.30, Q2: 7.33, Q3: 7.07,
Q4: 7.23, p for trend = 0.31
Change in TPR (%) across quartiles:
Q1: 2.91, Q2: 8.18, Q3: 9.55, Q4: 9.51,
p for trend = 0.03
Change in Stroke Volume (%) across
quartiles: Q1: 2.23, Q2: 0.91, Q3: -3.47,
Q4: -0.89, p for trend = 0.04
Blood Pb levels were not associated
with baseline cardiovascular levels.
Zhang et al.
(2011 a)
Prospective
457 mother child
pairs in a birth
cohort, born 1994
to 2003 in Mexico
City. Children
were evaluated
2008-2010 at ages
7-1 Syr
SBP
Cord blood Pb:
GM (GSD): 4.67 ug/dl_
(1.18)(N=323)
Concurrent blood Pb:
GM (GSD):
2.56ug/dL(1.16)(N=367)
Maternal post-partum
bone Pb:
Median (IQR):
Tibia Pb: 9.3(3.3,
16.1) ug/g
Patella Pb: 11.6(4.5,
19.9)ug/g
Multiple regression models
and generalized estimating
equations (log linear for cord
blood, linear for concurrent
blood and maternal bone).
The base model considered
maternal education, birth
weight, BMI, sex, and child
concurrent age as
covariates.
Prenatal Pb exposure may be
associated with higher BP in female
offspring.
Among girls, an IQR (13 ug/g) increase
in maternal tibia Pb was associated with
a 2.11 (95% Cl: 0.69, 3.52) mmHg
increase in SBP
IQR (16 ug/g) increase in maternal
patella Pb was associated with a 0.87
(95% Cl: -0.75, 2.49) mmHg increase in
SBP
IQR (4 ug/dL) increase in cord blood Pb
was associated with a 0.75 (95% Cl:
-1.13, 2.63) mmHg increase in SBP
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Table 4-22 (Continued): Studies of child cardiovascular endpoints and Pb biomarkers.
Study
(Ordered as they
appear in the text)
Study
Population/
Methodology
Parameters
Blood Pb Data3
Statistical Analysis
Effect Estimates/Results
Factor-Litvak et
al. (1999: 1996)
Cross-sectional SBP
260 children ages
5.5 yr in
K. Mitrovica and
Pristina,
Yugoslavia
Concurrent blood Pb
range:
4.1 to76.4ug/dl_
Linear regression analysis.
Potential confounders
considered: sex, maternal
education, birth weight,
HOME score, and BMI.
Per 1 ug/dL increase in concurrent
blood Pb level: 0.05 (95% Cl: -0.02,
0.13) mmHg higher SBP
Blood Pb level at birth and cumulative
blood Pb level were not as strongly
associated with SBP at age 5.5 yr.
Gerr et al.
(2002)
Cross-sectional
508 young adults
age 19-29 yr, born
1965-1975, male
and female; half of
the subjects had
grown up around
an active Pb
smelter in Silver
Valley, Idaho
BP
While the concurrent
mean blood Pb level was
3.15 ug/dLforthe highest
bone Pb category
(>10ug/g), early
childhood mean blood Pb
levels in this group were
substantially elevated for
all bone Pb level
categories and were
highest among
participants in the highest
bone Pb level category.
The mean blood Pb level
was 65 ug/dL among
participants with bone Pb
level >10 ug/g. Bone Pb
was measured at the time
of entry into this study.
Multiple linear regression
models always included 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 ug/g
to >10 ug/g) were tested
together.
Group in highest quartile of tibia Pb level
(>10 ug/g) had 4.26 (95% Cl: 1.36,
7.16) mmHg higher SBP and 2.80 (95%
Cl: 0.35, 5.25) mmHg higher DBP
compared to the lowest tibia Pb group
aBlood Pb data are estimates of geometric mean (GM) and geometric standard deviation (GSD) using the arithmetic mean and SD.
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4.4.4.2 Cardiovascular Functioning in Children
The relationship between cardiovascular functioning (TPR, BP, stroke volume, and
cardiac output,) and blood Pb levels was examined prospectively by Gump et al. (2007;
2005) in a cohort derived from the Oswego Children's Study discussed in Section 4.4.4.1
and born at a single New York hospital. Higher early childhood Pb levels (average age
2.6 years) were associated with greater TPR response to acute stress induced by mirror
tracing on a computer at age 9.5 years as shown in Figure 4-23. Testing blood Pb with
linear, quadratic, and cubic terms did not produce significantly different Pb-TPR
associations, and the authors suggested that these effects were concentration-dependent
and notably, were not emergent at a specific exposure threshold. TPR increased with
increasing quartile of blood Pb level. A mediational analysis indicated that Pb was a
significant mediator of the SES-TPR reactivity association; some evidence also suggested
moderation, whereby the inclusion of blood Pb into the model reduced the effect estimate
for SES. Observations that Pb exposure increases TPR in toxicological studies and
mechanistic evidence indicating that Pb-induced changes in SNS activity may mediate
such effects (Section 4.4.2.3) provide some biological plausibility for a role of Pb in
affecting the TPR response to acute stress in this child population. Additionally, higher
early childhood blood Pb level (average age 2.6 years) was associated with a smaller
stroke volume and cardiac output responses to acute stress at age 9.5 years (Gump et al.,
2007). In a further analysis in this cohort, Gump et al. (2009) examined the possibility
that Pb may mediate an association between SES and cortical responses to acute stress.
Elevated cortisol has been associated with hypertension (Whitworth et al.. 2000). Gump
et al. (2009) found that lower family income was associated with greater cortisol levels
following an acute stress task and that blood Pb was a mediator of this association, such
that after controlling for early childhood blood Pb levels, family income was no longer
predictive of cortisol levels.
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°- g
II
30
25
20
15
10
5
0
-5
-10
4 6 8 10
Blood lead level (ng/dL)
12
14
Source: Reprinted with permission of Elsevier (Gump et al., 2005)
Figure 4-23 Children's adjusted total peripheral resistance (dyn-s/cm5)
responses to acute stress tasks, as a function of early childhood
Pb levels.
In a different cohort of 140 children 9 to 11 years of age recruited from local pediatrician
offices and from mailings to homes with children in this age group, Gump et al. (2011)
used a similar acute stress-producing paradigm as in previous studies to examine the
cross-sectional associations of concurrent blood Pb with cardiovascular responses. TPR
significantly increased in a concentration-dependent relationship with blood Pb, with
most of the increase occurring between the first quartile blood Pb (0.14-0.68 (ig/dL) and
the second quartile blood Pb (0.69-0.93 (ig/dL). The first quartile contained samples from
children with blood Pb levels below the method detection limit (MDL) (0.28 - 0.34
(ig/dL). These values were used in the analyses and may be subject to greater
measurement error; however, the error is likely nondifferential. This result is consistent
with those of Gump et al. (2005). Also, these newer findings provided evidence of
associations with concurrent blood Pb levels and with lower blood Pb levels (Gump et al..
2011) than were previously examined by Gump et al. (2005) and in a large group of
children without higher Pb exposures earlier in childhood.
Studies in adults and animals indicate Pb-associated decreases in HRV (Section 4.4.3.4).
In Gump et al. (2011). cardiac autonomic regulation decreased in a
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concentration-dependent manner with increasing concurrent blood Pb quartile, with the
largest change relative to the first quartile (0.14-0.68 (ig/dL) measured in the highest
blood Pb quartile (1.21-3.76 (ig/dL). Also, high frequency HRV decreased more with
acute stress in the highest Pb quartile group (1.21-3.76 (ig/dL). In the earlier cohort, early
childhood (mean age at collection: 2.6 years) blood Pb level was associated with reduced
stroke volume and cardiac output measured at the age of 9.5 years (Gump et al.. 2005).
Blood Pb levels in Gump (2005) were determined in either venous (80%) or capillary
(20%) blood specimens. Capillary blood Pb determinations have greater potential for
contamination during collection resulting in greater measurement error, particularly at
concentrations approaching the MDL (1 (ig/dL). In a recent study, Gump et al. (2011)
also found reduced stroke volume and cardiac output but for concurrent (mean age: 10.2
years) blood Pb level and at lower blood Pb levels.
4.4.4.3 Blood Pressure in Children
Zhang et al. (2011 a) conducted a longitudinal study that examined changes in BP in 323
girls and boys aged 7 to 15 years in a Mexico City cohort and associations with maternal
bone Pb measured one month post-partum (a measure of cumulative exposure that could
expose fetuses to Pb through accelerated mobilization of bone Pb during pregnancy) and
with cord blood Pb at delivery. This was the first study to examine the association of
maternal bone Pb, as a marker of prenatal exposure, with offspring BP. The model
including both girls and boys (without adjustment for concurrent blood Pb) showed no
statistically significant association overall for any Pb biomarker with child BP. A
significant interaction was found between maternal tibia Pb and sex, and in models
stratified by sex, maternal tibia Pb was associated with adjusted systolic and diastolic BP
in females, but not males. Maternal post-partum median tibia Pb was 9.3 (ig/g (IQR: 3.3,
16.1 (ig/g) with no significant differences between mothers of male and female offspring.
Suboptimal growth in utero is associated with accelerated weight gain in offspring during
childhood and greater risk of later hypertension (Barker and Bagby. 2005; te Velde et al..
2004; Barker et al.. 1989). The relationship between birth weight and Pb biomarkers is
discussed in Section 4.8.2. These may represent biologically plausible mechanisms by
which prenatal Pb exposure may result in increased BP later in childhood as was
demonstrated in female offspring.
Gump et al. (2011; 2005) examined the relationship of blood Pb level with BP in their
two cohorts of contemporary children around age 10 years in New York State. Gump et
al. (2005) reported an association of cord blood levels with systolic BP (12.16 mmHg
[95% CI: 2.44, 21.88] increase per 1 (ig/dL increase in cord blood Pb level). Gump et al.
(2005) also reported cardiovascular outcomes associated (p<0.2) with other prenatal
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toxicants (i.e., mercury, DDE, HCB); however, the effects of postnatal and prenatal Pb
were found even after covariate control for all non-Pb toxicants that were related to the
outcome. Gump et al. (2011) found that with acute stress, children in higher quartiles of
concurrent blood Pb level (>0.69 (ig/dL) had larger increases in systolic BP. For
example, children with concurrent blood Pb levels between 1.21 and 3.76 (ig/dL had a
7.23 mmHg change, and children with blood Pb levels between 0.14 and 0.68 (ig/dL had
a 5.30 mmHg change. A linear trend was not observed across quartiles. Blood Pb levels
were not significantly associated with baseline cardiovascular levels. An interaction
between long-term perceived stress and bone Pb levels in association with BP and
hypertension also was reported in a study of adults (Peters et al., 2007) (described in
Section 4.4.2.1). An earlier study (Factor-Litvak et al.. 1999; Factor-Litvak et al.. 1996)
of children with higher blood Pb levels ranging from 4.1 to 76.4 (ig/dL found that a
1 (ig/dL increase in concurrent blood Pb was associated with a 0.05 (95% CI: -0.02, 0.13)
mmHg increase in systolic BP. An additional study (Gerr et al., 2002) reported that
systolic BP for young adults (ages 19-29 years) with bone Pb levels greater than 10 (ig/g
(mean concurrent blood Pb = 65 (ig/dL) was 4.26 mmHg higher (diastolic BP was 2.8
mmHg greater) compared with young adults with bone Pb levels < 1 (ig/g.
The pathogenesis of CVD has been hypothesized to begin in childhood (Kapuku et al..
2006). Early markers observable in youth in association with Pb biomarkers include
increased BP during stress, reduced HRV, increased IMT, and vascular endothelium
dysfunction. Kapuku et al. (2006) state that endothelial dysfunction is the center of the
CVD paradigm. The factors measured in childhood or as a cumulative burden since
childhood are predictors of outcomes in young adults who are still too young to
experience coronary events (Li et al.. 2003). and early-life exposures may induce changes
in arteries that contribute to the development of atherosclerosis (Raitakari et al.. 2003).
Berenson (2002) observed that the effects of multiple risk factors on coronary
atherosclerosis support evaluation of cardiovascular risk in young people. Thus, evidence
relating levels of biomarkers of Pb exposure in children to cardiovascular function in the
groups of studies presented in the preceding text when combined with the evidence for
the potential pathogenesis of CVD starting in childhood that yield effects in adulthood,
provides coherence with evidence in adults supporting the effects of long-term,
cumulative Pb exposures in the development of cardiovascular effects.
Few animal studies have examined the effect of Pb exposure during pregnancy and
lactation on BP in offspring as adults and those that have used high levels of exposure.
Recently, pups of Pb-exposed dams (1,000 ppm through pregnancy and lactation)
exhibited increased blood Pb level (mean blood Pb level: 58.7 (ig/dL) and increased
arterial systolic BP after weaning (Grizzo and Cordellini. 2008) suggesting a role for
childhood-Pb exposure leading to adult disease.
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4.4.4.4 Summary of Child Cardiovascular Studies
The 2006 Pb AQCD (U.S. EPA. 2006b) described three studies on the effects of Pb on
cardiovascular function in children; however, no conclusions were made as to the
strength of the evidence. Studies have reported cardiovascular changes antecedent to
CVD such as TPR responses to acute stress tasks as a function of childhood blood Pb
levels. Also, a study reported associations with acute stress-induced autonomic and
cardiovascular dysregulation responses. Biomarkers of prenatal Pb exposure (maternal
postpartum patella and tibia Pb levels) were related to later higher BP. Other lines of
evidence have linked increased intrauterine growth restriction to later accelerated weight
gain in childhood, and this may indicate greater risk of hypertension later in life. The
results are not uniform with respect to the important lifestages of Pb exposure and can
differ by sex and other factors. Uncertainties in these studies may be related to sample
size, single measures of BP, variation in the age of onset of puberty, and cross-sectional
design. However, some of these uncertainties may result in the attenuation of observed
associations rather than the generation of spurious associations. Overall, recent study
findings indicate that in children (ages 5-15 years) with mostly concurrent mean blood Pb
levels of 1-5 (ig/dL, increasing blood Pb level may be associated with small increases in
BP and changes in the cardiovascular system. It is uncertain to what extent these
associations relate to later-in-life effects, such as increased BP and the development of
CVD.
Factors may limit the ability of studies to detect statistically significant Pb-associated
changes with BP. The relatively young age of the subjects may have limited the ability of
these studies to detect significant BP effects (as opposed to early function effects) if
longer duration Pb exposure is necessary to produce the cardiovascular changes and
considering the lower prevalence and strength of compensatory mechanisms in children.
There is uncertainty in the shape of the concentration-response relationship for
cardiovascular endpoints at lower blood Pb levels since most studies modeled a linear
relationship. A nonlinear concentration-response relationship has been found for Pb with
other outcomes in children, most notably, decrements in cognitive function (see
Section 4.3.2).
Cardiovascular endpoints other than baseline BP may be more sensitive outcomes for
measuring Pb-associated cardiovascular effects in young children. The series of studies
by Gump et al. (2011; 2009; 2007; 2005) evaluating much smaller samples than did the
adult studies, was able to demonstrate statistically significant relationships of blood Pb
levels with cardiovascular responses such as TPR, related to acute stress. These results
suggest that the stress paradigm may be useful to detect associations of blood Pb levels
with effects on the cardiovascular system of children. Selection of the appropriate
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cardiovascular outcome in children is an important factor to consider. Rather than using
indicators of cardiovascular effects, such as BP, evaluation of cardiovascular changes that
are antecedent to increased BP and changes in other CVD-related endpoints that present
at later lifestages may be informative to understanding the time course of cardiovascular
changes potentially associated with early Pb exposure.
Overall this small body of evidence, based on different cohorts, locations, and study
designs, is suggestive of a relationship between biomarkers of Pb exposure and
cardiovascular effects in children. One longitudinal study ties in maternal bone Pb level,
and cord and concurrent blood Pb level for the children. Limitations exist in the studies.
While BP increases are more prevalent in older adults than in children, BP increases have
been related to higher blood Pb level in earlier studies of children and young adults (Gerr
et al.. 2002; Factor-Litvak et al.. 1999; Factor-Litvak et al.. 1996). The recent Gump
studies provide information in populations of children (ages 9-11) with mean blood Pb
levels in the range of 1-5 (ig/dL for BP and potential antecedents for CVD such as
increases in TPR and changes in cardiac autonomic regulation.
4.4.5 Mortality
The 2006 Pb AQCD (U.S. EPA. 2006b) stated that available evidence suggested an effect
of Pb on cardiovascular mortality in the general U.S. population but cautioned that these
findings should be replicated before these estimates for Pb-induced cardiovascular
mortality could be used for quantitative risk assessment purposes (U.S. EPA. 2006b).
Previous results involved NHANES II and III analyses that examined prospectively the
association of adult blood Pb measured at the time of the study with all cause and cause-
specific mortality ascertained 8-16 years later (Schober et al., 2006; Lustberg and
Silbergeld. 2002). As blood Pb levels in adults reflect contributions from both recent Pb
exposure and mobilization of historic Pb from bone, it is unclear to what extent recent,
past, or cumulative Pb exposures contributed to the observed associations. Given the
decline in ambient air Pb concentrations and population blood Pb levels, it is likely that
study subjects had a much higher past Pb exposure compared to exposure during the
study period. Using NHANES II (1976-1980) data, Lustberg and Silbergeld (2002) found
significant increases in all-cause mortality, circulatory mortality, and cancer mortality,
comparing adults with blood Pb levels of 20-29 (ig/dL to those with blood Pb levels less
than 10 (ig/dL (measured 12-16 years before ascertainment of vital status). Using
NHANES III data, Schober et al. (2006) found significant increased all-cause,
cardiovascular, and cancer mortality comparing adults with blood Pb levels
from 5-9 (ig/dL and above 10 (ig/dL to those with blood Pb levels less than 5 (ig/dL
(measured a median of 8.8 years before ascertainment of vital status).
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Recent studies substantially strengthen the evidence base for Pb-associated mortality. A
further analysis of the NHANES III database by a different research group using different
methods addressed uncertainties from earlier analyses by considering a greater number of
potential confounding factors and by characterizing concentration-response relationships.
Additionally, two longitudinal prospective studies in different U.S. cohorts conducted by
different researchers with different methods demonstrate consistency within the evidence
base for blood Pb and add new evidence for mortality associated with bone Pb levels.
Menke et al. (2006) examined all-cause and cause-specific mortality using NHANES III
data. Subjects at least 18 years of age were followed up to 12 years after their blood Pb
was measured, and 1,661 deaths were identified. Those with baseline blood Pb levels
from 3.63 to 10 (ig/dL had significantly higher risks of all-cause (HR: 1.25 [95% CI:
1.04, 1.51]), cardiovascular (HR: 1.55 [95% CI: 1.08, 2.24]), MI (HR: 1.89 [95% CI:
1.04, 3.43]), and stroke (HR: 2.51 [95% CI: 1.20, 2.26]) mortality compared to those with
baseline blood Pb levels less than 1.93 (ig/dL and increased risk of cancer mortality (HR:
1.10 [95% CI: 0.82, 1.47]). Effect estimates adjusted for demographic characteristics
were robust to the additional adjustment for factors such as smoking, alcohol
consumption, diabetes, BMI, hypertension, and level of kidney function. The consistency
of HRs across models with a varying number of control variables indicated little residual
confounding. Hazard ratios were not higher comparing adults with blood Pb levels from
1.94 to 3.62 (ig/dL to those with blood Pb levels <1.93 (ig/dL. However, tests for linear
trend were statistically significant for all mortality outcomes except for cancer mortality.
Menke et al. (2006) evaluated several of the model covariates (e.g., diabetes,
hypertension, and glomerular filtration rate [GFR]) in a subgroup analysis. The
comparisons for these are shown in Figure 4-24. The authors reported that there were no
interactions between blood Pb and other adjusted variables.
The results from Menke et al. (2006) generally were consistent with those from the
previous NHANES III analysis of the association of blood Pb with mortality by Schober
et al. (2006) that included participants greater than 40 years of age (N = 9,686) and
adjusted for covariates including age, sex, ethnicity, and smoking rather than the full suite
of covariates evaluated by Menke et al. (2006). Schober et al. (2006). which was
discussed in the 2006 Pb AQCD (U.S. EPA. 2006b). reported increased HRs comparing
adults with blood Pb levels > 10 (ig/dL to those with blood Pb levels <5 (ig/dL for all-
cause (HR: 1.59 [95% CI: 1.28, 1.98]), CVD (HR: 1.55 [95% CI: 1.16, 2.07]), and cancer
(HR: 1.69 [95% CI: 1.14, 2.52]) mortality. In general, HRs were higher but
nonsignificant, comparing adults with blood Pb levels from 5-9 (ig/dL to those with
blood Pb levels <5 (ig/dL. The median follow-up time between measurement of blood Pb
and death ascertainment was 8.55 years.
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Subgroup
Hazard ratio of all-cause mortality (95% Cl)
Hazard ratio of cardiovascular mortality (95% Cl)
Age (years)
<60
>=60
Race-ethnicity
Non-Hispanic white
Non-Hispanic black
Mexican-American
Sex and menopausai status
Male
Female
Pre-menopausal
Post-menopausal
Residence
Rural
Urban
Smoking
Never
Former
Current
Body mass index (kg/m2)
<25
>-25
Hypertension
No
Yes
Diabetes
No
Yes
Estimated glomerular filtration
rate (ml/min/1,73m2)
<60
>=60
Overall
1 .75 (1 .25
1.31 (1.08
1.32(1.09
1.23(0.99
1 .17 (0.86
1.41 (1.11
1.24(1.00
1 .02 (0.54
1.24(1.00
1.28(1.05
1.42(1.18
1.21 (0.93
1,61 (1.33
1 .34 (0.96
1,51 (1.16
1.28(1.03
1.31 (1.08
1.32(1.09
1.37 (1.19
1.12 (0.73
1 ,44 (1 .01
1,32(1.12
1.34(1.16
2,44)
1.58)
1.60)
1.52)
1 .60)
1.78)
1.54)
1 .95)
1.54)
1.54)
1.72)
1.58)
1.94)
1 .87)
1.96)
1.58)
1.58)
1.60)
1.58)
1.71)
2.06)
1.56)
1.54)
_,
— 1
•-
«™™,
• •
m*
B
, B_
•
1
-1
H
— 1
^
i —
• —
—
g
_^
•
1-
1—
>
0,5
1.49(1.12
1.49(1.12
1,63(1-25
1.49(1.15
1.59(1.31
1.49(1,18
1.53(1.21
3.22)
1.99)
1.99)
2.11)
1.94)
1.92)
1.89)
1.94)
0,5 1
i
_^,
— i
—
— •
-4
"•
1
*-
2 3
Note: Hazard ratios were calculated for a 3.4 ug/dL increase in blood Pb level with log-blood Pb as a continuous variable. This
increase corresponds to the difference between the 80th and 20th percentiles of the blood Pb distribution (4.92 ug/dL versus
1.46 ug/dL, respectively).
Source: Reprinted with permission of Lippincott Williams & Wilkins, Menke et al. (2006)
Figure 4-24 Multivariate adjusted relative hazards of all-cause and
cardiovascular mortality per 3.4 ug/dL increase in blood Pb.
Both Menke et al. (2006) and Schober et al. (2006) presented mortality curves that plot
the HRs against blood Pb level. Figure 4-25 shows the mortality HR curves (not absolute
cases of mortality) for both stroke and MI reported by Menke et al. (2006). Nonlinear
associations were modeled. The curves were fitted using predetermined restricted
quadratic splines with knots at the 10th percentile (1.00 (ig/dL), the 50th percentile
(2.67 (ig/dL), and the 80th percentile (5.98 (ig/dL) blood Pb levels. The authors did not
explain the shape of the blood Pb-mortality curves in detail; however, the knots
corresponded with the inflection points in the curve. In the tails of the blood Pb
distribution, HRs decreased with increasing blood Pb level. However, HRs remained
above 1 over most of the blood Pb distribution (blood Pb level greater than 2 (ig/dL and
between 2 and 3 (ig/dL for stroke and myocardial infarction, respectively), and in the
4-391
-------
most heavily populated portion of the blood Pb distribution, HRs increased with
increasing blood Pb level. Using a referent group of persons with blood Pb level less than
1.94 (ig/dL, the HR for persons with blood Pb level greater than 3.63 (ig/dL was
significant at the 5% level 1.51 (95% CI: 1.07, 2.14), but not significant for persons with
blood Pb level in the range of 1.94 to 3.62 (ig/dL. Hazard ratios peaked for all outcomes
at a blood Pb level of approximately 6 (ig/dL. Lower concentration-response functions at
higher blood Pb levels also have been found for blood Pb-cognitive function relationships
in children (Section 4.3.2).
2.0 -i
in
-------
linear trend for mortality was observed across blood Pb tertiles. The results of the spline
fit of the continuous blood Pb level term to relative hazard of all cardiovascular diseases
reported by Schober et al. (2006) are shown in Figure 4-26. Schober et al. (2006) shows
the upper 95% confidence band (dashed lines) of the relative risk for all cause mortality
spline is greater than 1 for all blood Pb levels greater than 1.5 (ig/dL using the referent
group of persons with blood Pb levels less than 1.5. The HR was fixed at 1.0 for the
referent blood Pb level of 1.5 (ig/dL. Also, the lower 95% confidence band is greater than
1 when the blood Pb level is greater than about 4.5 (ig/dL. Using a referent group of
persons with blood Pb levels less than 5 (ig/dL, they found statistically significant
relative risks of CVD mortality for persons with blood Pb levels in the range of 5 to 9,
and those with blood Pb levels greater than 10. In contrast to the curve presented by
Menke et al. (2006). Schober et al. (2006) found the relative hazard axis and the blood Pb
axis largely to be linear (solid line). Both Menke et al. (2006) and Schober et al. (2006)
agree that persons with blood Pb levels greater than 4.5 (ig/dL are at increased risk for
mortality; however, these studies report different shapes for the concentration-response
curves. Despite differences in the age groups included, follow-up time, categorization of
blood Pb levels, and differences in HR across the blood Pb range, results reported by
Menke et al. (2006) and Schober et al. (2006) find associations between higher blood Pb
and increased CVD mortality (see Figure 4-29).
2.0
1.8
1.6
•5 1.4
(0
| 1.2
g 1.0
^ 0.8
OJ
CC 0.6
0.4
0.2
0.0
1.0 2.0 3.0 4.0 5.0 6.0 7.0 8.0 9.0 10.0
Blood lead (ug/dL)
Source: Schober etal. (2006)
Note: The solid line shows the fitted five-knot spline relationship; the dashed lines are the point-wise upper and lower 95% CIs.
Figure 4-26 Relative risk of all cause mortality for different blood Pb levels
compared with referent level of 1.5 ug/dL (12.5th percentile).
4-393
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In addition to the NHANES analyses described above, studies of older adult, primarily
white, males (Weisskopf et al. 2009) and older adult females (Khalil. 2010; Khalil et al.
2009b) were conducted recently. Weisskopf et al. (2009) used data from the NAS to
determine the associations of blood, tibia, and patella Pb with mortality. The authors
identified 241 deaths over an average observation period of 8.9 years (7,673 person-
years). The strongest associations were observed between mortality and baseline patella
Pb concentration. Baseline tibia Pb levels were more weakly associated with CVD
mortality. Tibia bone Pb level is thought to reflect a longer cumulative exposure period
than is patella bone Pb level because the residence time of Pb in trabecular bone is shorter
than that in cortical bone. IHD contributed most to the relationship between patella Pb
and all CVD deaths with an individual HR of 2.69 (95% CI: 1.42, 5.08). Although there
was high correlation between tibia and patella Pb (Pearson r = 0.77), compared with
cortical bone Pb, trabecular bone Pb may have more influence on circulating blood Pb
level and thus, local organ concentration of Pb because of its shorter residence time in
bone. In contrast to the NHANES analyses, the NAS study found that baseline blood Pb
was not significantly related to cardiovascular mortality. This discrepancy may be related
to differences in sample size and resulting power, modeling strategies (e.g., linear versus
log-linear blood Pb level terms), or age range of the study populations. The duration of
follow-up was similar across studies. In the Weisskopf et al. (2009) study of NAS data,
the youngest subjects at baseline were approximately 50-55 years old, compared to the
youngest in the Menke et al. (2006) and Schober et al. (2006) NHANES studies, who
were 18 and 40 years, respectively. Further, the blood Pb tertile analysis of Weisskopf
et al. (2009) could have been affected if the majority of a hypothesized nonlinear
association was contained largely in the lowest (reference) blood Pb tertile.
Weisskopf et al. (2009) also conducted a concentration-response analysis. A linear trend
was observed for increasing HR across tertiles of both tibia and patella Pb levels. The
linear relationship using tertile patella Pb was confirmed in other models in which
continuous patella Pb and nonlinear penalized spline terms (higher order terms) were not
statistically significant. The number of knots and their placement within the Pb
distribution, which can influence these results, were determined by an iterative best fit
procedure. Concentration-response relationships shown in Figure 4-27 were
approximately linear for patella Pb on the log HR scale for all CVD, but appeared
nonlinear for IHD (p <0.10). The peak HR is shown around 60 (ig/g, beyond which the
HR tends to decrease. It is important to note the wide confidence limits, which increase
uncertainty at the lower and upper bounds of patella Pb levels.
4-394
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I'
CM
All-cause
IIINIIHIU IIIIM'y I I
CM
All cardiovascular
IIIHIIIIII mini (
0 50 100
Patella lead, pg/g
0 50 100
Patella lead, ug/g
CM
CM .
t
Ischemic Heart Disease
in mi until i i
0 50 100
Patella lead, yg/g
Note: The reference logHR = 0 at the mean of patella Pb concentration. The estimates are indicated by the solid line and the 95%
pointwise confidence intervals (Cis) by the dashed lines. The p values for significance of the nonlinear component for all-cause,
cardiovascular, and ischemic heart disease mortality were 0.42, 0.80, and 0.10 respectively. Patella Pb concentrations of all
individual participants are indicated by short vertical lines on the abscissa. Adjusted for age, education, smoking status, and pack-
years of smoking among participants without ischemic heart disease at baseline.
Source: Reprinted with permission of Lippincott Williams & Wilkins, Weisskopf et al. (2009)
Figure 4-27 Associations between patella bone Pb level and the log of hazard
ratio (logHR} for all-cause, cardiovascular, and ischemic heart
disease.
4-395
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The association of adult blood Pb with mortality has also been examined among women
enrolled in the Study of Osteoporotic Fractures (SOF) (Khalil et al.. 2009b). This
prospective cohort (N = 533) enrolled female volunteers (age 65-87 years) from two U.S.
locations, Baltimore, MD and Monongahela Valley, PA and followed women for an
average of 12 years after blood Pb measurement. All-cause mortality is significantly
higher comparing women with blood Pb levels >8 (ig/dL to those with blood Pb levels
<8 (ig/dL (HR: 1.59 [95% CI: 1.02, 2.49]). Hazard ratios for combined cardiovascular
disease mortality (HR: 1.78 [95% CI: 0.92, 3.45]), coronary heart disease mortality (HR:
3.08 [95% CI: 1.23, 7.70]), but not stroke mortality (HR: 1.13 [95% CI: 0.34, 3.81]) were
higher among the women enrolled in this study with blood Pb levels >8 (ig/dL. In
addition, analyses of blood Pb tertiles and quintiles indicated that blood Pb-mortality HRs
were consistently elevated in groups with blood Pb levels >7 (ig/dL (Khalil 2010). The
findings for elevated mortality HRs with the highest blood Pb levels are reinforced by the
results displayed in Figure 4-28. The HR curve for all-cause mortality is relatively flat
over most of population blood Pb distribution (represented by the blue dots) and
increases only in the upper tail of the blood Pb distribution where there are relatively few
subjects (i.e., fewer dots).
Other studies also reported Pb-associated increased in mortality but have limited
implications due to their weaker analytic methods. Two studies reported standardized
mortality ratios (SMRs) to compare observed deaths in a Pb-exposed population versus
expected deaths, calculated from a reference group (Neuberger et al.. 2009; Cocco et al..
2007). Mortality studies that compare populations by calculating SMRs based on an
"exposed group" versus the population within which the exposed group resides have
several drawbacks, including the ecologic nature of the analysis and the absence of Pb
exposure data or biological markers of Pb exposure. Neuberger et al. (2009) carried out a
retrospective mortality study of a Superfund site that was highly contaminated with heavy
metals, principally Zn, Pb, and Cd. Not knowing the metal concentrations in the study
area obscures interpretation of the significantly elevated county-state SMRs and the
insignificant or significantly lowered SMRs in the county comparison.
A retrospective study of causes of death among Pb smelter workers in Sardinia, Italy
followed 933 male production and maintenance workers (Cocco et al.. 2007). SMRs for
cardiovascular disease-related deaths were calculated based on age-specific and calendar-
year specific mortality of the entire region. Significantly reduced mortality was reported
in the worker groups. The authors attributed the results to the healthy worker effect based
on health criteria applied at hiring and the small size of the cohort. The usual caveats
regarding population comparison mortality studies apply.
4-396
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00 -
o -
5 10
LEAD VALUE
15
relative hazard
Median spline
Source: Khalil et al. (2010)
Figure 4-28 Multivariate adjusted relative hazard (left axis) of mortality as a
function of blood Pb levels between 1 ug/dL and 15 ug/dL.
4.4.5.1 Summary of Mortality
The mortality results in this review supported and expanded upon findings from the
2006 Pb AQCD (U.S. EPA. 2006b). which included NHANES mortality studies (Schober
et al.. 2006; Lustberg and Silbergeld. 2002). The recent NHANES mortality study
discussed above (Menke et al.. 2006) had many strengths over the earlier studies,
including control for a wider range of potential confounders, testing for interactions with
Pb, consideration of concentration-response relationships, extensive analysis of model
evaluations, and examination of mortality from specific CVDs. Further, an association
with increased mortality was observed at lower mean population blood Pb levels. The
mean blood Pb level of the NHANES III population was 2.58 (ig/dL. In the recent
analysis, the Pb risk of increased cardiovascular mortality increased with increasing
blood Pb level over the most heavily populated portion of the blood Pb distribution, with
4-397
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maximum blood Pb levels between 6 and 7 (ig/dL. It is important to note that the relative
contributions of recent, past, and cumulative Pb exposure to associations observed with
the baseline blood Pb levels is uncertain. In addition, the first evidence that bone Pb, a
metric of cumulative Pb exposure, is associated with increased mortality was reported
recently among NAS men (Weisskopf et al., 2009).
Quantitative differences in Pb-associated hazard for death between studies may be
influenced by age range of the study groups, follow up time to death, variation in model
adjustment, central tendency and range of the Pb biomarker levels, assumptions of
linearity in relationship with Pb biomarkers, and choice of Pb biomarker. Quantitative
differences in Pb-associated mortality across NHANES II and NHANES III studies or
between different NHANES III analyses may be explained by the use of continuous or
ordered blood Pb terms and different data selection strategies. Further, studies using
ordered categories of blood Pb level may obtain different results, as the range of blood Pb
level represented in the reference category will affect the calculated coefficients of the
remaining percentiles or groups.
Specifically, Menke et al. (2006) is the strongest study presently published for estimating
the effects of Pb on cardiovascular disease-related mortality. The study uses the
nationally representative NHANES III (1988-1994) sample of men and women. The
results corroborate earlier published NHANES studies but address some of the key
weaknesses noted in those studies. For example, Menke et al. (2006) examined potential
confounding by a large number of factors, including hypertension and kidney function.
Weisskopf et al. (2009) is the first published mortality study using bone Pb as an
exposure index. The study is a prospective study with nearly 100% successful follow-up
of deaths. This rigorous study found increased cardiovascular disease mortality in
association with patella bone Pb with weaker associations for tibia Pb level. The Khalil
et al. (2010; 2009b) study of SOF subjects provides supporting results for a cohort
consisting of white females aged 65-87 years. Further, a number of prior studies found
associations between accumulated Pb reflected in bone Pb measurements and higher
CVD morbidity (Sections 4.4.2.1 and 4.4.3). This evidence base is augmented with new
findings indicating that biomarkers of longer-term cumulative Pb exposure increases
CVD mortality. The NAS and SOF examine only men and women, respectively.
However, the consistency of findings between the two studies indicates that the results of
either study may be applicable widely. Despite the differences in design and methods
across studies, associations between higher levels of Pb bi4omarkers and higher risk of
mortality were generally observed (Figure 4-29 and Table 4-23). One exception is that
stroke mortality was not significantly elevated in the SOF study although it was positive.
Mortality from specific CVD causes, MI and IHD mortality, which are related to higher
BP and hypertension, were elevated with higher Pb biomarker levels.
4-398
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Reference
Outcome
Lustberg&Silbergeld{2002) All Cause
n = 4,190
mn age = 54 y CVD
blo
mn blood Pb = 14.0
Schoberetal. (2006)
n = 9,757
age > 40 y
Menkeetal. (2006)
n = 13,946
mn age = 44
mn blood Pb = 2.58
Khaliletal. (2009)
n = 533 women
mn age = 70 y
mn blood Pb =5.3
All Cause
CVD
All Cause
CVD
MI
Stroke
All Cause
CVD
CHD
Stroke
Weisskopfetal. (2009) AllCause
n = 868 men
mn age = 67.3 y CVD
mn blood Pb =5.7
IHD
All Cause
CVD
IHD
All Cause
CVD
IHD
Study
NHANES II
1976-1980
NHANES HI
1988-1994
NHANES III
1988-1994
SOF
1986-1988
NAS
1991-1999
Pb Biomarker
Comparison Groups
Blood Pb (u-g/dL)
20-29 vs. < 10
10-19 vs. < 10
20-29 vs. < 10
10-19 vs. < 10
>10vs. <5
5-9 vs. < 5
>10vs. <5
5-9 vs. < 5
>3.63vs.SL93
1.94-3.62 vs. S1.93
53.63 vs. <1.93
1.94-3.62 vs. <1.93
>3.63vs.<1.93
1.94-3.62 vs. Ł1.93
>3.63vs. S1.93
1.94-3.62 vs. S1.93
>8vs. < 8
>8vs. < 8
>8vs. <8
>8vs. < 8
>6vs. < 4
4-6 vs. < 4
>6vs. < 4
4-6 vs. < 4
>6vs. < 4
4-6 vs. < 4
Tibia Pb (ng/g)
TertileS vs. 1
Tertile 3 vs. 1
TertileS vs. 1
Patella Pb (ug/g)
>35vs. <22
22-35 vs. < 22
>35vs. <22
22-35 vs. < 22
>3Svs. <22
22-35 vs. < 22
34567
Hazard Ratio {95% Cl)
10
Note: Studies are categorized by Pb biomarker. Within each category, studies generally are presented in order of discussion in the
text. Hazard ratios represent the hazard in the higher blood or bone Pb group relative to that in the lowest blood or bone Pb group
(reference group).
Blood Pb (closed markers), or Bone Pb (open markers) associations with All-cause mortality (black diamonds) or
Cardiovascular mortality (blue circles).
Figure 4-29 Hazard ratios for associations of blood Pb or bone Pb with
all-cause mortality and cardiovascular mortality.
4-399
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Table 4-23 Additional characteristics and quantitative data for associations of blood and bone Pb with mortality
for studies presented in Figure 4-29.
Study
Study Population/
Methodology
Parameter
Pb Data
Statistical Analysis
Hazard Ratio or SMR (95% Cl)
Lustberg and
Silbergeld
(2002)
Longitudinal
4,190 adult
participants of
NHANES III
(1976-1980), ages 30
to 74 yr, Studied
through
December 31, 1992
All cause and cause-
specific mortality
Categorical blood Pb
level
Mean: 14.0(5.1)
Median: 13 ug/dL
1st fertile: <10 ug/dl_
(Reference)
2nd fertile: 10-19ug/dL
3rd fertile: 20-29 ug/dl_
Proportional Hazard model,
RRs adjusted for age, sex,
location, education, race,
income, smoking, BMI,
exercise
All-cause (2nd vs. 1st):
1.40(1.16-1.69)
All-cause (3rd vs. 1st):
2.02(1.62-2.52)
Circulatory (2nd vs. 1st):
1.27(0.97, 1.57)
Circulatory (3rd vs. 1st):
1.74(1.25,2.40)
Cancer (2nd vs. 1st):
1.95(1.28,2.98)
Cancer (3rd vs. 1st):
2.89(1.79,4.64)
4-400
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Table 4-23 (Continued): Additional characteristics and quantitative data for associations of blood and bone Pb with mortality
for studies presented in Figure 4-29.
Study
Study Population/
Methodology
Parameter
Pb Data
Statistical Analysis
Hazard Ratio or SMR (95% Cl)
Schober et al.
(2006)
Longitudinal
9,686 adult
participants of
NHANES III, ages
>40yr
All cause and cause-
specific mortality
Ordered categorical
blood Pb level,
measured a median of
8.55 yr prior to death
<5 ug/dL
5-9 ug/dl_
>10ug/dL
Survey-design adjusted Cox
proportional hazard adjusted
for sex, age, race/ethnicity,
smoking, education level. Did
not evaluate BMI nor
cormorbidities
All-cause (2nd vs. 1st):
1.24(1.05, 1.48)
All-cause (3rd vs. 1st):
1.59(1.28, 1.98)
CVD(2nd vs. 1st):
1.20(0.93, 1.55)
CVD(3rdvs. 1st):
1.55(1.16,2.07)
Cancer (2nd vs. 1st):
1.44(1.12, 1.86)
Cancer (3rd vs. 1st):
1.69(1.14,2.52)
Menke et al.
(2006)
Longitudinal
13,946 adult
participants of
NHANES III
(1988-1994), ages
>17yr, Studied
through December 31,
2000
All cause and cause-
specific mortality
CVD:ICD-9 390-434;
ICD-10 IOO-I99
Ml: ICD-9 41 0-414 and
429.2; ICD-10 I20-I25
Stroke: ICD-9 430-434
and 436-438;
ICD-10 I60-I69
Baseline blood Pb
(measured an average
of 12 yr before
mortality):
Mean: 2.58 ug/dL
Tertiles:
<1.93ug/dL,
1. 94-3.62 ug/dL,
> 3.63 ug/dL
Survey-design adjusted Cox
proportional hazard
regression analysis (up to 12
yr follow-up) adjusted for
Model 1: age, race/ethnicity,
sex, Model 2: Model 1 plus
urban residence, cigarette
smoking, alcohol
consumption, education,
physical activity, household
income, menopausal status,
BMI, CRP, total cholesterol,
diabetes mellitus, Model 3:
Model 2 plus hypertension,
GFR category
All-cause (3rd vs. 1st fertile):
1.25(1.04, 1.51)
CVD(3rdvs. 1st):
1.55(1.08,2.24)
Ml (3rd vs. 1st):
1.89(1.04, 3.43)
Stroke (3rd vs. 1st):
2.51 (1.20, 5.26)
Cancer (3rd vs. 1st):
1.10(0.82, 1.47)
4-401
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Table 4-23 (Continued): Additional characteristics and quantitative data for associations of blood and bone Pb with mortality
for studies presented in Figure 4-29.
Study
Study Population/
Methodology
Parameter
Pb Data
Statistical Analysis
Hazard Ratio or SMR (95% Cl)
Khaliletal.
(2009b)
Longitudinal
533 women, ages
65-87 yr, from Study of
Osteoporotic Fractures
cohort in Baltimore,
MD and Monongahela
Valley, PA
All cause and cause-
specific mortality
Blood Pb measured an
average 12 (SD; 3) yr
before death:
Mean (SD, range):
5.3(2.3, 1-21)ug/dL
Cox proportional hazards
regression analysis adjusted
forage, clinic, BMI, education,
smoking, alcohol intake,
estrogen use, hypertension,
total hip bone mineral density,
walking for exercise, and
diabetes
>8 ug/dLvs. <8 ug/dL
All cause:
1.59(1.02,2.49)
CVD: 1.78(0.92, 3.45)
Coronary Heart Disease:
3.08(1.23, 7.70)
Stroke: 1.13(0.34, 3.81)
Cancer: 1.64(0.73, 3.71)
4-402
-------
Table 4-23 (Continued): Additional characteristics and quantitative data for associations of blood and bone Pb with mortality
for studies presented in Figure 4-29.
Study
Study Population/
Methodology
Parameter
Pb Data
Statistical Analysis
Hazard Ratio or SMR (95% Cl)
Weisskopf et
al. (2009)
Longitudinal
868 elderly men
(mostly white, age >55
yr) enrolled in MAS in
Greater Boston, MA
area
All cause and cause-
specific mortality
Pb biomarkers
collected an average of
8.9 yr before death
Blood Pb:
Mean (SD):
5.6 (3.4) ug/dL
Patella Pb:
Mean (SD):
31.2 (19.4) ug/g
Tertiles:
<22 ug/g,
22-35 ug/g,
>35 ug/g
Tibia Pb:
Mean (SD):
21.8 (13.6) ug/g
Cox proportional hazard
regression analysis adjusted
forage, smoking, education.
Additional models adjusted for
alcohol intake, physical
activity, BMI, total cholesterol,
serum HDL, diabetes mellitus,
race, and hypertension
All-cause (3rd vs. 1st patella Pb
fertile):
1.76(0.95, 3.26)
All CVD (3rd vs. 1st fertile):
2.45(1.07, 5.60)
IHD(3rdvs. 1st):
8.37(1.29, 54.4)
Cancer (3rd vs. 1st):
0.59(0.21, 1.67)
After excluding 154 subjects
with CVD and stroke at
baseline:
All-cause (3rd vs. 1st):
2.52(1.17, 5.41)
All CVD (3rd vs. 1st):
5.63(1.73, 18.3)
All-cause (3rd vs. 1st blood Pb
fertile):
0.93(0.59, 1.45)
All CVD (3rd vs. 1st):
0.99(0.55, 1.78)
IHD(3rdvs. 1st):
1.30(0.54, 3.17)
4-403
-------
Table 4-23 (Continued): Additional characteristics and quantitative data for associations of blood and bone Pb with mortality
for studies presented in Figure 4-29.
Study
aNeuberger et
al. (2009)
Study Population/
Methodology
Ecological
Residents at or near
Parameter
Cause-specific
mortality
Pb Data
No biomarker
measurements
Statistical Analysis
Standardized mortality ratio
(SMR) based on 2000 U.S.
Hazard Ratio or SMR (95%
Heart disease:
Both sexes:
Cl)
Tar Creek Superfund
site, Ottawa County,
OK (exposed pop.
5,852, unexposed pop.
16,210)
Census data
114.1 (113.1, 115.2)
Men:
118(116.4, 119.6)
Women:
111 (109.5, 112.5)
Note: Studies are generally presented in order of discussion in the text.
"These references not included in Figure 4-29 because they reported standardized mortality ratios.
Stroke:
Both sexes:
121.6(119.2, 123.9)
Men:
146.7(107.4, 195.7)
Women:
106.5(80.2, 138.6)
aCocco et al.
(2QQ7)
Ecological
933 male Pb smelter
workers from Sardinia,
Italy (1973-2003)
All cause and cause-
specific mortality
No biomarker SMR
measurements
All cause: 56 (46, 68)
CVD: 37 (25, 55)
4-404
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4.4.6 Air Pb-PM Studies
4.4.6.1 Cardiovascular Morbidity
A relatively small number of studies used Pb measured in PMi0 and PM25 ambient air
samples to represent Pb exposures. However, given that size distribution data for Pb-PM
are fairly limited, it is difficult to assess the representativeness of these concentrations to
population exposure (Section 2.5.3). Moreover, data illustrating the relationships of
Pb-PMio and Pb-PM25 with blood Pb levels are lacking. A few available studies exposed
rats, dogs, or humans to concentrated ambient air particles (CAPs) in which Pb and
several other components were measured. Consistent with epidemiologic studies of blood
and bone Pb and with studies of animals exposed to Pb, these studies show that exposure
to Pb-containing CAPs resulted in various changes related to increased vasoconstriction
(Urch et al.. 2004; Wellenius etal.. 2003; Batalha et al.. 2002). While Pb-containing
CAPs indicate cardiovascular effects with short-term exposure (2-6 hours over
multiple days), they cannot be attributed specifically to the Pb component of the mixture.
It is important to note that Urch et al. (2004) estimated the Pb effect on brachial artery
diameter based on the ambient concentration of Pb, not direct exposure to Pb isolated
from CAPs.
A U.S. time-series study of almost 3 million pregnant women found that increases in
ambient Pb-TSP concentrations were associated with increased odds of pregnancy
induced hypertension (PIH) assessed at delivery (4% increase per 0.05 (ig/m3 increase in
seasonal average Pb at conception and birth) (Chen et al.. 2006c). The authors adjusted
for maternal age, race, education level, marital status, parity, and adequacy of prenatal
care measured at individual levels and stratified by maternal cigarette smoking. In
contrast, epidemiologic studies provide weak evidence for an association between short-
term changes (daily average) in ambient air concentrations of Pb- PM2 5 and
cardiovascular morbidity in adults adjusting for weather and time trends. Some of these
time-series studies analyzed Pb individually, whereas others applied source
apportionment techniques to analyze Pb as part of a group of correlated components. In a
time-series study of 106 U.S. counties, Bell et al. (2009) found that an increase in lag 0
Pb- PM2 5 was associated with an increased risk of cardiovascular hospital admissions
among adults ages 65 years and older. Quantitative results were not presented; however,
the 95% CI was wide and included the null value. In this study, statistically significant
associations were observed for other PM metal components such as nickel, vanadium,
and Zn. In the absence of detailed data on correlations among components or results
adjusted for copollutants, it is difficult to exclude confounding by ambient air exposures
4-405
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to these other components or copollutants. To address correlations among PM chemical
components, some studies applied source apportionment techniques to group components
into common source categories. In these source-factor studies, it is not possible to
attribute the observed association (Sarnat et al.. 2008) or lack of association (Andersen et
al.. 2007) specifically to Pb.
4.4.6.2 Mortality
Time-series epidemiologic studies of ambient air Pb- PM25 reported positive associations
with mortality. Although limited in number, these studies indicated associations in
multiple cities across the U.S. In the Harvard Six Cities Study, Laden et al. (2000) found
a 1.16% (95% CI: 0.20, 2.9%) increased risk in all-cause mortality per 461.4 ng/m3
(5th-95th percentile) increase in Pb-PM2 5. In six California counties, Ostro et al. (2007)
found that a 5 ng/m3 (interquartile range) increase in Pb-PM2 5 was associated with a
1.89% (95% CI: -0.57, 4.40%) increased risk of cardiovascular mortality and a 1.74%
(95% CI: 0.24, 3.26%) increased risk of all-cause mortality during the cool season. The
limitations of air-Pb studies were described in Section 4.4.6.1 above and also are relevant
to the interpretation of these findings for mortality.
4.4.7 Summary and Causal Determination
Large bodies of epidemiologic and toxicological evidence indicate effects of Pb exposure
on a range of related cardiovascular effects. For evaluation of causal relationships with
Pb exposure, evidence was grouped in categories using the U.S. Surgeon General's
Report on Smoking as a guideline (CDC. 2004). The categories include hypertension,
subclinical atherosclerosis, coronary heart disease (CHD), and cerebrovascular disease.
The causal determination for hypertension and increased BP is not only informed by
evidence for hypertension and BP, but also cardiovascular mortality. CHD is informed by
evidence in humans for HRV, MI, IHD, mortality from MI, IHD, and CHD, and in
animals, increased thrombosis, coagulation, and arrhythmia. The biological plausibility
and mode of action for these cardiovascular effects is provided by evidence for oxidative
stress, inflammation, and vascular cell activation or dysfunction. The sections that follow
describe the evaluation of evidence for these four groups of outcomes, hypertension,
subclinical atherosclerosis, coronary heart disease, and cerebrovascular disease, with
respect to causal relationships with Pb exposure using the framework described in
Table II of the Preamble. The key evidence as it relates to the causal framework is
summarized in Table 4-24.
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4.4.7.1 Evidence for Hypertension and Increased Blood Pressure
The 2006 Pb AQCD concluded that there was a relationship between higher blood Pb and
bone Pb and cardiovascular effects in adults, in particular increased BP and increased
incidence of hypertension (U.S. EPA. 2006b). and recent evidence strengthens this
conclusion. This conclusion is informed by the coherence of effects observed between
epidemiologic and toxicological findings and among related endpoints. Prospective
epidemiologic evidence and animal toxicology studies demonstrate the temporal
relationship of the exposure to effect, while meta-analyses provide indications of
consistency and strength. Cross-sectional epidemiologic evidence supports the
consistency of observed results. Consideration of numerous potential confounding factors
in both the prospective and cross-sectional studies limit uncertainty from bias and other
lines of evidence characterizing modes of action provide biological plausibility for the
associations.
Prospective studies clearly support the relationship between biomarkers of Pb exposure
and hypertension incidence and BP changes establishing the directionality of effects.
High-quality studies are replicated by different investigators using different designs and
in large cohorts in different locations (Peters et al., 2007; Glenn et al.. 2006; Cheng et al.,
2001). Bone Pb coupled with high perceived stress was associated with an increased risk
of developing hypertension in an originally nonhypertensive group of adults (Peters et al.,
2007). Cheng et al. (2001) examined subjects from the NAS cohort without hypertension
at baseline measurement and reported a significant increase for hypertension with higher
patella Pb as analyzed by linear models. A recent prospective study in Pb workers found
independent associations of both baseline blood Pb level and subsequent changes in
blood Pb over follow-up with changes in BP over follow-up and associations of bone Pb
level with hypertension (Glenn et al.. 2006). The results indicated that different
mechanisms may mediate short-term Pb-associated increases in BP and long-term
Pb-associated development of hypertension. Consideration for key potential confounding
factors was appropriate including baseline age, alcohol consumption, BMI, and use of BP
lowering medications. Other factors such as smoking and education were evaluated but
did not predict systolic BP. When subjects with hypertension were excluded from the
model, the predicted change was not altered. Thus, the positive, statistically significant
covariate-adjusted results consistently indicated in these studies help rule out chance,
bias, and confounding with reasonable confidence. Figure 4-17 and Figure 4-18 and the
meta-analysis indicate that results for effects of Pb exposure on BP and hypertension are
positive and precise. This provides more confidence in this relationship and reduces the
level of uncertainty.
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The prospective evidence is supported by meta-analyses that underscore the consistency
and reproducibility of Pb-associated increases in BP and hypertension across diverse
populations and different study designs (Navas-Acien et al., 2008; Nawrot et al., 2002).
Nawrot et al. (2002) found that each doubling of concurrent blood Pb level (between 1
and >40 ug/dL) was associated with a 1 mmHg increase in systolic BP and a 0.6 mmHg
increase in diastolic BP. Navas-Acien et al. (2008) found that all included studies showed
a relationship between higher bone Pb levels and higher BP. Also, all but one study that
characterized hypertension showed higher relative risks or odds ratios associated with
higher bone Pb levels.
Further support for a causal relationship between blood and bone Pb levels and increased
BP and hypertension is provided by many cross-sectional analyses conducted by
numerous researchers using different study designs and analyses in large, diverse cohorts
in different locations. Evidence presented in the 1990 Pb Supplement to the Addendum
(1990a) indicated that in individuals participating in NHANES II aged 40-59 years during
1976-1980, a time period corresponding with peak air Pb concentrations in the U.S., there
was no evident threshold found below which blood Pb level was not significantly related
to BP across a range of concurrent blood Pb level of 7-34 (ig/dL. A recent study in an
ethnically diverse community-based cohort of women and men aged 50-70 years found
hypertension risk to be associated with blood and tibia Pb levels (Martin et al., 2006).
Recent epidemiologic studies in adults found associations with hypertension in
populations with relatively low mean blood Pb levels. For example, a positive
relationship was found in the nationally representative NHANES III (1988-1994), in
which the population geometric mean concurrent blood Pb level was 1.64 (ig/dL
(Muntner et al.. 2005). Despite the extensive evidence for associations at relatively low
concurrent blood Pb levels, these cardiovascular outcomes were most often examined in
adults who have been exposed to higher levels of Pb earlier in life, and uncertainty
remains concerning the Pb exposure level, timing, frequency, and duration contributing to
the observed associations.
Further, recent cross-sectional epidemiologic studies also emphasized the interaction
between Pb biomarker levels and factors, such as genetic variants, race/ethnicity, and
metabolic syndrome, in modifying the association with BP or hypertension. Evidence
was presented for a larger blood Pb-associated increase in BP in carriers of the ALAD2
allele, which is associated with greater binding affinity for Pb in the bloodstream (see
Figure 4-17 for results) (Scinicariello et al.. 2010). Additionally, bone Pb level was
associated with larger increases in PP, which represents a good predictor of
cardiovascular morbidity and mortality and an indicator of arterial stiffness, among NAS
adults with the HFE H63D and/or C282Y variant (Zhang etal. 2010a) (Figure 4-17 for
results). Park et al. (2009a) provided further evidence of HFE and transferrin gene
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variants, related to Fe metabolism, impacting the associations of bone Pb levels with
cardiovascular effects, evaluated by QT interval changes in the NAS cohort.
Combined evidence from prospective and cross-sectional studies helps limit the level of
uncertainty for bias from confounding with reasonable confidence. While the adjustment
for specific factors varied by study, the collective body of evidence adjusted for multiple
potential key confounding factors, including age, diet, sex, BMI, blood pressure lowering
medication use, SES, race/ethnicity, alcohol consumption, cholesterol, smoking, pre-
existing disease (i.e., diabetes), measures of renal function, and copollutant exposures
(i.e., Cd).
Associations between biomarkers of Pb exposure and increased BP and hypertension
have been observed in a number of populations, including the large nationally
representative NHANES cohort (Menke et al., 2006; Muntner et al., 2005). In addition,
associations are found in other cohorts that include both men and women (Martin et al..
2006). Further, the meta-analyses assess cohorts both within the U.S. and international,
further supporting the generalizability of the relationship between Pb exposure and
increased BP and hypertension.
Cardiovascular effects of Pb exposure in children are discussed in Section 4.4.4. Overall
this body of evidence, based on different cohorts, locations, and study designs is
suggestive of a relationship between biomarkers of Pb exposure and cardiovascular
effects in children. Recent studies provide information for BP and cardiovascular changes
that may be antecedent to later-in-life effects, such as increases in TPR and changes in
cardiac autonomic regulation.
A causal relationship is further supported by coherence between epidemiologic and
toxicological evidence for the effects of long-term Pb exposure on BP. Collectively, all
animal toxicological studies providing blood Pb level and BP measurements reported
increases in BP with increasing blood Pb level in the range relevant to this ISA (Figure
4-20). Whereas the majority of studies examined long-term Pb exposures that resulted in
mean blood Pb levels >10 (ig/dL, one animal toxicological study found a continuous
monotonic increase in BP in animals with a mean blood Pb level from 0.05 to 29 (ig/dL
with no evidence of a threshold (Tsao et al., 2000). Thus, most evidence demonstrated
such effects in adult animals with blood Pb levels >10 ng/dL. Also, recent studies
demonstrated only partial reversibility of Pb-induced increased BP following Pb exposure
cessation or chelation and the possibility for short-term Pb exposure-induced increases in
BP. The short-term effects were found with routes of Pb exposure that may have
uncertain relevance to humans.
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Coherence for BP and hypertension evidence was also provided by epidemiologic
evidence indicating associations with related CV conditions. Studies in the medical
literature show that increasing BP, even within the nonhypertensive range, is associated
with increased rates of death and cardiovascular disease, including CHD, stroke, and
cardiac failure (Ingelsson et al., 2008; Chobanian et al., 2003; Pastor-Barriuso et al.
2003; Prospective Studies Collaboration. 2002; Kannel. 2000a. b; Neatonetal.. 1995).
Evidence for Pb-induced hypertension and increased BP is supported by consistently
observed associations between Pb biomarkers and both cardiovascular and all-cause
mortality in prospective studies with follow-up periods ranging between 8 and 12 years.
A recent analysis of the NHANES III sample reported associations of adult blood Pb
level with cardiovascular mortality (Menke et al. 2006). These findings were supported
by a community-based cohort study of women ages 65-87 years, in which higher effect
estimates were observed for mortality from cardiovascular disease than all-cause
mortality (Khalil et al. 2009b). Weisskopf et al. (2009) published the first mortality study
using bone Pb as an exposure index. This prospective study found that patella bone Pb
levels were associated with increased mortality from cardiovascular disease.
Animal toxicology studies further indicate coherence and strengthen the evidence for
causality by providing strong biological plausibility for Pb-associated increases in BP and
hypertension. Hypertension results from dysfunction in the regulation of blood flow and
vascular resistance. Many systems, including the central and sympathetic nervous
systems, the contractile processes in the vasculature, and various hormonal regulators,
contribute to the maintenance of BP, and disruption of these systems will alter BP
homeostasis. Studies demonstrate that oxidative stress produced following Pb exposure
inactivates the vasodilator NO which may lead to increased vasoconstriction and
increased BP, leading to hypertension. In addition, oxidative stress can damage the
endothelium, further disrupting endothelium-dependent vascular relaxation and
increasing the contractile response. Studies also suggest Pb exposure disrupts normal
contractile processes by altering the sympathetic nervous system, the renin-angiotensin-
aldosterone system, and the balance between production of vasodilators and
vasoconstrictors (Section 4.4.2.3). Changes in BP that have been associated with
biomarkers of Pb exposure indicate a modest change for an individual; however, these
modest changes can have a substantial public health implication at the population level.
The reported effects represent a central tendency of Pb-induced cardiovascular effects
among individuals; some individuals may differ in risk and manifest effects that are
greater in magnitude. For example, a small increase in BP may shift the population
distribution and result in considerable increases in the percentages of individuals with BP
values that are clinically significant, i.e., an indication of hypertension and medication
use.
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Overall, evidence in epidemiologic and toxicological studies demonstrates consistent
effects of long-term Pb exposure on increased BP and hypertension in adults; however,
uncertainty remains concerning the Pb exposure level, timing, frequency, and duration
contributing to the effects. The epidemiologic studies are of high-quality, have been
replicated by different researchers in different cohorts, and have adjusted for numerous
potential confounding factors. Thus, collectively, they help limit the level of uncertainty
for bias from confounding with reasonable confidence. In addition, a biologically
plausible potential mode of action is described in toxicological studies. Thus, the
combined evidence from epidemiologic and toxicological studies is sufficient to conclude
that there is a causal relationship between Pb exposure and hypertension.
4.4.7.2 Evidence for Subclinical Atherosclerosis
Measures of subclinical atherosclerosis provide the opportunity to assess the pathogenesis
of vascular disease at an earlier stage. Studies that examine markers of subclinical
atherosclerosis, such as PAD (i.e., ankle-brachial index) and generalized atherosclerosis
(i.e., IMT), are included in this category. A limited number of studies have evaluated
markers of subclinical atherosclerosis following Pb exposure in adult humans or animals.
One study described in the 2006 Pb AQCD (U.S. EPA. 2006b) indicated that Pb was
associated with PAD in the NHANES population, and co-exposure with Cd did not
confound the association (Navas-Acien et al.. 2004). Recent epidemiologic findings are
limited to cross-sectional analyses, so uncertainty exists as to the specific Pb exposure
level, timing, frequency, and duration that contributed to the observed associations. One
study reported an increasing trend in the odds of PAD across concurrent blood Pb level
groups in adults within the NHANES population (Muntner et al.. 2005). which is
consistent with the results from the previous Navas-Acien et al. (2004) analysis. An
occupational study presented evidence for increased measures of atherosclerosis analyzed
by Doppler ultrasound (i.e., greater IMT and atherosclerotic plaque presentation) in the
Pb-exposed population with a mean blood Pb level around 25 (ig/dL (Poreba et al.. 2011).
Similarly, toxicological studies have provided limited evidence to suggest long-term Pb
exposure may initiate atherosclerotic vessel disease. Pb exposure to human radial and
internal mammary arteries resulted in a concentration-dependent increase in arterial
intimal thickness (Zeller et al.. 2010). Also, exposure to Pb in rats increased aortic medial
thickness (Zhang et al.. 2009a).
Toxicological studies also present evidence to clearly describe a plausible biological
mechanism. Atherosclerosis is considered an inflammatory disease with a clear role for
oxidative stress in the pathogenesis of the disease. There is consistent evidence that Pb
exposure promotes oxidative stress and increased inflammation in animal and cell culture
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models (Section 4.4.2.3). In addition, there is evidence that Pb will stimulate vascular cell
activation and lead to endothelial cell dysfunction. Both events are key to the
development and progression of atherosclerosis. Also, epidemiologic and animal
toxicology studies have related higher blood Pb levels with higher cholesterol; high
cholesterol is one of the principal risk factors for atherosclerosis (Section 4.4.3.3).
In summary, the evidence includes one high-quality epidemiologic study with adjustment
for numerous potential confounders (Muntner et al.. 2005) and biological plausibility for
observations in humans provided by a few toxicological studies. Thus, the limited
combined epidemiologic and toxicological evidence for is suggestive of a causal
relationship between Pb exposure and subclinical atherosclerosis.
4.4.7.3 Evidence for Coronary Heart Disease
Coronary heart disease (CHD) results from interruption of the blood supply to a part of
the heart caused by atherosclerosis of the coronary arteries, with acute injury and scarring
leading to permanent damage to the heart muscles. A disrupted HRV has been associated
with a higher mortality after MI and is used as a predictor of the physiological processes
underlying CHD (Buccelletti et al.. 2009). Studies that examine incidence of MI, IHD,
HRV, and mortality from CHD, MI, or IHD are included in this category.
There were a small number of studies discussed in the 2006 Pb AQCD (U.S. EPA.
2006b) that indicated associations between Pb biomarker levels and increased risk of
cardiovascular outcomes associated with CHD. However, recent longitudinal studies in
cohorts in different locations with follow-up periods ranging between 8 and 12 years
report that biomarkers of Pb exposure are associated with risk of mortality from
cardiovascular disease, specifically MI, IHD, or CHD. A recent analysis of the NHANES
III sample reported associations of adult blood Pb level with cardiovascular mortality,
with stronger associations observed with MI mortality (Menke et al.. 2006). These
findings were supported by a community-based cohort study of women ages 65-87 years,
in which higher effect estimates were observed for mortality from CHD than all-cause
mortality (Khalil et al.. 2009b). Weisskopf et al. (2009) published the first mortality study
using bone Pb as an exposure index. This prospective study found that patella bone Pb
levels were associated with increased mortality from IHD. Despite the differences in
design and methods across studies, with few exceptions associations between higher
levels of Pb biomarkers and higher risk of CHD-related mortality were consistently
observed (Figure 4-29 and Table 4-23).
The body of evidence demonstrating associations with mortality from CHD is
substantiated by several findings indicating associations between biomarkers of Pb and
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incidence of CHD-related outcomes. A prospective analysis examined the incidence of
IHD (physician confirmed MI, angina pectoris) in the NAS cohort and reported findings
indicating that both blood and bone Pb levels contribute independently to IHD incidence
(Jain et al.. 2007). Earlier studies reported associations of increased Pb biomarkers with
increased risk of left ventricular hypertrophy (Schwartz. 1991). Coherence for the
associations in humans is provided by a recent animal study that suggested that Pb
exposure promotes a procoagulant state that could contribute to thrombus formation
which could reduce the blood supply to the heart (Shin et al.. 2007).
Further support for a relationship between Pb exposure and CHD is provided by evidence
from the NAS cohort for effects on disrupted HRV (Bum et al.. 2011: Park et al.. 2009a:
Park et al.. 2006). which has been associated with a higher mortality from MI and is used
as a predictor of the physiological processes underlying CHD. A prospective analysis
reported that higher tibia Pb, but not blood or patella Pb, was associated with increases in
QTc interval and QRSc duration (Bum et al.. 2011). Park et al. (2006) reported
associations of bone Pb with HRV measures and effect modification by increasing
number of Fe metabolism gene variants from 0 to 3. Park et al. (2006) reported the
strongest relationships between patella Pb levels and decreased HRV among adults with
three or more metabolic abnormalities.
As CHD is the result of vascular blockage, the suggestive evidence for subclinical
atherosclerosis supports the observations of increased CHD morbidity and mortality with
increased Pb exposure. In addition, the strong and consistent evidence for Pb-induced
hypertension serves as further biological plausibility for effects on CHD. Hypertension
may contribute to CHD development in a number of ways. Hypertension may lead to
thickening of the vascular wall or exacerbation of atherosclerotic plaque development
and thus contribute to plaque instability. In addition, hypertension may increase the
myocardial oxygen demand priming for potential myocardial ischemia (Olafiranye et al..
2011). Evidence for both subclinical atherosclerosis and hypertension is supported by
consistent evidence describing the mode of action including Pb-induced oxidative stress,
inflammation, cellular activation and dysfunction, altered vascular reactivity, RAAS
dysfunction, and vasomodulator imbalance.
Recent epidemiologic and toxicological studies substantiated the evidence that long-term
Pb exposure is associated with CHD in adults; however, uncertainty remains concerning
the Pb exposure level, timing, frequency, and duration contributing to the effects.
Overall, high-quality studies examining CHD morbidity and mortality and contributing
cardiovascular effects have been replicated by different researchers in different cohorts
and report consistent associations that increase the confidence that a relationship exists
between Pb exposure and CHD. In addition, both animal and human studies describe a
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biologically plausible potential mode of action. Thus, the overall evidence from primarily
epidemiologic studies and studies providing plausible mode of action information is
sufficient to conclude that there is a causal relationship between Pb exposure and
coronary heart disease.
4.4.7.4 Evidence for Cerebrovascular Disease
Cerebrovascular disease describes a group of conditions involving the cerebral blood
vessels that result in transient or permanent disruption of blood flow to the brain. These
conditions include stroke, transient ischemic attack, and subarachnoid hemorrhage. Both
hypertension and atherosclerosis are risk factors for Cerebrovascular disease and the
mechanisms for these outcomes also apply to Cerebrovascular disease. Despite strong
evidence for hypertension and CHD and long-term Pb exposure, very few studies have
examined the effects of Pb exposure on Cerebrovascular disease. Lee et al. (2009)
examined 153 patients in Taiwan cross-sectionally while adjusting for key potential
confounders and reported increased stenosis greater than 50% in the intracarotid system
related to urine Pb but not blood Pb level. Two epidemiologic studies prospectively
evaluated mortality from stroke. In the NHANES study, Menke et al. (2006) reported a
positive relationship between blood Pb levels and stroke mortality but with wider
confidence intervals compared to other outcomes examined. Khalil et al. (2009b)
reported a nonsignificant result with imprecise confidence intervals. These few studies
provide insufficient evidence to inform the causal relationship between Cerebrovascular
disease and long-term Pb exposure. Thus, the evidence at this time is inadequate to
determine that a causal relationship exists between Pb exposure and Cerebrovascular
disease.
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Table 4-24 Summary of evidence supporting cardiovascular causal determinations.
Attribute in Causal
Framework3
Key Evidence
References
Pb Exposure or Biomarker
Levels Associated with Effects0
Hypertension - Causal
Consistent associations
from multiple, high quality
epidemiologic studies with
relevant blood Pb levels
Longitudinal evidence for associations with incidence
of hypertension and increase in blood pressure in
adults.
Large body of supportive cross-sectional studies
applying differing designs across multiple cohorts of
adults in different locations.
Peters et al. (2007),
Glenn et al. (2006),
Cheng et al. (2001)
Martin et al. (2006).
Scinicariello et al. (2010).
Park et al. (2009b)
Adult, prospective :
Blood Pb level >20 ug/dL;
Bone Pb level >20 ug/g
Adult, concurrent:
Blood Pb level >2 ug/dL;
Bone Pb level >19 ug/g
Additional epidemiologic
evidence help rule out
chance, bias, and
confounding with reasonable
confidence
Associations found with adjustment for numerous
potential confounding factors including age, sex, diet,
BMI, SES, race/ethnicity, alcohol, cholesterol,
smoking, medication use, pre-existing disease, renal
function, and copollutants exposures.
Studies had population-based recruitment, most with
moderate to high follow-up participation not
conditional on blood or bone Pb level.
Meta-analyses demonstrate consistency of
association
Section 4.4.2.1
Table 4-18. Table 4-19
Navas-Acien et al. (2008), Nawrot
et al. (2002)
Consistent evidence from
well conducted animal
studies at relevant
exposures help rule out
chance, bias, and
confounding with reasonable
confidence
Well controlled studies in adult rodents with relevant
dietary long-term Pb exposure support the consistent
evidence for increases in BP in adults.
Rodents:
Rizzietal.(2009).
Bravo et al.(2007).
Chang etal.(2005).
Tsao et al. (2000)
Section 4.4.2.2
Rat, adult:
Blood Pb level >10 ug/dL
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Table 4-24 (Continued): Summary of evidence supporting cardiovascular causal determinations.
Attribute in Causal
Framework3
Key Evidence
References
Pb Exposure or Biomarker
Levels Associated with Effects0
Consistent associations
from multiple, high- quality
epidemiologic studies with
relevant Pb levels in blood
and/or bone and
cardiovascular mortality
Additional epidemiologic
evidence to rule out chance,
bias, and confounding with
reasonable confidence
Longitudinal studies find consistent associations of
blood and/or bone Pb levels in adults with risk of
cardiovascular mortality applying differing designs
across multiple cohorts in different locations
Associations found with adjustment for numerous
potential confounding factors.
Studies had population-based recruitment not
conditional on blood or bone Pb level.
Khalil et al. (2009b),
Weisskopf et al. (2009).
Menke et al. (2006)
Schober et al. (2006)
Lustberg and Silbergeld (2002)
Section 4.4.5
Adult, prospective:
Blood Pb level >4 ug/dL
Evidence clearly describes
mode of action
Oxidative Stress
Alteration of vascular
reactivity
Renin-angiotensin-
aldosterone system
dysfunction
Vasomodulator imbalance
Consistent evidence of increased oxidative stress
leading to inactivation of NO and downregulation of
sGC in animals with relevant dietary Pb exposures
and cultured vascular cells.
Toxicological evidence for activation of the
sympathetic nervous system, increased reactivity to
catecholamines, and activation of the adrenergic and
dopaminergic receptors in rats, isolated vessels, and
cultured cells.
Mixed evidence for reactivity to other pressor agents
(e.g., 5-HT) in rats.
Toxicological evidence that activation of the RAAS
may be involved in development of Pb-induced
hypertension
Evidence for increased RAAS activity in rats and
decreased BP following RAAS inhibition and Pb
exposure.
Limited available toxicological evidence reporting
vasomodulator imbalance in Pb-exposed rats and
cells.
Section 4.4.2.3
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Table 4-24 (Continued): Summary of evidence supporting cardiovascular causal determinations.
Attribute in Causal
Framework3
Key Evidence
References
Pb Exposure or Biomarker
Levels Associated with Effects0
Subclinical Atherosclerosis - Suggestive
Limited but high-quality
evidence in humans of an
association with subclinical
atherosclerosis and
peripheral artery disease
One NHANES analysis reported associations with
PAD at relevant adult blood Pb levels with adjustment
for numerous potential confounding factors.
Limited evidence for increased IMT or arterial
stiffness in adult human populations.
Occupational studies report increased IMT and
atherosclerotic plaque presentation in highly exposed
adult populations.
Muntner et al. (2005)
Ari et al. (2011)
Poreba et al. (2011: 2011a)
Sections 4.4.3.3 and 4.4.3.5
Adult, concurrent:
Blood Pb level >2.5 ug/dL
Adult, concurrent:
Serum Pb level >0.4 ug/dL
Adult workers:
Blood Pb level >24 ug/dL
Limited evidence in animals
of initiation or progression of
atherosclerosis after Pb
exposure
A few studies reported increased IMT, vascular
morphological changes, and endothelial and SMC
alterations in rats and human tissue.
Zeller et al. (2010).
Zhang et al. (2009a)
Section 4.4.3.3
Rat: 28.4 ug/dL
Human Tissue: 50 uM
Evidence clearly describes
mode of action
Oxidative Stress
Inflammation
Vascular Cell Activation
and Endothelial
Dysfunction
Consistent evidence of increased oxidative stress in
animals with relevant dietary Pb exposures and Section 4.4.2.3
cultured vascular cells.
Toxicological evidence of increased inflammation as
indicated by increased production of TNF-a, IL-6,
IL-8, and PGE2 by macrophages and vascular cells.
Toxicological evidence of VSMC stimulation and Section 4.4.3.1
endothelial dysfunction and damage in culture.
Limited available evidence of impaired flow-mediated
dilatation in Pb exposed workers.
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Table 4-24 (Continued): Summary of evidence supporting cardiovascular causal determinations.
Attribute in Causal
Framework3
Pb Exposure or Biomarker
Key Evidence13 References'3 Levels Associated with Effects0
Coronary Heart Disease - Causal
Consistent associations
from multiple, high quality
epidemiologic studies with
relevant bone and/or blood
Pb levels and mortality from
Ml, IHD, CHD, and
cardiovascular disease
Longitudinal studies find consistent associations of
bone and/or blood Pb levels in adults with risk of
cause-specific cardiovascular mortality applying
differing designs across multiple cohorts in different
locations.
Khalil et al. (2009b),
Weisskopf et al. (2009),
Menke et al. (2006)
Schober et al. (2006)
Lustberg and Silbergeld (2002)
Adult, prospective:
Blood Pb level >4 ug/dL
Limited but supporting
evidence in humans of an
association with IHD, Ml, or
HRV
One prospective study demonstrates an association
of adult blood and bone Pb levels with incidence of
IHD in the MAS cohort
Associations of Pb levels in adults and left ventricular
hypertrophy and Ml
Prospective evidence of association of HRV with tibia
bone Pb level in adults.
Evidence for interaction of markers of metabolic
syndrome and genetic polymorphisms with
Pb-induced HRV.
Jain et al. (2007)
Schwartz (1991)
Eum et al. (2011)
Park et al. (2009a; 2006)
Sections 4.4.3.4 and 4.4.3.6
Adult, prospective:
Blood Pb level >5 ug/dL
Adult, prospective:
Bone Pb level >23 ug/g
Additional epidemiologic
evidence help rule out
chance, bias, and
confounding with reasonable
confidence
Associations found while adjusting for numerous
potential confounding factors.
Studies had population-based recruitment not
conditional on blood or bone Pb level.
Section 4.4.5
Limited but supporting
evidence in animals of
increased thrombosis,
enhanced coagulation, and
arrhythmia
One study reporting increased thrombosis and
enhanced coagulation in rats and cells.
One study reporting increased incidence of
arrhythmia and atrioventricular conduction block in
rats.
Shin et al. (2007)
Reza et al. (2008)
Sections 4.4.3.4 and 4.4.3.6
Rat
Blood Pb level: 26.8 ug/dL
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Table 4-24 (Continued): Summary of evidence supporting cardiovascular causal determinations.
Attribute in Causal
Framework3
Key Evidence
References
Pb Exposure or Biomarker
Levels Associated with Effects0
Evidence clearly describes
mode of action
Oxidative Stress
Inflammation
Atherosclerosis
Hypertension
Consistent evidence of increased oxidative stress in
animals with relevant dietary Pb exposures and
cultured vascular cells.
lexicological evidence of increased inflammation as
indicated by increased production of TNF-a, IL-6,
IL-8, and PGE2 by macrophages and vascular cells.
Suggestive evidence of subclinical atherosclerosis in
humans and animals with relevant Pb exposure
resulting in narrowing of the blood vessels to the
heart.
Consistent evidence of increased BP and
hypertension following Pb exposure in humans and
animals at relevant Pb levels across numerous
studies with control for confounding.
Association of increased blood pressure with
manifestation of CHD has been well documented.
Section 4.4.2.3
Section 4.4.3.1
Sections 4.4.3.3 and 4.4.3.5
Section 4.4.2
Cerebrovascular Disease - Inadequate
Evidence for
cerebrovascular disease in
humans and animals is of
insufficient quality and
quantity
One study reported an association of intracranial
carotid stenosis with urinary Pb level in adults.
Lee et al. (2009)
Section 4.4.3.3
Adult, concurrent:
Blood Pb level >5 ug/dL
Limited evidence for
increased mortality from
stroke
Limited evidence for increased risk of mortality from
stroke across two cohorts in different locations.
Menke et al. (2006).
Khalil et al. (2009b)
Section 4.4.5
Adult, prospective:
Blood Pb level >4 ug/dL
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Table 4-24 (Continued): Summary of evidence supporting cardiovascular causal determinations.
Attribute in Causal
Framework3
Pb Exposure or Biomarker
Key Evidence13 References'3 Levels Associated with Effects0
Evidence for possible mode
of action
Hypertension
Atherosclerosis
Consistent evidence of increased BP and
hypertension following Pb exposure in humans and
animals at relevant Pb levels across numerous
studies with adjustment for numerous potential
confounding factors.
Association of increased blood pressure with
manifestation of CHD has been well documented.
Suggestive evidence of subclinical atherosclerosis in
humans and animals with relevant Pb exposure
resulting in narrowing of the blood vessels to the
heart.
Section 4.4.2
Sections 4.4.3.3 and 4.4.3.5
Described in detail in Table II of the Preamble.
""Describes the key evidence and references contributing most heavily to causal determination and where applicable to uncertainties and inconsistencies. References to earlier
sections indicate where full body of evidence is described.
°Describes the blood or bone Pb levels in humans with which the evidence is substantiated and blood Pb levels in animals most relevant to this ISA.
dBecause blood Pb level in nonoccupationally-exposed adults reflects both recent and past Pb exposures, the magnitude, timing, frequency, and duration of Pb exposure contributing
to the observed associations is uncertain.
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4.5 Renal Effects
4.5.1 Introduction
This section summarizes key findings with regard to effects of Pb on the kidney in animal
toxicology and epidemiologic studies. Toxicological studies report that long-term Pb
exposure is associated with pathological changes in the renal system such as proximal
tubule (PT) cytomegaly, renal cell apoptosis, mitochondrial dysfunction, aminoaciduria,
increased electrolyte excretion, ATPase dysfunction, oxidant redox imbalance, and
altered NO homeostasis with ensuing elevated BP. Epidemiologic studies provide
evidence for altered kidney function and chronic kidney disease (CKD) development.
The cardiovascular and renal systems are intimately linked. Homeostatic control at the
kidney level functions to regulate water and electrolyte balance via filtration,
re-absorption and excretion and is under tight hormonal control. Pb exposure has been
shown to damage the kidneys and its vasculature with ensuing effects on systemic
hypertension and effects on the cardiovascular (Section 4.4) and renal systems. Chronic
increases in vascular pressure can contribute to glomerular and renal vasculature injury,
which can lead to progressive renal dysfunction and kidney failure. In this manner,
Pb-induced hypertension has been regarded as one potential contributor of Pb-induced
renal disease. However, the relationship between BP and renal function is more
complicated. Not only does hypertension contribute to renal dysfunction but damage to
the kidneys can also cause increased BP. Long-term control of arterial pressure is
affected by body fluid homeostasis which is regulated by the kidneys. In examination of
the physiological definition of BP (i.e., mean BP equates to cardiac output multiplied by
total peripheral resistance [TPR]) the role of the kidneys in BP regulation is highlighted.
Cardiac output is driven by left ventricular and circulating blood volume. TPR is driven
by vasomodulation and electrolyte balance. Thus, it is possible to dissect the causes of
hypertension from features of primary kidney disease. Increased extracellular fluid
volume results in increased blood volume which enhances venous return of blood to the
heart and increases cardiac output. Increased cardiac output not only directly increases
BP, but also increases TPR due to a compensatory autoregulation or vessel constriction.
In addition, damage to the renal vasculature will alter the intra-renal vascular resistance
thereby altering kidney function and affecting the balance between renal function and BP.
The interactions between these systems can lead to further exacerbation of vascular and
kidney dysfunction following Pb exposure. As kidney dysfunction can increase BP and
increased BP can lead to further damage to the kidneys, Pb-induced damage to both
systems may result in a cycle of further increased severity of disease.
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In general, associations between bone Pb (particularly in the tibia) and health outcomes in
adults indicate chronic effects of cumulative Pb exposure. In adults without current
occupational Pb exposure, blood Pb level represents both recent and cumulative Pb
exposure. In particular, blood Pb level may represent cumulative exposure in
physiological circumstances of increased bone remodeling or loss (e.g., osteoporosis and
pregnancy) when Pb from bone of adults contributes substantially to blood Pb
concentrations. Blood Pb level in children is also influenced by Pb stored in bone due to
rapid growth-related bone turnover in children relative to adults. Thus, blood Pb in
children is also reflective of cumulative dose. Additional details on the interpretation of
Pb in blood and bone are provided in Section 3.3.5. The toxicokinetics of Pb in blood and
bone are important considerations in making inferences about etiologically-relevant Pb
exposures that contributed to associations observed between blood and bone Pb levels
and health outcomes.
4.5.1.1 Kidney Outcome Measures
The primary function of the kidneys is to filter waste from the body while maintaining
appropriate levels of water and essential chemicals, such as electrolytes, in the body.
Therefore, the gold standard for kidney function assessment involves measurement of the
glomerular filtration rate (GFR) through administration of an exogenous radionuclide or
radiocontrast marker (e.g., 1251-iothalamate, iohexol) followed by timed sequential blood
samples or, more recently, kidney imaging, to assess clearance through the kidneys. This
procedure is invasive and time-consuming. Therefore, serum levels of endogenous
compounds are routinely used to estimate GFR in large epidemiologic studies and clinical
settings. Creatinine is the most commonly measured endogenous compound; blood urea
nitrogen (BUN) has also been examined. Increased serum concentration or decreased
kidney clearance of these markers both indicate decreased kidney function. The main
limitation of endogenous compounds identified to date is that non-kidney factors impact
their serum levels. Specifically, since creatinine is derived from creatine in muscle,
muscle mass and diet affect serum levels resulting in variation in different population
subgroups (e.g., women and children compared to men), that are unrelated to kidney
function. Measured creatinine clearance, involving measurement and comparison of
creatinine in both serum and urine, can address this problem. However, measured
creatinine clearance utilizes timed urine collections, traditionally over a 24-hour period,
and the challenge of complete urine collection over an extended time period makes
compliance difficult.
Therefore equations to estimate kidney filtration that utilize serum creatinine but also
incorporate age, sex, race, and, in some, weight (in an attempt to adjust for differences in
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muscle mass), have been developed. Although these are imperfect surrogates for muscle
mass, such equations are currently the preferred outcome assessment method.
Traditionally, the Cockcroft-Gault equation (Cockcroft and Gault 1976), which estimates
creatinine clearance, a GFR surrogate, has been used. In the last decade, the abbreviated
Modification of Diet in Kidney Disease (MDRD) Study equation (Levey et al.. 2000;
Levey et al.. 1999). which estimates GFR, has become the standard in the kidney
epidemiologic and clinical communities. With widespread use of the MDRD equation, it
became clear that the equation underestimates GFR at levels in the normal range.
Therefore, the CKD-Epidemiology Collaboration (CKD-EPI) equation was recently
developed to be more accurate in this range (Levey et al.. 2009). This is a decided
advantage in nephrotoxicant research since most participants in occupational and many
even in general population studies have GFRs in a range that is underestimated by the
MDRD equation.
Both the MDRD and CKD-EPI equations use serum creatinine. Due to the inability to
adjust serum creatinine levels for muscle mass, alternative serum biomarkers have been
evaluated such as cystatin C, a cysteine protease inhibitor that is filtered, reabsorbed, and
catabolized in the kidney (Fried. 2009). It is produced and secreted by all nucleated cells
thus avoiding the muscle mass confounding that exists with serum creatinine (Fried.
2009). However, recent research indicates that serum cystatin C varies by age, sex, and
race (Kottgen et al.. 2008). Thus, a cystatin C-based eGFR equation was recently
developed that includes age, sex, and race (Stevens et al.. 2008).
Most of the kidney outcome measures discussed above were developed for use in the
clinical setting, but are insensitive for detection of early kidney damage. This lack of
sensitivity is evidenced by the fact that serum creatinine remains normal after kidney
donation. Therefore, in the last two decades, the utility of early biological effect (EBE)
markers as indicators of preclinical kidney damage has been of interest. These can be
categorized as markers of function (i.e., low molecular weight proteins that should be
reabsorbed in the PT such as |32-microglobulin and retinol-binding protein [RBP]);
biochemical alteration (i.e., urinary eicosanoids such as prostaglandin E2, prostaglandin
F2 alpha, 6-keto-prostaglandin Fj alpha, 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 kidney dysfunction. However, most of these
markers are research tools only, and their prognostic value remains uncertain since
prospective studies of most of these markers in nephrotoxicant-exposed populations are
quite limited to date. Recently, microalbuminuria has been identified as a PT marker, not
just glomerular as previously thought (Comper and Russo. 2009). Kidney EBE markers
are a major recent focus for research in patients with acute kidney injury (AKI) and
markers such as neutrophil gelatinase-associated lipocalin (NGAL) and kidney injury
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molecule-1 (Kim-1), developed in AKI research, may prove useful for chronic
nephrotoxicant work as well (Ferguson et al.. 2008; Devaraian. 2007).
4.5.2 Nephrotoxicity and Renal Pathology
A number of advances in the impact of Pb on the kidney in the 20 years following the
1986 Pb AQCD (U.S. EPA. 1986a) were noted in the 2006 Pb AQCD (U.S. EPA.
2006b). These advances included evidence that Pb is associated with kidney dysfunction
in both the general population and among CKD patients with mean blood Pb levels from
5-10 (ig/dL. Furthermore Pb exposure, at much lower doses than those causing chronic
Pb nephropathy, may act as a cofactor with other more established kidney risks to
increase the risk of CKD and disease progression in susceptible patients. Marie and Hall
(2011) note that data from basic and clinical studies suggest that obesity, hypertension,
hyperglycemia, hyperlipedemia, and other elements of the metabolic syndrome are highly
interrelated and contribute to the development and progression of diabetic nephropathy
and thus represent populations potentially at increased risk for kidney dysfunction.
Several key uncertainties in the relationship between Pb exposure and kidney function
were noted in the 2006 Pb AQCD (U.S. EPA. 2006b). First, because the effects on the
kidney are most often observed in adults with likely higher past Pb exposures, uncertainty
exists as to the Pb exposure level, timing, frequency, and duration contributing to the
associations observed with blood or bone Pb levels. In addition, there are uncertainties
relating to the impacts of Pb on the kidney in children, the use of paradoxical Pb-kidney
associations on risk assessment in the occupational setting, and the impact of co-exposure
to other environmental nephrotoxicants, such as Cd.
Another important uncertainty relates to the potential for reverse causality, i.e. the
potential for reduced excretion of Pb associated with diminished kidney function rather
than a causal association, to explain relationships observed between blood Pb level and
kidney function. Reverse causality is most important for the interpretation of
cross-sectional studies where the temporal relationship between Pb exposure and kidney
function is not well-established. Although reverse causality has been typically considered
plausible in the interpretation of studies of CKD patients whose kidneys are known to be
compromised, the reverse causality hypothesis can be broadened to include a physiologic
process in which Pb levels are influenced over the entire range of kidney filtration rates.
This expanded hypothesis implies that low blood and bone Pb levels may reflect kidney
function in addition to exposure. If so, this would increase misclassification bias, with Pb
biomarkers reflecting both exposure and kidney function. Various qualitative reviews and
editorials discuss the potential for reverse causality and reach different conclusions
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(Bellinger. 2011; Evans and Blinder. 2011; Kosnett et al., 2007; Marsden. 2003). This
uncertainty will be discussed in greater detail in Section 4.5.2.5.
4.5.2.1 Epidemiology in Adults
General Population Studies
The 2006 Pb AQCD (U.S. EPA. 2006b) included studies that examined associations
between indicators of Pb exposure and kidney function in the general population. The
body of evidence available in adults is summarized in Figure 4-30 and Table 4-25 and
discussed next. Although positive statistically significant associations were not observed
in each of the studies, collectively the findings represent a pattern of elevated risks for
reduced kidney function following long-term Pb exposure.
Two longitudinal analyses of the normative aging study (NAS) population report
associations between blood Pb and increased serum creatinine (Tsaih et al.. 2004; Kim et
al.. 1996). These analyses reduce uncertainties related to reverse causality because they
establish the temporal relationship between Pb exposure and kidney dysfunction.
Participants in NAS, who were originally recruited in the 1960s in the Greater Boston
area were male, age 21 to 80 years, and without chronic medical conditions. The first of
these studies included 459 men whose blood Pb levels from periodic examinations,
conducted every 3 to 5 years during 1979-1994, were estimated based on measurements
in stored packed red blood cell samples adjusted for hematocrit level (Kim et al., 1996).
Participants were randomly selected to be representative of the entire NAS population in
terms of age and follow-up. Kidney function was assessed with serum creatinine. Data
from four evaluations were available for the majority of participants. Mean (SD) age,
blood Pb level, and serum creatinine, at baseline, were 56.9 (8.3) years, 9.9 (6.1) ug/dL,
and 1.22 (0.2) mg/dL, respectively. In the longitudinal analysis with random-effects
modeling of repeated measures, logarithmic (In)-transformed blood Pb was statistically
significantly associated with an increase in serum creatinine from the previous to current
follow-up period in the 428 participants whose highest blood Pb level was < 25 ug/dL
((3= 0.027 mg/dL [95% CI: 0.0, 0.054] per unit increase in logarithmic (In)-transformed
blood Pb); effect estimates in the entire group and subsets with different peak blood Pb
levels (< 10 or 40 ug/dL) also were positive (and larger for blood Pb levels < 10 ug/dL).
This study made key contributions to the understanding of potential uncertainties. In
order to address the question of whether nephrotoxicity observed at current blood Pb
levels is due to higher blood Pb levels from past exposure, these authors performed a
sensitivity analysis in participants whose peak blood Pb levels, dating back to 1979, were
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< 10 ug/dL. A statistically significant positive association between blood Pb and
concurrent serum creatinine remained in a cross-sectional analysis. These authors
evaluated reverse causality, which attributes increased blood Pb levels to lack of kidney
excretion rather than as a causative factor for CKD, by showing in adjusted plots that the
association between blood Pb and serum creatinine occurred over the entire serum
creatinine range (0.7-2.1 mg/dL), including the normal range where reverse causality is
less likely.
In the subsequent NAS publication (Tsaih et al.. 2004). cortical and trabecular bone Pb
measurements were obtained in addition to whole blood Pb in evaluations performed in
the NAS between 1991 and 1995. Associations between baseline blood, tibia, and patella
Pb and change in serum creatinine over an average of 6 years in 448 men were reported.
Mean (SD) age, blood Pb level, and serum creatinine at baseline were 66.0 (6.6) years,
6.5 (4.2) (ig/dL, and 2.25 (0.2) mg/dL, respectfully. Mean blood Pb levels and serum
creatinine decreased significantly over the follow-up period in the whole study group. At
baseline, eligible participants were similar to nonparticipants with regard to age, BMI,
alcohol consumption, smoking status, diabetic status, hypertensive status, baseline serum
creatinine, and blood and bone Pb levels, indicating lack of selective follow-up by
blood/bone Pb level or kidney function. Notably, both patella and tibia Pb, which are
better markers of long-term exposure than blood Pb in populations that are not currently
exposed, but not baseline blood (cross-sectional analyses) Pb level, were positively but
nonsignificantly associated with serum creatinine. All three biomarkers of Pb exposure
from the entire study group (N =448) were positively associated with longitudinal
increases in serum creatinine: blood Pb (ig/dL ((3 = 0.009 mg/dL [95% CI: -0.0008,
0.0188]); tibia Pb jig/g ((3 = 0.007 mg/dL [95% CI: -0.0028, 0.0168]); and patella Pb jig/g
(P = 0.001 mg/dL [95% CI: -0.00684, 0.00884]).
At baseline 6 and 26% of subjects had diabetes and hypertension, respectively. Diabetes
was observed to be an effect modifier of the relationship between blood and tibia Pb with
10-year change in serum creatinine. Per unit increase in logarithmic (In)-transformed
blood Pb, the increase in serum creatinine between follow-up periods was substantially
stronger in diabetics ((3 = 0.076 mg/dL [95% CI: 0.031, 0.121]) compared to
non-diabetics ((3 = 0.006 mg/dL [95% CI: -0.004, 0.016]). 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.
A sensitivity analysis was conducted to characterize the potential for reverse causality by
examining participants whose serum creatinine was < 1.5 mg/dL. The authors reported
that longitudinal associations did not materially change, and stated that "These
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observations in combination with the prospective study design support the conclusion
that the direction of the association is lead dose resulting in renal dysfunction". However,
the potential for reverse causality still cannot be ruled out.
In modeling the association between blood Pb level and change in serum creatinine,
Tsaih et al. (2004) adjusted for baseline serum creatinine. Glymour et al. (2005) discusses
how such adjustment may introduce bias. If there is no interaction between Pb exposure
and unmeasured causes of kidney disease, the model is linear, and the slope does not
change direction prior to and during the study period, the bias should be to underestimate
the effect. However, Glymour (2012). noted that the direction of the bias is difficult to
predict when the model is nonlinear or the data are restricted to a specific stratum of
outcome.
In addition to these analyses of Pb-related increases in serum creatinine among the NAS
population, several cross-sectional studies evaluated GFR and creatinine clearance in
other populations. The Cadmibel Study was the first large environmental study of this
type that adjusted for multiple kidney risk factors, including urinary Cd (Staessen et al..
1992). It included 965 men and 1,016 women recruited from Cd exposed and control
areas in Belgium. Mean concurrent blood Pb was 11.4 ug/dL (range 2.3-72.5), and
7.5 ug/dL (range 1.7-60.3) in men and women, respectively. After adjustment for
covariates (Table 4-25). log transformed blood Pb was 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.
Blood Pb was also negatively associated with estimated creatinine clearance calculated
from the serum creatinine level.
In addition, a cross-sectional study of Swedish women reported that higher concurrent
blood Pb (median: 2.2 (ig/dL) levels were associated with lower eGFR based on serum
cystatin C alone (without age, sex, and race) after adjustment for socio-demographic and
CKD risk factors (Akesson et al.. 2005). Although GFR was estimated in this study,
associations observed in this study were comparable to those using creatinine clearance
as the kidney outcome for Pb. In addition, Payton et al. (1994) found a reduction in log-
transformed creatinine clearance (mL/min) with increasing blood Pb level among the
men enrolled in the NAS.
No significant associations between blood Pb level and kidney function were observed in
a population of adults (N = 168) and children (described in Section 4.5.2.2) living around
smelters in France; however, beta coefficients were not reported (De Burbure et al..
2003). Blood Pb levels were significantly increased in women (mean [SD] control:
4.2 [0.2] (ig/dL; polluted: 5.3 [0.18] (ig/dL) but not men (mean [SD] control: 7.1
[0.18] (ig/dL; polluted: 6.8 [0.17] (ig/dL) living in the polluted area.
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A longitudinal study of GFR in CKD patients that reported a reduction in GFR with
increasing Pb level (Yu et al.. 2004) is discussed below in the section Patient Population
Studies. Studies evaluating the associations between Pb levels and kidney function in
children in the general population are described in detail in Section 4.5.2.2.
The effect of Pb on the kidney has been examined in multiple NHANES datasets
obtained over the last few decades (Figure 4-31 and Table 4-25). NHANES data analyses
benefit from a number of strengths including large sample size, ability to adjust for
numerous potential confounding factors, and the fact that the study population is
representative of the U.S. non-institutionalized, civilian population. The results, covering
different time frames, have been consistent in providing support for Pb as a CKD risk
factor, including NHANES III, conducted from 1988-1994, in which adults with
hypertension and diabetes were observed to be potentially at-risk populations (Muntner et
al.. 2003) and NHANES 1999-2002 (Muntner et al.. 2005). However, because the various
NHANES analyses were cross-sectional in design, examining associations between
concurrent measures of kidney function and blood Pb levels, a common limitation is the
uncertainty regarding the temporal sequence between Pb exposure and renal function and
the magnitude, timing, frequency, and duration of Pb exposure that contributed to the
observed associations.
A recent publication examined NHANES data collected from 1999 through 2006 (Navas-
Acien et al., 2009). The geometric mean concurrent blood Pb level was 1.58 ug/dL in
14,778 adults aged > 20 years. After adjustment for survey year, sociodemographic
factors, CKD risk factors, and blood Cd, the odds ratios for albuminuria (> 30 mg/g
creatinine), reduced eGFR (< 60 mL/min/1.73 m2), and both albuminuria and reduced
eGFRwere 1.19 (95% CI: 0.96, 1.47), 1.56 (95% CI: 1.17, 2.08), and 2.39 (95% CI:
1.31, 4.37), respectively, comparing the highest (>2.4 ug/dL) to the lowest (< 1.1 ug/dL)
blood Pb quartiles. Thus, in the subset of the population with the most severe kidney
disease (both reduced eGFR and albuminuria), the magnitude of association with
concurrent blood Pb was greater. When blood Cd was included as a covariate, blood Pb
remained significantly associated with renal function. In fact, the most important
contribution of this recent NHANES analysis was the evaluation of joint Pb and Cd
exposure (discussed in Section 4.5.4.1).
An important contribution of all NHANES publications is that they provide evidence that
blood Pb remains associated with reduced kidney function (< 60 mL/min/1.73 m2 as
estimated with the MDRD equation cross-sectionally) despite steadily declining blood Pb
levels in the U.S. population during the time periods covered. Other studies of adults
participating in NHANES have also reported worse kidney function related to blood Pb
levels (Lai et al., 2008a; Hernandez-Serrato et al., 2006; Goswami et al., 2005).
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Study
Population
Blood Pb Median Blood Pb 10th-90th
(IQR)(ug/dL) Percentile (ug/dL)
Outcome
POSITIVE EFFECT ESTIMATESINDICATE POORER FUNCTION
LONGITUDINAL RESULTS
Kim et al. (1996) NAS men, blood Pb < 40 ug/dL
NAS men, blood Pb < 25 ug/dL
NAS men, blood Pb < 10 ug/dL
Tsiah et al. (2004) NAS men with diabetes
NAS men without diabetes
NAS men with hypertension
NAS men without hypertension
Entire Cohort
CROSS-SECTIONAL RESULTS
Kim et al. (1996) NAS men, blood Pb < 40 ug/dL
NAS men, blood Pb < 25 ug/dL
NAS men, blood Pb < 10 ug/dL
8.4(5.7,12.4) 4.0-17.5
3.9(2.8,5.6)
2.1-7.6
Tsaih et al. (2004)
NAS men with diabetes
NAS men without diabetes
NAS men with hypertension
NAS men without hypertension
8.4(5.7,12.4) 4.0-17.5
Baseline Blood Pb
5.5(3.7,8.1) 2.6-11.5
Change in Scr between visits
(mg/dL) x 10
Change in Scr peryear
(mg/dL) x 10
Scr (mg/dL)
Scr (mg/dL)
Repeated Measures Blood Pb
NAS men without diabetes
NAS men with hypertension
NAS men without hypertension
NEGATIVE EFFECT ESTIMATES INDICATE POORER FUNCTION
LONGITUDINAL RESULTS
Yu etal. (2004) CKD Patients
CROSS-SECTIONAL RESULTS
Akesson etal. (2005)
Staessen etal. (1992)
Payton etal. (1996)
Akesson etal. (2005)
Swedish Women
Belgian Women
NAS Men
Swedish Women
3.2(2.5,4.1) 2.0-5.1
2.2(1.7,3.0) 1.3-3.8
7.5(5.2,10.9) 3.7-15.1
7.3(5.4,9.9) 4.1-12.9
2.2(1.7,3.0) 1.3-3.8
Change in GFR over 4 yr/100
(mL/min)
Creatinine Clearance/100
(mL/min)
Creatinine Clearance/100
(mL/min)
LogCreatinine Clearance
(mL/min)
GFR/100 (mL/min)
'
-0.1
0.1
0.2
Change inKidneyFunction perl ug/dL Increase in
Blood Pb Level Withinthe 10th-90th Percentile Interval
aBlood Pb data are presented as median and (IQR) in |jg/dL for blood Pb. For uniform presentation, median and IQR were
estimated from the given distributional statistics by assuming normal distributions.
Note: Results are presented first for kidney function tests where an increase is considered impaired function (black circles) then for
tests where a decrease is considered impaired function (blue circles, outlined in box). Within a category, results are presented first
for longitudinal analyses followed by cross-sectional analyses. To compare results for linear and nonlinear modeling, effect
estimates were standardized to a 1 ug/dL increase in blood Pb level within the 10th-90th percentile interval. Magnitudes of the effect
should not be compared among different kidney metrics.
Figure 4-30 Concentration-response relationships for associations between
blood Pb level and kidney function outcomes in adults.
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Study
Quartiles of Blood Pb Distribution Used
Muntner et al. (2003)
Elevated Serum Creatinine
Hypertensive
Normotensive
Navas-Acien et al. (2009)
Albuminuria >30 mg/g creatinine
eGFR <60 mL/min/1.73m2
/I
1
ug/dL Blood Pb
-40 -10 10 50
% Change per ug/dL Blood Pb
Note: These results depicted are from studies that reported ORs of kidney function measures by grouping the population into
quartiles of blood Pb and then comparing each group to the quartile with the lowest blood Pb (reference group). The blood Pb
distribution of the examined group is shaded black and the reference group is shaded gray. To express these odds ratios in terms of
blood Pb concentration, a log normal distribution was fit to the statistics presented and then the medians of each group were
determined. The adjusted OR was the exponentiated quantity (log(OR) divided by the difference in the medians of the groups
compared). The resulting odds ratio is presented in terms of percent change=100*(OR-1).
Figure 4-31 Percent change in kidney outcomes across quartiles of blood Pb
level in NHANES.
Summary of General Population Studies
Collectively, the general population studies summarized in Figure 4-30. Figure 4-31 and
Table 4-25 provide critical evidence on the effects of Pb exposure on the kidney. Two
longitudinal studies provide high quality evidence for an association between reduced
kidney function and blood Pb level with adjustment for numerous potential confounding
factors. As presented in Table 4-25 the slopes and 95% CI limits per 1 (ig/dL increase in
blood Pb level are similar in magnitude for the entire population assessed by Tsaih et al.
(2004) and for the group of individuals with a peak blood Pb level ^ 10 (ig/dL assessed
by Kim et al. (1996). These studies show an overall pattern of elevated risks across all
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subgroups analyzed within the population. However, because blood Pb level in
nonoccupationally-exposed adults reflects both recent and past Pb exposures, the
magnitude, timing, frequency, and duration of Pb exposure contributing to the observed
associations is uncertain.
Although uncertainties remain, the longitudinal analyses of the NAS populations
included sensitivity analyses designed to address the potential for reverse causality to
explain findings. Furthermore, these general population studies included adjustment for
multiple key potential confounding factors including age, race, sex, education, household
income, smoking, alcohol use, and various health indicators such as diabetes, systolic BP,
BMI, and history of cardiovascular disease. Exposure to copollutants such as Cd was also
adjusted (Staessen et al., 1992).
Although the cross-sectional studies do not consistently report positive and significant
findings, they are compatible with evidence from higher quality studies of an effect of
long-term Pb exposure on kidney function. The strengths and limitations of studies
conducted in other populations are discussed in the sections that follow.
4-431
-------
Table 4-25 Additional characteristics and quantitative data for associations of blood and bone Pb with kidney
outcomes for results presented in Figure 4-30 and Figure 4-31.
Reference
Population
Study Location
Time Period
Pb Biomarker
Data Outcome
Statistical Analysis
Effect Estimate (95% Cl)
Results for Figure 4-30: Positive Effect Estimates Indicate Poorer Function
Per 1 ug/dL increase in blood Pb within the
10th-90th percentile interval
Longitudinal Results
Kim et al. (1996)
Adult males
N=459
Boston, MA;
Multiple
examinations
1979-1994
Median baseline
blood
Pb: 8.6 ug/dl_
10th-90th
percentile:
4.0-17.5
Change in
serum
creatinine
between visits
x10(mg/dL)
Random-effects modeling
adjusted for baseline age,
time since initial visit,
BMI, smoking status,
alcohol ingestion,
education level,
hypertension, baseline
serum creatinine, and
time between visits
Peak blood Pb < 40 ug/dl_: 0.012 (-0.0001, 0.025)
Peak blood Pb < 25 ug/dl_: 0.015 (0.0002, 0.03)
Peak blood Pb < 10 ug/dl_: 0.021 (-0.005, 0.048)
4-432
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Table 4-25 (Continued): Additional characteristics and quantitative data for associations of blood and bone Pb with kidney
outcomes for results presented in Figure 4-30 and Figure 4-31.
Reference
Tsaih et al.
(2004)
Population
Study Location
Time Period
Adult males
N~448
1 N — *T*T(J
Boston, MA;
8/1991-1995
with mean
6 year follow-up
Pb Biomarker
Data
Mean (SD)
baseline
Blood Pb: 6.5
(4.2) ug/dl_
10th-90th
percentile: 2.1-7.6
Tibia Pb: 21. 5
(13.5)ug/g
Patella Pb: 32.4
(20.5) ug/g
Outcome
Change in
serum
creatinine per
year
x10(mg/dL)
Statistical Analysis
Log linear regression
adjusted for age, age
squared, BMI,
hypertension, diabetes,
smoking status, alcohol
consumption, analgesic
use, baseline serum
creatinine, serum
creatinine squared
Effect Estimate (95% Cl)
Entire Cohort ( N=448) 0.0210 (-0.0019, 0.048)
With rliahpfpc;- 0 18 (0 07 0 9Q\
VVILII UldUCLCo. U. IO IU.U/ , \J . ^-<3 >
Without diabetes: 0.014 (-0.009, 0.037)
With hypertension: 0.019 (-0.027, 0.065)
Without hypertension: 0.021
(.-1 r\r\~7 r\ r\AC\\
-0.007, 0.049)
Per unit increase in logarithmic (In)-transformed tibia
Pb
Entire Cohort (N=448) 0.007 (-0.0028, 0.0168)
With diabetes: 0.082 (0.03, 0.14)
Without diabetes: 0.005 (-0.01, 0.02)
With hypertension: 0.023 (0.003, 0.04)
Without hypertension: 0.0004 (-0.01, 0.01)
Cross-Sectional Results
Kim et al. (1996) Adult males
N=459
Boston, MA;
Multiple
examinations
Median baseline
blood
Pb: 8.6 ug/dl_
1 0th-90th
percentile:
A n 17 ^
Serum
creatinine
(mg/dL)
Random-effects modeling
adjusted for baseline age,
time since initial visit,
BMI, smoking status,
alcohol ingestion,
education level, and
hypertension.
Peak blood Pb < 40 ug/dl_: 0.0017 (0.0005, 0.003)
Peak blood Pb < 25 ug/dl_: 0.0021 (0.0007, 0.0035)
Peak blood Pb < 10 ug/dl_: 0.0033 (0.0012, 0.0053)
4-433
-------
Table 4-25 (Continued): Additional characteristics and quantitative data for associations of blood and bone Pb with kidney
outcomes for results presented in Figure 4-30 and Figure 4-31.
Reference
Tsaih et al.
(2004)
Population
Study Location
Time Period
Adult males
N - 448
I N — T^U
Boston, MA;
8/1991-1995
with mean 6 yr
follow-up
Results for Fiqure 4-30: Neqative
Pb Biomarker
Data
Mean (SD)
baseline
Blood Pb:
6.5 (4.2) ug/dl_
10th-90th
percentile:
2.6-11.5
Repeated
measures
10th QOth
I ULI I C7ULI I
percentile: 2.1-7.6
Tibia Pb: 21. 5
(13.5)ug/g
Patella Pb: 32.4
(20.5) ug/g
Effect Estimates
Outcome
Serum
creatinine
(mg/dL)
Statistical Analysis
Log linear regression
adjusted for age, age
squared, BMI,
hypertension, diabetes,
smoking status, alcohol
consumption, analgesic
I |c:P
uoc
Effect Estimate (95% Cl)
Baseline blood Pb
With rliahptpc;- 0 OOQ 1 0 038 0 (]9fl\
v v in i u iciucico. \j.\j\j\y lu.uoo, u.u^ui
Without diabetes: -0.004 (-0.010, 0.003)
With hypertension: 0 (-0.013, 0.013)
Without hypertension: -0.005 (-0.011, 0.
Follow-up blood Pb
With diabetes: 0.053 (-0.032, 0.138)
Without diabetes: 0.034 (0.007, 0.061)
With hypertension: 0.083 (0.038, 0.128)
002)
Without hypertension: 0.014 (-0.016, 0.044)
Indicate Poorer
Function
Per 1 |jg/dL increase in blood Pb within the
Longitudinal Results:
Yu et al. (2004)
Adult CKD
patients
N = 121
Taipei, Taiwan;
4ft mrinth
HO 1 1 IUI III 1
longitudinal
study period
Mean (SD)
baseline blood
Pb: 4.2
(2.2) ug/dL
10th-90th
percentile: 2.0-5.1
Change in
MDRD eGFR
over 4 yr/1 00
(ml_/min/1.73m2
body surface
area)
Cox proportional hazard
model examined
whether a predictor was
associated with renal
function including age,
sex, BMI,
hyperlipidemia,
hypertension, smoking,
use of ACE inhibitor,
baseline serum
creatinine, daily protein
excretion, daily protein
intake, underlying kidney
disease
10th-90th percentile interval
-0.040 (-0.072, -0.008)3
4-434
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Table 4-25 (Continued): Additional characteristics and quantitative data for associations of blood and bone Pb with kidney
outcomes for results presented in Figure 4-30 and Figure 4-31.
Reference
Cross-Sectional
Akesson et al.
(2005)
Staessen et al.
(1992)
Payton et al.
(1994)
Population
Study Location
Time Period
Results:
WHILA,
adult women
N = 820
Sweden;
6/1999-1/2000
Adults
N -1 Q81
IN — I , \yO I
Belgium;
1985-1989
Adult males
N -744
IN — / *T*T
Boston, MA;
-IQftft 1QQ1
i \yoo i \y\y i
Pb Biomarker
Data
Median (5-95%
Cl) concurrent
blood Pb: 2.2 (1.1,
4 6) uq/dL
/ "13
10th-90th
percentile: 1.3-3.8
Concurrent blood
Pb mean (SD)
Males: 11.4ug/dL
Females:
7.5 ug/dL
10th-90th
percentile:
3.7-15.1
Mean (SD)
concurrent blood
Pb: 8.1
(3.9) ug/dL
10th-90th
percentile:
4.1-12.9
Outcome
Creatinine
clearance/100
(mL/min)
Cystatin C-
based eGFR
(Larsson et al.,
2004J/100
(mL/min)
Creatinine
clearance/100
(mL/min)
Log-transformed
creatinine
clearance
(mL/min)
Statistical Analysis Effect Estimate (95% Cl)
Linear regression -0.018 (-0.03, -0.006)
adjusted for age, BMI,
diabetes, hypertension,
nephrotoxic drug, -°-02 (-°-03' °-007)
smoking status
Log linear regression Females: -0.067 (-0.108, -0.027)3
adjusted for age, age Malp=;- 0 051 I 0 OQ7 0 047^
J \j i \j IVIalCo. \J.\J\J \ I \J.\J\yi , \J.\J^l I
squared, sex, BMI, BP,
ferritin level, smoking
status, alcohol ingestion,
rural/urban residence,
analgesic and diuretic
use, blood and urinary
Cd, diabetes,
occupational exposure
to heavy metals, and
gamma glutamyl
transpeptidase
Log linear regression -0.040 (-0.079, -0.0015)
adjusted forage, BMI,
analgesic and diuretic
use, alcohol
consumption, smoking
status, SBP, DBP
4-435
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Table 4-25 (Continued): Additional characteristics and quantitative data for associations of blood and bone Pb with kidney
outcomes for results presented in Figure 4-30 and Figure 4-31.
Population
Study Location
Reference Time Period
Results for Fiqure 4-31 : Analysis
Muntneretal. NHANES III,
(2003) adults
N = 4,813
U.S.; 1988-1994
Navas-Acien et NHANES III,
al. (2009) adults
N = 14,778
U.S.; 1999-2006
Pb Biomarker
Data
Outcome
Statistical Analysis
of Blood Pb Quartiles
Mean (SD)
concurrent blood
Pb:
With
Hypertension: 4.2
(0.14)ug/dL
Q1: 0.7 to 2.4
Q2: 2. 5 to 3. 8
Q3: 3.9 to 5. 9
Q4: 6.0 to 56.0
Without
Hypertension:
3.3(0.10)ug/dL
O1 • 0 7 to 1 6
\j< I . \J. 1 \.\J 1 .\J
Q2: 1.7 to 2.8
Q3: 2.9 to 4. 6
Q4: 4. 7 to 52. 9
Geometric mean
concurrent blood
Pb: 1.58ug/dL
Q1. < 1.1
Q2: 1.2 to 1.6
Q3: 1.7 to 2.4
Q4: >2.4
Elevated Serum
Creatinine
(99th percentile
of each race-sex
specific
distribution for
healthy young
adults)
CKD
eGFR <60
mL/minute/1.73
KV,2
m
Albuminuria and
eGFR <60
mL/minute/1 73
m2
Logistic regression
adjusted for age, race,
sex, diabetes, SBP,
smoking, history of CVD,
BMI, alcohol
consumption, household
income, education level,
marital status, health
insurance
Logistic regression
adjusted for survey year,
age, sex, race/ethnicity,
BMI, education,
smoking, cotinine,
alcohol intake,
hypertension, diabetes,
menopausal status
Effect Estimate (95% Cl)
% change in kidney outcome
Q1: Referent
With hypertension
Q2:47%(3, 110)
Q3: 80% (34, 142)
Q4: 141% (46, 297)
Without hypertension
Q2: 11% (-44, 121)
Q3: 19% (-38, 125)
Q4: 90/0 (.47, 122)
With hypertension
Q2:44%(0, 109)
Q3: 85% (32, 159)
Q4: 160% (52, 345)
Without hypertension
Q2:-10%(-63, 116)
Q3: 0%(-55, 122)
Q4: 9% (-59, 189)
Q1: Referent
Q2: 10% (-20, 51)
Q3: 36% (-1 , 85)
Q4: 56% (17, 108)
Q2: 53% (-15, 177)
Q3: 57% (-17, 198)
Q4: 139% (31, 337)
a95% Cl: estimated from given p-value.
4-436
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Patient Population Studies
CKD as defined by the National Kidney Foundation (NKF) - Kidney Disease Outcomes
Quality Initiative workgroup (NKF. 2002) is the presence of markers of kidney damage
or GFR < 60 mL/min/1.73 m2 for > 3 months. The MDRD equation is the most common
one used in the eGFR determination for this definition. Notably, decreased GFR is not
required for the first criterion and markers of kidney damage are not required for the
second criterion.
Several key studies in CKD patients provide prospective data that indicate that higher
baseline blood Pb level is associated with greater CKD progression overtime (kidney
function decline) in patient populations (Table 4-26). Yu et al. (2004). discussed in the
2006 Pb AQCD, followed 121 patients over a four year period. Eligibility required well-
controlled CKD with serum creatinine between 1.5 and 3.9 mg/dL. Importantly, EDTA-
chelatable Pb < 600 ug/72 hours, a level below that traditionally thought to indicate risk
for Pb-related nephrotoxicity, was required at baseline. Patients with potentially unstable
kidney disease were excluded (i.e., due to systemic diseases such as diabetes). Mean
blood Pb and EDTA-chelatable Pb levels were 4.2 ug/dL and 99.1 ug/72 hours,
respectively. Cox proportional hazard modeling indicated lack of significant association
between serum creatinine changes and various potential confounding factors (Table
4-25). examined one at a time. Only chelatable Pb (body Pb burden indicator) was
significantly associated with overall risk for the primary endpoint (doubling of serum
creatinine over the 4-year study period or need for hemodialysis). When the group was
dichotomized by EDTA chelatable Pb level, Kaplan-Meier analysis demonstrated that
significantly more patients (15/63) in the high-normal group (EDTA chelatable Pb level >
80 but < 600 ug/72 hours) reached the primary end point than did those in the lower
EDTA chelatable Pb levels (<80 ug Pb/72 hours) group (2/58). Associations between
baseline chelatable or blood Pb level and change in serial measurements of eGFR
(estimated by the MDRD equation (Levey etal. 1999)) were modeled separately using
generalized estimating equations. Based on these models, a 10 ug higher chelatable Pb
level or 1 ug/dL higher blood Pb level reduced the GFR by 1.3 and 4.0 mL/min/1.73 m2,
respectively, during the 4-year study period. The use of estimated GFR provides a better
estimate of progressive changes of renal function than creatinine clearance used in the
other related studies.
Recent studies expanded the CKD patient populations in which this effect was observed
to include those with diabetic nephropathy (Lin et al.. 2006b) and with the lowest blood
Pb levels studied to date (Lin et al.. 2006a). A potential limitation in these studies is lack
of researcher blinding. The treatment protocol, which included additional calcium
4-437
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disodium EDTA infusions as needed based on Pb levels during follow-up, may differ
from the placebo protocol; therefore, patients may have been able to discern their
treatment group. In addition, it is possible that the improved renal function following the
lowering of blood Pb level by chelation is a result of the secondary effect of reducing
ROS, blood Pb level or both. Results of these observational studies have been
summarized in Table 4-26 (Weaver and Jaar. 2010).
Table 4-26 Prospective patient population studies: kidney function decline.
Study
Lin et
al.
(2003)
Yuet
al.
(2004)
Lin et
al.
(2QQ6b)
Lin et
al.
(2006a)
Baseline
mean
(SD)
blood
Pb
N (ng/dL)
202 5.3(2.9)
121 4.2(2.2)
87 6.5(3.4)
108 2.9(1.4)a
Baseline mean
(SD) chelatable
Pb (ng/72 hours)
104.5(106.3)
99.1 (83.4)
108.5(53.8)
40.2(21.2)
(all <80)
Baseline
mean(SD)
eGFR
(mL/min
/1.73m2)
41.6(14.4)
36.0(9.8)
35.1 (9.0)
47.6(9.8)
Decline in
eGFR per
1 SD higher
Years of Pb dose at
follow- baseline
up per year
2 0.16
4 2.7
(chelatable)
2.2 (blood Pb)
1 3.87
2 1.1
Comments
Largest study to date
Longest follow-up;
1 |ig/dL higher blood
Pb, at baseline,
associated with
4.0mL/min/1.73m2
reduction in eGFR
over 4 years
Type II diabetics with
nephropathy
Lowest Pb exposed
CKD patients
aNotably, mean blood Pb level in this study was below that observed in a recent large general population study of 50- to 70-year
olds in Baltimore, MD (Martin et al., 2006).
Source: Reprinted with permission of UpToDate.com, Weaver and Jaar (2010)
A recent population-based case-control study examined occupational Pb exposure as a
risk factor for severe CKD (Evans et al.. 2010). The study included 926 cases with first
time elevations of serum creatinine >3.4 mg/dL for men and >2.8 mg/dL for women and
998 population-based controls. Occupational Pb exposure was assessed using an expert
rating method based on job histories; no biomarkers of Pb exposure were measured. In
multivariable logistic regression modeling, the OR for CKD (adjusted for age, sex,
smoking, alcohol consumption, diabetes, education, and BMI) was negative but not
significant (0.97 [95% CI: 0.68, 1.38] in Pb-exposed compared to non-exposed
participants). In addition, the CKD patients were followed prospectively for a mean of
2.5 years for the 70 Pb-exposed patients and 2.4 years for the 731 patients without past
4-438
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occupational Pb exposure. Mean eGFRs (using the MDRD equation) were 16.0 and 16.6
mL/min/1.73 m2 in exposed and non-exposed patients, respectively, indicating severe
disease in both groups. The authors caution that although the high serum creatinine limit
(exclusion criteria: 3.4 mg/dL for men and 2.8 for women), enough to ensure both high
sensitivity and specificity, is a strength in the case-control analyses, it is a limitation in
the prospective follow up analysis; they note the further limitations that serum creatinine
was not measured in controls and the absence of blood Pb measurements. The expert
ratings used in this study may have lower validity and reliability as compared to other
exposure assessment methods (Teschke et al.. 2002) including blood and bone
measurements used in the majority of well-conducted studies. The lack of blood or bone
Pb exposure estimates in this study compared to other studies considered suggests placing
reduced weight on this study.
Strengths included virtually complete case ascertainment and minimal loss to follow-up.
Exposure assessment was presented by the authors as both a strength and a limitation.
Expert rating methods are commonly used when biological monitoring is not an option
and in case-control studies where many occupational exposures are considered. In
Pb-kidney research, this approach is uncommon except in the case-control setting.
However, given the challenges of interpreting blood Pb in dialysis patients (discussed
below), this approach may have advantages in this study of such severe CKD. Other case-
control studies examining occupational risk factors for CKD found Pb exposure to be a
risk factor (Nuvtsetal.. 1995; Steenland et al.. 1990). Nuyts et al. (1995) found adults
with history of occupational Pb exposure to have elevated odds of CKD (OR for ever-
versus never-exposed: 2.11 [95% CI: 1.23, 4.36]). The association was weaker in
Steenland et al. (1990) (OR for ever- versus never-exposed: 1.73 [95% CI: 0.82, 3.65]).
Regular moonshine consumption, also a potential source of Pb exposure, was a stronger
risk factor for CKD (OR: 2.42 [95% CI: 1.10, 5.36]).
The prospective observational aspect of Evans et al. (2010) is similar in design to the
work of Lin and colleagues but differs in several important respects. In Evans et al.
(2010). only occupational Pb exposure was considered whereas the work in Taiwan
excluded occupational exposure and used blood and chelatable Pb measures. In the past
in developed countries, environmental exposures were substantial. For example, mean
tibia Pb levels were 21.5 and 16.7 |o,g/g bone mineral, in environmentally-exposed 50- to
70-year-old African-Americans and whites, respectively, in Baltimore (Martin et al..
2006). In Korean Pb workers, mean baseline tibia Pb level was only twofold higher (35.0
Hg/g) (Weaver et al.. 2003a) which illustrates the substantial body burden in middle- and
older-aged Americans from lifetime Pb exposure. Declines in blood Pb levels in Sweden
have been reported and attributed to the leaded gasoline phase-out (Stromberg et al..
1995; Blinder etal.. 1986). although blood Pb levels were lower than those noted during
4-439
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the U.S. phase-out. Finally, the severe degree of CKD among subjects in Evans et al.
(2010) creates a survivor bias at enrollment and limits the eGFR decline possible during
follow-up, thus limiting the ability to identify factors that influence that decline.
ESRD Patient Studies
End stage renal disease (ESRD) is a well-established public health concern, and is
characterized by the use of dialysis to perform the normal functions of the kidney.
Incidence and prevalence in the U.S. continue to increase resulting in rates that are the
third highest among nations reporting such data (U.S. Renal Data System. 2009). Studies
in patients with CKD requiring chronic hemodialysis have also been published in the past
five years. A study of 271 adult patients on regular thrice weekly dialysis reported much
higher blood Pb levels than had been appreciated by the treating clinicians (Davenport et
al.. 2009). Blood Pb levels ranged from 3 to 36.9 ug/dL; 25.5% had levels >20 ug/dL,
59% had values of 10-20 ug/dL, and 15.5% were <10 ug/dL. Few details on the statistical
analysis were provided which complicates interpretation of the findings. However, blood
Pb was positively correlated with hemodialysis vintage (months on dialysis; Spearman
r = 0.38, p <0.001); negatively correlated with urine output (r = -0.44, p <0.001) and
higher in patients using single carbon filter and reverse osmosis water purification
devices. Another recent publication reported higher Pb in dialysate than in the tap water
used in its preparation (Chen et al., 2009a). A systematic review of a wide range of trace
elements in hemodialysis patients reported higher Pb levels in patients compared to
controls although the difference was not large (Tonelli et al., 2009). These data suggest
that blood Pb monitoring in dialysis patients may be useful.
Interpretation of blood and bone Pb in patients on dialysis is challenging for several
reasons. First, renal osteodystrophy, the bone disease related to kidney disease, may
result in increased release of Pb from bone stores. Thus, interpretation of blood and even
bone Pb levels may require adjustment with one or more of a range of osteoporosis
variables. Secondly, as observed above (Davenport et al.. 2009), residual kidney function
may have a substantial impact on blood Pb levels in populations with such minimal
excretion. Third, as illustrated in the studies cited above (Chen et al.. 2009a; Davenport et
al.. 2009). water and concentrates used in dialysis may be variable sources of Pb. A
recent study reported decreased blood Pb in post-dialysis compared to pre-dialysis
samples (Kazi et al.. 2008). Thus, substantial fluctuations in blood Pb are possible while
on dialysis. Finally, anemia is common in CKD and Pb is stored in red blood cells. Thus,
measurement of blood Pb in anemia may require adjustment for hemoglobin; no
standardized approach to this currently exists.
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Given these caveats, a small cross-sectional pilot study observed higher median blood Pb
levels in 55 African-American dialysis patients compared to 53 age- and sex-matched
controls (6 and 3 ug/dL respectively; p <0.001) (Muntner et al., 2007). However, median
tibia Pb was higher in ESRD patients although the difference did not reach statistical
significance (17 and 13 ug/g bone mineral, respectively [p = 0.13]). Further, the authors
note the limitation related to the sample size based on too few cases needed to achieve
statistical significance from power calculations.
In order to determine the potential impact of renal osteodystrophy, median blood and
tibia Pb levels in the dialysis patients were compared by levels of serum parathyroid
hormone, calcium, phosphorus, and albumin and were not found to be significantly
different (Ghosh-Narang et al., 2007). A study of 211 diabetic patients on hemodialysis
(Lin et al.. 2008) found parathyroid hormone and serum creatinine to be associated with
blood Pb level in crude but not adjusted associations. In contrast, a study of 315 patients
on chronic peritoneal dialysis observed parathyroid hormone to be positively correlated
and residual renal function to be negatively correlated with logarithmic-transformed
blood Pb levels after adjustment (Lin etal.. 2010). In the prospective portion of this
study, blood Pb levels at baseline were categorized by tertile (range of 0.1 to 29.9 ug/dL
with cut points of 5.62 and 8.66 ug/dL). Cox multivariate analysis, after adjustment for
parathyroid hormone level, residual renal function, and 20 other variables, showed
increased all-cause mortality in the middle (5.62-8.66 ug/dL) and highest (>8.66 ug/dL)
compared to the lowest (<5.62 ug/dL) tertiles after 18 months of follow-up (hazard ratio=
2.1 [95% CI: 2.0, 2.2] and 3.3 [95% CI: 1.3, 13.5], respectively). A recent publication of
an 18-month follow-up of 927 patients on maintenance hemodialysis also reported
increased hazard ratios for all-cause (4.7 [95% CI: 1.9, 11.5]), cardiovascular-cause (9.7
[95% CI: 2.1, 23.3]), and infection-cause (5.4 [95% CI: 1.4, 20.8]) 18-month mortality in
the highest (>12.64 ug/dL) compared to the lowest tertile (<8.51 ug/dL) of baseline blood
Pb level, after adjustment for sex, urban residence, hemodialysis vintage, hemoglobin,
serum albumin, and ferritin (Lin etal.. 2011). Given other recent publications in
hemodialysis patients by this group, it would be valuable to examine these risks after
adjustment for hemoglobin A1C (Lin-Tan et al.. 2007a). and blood Cd (Yen etal.. 2011;
Hsu et al.. 2009a).
Clinical Trials in Chronic Kidney Disease Patients
Randomized chelation trials in CKD patients, uncommon in nephrotoxicant research,
provide unique information on the impact of Pb on the kidney. These studies have been
performed by Lin and colleagues at the Chang Gung Memorial Hospital in Taipei,
Taiwan and involve similar study designs. Initially, patients were observed to compare
CKD progression prior to chelation. Then, CKD patients whose diagnostic EDTA
4-441
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chelatable Pb levels were within certain ranges (generally 60-600 u.g/72 hours and thus
below the level commonly considered for chelation) were randomized. The treated group
received weekly chelation with 1 g EDTA intravenously for up to 3 months. The control
group received placebo infusions. In the follow-up period, chelation was repeated for
defined indications such as increased serum creatinine or chelatable Pb levels above
specified cut-offs. Placebo infusions were repeated in the controls as well. During the
intervention periods single-blind, randomized, placebo-controlled interventions were
conducted. The results of the most recent of these trials are summarized in Table 4-27
below.
Table 4-27 Clinical randomized chelation trials in chronic kidney disease
patients.
Reference Group
Lin et al. Chelated
(2003)
Control
Lin et al. Chelated
(2006b)
v ' Control
Lin et al. Chelated
(2006a)
Control
Lin-Tan et Chelated
al. (2007b)
Control
Baseline
mean(SD)
blood Pb
N (ng/dL)
32 6.1 (2.5)
32 5.9(3.0)
15 7.5(4.6)
15 5.9(2.2)
16 2.6(1.0)a
16 3.0(1.1)
58 5.0(2.2)
58 5.1 (2.6)
Baseline
mean(SD)
chelatable Pb
(H9/72 hr)
150.9(62.4)
144.5(87.9)
148.0(88.6)
131.4(77.4)
43.1 (13.7)
47.1 (15.8)
164.1 (111.1)
151.5(92.6)
Baseline
mean(SD)
eGFR
(mL/min
/1.73 m2)
32.0
(12.1)
31.5(9.0)
22.4 (4.4)
26.3(6.2)
41.2
(11.2)
42.6(9.7)
36.8
(12.7)
36.0
(11.2)
Change
Months in eGFR
of Per V
treatment (mL/min
/follow- /1.73
up m2)
27 + 1.07
-2.7
15 -3.5
-10.6
27 +3.0
-2.0
51 -0.3
-2.9
Comments
Subjects
with Type II
diabetes
and
nephropathy
Lowest Pb
exposed
and treated
range
Body Pb
Burden
(72 h
urinary Pb
excretion)
>20to
<80|ig
Subjects
without
diabetes
aNotably, mean blood Pb level in this study was below that observed in a recent large general population study of 50- to 70-year
olds in Baltimore, MD (Martin et al.. 2006).
4-442
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The unique body of work in patient populations by Lin and coworkers, both observational
and experimental, has numerous strengths including prospective study design,
randomization, Pb assessment that includes estimates of the bioavailable dose,
longitudinal statistical analysis, and control for multiple kidney risk factors. However, the
generalizability of the results to broader populations is unknown. In addition, the
association observed between Pb dose and decline in GFR has been variable; the annual
decline in eGFR per standard deviation higher Pb dose at baseline was much lower in the
2003 study than in subsequent publications (Table 4-27 above). Small sample sizes and
differences in renal diagnoses between groups may be factors in this variability.
The studies presented in Table 4-27 had a number of potential limitations. This included
small sample size. Also, the researchers were not blinded and the treatment protocol,
which included additional calcium disodium EDTA infusions as needed based on Pb
levels during follow-up, may differ from the placebo protocol; therefore, patients may be
able to discern their treatment group. Another possible limitation may be the shorter
follow-up time in some of the studies. The use of creatinine clearance to assess changes
in renal function may limit interpretation of results as discussed in Section 4.5.1.1. Also,
the effects observed following chelation therapy may result from removal of other ions
such as Zn, Cu, and Fe. In addition, changes in kidney function after treatment with Pb
chelating agents may be by mechanisms other than reduction in Pb body burden.
Chelating agents have been shown to act as antioxidants. DMSA abolished reactive
oxygen species formation (i.e., MDA and nitrotyrosine in interlobular arteries) and was
protective against non-Pb-induced nephrosclerosis in rats (Gonick et al.. 1996). EDTA
administration enhanced endothelial NO production and reduced kidney damage in a rat
model of ischemia-induced acute renal failure (Toglieni et al.. 2006). Improved renal
function following administration of chelating agents have been reported in rodent
models of Pb-induced nephrotoxicity (Sanchez-Fructuoso et al.. 2002a: Sanchez-
Fructuoso et al., 2002b; Khalil-Manesh et al., 1992a). Chelation did not appear to
improve Pb-induced structural damage (Khalil-Manesh et al.. 1992a): again suggesting
that improved hemodynamics may be a result of reduction in reactive oxidant species,
which could be due to reduced Pb level and/or directly to the chelating agent (Gonick et
al., 1996). Despite these uncertainties and limitations, the most plausible explanation for
the combination of the observational and experimental chelation work of Lin and
colleagues is that reduced Pb is the underlying reason for improved kidney function. The
limitations discussed suggest putting less weight on these clinical trials in CKD patient
studies. This study design requires replication in larger populations at multiple clinical
centers to confirm that the change in renal function may be due to removal of Pb.
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Occupational Studies
The vast majority of studies in the literature on the impact of Pb on the kidney have been
conducted in the occupational setting. In general, study size and extent of statistical
analysis are much more limited than those in general population studies. Publications in
few populations have reported adjusted results in occupationally exposed workers in the
five years since the 2006 Pb AQCD.
A study of Korean Pb workers included in the 2006 Pb AQCD reported inconsistent
results with higher Pb measures associated with worse renal function in some models and
better renal function in other models (Weaver et al., 2003a). Null associations were
observed when the entire population was studied. In models of effect modification by
age, a pattern emerged in which higher Pb exposure and dose measures were associated
with worse renal function in older workers and better renal function in younger workers
(Weaver et al., 2003a). This cohort was expanded into a two-year prospective cohort
study with data from Weaver et al. (2003a) serving as baseline Pb measures (Weaver et
al., 2009). Generalized estimating equations were used to model change in kidney
function between each evaluation in relation to tibia Pb and concurrent change in blood
Pb in 537 current and former Pb workers while adjusting for baseline Pb dose and other
covariates (Weaver et al.. 2009). Tibia Pb was evaluated at the beginning of each follow-
up period (yearly on average) and Pb biomarker levels were adjusted for baseline levels
and other covariates. In males, serum creatinine declined and calculated creatinine
clearance increased over the course of the study; these changes were largest in
participants whose blood Pb declined concurrently or whose tibia Pb was lower at the
beginning of the follow-up interval. In females, decreasing serum creatinine was
associated with declining blood Pb (as in males); however, increasing blood Pb was
associated with a concurrent increase in serum creatinine. Women (25.9% of the study
population) were older and more likely to be former Pb workers than were men which
may have been important factors in the effect modification observed by sex.
Chia and colleagues observed a significant, positive association between concurrent
blood Pb and urine NAG in linear regression models after adjustment for age, sex, race,
exposure duration, ALAD G177C polymorphism and the interaction between ALAD
genotype and blood Pb (Chiaetal.. 2006). Similar positive associations were observed
between blood Pb and a wider range of EBE markers in models that adjusted for age, sex,
race, exposure duration, and the HpyCH4 ALAD polymorphism (Chia et al., 2005).
Other studies published in the last 5 years also focused on ALAD polymorphisms but did
not find effect modification to be in a consistent direction (Gao etal., 2010a; Wang et al..
2009a: Weaver et al.. 2006; Weaver et al.. 2005b). In adults with the ALAD2 genotype,
Pb has been associated with better and poorer renal function in separate cohorts of Pb
workers.
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Two studies of occupationally-exposed adults have performed benchmark dose
calculations for the effect of Pb on the kidney. Both used only EBE markers and found
NAG to be the most sensitive outcome; reported lower confidence limits on the
benchmark doses were 10.1 ug/dL (Sun et al.. 2008b). and 25.3 ug/dL (Lin and Tai-vi.
2007).
A number of other publications in the five years since the 2006 Pb AQCD, have reported
significantly worse kidney outcomes in unadjusted analyses in occupationally-exposed
adults compared to unexposed controls (Onuegbu et al.. 2011; Patil et al., 2007) and/or
significant correlations between higher levels of Pb biomarkers and worse kidney
function (Alasiaet al.. 2010: Khan et al.. 2008: Garcon et al.. 2007: Lin and Tai-vi. 2007:
Alinovi et al.. 2005). A study of 155 male workers reported significant, positive
correlations between blood and urine Pb and urine NAG and albumin after controlling for
age and job duration (Sun et al.. 2008b). One small study found no significant differences
(Orisakwe et al.. 2007). In a study of 108 Pb workers with mean blood Pb level of
36.2 (ig/dL, no significant correlations were observed between blood Pb concentration
and GFR, creatinine clearance, uric acid clearance or uric acid excretion fraction
(Karimooy et al.. 2010). However, interpretation of this study is limited by the fact that
"only 30 subjects had a correct 24 hours urine volume" and no methods are described for
kidney outcome measurement or analysis.
Overall, the occupational literature published in the last five years on the kidney impact
of Pb exposure has been more consistent in reporting statistically significant associations
than were data reviewed for the 2006 Pb AQCD. This may reflect increased reliance on
EBE markers as more sensitive outcome measures, publication bias, or multiple
comparisons due to a greater number of outcomes assessed. The limitations and lack of
consistent results discussed above suggest putting less weight on these occupational
studies.
A small number of publications that include concentration-response information provides
evidence of Pb-related nephrotoxicity in the occupational setting across the blood Pb
ranges analyzed (Weaver et al.. 2003a: Ehrlich et al.. 1998). Data in 267 Korean Pb
workers in the oldest age tertile (mean age = 52 years) did not provide evidence of a
threshold for a Pb effect on serum creatinine levels (added variable plot shown in Figure
4-32) (Weaver et al.. 2003a). It is important to note the uncertainty regarding whether the
concentration-response information provided in these studies applies to lower blood Pb
levels or to populations with lower current environmental Pb exposures.
4-445
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0
•-Ł= o
o
_
0
CO
CD
CD
6
0
40
GO
Adjusted blood Pb level (|jg/dL)
Note: Both the adjusted regression line (straight line) and the line estimated by the smoothing method of the S-PLUS statistical
software function lowess (line with curves) are displayed. Both have been adjusted for covariates. For ease of interpretation, axes
have been scaled, so that the plotted residuals are centered on the means, rather than zero.
Source: Reprinted with permission of the BMJ Publishing Group, Weaver et al. (2003a)
Figure 4-32 Added variable plot of association between serum creatinine and
blood Pb in 267 Korean Pb workers in the oldest age tertile.
A major challenge in interpretation of the occupational literature is the potential for
Pb-related hyperfiltration. Hyperfiltration involves an initial increase in glomerular
hypertension which results in increased GFR. If persistent, the risk for subsequent CKD
increases. This pattern has been observed in diabetes, hypertension, and obesity (TSfenov
et al.. 2000). As discussed in the 2006 Pb AQCD (U.S. EPA. 2006b). findings consistent
with hyperfiltration have been observed in occupational populations (Weaver et al..
2003a; Hsiao et al.. 2001; Roels et al., 1994). a study of adults who were Pb poisoned as
children (Hu. 1991). and a study in European children (de Burbure et al.. 2006).
Longitudinal data in Pb-exposed rodents provide evidence of a hyperfiltration pattern of
increased, followed by decreased GFR, associated with Pb exposure and are critical in
interpretation of the human Pb-kidney literature (Khalil-Manesh et al.. 1992b). Pb could
induce glomerular hypertension resulting in hyperfiltration by several mechanisms
including increased ROS, changes in eicosanoid levels, and/or an impact on the renin-
4-446
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angiotensin system (Vaziri. 2008b; Roels et al., 1994). Whether hyperfiltration
contributes to pathology in humans is unclear; longitudinal studies are needed.
The 2006 Pb AQCD provided several explanations for this inconsistency (U.S. EPA.
2006b) (Chapter 6, pp 99):
"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."
Regardless, 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 Korean Pb workers (Weaver et al.. 2003a: Weaver etal..
2003b), significant associations in opposite directions were observed only when relevant
effect modifiers such as age or genetic variants in ALAD, VDR, and NOS were included
in the model. This is a valid concern for risk assessment, since the factors involved in
these inverse associations in Pb-exposed workers are not well defined at present.
4.5.2.2 Epidemiology in Children
Pb Nephrotoxicity in Children
Both the 2006 and 1986 Pb AQCDs noted that the degree of kidney pathology observed
in adult survivors of untreated childhood Pb poisoning in the Queensland, Australia
epidemic (Inglis etal.. 1978) has not been observed in other studies of childhood Pb
poisoning. Recent publications remain consistent with that conclusion; a recent study
observed an impact of childhood Pb poisoning on IQ but not kidney outcomes (Coria et
al., 2009). Chelation was raised as a potential explanation for this discrepancy in the
2006 Pb AQCD.
With declining Pb exposure levels, recent work has focused on studies in children with
much lower blood Pb levels. However, clinical kidney outcome (i.e., GFR) measures for
early kidney damage are insensitive in children who do not have many of the other
kidney risk factors that adults do, such as hypertension and diabetes. As a result, such
4-447
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studies have utilized EBE markers. However, data to determine the predictive value of
such biomarkers for subsequent kidney function decline in Pb exposed populations are
extremely limited (Coratelli et al., 1988) and may pose particular challenges in children
due to puberty-related biomarker changes (Sarasua et al.. 2003). The few studies included
in the 2006 Pb AQCD (U.S. EPA. 2006b) that analyzed clinical kidney outcomes in
children found associations with indicators of Pb exposure that were inconsistent in
direction. Pels et al. (1998) found no difference in mean serum creatinine between 62
children living near Pb-producing factories and 50 control children living in communities
without Pb emission sources. No significant associations between blood Pb level and
kidney function were observed in a population of adults (described in Section 4.5.2.1)
and children (N = 200) living around smelters in France; however, beta coefficients were
not reported (De Burbure et al., 2003). This study was expanded to include data from
300-600 European children (N varied by outcome) and reported higher concurrent blood
Pb level associated with lower serum creatinine and cystatin C (de Burbure et al., 2006).
Conversely, in a study of 200 Belgian adolescents aged 17 years, higher concurrent blood
Pb level was associated with higher serum cystatin-C (Staessen et al.. 2001).
Recent studies of children with elevated Pb exposure did not consistently indicate that Pb
exposure was associated with reduced kidney function. A study in 123 children of
workers in Pakistani Pb smelters and battery recycling plants and 123 control children,
ages 1-6 years, reported elevated blood Pb levels, serum creatinine and urea in children of
Pb-exposed workers compared to controls (medians: 8.1 versus 6.7 (ig/dL; 56 versus
52 (iM; and 4,500 versus 4,300 (iM, respectively (p < 0.01 for all) in unadjusted analyses
(Khan et al., 2010a). Blood Pb levels were correlated with serum creatinine (Spearman
r = 0.13; p = 0.05). However, a study of 77 participants, ages 10-25 years, who were
previously Pb poisoned through contaminated flour and chelated, reported no difference
in renal effects between children with blood Pb levels >48 (ig/dL and <43 (ig/dL although
lower IQ was observed in the subset who were exposed before the age of six years (Coria
et al.. 2009).
One of the key gaps identified in the 2006 Pb AQCD (U.S. EPA. 2006b) was limited data
in children and adolescents particularly with respect to GFR measures and in populations
without the elevated Pb exposure associated with Pb poisoning, living near a Pb source,
or having parents with occupational Pb exposures. A recently published NHANES
analysis in adolescents begins to fill this gap (Fadrowski et al.. 2010). Associations
between concurrent blood Pb and kidney function were investigated in 769 adolescents
aged 12-20 years in the U.S. NHANES III, conducted 1988-1994. Kidney function was
assessed with two eGFR equations. One utilized serum cystatin C and the other used the
more traditional marker, serum creatinine. Median concurrent blood Pb and cystatin C-
based eGFR levels were 1.5 ug/dL and 112.9 mL/min/1.73 m2, respectively. Cystatin C-
4-448
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based eGFR was lower (-6.6 mL/min/1.73 m2 [95% CI: -0.7, -12.6]) in participants with
blood Pb levels in the highest quartile (> 3.0 ug/dL) compared with those in the lowest
(<1 ug/dL). A doubling of blood Pb level was associated with a -2.9 mL/min/1.73 m2
(95% CI: -0.7, -5.0) lower eGFR. In contrast, the association between blood Pb and
creatinine-based eGFR, although in the same direction, was not statistically significant.
As these children were born between 1968 and 1982, some likely had higher Pb
exposures in earlier childhood, although notably, not as high or as long in duration as did
older adults examined in aforementioned studies. Nonetheless, in this study of NHANES
adolescents, there also is uncertainty regarding the magnitude, timing, frequency, and
duration of Pb exposure that contributed to the observed associations. Additional research
in children is warranted, in particular studies with longitudinal follow-up, multiple
outcome assessment methods, and examination of children born after Pb was banned
from gasoline.
4.5.2.3 Associations between Pb Dose and New Kidney Outcome
Measures
As noted above, in an effort to more accurately estimate kidney outcomes, new equations
to estimate GFR based on serum creatinine have been developed, and the utility of other
biomarkers, such as cystatin C, as well as equations based on them, are being studied.
However, few publications have utilized these state-of-the-art techniques when
evaluating associations between Pb or Cd dose and renal function. In addition to the
study in NHANES adolescents discussed above (Fadrowski et al.. 2010). a cross-
sectional study of Swedish women reported that higher concurrent blood Pb (median:
2.2 (ig/dL) and Cd (median: 0.38 (ig/L) levels were associated with lower eGFR based on
serum cystatin C alone (without age, sex, and race) after adjustment for socio-
demographic and CKD risk factors (Akesson et al.. 2005). Associations were comparable
to those using creatinine clearance as the kidney outcome for Pb; however associations of
Cd dose measures were stronger for the cystatin C based outcome. Staessen et al. (2001)
found a statistically significant association between concurrent blood Pb level and serum
cystatin C in a cross-sectional study of adolescents; creatinine-based measures were not
reported. However, in a cross-sectional study of 804 European children aged range 8.5 to
12.3 years, higher concurrent blood Pb levels were associated with lower serum cystatin
C and creatinine; these inverse associations were attributed to hyperfiltration (de Burbure
et al.. 2006). A recent publication compared associations of blood Pb and eGFR using the
traditional MDRD equation to those with four new equations: CKD-EPI, and cystatin C
single variable, multivariable, and combined creatinine/cystatin C, in 3,941 adults who
participated in the 1999-2002 NHANES cystatin C subsample (Spectoretal.. 2011).
Similar to the NHANES adolescent analysis, associations with the cystatin C outcomes
4-449
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were stronger. After multivariable adjustment, differences in mean eGFR for a doubling
blood Pb were -1.9 (95% CI: -3.2, -0.7), -1.7 (95% CI: -3.0, -0.5), and -1.4 (95% CI: -2.3,
-0.5) mL/min/1.73 m2, using the cystatin C single variable, multivariable and combined
creatinine/cystatin C equations, respectively, reflecting lower eGFR with increased blood
Pb. The corresponding differences were -0.9 (95% CI: -1.9, 0.02) and -0.9 (95% CI: -1.8,
0.01) using the creatinine-based CKD-EPI and MDRD equations, respectively.
4.5.2.4 Reverse Causality
As noted earlier, the reverse causality hypothesis suggests that the observed associations
between blood Pb and kidney function may be due to reduced excretion of Pb associated
with prior kidney damage rather than a causal association between Pb exposure and this
outcome. The potential for a bidirectional relationship because of reverse causality is
plausible in observational epidemiologic studies.
There are several techniques that can be used to assess the potential for reverse causality
to underlie associations between higher Pb dose and worse kidney function. Prospective
studies in which there are associations between baseline measurements of Pb biomarkers
and subsequent changes in renal function provide evidence to evaluate the possibility of
reverse causality. In the NAS, baseline blood Pb levels were associated with subsequent
declines in renal function over follow-up periods ranging from 3 to 6 years (Tsaih et al..
2004; Kim etal., 1996). Prospective data in CKD patients also revealed an association
between baseline Pb dose and decline in eGFR over follow-up periods as long as four
years (Yu et al., 2004). However, initial stages of renal decline are often difficult to
measure using clinical tests and baseline Pb levels may be accompanied by initial stages
of renal function decline. Another approach involves sensitivity analyses in which
associations are explored in participants with normal glomerular filtration. This approach
has been used in the analyses of the NAS; plots revealed that the association between
blood Pb level and serum creatinine was present across the entire range of serum
creatinine levels, including those in the normal range where excretion is not impaired
(Tsaih et al.. 2004; Kim et al.. 1996). Analyses restricted to the population with serum
creatinine below 1.5 mg/dL were conducted in a later publication and the authors
reported that associations were robust (Tsaih et al.. 2004). The use of serum creatinine
rather than eGFR to establish a normal kidney function cut-off is a limitation since there
can be substantial decrements in renal function with 'normal' serum creatinine. The
associations observed in both NAS studies were not limited to the segment of the
population with potentially clinically significant renal dysfunction in whom reduced Pb
excretion would be more likely.
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Additional evidence evaluating reverse causality was provided by findings among 153
adults with chronic kidney disease, in which renal failure was not associated with
increases in blood or bone Pb levels or chelatable Pb levels (Van de Vyver et al., 1988).
Batuman et al. (1983) found that chelatable Pb levels were similar in 27 adults with renal
disease of unknown and known non Pb-related causes where bone Pb levels (group
means: 18 and 19 (ig/g) were in the range of those measured in recent epidemiologic
studies. A pilot study of 55 ESRD cases at Tulane clinics (Muntner et al., 2007) reported
the median blood Pb level was significantly higher among the ESRD cases compared to
their control counterparts. For ESRD patients the distribution of blood Pb was shown by
pre-defined levels: 18.5% less than 5 (ig/dL; 66.7% of ESRD cases had blood Pb levels
between 5 and 9.9 (ig/dL; and 14.8% equal or greater 10 (ig/dL. Other studies have
reported elevated blood and bone Pb concentrations in patients with compromised renal
function, but prior elevated Pb exposure was often present. Hemodialysis patients
exposed to Pb in water (Davenport et al., 2009) and exposed occupationally or to alkyl Pb
from atmospheric contamination from gasoline additives (Colleoni et al.. 1993) had
higher blood Pb level compared to controls. Behringer et al. (1986) reported greater
blood Pb levels in patients with renal failure, but points to prior Pb exposure. Winterberg
et al. (1990) noted that bone Pb measured in hemodialysis patients without known Pb
exposure was elevated compared to chronic renal failure patients, patients after renal
transplantation, and a control group.
Although reverse causality has been typically considered plausible in the interpretation of
studies of CKD patients whose kidneys are known to be compromised, the reverse
causality hypothesis can be broadened to include a physiologic process in which Pb
levels are influenced over the entire range of kidney filtration rates. If this occurs, even
normal kidney function would impact blood Pb levels such that higher GFR would result
in greater Pb excretion and lower blood Pb levels. Serum creatinine levels are influenced
in this way over the entire range of kidney function; as a result, these levels are used to
estimate kidney function. However, creatinine is produced and excreted at a steady state
in the body which is one reason it was selected as a biomarker to assess kidney function.
This expanded hypothesis implies that low blood and bone Pb levels may reflect kidney
function in addition to exposure. If so, this would increase misclassification bias, with Pb
biomarkers reflecting both exposure and kidney function. Given the longstanding use of
blood Pb as a dose marker in research for many non-kidney outcomes this seems
unlikely. Thus, published research has not directly addressed this. One such approach
involves comparing associations of blood and urine Pb in models of kidney function. If
they are consistent, this hypothesis is not valid. However, urine Pb is rarely used and may
not be as reliable a biomarker as blood Pb (Gulson et al., 1998c).
4-451
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In summary, several lines of evidence support that reverse causality does not contribute
substantially to associations between higher blood Pb levels and worse kidney function.
These lines of evidence include prospective data observing that baseline Pb measures are
associated with subsequent declines in renal function, that associations in prospective
studies persist among adults with normal renal function, and that some studies indicate
renal failure does not increase Pb biomarker levels. Reverse causation remains a
completely plausible hypothesis. The existence of an association in adults with normal
renal function does not exclude reverse causation because it is not known how Pb
clearance varies within the range of normal kidney function. Thus, this bidirectional
relationship is possible and additional evidence is needed to fully elucidate the extent to
which diminished kidney function may itself result in increased Pb levels.
4.5.2.5 Toxicology
In animals, Pb has been found to induce changes in a wide range of indicators of renal
function. Most studies examined Pb exposure concentrations that resulted in higher blood
Pb levels (> 20 (ig/dL) than those in the current U.S. general population. While
toxicological information on renal dysfunction with blood Pb levels < 20 (ig/dL generally
is not available, dysfunction in kidney function measures, including urinary flow, ALP,
microalbumin, and NAG, was observed at blood Pb levels above 20 (ig/dL.
4-452
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Table 4-28 Animal toxicological studies reporting the effects of Pb exposure
Reference
Berrahal et
al. (2011)
Masso-
Gonzalez
etal.
(2009)
Roncal et
al. (2007)
Khalil-
Manesh et
al. (1993a)
Khalil-
Manesh et
al. (1992a)
Vyskocil et
al. (1995)
(as blood Pb level)
Species; Pb Dose;
Lifestage; Exposure
Sex Duration
Rat; Adult 50 ppm
Pb acetate in
drinking
water;
lactation day
1 to PND40 or
PND65
Rat; 300 ppm
Weanling Pb acetate in
pups drinking
water; GD1 to
PND21
Rat; Adult; 150 ppm
Male Pb acetate in
drinking
water;
16 weeks with
remnant
kidney
surgery
at week 4
Rat; Adult; 100 ppm
Male Pb acetate in
drinking
water;
12 months
Rat; Adult; 5,000 ppm
Male Pb acetate in
drinking
water; 1 , 6, or
9 months
Rat, Adult, 1,000 ppm
Female Pb acetate in
drinking
water; 2 or
4 months
on kidney
Blood Pb
Level with
Response
(ug/dL)
PND40: 12.7
PND65: 7.5
23
26
Mean at
3 months
29.4
Mean at
12 months
22
Range 9-34
At 1 month
7.9
At 6 months
-30
At 9 months
52
37.6
function.
Responses
Oxidative stress - Increased renal MDA (PND40 and
PND65), decreased renal SOD activity (PND40)
Morphology - Increased relative kidney weight (PND65)
Kidney function - Increased blood creatinine (PND40
and PND65), increased BUN (PND40), decreased uric
acid (PND65), increased kidney proteins (PND65).
Oxidative stress - Elevated TEARS and catalase
activity
Morphology - Elevated relative kidney weight at PND21
Inflammation - Increased the number of macrophages
& renal MCP-1 mRNA.
Morphology- Pb induced pre-glomerular vascular
disease of kidney (i.e., sclerosis, fibrosis, peritubular
capillary loss)
Kidney function - Decreased creatinine clearance,
increased serum creatinine, increased BUN, and
increased serum uric acid.
Morphology - Mild tubular atrophy and interstitial
fibrosis seen at 12 months, otherwise normal.
Kidney function - Increased GFR at 1 and 3 months,
increased NAG, and no change in GST.
At 1 month
Kidney function - no functional or pathological changes
At 6 months
Morphology - Prominent tubulointerstitial fibrosis and
segmental sclerosis. Increased kidney weight
Kidney function - Increased serum creatinine and SUN,
no change urinary NAG or GST
At 9 months
Morphology - Severe tubulointerstitial disease
Kidney function - Decreased GFR, increased serum
creatinine and SUN
Kidney function - No change in kidney function or
nephrotoxicity
4-453
-------
Table 4-28 (Continued): Animal toxicological studies reporting the effects of Pb exposure
(as blood Pb level) on kidney function.
Reference
Ademuyiwa
etal.
(2009)
Navarro-
Moreno et
al. (2009)
Khalil-
Manesh et
al. (1992b)
Wang et al.
(201 Oe)
Species; Pb Dose;
Lifestage; Exposure
Sex Duration
Rat; Adult 200, 300, and
400 ppm
Pb acetate in
drinking
water;
12 weeks
Rat; Adult; 500 ppm
Male Pb acetate in
drinking
water;
28 weeks
Rat; Adult; 5,000 ppm
Male Pb acetate in
drinking
water;
12 months
Rat; Adult; 300 ppm
Female Pb acetate in
drinking
water;
8 weeks
Blood Pb
Level with
Response
(ug/dL) Responses
200 ppm: 41 Kidney function - Renal phospholipidosis and depletion
300 ppm: 61 of renal cholesterol.
400 ppm: 39
43 Oxidative stress - Increased kidney lipid peroxidation
(i.e., TEARS)
Morphology- Electron micrography showed lumen
reduction, microvilli loss, brush border loss, and
mitochondrial damage
Kidney function - Elevated urinary pH and protein, and
glucose and blood in the urine.
Max 125 Morphology - Tubular atrophy and interstitial fibrosis
Mean 45 after 6 months. Increased urinary brush border
antigens.
Kidney function - Hyperfiltration at 3 months and
decreased GFR at 12 months. After 3 months, elevated
urinary NAG and GST.
20 (serum) Biomarker- Aberrant NAG, GGT, (32-microgobulin
expression
Oxidative Stress - Increased lipid peroxidation
(i.e., MDA production), elevated kidney antioxidant
enzymes (SOD, GPx, CAT), and depleted GSH
Morphology - Electron micrography showed Pb
damages mitochondria, basement membrane, and
brush border in kidney tissue. Some focal tubal necrosis
observed.
Kidney function - Elevated urinary total protein, urinary
albumin, and serum urea nitrogen.
Renal Function and Interstitial Fibrosis
Past studies have shown that chronic continuous or repeated Pb exposure can result in
interstitial nephritis and focal or tubular atrophy. A series of studies on Pb exposure in
rats (longitudinal 12-month exposure study to either 100 ppm or 5,000 ppm Pb in
drinking water) report an initially elevated GFR, consistent with hyperfiltration, and renal
hypertrophy (Khalil-Manesh et al.. 1993a: Khalil-Manesh et al.. 1992b: Khalil-Manesh et
al.. 1992a). After 6 months of exposure, GFR decreased, albuminuria was present, and
pathology ensued with focal tubular atrophy and interstitial fibrosis formation. This
pathology and functional decrement was persistent out to 12 months, and at 12 months
glomeruli developed focal and segmental sclerosis.
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The toxicological evidence for differences in GFR according to duration of Pb exposure
(i.e., hyperfiltration with 3-month exposure versus decreased GFR with 6- or 12-month
exposure) provides biological plausibility for epidemiological studies that observed a
similar phenomenon by age in adults in association with Pb biomarker levels. However,
the toxicological studies report blood Pb levels at some exposure durations much higher
than relevant to blood Pb levels in current human populations. Still, these duration-
dependent dichotomous changes in GFR are consistent between the toxicological and
epidemiologic literature.
At exposure concentrations resulting in blood Pb levels within one order of magnitude of
the upper range of current human population blood Pb level (< 30 (ig/dL), animal
toxicological studies present inconsistent results for the effects of Pb on kidney function
(Table 4-28). There are studies that have corroborated the previously observed increase in
serum creatinine following Pb exposure in rats. Berrahal et al. (2011) reported on the
effects of age-dependent exposure to Pb on nephrotoxicity in male rats (Table 4-29). Pups
were exposed to Pb lactationally (as a result of dams consuming water containing 50 ppm
Pb acetate) until weaning. Thereafter, the offspring were exposed to the same solution
from weaning (day 21) until sacrifice. Male pups were sacrificed at age 40 days (puberty;
blood Pb level 12.7 ug/dL) and at age 65 days (post-puberty; blood Pb level 7.5 ug/dL).
Serum creatinine was elevated at both 40 days and 65 days (0.54 and 0.60 mg/dL
compared to control values of 0.45 mg/dL [p O.001]). Various parameters of Pb-induced
renal dysfunction are listed in Table 4-29. The elevated serum creatinine in the
Pb-exposed animals compared to controls suggests that animals exposed to low dose
(i.e., 50 ppm) Pb from birth may develop renal abnormalities. However, the lack of
measurements of GFR or renal pathology weakens the conclusions.
Table 4-29 Indicators of renal damage in male rats exposed to 50 ppm Pb for 40
and 65 days, starting at parturition.
Biomarker
(Mean ± SD)
Blood Pb level (pg/dL)
Plasma Creatinine
(mg/L)
Plasma Urea (mg/L)
Plasma Uric Acid
(mg/L)
PND40 Control
1.8 ±0.33
4.5 ±0.21
0.37 ±0.019
7.51 ± 0.44
PND40 Pb
12.7 ± 1.7
5.35±0.25a
0.47±0.021a
7.65 ±0.32
PND65 Control
2.1 ±n0.35
4.55n± 0.27
0.29n± 0.009
9.39n± 0.82
PND65 Pb
7.5n± 0.78
6.04±0.29a
0.29n± 0.009
5.91n±0.53a
ap <0.001
Source: Modified with permission of John Wiley & Sons, Berrahal et al. (2011)
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Roncal et al. (2007) found that Pb accelerated renal function decrements,
tubulointerstitial injury, and arteriolopathy in non-Pb-related CKD. Sprague-Dawley rats
were administered Pb acetate at 150 ppm for 4 weeks, then subjected to remnant kidney
surgery (left kidney mass reduced by 2/3 and right kidney removed), and subsequently
exposed to Pb for an additional 12 weeks resulting in a blood Pb level of 26 (ig/dL.
Pb-treated rats had higher systolic BP, increased serum creatinine, lower creatinine
clearance, and higher proteinuria than did controls. Most striking was development of
worse arteriolar disease, peritubular capillary loss, tubulointerstitial damage, and
macrophage infiltration. Pb treatment was associated with significant worsening of pre-
glomerular vascular disease, as characterized by an increase in the media-to-lumen ratio.
There was also a higher percentage of segmental sclerosis within glomeruli and a
tendency for a higher number of sclerotic glomeruli. Additionally, a loss of peritubular
capillaries, as reflected by a reduction in thrombomodulin staining, was observed. This
was associated with worse tubular injury (osteopontin staining) due to more interstitial
fibrosis (type III collagen staining) and a greater macrophage infiltration in the
interstitium. The increase in macrophages was associated with higher renal MCP-1
mRNA. As a whole, these findings indicate that Pb exposure concomitant with existing
renal insufficiency due to surgical kidney resection accelerated vascular disease and
glomerular pathology. These findings are consistent with the previous work of Bagchi
and Preuss (2005) also showing that Pb-exposed animals with non-Pb-related CKD
(remnant surgery) had kidney dysfunction including impairment of the renin-angiotensin
system (Losartan challenge), elevated systolic BP, and alterations in renal excretion of
Pb, K+, and Na+. Thus, this model shows that Pb exposure may exacerbate pre-existing
underlying kidney disease.
Other investigators have shown that chronic Pb exposure has detrimental effects on renal
function at higher blood Pb levels. A number of studies report increased serum creatinine
following high level Pb exposure (e.g., blood Pb levels >55.6 (ig/dL) (Abdel Moneim et
al..2011b: Ozsov et al.. 2010: Javakumaretal.. 2009: Kharoubi et al.. 2008a). In
addition, studies reporting high blood Pb levels and high Pb exposure levels report
increased urine, serum, or blood urea nitrogen (Wang et al.. 2010e: El-Nekeetv et al..
2009: Javakumar et al.. 2009: Kharoubi et al.. 2008a). Jayakumar et al. (2009) reported
alterations in other markers of kidney toxicity, lysosomal marker and brush border
enzymes (i.e., ALP, ACP, y-GT, NAG, (3-D-glucuronidase), following Pb exposure
(2,000 ppm for 6 weeks). Similarly, Wang et al. (2010e) reported time-related increases
in urinary alkaline phosphatase, urinary GGT, urinary NAG, urinary total protein, urinary
(3-2 microglobulin, and urinary microalbumin following Pb exposure (300 ppm in
drinking water, serum Pb level 20 (ig/dL). Pb-exposed male rats (500 ppm Pb acetate in
drinking water for 7 months, blood Pb level 43 (ig/dL) had elevated urinary pH, urinary
glucose, and proteinuria (Navarre-Moreno et al.. 2009).
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Qiao et al. (2006) measured the effect of Pb on the expression of the renal nuclear factor-
kappa B (NF-KB), transforming growth factor (TGF-|3) and fibronectin in Sprague-
Dawley rat kidney. These growth (TGF-|3) and transcription (NF-KB) factors modulate
the progression of renal function decrements through promotion of extracellular matrix
(fibronectin) synthesis and promotion of fibrosis. Pb was administered at a dose of
5,000 ppm Pb acetate, continuously for either one, two, or three months. All factors
increased by the end of three months of treatment, but only NF-KB increased
progressively at each time period. These changes were hypothetically related to the
development of Pb-induced renal fibrosis in rats, but no histology was performed.
The renal effects of chronic Pb exposure as detailed above were partially mitigated in rats
following lowering of blood Pb level with chelation therapy (i.e., DMSA) (Khalil-
Manesh et al.. 1992a) and after treatment with antioxidants (Abdel Moneim et al.. 201 Ib:
Ozsovetal. 2010; Wang etal.. 2010e: El-Nekeetv et al.. 2009; Javakumaret al.. 2009;
Kharoubi et al.. 2008a). DMSA treatment improved renal function; however, Pb-induced
pathology remained (Khalil-Manesh et al.. 1992a). Improvements include increased GFR,
decreased albuminuria, and decreased inclusion body numbers but little change in
tubulointerstitial scarring. DMSA also acts as an antioxidant, so the protective effects
may not be entirely attributed to the lowering of blood Pb level. Similarly, several studies
found that treatment with antioxidant compounds could protect against Pb-induced
kidney dysfunction. Administration of flaxseed oil, L-carnitine, NAC, and several
medicinal plants including, Achyranthes aspera, Artemisia absinthium, and Aquilegia
vulgaris, to Pb-exposed rodents protected against injury to the kidney or restored kidney
function (Abdel Moneim et al.. 201 Ib: Ozsovetal.. 2010; Wang etal.. 2010e: EL
Nekeetv et al.. 2009: Javakumar et al.. 2009: Kharoubi et al.. 2008a). These studies
suggest that a reduction in reactive oxygen species may attenuate the effects of Pb on
kidney function implicating oxidative stress as a predominant mechanism for Pb-induced
reduced kidney function.
Histological Changes
Earlier studies discussed in previous Pb AQCDs have identified Pb-related renal damage
by the presence of dense intranuclear inclusion bodies, which are capable of sequestering
Pb (Goyeretal.. 1970b). Pb-induced formation of intranuclear inclusion bodies in the
proximal tubule (PT) is considered protective; Pb is sequestered such that it is not in its
bioavailable, free, lexicologically active form. Intranuclear inclusion bodies are found in
the kidney with short-term (i.e., <4 weeks) Pb exposure but present to a lesser degree
with chronic exposures (see Section 4.2.3 for further discussion). Chelators such as
CaNa2EDTA have removed these inclusion bodies from affected nuclei (Goyer et al..
1978).
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Multiple ultrastructural changes indicate dysfunction in the PT and nephropathy after Pb
exposure, including changes to the PT epithelium, endoplasmic reticulum dilation,
nuclear membrane blebbing, and autophagosome enlargement (Fowler et al., 1980; Goyer
et al.. 1970a). Indications similar to the PT transport-associated Fanconi syndrome appear
with Pb exposure, albeit often at high doses of Pb, i.e., Pb poisoning. These indications,
which include increased urinary electrolyte excretion (Zn), decreased Na+/K+ATPase
activity, mitochondrial aberrations, and aminoaciduria, also have been associated with
blood Pb levels in children.
Recent studies since the 2006 Pb AQCD are consistent with the earlier findings and build
upon the literature base by including the role of antioxidants. Jabeen et al. (2010) exposed
pregnant albino BALB/c mice to a daily oral dose of Pb acetate (10 mg/kg body weight,
daily throughout pregnancy) until GDIS, at which point the fetal kidneys were processed
for histological examination. Histology revealed Pb exposure induced decreased kidney
cortical thickness, decreased diameter of renal corpuscles, and increased renal tubular
atrophy (with desquamated epithelium and degenerated nuclei in the distal and proximal
tubules). Blood Pb levels were not reported in this study. Nonetheless, these data show
that in utero Pb exposure had significant histological effects on the fetal kidney, which
could contribute to altered renal function including clearance of waste products,
electrolyte balance, and vasoregulation.
Massanyi et al. (2007) reported on Pb-induced alterations in male Wistar rat kidneys after
single i.p. doses of Pb acetate (50, 25, and 12.5 mg/kg); kidneys were removed and
analyzed 48 hours after Pb administration. Qualitative microscopic analysis detected
dilated Bowman's capsules and dilated blood vessels in the interstitium with evident
hemorrhagic alterations. Quantitative histomorphometric analysis revealed increased
relative volume of interstitium and increased relative volume of tubules in the
experimental groups. The diameter of renal corpuscles and the diameter of glomeruli and
Bowman's capsule were significantly increased. Measurement of tubular diameter
showed dilatation of the tubule with a significant decrease of the height of tubular
epithelium compatible with degenerative renal alterations. These findings extend the
observations of Fowler et al. (1980) and Khalil-Manesh et al. (1992b; 1992a); in
particular, the enlarged glomeruli are consistent with the early hyperfiltration caused by
Pb.
Abdel Moneim et al. (20 lib) reported histological evidence of inflammation after Pb
treatment in rats (i.p. 20 ppm, 5 days). This evidence included increased inflammatory
cellular infiltrations, cytoplasmic vacuolation, and dilatation of some kidney tubules.
Inflammation was accompanied by an increase in apoptotic cells and increased oxidative
stress.
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A recent study has also reported inclusion body formation in the nuclei, cytoplasm, and
mitochondria of PT cells of Pb-treated rats (50 mg Pb/kg bw, i.p., every 48 hours for
14 days) (Navarre-Moreno et al., 2009). These inclusion bodies were not observed in
chronically Pb-exposed rats (500 ppm Pb in drinking water, 7 months). However, chronic
Pb exposure resulted in morphological alterations including loss of PT apical membrane
brush border, collapse and closure of the PT lumen, and formation of abnormal
intercellular junctions.
Vogetseder et al. (2008) examined the proliferative capacity of the renal PT (particularly
the S3 segment) following i.v. administration of Pb to juvenile and adult male Wistar
rats. Proliferation induction was examined by detection of Bromo-2'-deoxyuridine
(BrdU), Ki-67 (labels S, G2, and M phase cells), and cyclin Dl (an essential cell cycle
progression protein). The cycling marker Ki-67 revealed a much higher proliferation rate
in the S3 segment in control juvenile rats (4.8 ± 0.3%) compared with control adult rats
(0.4 ± 0.1%). Pb administration (3.8 mg/100 g bw) increased the proportion of Ki-67-
positive cells to 26.1 ± 0.3% in juvenile rats and 31.9 ± 0.3% in adult rats. Thus, the
increased proliferation caused by Pb was age independent. The proliferation induction
caused by Pb administration may be a result of reduced cell cycle inhibition by p27klp"1.
Acute Pb treatment increased the incidence of cyclin D1 labeling in the BrdU-positive
cells suggesting Pb was able to accelerate re-entry of cells into the cell cycle and cause
proliferation in the PT. Pb-induced cellular proliferation has also been reported in the
retina with gestational and early postnatal rodent Pb exposure (Giddabasappa et al..
2011).
Ademuyiwa et al. (2009) examined Pb-induced phospholipidosis and cholesterogenesis in
rat tissues. Sprague-Dawley rats were exposed to 200, 300, and 400 ppm Pb acetate for
12 weeks. The Pb exposure resulted in induction of phospholipidosis in kidney tissue,
accompanied by depletion of renal cholesterol. The authors suggested that induction of
cholesterogenesis and phospholipidosis in kidney may be responsible for some of the
subtle and insidious cellular effects found with Pb-mediated nephrotoxicity. Drug-
induced PT phospholipidosis is seen clinically with use of the potentially nephrotoxic
aminoglycoside drugs, including gentamicin (Baronas et al.. 2007).
Various antioxidants have been shown to attenuate Pb-induced histopathological changes
to the kidney. Ozsoy et al. (2010) found L-carnitine to be protective in a model of
experimental Pb toxicity in female rats. Markers of histopathological change in the
kidney, including tubule dilatation, degeneration, necrosis, and interstitial inflammation
were rescued by L-carnitine treatment in females. Male rats exposed to Pb (2,000 ppm for
6 weeks) also displayed tubular damage, whereas concomitant treatment with Pb and an
extract of Achyranthes aspera ameliorated the observed damage (Javakumar et al.. 2009).
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El-Sokkary et al. (2005) reported Pb-induced (100 ppm, s.c. for 30 days) tubular
degeneration with necrotic cells that could be prevented with melatonin treatment.
Melatonin is known to be an efficacious free radical scavenger and indirect antioxidant.
El-Nekeety et al. (2009) found an extract of the folk medicine plant Aquilegia vulgaris to
be protective against Pb acetate-induced tubular dilatation, vacuolar and cloudy epithelial
cell lining, interstitial inflammatory cell infiltration, hemorrhage, cellular debris, and
glomerulus hypercellularity. Concomitant exposure to Pb and extract produced histology
indiscernible from that in controls. Post treatment with extract partially rescued the
Pb-induced histopathology. El-Neweshy and El-Sayed (2011) studied the influence of
vitamin C supplementation (20 mg/kg pretreatment every other day) on histopathological
alterations in Pb-exposed male rats (20 mg/kg by intragastric feeding once daily for
60 days). Control rats showed normal histology, while Pb-treated rats exhibited
karyomegaly with eosinophilic intranuclear inclusion bodies in the epithelial cells of the
proximal tubules. Glomerular damage and tubular necrosis with invading inflammatory
cells were also found. Rats treated with Pb acetate plus vitamin C exhibited relatively
mild or no karyomegaly with eosinophilic intranuclear inclusion bodies in the proximal
tubules. Normal glomeruli were noted in animals exposed to Pb and vitamin C. These
findings consistently show that some antioxidants are capable of preventing or rescuing
Pb-induced renal histopathological changes, suggesting a role for oxidative stress in the
development of Pb-induced nephropathy.
Table 4-30 presents the acute and chronic renal effects of Pb exposure observed in recent
and past animal toxicology studies.
Table 4-30 Effects of Pb on the kidney/renal system related to exposure
duration: Evidence from animal (rodent) toxicology studies.
Effects with less than 3 months of exposure Effects with 6 or 12 months of exposure
Mitochondrial dysfunction Mitochondrial dysfunction
Renal cell apoptosis Renal cell apoptosis
Nuclear inclusion body formation Oxidant redox imbalance
Proximal tubule cytomegaly Altered NO homeostasis
Glomerular hypertrophy ATPase dysfunction
Increased GFR Aminoaciduria
Increased electrolyte excretion
Elevated blood pressure
Decreased GFR
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4.5.3 Modes of Action for Pb-lnduced Nephrotoxicity
4.5.3.1 Oxidative Damage
A role for ROS in the pathogenesis of experimental Pb-induced hypertension and renal
disease has been well characterized (Vaziri. 2008a. b; Vaziri and Khan. 2007). The
production of oxidative stress following Pb exposure is detailed in respect to modes of
action of Pb (Section 4.2.4). Past studies have shown that Pb treatment (single or three
daily i.p. injections) can elevate kidney GST levels, affecting glutathione metabolism
(Daggett et al.. 1998; Moseretal.. 1995; Oberlev et al.. 1995).
Animal studies continue to provide evidence for increased oxidative stress playing a role
in the pathogenesis of Pb-induced renal toxicity. Increased ROS, serum NO, and renal
NO were observed after Pb injections in rats (i.p. 20 mg/kg, 5 days) (Abdel Moneim et
al.. 20lib). Pb exposure to rat proximal tubular cells (0.5-1 (iM) also increased ROS
production, in a concentration-dependent manner (Wang et al.. 20 lib). Increased lipid
peroxidation (i.e., MDA) was demonstrated in serum and renal tissue after Pb exposure
(Abdel Moneim et al.. 201 Ib: Lodi etal.. 2011; Wang etal.. 20lib). Berrahal et al.
(2011) reported increased MDA in the kidney of Pb-exposed (50 ppm Pb acetate pre- and
post-natally) rats relative to controls at both 40 (puberty; blood Pb 12.7 ug/dL) and 65
(post-puberty; blood Pb 7.5 ug/dL) days of age. In addition, total sulfhydryl groups were
significantly decreased at 65 days. These increases in oxidative stress were accompanied
by age-dependent Pb nephrotoxicity in male rats.
Alterations in endogenous antioxidants and antioxidant enzymes that may lead to
oxidative stress have also been reported after Pb exposure. Pb treatment decreased the
activity of the renal antioxidant enzymes, CAT, SOD, GST, GPx, and GR (Abdel
Moneim et al.. 201 Ib) and protein levels of CAT and GSH (Lodi et al.. 2011).
Additionally, proteomic analysis of high-level Pb treated (1,500 ppm, 5 weeks; resulting
in blood Pb level of 53.4 (ig/dL) rat kidney identified decreased abundance of a rate-
limiting enzyme in the synthesis of GSH (glutamate cysteine ligase) (Chen etal.. 20 lib).
Conterato et al. (2007) examined the effect of Pb acetate on the cytosolic thioredoxin
reductase activity and oxidative stress parameters in rat kidneys. A single injection of
Pb acetate consisted of a single i.p. injection of 25 or 50 mg/kg Pb acetate, while repeated
injections consisted of one daily i.p. injection of Pb acetate (5 or 25 mg/kg) for 30 days.
Measured were thioredoxin reductase-1, a selenoprotein involved in many cellular redox
processes, SOD, 5-ALAD, GST, GPx, non protein thiol groups (NPSH), CAT, as well as
plasma creatinine, uric acid, and inorganic phosphate levels. The single injection at the
25 mg Pb dose level resulted in increased SOD and thioredoxin reductase-1 activity,
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while the 50 mg dose level increased CAT activity and inhibited 5-ALAD activity in the
kidney. Repeated injections at the 5 mg dose level of Pb inhibited 5-ALAD and increased
GST, NPSH, CAT, and thioredoxin reductase-1. Repeated injections at the 25-mg dose
level reduced 5-ALAD but increased GST, NPSH, and plasma uric acid levels. No
changes were observed in TEARS, GPx, creatinine or inorganic phosphate levels after
either single or repeated injection dosing. As both dosing regimens increased thioredoxin
reductase-1 activity, the authors suggest that this enzyme may be a sensitive indicator of
renal changes with low dose Pb treatment.
Jurczuk et al. (2006) published a study of the involvement of some low molecular weight
thiols in the peroxidative mechanisms of action of Pb in the rat kidney. Wistar rats were
fed a diet containing 500 ppm Pb acetate for a period of 12 weeks and were compared to
a control group receiving distilled water for the same time period. GSH, metallothionein
(MT), total and nonprotein SH groups (TSH and NPSH) were measured, as were the
blood activity and urinary concentration of 5-ALA. The concentrations of GSH and
NPSH were decreased by Pb administration, while MT concentration was unchanged.
5-ALAD in blood was decreased, whereas urinary 5-ALA was increased by Pb
administration. Negative correlations were found between the kidney GSH concentrations
and previously reported concentrations of Pb and MDA in kidneys of these rats. It is
apparent from graphical presentation of the data that GSH was reduced by more than
50% following Pb administration, while TSH was reduced by approximately 15%. No
values for either blood or kidney Pb levels or kidney MDA were reported in this article.
In 2007, the same authors (Jurczuk et al.. 2007) reported on the renal concentrations of
the antioxidants, vitamins C and E, in the kidneys of the same Pb-treated and control rats.
Exposure to Pb significantly decreased vitamin E concentration by 13% and vitamin C
concentration by 26%. The kidney concentration of vitamin C negatively correlated with
MDA concentration. The authors concluded that vitamins E and C were involved in the
mechanism of peroxidative action of Pb in the kidney, and their protective effect may be
related to scavenging of free radicals.
Studies have used antioxidant compounds to investigate the role of oxidative stress in
Pb-induced nephrotoxicity. Abdel Moneim et al. (20lib) reported that flaxseed oil
treatment protected rats from Pb-induced (i.p. 20 mg/kg, 5 days) oxidative stress,
inflammation, and apoptosis. However, the flaxseed oil also decreased the accumulation
of Pb in renal tissue making it difficult to ascertain whether the protection was due to
decreased oxidative stress or to altered Pb uptake kinetics.
El-Neweshy and El-Sayed (2011) studied the influence of vitamin C supplementation on
Pb-induced histopathological alterations in male rats. Rats were given Pb acetate,
20 mg/kg by intragastric feeding once daily for 60 days. Control rats were given 15 mg of
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sodium acetate per kg once daily, and an additional group was given Pb acetate plus
vitamin C (20 mg/kg every other day) 30 minutes before Pb feeding. Control rats showed
normal histology, while Pb-treated rats exhibited karyomegaly with eosinophilic
intranuclear inclusion bodies in the epithelial cells of the proximal tubules. Glomerular
damage and tubular necrosis with invading inflammatory cells were also seen in
Pb-treated animals. Among rats treated with Pb acetate plus vitamin C, five exhibited
relatively mild karyomegaly and eosinophilic intranuclear inclusion bodies of proximal
tubules and an additional five rats were normal. Normal glomeruli were noted in all.
Thus, vitamin C was shown to ameliorate the renal histopathological effects of Pb
intoxication, however no measures of Pb accumulation were provided to clarify the
mechanism of action of vitamin C.
Masso-Gonzalez and Antonio-Garcia (2009) studied the protective effect of natural
antioxidants (Zn, vitamin A, vitamin C, vitamin E, and vitamin B6) against Pb-induced
damage during pregnancy and lactation in rat pups. At weaning, pups were sacrificed and
kidneys were analyzed. Pb-exposed pups had decreased body weights. Blood Pb levels
were 1.43 ug/dL in the control group, 22.8 ug/dL in the Pb group, 21.2 ug/dL in the
Pb plus Zn plus vitamins group, and 0.98 ug/dL in the Zn plus vitamin group. The kidney
TEARS were significantly elevated in Pb exposed pups, while treatment with vitamins
and Zn returned TEARS to control levels. Kidney CAT activity was significantly
increased above control with Pb treatment; however supplementation with Zn and
vitamins reduced CAT activity toward normal. Pb exposure inhibited kidney Mn-
dependent SOD but not Cu-Zn-dependent SOD activity. Thus, supplementation with Zn
and vitamins during gestation and lactation was effective in attenuating the redox
imbalance induced by developmental, chronic low-level Pb exposure.
Bravo et al. (2007) reported further that mycophenolate mofetil (an immunosuppressive
agent used in renal transplantation which inhibits T and B cell proliferation)
administration reduced renal inflammation, oxidative stress and hypertension in
Pb-exposed rats. Thus, an inflammatory immune and oxidative stress component can be
seen as contributing to Pb-induced renal effects and hypertension.
Although the majority of studies of the effects of Pb exposure have been conducted in
male rats, a couple of studies have compared the response of male rats with female rats
(Sobekova et al.. 2009; Alghazal et al.. 2008a). Sobekova et al. (2009) contrasted the
activity response to Pb on the antioxidant enzymes, GPx and GR, and on TEARS in both
male and female Wistar rats of equal age. Males weighing 412 ± 47 g and females
weighing 290 ± 19 g were fed diets containing either 100 ppm or 1,000 ppm Pb acetate
for 18 weeks. In the male rats, kidney Pb content increased by 492% on the 100 ppm Pb
diet and by 7,000% on the 1,000 ppm Pb diet. In the female rats, kidney Pb content
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increased by 410% on the 100 ppm Pb diet and by 23,000% on the 1,000 ppm Pb diet.
There was virtually no change in GPx in the kidney of male rats given the 100 ppm Pb
diet but there was a significant reduction in GPx in the female rats on both the 100 ppm
diet and 1,000 ppm diet. In male rats, GR was increased from 182 units/gram of protein
in control kidneys to 220 units on the 100 ppm Pb diet and 350 units on the 1,000 ppm
diet. In female rats, kidney GR decreased from 242 units in control animals to 164 units
in animals on the 100 ppm Pb diet and 190 units in animals on the 1,000 ppm diet. In
male rats, kidney TEARS content increased from 7.5 units/gram protein to 10.0 units
(1,000 ppm Pb diet group). In female rats, there was a reduction in TEARS from 14.4
units per gram protein to 10.0 units in rats on the 100 ppm Pb diet and to 11 units in rats
on the 1,000 ppm Pb diet.
Alghazal et al. (2008a) compared the activity responses of the antioxidant enzyme, SOD
and the detoxifying enzyme, GST, of the same rats exposed to 100 ppm or 1,000 ppm
Pb acetate for 18 weeks. Similar to the previous study, kidney TEARS were increased
only in male rats given the higher dose of Pb. Kidney SOD activity, on the other hand,
was increased in both males and females at the higher dose of Pb, while GST activity was
increased in kidney of males at the higher dose of Pb and decreased at the lower dose, but
was decreased at both doses of Pb in females. Thus there were significant differences in
the responses of male and female rats to Pb exposure. Differences may be accounted for
in part due to the greater deposition of Pb in female rat kidneys. Another explanation,
offered by the authors, is that male rats are known to metabolize some foreign
compounds faster than do females, so the biological half-life of xenobiotics in the
females may be longer.
4.5.3.2 Hypertension and Alteration of Renal Vasculature and
Reactivity
As discussed in Section 4.5.1. changes in renal vasculature function or induction of
hypertension can contribute to further renal dysfunction. Pb exposure increases BP,
resulting in hypertension, through the promotion of oxidative stress and altered vascular
reactivity (Section 4.4). Antioxidants attenuated Pb-related oxidative/nitrosative stress in
the kidney and abrogated the Pb-induced increased BP (Vaziri et al.. 1999a). Chronic
increases in vascular pressure can contribute to glomerular and renal vasculature injury,
which can lead to progressive renal dysfunction and kidney failure. In this manner,
Pb-induced hypertension has been noted as one contributer to Pb-induced renal disease.
Also, Pb has been shown to act on known vasomodulating systems in the kidney. In the
kidney, two vascular tone mediators, NO and ET-1, are found to be affected by Pb
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exposure. Administration of the vasoconstrictor endothelin-1 (ET-1) affected mean
arterial pressure (MAP) and decreased GFR (Novak and Banks. 1995). Acute high-dose
Pb exposure (24 nmol/min, for 15 or 30 minutes) completely blocked this ET-1-mediated
GFR decrease but had no effect on MAP. Depletion of the endogenous antioxidant
glutathione using the drug buthionine sulfoximine, a GSH synthase inhibitor, increased
BP and increased kidney nitrotyrosine formation without Pb exposure, demonstrating the
importance of GSH in maintenance of BP (Vaziri et al., 2000). Multiple studies have
shown that Pb exposure depletes GSH stores. Catecholamines are vascular moderators
that are also affected by Pb exposure (Carmignani et al., 2000). The effect on BP with Pb
exposure is especially relevant to the kidney because it is both a target of Pb deposition
and a mitigator of BP. These earlier data detail the interaction of known modulators of
vascular tone with Pb.
The renin-angiotensin-aldosterone system (RAAS) plays an important role in kidney
homeostasis and alteration of this pathway may affect renal function. A discussion of the
RAAS can also be found in Section 4.4.2.3 in relation to the role of RAAS in Pb-induced
hypertension. A meta-analysis found that Pb exposure (resulting in blood Pb levels:
30-40 (ig/dL) increased plasma renin activity and renal tissue renin in young but not old
rats (Vander. 1988). Exposure of experimental animals to Pb also induced increases in
plasma, aorta, heart, and kidney angiotensin converting enzyme (ACE) activity; plasma
kininase II, kininase I, and kallikrein activities; and renal Angll positive cells (Rodriguez-
Iturbeetal.. 2005; Sharifi et al.. 2004; Carmignani et al.. 1999). Simoes et al. (2011)
reported that acute Pb treatment (Pb acetate i.v. bolus dose of 320 (ig/kg bw, blood Pb of
37 (ig/dL at 120 minutes after Pb administration) in adult male Wistar rats increased
serum angiotensin converting enzyme (ACE) activity. Systolic arterial pressure, but not
diastolic arterial pressure or heart rate, was also elevated 60 minutes after treatment. The
Pb-induced altered systolic BP attenuated in animals co-treated with Losartan (Ang II
receptor blocker) or Enalapril (ACE inhibitor), suggesting a regulatory role for the RAAS
(Simdes etal.. 2011). These data agree with earlier reports of Pb-related increases in
ACE activity in young rats exposed to Pb for 2-8 weeks (Sharifi et al.. 2004) and adult
rats exposed to Pb for 10 months (Carmignani et al.. 1999). Lower level Pb (100 ppm,
14 weeks; range of blood Pb levels: 23.7-27 (ig/dL) exposure increased renal cortical
Angll content and the number of tubulointerstitial Angll-positive cells (Bravo et al..
2007). This heightened intrarenal angiotensin corresponded with sodium retention and
increased systolic BP and was ablated by the anti-inflammatory antioxidant, MMF.
Recently, Vargas-Robles et al. (2007) examined the effect of Pb exposure (100 ppm
Pb acetate for 12 weeks) on BP and angiotensin II vasoconstriction in isolated perfused
kidney and interlobar arteries. Vascular reactivity was evaluated in the presence and
absence of the nitric oxide synthase inhibitor L-NAME in both Pb-treated and control
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animals. Pb exposure significantly increased BP (134 ± 3 versus 100 ± 6 mmHg), eNOS
protein expression, oxidative stress, and vascular reactivity to angiotensin II. L-NAME
potentiated the vascular response to angiotensin II in the control group, but had no effect
on the Pb-treated group. Conversely, passive microvessel distensibility, measured after
deactivation of myogenic tone by papaverine, was significantly lower in the arteries of
Pb-exposed rats. Nitrites released from the kidney under the influence of angiotensin II in
the Pb group were lower as compared to the control group whereas 3-nitrotyrosine was
higher in the Pb group. The authors concluded that Pb exposure increases vascular tone
through nitric oxide-dependent and -independent mechanisms, increasing renal vascular
sensitivity to vasoconstrictors.
4.5.3.3 Apoptosis and/or Ischemic Necrosis of Tubules and
Glomeruli
Apoptosis or programmed cell death in excess can cause cell atrophy while an
insufficiency can lead to uncontrolled cell proliferation, such as cancer. Pb exposure has
been shown to cause morphological changes to the kidney structure. Some of these
Pb-induced changes are a result of cellular apoptosis or necrosis. Past studies have shown
Pb-induced necrosis in proximal tubule cells (Fowler et al.. 1980). Pb-induced apoptosis
is known to act through the mitochondria (Rana. 2008). Pb-induced calcium overload
may depolarize the mitochondria, resulting in cytochrome c release, caspase activation,
and apoptosis. The apoptosis is mediated by Bax translocation to the mitochondria and
can be blocked by overexpression of Bcl-xl. Also, Pb-induced ALA accumulation can
generate ROS, which may damage DNA leading to apoptosis.
Mitochondria are targets of Pb toxicity and often involved in apoptosis. Pb can induce
uncoupling of oxidative phosphorylation, decreased substrate utilization, and
modification of mitochondrial ion transport. ATP energetics are affected when ATP-Pb
chelates are formed and ATPase activity is decreased. ROS formation can contribute to
these mitochondrial changes and to other changes within the kidney. Antioxidant
supplementation after Pb exposure can remedy some changes. All of these outcomes, in
conjunction with Pb-related depletion of antioxidants (e.g., GSH) and elevation of lipid
peroxidation point to possible susceptibility of the kidney to apoptosis or necrosis.
Rodriguez-Iturbe et al. (2005) reported that chronic exposure to low doses of Pb
(100 ppm in drinking water for 14 weeks) results in renal infiltration of immune cells,
apoptosis, NF-KB activation and overexpression of tubulointerstitial Ang(II). Similarly,
higher level Pb treatment in rats (i.p. 20 mg/kg, 5 days) induced inflammatory cellular
infiltrations and an increase in apoptotic cells, accompanied by more pronounced BAX
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staining in kidney tubule epithelial cells (Abdel Moneim et al., 201 Ib). Pb treatment
(0.5-1 (iM) of isolated rat proximal tubular cells increased cell death by apoptosis and
necrosis in a concentration- and time-dependent manner (Wang et al., 201 Ib). This was
accompanied by increased morphological changes typical of apoptosis such as
fragmented chromatin, condensed chromatin, and shrunken nuclei. These cells also
exhibited decreased mitochondrial membrane potential, decreased intracellular pH,
inhibition of Na+/K+ATPase and Ca2+ATPase activity, and increased intracellular Ca2+
following Pb treatment.
Navarro-Moreno et al. (2009) examined the effect of 500 ppm Pb in drinking water over
7 months on the structure (including intercellular junctions), function, and biochemical
properties of PT cells of Wistar rats. Pb effects in epithelial cells consisted of an early
loss of the apical microvilli, followed by a decrement of the luminal space and the
respective apposition and proximity of apical membranes, resulting in the formation of
atypical intercellular contacts and adhesion structures. Inclusion bodies were found in
nuclei, cytoplasm, and mitochondria. Lipid peroxidation (TEARS measurement) was
increased in the Pb-treated animals as compared to controls. Calcium uptake was
diminished and neither proline nor serine incorporation that was present in controls was
noted in the PT of Pb-exposed animals. The authors speculated that Pb may compete with
calcium in the establishment and maintenance of intercellular junctions.
Tubular necrosis was also observed in rats treated with Pb acetate (100 ppm, s.c.) for
30 days (El-Sokkary et al.. 2005). Histological sections of kidneys from Pb-treated rats
showed tubular degeneration with some necrotic cells. Similarly, El-Neweshy and
El-Sayed (2011) reported glomerular damage and tubular necrosis with invading
inflammatory cells after Pb treatment (20 mg/kg by intragastric feeding once daily for
60 days) to male rats. The incidence of necrosis was decreased in both of these studies by
pretreatment with either melatonin or vitamin C. Pretreatment with melatonin
(10 mg/kg), an efficacious free radical scavenger and indirect antioxidant, resulted in a
near normal tubular structure. The authors concluded that melatonin protected the liver
and kidneys from the damaging effects of exposure to Pb through inhibition of lipid
peroxidation and stimulation of endogenous antioxidative defense systems (El-Sokkary et
al.. 2005). Vitamin C supplementation (20 mg/kg pretreatment every other day) protected
the renal architecture and histology (El-Neweshy and El-Saved. 2011).
Wang et al. (2009c) examined the effect of Pb acetate (0.25, 0.5 and 1 (iM) on cell death
in cultured rat primary PT cells. A progressive loss in cell viability, due to both apoptosis
and necrosis, was observed in cells exposed to Pb. Apoptosis predominated and could be
ameliorated with concomitant N-acetylcysteine exposure, whereas necrosis was
unaffected. Elevation of ROS levels and intercellular calcium, depletion of mitochondrial
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membrane potential, and intracellular glutathione levels was observed during Pb
exposure. Pb-induced apoptosis was demonstrated morphologically (Hoechst 33258
staining) with condensed/fragmented chromatin and apoptotic body formation. CAT and
SOD activities were significantly elevated, reflecting the response to accumulation of
ROS.
4.5.3.4 Renal Gangliosides
Gangliosides are constituents of the plasma membrane that are important for control of
renal GFR because they can act as receptors for various molecules and have been shown
to take part in cell-cell interactions, cell adhesion, cellular recognition, and signal
transduction. Aguilar et al. (2008) studied changes in renal gangliosides following Pb
exposure (600 ppm Pb acetate in drinking water for 4 months) in adult male Wistar rats.
Pb exposure caused an increase in blood Pb from 2.1 to 35.9 (ig/dL. There was no change
in serum creatinine or in hemoglobin, but there was an increase in urinary 5-ALA. The
following renal gangliosides were measured by immunohistochemistry and by thin layer
chromatography: GM1, GM2, GM4, and 9-O-acetylated modified form of the
GD3 ganglioside (9-O-Ac-GD3). The ganglioside pattern was mainly characterized by a
decrease in the GM1 ganglioside as well as by a mild increase in GM4 and GM2
gangliosides, while the strongest alteration was observed in the 9-O-Ac-GD3, which was
overexpressed. The latter was observed only in the glomerular zone. This was associated
with a decrease in apoptotic glomerular cells, as assessed by the TUNEL assay. The
authors hypothesized that the increase in GD3-O-acetylation could represent a strategy to
attenuate the normal renal apoptotic process and therefore contribute to cell survival
during Pb exposure.
4.5.3.5 Altered Uric Acid
Higher occupational Pb exposure or blood Pb levels have been linked to increased risk
for both gout and kidney disease (Shadick et al., 2000; Batuman, 1993). Pb is thought to
alter uric acid homeostasis by affecting its kidney excretion (Emmerson and Ravenscroft.
1975; Ball and Sorensen. 1969; Emmerson. 1965). Behringer et al. (1986) suggests that
the appearance of gout in patients with renal failure points to prior Pb exposure. Colleoni
et al. (1993) note that in their uremic patients on regular hemodialysis blood Pb corrected
for hematocrit was significantly higher than blood Pb of healthy subjects living in the
same area and state that when they analyzed the possible different sources of Pb
accumulation in their population they found it to be higher in patients occupationally
exposed and in patients exposed to alkyl lead from atmospheric contamination from
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gasoline additives. Research during the past decade indicates that uric acid is nephrotoxic
at lower levels than previously recognized (Johnson et al.. 2003). Therefore, the
2006 Pb AQCD (U.S. EPA. 2006b) reviewed literature implicating increased uric acid as
a mechanism for Pb-related nephrotoxicity (Weaver et al.. 2005a: Shadick et al.. 2000).
However, this does not appear to be the only mechanism, since associations between
blood Pb and serum creatinine have remained significant even after adjustment for uric
acid (Weaver et al.. 2005a).
Alterations in serum uric acid have been studied in animal models exposed to Pb. In male
rats exposed to Pb in drinking water from lactation to puberty (40 days) or post-puberty
(65 days), Berrahal et al. (2011) found that plasma uric acid levels were unchanged after
40 days of exposure (puberty blood Pb level of 12.7 ug/dL) but decreased after 65 days of
Pb exposure (post-puberty blood Pb level of 7.5 ug/dL); plasma urea increased after
40 days of but were unchanged after 65 days of Pb exposure (Table 4-29). Serum uric
acid was increased in male rats following long-term Pb exposure (2,000 ppm for
6 weeks) (Jayakumar et al., 2009). Conterato et al. (2007) followed various parameters of
kidney function after single or multiple Pb injections in rats. The single dosing regimen
consisted of a single i.p. injection of 25 or 50 mg/kg Pb acetate, while the multiple
injections involved once daily i.p. injection of either vehicle or Pb acetate (5 or
25 mg/kg) for 30 days. Single and multiple injections at both dose levels increased
plasma uric acid levels. Similarly, Abdel Moneim et al. (20 lib) reported increased serum
uric acid and urea levels after 5 days of Pb acetate treatment (i.p. 20 mg/kg).
4.5.3.6 Role of Metallothionein
The metal-binding protein, metallothionein, may play a role in inclusion body formation
and thus block potential interaction of Pb with cellular targets. Yu et al. (2009) described
dichotomous effects of Pb acetate on the expression of MT in the liver and kidney of
mice. Male mice were i.p. injected with Pb acetate in doses of 100, 200, and 300 (imol/kg
and sacrificed 4, 8, and 24 hours after Pb treatment. Administration of Pb increased the
levels of MT-1 mRNA in the liver and kidneys but increased MT protein only in the
liver. Treatment of mouse PT cells in vitro with Pb also resulted in an increase in MT
mRNA but little increase in MT protein. Thus, Pb appears to exert a dual effect on MT
expression in the kidney: enhancement of MT gene transcription but suppression of MT
mRNA translation. Wang et al. (2009b) did not find a change in the distribution and
expression of MT-1 and MT-2 genes in the kidney of rats following long-term Pb
exposure (300 ppm in drinking water, 8 weeks, serum Pb level 10.9 (ig/dL), however did
report a synergistic effect of Pb and cadmium on metallothionein in the kidney (see
Section 4.5.4.1).
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Zuo et al. (2009) examined the potential role of a-Synuclein (Sena) and MT in
Pb-induced inclusion body formation. Unlike the parental wild type (WT) cells, MT-I/II
double knockout (MT-null) cells did not form inclusion bodies after Pb treatment;
however, transfection of MT-1 into MT-null cells allowed inclusion body formation after
Pb treatment. As inclusion bodies formed during Pb treatment, soluble MT protein in WT
cells was lost. As Sena is a protein with a natural tendency to aggregate into oligomers,
Sena was measured in WT cells and MT-null cells after Pb treatment. In both cell lines
Pb-induced Sena expression rapidly increased and then decreased over 48 hours as
Pb-induced inclusion bodies were formed. Pb exposure caused increased colocalization
of MT and Sena proteins and MT was localized to the surface of inclusion bodies in WT
renal cortex samples following Pb treatment. Thus, Sena may be a component of
Pb-induced inclusion bodies and, with MT, may play a role in inclusion body formation.
4.5.3.7 Summary
At blood Pb levels greater than 30 (ig/dL, there is clear evidence in animal toxicological
studies that Pb causes changes to the kidney morphology and function. Evidence for
functional changes in animals following lower Pb exposure levels resulting in blood Pb
levels < 20 (ig/dL is generally not available. However, human epidemiologic studies have
reported associations of increased blood Pb with elevated risk for kidney dysfunction in
human populations with a wider range of blood Pb levels. Animal toxicological and
epidemiologic studies provide a plausible biological mechanism for renal effects,
including increased BP and hypertension, increased oxidative stress and inflammation,
morphologic alterations, and increased uric acid.
As discussed in Section 4.5.1. the renal and cardiovascular systems are intimately linked
and chronic kidney disease can be a cause, as well as a consequence of hypertension. The
medical literature clearly demonstrates that renal function is inversely correlated with BP,
such that an increase in BP may be accompanied by a decrease in creatinine clearance or
increase in serum creatinine (Lindeman et al.. 1987). Pb exposure has been causally
linked to increased BP, hypertension, and other cardiovascular effects (Section 4.4).
Animal toxicology studies and epidemiologic studies consistently observe associations
between increases in BP and long-term Pb exposure, at relatively low (i.e., 2-10 (ig/dL)
concurrent blood Pb levels.
Animal studies provide evidence for concomitant vascular disease (e.g., elevated BP) and
glomerular pathology following Pb exposure (Roncal et al.. 2007; Bagchi and Preuss.
2005) providing a clear link between Pb exposure, cardiovascular disease, and renal
dysfunction. There are various mechanisms that explain the pathway from hypertension
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to kidney damage, including barotrauma to the glomerular capillaries caused by increased
capillary pressure, altered vascular reactivity and vasoconstriction, altered RAAS,
oxidative stress, and chronic tissue hypoxia (Palm and Nordquist 2011). Studies in
animals provide evidence for increased renal vascular reactivity (Vargas-Robles et al..
2007) and activation of the RAAS following Pb exposure (Bravo et al., 2007; Bagchi and
Preuss. 2005: Sharifi et al.. 2004: Carmignani et al.. 1999).
Damage to the kidneys has been closely associated with increased oxidative stress and
inflammation. ROS in the kidneys are responsible for preservation of routine cellular
physiology and are important for cell fate, regulation of blood flow, and regulation of
gene expression (Nistala et al.. 2008). However, loss of redox homeostasis and excessive
ROS contribute to proinflammatory and profibrotic pathways in the kidney. A number of
studies have characterized the role for ROS in the pathogenesis of experimental
Pb-induced renal dysfunction. For example, studies of long-term Pb exposure in animals
reported elevated TEARS at a blood Pb level of 23 (ig/dL (Masso-Gonzalez and Antonio-
Garcia. 2009) and elevated MDA levels at a blood Pb levels of 12.7 (ig/dL (Berrahal et
al.. 2011) (study details in Table 4-28). The increased MDA was accompanied by
nephrotoxicity including decreased plasma uric acid levels, increased blood creatinine,
and increased BUN. Studies have also found that treatment with antioxidant compounds
may protect against Pb-induced nephrotoxicity. ROS serve as essential mediators of
inflammatory signaling and thus oxidative stress is often accompanied by increased
inflammation. A few studies have reported increased inflammatory markers following Pb
exposure in animals, including leukocyte infiltration and increased expression of NF-KB
and MCP-1 (Bravo et al.. 2007: Roncal et al.. 2007: Rodriguez-Iturbe et al.. 2005).
Multiple ultrastructural and histological changes have been observed in the kidney
following Pb exposure, including changes to the PT epithelium, endoplasmic reticulum
dilation, nuclear membrane blebbing, autophagosome enlargement, and intranuclear
inclusion body formation. Some of these morphological changes may be the result of
cellular necrosis or apoptosis, often due to mitochondrial dysfunction, which has been
observed following Pb exposure (Navarre-Moreno et al.. 2009: Rodriguez-Iturbe et al..
2005: Fowler etal.. 1980).
Increased uric acid has also been proposed as a mechanism for Pb-related kidney damage
and may act through endothelial dysfunction, activation of RAAS, oxidative stress, and
proinflammatory and proliferative events (Filiopoulos et al.. 2012). Both increased and
decreased uric acid has been observed in animals after exposure to Pb (Berrahal et al..
2011: Roncal et al.. 2007). Epidemiologic studies provide some evidence for a positive
association between Pb and increased uric acid in certain populations (Weaver etal..
2005a: Shadick et al.. 2000).
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In summary, various mechanisms have been characterized to explain how exposure to Pb
may result in damage to the kidneys, including increased BP and hypertension, increased
oxidative stress and inflammation, morphologic alterations, and increased uric acid.
Studies have not investigated the dose-response relationship of all of these mechanisms;
however, the various mechanisms proposed appear to have the potential to occur at all
levels of exposure. This evidence supports the biological plausibility of associations
observed in epidemiologic studies between Pb levels and reduced kidney function.
4.5.4 Effects of Exposure to Pb Mixtures
The effect of Pb on other cations, specifically calcium, is well established in the kidney
literature. Calcium-mediated processes involving receptors, transport proteins, and
second messenger signaling among other endpoints have been shown to be significantly
affected by Pb exposure. The disposition of Pb in the soft tissues (kidney and spleen) can
change with exposure to Pb and other compounds. Pb plus Cd exposure changed Pb
disposition with increased blood Pb (versus Pb alone group) and decreased metal
concentration in the kidney and liver (versus Pb alone). An Fe deficient diet significantly
increased Pb deposition in adult animals (Hashmi et al.. 1989). pregnant dams, and
maternally-exposed fetuses (Singh etal. 1991). Dietary thiamine plus Zn slightly
reduced blood and kidney Pb in exposed animals (Flora etal.. 1989). Selenium (Se), a
cofactor for GPx, attenuated Pb-induced lipid peroxidation and abrogated the Pb-induced
attenuation of GR and SOD. Concomitant exposure to the cations aluminum and Pb
protected animals from ensuing nephropathy (Shakoor et al.. 2000). In summary, Pb has
been shown to affect processes mediated by endogenous divalent cations. In addition,
exposure to other metals or divalent cations can modulate Pb disposition and its effects in
the body.
4.5.4.1 Lead (Pb) and Cadmium (Cd)
Cd shares many similarities with Pb; it has been shown to be a ubiquitous PT
nephrotoxicant and accumulates in the body. Despite the similarities, few studies have
evaluated associations between Cd exposure and CKD or the impact of joint exposure of
Pb and Cd or other metals on CKD. As discussed in the 2006 Pb AQCD (U.S. EPA.
2006b). environmental exposure to Cd, at levels common in the U.S. and other developed
countries, has been shown to impact substantially associations between indicators of Pb
exposure and the kidney EBE marker, NAG, even in the presence of occupational level
Pb exposure. In an occupational study, mean NAG, although higher in the Pb-exposed
worker group compared to controls, was correlated with urine Cd but not blood or tibia
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Pb (Roels et al.. 1994). In another occupational population where both metals were
significantly associated with NAG, a 0.5 ug/g creatinine increase in Cd had the same
effect on NAG as did a 66.9 ug/g bone mineral increase in tibia Pb (Weaver et al..
2QQ3a).
The 2006 Pb AQCD noted that data examining the concentration-response relationship
between environmental Cd and the kidney were too scarce to determine the impact of Cd
exposure on the relationship between Pb exposure and other kidney outcomes. A recent
publication using NHANES data collected from 1999 through 2006 addresses this gap;
(results pertaining solely to Pb were discussed in Section 4.5.2.1) (Navas-Acien et al..
2009). Geometric mean concurrent blood Cd level was 0.41 ug/L in 14,778 adults aged
> 20 years. After adjustment for survey year, sociodemographic factors, CKD risk
factors, and blood Pb, the ORs for albuminuria (> 30 mg/g creatinine), reduced eGFR
(<60 mL/min/1.73 m2), and both albuminuria and reduced eGFR were 1.92 (95% CI:
1.53, 2.43), 1.32 (95% CI: 1.04, 1.68), and 2.91 (95% CI: 1.76, 4.81), respectively,
comparing the highest with the lowest blood Cd quartiles. Both Pb and Cd remained
significantly associated after adjustment for the other. Effect modification was not
observed; however, ORs were higher for adults in the highest quartiles of both metals,
compared with the ORs for the highest quartiles of concurrent blood Cd or Pb alone
Table 4-25). Compared with adults with blood Cd levels < 0.2 ug/L and blood Pb levels
< 1.1 ug/dL, adults with blood Cd levels >0.6 ug/L and blood Pb levels >2.4 ug/dL had
ORs (95% CIs) of 2.34 (95% CI: 1.72, 3.18) for albuminuria, 1.98 (95% CI: 1.27, 3.10)
for reduced eGFR, and 4.10 (95% CI: 1.58, 10.65) for albuminuria and reduced eGFR
together. These findings are consistent with other recent publications (Akesson et al..
2005; Hellstrom et al.. 2001). support consideration of both metals as independent CKD
risk factors in the general population, and provide novel evidence of increased risk in
those with higher environmental exposure to both metals.
However, a recent study suggests that interpretation of Cd associations with GFR
measures may be much more complex. Conducted in Pb workers to address the fact that
few studies have examined the impact of environmental Cd exposure in workers who are
occupationally exposed to other nephrotoxicants such as Pb, the study assessed Cd dose
with urine Cd, which is widely considered the optimal dose metric of cumulative Cd
exposure. In 712 Pb workers, mean (SD) blood and tibia Pb, urine Cd, and eGFR using
the MDRD equation were 23.1 (14.1) ug/dL, 26.6 (28.9) ug/g, 1-15 (0.66) ug/g
creatinine, and 97.4 (19.2) mL/min/1.73m2, respectively (Weaver et al.. 2011). After
adjustment for age, sex, BMI, urine creatinine, smoking, alcohol use, education, annual
income, diastolic BP, current or former Pb worker job status, new or returning study
participant, and blood and tibia Pb, higher urine Cd was associated with higher calculated
creatinine clearance, eGFR (P = 8.7 mL/min/1.73 m2 [95% CI: 5.4, 12.1] per unit
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increase in logarithmic (In)-transformed urine Cd) and logarithmic (In)-transformed
NAG, but lower serum creatinine. These unexpected paradoxical associations have been
reported in a few other publications (de Burbure et al., 2006; Hotz et al., 1999) and have
been observed in other populations. Potential explanations for these paradoxical results
included a normal physiologic response in which urine Cd levels reflect renal filtration;
the impact of adjustment for urine dilution with creatinine in models of kidney outcomes;
and Cd-related hyperfiltration.
Animal studies have also examined the effects from combined exposure to Pb and Cd.
Wang et al. (2009b) reported a greater than additive effect of Pb and Cd on
metallothionein expression in the kidney. Long-term Cd exposure (50 ppm in drinking
water, 8 weeks, serum Cd level 1.4 (ig/dL) increased the expression of MT-1 and MT-2
genes in the kidney of rats, whereas Pb exposure alone did not (300 ppm in drinking
water, 8 weeks, serum Pb level 10.9 (ig/dL). Combined exposure to Pb and Cd increased
metallothionein expression in the kidney above the level of either pollutant alone. Wang
et al. (2010d) studied the effects of Pb and/or Cd on oxidative damage to rat kidney
cortex mitochondria. In this study young female Sprague Dawley rats were fed for
8 weeks with either Pb acetate (300 ppm), Cd chloride (50 ppm), or Pb and Cd together
in the same dosage. Lipid peroxidation was assessed as MDA content. Renal cortex
pieces were also processed for ultrastructural analysis and for quantitative rtPCRto
identify the mitochondrial damage and to quantify the relative expression levels of
cytochrome oxidase subunits (COX-I/II/III). Cytochrome oxidase is the marker enzyme
of mitochondrial function, and COX-I, II, and III are the three largest mitochondrially-
encoded subunits which constitute the catalytic functional core of the COX holoenzyme.
Mitochondria were altered by either Pb or Cd administration, but more strikingly by Pb
plus Cd administration, as indicated by disruption and loss of mitochondrion cristae.
Kidney cortex MDA levels were increased significantly by either Pb or Cd, given
individually, but more so by Pb plus Cd. COX-I/II/III were all reduced by either Pb or Cd
administration, but more prominently by Pb plus Cd administration. This study adds to
knowledge of the combined effects of Pb and Cd on kidney mitochondria.
4.5.4.2 Lead (Pb), Cadmium (Cd), and Arsenic (As)
Wang and Fowler (2008) present a general review of the roles of biomarkers in
evaluating interactions among mixtures of Pb, Cd, and As. Past studies have found that
addition of Cd to treatment of rats with Pb or Pb and As significantly reduced the
histological signs of renal toxicity from each element alone, including swelling of the
proximal tubule cells and intranuclear inclusion body formation. On the other hand,
animals exposed to Cd in addition to Pb or Pb and As showed an additive increase in the
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urinary excretion of porphyrins, indicating that, although measured tissue burdens of Pb
were reduced, the biologically available fraction of Pb is actually increased (Mahaffev et
al.. 1981; Mahaffev and Fowler. 1977).
Stress proteins were examined after exposure to mixtures of Pb and other metals.
Induction of MT was strongest in groups with Cd treatment. However, co-exposure to Pb
and As induced higher levels of MT protein than did either Pb or As exposure alone in
kidney tubule cells. Heat shock proteins (Hsps) are commonly altered with exposure to
metal mixtures. A study found in vitro (low dose) and in vivo that Pb induced Hsps in a
metal/metalloid-, concentration- and time-specific manner (Wang et al.. 2005a). Additive
or more than additive interactions occurred among Pb, Cd, and As under combined
exposure conditions.
4.5.4.3 Lead (Pb) and Zinc (Zn)
Zinc has been investigated as a protective compound against the effects of Pb. Pb
treatment (35 mg/kg i.p. for 3 days) caused a significant fall in hemoglobin content,
significant increases in lipid peroxidation and decreased level of reduced glutathione in
liver, together with diminished total protein content in liver and kidney. Co-treatment of
Pb with Zn (10 mg/kg i.p.) or ascorbic acid (10, 20 and 30 mg/kg i.p.) showed a moderate
therapeutic effect when administered individually, but more pronounced protective
effects after combined therapy (Upadhyay et al.. 2009).
Jamieson et al. (2008) studied the effect of dietary Zn content on renal Pb deposition.
Weanling Sprague Dawley rats were assigned to marginal zinc (MZ, 8 mg Zn/kg diet),
zinc adequate control (CT, 30 mg Zn/kg), zinc-adequate diet-restricted (30 mg Zn/kg), or
supplemental zinc (SZn, 300 mg Zn/kg) groups, with or without Pb acetate (200 ppm for
3 weeks). Pb exposure did not result in nephromegaly or histological alterations. The MZ
rats had higher renal Pb (35%) and lower renal Zn (16%) concentrations than did CT rats.
On the other hand, SZn was more protective than the CT diet was against renal Pb
accumulation (33% lower). Standard procedures for indirect immunoperoxidase staining
were used to determine MT localization in the kidney. Pb had no effect on MT staining
intensity, distribution, or relative protein amounts. Western blot analysis confirmed that
MT levels were responsive to dietary Zn but not to Pb exposure.
4.5.4.4 Lead (Pb) and Mercury (Hg)
Stacchiotti et al. (2009) studied stress proteins and oxidative damage in a renal-derived
cell line exposed to inorganic Hg and Pb. The time course of the expression of several
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Hsps, glucose-regulating proteins, and MTs in a rat proximal tubular cell line (NRK-52E)
exposed to subcytotoxic doses of inorganic mercury (HgCl2, 1-40 uM) and Pb (PbCl2,
2-500 uM) were analyzed. ROS and reactive nitrogen species (RNS) were detected by
flow cytometric analysis. Endogenous total GSH content and the enzymatic activity of
GST were determined in cell homogenates. Western blot analysis and
immunohistochemistry were used for quantification of Hsps and MTs. Reverse
transcription PCR was used for quantification of metallothionein. The higher doses of Hg
(20 uM and 40 uM) were shown to markedly inhibit growth of the cell line while the
higher doses of Pb (60 uM to 500 uM) inhibited cell growth to a lesser degree. After 24
hours of exposure of 20 (iM Hg, the cells presented abnormal size and pyknotic nuclei,
swollen mitochondria and both apoptosis and overt necrosis. In the presence of 60 or
300 (iM Pb, the cells lost cell-cell and cell-matrix contacts, showed a round size, irregular
nuclear contour and often mitotic arrest, but no apoptosis or overt necrosis at 24 hours.
Mercury (Hg) induced a significant increase in both ROS and RNS, maximal RNS at
24 hours, and maximal ROS at 48 hours. Pb (60 or 300 (iM) did not cause an increase in
ROS or RNS beyond the levels measured in control cells. Total GSH significantly
increased in cells grown in the presence of Pb; the effect was concentration-dependent
and GSH reached its maximal value at a dose of 300 (iM Pb. The effect of Hg was
biphasic: 10 uM significantly enhanced GSH by 600%, while the amount of GSH
detected after 20 (iM Hg only increased by 50% compared to control levels. GST activity
was enhanced by both Pb and Hg. Hsp25 and Hsp72 were up-regulated by Hg but there
was no effect on Grp78 as compared to control. On the contrary, Pb treatment only
upregulated Grp78. Mercury (Hg) induced a time-dependent effect on MT mRNA
expression, which reached its maximal value 3 hours after beginning treatment and
reverted to control values at 24 hours. With Pb, on the other hand, mRNA transcription
was concentration- and time-dependent. The transcripts remained overexpressed
compared to controls up to 72 hours. The results of this study with regard to the Pb effect
on MT synthesis clearly differ from those of Jamieson et al. (2008). which found no
increase in MT following Pb exposure. This discrepancy remains to be clarified.
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4.5.5 Summary and Causal Determination
Evidence from epidemiologic studies of adults, including longitudinal analyses, and a
number of toxicological studies examining various exposure scenarios and lifestages
suggest Pb exposure leads to reduced kidney function. However, there is uncertainty
regarding the potential for reverse causality to explain findings in humans. As noted in
Section 1.1 most animal toxicology studies considered relevant to this assessment report
blood Pb levels up to 30 (ig/dL and studies describing the mode of action are sometimes
higher. The causal determination for reduced kidney function in adults is informed by
evidence in humans for reduced GFR, reduced creatinine clearance, and increased serum
creatinine. Biological plausibility and mode of action for these effects is provided by
evidence for hypertension, oxidative stress, inflammation, vascular reactivity and injury,
increased uric acid, morphological changes, and apoptosis or necrosis. The section that
follows describes the evaluation of evidence for reduced kidney function, with respect to
causal relationships with Pb exposure using the framework described in Table II of the
Preamble. The key evidence as it relates to the causal framework is summarized in Table
4-30.
4.5.5.1 Evidence for Reduced Kidney Function
The 2006 Pb AQCD (U.S. EPA. 2006b) concluded that "in the general population, both
circulating and cumulative Pb was found to be associated with a longitudinal decline in
renal function," evidenced by increased serum creatinine and decreased creatinine
clearance or eGFR over the follow-up of 4 to 15 years in association with higher baseline
blood and bone Pb levels (U.S. EPA. 2006b). Data in general and patient populations of
adults provided evidence of Pb-associated lower renal function in populations with mean
concurrent or baseline blood Pb levels of 2-10 ug/dL (Akesson et al., 2005; Tsaih et al.,
2004; Yu et al.. 2004; Kim et al.. 1996); associations with lower eGFR were observed in
adults with hypertension with a mean concurrent blood Pb level of 4.2 ug/dL (Muntner et
al.. 2003). Since these effects on the kidney were often observed in adults with likely
higher past Pb exposures, uncertainty exists as to the Pb exposure level, timing,
frequency, and duration contributing to the associations observed with blood or bone Pb
levels. The conclusion from the 2006 Pb AQCD was substantiated by the coherence of
effects observed across epidemiologic and toxicological studies. However, a number of
the animal toxicological studies were conducted at Pb exposure concentrations that
resulted in blood Pb levels higher than what is relevant to the general U.S. adult human
population. Both human and animal studies observed Pb-associated hyperfiltration. In
rodents during the first 3 months after Pb exposure, effects were characterized by
increased GFR and increased kidney weight due to glomerular hypertrophy. However,
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exposure for 6 or 12 months resulted in decreased GFR, interstitial fibrosis, and kidney
dysfunction. Additionally, toxicological studies found that early effects of Pb on tubular
cells were generally reversible, but continued exposure resulted in chronic irreversible
damage. Toxicological studies provided mechanistic evidence to support the biological
plausibility of Pb-induced renal effects, including oxidative stress leading to NO
inactivation. Despite the strong body of evidence presented in the 2006 Pb AQCD,
uncertainty remained on the contribution of past higher Pb exposures to associations
observed in adults, the impact in children, the implication of hyperfiltration, and reverse
causality.
Prospective epidemiologic studies in adult men in the general population (Tsaih et al..
2004; Kim et al.. 1996) support the relationship between long-term Pb exposure and
reduced kidney function at exposure levels considered relevant to this assessment
(^ 30 (ig/dL; see Section 1.1). These prospective population studies are high quality
community epidemiologic studies that show a longitudinal association after adjustment
for key potential confounding factors. As presented in Table 4-25 the slopes and 95% CI
limits per 1 (ig/dL increase in blood Pb level are similar in magnitude for the entire
population assessed by Tsaih et al. (2004) ((3 = 0.0210 [-0.0019, 0.048]) and for the group
of individuals with a peak blood Pb level ^ 10 (ig/dL assessed by Kim et al. (1996)
((3 = 0.021 [-0.005, 0.048]). In addition, these studies show an overall pattern of elevated
risks across all subgroups analyzed within the population.
Cross-sectional studies of the general adult population add support to the associations
observed in prospective epidemiologic studies (Section 4.5.2.1). The majority of cross-
sectional studies report associations between higher measures of Pb exposure and worse
renal function. These studies include analyses conducted from 1988-1994 (Muntner et al..
2003). 1999-2002 (Muntner etal.. 2005) and from 1999-2006 (Navas-Acien et al.. 2009)
of the NHANES cohort which provides a representative U.S. population sample that may
be generalizable to the total U.S. population. Re-examination of a study from the
2006 Pb AQCD (U.S. EPA. 2006b) provided data to conclude that in a population with
likely higher past exposures to Pb, a 10-fold increase in concurrent blood Pb was
associated with an 18 mL/min decrease in estimated creatinine clearance or a 25%
decrease from the mean, and that an increase in blood Pb from the 5th to the 95th
percentile (3.5 ug/dL) had the same negative impact on eGFR as did an increase of
4.7 years in age or 7 kg/m2 in body mass index (Akesson et al.. 2005). Overall, a
relationship between higher Pb exposure and various indicators of lower kidney function
is indicated by a set of high-quality covariate-adjusted cross-sectional studies that are
conducted with different designs in different cohorts by different researchers.
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A prospective study in CKD patients indicated that higher baseline blood Pb level was
associated with greater CKD progression overtime (i.e., reduced eGFR) (Yu et al..
2004). The use of eGFR provides a better estimate of progressive changes of renal
function than creatinine clearance used in the other related studies. Additionally, body Pb
burden assessed by EDTA chelatable Pb level was significantly associated with overall
risk for a doubling of serum creatinine over the 4-year study period or with the need for
hemodialysis.
Other studies in clinical trials of CKD patients treated with EDTA chelation provide
supportive results; however, uncertainty exists due to limitations noted in the studies.
Effects on renal function have also been observed in studies of CKD patient studies (Lin
et al., 2006b; Lin et al., 2006a; Lin et al., 2003). It is possible, however, that the CKD
patient studies, in which blood Pb level was lowered compared to control groups by
chelation, demonstrate an improvement in renal function by reducing ROS, blood Pb
level or both. The uncertainty related to this may reflect an involvement of both lowering
of blood Pb levels and a reduction of ROS following chelation as both are possible to a
varying extent. Other potential limitations in these studies include small sample size and
lack of researcher blinding. The treatment protocol, which included additional calcium
disodium EDTA infusions as needed based on Pb levels during follow-up, may differ
from the placebo protocol; therefore, patients may have been able to discern their
treatment group.
Research on the renal effects of Pb in the occupational setting has traditionally been far
less consistent than that in environmentally Pb-exposed populations (Section 4.5.2.1).
Limitations of the occupational evidence, including the potential for hyperfiltration, the
healthy worker effect, small sample sizes, and limited control for covariates may explain
this inconsistency. The actual cause of paradoxical or inverse associations (higher Pb
dose with lower serum creatinine, and/or higher eGFR or calculated or measured
creatinine clearance) in several of these studies is not understood. Overall, the recent
occupational studies on the kidney impact of Pb exposure has been relatively consistent
in reporting statistically significant associations than were data reviewed for the
2006 Pb AQCD, possibly reflecting increased reliance on EBE markers as more sensitive
outcome measures, publication bias, or multiple comparisons due to a greater number of
outcomes assessed.
Combined evidence from prospective and cross-sectional studies helps limit the level of
uncertainty for bias from confounding with reasonable confidence. While the adjustment
for specific factors varied by study, the collective body of evidence adjusted for multiple
potential key confounding factors, including age, sex, BMI, comorbid cardiovascular
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conditions and markers (e.g., BP, lipid levels), smoking and alcohol use, SES, medication
use, and copollutant exposure.
At blood Pb levels greater than 30 (ig/dL, there is clear evidence in animal toxicological
studies that long-term Pb exposure causes changes to the kidney morphology and
function (Khalil-Manesh et al.. 1992b: Khalil-Manesh et al.. 1992a). Evidence for
functional changes in animals following lower Pb exposure levels resulting in blood Pb
levels < 20 (ig/dL is generally not available. Yet at blood Pb levels between 20 and
30 (ig/dL studies with various exposure scenarios and in various lifestages provide some
evidence for dysfunction in kidney function measures (e.g., decreased creatinine
clearance, increased serum creatinine, increased BUN).
Animal toxicological and epidemiologic studies provide several plausible biological
mechanisms for Pb-induced renal effects, including increased BP and hypertension,
increased oxidative stress and inflammation, morphologic alterations, and increased uric
acid (Section 4.5.3). As kidney dysfunction can increase BP and increased BP can lead to
further damage to the kidneys, Pb-induced damage to either or both renal or
cardiovascular systems may result in a cycle of further increased severity of disease. Pb
exposure has been causally linked to both increased BP and other cardiovascular effects
(Section 4.4). Notably, animal studies provide evidence for concomitant vascular disease
and glomerular pathology following Pb exposure (Roncal et al.. 2007; Bagchi and Preuss.
2005) providing a clear link between Pb exposure, cardiovascular disease, and renal
dysfunction. Damage to the kidneys has been closely associated with increased oxidative
stress and inflammation. A number of studies have characterized the role for ROS in the
pathogenesis of experimental Pb-induced renal dysfunction (Berrahal et al.. 2011; Masso-
Gonzalez and Antonio-Garcia. 2009) and have reported increased inflammatory markers
following Pb exposure in animals (Bravo et al.. 2007; Roncal et al.. 2007; Rodriguez-
Iturbe et al.. 2005). Morphological changes resulting from cellular apoptosis or necrosis
and altered uric acid have also been characterized as mechanisms for Pb-induced kidney
damage. As noted in Section 1.1. animal toxicology studies considered relevant to this
assessment report blood Pb levels as high as 30 (ig/dL and studies that describe MOA are
sometimes higher. Lower concentration Pb exposures and lower blood Pb levels in
animals have not been widely examined.
Several important uncertainties limit the confidence in the relationship between Pb
exposure and reduced kidney function. First, because these studies report effects most
often observed in adults with likely higher past Pb exposures, uncertainty exists as to the
Pb exposure level, timing, frequency, and duration contributing to the associations
observed with blood or bone Pb levels. Second, the potential for a bidirectional
relationship because of reverse causality is plausible in observational epidemiologic
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studies (see Section 4.5.2.4). While several lines of evidence support that reverse
causality does not contribute substantially to associations between higher blood Pb levels
and worse kidney function, reverse causation remains a plausible hypothesis. The
existence of an association in adults with normal renal function does not exclude reverse
causation because it is not known how Pb clearance varies within the range of normal
kidney function. Thus, this bidirectional relationship is possible and additional evidence
is needed to fully elucidate the extent to which diminished kidney function may itself
result in increased Pb levels.
There are multiple high quality epidemiologic studies and clear biological plausibility;
however, uncertainty remains regarding the potential for reverse causality to explain
findings in humans. Among epidemiologic studies, key evidence is provided by the few
available longitudinal studies which better characterized the temporal sequence between
Pb exposure and changes in renal function by showing associations between baseline
bone or blood Pb levels and reduced kidney function over time in men in the Boston, MA
area and greater progression of kidney disease in CKD patients. Evidence from
longitudinal studies also addressed the potential for reverse causality by showing the
persistence of blood Pb-associated decrements in kidney function in the range of normal
kidney function and demonstrated stronger associations than cross-sectional analyses of
the same data. Cross-sectional adult studies provide supportive evidence but are weighed
less than the prospective studies in conclusions because by design, they do not inform
directionality (i.e., reverse causality). Inconsistencies were noted in occupational studies
and studies of children, and important study design limitations were noted in clinical
trials of chelation in CKD patients. These inconsistencies and limitations preclude strong
inferences from the results of these three study groups. Longitudinal studies found Pb-
associated decrements in renal function in populations with mean blood Pb levels of 7
and 9 (ig/dL. However, the contributions of higher past Pb exposures cannot be excluded.
Animal toxicological studies provide clear biological plausibility with evidence for Pb-
induced kidney dysfunction at blood Pb levels greater than 30 (ig/dL; however, evidence
in animals with blood Pb levels < 20 (ig/dL is generally not available. Still at blood Pb
levels between 20 and 30 (ig/dL studies provide some evidence for dysfunction in kidney
function measures (e.g., decreased creatinine clearance, increased serum creatinine,
increased BUN). Animal studies also provide biological plausibility for the associations
observed between blood Pb levels and reduced kidney function with evidence for
Pb-induced hypertension, renal oxidative stress and inflammation, morphological
changes, and increased uric acid. Collectively, the evidence integrated across
epidemiologic and toxicological studies, with uncertainties related to the potential for
reverse causation, is suggestive of a causal relationship between Pb exposures and
reduced kidney function among adults.
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Table 4-31 Summary of evidence supporting renal causal determinations.
Attribute in Causal
Framework3
Key Evidence
References
Pb Exposure or
Biomarker Levels
Associated with Effects0
Reduced Kidney Function - Suggestive
Consistent
associations from
multiple, high
quality
epidemiologic
studies with
relevant blood Pb
levels
Longitudinal epidemiologic analyses in
nonoccupationally Pb exposed adults
for associations with change in serum
creatinine and GFR in MAS
Support from prospective analyses of
CKD patients and Korean workers
Supporting cross-sectional evidence
for associations between concurrent
blood Pb level and serum creatinine,
Tsaih (2004),
Kim (1996)
Yu (2004).
Akesson (2005).
Staessen (1992),
Payton (1994)
Median baseline blood Pb
levels:
Adults, 6.5 - 8.6 ug/dl_d
Mean baseline blood Pb
level: 4.2 ug/dL
Median concurrent blood
Pb levels:
Adults, 2.2 -11.4 ug/dl_
creatinine clearance, and GFR
applying differing designs across
multiple cohorts of adults in different
locations.
Associations found with adjustment for
potential confounding factors including
age, pre-existing cardiovascular
disease, baseline kidney function, and
Studies had population based
recruitment (NHANES) with high
follow-up participation
Section 4.5.2.1
Muntner (2003).
Navas-Acien (2009)
Mean concurrent blood
Pb levels:
Adults, 1.58- 4.2 ug/dl_
Important Uncertainty related to reverse
uncertainties causality; the bidirectional association
remain is possible
Uncertainty related to exposure
patterns, with likely higher past Pb
exposure
Uncertainty due to baseline serum
creatinine adjustment
Section 4.5.2.4
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Table 4-31 (Continued): Summary of evidence supporting renal causal determinations.
Attribute in Causal
Framework3
Key Evidence
References
Pb Exposure or
Biomarker Levels
Associated with Effects0
Limited evidence in
animals of reduced
kidney function
after Pb exposure
Limited evidence in animals report
decreased creatinine clearance,
increased serum creatinine, and
decreased GFR at both relevant and
high level long-term Pb exposures.
Decreased function:
Berrahal et al. (2011)
Roncal et al. (2007)
Khalil-Manesh et
al.(1992a)
Mean blood Pb levels:
Rodents
>7.5 ug/dL (<65 days
from birth)
26 ug/dL (12 weeks as
adults with pre-existing
kidney disease),
-30 ug/dL (6 months as
adults)
Hyperfiltration:
Khalil-Manesh et al.
(1993a)
29 ug/dL (3 months as
adults)
No functional change:
Vyskocil et al. (1995)
Ademuyiwa et al.
(2009)
38 ug/dL (4 months as
adults)
41 ug/dL (12 weeks as
adults)
Evidence clearly
describes mode of
action
Hypertension
Oxidative Stress
Inflammation
Consistent evidence for increased BP
and hypertension following Pb
exposure in humans and animals at
relevant Pb levels across numerous
studies with adjustment for potential
confounding factors.
Association of increased BP with
reduced kidney function has been well
documented
Consistent evidence for increased
ROS, enhanced lipid peroxidation, and
antioxidant enzyme disruption in Pb
exposed animals.
Lymphocyte and macrophage
infiltration
Increased NF-KB and MCP-1
expression
Section 4.5.3.7
Sections 4.4 and
4.5.3.2
Table 4-24
(Lindeman et al..
1987)
Section 4.5.3.1
Sections 4.5.2.5
and 4.2.5
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Table 4-31 (Continued): Summary of evidence supporting renal causal determinations.
Attribute in Causal
Framework3
Key Evidence
References
Pb Exposure or
Biomarker Levels
Associated with Effects0
Evidence clearly
describes mode of
action (continued)
Morphology
Increased
Uric Acid
Glomerular hypertrophy
Cellular apoptosis and necrosis
leading to PT damage
Evidence of increased and decreased
plasma uric acid in animals.
Limited evidence in humans
Sections 4.5.2.5 and
4.5.3.3
Section 4.5.3.5
"Described in detail in Table II of the Preamble.
bDescribes the key evidence and references contributing most heavily to causal determination and where applicable to uncertainties
and inconsistencies. References to earlier sections indicate where full body of evidence is described.
""Describes the blood Pb levels in humans with which the evidence is substantiated and blood Pb levels in animals most relevant to
this ISA.
dBecause blood Pb level in nonoccupationally-exposed adults reflects both recent and past Pb exposures, the magnitude, timing,
frequency, and duration of Pb exposure contributing to the observed associations is uncertain.
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4.6 Immune System Effects
4.6.1 Introduction
With respect to studies conducted in laboratory animal and in vitro models, the immune
effects of Pb exposure have been extensively examined over several decades. Animal
studies of the effects of Pb exposure on host resistance date back to the 1960s while those
examining Pb-induced immune functional alterations, including developmental
immunotoxicity, were first conducted during the 1970s. Despite this long history of
research, Pb-associated immune effects in animals with blood Pb levels considered
relevant to this ISA (i.e., < 30 (ig/dL, Section 4.1). particularly early in life, have been
observed mostly within the last 10-15 years (Dietert and McCabe. 2007). Recent findings
of Pb-associated changes in immunological parameters in humans have increased
understanding of the immune effects of environmental exposure to Pb.
The pathways by which Pb exposure potentially may alter immune cell function and
consequently increase the risk of immune-related diseases, are presented in Figure 4-33.
Rather than producing overt cytotoxicity, Pb exposure has been associated with
functional alterations in cellular and humoral immunity. In the 2006 Pb AQCD (U.S.
EPA. 2006b). the hallmarks reported for Pb-induced changes in immune functional
pathways were: (1) suppression of T cell-derived helper (Th)l-mediated immunity
(i.e., suppressed Thl cytokine production and delayed type hypersensitivity [DTH]
response); (2) stimulation of Th2 immunity (i.e., increased production of Th2 cytokines
and immunoglobulin (Ig)E antibody); and (3) altered macrophage function. The latter
was characterized by increased production of reactive oxygen species (ROS),
prostaglandin E2 (PGE2), and inflammatory cytokines such as tumor necrosis
factor-alpha (TNF-a) and interleukin (IL)-6 and decreased production of nitric oxide
(NO). Changes in immune cells can alter cell-to-cell interactions, multiple signaling
pathways, and inflammation, that in turn, can influence the risk of developing infectious,
allergic, and autoimmune diseases and exacerbate inflammatory responses in other organ
systems. Studies conducted in animal and in vitro models provided consistent evidence
for Pb exposure inducing effects on the range of immune effects presented in this
continuum. In the much smaller epidemiologic evidence base, most studies examined
Pb-exposed male workers and a limited range of immune-related endpoints.
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Reduced:
Phagocytosis
Nitric oxide production
Peroxynitrite production
Lysosomal activity
Macrophages and
Other Innate Immune Cells
Elevated IL-4 and IL-5
Suppressed INF-y
Elevated IL-10
Suppressed IL-12
Elevated TNF-a
Overproduction of ROS
Depleted antioxidant
efenses
Increased lipid
and DNA oxidation
in tissues
Skewed
Th2-biased
responses
Removal of normal
myelomonocytic
suppression
Increased tissue
inflammation
(e.g. lung, gut, skin)
Damaged epithelia
and mucosal barriers
Increased tumor
cell formation
Tissue damage
and de novo
antigen
appearance
Inappropriate
Tcell proliferation
activation
IncreasedIgE
production
Increased risk of
other inflammatory
diseases
Increased risk of
Suppressed
Thl- mediated
anti-tumor
host defense
Increased risk of
later-life cancer
Increased risk of atopic and
inflammatory diseases
Note: Conditions shown in boxes at the bottom are evaluated in detail in this section. As shown in the figure, immunological
pathways may increase risk of diseases such as cancer (Section 4.10) and inflammatory diseases in the cardiovascular
(Section 4.4). renal (Section 4.5). and hepatic (Section 4.9.1) systems. These effects are evaluated explicitly in other sections, as
specified.
Figure 4-33 Immunological pathways by which Pb exposure potentially may
increase risk of immune-related diseases.
Reflecting suppressed Thl activity, toxicological evidence presented in the
2006 Pb AQCD linked Pb exposure of animals to impaired host resistance to viruses and
bacteria (U.S. EPA. 2006b). Indicating a hyperinflammatory state and local tissue
damage, a few available toxicological studies found Pb exposure-induced generation of
auto-antibodies, suggesting an elevated risk of autoimmune reactions. Additionally, the
shift toward Th2 responses suggested that Pb could elevate the risk of atopic and
inflammatory responses. While the biological plausibility of such effects was supported
by toxicological evidence for Pb-induced increases in Th2 cytokines, IgE, and
inflammation, epidemiologic evidence was too sparse to draw conclusions about the
effects of Pb exposure on these broader indicators of immune dysfunction in humans.
However, in concordance with toxicological evidence, a shift to a Th2 phenotype in
humans was indicated in the few available studies by associations observed between
higher concurrent blood Pb level and higher serum IgE levels in children. Because of lack
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of examination, the immune effects of Pb exposure in adults without occupational
exposures were not well characterized.
Changes in the spectrum of immune endpoints were found in association with a wide
range of blood Pb levels in animals. Juvenile and adult animals with blood Pb levels in
the range of 7-100 (ig/dL were found to have suppressed DTH, elevated IgE, and changes
in cytokine levels. Most epidemiologic studies examined and found lower T cell
abundance and higher serum IgE levels in association with population mean (or group)
concurrent blood Pb levels >10 (ig/dL.
With respect to important lifestages of Pb exposure, animal studies provided strong
evidence for immune effects in juvenile animals induced by prenatal Pb exposures and in
adult animals by postnatal exposures. There was uncertainty regarding important
lifestages of Pb exposure in humans as epidemiologic studies of children primarily were
cross-sectional and examined concurrent blood Pb levels. Several other limitations of
epidemiologic studies were noted, including small sample sizes; little consideration for
potential confounding factors such as age, sex, smoking, SES indicators, and allergen
exposures; and comparisons of immune endpoints among groups with different blood Pb
levels that provided little information on the concentration-response relationship.
Collectively, the small numbers of toxicological and epidemiologic studies published
since the 2006 Pb AQCD, supported the previous findings of Pb-associated immune
effects. Epidemiologic studies supported previous findings in children and provided new
evidence for effects in nonoccupationally-exposed adults. Recent studies also expanded
on the array of immunological parameters potentially affected by Pb exposure as
presented in Figure 4-33. For example, a recent toxicological study indicated that Pb may
modulate the function of dendritic cells. Results from recent toxicological and
epidemiologic studies supported the link between Pb-associated effects on immune cells
and immune- and inflammatory-related diseases by providing evidence for changes in
intermediary signaling and inflammatory pathways (Figure 4-33). Several recent
epidemiologic studies examined signaling molecules such as pro-inflammatory cytokines
to provide coherence with toxicological findings. Recent toxicological studies further
supported the broader role of Pb-associated immune modulation in mediating Pb effects
in nonlymphoid tissues (e.g., nervous, reproductive, respiratory systems). Recent
epidemiologic studies improved on the design of previous studies through greater
examination of children and adults with blood Pb levels more comparable to current
levels in the U.S. population and greater consideration for confounding by age, sex,
smoking, SES indicators, and allergen exposures. This epidemiologic evidence
particularly that from prospective studies, combined with the extensive toxicological
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evidence formed the basis of conclusions about the immune effects of Pb exposure
(Section 4.6.8).
4.6.2 Cell-Mediated Immunity
4.6.2.1 T Cells
T cells have a central role in cell-mediated immunity, influencing maturation of B cells,
activation of cytotoxic T cells and macrophages, and interactions with antigen presenting
cells (APCs). A majority of the evidence for the effects of Pb exposure on T cells was
provided by toxicological and epidemiologic studies reviewed in the 2006 Pb AQCD
(U.S. EPA. 2006b). Toxicological evidence consistently links Pb exposure with
alterations in T cells with observations of Pb-induced shifts in the partitioning of CD4+
(T helper) cell populations to favor Th2 cells in vitro (25-50 (iM Pb chloride, 3-5 days)
(Heoetal.. 1998; 1996). proliferation of Th2 cells over Thl cells in vitro (10, 100 (iM
Pb chloride, 7 days) (McCabe and Lawrence. 1991). and the production of Th2 cytokines
over Thl cytokines in vitro and in vivo (wide range of Pb concentrations,
Section 4.6.6.1). Epidemiologic findings are limited largely to associations observed
between higher concurrent blood Pb level and lower T cell abundance in children.
In vitro results indicated various mechanisms by which Pb exposure may induce a shift to
Th2 responses including activation of transcription factor NF-KB (regulates T cell
activation) in cultures of human CD4+ T cells (1 (iM Pb acetate, 30 minutes) (Pvatt et al..
1996) and a concentration-dependent (10, 50 (iM Pb chloride, 24 hours) increased
expression of MHC class II surface antigens (e.g., HLA-DR), which mediate the CD4+
response to exogenous antigens (Guo etal.. 1996b). The few available recent
toxicological studies described T cell-dependent and -independent pathways. Kasten-
Jolly et al. (2010) provided evidence in vivo and with dietary Pb exposures relevant to
this ISA. While results were based on amicroarray analysis of hundreds of genes, which
is subject to higher probability of chance findings, they were supported by other lines of
evidence. In this study, gestational-lactational Pb acetate exposure of BALB/c mice
(100 (iM in drinking water of dams GD8-PND21, resultant spleen homogenate Pb level
13.8 (ig/g) altered splenic cell gene expression of cytokines well documented in the
literature to be affected by Pb, including the Th2 cytokine IL-4 and the Thl cytokine
interferon-gamma (IFN-y). These changes occurred with increases in adenylate cyclase 8
and phosphatidylinositol 3-kinase in the absence of signaling molecules STAT4 or
STAT6, which comprise the preferential signaling pathway for T cells. Similarly, in
cultures of stimulated mouse T cells, Heo et al. (2007) showed that Pb chloride (25
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12-24 hours) decreased the IFN-y to IL-4 ratio (indicating a shift to Th2) in the absence
of STAT6. Additionally, Pb blocked production of IFN-y not by affecting gene
expression but by suppressing translation of the protein. This blockage was rescued with
the addition of IL-12, which is a T cell stimulating factor. The STAT results indicated a T
cell-independent pathway to skewing toward Th2 responses whereas the IL-12 results
pointed to a T cell-dependent pathway.
While a few available recent epidemiologic studies found associations of blood Pb levels
with Thl and Th2 cytokines in humans (Section 4.6.6.1). the extant evidence for effects
on T cells in humans is derived largely from previous cross-sectional studies describing
differences in the abundance of several T cell subtypes that mediate acquired immunity
responses to antigens. Most studies of children found that higher blood Pb levels were
associated with lower T cell abundance in serum, primarily CD3+ cells. These
associations were observed in studies that adjusted for some potential confounding
factors (as described below) (Karmaus et al.. 2005; Sarasua et al.. 2000) and studies
without consideration for potential confounding (Zhao et al.. 2004; Lutz et al.. 1999). In
children, blood Pb level was less consistently associated with abundance of other T cell
subtypes such as CD4+ (helper T) or CD8+ (cytotoxic T). A few studies did not provide
evidence of blood Pb-associated decreases in T cell abundance but did not consider
potential confounding (Hegazy etal., 2011; Belles-Isles et al., 2002).
Associations between blood Pb level and T cell abundance were found in studies that
generally had population-based subject recruitment. Most studies did not report sufficient
information to assess the potential for biased participation by Pb exposure and immune
conditions. Most studies had multiple comparisons; however, associations were not
isolated to T cell abundance. Most studies found lower T cell abundance in groups of
U.S. and non-U.S. children (ages 6 months-10 years, N = 73-331) with concurrent blood
Pb levels >10 (ig/dL (Zhao et al.. 2004; Sarasua etal.. 2000; Lutzetal.. 1999).
Associations were inconsistent in comparisons of children with lower blood Pb levels.
Among 331 children in Germany living near (15 km) and distant from industrial facilities
(some associated with organochlorine pollution), Karmaus et al. (2005) found that
children (ages 7-10 years) with concurrent blood Pb levels 2.2-2.8 (ig/dL (2nd quartile)
had a 9 to 11% lower abundance of several T cell subtypes (for some subtypes, p <0.05,
t-test) compared with children with blood Pb levels <2.2 (ig/dL (lowest quartile). The
detection limit for blood Pb was 0.9 (ig/dL. This study recruited children from schools
and examined multiple exposures, reducing the likelihood of biased participation by
specifically by Pb exposure. Compared with other studies of T cells, Karmaus et al.
(2005) had greater consideration for potential confounding, adjusting for sex, age,
number of infections in the previous 12 months, number of cigarettes/day smoked in the
home in the previous 12 months, serum lipids, and blood organochlorine levels. But, SES
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was not examined. Cord blood levels of organochlorine and Hg but not Pb were
associated with T cell abundance in 96 newborns from a subsistence fishing community
and an urban center in Quebec, Canada with a population mean cord blood Pb level
<2 (ig/dL (Belles-Isles et al.. 2002).
Another study of children from multiple unspecified U.S. locations with and without
mining and smelting operations that considered confounding also found an association
between higher concurrent blood Pb level and lower T cell abundance, albeit limited to
the youngest subjects (Sarasua et al., 2000). Among 241 children ages 6-35 months, a
1 (ig/dL higher blood Pb level was associated with a 0.18% lower (95% CI: -0.34, -0.02)
CD3+ cell count, a 0.10% lower (95% CI: -0.24, 0.04) CD4+ cell count, and a 0.04%
lower (95% CI: -0.15, 0.07) CD8+ cell count, with adjustment for location of residence,
age, and sex. In older age groups (36-71 months, 6-15 years), many effect estimates were
positive. Analysis of blood Pb level categories indicated that associations were influenced
by lower T cell abundance (3-6%) among children ages 6-35 months with concurrent
blood Pb levels >15 (ig/dL. Notably, 76% of subjects lived near a Pb smelting operation,
were likely to have higher blood Pb levels, and may have largely influenced the observed
associations. Neither Karmaus et al. (2005) nor Sarasua et al. (2000) found a monotonic
decrease in T cell abundance across blood Pb level groups. Neither of these studies
adjusted for SES, which has been associated with blood Pb levels and immune-related
conditions such as asthma, allergy, and respiratory infections.
In the few studies with nonoccupationally-exposed adults (U.S, Italy), higher concurrent
blood Pb levels were associated with higher T cell abundance (Boscolo et al.. 2000;
Sarasua et al.. 2000; Boscolo etal.. 1999). the functional relevance of which is unclear.
These studies included healthy subjects and those with allergies, a wide range of samples
sizes (17-433), ages (16-75 years), and mean blood Pb levels (4.3-11.4 (ig/dL).
Pb-exposed workers in the U.S. and Asia did not consistently have lower or higher
abundance of various T cell subtypes than unexposed controls (Mishra et al.. 2010;
Pinkerton et al.. 1998; Yiicesoy et al.. 1997b; UndegeretaL 1996; Fischbein et al..
1993). The inconsistency among studies was not related to differences in sample size
(20-145), age (means: 22-58 years), or blood Pb levels (14.6-132 (ig/dL) among Pb
workers. None of the studies of adults considered potential confounding factors,
including other workplace exposures in occupational studies.
In summary, toxicological studies provided clear evidence for the effects of Pb exposure
on T cells by demonstrating Pb-induced expansion of Th2 cells and increased Th2
cytokine production. Providing mechanistic evidence, a few recent toxicological studies
found that Pb-induced Th2 skewing may occur via T cell-dependent (Heo et al.. 2007)
and -independent pathways (Kasten-Jolly et al.. 2010; Heo et al.. 2007). The most
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consistent epidemiologic findings were associations between higher concurrent blood Pb
level (>10 (ig/dL) and lower T cell abundance observed in children ages 6 months to 10
years (Karmaus etal.. 2005; Zhao et al.. 2004; Sarasua et al.. 2000; Lutzetal. 1999). An
association was found with lower blood Pb levels, i.e., <3 (ig/dL (Karmaus et al.. 2005).
albeit in children ages 7-10 years who likely had higher earlier childhood blood Pb levels.
The uncertainties regarding blood Pb levels and the timing and duration of Pb exposure
contributing to the associations with T cell abundance apply to the evidence as a whole
since concurrent blood Pb levels in children also reflect past Pb exposures. There are
several other uncertainties in this evidence, including the temporal sequence between Pb
exposure and T cell changes, potential selection bias, the concentration-response
relationship, and potential confounding by factors such as SES and other environmental
exposures. The implications of Pb-associated changes in T cell abundance in children are
limited further by the uncertain functional relevance of small magnitudes of change in T
cell abundance (1-9% lower CD3+ abundance in groups of children with higher blood Pb
levels) to downstream immune responses. Because toxicological studies examined effects
related specifically to Thl or Th2 responses, which have greater implications for
downstream immune effects, toxicological evidence is the major consideration in
drawing conclusions about the effects of Pb exposure on T cells.
4.6.2.2 Lymphocyte Activation
Lymphocytes (T, B, and natural killer [NK] cells) are activated by reversing the normal
suppression mediated by macrophage-like cells. Their activation is an indicator of
response to antigens. A majority of data on the effects of Pb exposure on lymphocyte
activation is provided by toxicological studies reviewed in the 2006 Pb AQCD that
showed both mitogen-induced expansion and suppression of alloreactive B and T
lymphocytes proliferation with Pb exposures in vivo and in vitro (U.S. EPA. 2006b).
Adding to the mixed nature of evidence, a recent study found that 4-week oral exposure
of 7 week-old Wistar rats to 200 ppm Pb acetate induced proliferation of lymphocytes
within the thymus and submaxillary lymph nodes, primarily by affecting B cells (Teijon
et al.. 2010). Total T cell proliferation did not change or was suppressed. Specific T cell
subtypes, CD4+, CD8+ (decreased), CD4-CD8- (elevated) were affected only with i.p.
Pb dosing (p <0.05) but not oral exposure. Using the local lymph node assay, Carey et al.
(2006) found that Pb chloride increased antigen-induced (ovalbumin, OVA) T cell
proliferation in adult female BALB/c mice but administered Pb via injection (25-50 (ig).
The mechanistic basis for Pb effects on lymphocyte activation is not well characterized.
As discussed in Section 4.6.6.2. changes in NO production appear to be involved (Farrer
et al.. 2008). Gao et al. (2007) described a potential role for dendritic cells. Dendritic
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cells that matured in the presence of 25 (iM Pb chloride enhanced alloreactive T cell
proliferation in vitro compared to control dendritic cells.
A few available cross-sectional epidemiologic studies in children and nonoccupationally-
exposed adults provided indirect evidence for Pb-associated lymphocyte activation.
Instead of directly measuring lymphocyte proliferation, these studies measured the
abundance of cells that expressed HLA-DR, a cell surface marker that indicates both
activated lymphocytes and monocytes. These studies had little consideration for potential
confounding, which also limits the implications of findings. In the study of children (ages
9 months-6 years, Missouri), the mean percentage of HLA-DR+ cells was about 2-fold
higher (p >0.05, Kruskal-Wallis) in the 19 children with concurrent blood Pb levels
15-19 (ig/dL than in children with blood Pb levels 10-14 (ig/dL (n = 61) or <10 (ig/dL
(n = 178) with adjustment for age (Lutz etal. 1999). However, activated cells were not
elevated in 16 children with blood Pb levels 20-44 (ig/dL. Small studies of adults without
occupational Pb exposure in Italy found that concurrent blood Pb level was correlated
positively with the percentage of HLA-DR expressing cells in men with and without
allergies (Spearman r = 0.51, p <0.002, n=17 each, overall median blood Pb level:
11 (ig/dL, ages 19-52 years) (Boscolo etal.. 1999) but only in women without allergies
(Spearman r = 0.44, p <0.05, n=25, median blood Pb level: 5.5 (ig/dL, ages 19-49 years)
(Boscolo et al., 2000). Associations also were found with other metals.
Comparisons of Pb-exposed workers and unexposed controls indicated similar levels of
lymphocyte proliferation (< 1% difference) (n = 10-33, mean age: 32-40 years, blood Pb
level range: 12-80 ng/dL) (Queiroz et al.. 1994b: Cohen etal.. 1989) or lower
lymphocyte proliferation (8-25%) among Pb-exposed workers (n = 15-39, mean age:
30-49 years, mean blood Pb level: 14.6-129 (ig/dL) (Mishra et al.. 2003; Fischbein et al..
1993; Alomran and Shleamoon. 1988; Kimberetal.. 1986). In the combined
epidemiologic and toxicological evidence, Pb was associated with both activation and
suppression of lymphocyte activation. The epidemiologic studies had little consideration
for potential confounding factors, including other workplace exposures in occupational
studies, and inconsistency among studies could not be explained by differences in sample
size, age of subjects, or blood Pb level either. None of the studies provided concentration-
response information. Toxicological studies have demonstrated the selective expansion of
Th2 cells and suppression of Thl cells (Section 4.6.2.1). Therefore, the differential
activation of specific subtypes may not be discernible in studies that measure overall
lymphocyte proliferation.
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4.6.2.3 Delayed-type Hypersensitivity
Although not widely examined recently, several toxicological studies reviewed in the
2006 Pb AQCD (U.S. EPA. 2006b) and recent reviews (Mishra. 2009; Dietert and
McCabe. 2007) identified a suppressed DTH response as one of the most consistently
observed immune effects of Pb exposure in animal models. A recent study indicated that
this effect may be mediated by dendritic cells. The DTH assay commonly is used to
assess the T cell-mediated response to antigens, i.e., induration and erythema resulting
from T cell activation and recruitment of monocytes to the site of antigen deposition. The
DTH response is largely Thl dependent in that Thl cytokines induce the production of T
cells specifically directed against the antigen (sensitization) and recruitment of antigen-
specific T cells and monocytes to the site of antigen deposition (elicitation phase).
Previous studies demonstrated suppressed DTH in animals after gestational (Chen et al.,
2004: Bunnetal. 200la: 2001b. c; Lee et al.. 200Ib: Chenetal.. 1999: Miller et al..
1998: Faith etal.. 1979) and postnatal (McCabe et al.. 1999: Laschi-Loquerie et al.. 1984:
Muller et al.. 1977) Pb acetate exposures. Such observations were made in F344 and CD
rats, BALB/c and Swiss mice, and chickens. Most studies exposed animals to Pb in
drinking water and found suppressed DTH in animals with blood Pb levels relevant to
this ISA (means: 6.75, 25 (ig/dL) (Chen et al.. 2004: Bunn etal.. 200 la) and higher (51 to
>100 (ig/dL) (Bunnetal.. 200 Ib: Chenetal.. 1999: McCabe et al.. 1999). The
associations of DTH with lower blood Pb levels occurred with gestational Pb exposure.
In some studies that examined Pb exposures at multiple stages of gestation, exposures
later in gestation suppressed DTH in animals (Bunnetal.. 2001c: Lee etal.. 200 Ib).
These latter findings may reflect the status of thymus and T cell development. A recent
study contributed to the robust evidence base by indicating a role for dendritic cells in the
Pb-induced suppression of the DTH response. Gao et al. (2007) exposed bone marrow-
derived dendritic cells in vitro to Pb chloride (25 (iM, 10 days) then the antigen OVA and
injected the cells into Pb-nai've adult mice. Mice treated with Pb-exposed dendritic cells
had a diminished OVA-specific DTH footpad response compared with mice treated with
dendritic cells not exposed to Pb.
Evidence indicates Pb-induced suppression of DTH in animals, some with blood Pb
levels relevant to this ISA (6.75, 25 (ig/dL) produced by gestational Pb exposure via
drinking water of dams. The mode of action is strongly supported by observations that Pb
suppresses production of the Thl cytokine IFN-y (Section 4.6.6.1). IFN-y is the primary
cytokine that stimulates recruitment of macrophages, a key component of the DTH
response. In animal studies that also examined IFN-y, the suppressed DTH response was
accompanied by a decreased production of IFN-y (Lee etal.. 2001b: Chen et al.. 1999).
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Observations of a concomitant decrease in IFN-y strengthen the link between Pb-induced
inhibition of Thl functional activities and suppression of the DTH response.
4.6.2.4 Macrophages and Monocytes
As reported in the 2006 Pb AQCD, based on a large body of toxicological evidence and
some supporting epidemiologic evidence, Pb-induced alteration of macrophage function
was considered to be a hallmark of Pb-associated immune effects (U.S. EPA. 2006b).
Macrophages, which are produced by the differentiation of blood monocytes in tissues,
mediate host defense through their role in phagocytosing pathogens and stimulating other
immune cells. Pb exposure was found to induce macrophages into a hyperinflammatory
phenotype as indicated by enhanced production of TNF-a, IL-6, and ROS and increased
metabolism of arachidonic acid into PGE2. Observations in macrophages of Pb-induced
enhanced production of ROS, suppressed production of NO, impaired growth and
differentiation, and potentially altered receptor expression [e.g., toll-like receptors])
provided coherence for the effects of Pb observed on tissue damage and diminished host
resistance in animals. Several of these findings are described in detail in Sections 4.6.6.2
and 4.6.6.3. Because macrophages are major resident populations in most tissues and
organs and also are highly mobile in response to microbial signals and tissue alterations,
their functional impairment in response to Pb exposure may serve as a link between
Pb-induced immune effects and impaired host defense, tissue integrity, and organ
homeostasis in numerous physiological systems.
Some rodent studies indicated reduced macrophage generation or phagocytosis with
gestational or postnatal dietary Pb acetate exposure that produced blood Pb levels (upon
cessation of exposure) relevant to this ISA, i.e., 8.2 (ig/dL in F344 rats (Bunn et al.,
200Ic) and 18 (ig/dL in CBA/J mice (Kowolenko et al.. 1991). Similarly, Knowles and
Donaldson (1997) found that Pb acetate trihydrate given PND1 to PND21 induced a
decrease in macrophage phagocytosis in turkeys but with higher blood Pb levels,
42 (ig/dL. In one set of experiments, CBA/J mice exposed to Pb acetate in drinking water
for 2 weeks, with blood Pb levels of 18 (ig/dL, had reduced macrophage generation
(Kowolenko et al., 1991) in response to Listeria monocytogenes inoculation (via tail vein
i.v. injection), but no change in macrophage phagocytosis (Kowolenko et al.. 1988).
Other animal studies administered Pb through routes that may not be directly relevant to
exposure routes of humans. Effects such as decreased macrophage yield, viability,
phagocytosis, chemotaxis, and killing ability were reported in Swiss mice following
bacterial infection and Pb treatment by oral gavage (40 mg/kg Pb nitrate, 40 days) (Lodi
et al.. 2011) or injection (10 mg/kg, i.p., 15 days) (Bishayi and Sengupta. 2006). Lee et
al. (2002) found no change in monocyte abundance in 5-6 week-old chickens treated with
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200 (ig Pb acetate via the air sac in ovo at embryonic day 5 or 12. Some (Bussolaro et al..
2008; Zhouetal.. 1985) but not all in vitro studies (De Guise et al.. 2000) also found
Pb-induced (0.2-1,000 (iM Pb chloride or Pb nitrate) reduced phagocytosis. In particular,
Bussolaro et al. (2008) found such effects with a relatively low concentration of Pb
exposure (0.2 (iM Pb nitrate, 72 hours).
The effects of Pb exposure on macrophages in humans have not been widely examined.
Pineda-Zavaleta et al. (2004) was unique in examining the hyperinflammatory state
specifically in macrophages, and consistent with the large body of toxicological studies,
found associations of higher concurrent blood Pb level with lower NO release and higher
superoxide anion release from macrophages isolated from child sera (Sections 4.6.6.2 and
4.6.6.3). These results were adjusted only for age and sex. Other studies in humans
examined macrophage abundance in Pb-exposed workers, and evidence overall did not
clearly indicate an association with concurrent blood Pb level. Pinkerton et al. (1998)
considered potential confounding and found a
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macrophages as well as microglia and astrocytes in the brain, osteoclasts in the bone, and
testicular macrophages.
Fan et al. (2009b) reported major changes in phenotypic expression (e.g., CD68 and
ferritin light chain), organization, and functional activity of Kupffer cells connected to
apoptosis in the liver of 6-7 week old Sprague-Dawley rats treated with a single 2.5* 104
(iM Pb nitrate injection. Dosing of juvenile Wistar rats with Pb injections (15 mg/kg of
Pb acetate daily for 2 weeks, resulting in a mean blood Pb level of 30 (ig/dL) during early
postnatal maturation resulted in an increase in GFAP, indicating chronic activation of
glial cells; an increase in pro-inflammatory cytokines (i.e., IL-lp, TNF-a and IL-6); and a
decrease in synaptophysin (component of presynaptic vesicles), indicating
neurodegeneration in brain tissue (Struzvnska et al.. 2007). Bone osteoblasts have been
shown to be affected by Pb exposure (Section 4.9.4). which, given the interactions
between osteoblasts and osteoclasts (Chang et al., 2008a). could have implications for
development of arthritis [reviewed in Zoeger et al. (2006)]. In vitro, 1 (iM Pb acetate
elevated TGF-(3 production and cartilage formation in limb bud mesenchymal cells
(Zuscik et al.. 2007). Kaczynska et al. (2011) reported effects on alveolar macrophages
after Pb acetate treatment (i.p. 25 mg/kg, 3 days, resulting in mean blood Pb level of
2.1 (ig/dL) in Wistar rats. Macrophages infiltrated airways, limiting air space available to
gas exchange and contained parts of phagocytized surfactant and alveolar lining. Resident
immune cells in reproductive organs have been shown to be affected by high
concentration Pb exposure. In male BALB/c mice, Pace et al. (2005) found that 0.1 ppm
Pb acetate exposure in drinking water PND1-PND42 (mean peak blood Pb level:
59.5 (ig/dL) resulted in sterility concomitantly with a decrease in the testicular
macrophage population and an increase in apoptotic testicular cells.
In summary, an extensive toxicological evidence base demonstrates that Pb exposure
decreases functionality of macrophages and promotes a hyperinflammatory phenotype.
Animals with dietary Pb exposure resulting in blood Pb levels (upon cessation of
exposure) relevant to this ISA, 8.2 and 18 (ig/dL, had reduced macrophage generation
and phagocytosis (Bunn etal.. 2001c: Kowolenko et al.. 1991). Some in vitro studies
(Bussolaro et al.. 2008; Zhou etal.. 1985) provided supporting evidence. Several
observations link Pb exposure to impaired function and/or structure of specialized
macrophages in nonlymphoid tissue, including liver Kupffer cells and alveolar
macrophages. The results suggest that immune dysfunction may contribute to the effects
of Pb on dysfunction in nonlymphoid tissues and provides a link between immune
dysfunction and impairments in other organ systems. However, the implications of
findings to effects in humans are uncertain because in several studies, animals were
treated with Pb by injection and/or had high blood Pb levels (i.e., >30 (ig/dL). Evidence
for Pb-induced decreases in macrophage functionality provides mode of action support
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for observations of Pb-induced decreased host resistance in animals (Section 4.6.5.1).
The sparse epidemiologic evidence is not conclusive but was a lesser consideration than
the toxicological evidence in drawing conclusions about the effects of Pb exposure on
macrophages because most epidemiologic studies did not examine the functional state of
macrophages. The study that did examine the functional state of macrophages,
i.e., mediators of host defense or inflammation, found an association with concurrent
blood Pb level in children (Pineda-Zavaleta et al., 2004). These results were consistent
with toxicological evidence but were adjusted only for potential confounding by age and
sex. Occupational studies examined abundance of monocytes or markers of activated
macrophages and other antigen presenting cells, and evidence did not clearly indicate a
difference in Pb-exposed workers. Inconsistency among studies was not related to
differences in sample size, age, or blood Pb levels of Pb-exposed workers. None of the
occupational studies considered potential confounding by other occupational exposures
or provided concentration-response information.
4.6.2.5 Neutrophils
In the 2006 Pb AQCD, Pb exposure was not judged to have a strong effect on
neutrophils, which comprise the majority of polymorphonuclear cells (PMNs) (U.S. EPA.
2006b). This conclusion was based on the limited available toxicological evidence as
compared with that for effects on other immune cells. The modulation of neutrophil
activity may have important consequences on the dysregulation of inflammation and
ability of organisms to respond to infectious agents. Studies of cultured human PMNs
(Governa et al.. 1987) and occupationally-exposed adults (Oueiroz et al.. 1994a; Oueiroz
et al.. 1993; Valentino et al.. 1991; Bergeret et al.. 1990) found Pb-associated reductions
in PMN functionality, as indicated by reduced chemotactic response, phagocytic activity,
respiratory oxidative burst activity, or ability to kill ingested antigen. These observations
were made in groups of Pb-exposed workers (n = 10-60) with mean ages ranging 34 to 41
years and blood Pb levels ranging 14.8 to 91.4 (ig/dL. These studies reported specifically
on neutrophil function not a multitude of other immune parameters. Thus, the evidence
could have been influenced by publication bias. In these cross-sectional studies, the
temporal sequence between Pb exposure and neutrophil function cannot be determined.
Other limitations across all studies include the lack of consideration for potential
confounding factors, including other workplace exposures, and high blood Pb levels of
Pb-exposed workers.
Instead of examining neutrophil functional activities, the few available recent studies of
animals and occupationally-exposed adults examined neutrophil counts, an increase in
which has been interpreted by some investigators to be a hyperinflammatory response
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and a compensatory response to Pb-induced impairment in neutrophil chemotactic
activity. A study in male Wistar rats found that 12 mg Pb spheres implanted in brains
(compared with control glass spheres) resulted in greater neutrophil infiltration from day
7 to 28 with inflammatory-related damage that included apoptosis and indications of
neurodegeneration (Nakao et al., 2010). However, the route of Pb administration has
uncertain relevance to the typical routes of Pb exposure in humans.
Occupational studies produced contrasting results that were not explained by differences
in the blood Pb levels of workers. DiLorenzo et al. (2006) found a Pb-associated higher
absolute neutrophil count (ANC) in analyses that adjusted for some potential
confounding factors and showed a concentration-dependent relationship. In an analysis
combining 68 ceramic, Pb recycling, or bullet manufacturing workers and 50 control food
plant workers in Italy, a 1 (ig/dL higher concurrent blood Pb level was associated with a
21.8 cells/jiL (95% CI: 11.2, 32.4 cells/jiL) higher ANC. While these results were
adjusted for age, current BMI, and current smoking status, other workplace exposures
were not examined. Pb-exposed workers had a mean age 44 years and a geometric mean
concurrent blood Pb level of 20.5 (ig/dL. Controls had a mean age 46.8 years and mean
blood Pb level of 3.5 (ig/dL. Neutrophilia (n >7,500 cells/mm3) was found in 8 workers
described to have medium to high Pb exposures (exact blood Pb levels not reported) but
in no controls suggesting that long-term, higher-level Pb exposures can lead to a
biologically meaningful excess of circulating neutrophils. Additionally, a blood Pb
concentration-dependent relationship was indicated by observations of a monotonic
increase in ANC across increasing blood Pb level groups: controls, workers with blood
Pb levels < 30 (ig/dL, and workers with blood Pb levels >30 (ig/dL. Results further
indicated an interaction between concurrent blood Pb level and current smoking. ANC
increased across the three blood Pb groups among current smokers but not nonsmokers.
In contrast, in a study that did not consider any potential confounding factors, Conterato
et al. (2013) found lower neutrophil concentrations among 23 battery workers and 50
painters in Brazil with mean concurrent blood Pb levels of 50.0 and 5.4 (ig/dL,
respectively, than among 36 controls with a mean blood Pb level of 1.5 (ig/dL.
Pb-exposed workers did not consistently have higher levels of eosinophils, basophils,
monocytes, or total lymphocytes either.
Support for the decreased neutrophil function found in Pb-exposed workers is provided
by findings of Pb-associated decreases in IFN-y (Section 4.6.6.1) and complement. IFN-y
gradients serve to direct neutrophil migration. The complement system is a component of
the innate immune system that is involved in chemotaxis of macrophages and neutrophils
and phagocytosis of antigens. The few available occupational studies found lower
complement in Pb-exposed workers with mean blood Pb levels >60 (ig/dL (Undeger et
al.. 1996; Ewers et al.. 1982). higher than those relevant to this ISA. Neither study
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considered potential confounding factors, including other workplace exposures. The
evidence has limited implications also because the cross-sectional nature of studies
cannot establish the temporal sequence between Pb exposure and complement. Pb has
been associated with increased TNF-a (Section 4.6.6.1). which has been linked to both
neutrophil stimulation and apoptosis. Therefore, it is unclear whether the evidence for
TNF-a describes a potential mode of action for the effects of Pb observed on neutrophils
in Pb-exposed workers.
In summary, previous occupational studies provided evidence for the effects of Pb
exposure on neutrophils by finding that compared with controls, Pb-exposed workers had
lower neutrophil functionality (Queiroz et al.. 1994a: Queiroz et al.. 1993; Valentino et
al.. 1991; Bergeret et al., 1990) and lower complement (Undeger et al., 1996; Ewers et
al.. 1982). which is a mediator of phagocyte functionality. The limited number of recent
epidemiologic studies examined only neutrophil abundance and conclusively did not find
Pb-exposed workers to have higher or lower neutrophil abundance. While there is
evidence for Pb-associated reduced neutrophil functionality, firm conclusions are not
warranted because the results are based on cross-sectional examination of male workers
with relatively high blood Pb levels (range: 18.6-100 ng/dL), and they lack consideration
for potential confounding factors including other occupational exposures, concentration-
response information, and analogous toxicological evidence.
4.6.2.6 Dendritic Cells
Whereas as research reviewed in the 2006 Pb AQCD (U.S. EPA. 2006b) focused on
examining T cells (Section 4.6.2.1). recent ex vivo and in vitro results suggest that the
effects of Pb on suppressing Thl activity and promoting Th2 activity may be a
consequence of the direct action of Pb on the function of dendritic cells, a major APC.
Gao et al. (2007) found that 25 (iM Pb chloride exposure for 10 days stimulated dendritic
cell maturation in bone marrow cultures as shown by the increased expression of MHC
class II, a cell surface marker that is used to distinguish between immature and mature
dendritic cells. Additionally, upon activation with LPS, Pb-matured dendritic cells
produced less IL-6, TNF-a, and IL-12 (stimulates growth and differentiation of T cells)
than did control cells but the same amount of IL-10 (inhibits production of Thl
cytokines). The effect of Pb on altering the cytokine expression profile of dendritic cells,
in particular, the lower IL-12/IL-10 ratio and the effect on increasing the ratio of cell
surface markers CD86/CD80 were indications that Pb exposure preferentially induced
development of dendritic cells that signal naive T cell populations to shift toward a Th2
phenotype. Supporting a role for dendritic cells in skewing to a Th2 phenotype, ex vivo
results from the same study showed that Pb-naive adult BALB/c mice implanted with
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Pb-exposed dendritic cells had suppressed DTH (Section 4.6.2.3) and IgG2a antibody
(Section 4.6.3) responses (Gao et al.. 2007).
4.6.2.7 Natural Killer Cells
Based mostly on studies reviewed in the 2006 Pb AQCD, evidence does not clearly
indicate that Pb exposure affects the innate immune NK cells, which mediate host
defense by killing infected cells. Some epidemiologic studies adjusted for factors such as
age, sex, and smoking but did not find differences in NK cell abundance or level of
functional activity by blood Pb level in children or adults with or without occupational
exposures (N = 145-675) (Karmaus et al.. 2005; Sarasua et al., 2000; Pinkerton et al.,
1998). Studies in children did not examine potential confounding by SES, and the study
in Pb-exposed workers did not examine other workplace exposures. Other smaller
(N = 30-108) studies that did not consider potential confounding factors found a positive
correlation between concurrent blood Pb level and NK cell abundance in adults in Italy
(Boscolo et al.. 2000; 1999) or reported no statistically significant association in children
(quantitative results not reported) (Belles-Isles et al., 2002). Pb-exposed workers in the
U.S., Europe, and Asia (N = 25-141, mean ages: 26-49 years) with higher concurrent
blood Pb levels (means: 6.5-128 ng/dL) had similar means of NK cell abundance or
functional activity as did unexposed controls (N = 10-84, mean blood Pb levels:
<2-16.7 (ig/dL, mean ages: 28-47 years) (Garcia-Leston et al.. 2011; Mishra et al.. 2003;
Pinkerton et al.. 1998; Yucesov et al.. 1997b; Undegeretal.. 1996; Fischbein et al.. 1993;
Kimber et al.. 1986).
The epidemiologic evidence is not sufficiently informative for drawing conclusions about
the effects of Pb exposure on NK cells because of its many limitations including cross-
sectional nature, limited consideration for potential confounding, and lack of
concentration-response information. However, toxicological evidence equally does not
clearly indicate an effect of Pb on NK cells. A decrease in NK cell activity was found in
6-8 week-old BALB/c mice but with higher Pb exposure than that relevant to this ISA
(1,300 ppm Pb acetate in drinking water, 10 days, blood Pb level -100 (ig/dL) (Queiroz
et al., 2011). In an in vitro study, Fortier et al. (2008) found that Pb chloride
(7.5-20.7 (ig/dL) did not affect NK cytotoxicity compared with the control. However,
Pb chloride was not found to affect monocytes or lymphocytes either.
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4.6.3 Humoral Immunity
The 2006 Pb AQCD (U.S. EPA. 2006b) described another hallmark effect of Pb on the
immune system to be an enhanced humoral immune response as characterized by
increased production of IgE antibodies. Several animal and epidemiologic studies (Table
4-32) demonstrated Pb-associated increases in IgE, which mediates inflammation in
allergic and allergic asthma responses by binding to mast cells and releasing histamines,
leukotrienes, and interleukins upon exposure to an allergen. Neither toxicological nor
epidemiologic evidence (Table 4-32) consistently indicated that Pb exposure was
associated with changes in other classes of Igs including IgG, IgM, and IgA, which
function in complement activation and host resistance or activation of immune cells.
4.6.3.1 Immunoglobulin E Antibody and B Cells
In toxicological evidence, there was a lack of coherence between results for IgE and
activation of B cells, which regulate IgE production through differentiation into antibody-
producing cells. Pb chloride exposure in vitro (10 (iM up to 5 days) was found to increase
markers of B cell activation, including cell surface markers and levels of plaque forming
cells (PFCs), which are a measure of antibody-forming cells (McCabe and Lawrence.
1990; Lawrence. 198 la). However, several studies in animals (Swiss mice and rabbits)
found a wide range of Pb exposure concentrations (0.5 to 250 ppm Pb acetate or
tetraethyl Pb for 3-10 weeks, postnatal via drinking water) to decrease levels of PFCs
(Blaklev et al.. 1980; Koller and Kovacic. 1974; Koller. 1973). Among many mice strains
tested, Mudzinski et al. (1986) found an increase in PFCs only in BALB/c mice with 8-
week postnatal dietary Pb acetate exposure that produced high blood Pb levels, 70 (ig/dL.
Epidemiologic studies of children and adults [Table 4-32 with group comparisons and
(Boscolo et al., 2000; Boscolo etal.. 1999) with correlation analyses] did not find a
consistent association between blood Pb level and the abundance of B cells. Cell
abundance may not reflect activation. Inconsistencies among studies did not appear to be
related to differences in age, group blood Pb levels, or the extent of consideration for
potential confounding (Table 4-32).
Most animal studies found Pb-induced increases in IgE, with key evidence provided by
studies that examined Pb acetate exposure through drinking water during the gestation
and/or lactation period (Snyder et al., 2000; Miller et al.. 1998). In particular, Snyder et
al. (2000) found elevated IgE in juvenile BALB/c mice with relevant blood Pb levels,
means 5-20 (ig/dL measured 0-1 week after gestational and/or lactational Pb exposure. In
Miller et al. (1998). elevated IgE was found in adult mice exposed gestationally to Pb via
drinking water of dams that had blood Pb levels of 30-39 (ig/dL. Chen et al. (2004) did
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not find gestational dietary Pb acetate exposure to result in an increase in IgE in adult
F344 rats who had blood Pb levels of 6.75 and 8 (ig/dL, measured one week
postweaning. In BALB/c and OVA-transgenic (produce OVA-specific T cells) mice, Heo
et al. (1997; 1996) found concomitant Pb-induced increases in IgE and IL-4, consistent
with the mode of action of IL-4 to induce class switching of B cells to IgE producing
cells. However, these effects were observed with Pb administered via s.c. injection
(50 (ig/100 (iL, 3 times per week for 3 weeks) and with higher blood Pb levels, 38 (ig/dL,
than those relevant to this ISA. Some of these animal studies examined multiple immune
endpoints; however, chance findings due to multiple comparisons likely are not
responsible for the IgE findings because the pattern of results consistently pointed to a
shift from Thl to Th2 responses (Miller et al.. 1998: Heoetal.. 1997: 1996).
Observations of Pb-induced increases in IgE in animals provide biological plausibility for
associations observed between higher Pb biomarkers levels and higher serum IgE levels
in various populations of children (Hegazy et al.. 2011: Hon et al.. 2010: Hon et al.. 2009:
KarmausetaL 2005: Annesi-Maesano et al.. 2003: Sun etal., 2003: LutzetaL 1999)
(Table 4-32). The evidence was based on cross-sectional analyses which preclude
establishing the temporal sequence between Pb exposure and IgE. Also, a monotonic
blood Pb concentration-dependent increase in IgE was not consistently observed.
Associations between blood Pb level and IgE were found in studies that generally had
population-based recruitment. Most studies did not report sufficient information to assess
the potential for biased participation by Pb exposure and immune conditions. Most
studies examined multiple immune endpoints; however, associations were not isolated to
IgE.
4-502
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Table 4-32
Study
Comparison of serum immunoglobulin levels and B cell abundance among various blood
Study Population and Methodological Details
Blood Pb
Level (ug/dl_) |gEa
igGb
igMb
igAb
Pb groups.
B cells0
Children
Karmaus et al.
(2QQ5)
Sarasua et al.
(2000)
' _'— " " " /
Sarasua et al.
(2000)
' f—v v v i
331 children, ages 7-10 yr, Hesse, Germany
Cross-sectional. School-based recruitment.
Moderate participation rate. Some subjects live
near industrial facilities. Multiple exposures
examined. Results adjusted for age, sex,
#cigarettes/day smoked in home in previous
12 months, # infections in the previous
12 months, serum lipid level, blood
organochlorine level. No consideration for
potential confounding by SES, allergens. No
monotonic C-R.
382 children, ages 6-35 mo, Multiple U.S.
locations
Cross-sectional. No information on participation
rate. Large proportion lives near Pb and other
metal sources. Comparison group age- and
demographically-matched. Results adjusted for
age, sex, and study location. No consideration for
potential confounding by SES, allergens.
Inconsistent C-R.
562 children, ages 36-71 mo, Multiple U.S.
locations
Same methodology as above.
<2.2 46(1.0)
2.21-2.83 30(0.65)
2.84-3.41 59(1.28)
>3.41 59(1.28)d
Means Boys:
2.8 Girls: 2.5
0.6-4.9
5-9.9
10-14.9
>15
Mean(SD):
7.0 (5.2)
0.6-4.9
5-9.9
10-14.9
>15
Mean (SD):
6.0(4.3)
1210
1,214
1,241
1,201
609
666d
680d
630
817
813
856
835
150
143
153
148
103
108
105
124
120
116
125
121
123
121
133
136
50.1
55.0
58.2
61.4
88.6
90.9
96.3
94.1
418e(1.0)
353 (0.84)
389 (0.93)
393 (0.94)
19.1 (1.0)
20(1 05)
\ w/
20.4(1.07)
22.2(1.16)
18.4(1.0)
17.6(0.96)
19.2(1.04)
18.6(1.01)
4-503
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Table 4-32 (Continued): Comparison of serum iniiminoglobulin levels and B cell abundance among various blood Pb groups.
Study
Sarasua et al.
(2000)
Lutz et al.
(1999)
Hegazy et al.
(2011)
Zhao et al.
(2004)
Study Population and Methodological Details
675 children ages 6-15 yr, Multiple U.S. locations
Same methodology as above.
279 children, ages 9 mo-6 yr, Springfield, MO
Cross-sectional. Recruitment from public
assistance and Pb poisoning prevention program.
No information on participation rate. Results
adjusted for age. Lack of rigorous statistical
methods. No consideration for potential
confounding by SES, allergens. Monotonic C-R
for IgE except in highest blood Pb group.
318 children, ages 6 mo-7 yr, Egypt
Cross-sectional. Clinic-based recruitment. No
information on participation rate. Lack of rigorous
statistical methods. Potential confounding not
considered. No monotonic C-R.
73 children, ages 3-6 yr, Zhejiang Province, China
Cross-sectional. No information on recruitment
method or participation rate. Lack of rigorous
statistical methods. Potential confounding not
considered.
Blood Pb
Level (ug/dL)
0.6-4.9
5-9.9
10-14.9
>15
Mean (SD):
4.0(2.8)
<10
10-14
15-19
20-69
64% with blood
Pb level <10
<5
5-9
10-14
15-19
20-44
45-69
63% with blood
Pb level < 10
<10
>10
Mean (SD):
9.5(5.6)
igŁa
51.8(1.0)
74.0(1.43)
210.7(4.07)
63.7(1.23)d
13.0(1.0)
12.0(0.92)
20.8(1.60)
14.9(1.15)
20.4(1.57)
10.2(0.78)d
lgGb lgMb lgAb B cells0
1,031 128 140 16.1(1.0)
1,094d 131 143 15.8(0.98)
1,048 136 140 15.3(0.95)
1,221 106 108 20.1(1.25)
13.4(1.0)
12.6(0.94)
16.9(1.26)
11.1 (0.83)
16.6(1.0)
16.8(1.01)
4-504
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Table 4-32 (Continued): Comparison of serum iniiminoglobulin levels and B cell abundance among various blood Pb groups.
Study
Sun et al.
(2003)
Adults without
Sarasua et al.
(2000)
Study Population and Methodological Details
73 children, ages 3-6 yr, Zhejiang Province, China
Cross-sectional. No information on recruitment
method or participation rate. Lack of rigorous
statistical methods. Potential confounding not
considered.
Occupational Pb Exposures
433 children and adults, ages 16-75 yr, Multiple
U.S. locations
Same methodology as that in children.
Blood Pb
Level (ug/dl_) |gEa
<10 32.30
>10 41.33f
Mean (SD):
9.5(5.6)
0.6-4.9
5-9.9
10-14.9
>15
Mean (SD):
4.3(3.9)
lgGb lgMb lgAb B cells0
40.53 41.53
34.76f 31.74f
1,099 175 252 13.9(1.0)
1,085 175 242 13.0(0.94)
1,231 262d 283 12.4(0.89)
1,169 139 190 14.8(1.06)
Adults with Occupational Pb Exposures
Pinkerton et al.
(1998)
Fischbein et al.
(1993)
Kimber et al.
84 hardware factory controls, mean age 30 yr
145 male Pb smelter workers, mean age 33 yr,
U.S., exact location NR
Cross-sectional. Results adjusted for age, race,
current smoking status, workshift. No
consideration for potential confounding by other
workplace exposures, SES.
36 industrial worker controls, mean age 47 yr
36 firearms instructors, mean age 49 yr
15 firearms instructors, mean age 48 yr
New York metropolitan area
Cross-sectional. Lack of rigorous statistical
methods. Potential confounding not considered.
21 unexposed male controls, ages 20-60 yr
Mean: <2
Mean: 39
(cumulative
over 0.5-1 8 yr)
NR
Mean: 14.6
Mean: 31.4
Mean: 11.8
1,090 94.5 180 14.6(1.0)
1,110 106.2 202 13.2(0.90)
8.6(1.0)
10.5(1.22)
11.2(1.30)d
1,062 129.4 223.5
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Table 4-32 (Continued): Comparison of serum iniiminoglobulin levels and B cell abundance among various blood Pb groups.
Study
(1986)
Heo et al.
(2QQ4)
Anetor and
Adeniyi (1998)
Ewers et al.
(1982)
Undeger et al.
(1996)
Study Population and Methodological Details
39 male tetraethyl Pb plant workers, ages 25-61
yr, U.K.
Cross-sectional. Lack of rigorous statistical
methods. Potential confounding not considered.
606 Pb battery plant workers, ages not reported,
Korea
Cross-sectional. Monotonic C-R found; lack of
rigorous statistical methods. Potential
confounding not considered.
50 male controls, ages 22-58 yr
80 male Pb-exposed workers, ages 21-66 yr,
Nigeria
Cross-sectional. Lack of rigorous statistical
methods. Potential confounding not considered.
53 male various occupation controls, ages 21-54
yr
72 male Pb battery/smelter workers, ages 16-58
yr, Germany
Cross-sectional. Lack of rigorous statistical
methods. Potential confounding not considered.
25 male university worker controls, ages 22-56 yr
25 male Pb battery plant workers, ages 22-55 yr,
Turkey
Cross-sectional. Lack of rigorous statistical
methods. Potential confounding not considered.
Blood Pb
Level (ug/dl_) |gEa
Mean: 38.4
<10 112.5(1.0)
10-29 223.3(1.99)
> 30 535.8 (4.76)d
Mean: 23
Mean: 30.4
Mean: 56.3
Mean: 11.6
Mean: 51.4
Mean: 16.7
Mean: 74.8
lgGb lgMb lgAb B cells0
1,018 104.0 242.5
1,997 215 188
1,188d 191 144d
190g 156g 137g
173 127 129
1,202 140 210 636e(1.0)
855d 93.3d 168 546e (0.86)
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Table 4-32 (Continued): Comparison of serum iniiminoglobulin levels and B cell abundance among various blood Pb groups.
Study
Alomran and
Shleamoon
(1988)
Study Population and Methodological Details
18 management personnel controls
39 Pb battery workers, mean age 35 yr Iraq
Cross-sectional. Controls age matched. Lack of
rigorous statistical methods. Potential
confounding not considered.
Blood Pb
Level (ug/dl_) |gEa
NR
64 (previously
reported)
igGb
1,713
1,610
lgMb lgAb B cells0
183
170
Note: Results are grouped according to strength of study design and methodology including the extent of potential confounding considered and the relevancy of the examined blood Pb
levels to current levels in the general U.S. population.
algE data are presented as lU/mL unless otherwise specified. In parentheses are ratios of IgE in the higher blood Pb group to IgE in the lowest blood Pb group.
bOther Ig data are presented as mg/dL unless otherwise specified.
°B cell data are presented as the percentage of B cells among all lymphocytes unless otherwise specified. In parentheses are the ratio of B cells in the higher blood Pb group to B cells
in the lowest blood Pb group.
dp <0.05 for group differences.
eData represent the number of cells/uL serum.
'Data represent the mean rank for Mann-Whitney U test, p = 0.07 for IgE.
9Data are presented as Ill/mL
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Karmaus et al. (2005) had greater adjustment for potential confounding factors and
examined associations with lower blood Pb levels than did other studies. The detection
limit for blood Pb was 0.9 (ig/dL. Compared with 82 children (ages 7-10 years) in
Germany with concurrent blood Pb level <2.2 (ig/dL, children with blood Pb levels
2.84-3.41 (ig/dL (n = 86) and >3.4 (ig/dL (n = 82) had 28% higher serum IgE levels
(p = 0.03, F-test). These differences were found with the adjustment for age, serum lipid
levels, number of infections in the previous 12 months, number of cigarettes/day smoked
in the home in the previous 12 months, and blood organochlorine levels. Some children
lived near organochlorine-polluting facilities. SES was not examined. Although IgE was
elevated in children with relatively low blood Pb levels (>2.84 (ig/dL), in these children
ages 7-10 years, the contribution of higher past Pb exposures cannot be excluded. Similar
differences in IgE count on basophils were not found among the blood Pb quartiles.
Although serum IgE and basophil-bound IgE have been correlated in adults (Malveaux et
al., 1978; Conroyetal.. 1977). few data are available in children (Dehlink et al.. 2010). A
study in children found that serum IgE levels were not correlated with basophil-bound
IgE (Spearman r = -0.003) but were correlated with other IgE receptor-expressing cells
such as dendritic cells and monocytes (Spearman r = 0.43 to 0.65, p <0.05) (Dehlink et
al.. 2010). The number of IgE-bound basophils also has been highly variable across
individuals, particularly children (Hausmann et al.. 2011; Dehlink etal. 2010). Thus, it is
not unexpected that higher blood Pb level was associated with higher serum IgE but not
basophil-bound IgE counts in Karmaus et al. (2005). In this study, blood Pb level was not
associated with serum levels of IgG, IgA, or IgM or B cell abundance. Lutz et al. (1999)
found higher serum IgE in low SES children (9 months-6 years) on public assistance in
Springfield, MO after only adjusting for age, albeit with concurrent blood Pb levels
>10(ig/dL(n= 105/279).
Recent studies in children (non-U.S.) also reported associations between concurrent
blood Pb level and elevated serum IgE but did not adjust for potential confounding
factors (Hesazv etal.. 2011: Hon etal.. 2010: Hon et al.. 2009). Ron et al. (2010: 2009)
demonstrated associations in 110 children (mean age 9.9 years) with atopic dermatitis in
Hong Kong with low blood Pb levels (range: 1.4-6.0 (ig/dL) and found that blood Pb
level also was correlated with severity of atopic dermatitis, a condition commonly
characterized by elevated IgE levels. Among 318 children, ages 6 months to 7 years, in
Egypt, Hegazy et al. (2011) did not find a monotonic increase across the blood Pb groups.
While most studies of children examined concurrent blood Pb, Annesi-Maesano et al.
(2003) found that Pb level in infant hair (mean: 1.38 ppm, n = 234) but not cord blood
(mean: 6.7 (ig/dL, n = 326) or placenta (mean: 9.6 (ig/dL, n = 332) was associated with
cord serum IgE in newborns in Paris. While this study better indicated the temporal
sequence between Pb exposure and an increase in IgE, an important limitation was a lack
4-508
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of consideration for potential confounding. The authors inferred a stronger effect of Pb
exposure integrated over the entire prenatal period than exposure closer to birth. But,
there is no empirical basis for interpreting hair Pb levels (Section 3.3.4.2). Cotinine was
not associated significantly with hair Pb or IgE. The hair Pb-cord blood IgE correlation
was larger among the 67% infants with mothers without allergies (Spearman r = 0.21,
p <0.01) than infants with maternal allergies (Spearman r = 0.12), pointing to a possible
masking of a Pb association by the stronger association of family history of allergy.
Blood Pb level also was associated with IgE in adults without (Pizent et al., 2008) and
with occupational Pb exposure (Heo et al.. 2004). Pizent et al. (2008) adjusted for
potential confounding by age, pack years smoking, and alcohol consumption and found
that higher concurrent blood Pb level was associated with higher IgE in 166 women in
Zagreb, Croatia of similar SES (i.e., white-collar office workers) (Pizent et al.. 2008).
Among women not on hormone replacement therapy or oral contraceptives, a 1 (ig/dL
higher blood Pb level was associated with a 0.60 (95% CI: 0.58, 1.18) higher log of IgE.
Concurrent blood Pb levels were low in these women who were aged 19-67 years (mean:
2.16 (ig/dL, range: 0.56-7.35 (ig/dL); however, the cross-sectional study design makes it
difficult to characterize the temporal sequence between exposure and outcome or the
timing, level, frequency, and duration of Pb exposure that contributed to the observed
association. Authors did not report an effect estimate in men because it did not attain
statistical significance. Without quantitative results, it is difficult to ascertain whether
there was suggestion of association in men but insufficient power to indicate statistical
significance due to the smaller number of men examined (50 men versus 166 females).
Another study of 34 men with and without allergy in Italy also did not report quantitative
results and only indicated a lack of statistically significant correlation between concurrent
blood Pb level (median: 11 (ig/dL) and IgE without considering potential confounding
(Boscolo et al.. 1999).
Limitations of the collective epidemiologic evidence for IgE include the cross-sectional
analyses with limited adjustment for potential confounding. Karmaus et al. (2005) found
a blood Pb-IgE association with adjustment for age, blood organochlorine levels, serum
lipid levels, number of infections in the previous 12 months, and number of cigarettes
a day smoked in the home in the previous 12 months. Lutz et al. (1999) comprised a low
SES population on public assistance; however, none of the studies adjusted for SES or
allergen exposure. Lower SES has been associated with poorer housing conditions,
higher exposures to Pb, allergens, and other factors associated with allergy and asthma.
Allergen exposure and lower SES are associated with higher IgE and related conditions
such as allergy and asthma (Bryant-Stephens. 2009; Dowd and Aiello, 2009; Aligne et
al.. 2000). Most studies did not provide detailed demographic or residential information.
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Thus, there is uncertainty as to the extent to which the blood Pb-IgE associations
observed in children may be confounded by unmeasured SES and/or allergen exposure.
For the cross-sectional evidence for IgE, consideration is given to the possibility of
associations being influenced by reverse causation, i.e., the possibility that higher IgE
results in greater inflammation and consequently, greater lung permeability and greater
absorption of Pb into the blood. Studies have not directly compared Pb uptake and
transfer into blood in groups with different IgE levels but have assessed lung epithelial
permeability by examining lung clearance of inhaled very small, hydrophilic radiolabeled
solutes. Compared with controls, animals sensitized with allergens to produce higher IgE
had similar clearance from lungs (Turi et al.. 2011). Histamine was shown only
transiently to increase lung clearance of 99mTc-DTPA particles (492 daltons, 0.57 nm) in a
baboon (Yeates and Hameister. 1992) and in humans, in both those with and without
asthma (ReesetaL 1985; Braudeetal. 1984; O'Byrne et al., 1982). Histamine is
released by IgE-bound mast cells and basophils upon exposure to sensitized antigens and
leads to inflammation. Increased lung clearance of inhaled 99mTc-DTPA particles, and in
some cases, increased presence in blood, has been found in some (Ilowite et al.. 1989;
ReesetaL. 1985; O'Byrne et al.. 1984) but not all (Del Donno et al.. 1997; Elwood et al..
1983) studies of subjects with asthma (n = 9-13 across studies) and also with airway
inflammation associated with fibrosis, smoking, and pulmonary edema (U.S. EPA.
2009a). The available evidence does not inform Pb absorption directly but provides some
indication of higher inflammation resulting in increased lung permeability to very fine
particles. However, a role for reverse causation may apply particularly to very fine
particulate Pb and exposure via inhalation. Evidence does not indicate inhalation to be a
significant route of exposure compared with ingestion (Section 3.1). and a relationship
between IgE-mediated inflammation and gastrointestinal permeability has not been
indicated. Thus, there is not an empirical basis for attributing the blood Pb-IgE
associations observed in children or adults to reverse causation.
4.6.3.2 Other Immunoglobulin Antibodies
A small number of available recent toxicological studies examined IgG subtypes, and as
in previous studies, found inconsistent effects of Pb exposure. Kasten-Jolly et al. (2010)
examined 100 (iM Pb acetate in drinking water of BALB/c dams GD8-PND21 because it
produced relevant blood Pb levels in another study, i.e., 10-30 (ig/dL (Snyder et al..
2000). Pb-exposed pups had increases in the expression of genes encoding Ig antibodies
and those involved in B lymphocyte function and activation. These included genes for the
heavy chain of IgM, IL-4, IL-7 and IL-7 receptor, IL-21, RAG-2, CD antigen 27, B-cell
leukemia/lymphoma 6, RNA binding motif protein 24, Histocompatibility class II antigen
4-510
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A (beta 1), Notch gene homolog 2, and histone deacetylase 7A. These results were
produced by a microarray analysis of hundreds of genes, which is subject to a higher
probability of finding an effect by chance.
Other recent animal studies examined specific IgG subtypes, did not find Pb-induced
changes in a consistent direction, and thus did not clearly indicate a shift to Th2
responses. All animal studies were limited by the higher blood Pb levels of animals than
those relevant to this ISA. Fernandez-Cabezudo et al. (2007) reported evidence for a shift
to Th2 responses following Salmonella infection in C3H/HeN mice exposed postnatally
to 1 x 104 (iM Pb acetate in drinking water for 16 weeks (resultant mean blood Pb level:
106 (ig/dL). Relative to control mice, production of the Th2 cytokine IL-4 increased in
spleen cells of Pb-exposed mice after infection as did serum levels of Salmonella-specific
IgGl. Infection increased Thl-mediated IgG2a levels in control but not Pb-exposed mice.
In contrast, Gao et al. (2006) found a Pb-induced increased IgG2a/IgGl ratio, albeit via
i.p. injection (50 (ig Pb chloride, 3 times per week for 3 weeks), high blood Pb levels
(65 (ig/dL), and in a highly-specialized strain of adult knockout mice without ability to
produce IFN-y. This result was surprising given evidence that IFN-y usually directs
secretion of IgG2a; however, the authors suggested that in these knockout mice, Pb may
initiate a Thl response via an IFN-y independent pathway to enhance IgG2a production.
Carey et al. (2006) found concentration-dependent increases in both IgG2a- and IgGl-
producing cells (after 7 days) in adult BALB/c mice treated with subsensitizing doses of a
T cell-independent (Trinitrophenyl-Ficoll [TNP-Ficoll]) or T cell-dependent (TNP-
ovalbumin [TNP-OVA]) hapten-protein conjugate and 25-50 (ig Pb chloride by injection.
These results indicated stimulation of both Thl- and Th2-mediated mechanisms. Pb
treatment also increased the numbers of T and B cells and IgM-producing cells in the
lymph node against both TNP-Ficoll and TNP-OVA. The increase in IgM-producing
cells against TNP-Ficoll indicated a T-cell independent mechanism. Despite finding
increases in both IgGl- and IgG2a-producing cells, the authors concluded that Pb
induced Th2 skewing based on observations of Pb-induced increases in T and B cells and
suppression of DTH. Thus, results indicated the potential for Pb to promote allergic
sensitization against T-dependent antigens.
Some (Belles-Isles et al.. 2002; Sarasua et al.. 2000) but not all (Karmaus et al.. 2005).
studies in children found blood Pb-associated increases in IgG but had limited
implications because of the lack of extensive consideration for potential confounding
factors. Sarasua et al. (2000) adjusted for age, sex, and location and found associations of
higher concurrent blood Pb level with higher IgA, IgG, and IgM in 372 U.S. children
ages 6-35 months but not older (36-71 months, 6-15 years, 16-75 years, n = 433-673). In
the youngest age group, a 1 (ig/dL higher blood Pb level was associated with a 0.8 [95%
CI: 0.2, 1.4], 4.8 [95% CI: 1.2, 8.4], and 1.0 [95% CI: 0.1, 1.9] mg/dL higher IgA, IgG,
4-511
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and IgM, respectively. In the youngest children, serum levels of all three examined Igs
were higher among the 24 children with concurrent blood Pb levels > 15 (ig/dL than
among the 165 children with blood Pb levels <5 (ig/dL. Examining cord blood Pb,
Belles-Isles et al. (2002) better indicated the temporal sequence between Pb exposure and
IgG levels. Among 97 newborns in Quebec, Canada from subsistence fishing
communities and two other towns (geometric mean blood Pb levels: 1.64 and 1.32 (ig/dL,
respectively), higher cord blood Pb level was associated with higher cord serum IgG
adjusted for prenatal maternal smoking status. However, IgG also was associated with
cord plasma organochlorines, which also are elevated with high fish diets.
Most examination of the effects of Pb on IgA, IgG, and IgM in humans is provided by
previous studies of Pb-exposed workers (mostly males) from various industries with
mean ages 32-36 years and mean blood Pb levels 38-74.8 (ig/dL (Anetor and Adeniyi.
1998; Pinkerton et al.. 1998; Undeger et al.. 1996; Queiroz et al.. 1994b; Alomran and
Shleamoon. 1988; Kimberetal.. 1986; Ewers et al.. 1982). As are toxicological findings
for these other Ig classes, epidemiologic evidence is mixed, with higher, lower, and
similar Ig levels found in Pb-exposed workers (n = 25-145) compared with unexposed
controls (n = 18-84). Some studies reporting lower Ig levels in Pb-exposed workers
included workers with the highest mean blood Pb levels (>50 (ig/dL) (Anetor and
Adeniyi. 1998; Undeger et al.. 1996; Ewers et al., 1982). The lack of analysis of potential
confounding factors, including other workplace exposures, precludes characterization of
others factors that may contribute to inconsistent associations in occupational studies.
4.6.3.3 Summary of Humoral Immunity
Evidence in support of the effects of Pb exposure on humoral immunity largely
comprises consistent toxicological and epidemiologic observations of Pb-associated
increases in IgE. The combined toxicological and epidemiologic results do not clearly
indicate the effects of Pb exposure on IgG, IgM, or IgA.
Collectively, epidemiologic studies conducted in the U.S., Europe, and Asia, some with
large study populations (n = 279-331) indicate higher serum IgE in children with
concurrent blood Pb levels >10 (ig/dL. There was some evidence of associations in non-
11.S. children with lower blood Pb level; however, the contribution of higher Pb
exposures earlier in childhood cannot be excluded (Honetal.. 2010; Karmaus et al..
2005). Pb-associated increases in IgE were found in children with atopic dermatitis (Hon
et al.. 2010; 2009) and children without immune conditions (Sun et al.. 2003). Other
studies did not report the health status of subjects. Thus, sufficient information was not
reported to assess the potential for selection bias. Among studies that provided
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concentration-response information, most did not find serum IgE to increase
monotonically across blood Pb groups (Hegazy et al.. 2011; Karmaus et al.. 2005; Lutz et
al., 1999). All evidence in humans is based on cross-sectional analyses, making it
difficult to establish the temporal sequence between Pb exposure and increase in IgE.
Some but not all evidence points to increased lung permeability to very small particles in
the presence of acute airway inflammation or inflammatory conditions, but there is a lack
of evidence for increased gastrointestinal permeability for the more dominant ingestion
route of Pb exposure. Thus, reverse causation seems unlikely as the primary explanation
for the observed blood Pb-IgE associations. Most studies did not consider potential
confounding, and none adjusted for SES. A study in children in Germany and a study in
adults in Croatia adjusted for age and smoking, with additional adjustment for blood
organochlorine levels in children (Karmaus et al., 2005) and alcohol consumption in
adults (Pizent et al.. 2008). Blood Pb level was associated with IgE in a low SES
population of children in Michigan (Lutz et al.. 1999) and in a population of female
office workers of similar SES in Croatia (Pizent et al.. 2008). However, uncertainty
remains regarding confounding in the associations observed between blood Pb level and
IgE in humans. Biological plausibility for the epidemiologic evidence is provided by the
Pb-induced increases in IgE observed in most of the animal studies, with some evidence
at blood Pb levels relevant to this ISA (Snvder et al.. 2000: Miller et al.. 1998).
Toxicological evidence indicates increases in IgE with gestational and/or postnatal
juvenile Pb exposure, whereas epidemiologic evidence points to associations with
concurrent blood Pb levels. Because concurrent blood Pb levels in children reflect both
recent and past Pb exposures, the combined evidence indicates that Pb exposures during
childhood may affect IgE levels. While toxicological evidence for B cell activation is
inconsistent, the mode of action for Pb-induced IgE production is well supported by
extensive toxicological evidence for Pb-induced increases in the Th2 cytokine, IL-4
(Section 4.6.6.1). The coherence between epidemiologic and animal findings for IgE and
evidence describing modes of action for increases in IgE supports a relationship between
Pb exposure and increases in IgE.
4.6.4 Inflammation
The 2006 Pb AQCD (U.S. EPA. 2006b) identified misregulated inflammation a major
immune-related effect of Pb based primarily on consistent toxicological evidence for
Pb-induced increases in pro-inflammatory cytokines (Section 4.6.6.1). PGE2, and ROS
(Section 4.6.6.3). Inflammation has been characterized as a major mode of action for Pb
effects in multiple organ systems such as the liver, kidney, and vasculature given that
immune cells make up permanent residents and infiltrating cell populations of these other
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organ systems (Section 4.2.5). Inflammation also provides a link between the evidence
for the effects of Pb on modulating function of immune cell and production of cytokines
and IgE and the evidence for the effects of Pb on immune-based conditions such as
infections and asthma and allergy. For example, IL-4-induces increases in IgE, which
primes basophils and mast cells to secrete histamine, leukotrienes, and cytokines, which
in turn, produce the inflammatory-related effects associated with asthma and allergy,
i.e., airway responsiveness, mucus secretion, respiratory symptoms. Pb-induced
inflammation also has been associated with diminished host resistance by inducing local
tissue damage. As described in Section 4.6.6. the few available recent toxicological
studies support the effects of Pb exposure on inflammation with findings of Pb-induced
increases in pro-inflammatory cytokines, ROS and PGE2.
The few available epidemiologic studies have found Pb-associated changes in ROS
release from macrophages (Section 4.6.6.3) and cytokine levels (Section 4.6.6.1) in
children and adults. Adding to this evidence, recent cross-sectional studies found
associations between blood Pb level and indicators of inflammation that may be related to
multisystemic effects. As discussed in Section 4.6.3.1. some but not all evidence has
indicated increased lung permeability in the presence of acute inflammation or various
inflammatory conditions, but evidence is lacking for a relationship with gastrointestinal
permeability for the more dominant ingestion route of Pb exposure. Thus, reverse
causation seems unlikely as the primary explanation for the observed blood Pb-
inflammation associations. However, because of the cross-sectional design of studies, the
temporal sequence between Pb exposure and inflammation cannot be established. The
most compelling epidemiologic evidence was provided by studies in adults, which were
larger and had greater consideration for potential confounding. The consistent pattern of
association observed across the multiple comparisons made reduces the likelihood of
finding associations due to chance alone.
Strengths of the 1999-2004 NHANES analysis (age > 40 years, N = 4,663-7,342)
included the examination of several potential confounding factors, multiple exposures
and outcomes in predominately healthy adults and statistical analyses to provide
nationally-representative results. Higher concurrent blood Pb level was associated with
higher serum inflammation markers, C-reactive protein (CRP), fibrinogen, and white
blood cell (WBC) count among men (Songdej et al.. 2010). Results were adjusted for
age; sex; race/ethnicity; education; current income; physical activity; and several factors
related to inflammation including, BMI, smoking status, and history of diabetes or
inflammatory or cardiovascular disease. For women, most ORs for associations between
quintiles of blood Pb and tertiles of CRP, fibrinogen, and WBC count were <1.0 whereas
corresponding ORs in men mostly were >1.0. For example, compared with men with
concurrent blood Pb level <1.16 (ig/dL (detection limit: 0.6 (ig/dL), men with blood Pb
4-514
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levels of 1.16-<1.63 ng/dL, 1.63-<2.17, 2.17-<3.09 ng/dL, and > 3.09 (ig/dL had
elevated odds of higher CRP (OR [95% CI]: 2.22 [1.14, 4.32], 1.67 [0.85, 3.28], 2.12
[1.07, 4.21], and 2.85 [1.49, 5.45], respectively). Although the highest ORs were found in
the highest blood Pb quintile (> 3.09 (ig/dL), monotonic concentration-dependent
increases were not observed. Consistent with NHANES findings, higher concurrent blood
Pb level was associated with higher levels of WBCs and IL-6 with adjustment for age,
BMI, and current smoking status among 300 university students in Incheon, Korea (Kim
et al.. 2007). Adults with allergic conditions or using anti-inflammatory medication were
excluded; however, sufficient information was not reported to assess potential selection
bias. Larger effects were estimated for the 147 men in the upper two quartiles of blood Pb
levels, 2.51-10.47 (ig/dL than for the full range of blood Pb levels (N = 300).
Low blood Pb levels also were associated with inflammation in a small genome-wide
association study that included 37 children with autism and 15 children without autism
(ages 2-5 years; blood Pb levels: 0.32-5.2 (ig/dL, detection limit: 0.06 (ig/dL) in
California. The subjects' ages overlap with those associated with peak blood Pb levels,
reducing the likelihood of subjects having much higher past blood Pb levels. In models
that included age, sex, and autism diagnosis, concurrent blood Pb level was associated
with the expression of several immune and inflammation genes, including HLA-DRB and
MHC Class II-associated invariant chain CD74 (influences antigen presentation) (Tian et
al.. 2011). Blood Pb levels were similar between children with and without autism.
Correlations with gene expression were observed in both groups, but they were in
opposite directions (positive in children with autism and negative in children without
autism). With gene expression arrays, there is a higher probability of finding associations
by chance. Further, there was limited consideration for potential confounding, and the
representativeness of findings in children with autism may be limited. However, the
results are supported by observations that Pb chloride (10-100 (iM) increases MHC
expression in mouse and human HLA APCs (Guo et al., 1996a; McCabe and Lawrence.
1991V
In summary, Pb-associated increases in indicators of inflammation such as CRP, WBCs,
and IL-6 were found in populations mostly comprising healthy adults with concurrent
blood Pb levels 1.16-10 (ig/dL with adjustment for several potential confounding factors
(Songdej et al., 2010; Kim et al.. 2007). The analysis of adults participating in NHANES
was particularly noteworthy for its representative population and adjustment for age, sex,
race/ethnicity, education, current income, physical activity, and several inflammatory
conditions (Songdej et al.. 2010). Because all evidence was based on cross-sectional
analyses, the relative contributions of recent and past Pb exposures to the observed
associations are uncertain. Other lines of evidence do not strongly support reverse
causation as the primary explanation for blood Pb-inflammation associations given the
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multiple routes of Pb exposure; however, the temporal sequence between Pb exposure
and inflammation is difficult to establish. Despite the limited extent and cross-sectional
nature of epidemiologic evidence, the biological plausibility is provided by findings of
Pb-induced increases in Th2 cell partitioning (Section 4.6.2.1) and IL-6 (Section 4.6.6.1)
in toxicological studies. Th2 cells produce IL-6 which is the primary stimulus for
expression of CRP and fibrinogen (Hage and Szalai. 2007; Fuller and Zhang. 2001).
4.6.5 Immune-based Diseases
4.6.5.1 Decreased Host Resistance
The capability of Pb to reduce host resistance of animals to bacteria has been recognized
for almost 40 years and was supported by several animal studies described in the
2006 Pb AQCD. These studies demonstrated increased mortality in animals following Pb
exposure through drinking water and infection with Listeria monocytogenes. Multiple
investigations in the same laboratory indicated increases in body burdens of viable
bacteria, mortality, and sickness behavior induced by Listeria exposure in juvenile or
adult BALB/c or CBA/J mice exposed postnatally to 500 to 2,000 (iM Pb acetate in
drinking water for 3 to 8 weeks (Dvatlov and Lawrence. 2002; Kim and Lawrence. 2000;
KishikawaetaL 1997; Lawrence. 198 Ib). Decreased bacterial resistance was observed
in mice with blood Pb levels (upon cessation of Pb exposure) relevant to this ISA,
i.e., 25 (ig/dL in BALB/c mice exposed PND1-PND22 (Dvatlov and Lawrence. 2002)
and 20 (ig/dL in adult C3H/HeN mice with 16-week exposure (Fernandez-Cabezudo et
al.. 2007). Other studies found mortality from Salmonella or E. coli or reduced clearance
of Staphylococcus in mice or rats administered Pb acetate or nitrate via injection, a route
of Pb exposure less relevant to humans (Bishavi and Sengupta. 2006; Cook et al.. 1975;
Hemphill et al., 1971; Serve et al.. 1966). Although not examined as much, increased
mortality from viral infection was found in mice and chickens with postnatal dietary Pb
(mostly Pb acetate) exposure for 4-10 weeks (Gupta et al., 2002; Youssef etal.. 1996;
Exon et al.. 1979; Thind and Khan. 1978; Gainer. 1977). These effects were observed in
animals with high blood (71-313 (ig/dL) (Gupta et al.. 2002; Thind and Khan. 1978) or
tissue (0.12-0.71 ppm) (ExonetaL 1979) Pb levels.
The mode of action for Pb-induced decreased host resistance is well characterized by
observations that Pb suppresses Thl-driven acquired immune responses and increases
inflammatory responses in target tissue, which may compromise host protective barriers.
Host resistance to bacteria such as Listeria requires effective Thl-driven responses
including the production of IL-12 and IFN-y (Lara-Tejero and Pamer. 2004) and these
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have been found to be inhibited by Pb exposure (Section 4.6.6.1). The lack of IFN-y can
inhibit appropriate and timely macrophage activation. Strengthening the evidence for
Pb-induced decreased host resistance, Fernandez-Cabezudo et al. (2007) found both
increased mortality and decreased production of IL-12 and IFN-y (ex vivo in spleen cells)
in Salmonella- plus Pb-exposed CH3/HeN mice with blood Pb levels of 20 (ig/dL. Nitric
oxide is produced by activated macrophages and has been found to be suppressed by Pb
exposure (Section 4.6.6.2). Pb-induced decreases in bacterial clearance have been found
in conjunction with reduced NO and macrophage functionality (Bishayi and Sengupta.
2006). Further, Pb-induced inflammation has been demonstrated as increases in ROS and
PGE2 (Section 4.6.6.3). Additional mode of action evidence was provided recently with
observations that developmental Pb exposure of BALB/c mice (100 (iM Pb acetate in
drinking water of dams from GD8 to PND21) upregulated splenic gene expression of
caspase-12 (Kasten-Jolly et al.. 2010). Caspase-12 is a cysteine protease that functions in
apoptosis and activation of pro-inflammatory cytokines and has been linked with a role in
the inhibition of bacterial clearance both systemically and in the gut mucosa (Saleh et al..
2006).
In the few available epidemiologic studies, various Pb exposure indicators (i.e., cord or
concurrent blood Pb, Pb content in total deposition samples or lichen) were associated
with viral and bacterial infections in children. An increase in infections was associated
with cord blood Pb levels > 10 (ig/dL in children in Boston, MA (n = 283) (Rabinowitz et
al.. 1990) and a mean concurrent blood Pb level of 3.34 (ig/dL in children in Germany
(n = 311) (Karmaus et al.. 2005). Similarly, a study found higher frequency of self-
reported colds or influenza among 66 Pb battery or smelter plant workers with blood Pb
levels 21.3-85.2 (ig/dL than among 53 controls with blood Pb levels 6.6-20.8 (ig/dL
(Ewers et al.. 1982). Because of the many limitations, including the lack of consideration
for potential confounding (Karmaus et al.. 2005; Rabinowitz et al.. 1990; Ewers et al..
1982). lack statistical rigor in comparisons of mean blood Pb levels by number of
infections (Karmaus et al.. 2005; Ewers et al.. 1982). and ecological design (Carreras et
al.. 2009). conclusions about the effects of Pb exposure on viral or bacterial infections
cannot rely on epidemiologic evidence and are based primarily on animal evidence. The
weak epidemiologic data do not detract from the consistent findings in animals for
Pb-induced decreased host resistance, including those in animals with relevant blood Pb
levels, and evidence for modes of action including decreased macrophage function and
Thl cytokine production.
With few studies available, the effect of Pb on resistance to eukaryotic parasites is not
clear. High concentration Pb acetate (> 30 \\M) diminished the ability of macrophages to
kill Leishmania enrietti protozoa parasites in vitro (Mauel et al.. 1989). Survival of
malaria-infected mice was enhanced with 100 (iM Pb nitrate exposure via drinking water
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(Koka et al.. 2007). which was attributed to Pb inducing eryptosis and removal of
infected erythrocytes and not to Pb-induced alterations in immune function. Nriagu et al.
(2008) found that higher concurrent blood Pb level was associated with lower malaria
prevalence among 653 children (ages 2-9 years) from three Nigerian cities with a mean
blood Pb level of 8.9 (ig/dL. Results were adjusted for age, sex, number of siblings, and
comorbidities such as depressed mood, headaches, and irritability. Given the well-
characterized effect of Pb in promoting Th2 activity, it is plausible for Pb to enhance host
resistance to parasites that require robust Th2 responses such as helminths. However, this
relationship has not been well examined.
4.6.5.2 Atopic and Inflammatory Conditions
Toxicological evidence and to a lesser extent, epidemiologic evidence, have supported
the effects of Pb exposure on stimulating Th2 activity, including increasing production of
Th2 cytokines such as IL-4 (Section 4.6.6.1). IgE antibody (Section 4.6.3). and
inflammation (Section 4.6.4). These endpoints comprise a well-recognized mode of
action for the development and exacerbation of atopic and inflammatory conditions such
as asthma and allergy. Thus, this mechanistic evidence provides support for the small
body of epidemiologic evidence indicating associations of blood Pb levels with asthma or
allergy in children (Figure 4-34 and Table 4-33). Whereas such evidence reviewed in the
2006 Pb AQCD was too sparse to permit conclusions, findings from recent studies add
supporting evidence. Children examined in studies of asthma and allergy encompassed a
wide age range (i.e., <1-12 years) and across studies, blood Pb was measured at different
lifestages and ages. Studies ascertained outcomes with parent report of doctor diagnosis
but also more objectively using a surveillance database or clinical testing. This variability
could contribute to between-study heterogeneity in results; however, the objective
assessment of outcomes in some studies is not likely to produce a spurious association.
Further, some of the evidence was provided by large studies that prospectively
ascertained outcome incidence after the measurement of blood Pb levels, did not indicate
selection bias, and considered potential confounding by SES and other environmental
exposures. These strengths reduce the likelihood of reverse causation and the influence of
other risk factors and increase confidence that the observed associations reflect a direct
relationship with Pb exposure.
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Study
Outcome
Blood Pb Level Mean
or Group (ug/dL)
Jedrychowski et al. (2011) Positive skin prick test prenatal: 1.16
concurrent: 2.04
Rabinowitz et al. (1990) Eczema
Joseph et al. (2005) Incident asthma
requiring medical
care
Rabinowitz et al. (1990) Asthma
Pugh Smith and Nriagu . ..
(2011) Asthma
>10vs. <10
Caucasian, >5vs. <5
African American, >5vs. <5
African American, >10 vs. <5
>10vs. <10
>10vs. <10
-O—*•
01 234567
Odds ratio (95% Clp
Note: Results are presented first for allergy-related outcomes then for asthma. All results are from prospective analyses, except for
Pugh Smith and Nriagu (2011). Black diamond represents associations with concurrent blood Pb levels, green triangles represent
associations with prenatal (cord) blood Pb levels, and blue circles represent associations with blood Pb levels measured in
childhood up to 12 months prior to outcome assessment. Odds ratios in Jedrychowski et al. (2011) are standardized to a 1 |jg/dL
increase in blood Pb level.
Figure 4-34 Associations of blood Pb levels with asthma and allergy in
children.
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Table 4-33 Additional characteristics and quantitative results for studies presented in Figure 4-34.
Study
Study Population and
Methodological Details
Blood Pb Level Data (ug/dL)
Outcome
Odds Ratio or
Relative Risk
(95% Cl)
Jedrychowski et 224 children followed prenatally to age 5 yr,
al. (2011) Krakow, Poland
Prospective. Study of multiple exposures and
outcomes. Clinical assessment of atopy. No
information on follow-up participation but no
selective attrition. Excluded subjects with cord
blood level > 2.5 ug/dL. Logistic regression
adjusted for sex, parity, maternal age, education,
and atopy, cord blood cotinine, smoker in home
during follow-up. Also considered potential
confounding by breastfeeding and allergen levels in
house dust.
Prenatal (cord) geometric mean: 1.16
(95% Cl: 1.12, 1.22)
Concurrent geometric mean: 2.04 (95%
Cl: 1.95,2.12)
Positive Skin Prick 2.3 (1.1, 4.6)a
Test
1.1 (0.7, 1.6)a
Rabinowitz et al. 159 children followed from birth to unspecified age,
(1990) Boston area, MA
Prospective. Low participation among eligibles. No
information on differences with nonparticipants.
Logistic regression with parental report of eczema.
No consideration of potential confounding factors.
Prenatal (cord) > 10 vs. <10
Mean not reported, approximate 90th
percentile = 10
Eczema
1.0(0.6, 1.6)"
Hon et al. (2010: 110 children with atopic dermatitis, mean (SD) age:
2009)° 9.9 (4.6) yr, Hong Kong, China,
Not clear whether subjects were free of atopic
dermatitis at time of blood Pb measurement.
Recruitment from dermatology clinic. Clinical
assessment of atopic dermatitis. No information on
participation rate. Examination of multiple metals.
Lack of rigorous statistical methods. No
consideration of potential confounding factors.
Serum Pb mean (SD): 1.86 (0.83)
Atopic dermatitis
severity
r= 0.329, p<0.001
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Table 4-33 (Continued): Additional characteristics and quantitative results for studies presented in Figure 4-34.
Study
Joseph et al.
(2QQ5)
Study Population and
Methodological Details
4,634 children, ages 4 mo-3 yr followed for
12 months, Southeastern Ml
Prospective. Large sample size. Indirect
assessment of asthma diagnosis but ascertained
Blood Pb Level Data (ug/dL)
Caucasian, > 5 vs. <5
African American, > 5 vs. <5
African American, > 10 vs. <5
Outcome
Incident asthma
requiring medical
care
Odds Ratio or
Relative Risk
(95% Cl)
2.7(0.9, 8.1)d
1.1 (0.8, 1.7)d
1.3(0.6, 2.6)d
from managed care organization claims database.
Logistic regression adjusted for sex, birth weight,
and average annual income available only at
census block level. Lack of information on other
environmental exposures.
Measured up to 12 mo before outcome
Mean (SD)
Caucasian: 3.2 (2.5)
African American: 5.5 (4.3)
Rabinowitz et al. 204 children followed from birth to unspecified age,
(1990) Boston area, MA
Same methodology as that described for eczema
above. Parental report of asthma.
Prenatal (cord) > 10 vs. <10
Mean not reported, approximate 90th
percentile = 10
Prevalent asthma 1.3 (0.8, 2.0)b
Pugh Smith and 356 children, ages 0-12 yr (71% age < 6 yr),
Nriagu (2011) Saginaw, Ml
Cross-sectional. Recruitment from state blood Pb
database. Moderate participation rate. Parental
report of asthma diagnosis. Logistic regression
adjusted for age, sex, family income, #stories in
unit, cat in home, dog in home, cockroach problem,
# persons in home, smoker in home, clutter,
highest blood Pb level at address, candles or
incense, months of residency, housing tenure,
cooking stove type, main heating source, air
conditioning type, peeling paint, ceiling/wall
damage, housing age, water
dampness/mildew/musty odor.
Blood Pb level > 10 vs. <10
Levels ascertained from statewide
database, specific timing unreported but
varied among subjects
Prevalent asthma
diagnosed within
previous
12 months
7.5(1.3, 42.9)b
Note: Results are presented first for allergy-related outcomes then for asthma. Within each category, results are grouped according to prospective or cross-sectional study design.
aOdds ratio presented per 1 ug/dL increase in blood Pb level.
bOdds ratio in children with blood Pb level > 10 ug/dL with children with blood Pb level <10 ug/dL serving as the reference group.
°Results are not included in Figure 4-34 because only correlations are presented.
dRelative risk in each specified group with children of the same race with blood Pb level <5 ug/dL serving as the reference group.
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Key evidence for Pb-associated effects on allergy-related outcomes was provided by a
prospective study in 224 children in Poland that compared associations of prenatal (cord
and maternal) and concurrent blood Pb levels with incidence of allergic sensitization at
age 5 years, as ascertained by skin prick test (SPT) (Jedrychowski et al.. 2011). Subjects
assessed at age 5 years did not differ from the full cohort, indicating lack of selective
attrition of subjects by blood Pb level or health status. The potential for selection bias
specifically by Pb exposure also is reduced because multiple exposures and outcomes
were examined in this cohort. A 1 (ig/dL increase in prenatal cord blood level was
associated with greater risk of positive skin SPT (rash/inflammatory reaction) to dust
mite, dog, or cat allergen with an RR of 2.3 (95%CI: 1.1,4.6). Concurrent blood Pb level
was only weakly associated with positive SPT (Figure 4-34 and Table 4-33). For prenatal
Pb biomarkers, similar effect estimates were obtained before and after adjustment for sex,
parity, maternal age, education, and atopy, and prenatal (cord blood cotinine) and
postnatal (smoker in the home) smoking exposure. Results were not altered by additional
adjustment for house dust allergen levels. Cord and concurrent blood Pb levels were
weakly correlated (r = 0.29), providing support for an independent association for
prenatal Pb biomarkers. A relationship with Pb was substantiated by observations that
indicators of other exposures, including blood levels of Hg, poly cyclic aromatic
hydrocarbon DNA adducts, and residential levels of dust mite or pet allergen were
associated with lower risks of SPT than was blood Pb level. These associations were
observed with relatively low cord blood Pb levels (geometric mean: 1.16 (ig/dL [95% CI:
0.12, 1.22], detection limit not reported). However, cord blood Pb levels reflect the
pregnancy blood Pb levels of mothers. Because adult blood Pb levels are influenced by
both recent and past Pb exposures, there is uncertainty regarding the specific Pb exposure
scenarios that contributed to associations between cord blood Pb level and allergic
sensitization in children examined in Jedrychowski et al. (2011).
Contrasting results were produced by prospective studies of eczema or atopic dermatitis,
which are reactions of the skin to sensitized allergens; however, neither study considered
potential confounding factors. Rabinowitz et al. (1990) found no elevated risk of parental
reported eczema in children (n = 159) in the Boston, MA area with cord blood Pb levels
> 10 (ig/dL. Hon et al. (2010; 2009) found a correlation between concurrent serum Pb
levels (mean: 1.86 (ig/dL, detection limit not reported) and clinically assessed atopic
dermatitis severity (e.g., skin area affected, intensity of rash and inflammation,
symptoms) in 110 children (mean age 10 years) in Hong Kong (Spearman r = 0.33,
p <0.005). Various other metals that were examined were negatively correlated with
atopic dermatitis. Although Hon et al. (2010; 2009) examined incidence of atopic
dermatitis, the subjects were patients referred to a dermatology clinic. The
representativeness of the children to the source population is uncertain, and allergy may
already have developed by the time serum Pb levels were measured.
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Prospective and cross-sectional evidence indicated associations of blood Pb levels with
asthma in children. Among prospective studies, Joseph et al. (2005) accounted for
potential confounding factors. Asthma-free children, ages 1-3 years, (N = 4,634) all
members of the same managed care organization in southeastern Michigan were selected
based on availability of blood Pb level data in a statewide database and then tracked for
the following 12 months for incidence of asthma. Incident asthma requiring a doctor visit
or medication was defined from the medical claims database as >4 asthma medication
dispensing events and >1 asthma emergency department visit, >1 hospitalizations, or >4
outpatient visits with at least two asthma medication dispensing events in the previous
12 months. While this definition is not a direct diagnosis, it is used to define persistent
asthma by the Healthcare Effectiveness Data and Information Set, which most U.S.
health plans use to measure health care performance. The records-based analysis
precluded bias due to selective participation of subjects by blood Pb level and health
status. However, because a blood Pb measurement was required, it is uncertain whether
the study population is representative of the managed care organization population. In
analyses that adjusted for average annual income at the census block level, birth weight,
and sex, an elevated risk of incident asthma requiring a doctor visit or medication was
associated with blood Pb levels > 5 (ig/dL in Caucasian children (RR: 2.7 [95% CI: 0.9,
8.1] compared with Caucasian children with blood Pb levels <5 (ig/dL) (Figure 4-34 and
Table 4-33). In analyses restricted to African Americans, children with blood Pb levels
> 10 (ig/dL had an elevated risk of incident asthma requiring medical care (RR: 1.3 [95%
CI: 0.6, 2.6] compared with children with blood Pb level <5 (ig/dL); however the
association was imprecise. There were small numbers of children with asthma requiring
medical care in the higher blood Pb level categories, which could have accounted for the
wide 95% CIs (5 Caucasian children with blood Pb > 5 (ig/dL and 9 African American
children with blood Pb level > 10 (ig/dL). In analyses that used Caucasian children with
blood Pb level <5 (ig/dL as the reference group, blood Pb level was associated with
increased risk of asthma requiring medical care among African American children in all
blood Pb level categories, which indicated a stronger association with race. Nonetheless,
results within the Caucasian group pointed to an association with blood Pb level. The
prospective study of 204 children in the Boston, MA area also found a Pb-associated
increased risk of parental-reported asthma (age of assessment not reported), specifically
in children with cord blood Pb levels >10 (ig/dL relative to those with cord blood Pb
levels < 10 (ig/dL (Rabinowitz et al.. 1990) (Figure 4-34 and Table 4-33). However,
potential confounding factors were not examined.
Supporting the prospective evidence, a cross-sectional study conducted in Saginaw, MI
found a higher prevalence of parental report of doctor-diagnosed asthma in children (71%
ages < 6 years) with blood Pb levels > 10 (ig/dL (Pugh Smith and Nriagu. 2011). Similar
to Joseph et al. (2005), the study population was predominately African American (78%
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of 356). Children were randomly selected from a statewide database of initial blood Pb
measurements collected at unspecified ages, and a positive bias is possible if parents of
children with higher blood Pb levels and asthma were more likely to participate or recall
an asthma diagnosis. Data were collected on asthma diagnosis in the previous 12 months;
thus, for some subjects, blood Pb measurement likely preceded asthma diagnosis. A
strength of this study was the adjustment for a large number of potential confounding
factors such as age, sex, household pets, housing characteristics, household smoking, and
family income. Compared with children with initial blood Pb levels <10 (ig/dL, children
with initial blood Pb levels > 10 (ig/dL had a higher odds of having a doctor diagnosis of
asthma within the past 12 months (OR: 7.5 [95% CI: 1.3, 42.9]). The results were
imprecise, and while this study had more children with blood Pb levels > 10 (ig/dL
(18.6%) than did Joseph et al. (2005). the analyses considered a much larger number of
covariates, some of which may be predictors of Pb exposure.
As discussed in Section 4.6.3.1. one may speculate that cross-sectional associations with
asthma and allergy could be attributed to reverse causation. Humans with asthma and
animal models of asthma have been shown to have epithelial cell damage and exudation
of cells and fluids into airways, which are signs of increased lung permeability. With lung
permeability, one may speculate the potential for greater uptake of Pb from airways into
blood. Some but not all evidence indicates greater lung clearance of very small particles
(99mTc-DTPA) or uptake into the blood of subjects with asthma than healthy subjects (Del
Donnoetal.. 1997; Reesetal. 1985; O'Byrne et al.. 1984; Elwoodetal.. 1983) and also
with airway inflammation associated with fibrosis, smoking, and pulmonary edema
[2009 PM ISA (U.S. EPA. 2009a)1. The available evidence does not inform Pb
absorption with asthma and allergy directly but provides some indication for higher
inflammation resulting in increased lung permeability to very fine particles. However, a
role for reverse causation may apply mostly to fine particle Pb and exposure via
inhalation, which is not found to be a significant route of exposure (Section 3.1). A
relationship between IgE-mediated inflammation and gastrointestinal permeability, a
more dominant route of Pb exposure, has not been indicated. These lines of evidence
combined with that from prospective studies in children and that characterizing modes of
action (i.e., increases in IgE, Th2 cytokines, and inflammation) do not suggest that the
associations observed for blood Pb level with asthma and allergy in children are due
primarily to reverse causation.
Among the studies in children that found associations of blood Pb level with asthma and
allergy, several adjusted for potential confounding by indicators of SES and other
environmental exposures. Joseph et al. (2005) adjusted for census block annual income,
which may represent family-level income with error. However, Pugh Smith and Nriagu
(2011). which examined a primarily low SES population of children in Michigan,
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adjusted for family annual income. In this study, a Pb-associated higher asthma
prevalence was found with adjustment for various additional factors associated with SES
and allergen exposure, including multiple indices of housing condition and presence of
pets and cockroaches in the home. Jedrychowski et al. (2011) adjusted for maternal
education, and found similar magnitudes of association between cord blood Pb level and
positive SPT as those in the unadjusted analysis. Further, residential levels of dust mite or
pet allergen were associated with lower risks of SPT than was blood Pb level. Blood Pb
level also was associated with asthma or allergy after adjusting for concurrent exposure
to smoking in the home (Jedrychowski et al., 2011; Pugh Smith and Nriagu. 2011). with
Jedrychowski et al. (2011) additionally adjusting for prenatal smoke exposure assessed as
cord blood cotinine levels. The studies varied in the specific confounding factors
considered, the method of measurement, and the method of control. Some studies
examined multiple potential confounding factors (Jedrychowski et al.. 2011; Pugh Smith
and Nriagu. 2011). which increases confidence that the associations observed with
asthma and allergy, reflect a relationship with Pb. However, in the small evidence base,
uncertainty remains regarding residual confounding, particularly by SES. While there is
no single complete measure of SES, these studies adjusted for different indicators of SES
that may vary in the adequacy of control for confounding. Residual confounding also is
possible by factors not examined.
Cross-sectional evidence did not strongly indicate associations of biomarkers of Pb
exposure with asthma or allergy in adults (Mendy etal.. 2012; Pizent et al., 2008; Bener
et al.. 200la). However, study limitations make the evidence less informative for drawing
conclusions about the effects of Pb on asthma compared with evidence in children.
Higher Pb level in a spot urine sample (geometric mean: 0.59 (ig/g creatinine, detection
limit not reported) was not associated with an increase in asthma prevalence (OR: 0.72
[95% CI: 0.46, 1.12] per 1 (ig Pb/g creatinine increase in urine) in the large U.S.
NHANES 2007-2008 analysis of 1,857 adults ages 20 years and older (Mendv et al..
2012). Associations were not found with other respiratory conditions such as emphysema
or chronic bronchitis either. This study adjusted for several potential confounding factors:
sex, race/ethnicity, education, income to poverty ratio, number of alcoholic drinks/day,
and current smoking status. However, spot urinary Pb level has an uncertain relationship
with long-term Pb exposure (Section 3.3.3). In a study of adult (ages 19-67 years) office
workers in Zagreb, Croatia, Pizent et al. (2008) found a concurrent blood Pb-associated
lower odds of positive SPT to common inhaled allergens among 50 men (median blood
Pb level: 3.2 (ig/dL) and lack of a statistically significant association among 166 women
(median blood Pb level: 2.2 (ig/dL), adjusting for age, smoking (current and history), and
number of alcoholic drinks/day. The findings in women appeared to be discordant
because there was an association between concurrent blood Pb level and serum IgE,
which commonly mediates the acute inflammatory response to allergens. However, the
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interpretation of the findings is difficult because only statistically significant effect
estimates were reported; thus, it is not known whether odds ratios were in the same
direction for SPT and IgE in women. Bener et al. (200la) found higher prevalence of
asthma and allergy-related conditions in 110 Pb industrial workers (mean age: 35.5 years)
than in 110 age-matched controls. However, the implications are limited because blood
Pb levels in both Pb-exposed workers and controls (geometric means: 77.5 and
19.8 (ig/dL, respectively) were higher than those in most of the current U.S. adult general
population, and potential confounding factors, including other occupational exposures,
were not evaluated.
In summary, studies support associations of higher blood Pb levels with asthma and
allergy in children (Jedrychowski et al., 2011; Pugh Smith and Nriagu. 2011; Joseph et
al.. 2005). Because of study limitations in the Pb biomarker analyzed and confounding
considered, the evidence in adults is largely uninformative (Mendy et al.. 2012; Pizent et
al.. 2008; Bener etal.. 200la). In children, evidence was limited to a few populations.
Because of the heterogeneity in the relatively small body of evidence, it was difficult to
identify whether the strength of association with asthma and allergy differed by age of
children, lifestage or time period of blood Pb measurement (prenatal, sometime in
childhood prior to outcome assessment, concurrent), or blood Pb level. The few
prospective analyses examining incidence of allergy or asthma (Jedrychowski et al..
2011; Joseph et al.. 2005) do not suggest that the observed associations are due to reverse
causation. Some but not all evidence links asthma and other inflammatory conditions
with increased lung permeability to very small particles, but evidence is lacking for a
relationship with increased gastrointestinal permeability, the more dominant ingestion
route of Pb exposure. These various lines of evidence indicate that associations observed
between blood Pb level and asthma and allergy in children are not attributable primarily
to reverse causation. In the epidemiologic studies, the lack of selective participation of
subjects and objective assessment of outcomes indicates lack of biased reporting of
asthma and allergy for children with higher blood Pb levels. Although not affecting
internal validity, in some studies, the recruitment of subjects from blood Pb surveillance
databases may limit generalizability of findings. The adjustment for maternal education
and exposure to smoking or allergens in Jedrychowski et al. (2011) and family income,
smoking, housing conditions, pets, or pests in Pugh Smith and Nriagu (2011) increase
confidence that the associations observed in these studies are not solely due to
confounding by SES, smoking, or allergen exposure. Further, biological plausibility is
well supported by evidence describing modes of action for asthma and allergy, including
Pb-associated increases in IgE (Section 4.6.3). Th2 cytokines (Section 4.6.6.1). and
inflammation (Section 4.6.4). However, because there are few studies of asthma and
allergy, uncertainty remains regarding residual confounding, particularly by SES, which
was examined with different indicators across studies.
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4.6.5.3 Other Respiratory Effects
Potential relationships between Pb exposure and respiratory effects such as bronchial
responsiveness (BR), pulmonary function or morphology, and respiratory symptoms were
not well characterized in the 2006 Pb AQCD (U.S. EPA. 2006b) but have been examined
recently in a small number of studies. Because these outcomes are not specific to atopic
and inflammatory conditions, they are evaluated separately here and in Section 4.9.6.
Associations between blood Pb level and indicators of lung function in adults are
inconsistent. Studies were cross-sectional, included similarly aged subjects, and
considered similar confounding factors. Increased BR is a characteristic feature of asthma
and other respiratory diseases and can result from the activation of innate immune
responses and increased airway inflammation. In the larger study of 523 adults (ages
19-58 years) in Seoul, Korea, Min et al. (2008a) found an association between concurrent
blood Pb level and BR. A 1 (ig/dL higher concurrent blood Pb level was associated with a
higher BR index (log [(% decline in forced expiratory volume in 1 second [FEVi]/log of
final methacholine concentration in mg/dL)+10]) of 0.018 (95% CI: 0.004, 0.03), with
adjustment for age, sex, height, smoking status, baseline FEVi, and presence of asthma
(Min et al.. 2008a). The concurrent blood Pb levels in these adults were low (mean [SD]:
2.90 [1.59] (ig/dL); however, it is uncertain what timing, level, frequency, and duration of
Pb exposures contributed to the observed association. In contrast to Min et al. (2008a),
Pizent et al. (2008) found that higher concurrent blood Pb level was associated with
lower BR in 47 men (decrease of 2.4 [95% CI: 0.52, 4.2] in log of percentage change in
FEVi post-histamine challenge/cumulative dose histamine per log increase in blood Pb
level adjusted for age and serum Se). Smoking intensity and alcohol consumption were
excluded as covariates by stepwise regression. Similarly, among these men, higher blood
Pb level was associated with lower odds of positive SPT.
Pb-associated respiratory effects were not clearly indicated in adults with occupational Pb
exposures either. However, the lack of direct analysis of blood Pb levels and
consideration for potential confounding limit the utility of this evidence in drawing
conclusion about the respiratory effects of Pb. In bus drivers (mean age: 46 years) in
Hong Kong (Jones et al.. 2008; Jones et al.. 2006). 129 drivers of non-air conditioned
buses had lower exposures to PMi0, lower blood Pb levels (mean 3.7 (ig/dL versus
5.0 (ig/dL in air conditioned buses) but lower indices of lung function than did 358
drivers of air conditioned buses (Jones et al.. 2006). The authors attributed the slightly
higher blood Pb levels of air conditioned bus drivers to the poor efficiency in the filters
resulting in higher PM10 levels on those buses. Blood Pb levels and various lung function
parameters were similar between 33 roadside vendors and 31 adjacent shopkeepers
(respectively, mean ages: 45.1 and 42.8 years and mean blood Pb levels: 5.61 and
5.14 (ig/dL) (Jones et al.. 2008). Pb industrial workers (N = 100, mean age: 34.6 years) in
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the United Arab Emirates had higher prevalence of respiratory symptoms such as cough,
phlegm, shortness of breath, and wheeze than did 100 age- and sex-matched unexposed
controls (Bener et al., 200la). Blood Pb levels in both the Pb industrial workers and the
control group (geometric means: 77.5 and 19.8 (ig/dL, respectively) were higher than
those in most of the current U.S. adult general population.
An effect of Pb specifically on the lung was demonstrated in a recent study of Wistar rats
with low blood Pb levels (2.1 (ig/dL) but produced by Pb acetate given by injection
(25 mg/kg, 3 consecutive days) (Kaczynska et al., 2011). The lungs of Pb-treated rats
exhibited pulmonary fibrosis, epithelial cell damage, an increase in mast cells, increased
recruitment of monocytes and thrombocytes into capillaries, and increased macrophage
accumulation in the alveolar space. While these pulmonary changes have been linked
with functional pulmonary decrements and inflammation in other studies (unrelated to Pb
exposure), the implications are uncertain because the results were obtained with a route
of Pb exposure less relevant to those in humans.
Air-Pb Studies
The 2006 Pb AQCD (U.S. EPA. 2006b) did not review respiratory effects studies that
represented Pb exposure by Pb measured in ambient air samples. However, there are
studies that have examined relationships between respiratory effects and airborne Pb by
analyzing the Pb component of PMi0 or PM25 air samples individually or as part of a
group of correlated components within these samples using source apportionment or
principal component analysis. Daily ambient air Pb-PM concentrations were associated
with daily respiratory morbidity in children (Gent et al.. 2009; Hong et al.. 2007b). Gent
et al. (2009) found that increases in Pb-PM2 5 (same-day or 3-day average) were
associated with increases in respiratory symptoms and asthma medication use in 149
children with asthma in Southern New England (ages 4-12 years), adjusting for
season, day of the week, and date. Hong et al. (2007b) found that an increase in previous-
day Pb-PM 10 was associated with a decrease in lung function in 43 mostly healthy
children in Korea, adjusting for age, sex, height, weight, household smoking, and
weather. Toxicological studies found Pb-containing CAPs to induce pulmonary
inflammation. Uzu et al. (2011) found that Pb-rich PM from a Pb recycling plant
increased the release of the cytokine granulocyte-macrophage colony-stimulating factor
from human epithelial cells. Pulmonary inflammation was found in animals exposed to
CAPs in which Pb was one of numerous components (Wei et al., 2011; Duvall et al..
2008: Godleski et al.. 2002: Saldiva et al.. 2002).
As with blood Pb, ambient air Pb-PM2 5 concentrations were not consistently associated
with respiratory effects in older adults with adjustment for weather and temporal trends.
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Among adults ages 65 years and older in 6 California counties, an increase in Pb-PM2 5
was associated with an elevated risk of respiratory mortality 3 days later in an all-year
analysis (RR: 1.01 [95% CI: 0.99, 1.03] per 4 ng/m3 increase in Pb-PM25) and summer-
only analysis (RRnot reported) (Ostro et al.. 2007). However, among adults ages 65
years and older in 106 U.S. counties, Bell et al. (2009) found that an increase in same-day
Pb-PM25 was associated with a decrease in respiratory hospital admissions.
Despite limited available results consistent with a relationship between respiratory effects
in children and short-term (over a few days) changes in ambient air Pb-PM
concentrations, uncertainties limit the utility of these findings in evaluating Pb-associated
respiratory effects. Few data on size distribution of Pb-PM are available, so it is difficult
to assess the representativeness of these concentrations to population exposure
(Section 2.5.3). Moreover, few data are available on the relationship between blood Pb
and air Pb for the varying Pb-PM size distributions (see Section 3.5.1). In several air-Pb
studies, other PM components such as elemental carbon, copper, and zinc also were
associated with respiratory effects. In the absence of data on correlations among PM
components, measurements on co-occurring ambient pollutants, or results adjusted for
copollutants, it is difficult to exclude confounding by ambient air exposures to other PM
components or ambient pollutants. In several studies that analyzed PM component
mixtures, of which Pb-PM comprised one component, it is not possible to attribute the
observed associations or lack of associations specifically to Pb (Sarnat et al.. 2008;
Andersen et al., 2007; Veranth et al., 2006; Maciejczyk and Chen. 2005).
In summary, while a few studies found air Pb-PM to be associated with respiratory
effects in children, limitations of these recent results include potential confounding by
other air pollutants and the uncertain representativeness of Pb-PM to population
exposures. In adults, neither blood Pb nor air Pb-PM was consistently associated with
respiratory effects. Blood Pb studies of nonoccupationally-exposed adults were similar in
cross-sectional design, age of subjects, potential confounding factors examined, and
examination of respiratory endpoints that can exhibit short-term changes. Studies of
Pb-exposed workers were similarly limited by lack of rigorous statistical analysis with
blood Pb levels and lack of consideration for potential confounding factors, including
other occupational exposures.
4.6.5.4 Autoimmunity
Autoimmunity is an immune response against self (e.g., generation of antibodies against
self antigens) and is linked with diseases such as lupus and rheumatoid arthritis. Evidence
for the effects of Pb on increasing the risk of autoimmunity is provided largely by animal
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studies reviewed in the 2006 Pb AQCD in which pre- or post-natal Pb acetate treatment,
most by injection, was associated with the generation of auto-antibodies in juvenile and
adult animals (Hudson et al., 2003; Bunn et al., 2000; El-Fawal et al., 1999; Waterman et
al.. 1994). El-Fawal et al. (1999) found elevated auto-antibodies in F344 rats with blood
Pb levels 11-50 (ig/dL. Results did not indicate whether the mechanism involved a shift
toward Th2 or Thl responses. While recent studies did not examine production of auto-
antibodies by Pb, some provided indirect evidence by showing the potential of Pb to
induce formation of neo-antigens which in turn could induce the formation of auto-
antibodies. For example, Kasten-Jolly et al. (2010) found that developmental Pb acetate
exposure of BALB/c mice (100 (iM in drinking water, GD8-PND21) upregulated genes
for digestive and catabolizing enzymes, which could lead to the generation of self-
peptides, which combined with other Pb-induced immune effects, had the potential to
induce generation of auto-antibodies. The potential for auto-antibody generation also was
indicated by the activation of neo-antigen-specific T cells in adult BALB/c mice injected
once with 25-50 (ig Pb chloride (Carey et al.. 2006). Evidence of Pb-associated
autoimmune responses in humans is limited to findings of higher levels of IgM and IgG
auto-antibodies to neural proteins in male battery-plant workers (n = 20, 56) with blood
Pb levels 10 to 40 (ig/dL compared with controls (n = 7, 15, blood Pb levels not reported)
(El-Fawal et al.. 1999). Pb workers and controls were matched by demographic and SES
characteristics, but potential confounding by other workplace exposures was not
examined. Consistent with findings in Pb-exposed workers, modified neural proteins
were found in CBA/J rats injected with native protein altered by Pb acetate in vitro
(Waterman et al.. 1994).
4.6.5.5 Tumors
Toxicological evidence indicates that high concentration Pb exposures directly promote
tumor formation or induce mutagenesis and genotoxicity and is evaluated in detail in
Section 4.10. Kerkvliet and Baecher-Steppan (1982) provided evidence for involvement
of the immune system. Postnatal exposure of 6-8 week old male C57BL/6 mice to 130
and 1,300 ppm Pb acetate in drinking water for 10-12 weeks transiently enhanced
moloney sarcoma virus-induced tumor growth compared with control animals but did not
prevent subsequent tumor regression. The Pb-induced tumor growth was accompanied by
impaired macrophage phagocytosis (indicating suppressed Thl responses) but not
cytotoxicity. Cancer promotion is a relatively common outcome in chemical-induced
immunotoxicology, particularly when early life exposures are involved (Dietert. 2011).
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4.6.6 Modes of Action for Pb Immune Effects
4.6.6.1 Cytokine Production
As referenced in preceding sections, cytokines are signaling molecules that affect
immune cell function. For example, IL-4 induces B cells into IgE-producing cells, and
IFN-y induces macrophage recruitment and antigen presenting activity. The
2006 Pb AQCD (U.S. EPA. 2006b) presented a large body of evidence that clearly
demonstrated that pre- or post-natal Pb exposure of animals such as rodents and chickens
suppressed the production of Thl cytokine IFN-y and/or increased production of Th2
cytokines such as IL-4 [Table 5-7 of the 2006 Pb AQCD (U.S. EPA. 2006h)1. The
combined evidence for Pb-induced cytokine changes in multiple cell types, including T
cells and macrophages, indicates a shift of acquired immunity responses away from Thl
responses and toward Th2 responses. In turn, the Thl to Th2 shift provides mode of
action support for the associations of Pb with downstream events (Figure 4-34) such as
inflammation, ROS production, impaired macrophage function, decreased host resistance
observed primarily in toxicological studies, increased IgE production observed in both
epidemiologic and toxicological studies, and asthma and allergy observed in
epidemiologic studies. Previous toxicological studies found Pb to affect cytokine
production via action on T cells and macrophages, and a recent study provided new
evidence that Th2 skewing may be mediated via effects on dendritic cells.
Many studies found a shift to Th2 cytokine production in animals with long-term
(>4 weeks) dietary Pb exposure, and in some studies, the effects of prenatal exposure on
cytokine production persisted to the adult lifestage (Chen et al., 2004; Miller et al., 1998).
In the few studies that measured blood Pb levels shortly after cessation of Pb exposure
(gestational plus postnatal or postnatal only), higher IL-4 and/or lower IFN- y were found
in rodents with relevant blood Pb levels, means 6.75 and 17 (ig/dL (Chen et al.. 2004;
Dyatlov and Lawrence. 2002). Some studies found an increase in IL-4 or decrease in
IFN-y concomitantly with an increase in IgE (Heo et al.. 1997; 1996) or decrease in host
resistance (Fernandez-Cabezudo et al.. 2007) further supporting changes in cytokine
production as a mode of action for Pb-induced effects on downstream immune endpoints.
A recent study found a shift to Th2 cytokine production in mice over a wide range of Pb
exposures and provided evidence of effects at lower Pb exposures and a concentration-
dependent relationship. In this study, lifetime (gestation through adulthood) exposure of
Swiss mice to 0.06-400 ppm dietary Pb acetate produced blood Pb levels (upon
termination of exposure) of 1.23 to 61.48 (ig/dL (lavicoli et al., 2006b). For IL-2 and
IL-4, nonlinear concentration-response relationships were found, with the largest
decrease and increase, respectively, found between animals with blood Pb levels of
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0.8 (ig/dL (0.02 ppm Pb acetate, controls) and 1.23 (ig/dL. A linear concentration-
dependent decrease in IFN-y was observed in animals with blood Pb levels
0.8-61.48 (ig/dL. Although examined in few in vivo studies, increases in IL-6 and IL-10,
also Th2 cytokines, were reported in juvenile and adult rodents exposed to Pb
gestationally or postnatally in drinking water [Table 5-7 of the 2006 Pb AQCD (U.S.
EPA.2006h)1.
In vitro studies also reported a Pb-induced shift to production of Th2 cytokines. In
concordance with other indicators of Th2 skewing (i.e., suppressed DTH) in BALB/c
mice treated with Pb-exposed dendritic cells, Gao et al. (2007) observed that 10-day
25 (iM Pb chloride exposure lowered the ratio of IL-12:IL-10 production by dendritic
cells in vitro. Pb did not affect dendritic cell production of IL-6, IL-10, or TNF-a;
however, in co-cultures of Pb-treated dendritic cells and T cells, most results indicated
that dendritic cells stimulated T cells to produce Th2 cytokines. For example, although T
cell production of the Thl cytokine IL-2 increased, production of Th2 cytokines, IL-6
and IL-10 increased. Further, Pb-treated dendritic cells increased IL-4 production in
OVA-specific T cells, indicating that Pb affected the antigen presenting cell function of
dendritic cells. In another in vitro study, 24-hour Pb acetate exposures of 0.15 (ig/dL and
higher suppressed expression of Thl cytokines, IFN-y, IL-1(3, and TNF-a, and increased
secretion of Th2 cytokines, IL-5, IL-6, and IL-10 in cultures of human PMNs activated
with Salmonella enteritidis or with monoclonal antibodies of CDS, CD28, and CD40,
(Hemdan et al.. 2005).
Several toxicological studies found Pb-induced increases in the cytokine TNF-a, in some
cases, specifically from macrophages (Khan et al.. 2011; Cheng et al.. 2006; Flohe et al..
2002; Zelikoff et al.. 1993). This provides mode of action support for toxicological
evidence indicating Pb-induced decreases in resistance to bacterial infection since TNF-a
is produced primarily by activated macrophages, is increased in response to infection, and
induces inflammation. Among the in vivo studies, increases in TNF-a were found with
prenatal dietary Pb acetate exposure (250 ppm) of F344 rats (Chen et al., 1999; Miller et
al.. 1998). postnatal Pb oxide air exposure (31 (ig/m3, 3 hours/day, 4 days) of rabbits
(Zelikoff etal.. 1993). and postnatal i.p. Pb acetate treatment (5.0 mg) of Swiss mice
(Dentener et al.. 1989). The effects of prenatal dietary Pb exposure were found to persist
to adulthood. In animals, the Pb-induced increases in TNF-a were accompanied by
functional changes in host responses such as decreased macrophage phagocytosis
(Zelikoff etal.. 1993). suppressed DTH (Miller etal.. 1998). and increased mortality to
E.coli endotoxin (Dentener et al.. 1989). Blood Pb levels of animals were infrequently
reported; however, Chen et al. (1999) found increased TNF-a in rats with embryonic
blood Pb levels of 149 (ig/dL. In addition to finding Pb-induced increases in TNF-a
(Khan etal.. 2011; Gao et al.. 2007; Cheng et al.. 2006; Flohe et al.. 2002; Krocova et al..
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2000; Guo etal.. 1996a). some in vitro studies provided mechanistic explanation by
finding that Pb acetate or chloride (10-50 (iM, 1.5 hours-10 days) induced
phosphorylation of mitogen-activated protein kinase (MAPK) signaling molecules (Khan
etal.. 2011: Gao et al.. 2007: Cheng et al. 2006V Further, Cheng et al. (2006) found that
blocking protein kinase C or MAPK reduced TNF-a production by macrophages in vitro,
which in turn, protected against Pb acetate + LPS-induced liver injury in A/J mice.
The few available epidemiologic studies that examined cytokines found higher
concurrent blood Pb levels in children and adults to be associated with higher Th2
cytokine and lower Thl cytokine levels in serum. The epidemiologic evidence overall
was based on cross-sectional analyses, which precludes identifying the temporal
sequence between Pb exposure and cytokine changes. Other limitations include the lack
of rigorous statistical analysis and limited consideration of potential confounding.
Because of these limitations, the epidemiologic evidence is not a primary consideration in
drawing conclusions about Pb-associated cytokine changes. However, it does not detract
from the compelling toxicological evidence. Among children ages 9 months to 6 years in
Missouri recruited from a public assistance or Pb poisoning prevention program, Lutz et
al. (1999) found that 8 children with concurrent blood Pb levels 15-19 (ig/dL had 4-5 fold
higher serum levels of IL-4 (p = 0.08, Kruskal Wallis) and 3-fold higher IgE
(Section 4.6.3) than did 90 children with lower blood Pb levels. IL-4 levels in 9 children
with blood Pb levels 20-44 (ig/dL were lower than those in 90 children with blood Pb
levels <15 (ig/dL. The elevated IL-4 and IgE in children with blood Pb levels
15-19 (ig/dL were consistent with the action of IL-4 to activate B cells to induce class
switching to IgE. In another study of 214 children in grades 5 and 6 in Taiwan,
investigators compared cytokine levels not by blood Pb level groups but by potential for
Pb exposures due to age of home and location of residence (Hsiao etal.. 2011). Elevated
concurrent blood Pb levels were found only among 64 children living near an oil refinery,
in particular, among 34 children with known respiratory allergies (mean: 8.8 (ig/dL,
versus 3.2-3.8 (ig/dL in urban and rural groups). Children with allergies near the oil
refinery also had the lowest serum levels of IFN-y (45-fold) and highest levels of IL-4
(6-fold) (p <0.05 for comparisons with any subgroup). While the results suggested that
residence near the oil refinery contributed to differences in cytokine levels between
healthy and allergic children, they do not identify an independent effect of Pb, other
exposures or co-occurring factors, or a combination of factors.
Evidence of an association between blood Pb levels and cytokine levels in adults is
unclear. However, a relatively large study that considered potential confounding found an
association. Sufficient information was not reported to assess potential selection bias.
Among 300 (93% male, mean age 24 years) healthy university students in Incheon,
Korea, higher concurrent blood Pb level was associated with higher serum levels of TNF-
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a and IL-6, adjusting for age, BMI, and smoking status (Kim et al., 2007). Associations
were larger in magnitude among the 147 males in the upper two quartiles of blood Pb
levels, 2.51-10.47 (ig/dL. Associations with these blood Pb levels may reflect
contributions of higher past Pb exposures. A 1 (ig/dL higher blood Pb level was
associated with a 0.75 (95% CI: 0.14, 1.36) pg/mL higher TNF-a and a 0.18 (95% CI:
-0.02, 0.38) pg/mL higher IL-6. The association between levels of blood Pb and plasma
TNF-a was greater among men who were GSTM1 null (n = 77) than men who were
GSTM1 positive and men who had the TNF-a GG genotype (n = 131) than men who had
the GA or AA genotype. For the association between blood Pb level and plasma IL-6, the
effect estimate was slightly elevated in TNF-a GG genotype but similar between GSTM1
genotypes. In this study, there were multiple comparisons, but a consistent pattern of
association was observed across the immune endpoints examined. Subgroup analyses are
prone to higher probability of findings by chance, but these analyses had fairly large
sample sizes, and some results had biological plausibility. Pb has been shown to increase
ROS (Section 4.2.4). and cytokine expression has been shown to be modulated by ROS-
sensitive transcription factors. Thus, it is biologically plausible that the null variant of
GSTM1, which is associated with reduced elimination of ROS, may be associated with
increased cytokine levels. The results for the TNF-a variant are difficult to interpret. The
GG genotype is associated with lower expression of TNF-a, but the literature is mixed
with respect to which variant increases risk of inflammation-related conditions.
Much smaller studies of adults that did not consider potential confounding, did not report
quantitative results but only reported lack of statistically significant correlations between
concurrent blood Pb level and serum Th2 and Thl cytokine levels in men (n = 17 with
and 17 without allergy, ages 19-52 years, median blood Pb levels: -11 (ig/dL) (Boscolo
et al., 1999) and women in Italy (n = 23 with and 25 without allergy, ages 19-49 years,
median blood Pb levels: 6.4 and 5.5 (ig/dL, respectively) (Boscolo et al.. 2000).
Results from studies of occupationally-exposed adults also suggested that Pb exposure
may be associated with decreases in Thl cytokines and increases in Th2 cytokines (Di
Lorenzo et al.. 2007; Valentino et al.. 2007; Yiicesov et al.. 1997a). Valentino et al.
(2007) had the most rigorous statistical methods comprising regression analyses with
adjustment for age, BMI, smoking status, and alcohol consumption status but not other
occupational exposures. Regression coefficients describing the concentration-response
functions were not reported; however, 44 male foundry workers in Italy (mean blood Pb
levels: 21.7 (ig/dL) and 14 pottery workers (mean blood Pb level: 9.7 (ig/dL, ages of all
workers 30-61 years) had higher plasma IL-10 (ANOVA, p <0.05) than did the 59
unexposed controls (mean blood Pb level: 3.9 (ig/dL, ages 25-61 years). Levels of Th2
cytokines IL-2, IL-6, and IL-10 also increased from the lowest to highest blood Pb group
(ANOVA, p >0.05). In contrast with most other studies, both exposed worker groups had
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lower IL-4 levels compared with controls (ANOVA, p >0.05). In Yiicesoy et al. (1997a).
serum levels of the Thl cytokines, IL-1(3 and IFN-y, were lower in 20 Pb-exposed
workers (mean blood Pb level: 59.4 (ig/dL, ages 19-49 years) than in the 12 age-matched
controls in Turkey (mean blood Pb level: 4.8 (ig/dL). Some (Di Lorenzo et al.. 2007;
Valentino et al.. 2007) but not all (Yiicesoy et al.. 1997a) studies found higher serum
TNF-a in Pb-exposed workers. DiLorenzo et al. (2007) found a monotonic increase from
the 28 unexposed (blood Pb levels not reported, mean age: 48.2 years) to 17 intermediate
worker (blood Pb: 9.1-29.4 (ig/dL) and 19 high worker (blood Pb: 29.4-81.6 (ig/dL)
blood Pb level groups (mean age of Pb-exposed workers: 45.3 years) in Italy. Results also
indicated a potential interaction between blood Pb level and smoking. Among current
smokers (n = 9 to 20), a 13- to 16-fold difference in TNF-a levels was observed among
blood Pb groups. Among nonsmokers (n = 2 to 8), the differences were less than two
fold.
In summary, the few epidemiologic studies indicate associations of higher concurrent
blood Pb level with higher levels of IL-4 and/or lower levels of IFN-y in children (Hsiao
et al.. 2011; Lutzetal.. 1999) and occupationally-exposed adults (Di Lorenzo et al..
2007; Valentino et al.. 2007; Yiicesoy et al.. 1997a). Because quantitative results were
not reported in each study of nonoccupationally-exposed adults, implications of findings
are difficult to assess. Limitations of the epidemiologic evidence overall include the
cross-sectional design of studies and lack of rigorous statistical analysis that considered
potential confounding factors. Sufficient data were not reported to assess potential
selection bias. Because of the many limitations, the epidemiologic evidence alone is not
used to draw conclusions about Pb-associated cytokine changes. However, they are
useful in supporting the relevance of toxicological evidence to humans. Biological
plausibility for an effect of Pb on cytokine production is provided by a large body of
toxicological evidence that clearly demonstrates a Pb-induced shift to a Th2 phenotype
with increases in the Th2 cytokine IL-4 and decreases in the Thl cytokine IFN-y. Several
of these observations were made in juvenile and adult animals exposed prenatally or
postnatally via diet that resulted in blood Pb levels (upon cessation of Pb exposure)
relevant to this ISA, 1.23-17 (ig/dL qavicoli et al.. 2006b: Chen et al.. 2004: Dvatlov and
Lawrence. 2002). While results were not uniform for other cytokines (i.e., Thl cytokine
IL-2), most available results pointed to increases in Th2 cytokines, specifically, IL-6 and
IL-10, in Pb-exposed animals. Results from Gao et al. (2007) pointed to a role for
dendritic cells in skewing T cells to Th2 cytokine production. Several studies
demonstrated Pb-induced increases in TNF-a but in animals with high prenatal dietary or
postnatal air Pb exposure or blood Pb levels (e.g., 149 (ig/dL) (Chen et al.. 1999; Miller
et al.. 1998; Zelikoff et al.. 1993). In vitro evidence indicates that Pb may induce
increases in TNF-a via MAPK signaling pathways (Khan et al.. 2011; Gao et al.. 2007;
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Cheng et al.. 2006). Animal evidence indicates effects on cytokine production by prenatal
and postnatal Pb exposures, whereas epidemiologic studies examined only concurrent
blood Pb level. Because concurrent blood Pb level in children and adults reflects both
previous and recent Pb exposures, associations with concurrent blood Pb level may
reflect an effect of cumulative Pb exposure. Overall, the consistent toxicological evidence
for Pb-induced decreases in Thl cytokines and increases in Th2 cytokines and pro-
inflammatory cytokines such as TNF-ot provides clear mode of action support for the
evidence indicating the effects of Pb on increases in IgE, inflammation, atopic and
inflammatory conditions, and decreases in host resistance.
4.6.6.2 Decreased Nitric Oxide
As described in the 2006 Pb AQCD (U.S. EPA. 2006b). key mode of action support for
the effects of Pb on impairing macrophage function and decreasing host resistance was
provided by consistent toxicological findings for Pb-induced decreases in NO, which is
involved in the cytotoxic activity of macrophages in host defense processes [see 2006 Pb
AQCD Annex Table AX5.9.6 (U.S. EPA. 20061)1. In adult rodents, decreases in NO from
macrophages were observed with short-term Pb acetate exposures (1 or 6 days) during
early gestation (Bunnetal.. 200Ib: Lee etal. 200Ib) but not long-term exposures
occurring during the full gestational period (Bunn etal.. 200 Ic: Chen etal.. 1999; Miller
et al.. 1998). With short-term exposure, decreases in NO were found in Sprague-Dawley
rats with blood Pb level 4.5 (ig/dL in males and 5.3 (ig/dL in females measured 2 weeks
after Pb exposure in drinking water of dams was terminated (Bunnetal.. 200 Ib) and in
chicks with blood Pb levels that did not exceed 11 (ig/dL but with Pb injected into eggs
embryos (Lee etal.. 200Ib).
The short-term in vivo findings are supported by several in vitro observations of
decreased release of NO from macrophages and splenocytes induced by a wide range of
Pb exposure concentrations (0.625-5 (iM) and durations (2 hours-6 days) (Farrer et al..
2008: Mishra et al.. 2006a: Krocova et al.. 2000: Chen etal.. 1997: Tian and Lawrence.
1996. 1995). Farrer et al. (2008). with 5 (iM Pb glutamate, further indicated that the mode
of action for Pb may involve a decrease in inducible NO synthase function in myeloid
cells without a change in its gene expression. Also in this study, Pb abrogated the
myeloid cell (CD1 lb+)-mediated suppression of CD4+ T cell proliferation, and
exogenous NO restored suppression. Together, these findings indicated that Pb may
indirectly enhance T cell proliferation through its effect on decreasing NO production.
Combined with the observation that Pb can alter antigen processing (Farrer et al.. 2005)
and, hence, the quality and magnitude of the acquired immune response against pathogen
exposure, evidence indicated that multiple arms of host defense against infectious
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challenge can be compromised. Diminished production of NO in innate immune cells
such as macrophages could affect other physiological systems (e.g., neurological,
cardiovascular, endocrine) that require NO signaling cascades.
Consistent with the toxicological evidence, cross-sectional studies found associations
between concurrent blood Pb level with lower NO in populations living near Pb sources
(Barbosa et al., 2006c; Pineda-Zavaleta et al., 2004). These studies did not report
sufficient information on participation rates; however, examination of populations near
Pb sources could limit generalizability of findings. Additional limitations include the
cross-sectional design, which impedes determination of the temporal sequence between
Pb exposure and NO suppression and the limited consideration for potential confounding.
Because of these limitations, epidemiologic evidence is not a major consideration in
drawing conclusions about the effects of Pb on NO. However, it is useful in that it shows
Pb-associated decreases in NO in humans, similar to toxicological studies.
In a study of 65 children (ages 6-11 years) in Region Lagunera, Mexico, mean concurrent
blood Pb levels increased (7.02 to 20.6 to 30.38 (ig/dL) with increasing school proximity
(8,100-650 meters) to a Pb smelter (Pineda-Zavaleta et al., 2004). With adjustment for
age and sex, a 1 (ig/dL higher blood Pb level was associated with a 0.00089 (95% CI:
-0.0017, -0.00005) log lower NO release from macrophages activated by
phytohemagglutinin (PHA). Because PHA activates macrophages indirectly through the
activation of lymphocytes, the results indicated that Pb suppressed T cell-mediated
macrophage activation. Blood Pb group comparisons indicated that associations were due
largely to the lower NO in the 23 children closest to the smelter who had blood Pb levels
10.31-47.49 (ig/dL. Though not described in detail, higher blood Pb level was not
associated with lower NO in girls.
Among 104 adults (ages 18-60 years) in Sao Paolo, Brazil residing near a closed battery
plant, Barbosa et al. (2006c) observed an association between higher concurrent blood Pb
level and lower plasma NO in the 69 adults (mean blood Pb level: 6.4 (ig/dL) with the TC
or CC eNOS genotype (r = 0.23, p = 0.048). The results are consistent with the reduced
promoter activity and potentially reduced gene expression of the TC/CC variants. Results
were not adjusted for potential confounding factors, but subjects were nonsmoking,
nonalcohol drinking with normal mean BMI and SEP. The inclusion criteria may further
limit the generalizability of findings. Because NO was measured in plasma, immune cells
could not be identified as the source of NO. In contrast, Valentino et al. (2007) found
similar plasma NO levels in 44 male foundry workers (mean blood Pb level: 21.7 (ig/dL),
14 male pottery workers (mean blood Pb level: 9.7 (ig/dL), and 59 male unexposed
workers of similar age (mean blood Pb level: 3.9 (ig/dL, ages 25-61 years). Quantitative
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results were not reported, but blood Pb level was reported not to be correlated with NO.
Potential confounding factors, including other workplace exposures were not examined.
In summary, short-term dietary Pb exposure early in gestation but not long-term exposure
for the full gestational period was found to reduce NO production by macrophages in
adult animals. In vitro evidence consistently demonstrates Pb-induced decreases in NO
from macrophages and splenocytes. The relevance of toxicological evidence is supported
by observations of an association between higher concurrent blood Pb level and lower
release of NO from macrophages of children in Mexico. The association in children was
due largely to lower NO in macrophages from children living near a Pb smelter with
concurrent blood Pb levels >10 (ig/dL, higher than those in most of the current U.S.
population of children. This study was cross-sectional and had limited consideration for
potential confounding factors. However, the toxicological evidence provides clear mode
of action support for the effects of Pb on decreasing host resistance given the role of NO
in mediating the cytotoxic activity of macrophages.
4.6.6.3 Increased Reactive Oxygen Species and Prostaglandins
ROS are released from macrophages during phagocytosis and are involved in killing
invading bacteria. ROS and PGE2 are important mediators of inflammation which can
result in local tissue damage (Figure 4-33). The roles of ROS and PGE2 in both host
defense and injury may explain some of the inconsistencies in the evidence as reported in
the 2006 Pb AQCD. In activated macrophages undergoing phagocytosis, high
concentration (10-1,000 (iM, 15 minutes-20 hours) Pb chloride or acetate exposures were
found to reduce release of ROS (Hilbertz et al.. 1986; Castranova et al. 1980). consistent
with observations of Pb-induced decreased resistance to bacterial or viral infection. Also
in stimulated macrophages, Hilbertz et al. (1986) found an increase in ROS with 1-hour
Pb chloride exposure but a decrease in ROS (and phagocytosis) with 20-hour exposure
indicating differential responses by exposure duration. Similarly, Shabani and Rabbani
(2000) found 3-hour Pb nitrate exposure (240 (iM) to increase ROS from alveolar
macrophages and induce their apoptosis, consistent with impaired host defense. Also
similar to Hilbertz et al. (1986). Chen et al. (1997) found 18-hour Pb glutamate exposure
(4 (iM) to induce a decrease in ROS but did not indicate the functional state of
macrophages. Other studies reported reductions in antioxidants such as glutathione and
catalase in conjunction with reduced macrophage function in Swiss mice treated with
Pb nitrate by oral gavage (40 mg/kg per day, 30 days) (Lodi et al.. 2011) or increases in
PGE2 and apoptosis in cultures of neuroblastoma cells with 0.01-10 (iM Pb nitrate (3
hours) (Chetty et al.. 2005). While several processes have been proposed to explain the
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mechanisms of Pb-induced oxidative damage, the exact combination of processes
involved remains to be determined (Section 4.2.4).
In adult animals, Pb exposure increased ROS release from macrophages immediately
upon cessation of exposure (Bavkov et al.. 1996; Zelikoff etal.. 1993) but not 9-10 weeks
after exposure (Miller et al.. 1998). consistent with in vitro findings. These Pb exposures
occurred through relevant routes of exposure, i.e., diet or air, but with high
concentrations, e.g., 31 (ig/m3 Pb oxide in air for 3 hours/day for 4 days in rabbits
(Zelikoff et al.. 1993) and 1.5 mg/kg Pb acetate in diet for 30 days in Swiss mice (Bavkov
et al.. 1996). Neither study measured the blood Pb levels of animals.
Pb-associated increases in ROS also were found in macrophages of children. However,
the findings are based on a cross-sectional analysis with limited consideration for
potential confounding and higher blood Pb levels (means: 7.02, 20.6, and 30.38 (ig/dL)
than most of those in the current U.S. population. In addition to finding suppressed NO
production (Section 4.6.6.2). Pineda-Zavaleta et al. (2004) found a Pb-associated increase
in ROS production from macrophages in children in Mexico living in varying proximities
to a Pb smelter. With adjustment for age and sex, a 1 (ig/dL higher concurrent blood Pb
level was associated with a 0.00389 (95% CI: 0.00031, 0.00748) log higher release of
superoxide anion from macrophages directly activated by IFN-y/LPS. The blood
Pb-associated superoxide anion release was larger from macrophages of males. Because
IFN-y directly activates macrophages, these results indicated that Pb stimulated cytokine-
induced macrophage activation. Blood Pb level was not associated with ROS from
neutrophils in a study of male Pb recycling workers (ages 19-45 years) in India. Despite
large differences in blood Pb levels between 30 Pb workers (mean: 106 (ig/dL) and 27
unexposed controls (mean: 4.5 (ig/dL), levels of ROS released from neutrophils
(indicators of respiratory burst) were similar between groups (Mishra et al.. 2006a).
However, evidence does not clearly indicate that neutrophils are a major responding cell
to Pb exposure (Section 4.6.2.5).
PGE2 is produced from the metabolism of cell membrane phospholipids and may be
released by macrophages to modulate their function in a paracrine or autocrine manner.
An increase in PGE2 from macrophages was found in turkeys exposed to Pb acetate in
feed during the early postnatal period (PND1-PND21) which produced a high blood Pb
level, i.e., mean of 42 (ig/dL (Knowles and Donaldson. 1997). However, dietary
Pb acetate exposure (250-2,000 ppm) of chicks during the same developmental period
(PND1-PND19) resulted in an increase in serum arachidonic acid but not PGE2 or other
prostaglandins (Knowles and Donaldson. 1990). In vitro studies also used high Pb
exposure concentrations, >20 (iM Pb chloride (Flohe et al.. 2002; Lee and Battles. 1994).
A recent in vitro study with human neuroblastoma cells found increases in PGE2 with
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lower Pb concentrations (0.01-1 (iM Pb acetate) than previously reported (Chetty et al..
2005).
In summary, ROS and PGE2 function in modulating macrophage function, aiding in
bacterial killing, and inducing tissue damage as part of an inflammatory response.
Consistent with these diverse roles, toxicological studies have found both Pb-associated
increases and decreases in ROS and PGE2. Consistent with findings in animal and in
vitro models, a cross-sectional study found an association between higher concurrent
blood Pb level (>10 (ig/dL) and higher ROS release from macrophages in children in
Mexico with adjustment for potential cofounding only by age and sex. In animals,
Pb-induced increases in macrophage production of ROS and PGE2 occurred
concomitantly with functional alterations such as impaired macrophage phagocytosis and
apoptosis. Toxicological results were based on examination of high Pb concentrations but
are considered to be relevant in the ISA for informing mode of action. The toxicological
observations of Pb-induced effects on ROS and PGE2 provide evidence for modes of
action underlying the effects of Pb on reduced macrophage function and decreased host
resistance.
4.6.6.4 Cellular Death (Apoptosis, Necrosis)
The 2006 Pb AQCD reported contrasting effects of Pb on the apoptosis of macrophages
in vitro (U.S. EPA. 2006b). However, with a few recent studies in mice, evidence
suggests that Pb exposure may induce cell death of immune cells via apoptosis or
necrosis. Xu et al. (2008) found that 4-week dietary exposure of juvenile ICR mice to
Pb acetate (50-100 mg/kg) induced DNA damage in peripheral blood lymphocytes,
increased p53 and Bax expression in the liver, but did not change Bcl-2 expression
(creating a Bax/Bcl-2 imbalance). Bax promotes apoptosis, whereas Bcl-2 inhibits
apoptosis. Concomitant increases in indicators of oxidative stress in liver homogenate
suggested that oxidative stress may mediate the effects of Pb in promoting apoptosis. In
Swiss mice, Bishayi and Sengupta (2006) found splenic macrophages to have elevated
DNA fragmentation, a key event in apoptosis, but with i.p. Pb acetate treatment
(10 mg/kg). Consistent with in vivo findings, Gargioni et al. (2006) found 20 and 40 (i
Pb nitrate to induce cell death in mouse peritoneal macrophages in vitro which occurred
concomitantly with a loss of cell membrane integrity, indicating that Pb primarily
induced macrophage necrosis or cell lysis. While evidence for Pb-induced cell death of
immune cells with routes of Pb exposure relevant to this ISA is sparse, the evidence
suggests that the induction of cell death may be a potential mode of action by which Pb
affects macrophage function and decreases host resistance.
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4.6.7 Immune Effects of Pb within Mixtures
In the 2006 Pb AQCD (U.S. EPA. 2006b). the immune effects resulting from Pb within
metal mixtures were not well characterized; however with recent studies, there is some
evidence for greater immune effects resulting from co-exposure to Pb and other metals.
In Swiss mice treated with Pb acetate (10 mg/kg i.p., daily for 15 days), As (0.5 mg/kg
i.p., daily for 15 days), or both, Bishayi and Sengupta (2006) reported a greater than
additive effect of co-administered Pb and As on decreasing bacterial resistance,
decreasing macrophage myeloperoxidase release, and NO production. Epidemiologic
studies have not widely examined interactions between Pb and other metals. However,
consistent with Bishayi and Sengupta (2006). Pineda-Zavaleta et al. (2004)
(Section 4.6.6.2) found interactions between Pb and As among 65 children in Mexico
ages 6-11 years. Contamination of drinking water by both Pb and As was a concern in the
study area; however, urinary As levels were higher in children who had lower blood Pb
levels. Higher urinary As was associated with lower NO release from macrophages
(similar to blood Pb). Higher internal dose of both As and Pb was associated with a larger
decrease in NO (p for interaction = 0.037) than was dose of either metal alone. A positive
interaction (p = 0.042) between blood Pb and urinary As was found for superoxide anion
release from indirectly activated macrophages. But each metal alone was significantly
associated with superoxide anion release only from directly activated macrophages.
Because of the high blood Pb levels in these children (means in three groups at varying
distances from a Pb smelter: 7, 20.6, 30.4 (ig/dL), it is not clear whether these
relationships would apply to the current U.S. population of children.
Institoris et al. (2006) found that Cd or Hg co-exposure potentiated the effects of Pb.
Lymph node weight decreased in 4 week-old Wistar rats exposed to 20 mg/kg Pb acetate
by drinking water plus another metal but not with Pb alone. In contrast, Fortier et al.
(2008) did not find metal co-exposure to increase the effects of Pb. Pb chloride
(7.5-20.7 (ig/dL) did not alter lymphocyte proliferation, monocyte phagocytosis, or NK
cell activity in human leukocyte cultures. A mixture of 20.7 (ig/dL Pb chloride plus
12.0 (ig/dL methylmercuric chloride (MeHgCl) decreased lymphocyte proliferation;
however, these effects were attributed to MeHgCl, which singly had a stronger
suppressive effect. Other toxicological studies found metal mixtures that included Pb to
decrease antibody titers or increase neutrophil counts (Jadhav et al.. 2007; Massadeh et
al.. 2007) but did not test each metal individually. The latter findings cannot necessarily
be attributed to interactions between Pb and other components within the mixture.
Overall, several results indicated that exposures to Pb-containing metal mixtures are
associated with immune effects. Some but not all results showed greater immune effects
with co-exposures to Pb and metals such as As, Cd, or Hg than with Pb exposure alone.
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4.6.8 Summary and Causal Determination
The cumulative body of epidemiologic and toxicological evidence describes several
effects of Pb exposure on the immune system related to a shift from Thl- to Th2-type
responses, including an increase in atopic and inflammatory conditions and a decrease in
host resistance. Outcomes related to an increase in atopic and inflammatory conditions
include asthma, allergy, increased IgE, and mode of action endpoints such as selective
differentiation of Th2 cells, increased production of Th2 cytokines, B cell activation, and
increased inflammation. Outcomes related to decreased host resistance include enhanced
susceptibility to bacterial and viral infection, suppressed DTH, and those describing
mode of action, i.e., decreased production of Thl cytokines, reduced phagocyte function,
and increased inflammation. A small body of studies indicates the effects of Pb exposure
on autoimmunity. The sections that follow describe the evaluation of evidence for these
three groups of outcomes, increased atopic and inflammatory conditions, decreased host
resistance, and autoimmunity, with respect to causal relationships with Pb exposure using
the framework described in Table II of the Preamble. The application of key evidence,
supporting or contradicting, to the causal framework is summarized in Table 4-34.
4.6.8.1 Evidence for an Increase in Atopic and Inflammatory
Conditions
Epidemiologic and toxicological evidence together indicate that a causal relationship is
likely to exist between Pb exposure and atopic and inflammatory conditions. This
relationship is supported by evidence for associations of blood Pb levels with asthma and
allergy in studies in children (Jedrychowski et al.. 2011; Pugh Smith and Nriagu. 2011;
Joseph et al., 2005). Pb-associated increases in IgE in children and animals, and evidence
describing modes of action including increases in Th2 cytokines and inflammation.
Recent studies on asthma and allergy expand upon the evidence presented in the
2006 Pb AQCD by providing additional evidence from prospective analyses, and by
better addressing uncertainties regarding potential confounding by factors such as SES
and smoking and residential allergen exposures (Table 4-34). Findings from studies that
prospectively ascertained outcomes provide support for the directionality of associations
(Jedrychowski et al.. 2011; Joseph et al.. 2005). In these studies, the lack of selective
participation and objective assessment of outcomes of asthma and allergy through
medical records or clinical testing, respectively, indicates lack of biased reporting of
asthma and allergy in children with higher blood Pb levels (Section 4.6.5.2 and Table
4-34). Among children age 5 years in Poland, Jedrychowski et al. (2011) found that a
1 (ig/dL increase in prenatal cord blood Pb level was associated with an increased risk of
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allergic sensitization of 2.3 (95% CI: 1.1, 4.6). The magnitude of risk did not differ with
and without adjustment for maternal education or residential allergen levels. An
additional strength of this study was the adjustment for prenatal and postnatal smoking
exposure. A large study of 4,634 children in Michigan ages 1-3 years found that
compared with Caucasian children with blood Pb levels <5 (ig/dL measured up to
12 months before asthma assessment, Caucasian children with blood Pb levels > 5 (ig/dL
had an increased risk of incident asthma of 2.7 (95% CI: 0.9, 8.1) (Joseph et al.. 2005).
Adjustment was made for census block average income, which may not adequately
control for potential confounding by individual subject-level SES.
Supporting evidence was provided by a cross-sectional study of 356 children (71% ages
<6 years) in Michigan, which found that compared with children with concurrent blood
Pb levels <10 (ig/dL, children with concurrent blood Pb level > 10 (ig/dL had increased
parental report of an asthma diagnosis in the previous 12 months with an OR of 7.5 (95%
CI: 1.3, 42.9) (Pugh Smith and Nriagu. 2011). This study was cross-sectional and
produced an imprecise effect estimate; however, a strength of the study was the relatively
extensive consideration for potential confounding, including adjustment for family-level
income. As with Jedrychowski et al. (2011). Pugh Smith and Nriagu (2011) found an
association with adjustment for smoking exposures in the home plus other indicators of
housing exposures and condition (Table 4-34). Some covariates may be predictors of Pb
exposure and control for Pb exposure itself. The studies of asthma and allergy differed in
which and how potential confounding factors were considered, particularly SES. While
there is no single complete measure of SES, the heterogeneity in SES indicators used
across these few studies produces uncertainty regarding residual confounding. Residual
confounding also is possible by factors not examined. The examination of maternal
education and exposure to smoking or allergens in Jedrychowski et al. (2011) and family
income, smoking, housing conditions, pets, or pests in Pugh Smith and Nriagu (2011)
increase confidence in the associations observed for blood Pb levels. However, because
evidence is limited to a few populations, there is uncertainty regarding potential
confounding of Pb associations by SES and other exposures well characterized in the
literature to be associated with asthma and allergy.
With respect to blood Pb levels associated with atopic and inflammatory conditions,
Joseph et al. (2005) found elevated incidence of asthma in Caucasian children with earlier
childhood blood Pb levels > 5 (ig/dL. Pugh Smith and Nriagu (2011) found higher asthma
prevalence in children with concurrent blood Pb levels > 10 (ig/dL. Jedrychowski et al.
(2011) found increased allergic sensitization in association with cord blood Pb levels that
were low (geometric mean: 1.16 (ig/dL). However, because cord blood Pb levels are
influenced by maternal blood Pb levels, the associations may reflect higher past maternal
Pb exposures.
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Biological plausibility for the relationships found between blood Pb levels and asthma
and allergy in children is provided by evidence characterizing modes of action, namely, a
Pb-associated shift in production from Thl cytokines (e.g., IFN-y) to Th2 cytokines
(e.g., IL-4) and increase in Th2-dependent IgE levels (Table 4-34). A majority of this
evidence was available in the 2006 Pb AQCD (U.S. EPA. 2006b). The shift from Thl to
Th2 cytokine production in animals was found with prenatal or postnatal (4 weeks in
juveniles, 3 weeks or 8 weeks in adults) dietary Pb exposures. In the studies available in
humans (Table 4-34). higher concurrent blood Pb levels were associated with higher
serum IL-4 in children (Lutz et al.. 1999) and higher serum IL-6 in nonoccupationally-
exposed adults (adjusted for age, BMI, and current smoking status and additionally
adjusted for income, physical activity, education, and history of inflammatory conditions
in the large NHANES analysis) (Songdei etal.. 2010: Kim et al.. 2007). Because of the
limitations in the small body of epidemiologic studies, i.e., the cross-sectional design of
studies and inconsistent consideration for potential confounding, the epidemiologic
evidence is a lesser consideration in drawing conclusions about Pb-associated cytokine
changes. However, epidemiologic evidence does not detract from the clear toxicological
evidence for Pb-induced increases in Th2 cytokine production. Coherence for a shift from
Thl to Th2 cytokine production is found in the in vitro evidence for Pb-induced selective
differentiation of naive T cells to a Th2 subtype (Heo etal.. 1998: 1996: McCabe and
Lawrence. 1991). A recent study in adult mice and in vitro provided new evidence that
Pb may promote the shift to Th2 responses by increasing production of Th2 cytokines in
dendritic cells, the major effector in antigen responses (Gao et al., 2007).
Additional mode of action support is provided by associations observed between higher
concurrent blood Pb levels and higher serum IgE in several different populations of
children (Section 4.6.3. Table 4-34). While most studies found elevated IgE in groups of
children with concurrent blood Pb levels >10 (ig/dL, Karmaus et al. (2005) found higher
serum IgE in children ages 7-10 years in Germany with blood Pb levels 2.8-3.4 (ig/dL
compared with children with lower blood Pb levels. A monotonic increase in IgE was
found across increasing blood Pb groups, except in the highest group in Lutz et al. (1999)
but not in other studies (Hegazy etal.. 2011: Karmaus et al.. 2005). Lutz et al. (1999)
recruited children in Michigan from a public assistance or Pb poisoning prevention
program, and Karmaus et al. (2005) recruited schoolchildren but excluded those from
homes where more than 12 cigarettes were smoked per day. The nature of recruitment
may limit generalizability of findings. Sufficient information was not reported to assess
biased participation by Pb exposure and history of allergy or asthma. The limited
consideration for potential confounding comprised adjustment for age (Karmaus et al..
2005: Lutz etal.. 1999). smoking exposure, serum lipids, blood organochlorine levels,
and previous infections (Karmaus et al.. 2005) but not SES or allergen exposure.
Although these findings were based on cross-sectional analyses and had limited
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consideration for potential confounding, they were supported by similar findings in
animals, which are not subject to reverse causation or confounding bias. Despite clear
evidence in animals overall, there was some inconsistency for Pb-induced increases in
IgE in animals with gestational or gestational/lactational dietary Pb exposures that
resulted in blood Pb levels 5-20 (ig/dL, which are more relevant to this ISA (Chen et al..
2004: Snvderetal.. 2000). Miller et al. (1998) found elevated IgE in adult F344 rats after
gestational Pb exposure via drinking water of dams whose blood Pb levels peaked at
30-39 (ig/dL. There is lack of coherence between the consistent results for IgE and the
inconsistent findings for Pb-induced activation of B cells, which differentiate into allergic
antibody-producing cells (Section 4.6.3). However, there is strong mode of action support
in animals for Pb-induced increases in IL-4, which stimulates differentiation of B cells.
Further support for the effects of Pb exposure on increasing risk of atopic and
inflammatory conditions is provided by evidence of Pb-associated inflammation
(Section 4.6.4 and Table 4-34). Coherence for this evidence is found with findings for
Pb-induced increases in IgE which primes basophils and mast cells to release pro-
inflammatory mediators. Pb-induced inflammation is clearly demonstrated by a large
toxicological evidence base for the effects of Pb exposure on inducing macrophages into
a hyperinflammatory state as characterized by enhanced production of TNF-a, PGE2, and
ROS. Inflammation was observed in rabbits exposed to Pb via air for 4 days (31 (ig/m3)
(Zelikoff et al.. 1993) and rodents exposed via diet (250 ppm drinking water during
gestation, 1.5 mg/kg food postnatally for 30 days) (Miller et al.. 1998; Bavkov et al..
1996). Epidemiologic evidence is sparse and therefore is a lesser consideration in
drawing conclusions regarding the effects of Pb exposure on inflammation. But,
consistent with previous toxicological evidence, a large analysis of adults participating in
NHANES found an association in 4,278 men between concurrent blood Pb levels and
serum CRP, an indicator of systemic inflammation, with adjustment for age, BMI,
income, physical activity, education, history of inflammatory conditions, cardiovascular
disease, diabetes, and smoking status (Songdei et al.. 2010). Because only concurrent
blood Pb levels were examined, there is uncertainty regarding the temporal sequence
between Pb exposure and inflammation and the magnitude, timing, frequency, and
duration of Pb exposures that contributed to the observed association.
With respect to important lifestages of Pb exposure, gestational or gestational/lactational
Pb exposures, producing blood Pb levels 8 (ig/dL 2 weeks postweaning and 20 (ig/dL at
age 1-2 weeks, were found to affect endpoints such as IgE and/or cytokine levels in
juvenile and adult rodents (Chen et al.. 2004; Snyder et al.. 2000). However, increases in
Th2 cytokines also were found in adult animals with lifetime Pb exposures beginning in
gestation and producing blood Pb levels 1-12 (ig/dL upon cessation of exposure (lavicoli
et al., 2006b). The Pb exposure lifestage, magnitude, frequency, and duration associated
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with atopic and inflammatory conditions are not well characterized in humans. Cord
blood Pb level was associated with allergic sensitization in children (Jedrychowski et al..
2011). whereas other studies of asthma, allergy, IgE, cytokines, and inflammation in
children and adults examined blood Pb levels concurrently with outcomes or at some
unspecified time before outcome assessment (Table 4-34). In children and adults,
concurrent blood Pb levels are influenced by cumulative Pb exposure history (from
remodeling of bone) and recent Pb exposures. Thus, the evidence overall indicates that
prenatal only or cumulative Pb exposure may influence atopic and inflammatory
conditions. However, neither toxicological nor epidemiologic evidence clearly identifies
an individual critical lifestage or duration of Pb exposure that is more strongly associated
with atopic and inflammatory conditions.
In conclusion, prospective studies in a few populations of children indicate associations
of asthma and allergy with blood Pb levels measured prenatally or sometime before the
outcome, with a cross-sectional study providing supporting evidence with associations
for concurrent blood Pb. Prospective design, lack of selective participation of subjects,
and objective assessment of outcomes reduce the likelihood of undue selection bias.
These few studies varied in their consideration for potential confounding by SES and
exposure to smoking or allergens. Thus, uncertainty remains regarding residual
confounding in associations observed between blood Pb levels and asthma and allergy in
children. The evidence for asthma and allergy is supported by cross-sectional associations
found between higher concurrent blood Pb levels in children and higher IgE, an important
mediator of asthma and allergy. The biological plausibility for the effects of Pb on IgE is
provided by consistent findings in animals with gestational or gestational-lactational Pb
exposures, with some evidence at blood Pb levels relevant to this ISA. In children, higher
IgE and higher asthma prevalence were examined and found mostly in groups with
concurrent blood Pb levels >10 (ig/dL. Coherence for the evidence of Pb-associated
increases in asthma, allergy, and IgE is found with evidence for most of the examined
endpoints related to mode of action, i.e., Pb-induced increases in Th2 cytokine production
and inflammation in animals. The biological plausibility from toxicological evidence,
evidence from prospective studies in children, and lack of evidence for inflammation-
related permeability in major routes of Pb exposure such as the gastrointestinal system
together suggest that reverse causation is not the primary explanation for associations
observed between Pb and atopic and inflammatory conditions. Neither toxicological nor
epidemiologic evidence clearly identifies an individual critical lifestage or duration of Pb
exposure associated with atopic and inflammatory conditions but points to an influence of
prenatal only or cumulative Pb exposures. The epidemiologic evidence in a few
populations and toxicological evidence together supporting a relationship between Pb
exposure and asthma, allergy, and shift to a Th2 phenotype as an underlying mode action
but some uncertainty regarding potential confounding is sufficient to conclude that a
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causal relationship is likely to exist between Pb exposures and an increase in atopic and
inflammatory conditions.
4.6.8.2 Evidence for Decreases in Host Resistance
Evidence indicates that a causal relationship is likely to exist between Pb exposure and
decreased host resistance based primarily on consistent observations that relevant Pb
exposures decrease responses to antigens (i.e., suppresses DTH) and increase bacterial
titers and subsequent mortality in rodents (Table 4-34. Sections 4.6.2.3 and 4.6.5.1). A
majority of this evidence was available in the 2006 Pb AQCD (U.S. EPA. 2006b). The
studies that reported blood Pb levels demonstrated increased bacterial titers and mortality
with adult-only 16 week Pb exposure via drinking water in adult mice with Salmonella
typhimurium infection and blood Pb level 20 (ig/dL (Fernandez-Cabezudo et al.. 2007)
and with lactational (PND1-PND22) Pb exposure in juvenile mice with Listeria
monocytogenes infection and blood Pb level 25 (ig/dL (Dyatlov and Lawrence. 2002).
DTH was suppressed in adult rats with blood Pb levels 6 and 25 (ig/dL after gestational
Pb exposure in drinking water (Chen et al., 2004; Bunn et al.. 200la). While a few
epidemiologic studies found higher prevalence of respiratory infections in children with
higher concurrent blood Pb levels (Karmaus et al.. 2005; Rabinowitz et al.. 1990) and
Pb-exposed workers (Ewers etal.. 1982). the implications are limited by the lack of
rigorous statistical analysis (i.e., regression) and consideration for potential confounding.
These limitations also apply to the recent ecologic study finding increased prevalence of
child respiratory infections in areas with higher lichen or soil Pb (Carreras et al.. 2009)
(Table 4-34). These limitations produce uncertainty about the effects of Pb exposure on
decreased host resistance in humans but do not detract from the consistent evidence in
animals.
The effects of Pb on decreased host resistance are well supported by evidence describing
underlying modes of action (Table 4-34). Evidence in animals indicates Pb-induced
functional impairment of macrophages, which phagocytize pathogens. Decreased
macrophage colony formation was found in rats after gestational Pb exposure (Bunn et
al.. 2001b). and decreased phagocytic activity was found in mice and turkeys after
lactational or 2-week juvenile Pb exposure (Knowles and Donaldson. 1997; Kowolenko
et al.. 1991). Additional coherence for Pb-induced decreased host resistance is found with
observations in animals that gestational Pb exposure suppressed macrophage production
of NO which is involved in bacteria killing (Section 4.6.6.2). and postnatal Pb exposure
(air for 4 days, food for 30 days) increased production of ROS and PGE2, which mediate
tissue damage (Section 4.6.6.3). Similarly, a cross-sectional epidemiologic study found a
smaller release of NO and larger release of superoxide anion from macrophages of
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children ages 6-11 years with higher concurrent blood Pb levels (10.3-47.5 versus
<10.3 (ig/dL) (Pineda-Zavaleta et al.. 2004) after adjustment for age and sex. Because of
the limited consideration for potential confounding in this study and examination of
higher blood Pb levels than those in most current U.S. children, the results are a lesser
consideration in drawing conclusions about the effects of Pb on macrophages. However,
they do suggest the relevancy of toxicological observations to humans. The killing
capability of macrophages is enhanced by the Thl cytokine IFN-y. Therefore, an effect of
Pb exposure on decreased host resistance is additionally supported by clear evidence in
animals for the effects of Pb exposure on suppressing IFN-y production (Section 4.6.6.1).
A recent study in mice indicated that Pb-induced suppression of DTH may be mediated
by a shift in production from Thl to Th2 cytokines specifically in dendritic cells (Gao et
al.. 2007).
Some evidence did not contribute strong support for the mode of action for Pb-induced
decreased host resistance. Pb-exposed workers were found to have reduced functionality
of neutrophils, which respond to bacterial infection (Table 4-34. Section 4.6.2.5) but
without consideration for potential confounding or analogous toxicological evidence.
Neither epidemiologic nor toxicological evidence clearly demonstrated an effect of Pb
exposure on NK cells, which respond to viral infection (Section 4.6.2.7).
With respect to important lifestages of Pb exposure, animal studies found that gestational
Pb exposures, producing blood Pb levels of 6 and 25 (ig/dL, resulted in decreases in Thl
cytokines, suppression of DTH, and greater susceptibility to bacterial infection (Chen et
al., 2004; Bunnetal.. 200la). However, these effects related to decreased host resistance
also were affected in mice by lactational (Dyatlov and Lawrence. 2002). adult-only
(>4 weeks) (Fernandez-Cabezudo et al.. 2007). and lifetime gestation to adult Pb
exposures (lavicoli et al.. 2006b) that produced blood Pb levels 1-25 (ig/dL. Thus, the
animal evidence does not clearly identify an individual lifestage or time period of Pb
exposure that is more strongly associated with decreased host resistance.
In conclusion, animal toxicological observations are the primary contributors to the
evidence for Pb-induced decreased host resistance. Several studies in rodents show that
dietary Pb exposures producing relevant blood Pb levels result in increased susceptibility
to bacterial infection and suppressed DTH. Further, coherence is found with evidence
describing modes of action, including suppressed production of Thl cytokines and
decreased macrophage function in animals. These effects were found with gestational,
lactational, adult-only, and lifetime Pb exposures of animals, without an individual
critical lifestage of exposure identified. A few cross-sectional epidemiologic studies
indicate Pb-associated increases in respiratory infections but limitations, including the
lack of rigorous methodology and consideration for potential confounding produce
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uncertainty in the epidemiologic evidence for decreased host resistance in humans. The
consistent toxicological evidence in animals for increased susceptibility to bacterial
infection, suppressed DTH, and related modes of action but uncertainty in the
epidemiologic evidence in humans is sufficient to conclude that a causal relationship is
likely to exist between Pb exposure and decreased host resistance.
4.6.8.3 Evidence for Autoimmunity
Animal toxicological evidence describes the potential of Pb to increase autoimmunity,
with a few previous studies showing Pb-induced generation of auto-antibodies (Hudson
etal.. 2003; Bunn et al. 2000; El-Fawal et al.. 1999; Waterman et al.. 1994) and recent
studies providing indirect evidence by showing formation of neoantigens that could result
in the formation of auto-antibodies (Table 4-34). Several observations were made in
animals injected with Pb, which is a route of exposure with uncertain relevance to
humans. Higher levels of auto-antibodies also were found in Pb-exposed battery workers;
however, implications are limited because of the high blood Pb levels (range:
10-40 (ig/dL) of some of the workers and lack of consideration for potential confounding
by several factors, including other occupational exposures (El-Fawal et al., 1999).
Because the results combined from available toxicological and epidemiologic studies do
not sufficiently inform Pb-induced generation of auto-antibodies with relevant Pb
exposures, the evidence is inadequate to determine if there is a causal relationship
between Pb exposure and autoimmunity.
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Table 4-34 Summary of evidence supporting immune causal determinations.
Attribute in Causal
Framework3
Key Evidence
References
Pb Biomarker Levels
Associated with Effects0
Increase in Atopic and Inflammatory Conditions - Likely Causal
Multiple
epidemiologic
studies show
associations with
relevant blood Pb
levels
A few high-quality prospective studies indicate higher asthma
and allergy incidence in association with early childhood or
prenatal blood Pb levels in children (ages 1-5 yr) (U.S.,
Poland)
Some studies report high participation and/or follow-up
retention, not conditional on Pb exposure or outcome.
Some studies objectively assessed outcomes with clinical
testing, medical records.
Other supporting cross-sectional evidence in children (ages 6
mo-10 yr) for increases in IgE but with limited consideration
for potential confounding factors, particularly SES.
Associations observed in studies in U.S., Europe, Asia;
insufficient information to assess potential selection bias.
Mixed evidence for monotonic C-R varies for IgE. Some
studies show increasing IgE across blood Pb groups, except
in highest group.
Joseph et al. (2005), Jedrychowski et al.
(2011),
Section 4.6.5.2
Joseph et al. (2005). Jedrychowski et al.
(2011)
Jedrychowski et al. (2011)
Pugh Smith and Nriagu (2011)
Karmaus et al. (2005).
Hegazy et al. (2011).
Lutz et al. (1999).
Hon et al. (2010: 2009).
Sun et al. (2003)
Section 4.6.3
C-R found: Lutz et al. (1999)
No C-R: Karmaus et al. (2005), Hegazy et
al. (2011)
Groups (ages 1-3 yr) with
blood Pb levels measured
earlier in childhood
>10ug/dL
Prenatal (cord): geometric
mean 1.16 ug/dL
Group (71% age <6 yr)
with concurrent blood Pb
levels >10 ug/dL
Groups (ages 6 mo-10 yr)
with concurrent blood Pb
levels >10 ug/dL
But, potential
confounding difficult
to address
A few studies adjust or consider potential confounding by
SES, exposure to smoking, and/or allergen.
Heterogeneity in evaluation of potential confounding among
the few available studies produces uncertainty.
Jedrychowski et al. (2011),
Pugh Smith and Nriagu (2011)
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Table 4-34 (Continued): Summary of evidence supporting immune causal determinations.
Attribute in Causal
Framework3
Key Evidence
References
Pb Biomarker Levels
Associated with Effects0
Coherence with
toxicological
evidence at relevant
exposures
Most animal studies show elevated IgE in animals with
prenatal and postnatal dietary Pb exposures. Some
inconsistency in animals with relevant Pb concentrations.
Increase in IgE:
Snyder et al. (2000).
Miller et al. (1998)
No increase in IgE:
Chen et al. (2004)
Section 4.6.3
Increase with gestational-
lactational exposure, blood
Pb means 5, 20 ug/dL
Gestational exposure,
maternal blood Pb peak:
30-39 ug/dl_
No increase with
gestational Pb exposure,
blood Pb means 7-8 ug/dL
Most evidence
clearly describes
mode of action
Stimulation of Th2
phenotype
Inflammation
Extensive, consistent evidence of increased production of
Th2 cytokines (e.g., IL-4, IL-6, IL-10) in animals with prenatal
and postnatal (4 weeks juvenile, 3, 8 weeks adults) dietary
Pb exposures. Recent evidence for the role of dendritic cells
in mediating Th2 shift.
Limited available cross-sectional evidence in children, adults
but with limited consideration for potential confounding. Is not
a major consideration in conclusions.
The few available in vitro studies indicate activation Th2 cells
from naTve T cells or over Th1 cells.
Extensive evidence for increased production of TNF-a, IL-6,
ROS, PGE2 by macrophages from animals with prenatal and
postnatal (dietary 4 weeks juvenile, dietary 3, 8 weeks adults,
air 4 days adults) Pb exposure. Supported by in vitro
evidence.
Cross-sectional association observed in children living near
Pb source, adjusted for confounding by age and sex but not
other factors such as SES. Cross-sectional evidence in adults
in NHANES adjusted for inflammatory conditions, smoking
and SES. Is not a major consideration in conclusions.
Section 4.6.6.1
Table 5-7 of the 2006 Pb AQCD (U.S.
EPA, 2006h)
Children: Lutz et al. (1999)
Adults: Kim et al. (2007)
Section 4.6.6.1
Section 4.6.2.1
Sections 4.6.6.1 and 4.6.6.3
Children: Pineda-Zavaleta et al. (2004)
Sections 4.6.6.2, and 4.6.6.3
Adults: Songdej et al. (2010)
Section 4.6.4
Children (ages 6 mo-6 yr):
Concurrent blood Pb group
range 15-19 ug/dL
Adults: Group range
2.5-10.5 ug/dL
Children: Group (ages
6-11 yr)with concurrent
blood Pb levels >10 ug/dL
Adults: Group with
concurrent blood Pb levels
>1.16 ug/dL
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Table 4-34 (Continued): Summary of evidence supporting immune causal determinations.
Attribute in Causal
Framework3
Most evidence
clearly describes
mode of action
(continued)
B cell activation
Key Evidence13
Inconsistent toxicological evidence in animals for B cell
activation by Pb exposure concentration and duration and
strain.
Inconsistent epidemiologic evidence for B cell abundance. B
cell activation not examined.
Pb Biomarker Levels
References'3 Associated with Effects0
Section 4.6.3
Table 4-32 and Section 4.6.3
Decreases in Host Resistance - Likely Causal
Multiple animal
toxicological studies
demonstrate effects
with relevant
exposures
The few studies with relevant dietary Pb exposures
demonstrate increased bacterial infection, sickness behavior,
and mortality in mice. Similar observations in several other
studies with higher Pb exposures.
The few studies with relevant prenatal dietary Pb exposures
show suppressed DTH in rodents. Similar observations in
several other studies with higher Pb exposures.
Dyatlov and Lawrence (2002). Fernandez-
Cabezudo et al. (2007)
Section 4.6.5.1
Chen et al. (2004).
Bunn et al. (2001 a)
Section 4.6.2.3
Blood Pb means
20 ug/dl_ after adult 16-
week exposure, 25 ug/dL
after lactational exposure
Blood Pb means 6.75,
25 ug/dL after gestational
exposure
But, epidemiologic
evidence is limited
and not sufficiently
informative
Epidemiologic studies found associations with increased
respiratory infections but limitations include lack of
consideration for potential confounding or rigorous statistical
analysis, and/or ecologic study design
Children:
Karmaus et al. (2005). Rabinowitz et al.
(1990).
Carreras et al. (2009)
Pb-exposed workers:
Ewers et al. (1982)
Section 4.6.5.1
Children (ages 7-10 yr):
Group with concurrent
blood Pb levels
>3.34 ug/dl_, Group with
cord blood Pb levels
>10ug/dL
Pb-exposed workers:
Concurrent blood Pb levels
21-85ug/dL
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Table 4-34 (Continued): Summary of evidence supporting immune causal determinations.
Attribute in Causal
Framework3
Key Evidence
References
Pb Biomarker Levels
Associated with Effects0
Most evidence
clearly describes
mode of action
Decreased
macrophage
function
Decreased
Th1 cytokine
(IFN-y) production
Decreased
neutrophil or NK
cell function
Decreased macrophage colony formation in animals with
dietary prenatal and postnatal Pb exposure; not widely
examined.
Decreased macrophage phagocytosis in animals and in cell
culture, not widely examined.
Several studies demonstrate decreased NO production by
macrophages from animals with prenatal Pb exposure.
Supported by in vitro evidence.
Cross-sectional association of decreased NO in
macrophages of children living near Pb source with higher
concurrent blood Pb level, adjusted for age and sex but not
Inconsistent evidence in Pb-exposed workers but for
macrophage abundance, not function.
Consistent evidence from a large body of toxicological
studies with prenatal and postnatal (4 weeks juvenile, 3,
8 weeks adults) dietary Pb exposures of animals.
Evidence for reduced neutrophil function in Pb-exposed
workers but lacks consideration for potential confounding and
analogous toxicological evidence.
Toxicological and epidemiologic evidence does not strongly
indicate effects on NK cells.
Section 4.6.2.4
Section 4.6.2.4
Section 4.6.6.2
Pineda-Zavaleta et al. (2004)
Section 4.6.6.2
Pinkerton et al. (1998).
Fischbein et al. (1993).
Conterato et al. (2013)
Section 4.6.6.1 and
Table 5-7 of the 2006 Pb AQCD (U.S.
EPA, 2006h)
Section 4.6.2.5
Section 4.6.2.7
Children (ages 6-11 yr):
Group with concurrent
blood Pb levels >10 ug/dL
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Table 4-34 (Continued): Summary of evidence supporting immune causal determinations.
Attribute in Causal
Framework3
Key Evidence
References
Pb Biomarker Levels
Associated with Effects0
Autoimmunity - Inadequate
Available
toxicological and
epidemiologic
evidence is of
insufficient quantity
or quality
A study in rats shows generation of auto-antibodies with
relevant adult-only dietary Pb exposure for 4 days. Several
other studies have Pb exposure concentrations and/or routes
(e.g., i.p.) with uncertain relevance to humans.
Evidence for increased auto-antibodies in Pb-exposed
workers with high blood Pb levels and limited consideration
for potential confounding, including other workplace
exposures.
Rats: EI-Fawal et al. (1999)
Section 4.6.5.4
Pb-exposed workers:
EI-Fawal et al. (1999)
Section 4.6.5.4
Rats: Blood Pb level
range 11-50 ug/dL
Pb-exposed workers:
Blood Pb level range:
10-40ug/dL
"Described in detail in Table II of the Preamble.
bDescribes the key evidence and references, supporting or contradicting, that contribute most heavily to causal determination. References to earlier sections indicate where full body of
evidence is described.
""Describes the blood Pb levels in children with which the evidence is substantiated and blood Pb levels in animals most relevant to this ISA.
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4.7 Hematological Effects
4.7.1 Introduction
The effects of Pb exposure on red blood cell function and heme synthesis have been
extensively studied over several decades in both human and animal studies. The 1978
National Ambient Air Quality Standard for Lead was set to protect blood Pb levels in
children from exceeding 30 (ig/dL as such levels were associated with impaired heme
synthesis, evidenced by accumulation of protoporphyrin in erythrocytes (U.S. EPA.
1978).
The 2006 Pb AQCD (U.S. EPA. 2006b) reported that Pb exposure significantly decreases
several hematological parameters including hemoglobin (Hb), hematocrit (Hct), mean
corpuscular volume (MCV), mean corpuscular hemoglobin (MCH), and mean
corpuscular hemoglobin concentration (MCHC). Further, the 2006 Pb AQCD reported
that Pb affects developing red blood cells (RBCs) in children and occupationally exposed
adults as noted by anemia observed with concurrent blood Pb >40 (ig/dL. Pb-induced
anemia is thought to occur due to decreased RBC life span and effects on Hb synthesis.
The exact mechanism for these effects is not known, although Pb-induced changes in
Fe uptake or inhibition of enzymes in the heme synthetic pathway may be responsible.
Once Pb enters the cells, it is predominantly found in protein-bound form, with Fib and
aminolevulinic acid dehydratase (ALAD) both identified as targets.
Consistent with the majority of human evidence that high Pb blood levels (i.e., >20 (ig/dL
in adults) are associated with decreased hematological indices, blood Pb levels
>100 (ig/dL were associated with decreased RBC survival in laboratory animals. Effects
on RBC membrane mobility were observed in rats at blood Pb levels as low as 10 (ig/dL,
although the precise mechanisms underlying these effects of Pb are not known. Studies
conducted in animal and in vitro models provide evidence of multiple other effects on
RBC membranes, including altered microviscosity and fluidity, decreased sialic acid
content, decreased lamellar organization, decreased lipid resistance to oxidation (possibly
mediated by perturbations in RBC membrane lipid profiles), and increased permeability.
These alterations to RBC membranes may indicate potential modes of action by which Pb
induces RBC fragility, abnormal cellular function, RBC destruction, and ultimately
anemic conditions. Pb exposure also has been shown to result in increased activation of
RBC scramblase, an enzyme responsible for the expression of phosphatidylserine (PS) on
RBC membranes. This expression of PS decreases the life span of RBCs via phagocytosis
by macrophages. Pb exposure has been observed to alter the phosphorylation profiles of
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membrane proteins, which may influence the activity of membrane enzymes and the
functioning of receptors and channels located on the membrane.
The 2006 Pb AQCD reported that Pb exposure affects heme synthesis in humans and
animals through the inhibition of multiple key enzymes, most notably ALAD, the enzyme
that catalyzes the second, rate-limiting step in heme biosynthesis (Figure 4-35 presents a
schematic representation of the heme biosynthetic pathway). The 2006 Pb AQCD (U.S.
EPA. 2006b) further reported that decreased RBC ALAD activity is the most sensitive
measure of human Pb exposure, in that measurement of ALAD activity is correlated with
blood Pb levels. Concentration-response changes in the ratio of activated/nonactivated
ALAD activity in avian RBCs were observed to be not dependent on the method of Pb
administration. The Pb-associated inhibition of the ALAD enzyme was consistently
observed in RBCs from multiple species, including birds, cynomolgus monkeys, and
humans. Pb was also observed to inhibit other enzymes responsible for heme
biosynthesis, including ferrochelatase, porphobilinogen (PEG) deaminase, and
coproporphyrinogen oxidase. Pb also potentially alters heme biosynthesis through
inhibition of transferrin (TF) endocytosis and Fe transport.
Pb has been found to alter RBC energy metabolism through inhibition of enzymes
involved in anaerobic glycolysis and the pentose phosphate pathway. Pb was also found
to inhibit pyrimidine 5'-nucleotidase (P5N) activity, and the 2006 Pb AQCD indicated
that this might be another biomarker of Pb exposure. Inhibition of P5N results in an
intracellular increase in pyrimidine nucleotides leading to hemolysis and potentially
ultimately resulting in anemic conditions. The 2006 Pb AQCD indicated that
perturbations in RBC energy metabolism may be related to significant decreases in levels
of nucleotide pools, including nicotinamide adenine nucleotide (NAD), possibly due to
decreased NAD synthase activity, and nicotinamide adenine nucleotide phosphate
(NADP) accompanying significant increases in purine degradation products.
The 2006 Pb AQCD identified oxidative stress as an important potential mode of action
by which Pb exposure induced effects on RBCs. Increased lipid peroxidation and
inhibition of antioxidant enzymes in RBCs (e.g., superoxide dismutase [SOD], catalase
[CAT]) were observed following exposure to Pb.
The epidemiologic and toxicological studies published since the 2006 Pb AQCD, largely
support the reported Pb-associated effects on RBC function and heme synthesis.
Epidemiologic studies support previous observations that occupationally-exposed adults
with higher blood Pb levels than the current U.S. general population (>26 (ig/dL) have
decreased RBC numbers. However, a few epidemiological studies investigating
occupationally-exposed adults and pregnant women provide some evidence that more
relevant blood Pb levels, <10 (ig/dL, are associated with decreased RBC numbers,
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possibly through decreased survival of the RBCs. Effects seen in children are largely
consistent with those observed in adults, and a number of toxicological studies support
findings observed in human populations. Recent epidemiologic and toxicological studies
also support previous findings that Pb biomarker levels or exposure in adults, children,
and laboratory animals decreases ALAD activity, as well as the activity of other enzymes
in the heme biosynthetic pathway. Recent epidemiologic and toxicological studies
expand upon the evidence that Pb exposure results in oxidative stress in RBCs. Although
the epidemiologic studies included below are cross-sectional in study design and are
limited to observed associations between concurrent blood Pb levels and hematological
effects, they do improve upon earlier studies as more studies characterize the effects in
children, and they investigate effects in populations with blood Pb levels more
comparable to those in the current U.S. population. Additionally, the associations
observed in these cross-sectional studies are supported by a large number of animal
toxicology studies.
4.7.2 Red Blood Cell Function
As stated in the 2006 Pb AQCD (U.S. EPA. 2006b). Pb poisoning in children has been
associated with anemia. As of 2006, the mechanism for this was not clear, but it was
determined not to be due to Fe deficiency, which can be found to occur independently of
Pb exposure. However, Zimmerman et al. (2006) found that administration of an
Fe-fortified diet to non-anemic (or mildly anemic) Fe-deficient Indian children aged 5 to
9 years (baseline median BLL [range] = 11.6 [2.1-24.1] (ig/dL) reduced median blood Pb
levels to 8.1 (ig/dL (range: 3.1-21.9 (ig/dL). This reduction in blood Pb levels was
statistically significantly greater than in children fed a non-fortified diet for 30 weeks
(median blood Pb level [range]: 10.2 [4.4-25.3] (ig/dL). Differences in blood Pb levels
between diet intervention groups were not different after only 14 weeks. Although a
number of epidemiologic studies found decreases in RBCs and/or Hct levels associated
with higher blood Pb levels, it is not known whether this is due to reduced RBC survival
or a decrease in RBC production. Regardless, decreased RBC survival and hematopoiesis
can be expected to occur simultaneously, and any effect on RBC numbers is likely a
combination of the two modes of action.
4.7.2.1 Pb Uptake, Binding, and Transport into Red Blood Cells
The 2006 Pb AQCD (U.S. EPA. 2006b) reported that Pb uptake into human RBCs occurs
via passive anion transport mechanisms. Although Pb can passively cross the membrane
in both directions, little of the Pb is found to leave the cell after entry. Simons (1993b)
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found that in vitro uptake of 203Pb (1-10 (iM) occurred via an anion exchanger while the
efflux occurred via a vanadate-sensitive pathway. After entry into the RBC, radioactive
Pb was found to partition with Hb at a ratio estimated to be about 6000:1 bound to
unbound (Simons. 1986). However, Bergdahl et al. (1997a) suggested that ALAD was
the primary Pb-binding protein and not Hb. The 2006 Pb AQCD also reported that the
majority (approximately 98%) of Pb accumulates in RBC cytoplasm bound to protein,
and only about 2% is found in the membrane. This is related to the high ratio of Pb in
RBCs compared to plasma Pb. Further information on Pb binding and transport in blood
can be found in the kinetics section of Chapter 3 (Section 3.2).
Although no recent studies were identified that examined transport of Pb into RBCs, Lind
et al. (2009) recently observed that several Zn ionophores (8-hydroxyquinoline
derivatives and Zn and Na pyrithione) were able to effectively transport Pb out of RBCs
into the extracellular space.
4.7.2.2 Red Blood Cell Survival, Mobility, and Membrane Integrity
A number of cross-sectional studies have investigated the effect of exposure to Pb on
various inter-connected and related hematological parameters in children and adults. As
these studies were cross-sectional in design, there is uncertainty regarding the
directionality of effects and the magnitude, timing, frequency, and duration of Pb
exposure that contributed to the observed associations. Additionally, unless explicitly
noted, potential confounding by subject characteristics and other workplace or residential
exposures was not accounted for in these studies. Adults and children exposed to Pb may
also have been co-exposed to other contaminants that can affect the hematological
system, and the potential for co-exposure was not assessed in most studies.
In an earlier cross-sectional study of children in Idaho (aged 1-9 years) with blood Pb
levels ranging from 11 to 165 (ig/dL (approximately 40% were >40 (ig/dL), a 10%
probability of anemia (Hct <35%) was predicted (in association with blood Pb levels of
-20 (ig/dL [age 1 year], 50 (ig/dL [age 3 years] and 75 (ig/dL [age 5 years]) (Schwartz et
al.. 1990). More recent studies have also demonstrated negative effects on hematological
parameters in children due to the Pb exposure. Ahamed et al. (2006) studied 39 male
urban adolescents in India who were separated into groups according to their blood Pb
level (Group 1: <10 (ig/dL [mean 7.4 (ig/dL], Group 2: >10 (ig/dL [mean 13.27 (ig/dL]).
Although the groups were similar in age (mean [SD]: 16.59 [0.91] versus 16.76 [0.90]
years, respectively), height, weight, and BMI (therefore, not considered to be potential
confounders), Group 2 had a significantly lower packed cell volume (PCV) compared to
Group 1. In a related study, Ahamed et al. (2007) investigated the relationship between
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blood Pb level, anemia, and other hematological parameters in urban children in India
(N = 75). Children were split into two groups as above: Group 1 had blood Pb levels
<10 (ig/dL (mean [SD]: 6.89 [2.44] (ig/dL, n = 19), whereas Group 2 had blood Pb levels
>10 (ig/dL (mean [SD]: 21.86 [7.58] (ig/dL, n = 56). As with the earlier study, ages were
similar between the two groups: mean [SD]: 4.68 [1.49] and 4.11 [1.77] years,
respectively. Hb and Hct were significantly decreased in Group 2, compared to Group 1
after adjusting for age, sex, and area of residence. Children in Group 2 had an increased
odds ratio of anemia (OR: 2.87 [95% CI: 1.60, 2.87]) compared to Group 1 after
adjustment for age, sex, and area of residence.
In a cross-sectional study measuring blood Hb as the independent variable, blood Pb
levels were observed to decrease with increasing blood Hb. Riddell et al. (2007) found
that 21% of children, who were 6 months to 5 years of age living in the rural Philippines,
had concurrent blood Pb levels >10 (ig/dL (total population mean: 6.9 (ig/dL). After
controlling for potential confounding by age, sex, birth weight, and history of
breastfeeding, Hb levels were inversely related to blood Pb level, with a decrease of 3%
blood Pb associated with every 1 g/dL increase in Hb. Among children aged 6-36 months
(N = 222) living in Montevideo, Uruguay, 32.9% had blood Pb levels greater than
10 (ig/dL (population mean [SD]: 9.0 [6.0] (ig/dL) (Queiroloetal.. 2010). The mean
[SD] Hb concentration was 10.5 [1.5] g/dL, and 44.1% of children were diagnosed as
anemic (Hb <10.5 g/dL). Blood Pb levels were significantly higher in anemic children
compared to non-anemic (mean [SD]: 10.4 [6.8] versus 7.9 [5.1] (ig/dL), and anemic
children were more likely to have elevated blood Pb after controlling for age and
mouthing behavior (OR = 1.9, 95% CI: [1.098, 3.350]). The likelihood of elevated blood
Pb was more pronounced in anemic children younger than 18 months (OR = 3.1, 95% CI:
[1.3,7.4]).
Similarly, in a cross-sectional study of 340 children (aged 1-5 years) from Karachi,
Pakistan, mildly-anemic and severely-anemic children (mean [SD] Hb levels: 8.9 [0.9]
and 7.4 [0.5] g/dL, respectively) had lower Hb levels but higher blood Pb levels
compared to non-anemic children (mean [SD] Hb: 12.1 [1.3] g/dL). Mean [SD] blood Pb
levels in the mildly-anemic, severely-anemic, and non-anemic children were 14.9 [0.81],
21.4 [2.7], and 7.9 [1.7] (ig/dL, respectively (p <0.01) (Shah etal.. 2010). Additionally,
Hct, RBC count, and MCV were all decreased in anemic children versus non-anemic
children. Although statistical analyses were not reported, the levels of Hb, Hct, RBC
count, and MCV in anemic children all fell outside of the reported normal range for these
parameters, whereas the reported values in non-anemic children did not. Blood Pb level
was negatively correlated with Hb level in all groups, with the magnitude of negative
correlation increasing with increasing severity of anemia: r = -0.315 (non-anemic
children), -0.514 (mildly-anemic), and -0.685 (severely- anemic). In Fe-deficient anemic
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children (n = 23) from Denizli, Turkey, mean (SD) serum Pb levels were statistically
(p <0.05) increased compared to healthy children (n = 179): 0.013 (0.004) versus 0.008
(0.001) (ig/dL, respectively (Turgut et al., 2007). The Fe-deficient children were
observed to have decreased Hb, MCV, RBC, and ferritin compared to controls, but
increased RBC distribution width. In 140 children from southern Brazil aged 2-11 years,
living within 25 km of a Pb smelter, blood Pb levels were not observed to differ between
anemic and non-anemic children (mean [SD]: 10.36 [6.8] versus 9.73 [5.8] (ig/dL,
p = 0.98) (Rondo et al.. 2006). However, blood Pb levels were significantly negatively
correlated with Hb in anemic children (r = -0.41,p = 0.01); this relationship was not
observed in non-anemic children (r = 0.018, p = 0.84).
In children aged 5-9 years (n = 189) without anemia living in Cartagena, Columbia, a
smaller percentage (4.7%) of children had blood Pb >10 (ig/dL (mean [SD]: 5.49
[0.23] (ig/dL). The only hematological parameters that fell outside of their reference
values were MCV and MCH, which were negatively correlated with blood Pb levels
(r = -0.159 [p = 0.029] and -0.171 [p = 0.019], respectively) (Olivero-Verbel et al.. 2007).
RBC count, which was not observed to differ from reference values, was positively
correlated with blood Pb level (r = 0.208, [p = 0.004]). In a group of 268 Lebanese
children, children aged 11-23 months with blood Pb levels >10 (ig/dL had increased
likelihood of having Fe-dependent anemia and transferrin saturation (TF <12%)
compared to age-matched children with blood Pb levels <10 (ig/dL (OR = 4.59, 95% CI:
[1.51, 13.92]) (Muwakkit et al.. 2008). In children aged 24-35 months, higher blood Pb
level was not associated with increased likelihood of either effect. Huo et al. (2007)
found that children (less than 6 years of age) living near an area where electronic waste
was recycled in China had significantly higher mean blood Pb levels than did children in
the neighboring town with no waste recycling (15.3 versus 9.94 (ig/dL). However,
contrary to the findings above, no difference was detected in the mean Hb levels of the
children in the two towns (12.76 g/dL in children from the waste recycling town versus
12.35 g/dL in children from the town with no recycling).
In adult, occupationally exposed populations, decreased erythrocyte numbers and Hb
were observed in multiple, earlier cross-sectional studies investigating workers with
blood Pb levels >40 (ig/dL (Solliwav et al.. 1996; Horiguchi et al.. 1991; Poulos et al..
1986). However, a larger, longitudinal study (Hsiao et al.. 2001) observed that
occupationally-exposed adults exhibited erythrocyte counts and Hct that were positively
associated with blood Pb levels. Most of the recent occupationally-exposed groups
represent populations highly exposed to Pb, with mean blood Pb levels ranging from 26-
74 (ig/dL. Although effects observed within these groups may not be generalizable to the
general population as a whole, they are useful in demonstrating consistent effects on a
number of hematological parameters, including Hb, MCV, MCH, MCHC, total RBCs,
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and packed cell volume (PCV) (Khan et al. 2008; Patil et al.. 2006a: Patil et al.. 2006b:
Karita et al. 2005).
A few recent cross-sectional occupational studies did investigate the effect of moderate
occupational Pb exposure on hematological parameters. In gas station attendants in
Sarajevo, Bosnia and Herzegovina, workers (mean [SD] duration of exposure: 12.1 [9.1]
years) had significantly increased blood Pb levels (mean: 5.96 (ig/dL) in 2008, compared
to the same population that was previously examined in 2003 (mean: 4.07 (ig/dL; mean
[SD] duration of exposure: 10.4 [5.5]). Levels of MCH and MCHC were significantly
decreased when assessed in 2008, compared to the 2003 measurements, although RBC
numbers, Hb, Hct, and MCV were increased in 2008 compared to 2003. Positive
correlations were observed in all subjects between blood Pb and RBC count, Hb, and
MCH (r = 0.241, 0.201, and 0.213, respectively; p <0.05). No control group was included
in this study (Cabaravdic et al.. 2010). Ukaejiofo et al. (2009) studied the hematological
effects of Pb in 81 male subjects moderately exposed to Pb at three different
manufacturing companies in Nigeria for durations between six months and 20 years. The
exposed individuals had a mean blood Pb level of 7.00 (ig/dL (± 0.07 (ig/dL) compared to
3 (ig/dL (±0.19 (ig/dL) in controls drawn from industries not involved in Pb handling
(control group I) and 2 (ig/dL (±0.04 (ig/dL) in controls drawn from the general
population (control group II). It should be noted that the range of reported blood Pb levels
in exposed workers ranged from 30 to 0 (ig/dL. The limit of detection for blood Pb was
reported as 0.05 (ig/dL, meaning that there possibly is some bias towards zero in the
reported mean blood Pb levels. Pb-exposed workers had significantly reduced Hb and
PCV levels and increased percentage of reticulocytes compared to controls. Although the
differences were statistically significant between the exposed and control subjects, the
study authors stated that the levels in the exposed subjects were at the lower range of
normal for Nigerians. The percent cell lysis did not significantly differ between controls
and exposed workers; however, when workers and controls were stratified by age, there
was a significant increase in cell lysis in workers under age 30 compared to similarly
aged controls in group II (p <0.01). Similarly, stratification of subjects by duration of
exposure revealed that MCHC was decreased in exposed workers (6-60 months of
exposure). Conterato et al. (2013) investigated hematological parameters in automotive
painters exposed to Pb in Brazil. Exposed painters had a mean [SEM] blood Pb
concentration of 5.4 [0.4] (ig/dL compared to 1.5 [0.1] (ig/dL in controls. The mean
[SEM] duration of exposure to Pb in painters was 133.9 [14.5] months, whereas the
controls were not occupationally exposed to Pb. Although Hct, Hb concentration, and the
number of RBCs were significantly decreased in painters compared to controls, they were
not correlated with blood Pb levels; however, these parameters were correlated with
blood Cd2+ levels, which were also significantly elevated in painters compared to
controls.
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Taken together, the above occupational studies provide consistent evidence that high
(mean blood Pb >26 (ig/dL) occupational exposure to Pb reduces the number of RBCs in
circulation. Additionally, the Ukaejiofo et al. (2009) study provides suggestive evidence
that concurrent blood Pb levels below 10 (ig/dL (7.0 (ig/dL) may also be associated with
decreased RBC survival. Although the decrease in RBCs observed in highly exposed
worker populations may be explained by both decreased RBC survival and/or disruption
of hematopoiesis, the observation of increased reticulocytes in Ukaejiofo et al. (2009)
seems to represent compensation for decreased RBC survival due to Pb exposure.
In a non-occupational study, the associations between blood Pb levels, Ca2+, Fe, and Hb
were investigated in 55 pregnant Brazilian women (21.9% were 14-19 years old, 74.5%
were 20-34 years olds, and 3.6% were > 35 years old) (Zentner et al.. 2008). The
majority of women (across all age groups) had concurrent blood Pb levels <5 (ig/dL
(58.2%), although the mean blood Pb level was not reported; only 5.4% of women had
blood Pb levels >10 (ig/dL. The vast majority of the women (78.2%) were also observed
to have adequate levels of Hb (> 11 g/dL). In a multiple linear regression model, blood
Pb level was observed to be negatively associated with Hb ((3 = -0.359 per 1 (ig/dL Pb),
when controlling for age, BMI, income, energy intake, Ca2+ intake, vitamin C intake, and
Fe intake.
The associations of blood Pb levels with hematological parameters observed in
epidemiologic studies are clearly supported by a number of animal toxicology studies
reporting blood Pb levels relevant to this ISA, i.e., <10 (ig/dL. Hb concentrations in
plasma (a marker of RBC hemolysis) was significantly increased in rats exposed to
Pb acetate (1,000 ppm in drinking water for 9 months; blood Pb level: 7.1 (ig/dL)
compared to controls (Baranowska-Bosiacka et al., 2009). In a complementary in vitro
experiment, a concentration-dependent increase in the amount of hemolysis was observed
in human RBCs exposed to Pb at concentrations ranging from 0.1-100 (iM for 5-30
minutes. Hemolysis was increased even at the lowest concentration tested (i.e., 0.1 (iM).
Pb-induced hemolysis in these experiments may be due to inhibition of RBC
phosphoribosyltransferases (Section 4.7.2.5). In weanling rats (PND25 days, n = 10)
whose dams were exposed to Pb acetate in drinking water (2.84 mg/mL, approximating
mean [SD] daily exposures of 342.57 [28.11] and 744.47 [29.27] mg/kg [dam weight]
during gestation and lactation, respectively), blood Pb level was significantly elevated
compared to controls (mean [SE]: 69.8. [7.82] versus 0.54 [0.08] (ig/dL). The only
hematological parameter affected by Pb exposure was Hct, which was decreased in
exposed rats (mean [SE]: 27.3 [0.5]%) versus controls (33.4 [0.3]%) (Molina et al..
2011). In rats treated with 25 mg Pb/kg by oral gavage for 4 weeks, the average plasma
Pb concentration was 6.5 (ig/dL (9.6-fold higher than controls), and statistically
significant decreases in Hct, Hb, and RBCs were observed (Lee et al.. 2005). Effects on
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erythrocyte survival were similar in mice treated with Pb nitrate (50 mg/kg via gavage for
40 days): mean [SD] blood Pb levels were 1 .72 [0.02] (ig/dL versus 0.09 [0.01 1] (ig/dL
in control mice, and exposed mice had significantly reduced total RBC counts, total
leukocyte counts, Hb, lymphocytes, and monocytes compared to controls (p <0.001)
(Sharma et al.. 201 Ob).
A number of toxicological studies also reported similar hematological effects, but did not
report final blood Pb concentrations. Rats exposed to Pb acetate (2 g/L in drinking water
for 30 days) had significantly decreased RBCs, Hb, PCV, MCH, and MCHC compared to
controls (p <0.05) (Simsek et al.. 2009). but not a disruption of hematopoiesis. Mice
exposed to Pb acetate (1 g/L in drinking water for 90 days, but not those exposed for 15
or 45 days), had significantly decreased RBC counts and Hct compared to controls
(p<0.05) (Marques et al.. 2006). Spleen weights were observed to be increased relative to
body weight in animals exposed to Pb for 45 days. Mice injected daily with Pb acetate
(50 mg/kg subcutaneously) had significantly reduced Hb, MCV, MCH, and MCHC
compared to controls injected with 5% dextrose (Wang et al.. 2010h).
Some toxicological studies found no evidence of hematological effects in animals
following exposure to Pb. Male rats exposed to Pb acetate in their drinking water for
4 weeks at concentrations ranging from 100-1,000 ppm had a concentration-dependent
increase in blood Pb levels (range: 6.57-22.39 (ig/dL) compared to controls (0.36 (ig/dL),
but there were no significant changes in any of the hematological parameters (complete
blood cell count performed) measured at the end of treatment (Lee et al.. 2006b). Slight,
statistically nonsignificant increases in PS expression on RBC membranes were also
observed. Similarly, exposure of male rats to 5,000 ppm Pb nitrate in drinking water
(blood Pb not reported) for three weeks had no affect on any measured hematological
parameter (Gautam and Flora. 2010). In vitro experiments with rat and human blood did
not demonstrate a significant increase in hemolysis after 4 hours of treatment with
Pb acetate at concentrations up to 10
Although Pb exposure has been consistently shown to shorten RBC life span and alter
RBC mobility, as of the 2006 Pb AQCD, the mechanism of this was not well understood.
While the mechanism is still not fully understood, there has been some indication for a
role of free Ca2+. Occupational studies investigating highly Pb-exposed worker
populations (mean blood Pb >28 (ig/dL) observed increased intracellular free Ca2+ levels
([Ca2+];) in RBCs, and decreased RBC membrane Ca2+/Mg2+ATPase activity in workers
compared with unexposed controls (Abam et al.. 2008; Quintanar-Escorza et al.. 2007).
[Ca2+]j levels were highly correlated with blood Pb levels even among unexposed control
populations with mean blood Pb levels of approximately 10 (ig/dL (9.9 ± 2 (ig/dL)
(Ouintanar-Escorza et al.. 2007). Changes in [Ca2+]j were associated with increased
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fragility of the RBCs and dramatic morphological alterations, including the increased
presence of echinocytes (cells without normal biconcave shape) and crenocytes
(speculated cells) in Pb-exposed workers.
Similar to the associations observed in Quintanar-Escorza et al. (2007). [Ca2+]j increased
in a concentration-dependent manner when RBCs from healthy human volunteers were
exposed (in vitro) to 0.2 or 0.4 (iM Pb nitrate for 24 or 120 hours (0.4 (iM Pb nitrate
roughly approximates 10 (ig/dL Pb, although concentrations in exposure media are not
directly comparable to blood Pb levels) (Quintanar-Escorza et al., 2010). The increase in
[Ca2+]j levels was observed to be related to increased Ca2+ influx and decreased efflux.
As was observed among highly Pb-exposed workers, changes in [Ca2+]j were associated
with increased fragility of the RBCs and dramatic morphological alterations following
exposure to 0.4 (iM Pb. Similarly, Ciubar et al. (2007) found that RBC morphology was
disrupted, with > 50% RBCs having lost the typical discocytic morphology and
displaying moderate to severe echinocytosis following exposure to Pb nitrate
concentrations of 0.5 (iM or higher for 24 hours. Exposure of RBCs to higher
concentrations (concentrations not stated) of Pb nitrate resulted in cell shrinkage. In rats
exposed to 200 ppm Pb acetate via drinking water for three months (mean [SD] blood Pb
level: 40.63 [9.21] (ig/dL), the cholesterol/phospholipid ratio of RBC membranes was
increased, indicating that RBC membrane fluidity was decreased.
Khairullina et al. (2008) observed that the surface profiles of RBC membrane shadows
incubated with 0.5-10 (iM Pb acetate for three hours were much smoother than were
untreated RBC membranes when examined by atomic force microscopy. The authors
postulated that the observed smoothing in Pb-treated RBC membranes may be due to
clusterization of band 3 protein. Band 3 (anion exchanger 1 [AE1]), is a
chloride/bicarbonate (C1~/HCO3~) exchanger and is the most abundant protein in RBC
membranes. AE1 is integral in carbon dioxide (CO2) transport and linkage of the cellular
membrane to the underlying cytoskeleton (Akel et al.. 2007; Su et al.. 2007). The
observed smoothing of the RBC membrane may due to Pb interfering with how the
membrane attaches to the cytoskeletal structure of the RBC through perturbation of the
normal activity of AE1.
Eryptosis
Eryptosis is the suicidal death of RBCs. It is characterized by cell shrinkage, membrane
blebbing, and cell membrane phospholipid scrambling associated with PS exposure on
the cell membrane that leads to cell destruction via macrophages (Foller et al.. 2008;
Lang et al.. 2008). As previously reported in the 2006 Pb AQCD (U.S. EPA. 2006bX
Kempe et al. (2005) found that exposing human RBCs to Pb at concentrations ranging
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from 0.3 (iM to 3 (iM caused increased activation of K+ channels that led to cell
shrinkage and scramblase activation. The activation of scramblase increased the exposure
to PS on the cell membrane, which causes an increase in the destruction of the RBCs by
macrophages.
Consistent observations were made in recent studies that included in vitro and in vivo
evidence. Shin et al. (2007) found that in vitro exposure of human RBCs to 1-5 (iM
Pb acetate increased PS expression in a time- and concentration-dependent manner. The
maximum mean [SE] increase in expression of PS was 26.8% [3.15] (compared to
deionized water), following exposure to 5 (iM Pb for four hours. Scramblase activity was
increased in Pb-exposed RBCs, and [Ca2+]l5 which regulates scramblase activation, was
also increased in exposed RBCs. Flippase, which translates PS exposure to inner
membranes, is inhibited by high levels of [Ca2+]j and was observed to exhibit reduced
activity following Pb exposure. The inhibition of flippase is additionally influenced by
the depletion of cellular adenosine triphosphate (ATP). ATP levels were decreased in a
concentration-dependent manner following exposure to Pb. To corroborate these findings
in vivo, Shin et al. (2007) treated male rats with Pb acetate (i.p. to 25, 50, or 100 mg/kg;
blood Pb not reported). Expression of PS was observed to increase in a concentration-
dependent manner at concentrations > 50 mg/kg, confirming the in vitro results. No
hemolysis or microvesicle formation was observed in the in vitro or in vivo experiments.
In a follow-up study, the same laboratory observed that in vitro exposure of human RBCs
to much lower concentrations of Pb acetate (0.1, 0.25, and 0.5 (iM) also induced PS
expression. Most notably, exposure to 0.1 (iM Pb for 24 hours increased PS expression
on RBC membranes by approximately 20% (Jang et al.. 2011). Accompanying the
increased expression of PS (associated with Pb exposure) was the presence of abnormal
echinocytic RBCs. Unlike the Shin et al. (2007) study described above, incubation of the
RBCs with low concentrations of Pb (0.1 (iM) induced the generation of microvesicles,
which also expressed PS on their membranes in this (Jang et al.. 2011) study. At 0.5 (iM,
Pb-exposed RBCs with externalized PS were observed to be targeted and engulfed by
differentiated macrophages. Similar ex vivo effects were observed in rat erythrocytes four
hours after oral exposure (0, 10 and 50 mg/kg) to Pb, although higher concentrations
were generally required. PS expression on the rat erythrocytes was also observed. To
corroborate these in vitro and ex vivo findings, rats were also exposed in vivo to 0, 50,
250, or 1,000 ppm Pb acetate in drinking water for 4 weeks. At 1,000 ppm, Hb and Hct
were significantly decreased relative to control, and liver and spleen weights were
increased. At the two highest doses, Fe accumulation was observed in the spleen, a clear
sign of increased RBC clearance via phagocytosis.
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Ciubar et al. (2007) also found that exposure to Pb nitrate (0.5-2 (iM) resulted in an
increase in PS exposure in RBCs and cell shrinkage, which the authors stated were
indicators of cell apoptosis. As reported above, Khairullina et al. (2008) observed
Pb-induced RBC membrane smoothing that may be due to alterations in AE1 activity.
Disruptions in AE1 activity may also result in enhanced PS exposure and premature cell
death. Akel et al. (2007) observed that in AE1"" knockout mice, Pb-induced PS exposure
was much greater than that in wild type mice. Decreased RBCs and increased
reticulocytes were also observed, an indication of high cell turnover.
4.7.2.3 Red Blood Cell Hematopoiesis
Erythropoietin (EPO) is a glycoprotein hormone excreted by the kidney to promote the
development of RBCs in the bone marrow. As reported in the 2006 Pb AQCD, analyses
of the cohort of children in Yugoslavia observed that EPO was increased in children aged
4.5 and 6.5 years of age living in a town near Pb sources (blood Pb levels >30 (ig/dL)
compared to children living in more distant town (blood Pb levels <10 (ig/dL), when
stratified by Hb concentrations (Graziano et al.. 2004; Factor-Litvak et al.. 1999; Factor-
Litvak et al., 1998). These differences were not observed in children aged 9.5 or 12 years.
With adjustment for Hb concentrations, blood Pb levels were observed to be significantly
associated with EPO levels at ages 4.5 and 6.5 years when considering all children
together. No significant association was observed at ages 9.5 and 12 years. Hb was not
observed to differ at any age between towns, thus possibly indicating that
hyperproduction of EPO is necessary to maintain Hb levels in young children living near
Pb sources. The authors postulated that increases in EPO in younger children reflect bone
marrow hyperactivity to counteract RBC destruction, whereas the lack of EPO elevation
in older children may reflect a transitional period where increasing renal and bone
marrow toxicity leads to decreases in EPO observed later in life, as observed in anemic,
pregnant women (Graziano et al.. 1991). Decreased EPO concentrations were also
observed in association with Pb exposure in adults in two cross-sectional studies cited in
the 2006 Pb AQCD (Osterode et al.. 1999; Romeo etal. 1996).
Consistent with findings that EPO is negatively associated with blood Pb levels in adults,
Sakata et al. (2007) observed that non-anemic tricycle taxi drivers (n=27) working in
Kathmandu, Nepal (blood Pb level: 6.4 (ig/dL) had significantly lower levels of EPO
(12.7 versus 18.8 mU/mL) compared to non-driver controls (blood Pb level: 2.4 (ig/dL).
In taxi drivers, there was an inverse relationship between the level of serum
erythropoietin and blood Pb level (r = -0.68, p <0.001). Blood Pb level was not associated
with any other hematological effects.
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Recent toxicology studies of cytotoxicity and genotoxicity in RBC precursor cells
support the observations that Pb exposure disrupts normal hematopoiesis. Cytotoxicity
and genotoxicity in RBC precursor cells are strong indications of altered hematopoiesis in
bone marrow. Celik et al. (2005) observed that treatment of female rats with Pb acetate
(140, 250, or 500 mg/kg via gavage once per week for 10 weeks; blood Pb not reported)
resulted in decreased numbers of polychromatic RBCs (PCE) and increased numbers of
micronucleated PCEs, compared to controls (p <0.001). Alghazal et al. (2008b) exposed
male and female rats to 100 mg/L Pb acetate daily in drinking water for 125 days (blood
Pb not reported) and observed increases in micronucleated PCEs in female rats (p = 0.02)
but no significant reduction in the ratio of PCEs to normochromic RBCs (NCE). In male
rats, an increase in micronucleated PCEs was observed (p <0.001) along with a decrease
in the PCE/NCE ratio (p = 0.02). While the results from Alghazal et al. (2008b) indicate
that Pb is cytotoxic in male rats only, but is genotoxic in both sexes, results from Celik et
al. (2005) indicate that Pb is cytotoxic in female rats as well. Mice exposed to Pb acetate
(1 g/L in drinking water for 90 days; blood Pb not reported) had statistically significant
increases in micronucleated PCEs; a small, but statistically nonsignificant decrease in the
PCE/NCE ratio was also observed (Marques et al.. 2006).
4.7.2.4 Membrane Proteins
While there have been few studies, evidence included in the 2006 Pb AQCD indicated
there are effects of Pb on changes in RBC proteins. Huel et al. (2008) found that newborn
hair and cord blood Pb levels (mean [SD]: 1.22 [1.41] (ig/g and 3.54 [1.72] (ig/dL
respectively) were negatively associated with Ca2+ATPase activity in plasma membranes
of RBCs isolated from cord blood after controlling for gestational age and maternal Ca
pump activity. However, newborn hair Pb levels were more strongly associated with cord
Ca2+-pump activity than were cord blood Pb (p <0.0001 versus p <0.05). Maternal blood
Pb levels were not correlated with Ca2+-pump activity in maternal or newborn cord blood.
Pb-induced disruptions in Ca2+ homeostasis in RBCs can lead to cytotoxicity and
necrosis, and these effects may be representative of cellular dysfunction in other organ
systems.
In RBC membranes from Pb-exposed workers, Fukumoto et al. (1983) used
polyacrylamide electrophoresis analysis and found increased levels of polypeptides in
bands 2, 4, 6, and 7 and decreased levels of polypeptides in band 3. Apostoli et al. (1988)
found changes in RBC membrane polypeptides, including a significant decrease in band
3, in occupationally exposed adults with blood Pb levels greater than 50 (ig/dL. Apostoli
et al. (1988) suggested that band 3 may represent an anion channel protein, whereas
Fukumoto et al. (1983) suggested that the changes in the RBC membrane polypeptides
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may cause changes in membrane permeability. Exposure to Pb acetate at concentrations
above 0.1 (iM for 60 minutes has also been found to increase the phosphorylation of
proteins in human RBC membranes in vitro (Belloni-Olivi et al., 1996). Phosphorylation
did not occur in cells depleted of protein kinase C (PKC), indicating a PKC-dependent
mechanism.
4.7.2.5 Red Blood Cell Energy Metabolism Enzymes
RBCs use high energy purine nucleotides (i.e., ATP and guanine triphosphate [GTP]) to
support basic metabolic functions. In mature RBCs, these nucleotides are synthesized via
salvage reactions through either an adenine pathway, which requires adenine
phosphoribosyltransferase (APRT), or an adenosine pathway, which requires adenosine
kinase. The 2006 Pb AQCD (U.S. EPA. 2006b) reported that Pb significantly reduces the
nucleotide pool (including NAD and NADP), as well as increases purine degradation
products resulting in altered RBC energetics. Since the 2006 Pb AQCD was published,
there have been few studies examining Pb effects on energy metabolism. Baranowska-
Bosiacka et al. (2009) examined the effects of Pb on RBC APRT and hypoxanthine-
guanine phosphoribosyltransferase (HPRT). In an in vitro experiment, APRT and HPRT
were measured in lysate of human RBCs after exposure to Pb at a concentration range
from 0.1 to 100 (iM for 5-30 minutes. Complementary in vivo tests measured APRT and
HPRT in RBC lysate from rats exposed to Pb acetate (1,000 ppm) in drinking water for
9 months. Both the in vivo and vitro studies found a significant decrease in both HPRT
and APRT levels. The levels in human RBCs were significantly decreased in vitro after
only 5 minutes of exposure to the 0.1 (iM concentration, and the decrease was also
concentration-dependent. However, the study authors considered the inhibition moderate
(30-35%) even with the highest Pb levels used in vitro. Shin et al. (2007) found a
concentration-dependent decrease in intracellular ATP in human RBCs in vitro with
significant decreases, found even with the lowest concentration (i.e., 1
4.7.2.6 Other Enzymes
The 2006 Pb AQCD (U.S. EPA. 2006b) reported that K+ permeability was increased by
Pb exposure due to altered sensitivity of the membrane Ca2+-binding site that caused
selective efflux of K+ ions from the RBC membrane. However, inhibition of the RBC
Na+/K+ATPase is more sensitive to Pb exposure than is the inhibition of
Ca2+/Mg2+ATPase. Few recent studies were found that examined the effects of Pb
exposure on other enzymes. Ekinci et al. (2007) tested the effects of Pb exposure on two
carbonic anhydrase isozymes (I and II) isolated from human RBCs. Carbonic anhydrases
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are metalloproteins that use Zn to catalyze the equilibrium between CO2 and bicarbonate
in the cells of higher invertebrates. Although investigators found that Pb nitrate inhibited
both carbonic anhydrase isozymes in a concentration-dependent manner, the
concentrations used (i.e., 200-1,000 (iM) were above those that would be physiologically
relevant. Inhibition of isozyme I was noncompetitive, while the inhibition for isozyme II
was uncompetitive. Bitto et al. (2006) examined the mechanisms of action of Pb-induced
inhibition of P5N, an enzyme important in the pyrimidine salvage pathway that requires
Mn for normal activity. Pb was observed to bind directly to the active site of the enzyme
in a different position than the Mn, thus possibly resulting in improper protein folding
and inhibition of activity.
4.7.2.7 Red Blood Cell Oxidative Stress
It has been suggested that the Pb-associated decreases in ALAD activity result in
increased oxidative stress, owing to the buildup of ALA. ALA can act as an electron
donor in the formation of reactive oxygen species (ROS) (Nemsadze et al.. 2009;
Ahamed and Siddiqui. 2007). Many epidemiologic and toxicological studies have found
an association between the level of blood Pb and lipid peroxidation, antioxidant levels, or
indicators of ROS production. The same limitations regarding cross-sectional studies
listed in Section 4.7.2.2 (including uncertainty in directionality of effects and specific
information regarding exposure) apply to the epidemiologic studies investigating RBC
oxidative stress. Additionally, potential confounders and co-exposures were not
considered in the majority of these studies. However, in studies where confounders were
considered, they are explicitly delineated in the text.
Oxidative Stress, Lipid Peroxidation, and Antioxidant Enzymes
Malondialdehyde (MDA) is an end product of lipid peroxidation and is commonly
measured as an indicator of oxidative stress. Evidence of lipid peroxidation has been
observed in children moderately exposed to Pb. Ahamed et al.(2008; 2006. 2005)
investigated the relationship between blood Pb levels and antioxidant enzyme levels and
lipid peroxidation in children in India. Among children (N = 62) aged 4-12 years in
Lucknow, India, children with mean blood Pb levels of 11.39 (SD: 1.39) (ig/dL had
increased measures of lipid peroxidation and decreased glutathione (GSH) levels
compared to children with mean blood Pb levels of 3.93 (SD: 0.61) or 7.11 (SD:
1.25) (ig/dL (Ahamed et al., 2005). Catalase activity was decreased in children with a
mean 7.11 (SD: 1.25) (ig/dL blood Pb level, compared to children with mean 3.93 (SD:
0.61) (ig/dL blood Pb level. Additionally, blood Pb levels were found to be significantly
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positively correlated with MDA and CAT, and negatively correlated with GSH. In a
similar study, Ahamed et al. (2006) observed significantly higher levels of CAT and
J ' V / O J O
MDA in children with a mean 13.27 (ig/dL blood Pb level compared to children with a
mean 7.40 (ig/dL blood Pb level, with other characteristics such as age, height, wieght,
and BMI not differing between the two groups and thus, not considered as potential
confounders. Examining all the study subjects together, investigators found a correlation
between blood Pb level and blood MDA and RBC CAT levels, as well as an inverse
relationship between ALAD activity and MDA and CAT levels. Among Indian children
with neurological disorders, blood Pb levels were significantly increased compared to
healthy control children (18.60 versus 10.37 (ig/dL respectively) (Ahamed et al., 2008).
Potential confounding characteristics such as age, sex, area of residence, and SES were
not observed to be statistically different between the two groups, and therefore, were not
included in statistical analyses. In addition, the following indicators of oxidative stress
were elevated among case children: increased blood MDA, RBC SOD and CAT levels,
and decreased blood GSH levels. GPx levels were similar between the two groups.
Typical indicators of Pb exposure (active/nonactive ALAD ratio) were found to be
correlated with lipid peroxidation and oxidative stress. Children aged 3-6 years old living
near a steel refinery in China with blood Pb levels > 10 (ig/dL also had a significant
increase in plasma MDA compared to children with blood Pb levels <10 (ig/dL.
However, levels of RBC SOD, GSH, and GPx were not different from those in controls
(Jin et al.. 2006).
Evidence of lipid peroxidation was also observed in occupational cohorts moderately
exposed to Pb. In auto repair apprentices in Turkey (mean [SD]: 16.8 [1.2] years of age,
3.8 [1.8] years duration of exposure) with mean blood Pb levels of 7.9 (ig/dL (Ergurhan-
Ilhan et al.. 2008). increases in glutathione peroxidase (GPx) and MDA, as well as
decreases in a-tocopherol and p-carotene were observed compared with controls (mean
[SD] age: 16.3 [1.0] years, mean blood Pb level: 2.6 (ig/dL). Decreases were observed in
SOD and CAT, but the results did not attain statistical significance. Statistically
significant alterations in measures of oxidative stress were also observed in other
occupationally exposed populations. SOD, GSH, and CAT were decreased; while
oxidized GSH (i.e., GSSG) and thiobarbituric acid reactive species (TEARS, expressed
in terms of MDA) were increased in painters in India (mean [SD] duration of exposure:
126.08 [49.53 months], mean blood Pb level: 21.92 (ig/dL, compared to 3.06 (ig/dL in
controls) (Mohammad et al.. 2008). Glutathione-S-transferase, GPx, and SOD were
positively correlated with blood Pb levels (mean: 5.4 (ig/dL, r = 0.34, 0.38, and 0.32,
respectively; p <0.05) in automotive painters in Brazil (Conterato et al.. 2013).
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Numerous cross-sectional, occupational studies have also demonstrated increased lipid
peroxidation in highly-exposed worker populations (blood Pb levels ranging from 29.0 to
74.4 (ig/dL) (Kasperczvk et al.. 2009; Khan et al.. 2008; Quintanar-Escorza et al.. 2007;
Patil et al.. 2006a; Patil et al.. 2006b). There was a correlation between MDA levels and
blood Pb levels, even in the unexposed workers who had lower (i.e., <12 (ig/dL) blood Pb
levels, although the magnitude of correlation in exposed workers was greater (Quintanar-
Escorza et al.. 2007). Increases in C-reactive protein and decreases in RBC SOD, CAT,
and plasma ceruloplasmin were also observed in these workers, further indicating
increased RBC oxidative stress due to higher Pb exposure.
Oral administration of Pb (25 mg/kg) to rats once a week (i.e., bolus gavage) for 4 weeks,
which produced a plasma Pb level of about 6.5 (ig/dL, caused a significant increase in
RBC MDA levels (Lee et al.. 2005). Other indications of Pb-induced oxidative stress
included significant increases in RBC SOD and CAT levels accompanied by significant
decreases in GSH and GPx. Exposure of rats to Pb acetate (750 mg/kg in drinking water
for 11 weeks) resulted in decreased concentrations of plasma vitamin C, vitamin E,
nonprotein thiol, and RBC-GSH, with simultaneous increased activity of SOD and GPx
(Kharoubi et al., 2008b). CAT activity was also slightly elevated in RBCs from the
Pb-exposed rats, but the increase failed to reach statistical significance. Exposure of male
rats to 5,000 ppm Pb nitrate in drinking water (blood Pb not reported) for three weeks
decreased GSH levels compared to that in controls (mean [SE]: 1.91 [0.02] versus 2.44
[0.09] mg/mL, respectively) (Gautam and Flora. 2010). SOD activity was significantly
decreased in rats injected with Pb acetate (15 mg/kg, i.p. for seven days, but not rats
injected with 5 mg/kg) (Berrahal et al., 2007). GPx activity and MDA concentrations
were slightly elevated in the two exposed groups, but differences with the control
(15 mg Na acetate/kg) group failed to reach statistical significance. Effects on indices of
oxidative stress were also observed in in vitro studies: increased MDA and decreased
SOD and CAT in RBCs exposed to 2 (iM Pb (Ciubar et al., 2007), decreased glutathione
reductase (GR) activity in human RBCs incubated with 5-18 (iM Pb (Coban et al.. 2007).
and decreased GSH and increased GSSG and lipid peroxidation in RBCs from healthy
volunteers (with no history of Pb exposure) incubated with 0.4 (iM Pb for 24-120 hours
(Quintanar-Escorza et al.. 2010).
Antioxidant Defense
In addition to the studies listed above that examined lipid peroxidation and oxidative
stress, there have been toxicological studies that indicate that the use of antioxidants and
free radical scavengers is protective against Pb-induced RBC oxidative stress. Rats
treated with 500 ppm Pb acetate in drinking water for 15 or 30 days had a significant
increase in free RBC protoporphyrin and TEARS levels that was related to length of
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exposure and blood Pb levels (Rendon-Ramirez et al.. 2007). Vitamin E administration
after exposure to Pb significantly reduced the rat RBC TEARS levels and increased
ALAD activity, compared to exposure to Pb alone. Co-exposure to vitamin E and Pb
simultaneously and exposure to vitamin E before Pb exposure also prevented Pb-induced
oxidative stress. In vitro studies by Casado et al. (2007) found that Pb-induced hemolysis
using blood from non-occupationally exposed volunteers indicated that RBC membrane
damage was mediated via oxidative stress. The in vitro studies demonstrated a
concentration- and time-dependent formation in lipid peroxide that was inhibited with a
number of antioxidants, including desferrioxamine (Fe chelator), trolox (chain breaking
antioxidant), and mannitol and Na formate (OH scavengers). Results suggested the role
of singlet oxygen in Pb-mediated membrane damage and hemolysis of exposed RBCs. In
rats exposed to 2,000 ppm Pb in drinking water for 5 weeks, MDA levels were
significantly increased, whereas vitamin E concentrations were significantly decreased
(Caylak et al.. 2008). In the case of MDA, co-exposure to Pb and a number of sulfur-
containing antioxidants (e.g., L-methionine, N-acetylcysteine, and L-homocysteine)
reduced concentrations to a level not significantly different from that in controls, but
were significantly smaller than concentrations observed with Pb alone. Exposure to
L-methionine and N-acetylcysteine also reduced Pb-induced depletion of vitamin E.
4.7.2.8 Summary of Effects on RBC Survival and Function
In summary, Pb exposure has been shown to affect multiple hematological outcomes that
are related to RBC survival and function, as demonstrated in both cross-sectional
epidemiologic studies and toxicological studies. Pb exposure has been shown to decrease
RBC survival, either through direct effects on RBC membranes leading to increased
fragility, or through the induction of eryptosis and eventual phagocytosis by
macrophages. Limited evidence that Pb can negatively affect hematopoiesis is also
available. Consistent evidence also exists demonstrating that Pb exposure increases
oxidative stress in exposed adults and children, as well as in laboratory animals. The
epidemiologic studies demonstrating these effects are cross-sectional in design, therefore
there is some uncertainty regarding the direction of effects and the magnitude, timing,
frequency, and duration of Pb exposure that contributed to the observed observations.
Also, the majority of epidemiologic studies did not account for potential confounding,
although the effects observed in these studies are consistent with effects from studies that
did consider potential confounding. The coherence with effects observed in animal
toxicology studies supports the conclusion that Pb exposure affects both the survival and
function of RBCs.
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4.7.3
Red Blood Cell Heme Metabolism
Pb exposure has been found to inhibit several enzymes involved in heme synthesis,
namely ALAD (a cytoplasmic enzyme catalyzing the second, rate-limiting, step of the
heme biosynthesis pathway), coproporphyrinogen oxidase (catalyses the sixth step in
heme biosynthesis converting coproporphyrinogen III into protoporphyrinogen IX), and
ferrochelatase (catalyses the terminal [eighth] step in heme synthesis converting
protoporphyrin IX into heme) (Figure 4-35). The observations of decreased Hb
(measured as total Hb, MCH, or MCHC) in occupationally-exposed adults (Ukaejiofo et
al.. 2009; Khan et al.. 2008; Patil et al.. 2006b: Karita et al.. 2005) and Pb-exposed
experimental animal models (Sharma et al.. 2010b: Baranowska-Bosiacka et al.. 2009;
Simsek et al.. 2009; Marques et al.. 2006; Lee et al.. 2005) and associations with blood
Pb levels in children (Oueirolo et al.. 2010; Shah etal. 2010; Olivero-Verbel et al.. 2007;
Riddell et al.. 2007) are supporting lines of evidence for decreased heme synthesis due to
Pb exposure.
Cytosol
2x6-aminolevulpnic acid
1
ALA dehycfratase
(porpfiobitinagei
synthase
Porphobilinogen 1PBG)
3 | | P8G deomtnase
hydroxymethylbila
Uroporphyrin ogen
fit syinht-'tase
^> Uroporphyrinogen II
Copiopotphyrinogcn III
Uroporphyrinogen
decarboxylase
Note: Steps in the pathway potentially affected by Pb are indicated with curved arrows pointing to the affected enzyme, and the
directions of effects are represented by f and \, arrows.
Figure 4-35 Schematic representation of the enzymatic steps involved in the
heme synthetic pathway.
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4.7.3.1 Red Blood Cell 5-Aminolevulinic Acid Dehydratase
Decreases in RBC 5-aminolevulinic acid dehydratase (ALAD) levels are strongly
associated with Pb exposure in humans; to such an extent that RBC ALAD activity is
used as a biomarker to assess Pb toxicity. Several epidemiologic studies published since
the 2006 Pb AQCD evaluated the relationship between Pb exposure, blood Pb levels and
ALAD activity in adults and children (see below). These studies were cross-sectional in
nature. This limits their utility in assessing the direction of effects and the magnitude,
timing, frequency, and duration of Pb exposure necessary to contribute to the observed
associations. In studies that considered potential confounders, those confounding
variables are listed in the text. However, potential confounding was not accounted for in
the majority of these studies.
Wang et al. (2010g) found that, after controlling for sex, age, alcohol consumption and
smoking (adults only), there was a blood Pb concentration-dependent decrease in ALAD
activity in both children (4-13 years old) and adults (16-77 years old) (mean blood Pb
levels: 7.1 and 6.4 (ig/dL, respectively) in rural southwest China. Further, Wang et al.
(2010g) observed that the relationship between blood Pb level and ALAD activity was
nonlinear and exponential, with larger decreases in ALAD activity occurring with blood
Pb levels >10 (ig/dL. No correlation was observed between urinary ALA levels and blood
Pb levels. Ahamed et al. (2006) studied male urban adolescents in India. The 39
adolescents were separated into two groups according to their blood Pb levels (Group 1:
<10 (ig/dL [mean 7.4 (ig/dL], Group 2: >10 (ig/dL [mean 13.27 (ig/dL]). Although
Groups 1 and 2 were similar in age (mean [SD]: 16.59 [0.91] versus 16.76 [0.90] years,
respectively), height, weight, and BMI (therefore not considered potential confounders),
Group 2 (with the higher blood Pb levels) had lower ALAD activity than did Group 1
(p <0.001). When all 39 adolescents were examined together, an inverse relationship was
found between blood Pb and ALAD activity. Similar decreases in ALAD activity were
observed in other populations of children from India (aged 4-12 and 1-7 years) with
elevated blood Pb levels (mean [SD]: 11.39 [1.39] and 21.86 [7.58] (ig/dL respectively)
compared to the two age ranges of the children with lower blood Pb levels (mean [SD]:
3.93 [0.61] and 6.89 [2.44] (ig/dL respectively) (Ahamed et al.. 2007: Ahamed et al..
2005). While Ahamed et al. (2005) did not address potential confounding, Ahamed et al.
(2007) observed decreases in ALAD activity after controlling for age, sex, and area of
residence. Decreases in ALAD activity were also observed in children 3-6 years of age
with Pb blood levels >10 (ig/dL, compared to children <10 (ig/dL (mean blood Pb
concentration for groups not reported) in northeastern China (Jin et al.. 2006).
As was seen with epidemiologic studies investigating Pb-associated deficits in
hematological parameters, most occupational studies investigating ALAD levels may not
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be generalizable to the population as a whole; however, they are useful in demonstrating
consistent and negative effects of Pb exposure on the activity of this enzyme (Quintanar-
Escorzaetal.. 2007; Patil et al. 2006a: Patil et al.. 2006b: Ademuyiwa et al.. 2005b).
Occupationally-exposed adults (mean blood Pb levels: 47.4 (ig/dL) had levels of
inhibition of ALAD that were as great as 90% relative to control (blood Pb levels: 9.9
[ig/dL)(Quintanar-Escorza et al.. 2007). There were few studies that investigated
Pb-associated decrements in ALAD levels among moderately-exposed workers. Painters
in India with a mean blood Pb level of 21.92 (ig/dL (mean [SD] duration of exposure:
126.08 [49.53] months) had lower ALAD levels (p <0.01) compared to controls whose
mean blood Pb level was 3.06 (ig/dL (Mohammad et al.. 2008). Stoleski et al. (2008)
observed that workers in a Pb smelter in Macedonia (mean [SD]: 16.4 [8.5] (ig/dL blood
Pb; 18.8 [7.5] years employment) had lower ALAD activity (p <0.001) and higher ALA
levels (p <0.0005) compared to workers with no history of exposure to Pb (mean [SD]
blood Pb: 7.0 [5.4] (ig/dL). In automotive painters exposed to Pb in Brazil (mean [SD]:
5.4 [0.4] (ig/dL blood Pb level; 133.9 [14.5] months duration of exposure), the ALAD
reactivation index was increased over that in controls, although ALAD activity did not
differ between groups (Conterato et al.. 2013). However, ALAD activity was negatively
correlated with blood Pb levels (r = -0.59, p <0.05) but not blood Cd levels, whereas
ALAD reactivation index was positively correlated with blood levels of both metals (Pb:
r = 0.84, p <0.05; Cd: r = 0.27, p <0.05). In a benchmark dose (BMD)-based analysis
(benchmark response = 5% using the hybrid approach and a 5% adversity cut-off value),
Murata et al. (2009) calculated the BMD and 95% lower confidence limit of the BMD
(BMDL) for decreased ALAD activity in RBCs of exposed Pb workers. The calculated
BMD (2.7 (ig/dL) and BMDL (2.3 (ig/dL) values for ALAD, respectively, were
substantially lower than the BMDs (28.7-44.2 (ig/dL) and BMDLs (19.4-29.6 (ig/dL) for
decreased Hb, Hct, and RBC count in similarly exposed workers, indicating decreases in
ALAD activity can occur at blood Pb levels that do not decrease RBC survival.
Decreased ALAD activity in response to Pb exposure has also been observed in
toxicological studies. Rats administered 500 ppm Pb acetate in drinking water for 15 or
30 days had decreased blood ALAD activity, which was related to duration of exposure
and blood Pb levels (Rendon-Ramirez et al.. 2007). Oral administration of Pb (25 mg/kg)
to rats once a week for 4 weeks achieved a blood Pb level of 6.5 (ig/dL, which was
associated with statistically significant decreases (approximately 50% lower than control
levels) in RBC ALAD activity (Lee et al.. 2005). Exposure of male Wistar rats to
5,000 ppm Pb acetate via drinking water for three weeks significantly decreased ALAD
activity by 72% (mean [SD]: 7.35 [0.35] versus controls: 26.14 [2.19] nmol/min/mL
RBCs [nanomoles of porphobilinogen (PEG) formed per minute, per 1 mL blood])
(Gautam and Flora. 2010).
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4.7.3.2 Other Heme Metabolism Enzymes
The 2006 Pb AQCD (U.S. EPA. 2006b) indicated that Pb affects RBC PEG synthase
(Simons. 1995; Farant and Wigfield. 1990. 1987). PEG deaminase (Tomokuni and
Ichiba. 1990). and TF endocytosis and Fe transport across membranes (Qian and Morgan.
1990). all of which are directly or indirectly involved in heme synthesis. Although there
are no recent studies that examine the effect Pb has on the activities of other heme
metabolism enzymes, a number of studies investigated associations of blood Pb level
with concentrations of various intermediate products in the heme biosynthetic pathway.
Pb intoxication has been shown to inhibit the function of ferrochelatase, the enzyme that
catalyzes the last (eighth) step in the heme biosynthetic pathway. Under normal
conditions, ferrochelatase incorporates ferrous iron (Fe2+) into protoporphyrin IX,
converting it into a heme molecule (Figure 4-35). However, Pb has been shown to inhibit
this insertion of Fe2+ into the protoporphyrin ring and instead, Zn is inserted into the ring
creating ZPP. A number of recent studies have shown that blood Pb level is significantly
associated with increased RBC ZPP levels in adults occupationally exposed to high levels
of Pb (blood Pb levels: 27-54 jig/dL) (Patil et al.. 2006b: Ademuviwa et al.. 2005b).
workers exposed to moderate levels of Pb (blood Pb level = 21.92 (ig/dL) (Mohammad et
al.. 2008). children aged 1-21 years (blood Pb levels: 18-23 (ig/dL) (Counter et al.. 2009.
2008; Counter et al.. 2007). and animals exposed to 500 ppm Pb via drinking water for 15
or 30 days (blood Pb levels: 24.7 or 31.8, respectively) (Rendon-Ramirez et al.. 2007).
Interestingly, Wang et al. (2010g) found that in children and adults living in a rural area
of Southwest China, ZPP levels were negatively correlated with blood Pb at blood Pb
levels <10 (ig/dL and were only positively correlated with blood Pb at higher blood Pb
concentrations (i.e., >10 (ig/dL). The authors suggested that this may be representative of
ALAD activities at low blood Pb levels, which contributes to lower ZPP levels.
Scinicariello et al. (2007) performed a meta-analysis and observed that Pb-exposed
individuals who carried the ALAD2 allele had slightly lower concentrations of blood ZPP
levels compared to carriers of the ALAD1 allele (overall pooled standardized mean
estimate: -0.09 [ZPP value for ALAD 1-2/2-2 carriers vs ALAD1-1 carriers]; 95% CI:
-0.22, 0.03, p = 0.13).
4.7.3.3 Hematological Effects
In summary, Pb exposure has been shown in both cross-sectional epidemiologic studies
and toxicological studies to alter heme synthesis. Pb exposure has been shown to inhibit
the activities of two major enzymes in the heme biosynthetic pathway: ALAD and
ferrochelatase. Evidence for the inhibition of ALAD comes from direct measurements of
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its activity in exposed human populations, whereas evidence for inhibition of
ferrochelatase comes from the observation of increased ZPP levels in association with
occupational exposure or blood Pb levels. Animal toxicology and ecotoxicology studies
provide evidence of coherent effects in animals. The epidemiologic studies demonstrating
these effects are cross-sectional in design, therefore there is some uncertainty regarding
the direction of effects and the magnitude, timing, frequency, and duration of Pb
exposure that contributed to the observed observations. Also, the majority of
epidemiologic studies did not account for potential confounding, although the effects
observed in these studies are consistent with effects from studies that did account for
confounding. The coherency of effects observed in animal toxicology and ecotoxicology
studies support the conclusion that Pb exposure alters the synthesis of heme in RBCs.
4.7.4 Summary and Causal Determination
Recent toxicological and epidemiologic evidence substantiates evidence presented in the
2006 Pb AQCD that exposure to Pb affects hematological endpoints, and supports a
causal relationship between Pb exposure and decreased RBC survival and function and
altered heme synthesis. Outcomes related to decreased RBC survival and function
included alterations in multiple hematological parameters (e.g., Hb, Hct, PCV, MCV,
MCH), oxidative stress (altered antioxidant enzyme activities [SOD, CAT, GPx],
decreased cellular GSH, and increased lipid peroxidation), increased cytotoxicity in RBC
precursor cells, and mode of action endpoints such as decreased intracellular calcium
concentrations, decreased ATPase activity, and increased phosphatidylserine expression.
Outcomes related to altered heme synthesis included decreased activities of ALAD and
ferrochelatase, and decreased levels of Hb. The sections that follow describe the
evaluation of evidence for decreased red blood cell (RBC) survival and function and
heme synthesis, with respect to causal relationships with Pb exposure using the
framework described in Table II of the Preamble. The application of the key evidence to
the causal framework is summarized in Table 4-35.
4.7.4.1 Evidence for Decreased RBC Survival and Function
The 2006 Pb AQCD (U.S. EPA. 2006b) reported that Pb exposure is associated with
multiple measures of decreased RBC survival and function. Epidemiologic evidence
included the observation of a 10% probability of anemia with blood Pb levels of
approximately 20 (ig/dL at age 1 year, and perturbed hematopoiesis in children and adults
at blood Pb levels between 10-40 (ig/dL. Oxidative stress was also identified by the
2006 Pb AQCD as a potential mode of action for Pb-induced effects in RBCs. A causal
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relationship between Pb exposure and decreased RBC survival and function is strongly
supported by the available, recent toxicological and epidemiologic data. Among the
strongest evidence for Pb-induced decreases in RBC survival and function is the
consistent observation of alterations in hematological parameters (e.g., Hb, Hct, PCV,
MCV, MCH), oxidative stress (altered antioxidant enzyme activities [SOD, CAT, GPx],
decreased cellular GSH, and increased lipid peroxidation), and increased cytotoxicity in
RBC precursor cells in rodents exposed to various forms of Pb via drinking water (Jang
et al.. 2011; Molina et al. 2011; Gautam and Flora. 2010; Baranowska-Bosiacka et al..
2009; Simsek et al.. 2009; Alghazal et al.. 2008b: Kharoubi et al.. 2008b: Marques et al..
2006). Some of these effects have been observed in toxicological studies reporting blood
Pb levels of 2-7 (ig/dL, and therefore occur at blood Pb levels that are relevant to this
ISA. These effects at relevant blood Pb levels were found primarily in adult animals with
Pb exposure durations of 4 weeks to 9 months. Although not as representative of
potential human exposure pathways as exposure via drinking water, numerous
toxicological studies utilizing oral gavage have also observed effects on hematological
parameters, oxidative stress, and hematopoiesis (Sharma et al.. 2010b; Celiket al.. 2005;
Lee et al.. 2005). The animal toxicological evidence for decreased RBC survival and
function is particularly important to the weight of evidence as it establishes clear
temporality of exposure to Pb and induction of effects on red blood cells.
Associations between increased blood Pb levels and decreased RBC survival and
function, are also evident in diverse populations of human adults and children. Cross-
sectional studies in children measuring concurrent blood Pb levels are consistent
regarding effects on hematological parameters (Oueirolo et al.. 2010; Shah et al.. 2010;
Ahamed et al.. 2007; Huo et al.. 2007; Olivero-Verbel et al.. 2007; Riddell et al.. 2007;
Turgut et al.. 2007; Ahamed et al.. 2006; Jin et al.. 2006; Rondo et al.. 2006).
Associations between altered indices of RBC oxidative stress and blood Pb levels were
also seen in adolescents and children (Ahamed et al.. 2008; Ahamed et al.. 2006; Jin et
al.. 2006). The blood Pb levels observed in cross-sectional studies of children tended to
be lower than those observed in adult populations (see below), with the majority of
studies in children (ages 5 months to 5 years old) reporting mean concurrent blood Pb
levels <15 (ig/dL (range: 6.9-21.86 (ig/dL). The difference in blood Pb levels may reflect
the comparatively shorter duration and lower magnitude of Pb exposure experienced by
children compared to adults.
For adult populations, the largest body of evidence consists of Pb-exposed workers in
which measures of RBC survival (e.g., Hb, Hct, PCV, MCV, MCH) are altered when
compared with unexposed control populations in cross-sectional studies (Conterato et al..
2013; Cabaravdic et al.. 2010; Ukaeiiofo et al.. 2009; Khan et al.. 2008; Patil et al..
2006a; Patil et al.. 2006b; Karitaetal.. 2005). Only one non-occupational study was
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found investigating the association of Pb with hematological parameters; in pregnant
women, concurrent blood Pb levels were found to be negatively correlated with Hb
concentrations. Cross-sectional studies have also observed consistent increases in lipid
peroxidation in occupationally-exposed adult populations (Ergurhan-Ilhan et al.. 2008;
Khan et al.. 2008; Mohammad et al., 2008; Quintanar-Escorza et al., 2007; Patil et al..
2006a: Patil et al.. 2006b). and have observed changes in oxidative stress parameters,
including lowered activities of antioxidant enzymes such as SOD, GR, and CAT. Recent
evidence of disrupted hematopoiesis, including the observation of decreased serum EPO
in occupationally-exposed adults with a mean blood Pb level of 6.4 (ig/dL (Sakataetal..
2007). was consistent with previous findings of decreased EPO in exposed adults
reported in the 2006 Pb AQCD. Although the mean blood Pb level in most
occupationally exposed populations was >20 (ig/dL, multiple studies observed negative
effects in occupationally-exposed populations with mean blood Pb levels <10 (ig/dL,
including significant correlations PCV (7 (ig/dL), significant correlations between RBC
distribution width and MCHC (5.4 ng/dL), and decreased EPO (6.4 (ig/dL). Any
differences in the effects on specific hematological and oxidative stress parameters
between adult populations and children may reflect differences in exposure durations or
patterns of exposure, although there is greater uncertainty regarding the timing and
duration of exposure associated with these effects in adults.
The evidence for Pb-associated decrements in RBC function and survival in adults and
children comes from cross-sectional studies measuring concurrent blood Pb levels, and
thus, the temporality of effects and the timing and duration of exposure associated with
altered RBC survival and function in RBCs is unclear. This uncertainty is greatest in
adults and older children as concurrent blood Pb levels also reflect higher past Pb
exposures. Additional limitations of the epidemiologic database include the general lack
of controlling for potential confounders or other possible co-exposures to contaminants
that can affect the hematological system. Although most studies did not control for
potential confounders, a few studies investigating effects in children did adjust for
potential confounders such as age, sex, area of residence, breastfeeding, mouthing
behavior, family structure, anemia, dairy product consumption, maternal age, education,
employment, marital status, and SES-related variables; and still observed negative effects
on RBC survival and function. However, no studies controlled for nutritional status,
including Fe intake. Further, while the epidemiologic database may be limited for the
above reasons, the findings in these studies demonstrated coherence with findings from
multiple toxicological studies that either reported blood Pb levels that are relevant to this
ISA, i.e., <10 (ig/dL (drinking water and gavage studies) or utilized a relevant route of
exposure (drinking water), and reported clear evidence for decreased RBC survival and
function.
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The causal relationship between Pb exposure and decreased RBC survival and function is
further supported by epidemiologic and toxicological evidence characterizing mode of
action and biological plausibility. Pb was shown to reduce Ca2+ATPase and
Ca2+/Mg2+ATPase activities in RBC membranes, which leads to an increase in RBC
[Ca2+]l5 increased membrane fragility, and abnormal morphological changes in studies of
occupationally exposed adults (Quintanar-Escorza et al.. 2007) and in in vitro studies
(Quintanar-Escorza et al., 2010; Ciubar et al., 2007). Heul et al. (2008) observed a
reduction in plasma membrane Ca2+ATPase pump activity in newborn children's RBC
membranes in association with cord blood Pb level in a group with a mean blood Pb of
3.54 (ig/dL. Pb exposure has also been observed to increase PS expression on RBC
membranes, leading to cell shrinkage, erythropoiesis, and destruction of the RBCs by
macrophages (Jang etal.. 2011; Ciubar etal.. 2007; Shin et al.. 2007).
Experimental animal studies demonstrate that Pb exposures via drinking water and
gavage, resulting in blood Pb levels relevant to this ISA, alter several hematological
parameters, increase measures of oxidative stress, and increase cytotoxicity in RBC
precursor cells. These effects were found primarily in adult animals with Pb exposure
durations of 4 weeks to 9 months. Support for these findings is provided by biologically
plausible modes of action, including decreased intracellular calcium concentrations,
decreased ATPase activity, and increased phosphatidylserine expression. Epidemiologic
studies demonstrate evidence in both adults and children that concurrent blood Pb levels
are associated with altered hematological endpoints and increased measures of oxidative
stress, and altered hematopoiesis. However, the majority of these studies are limited by
the lack of rigorous methodology and consideration for potential confounding. Although
there are limitations in the epidemiologic evidence, some studies in children did control
for or considered potential confounding and effects in adults and children are coherent
with effects observed in exposed animals. Because epidemiologic evidence is limited to
associations with concurrent blood Pb levels, there is uncertainty regarding the timing,
duration, magnitude, and frequency of Pb exposure associated with decreased RBC
survival and function. Collectively, the strong evidence from toxicological studies that is
supported by findings from mode of action and epidemiologic studies is sufficient to
conclude that there is a causal relationship between Pb exposures and decreased RBC
survival and function.
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4.7.4.2 Evidence for Altered Heme Synthesis
The 2006 Pb AQCD (U.S. EPA. 2006b) reported that Pb exposure affects heme synthesis
in humans and animals through the inhibition of multiple key enzymes in the heme
biosynthetic pathway, including ALAD and ferrochelatase. A causal relationship between
Pb exposure and altered heme synthesis is strongly supported by the available
toxicological, ecotoxicological, and epidemiologic data (Table 4-35). The greatest weight
of evidence for Pb-induced alterations in heme synthesis lies primarily in the
toxicological and ecotoxicological literature. A small, but coherent, body of recent
toxicological evidence demonstrates decreased ALAD activity (Gautam and Flora. 2010;
Lee et al.. 2005) and ferrochelatase (Rendon-Ramirez et al.. 2007) in adult rats exposed
to Pb via drinking water and oral gavage for 3-4 weeks. Lee et al. (2005) observed effects
on ALAD activity in Wistar rats at mean blood Pb levels of 6.5 (ig/dL after Pb
administration by oral gavage once per week for 4 weeks. Evidence from previous studies
cited in the 2006 Pb AQCD consistently observed Pb-induced ALAD inhibition in
multiple species, including birds, primates, and humans, further supporting the causal
association between Pb exposure and altered heme synthesis.
Similar to the earlier and more recent toxicological studies that demonstrate an
association between Pb exposure and hematological effects in humans and laboratory
animals, the ecological literature has consistently reported on hematological responses in
Pb-exposed aquatic and terrestrial invertebrates and vertebrates (Sections 6.3.12.5.
6.4.12.5. and 6.4.21.5). The most consistently observed effect in metal impacted
environments is decreased RBC ALAD activity. This effect has been observed across a
wide range of taxa, including bivalves, fish, amphibians, birds, and mammals. More
limited evidence exists regarding deleterious effects of Pb exposure on serum enzyme
levels and white blood cell counts in birds and mammals.
Consistent associations between increased Pb blood levels and decreased activity of
multiple enzymes involved in the heme synthetic pathway have also been observed in
diverse populations of adults and children. The strongest evidence for altered heme
synthesis in adults and children come from cross-sectional epidemiologic studies
measuring concurrent blood Pb and reporting associations with decreases in RBC ALAD
levels and activity (Conterato et al.. 2013; Wang et al.. 2010g: Mohammad et al.. 2008;
Ahamed et al.. 2007; Quintanar-Escorza et al.. 2007; Ahamed et al.. 2006; Patil et al..
2006a; Patil et al.. 2006b; Ademuviwa et al.. 2005b; Ahamed et al.. 2005). In addition to
ALAD inhibition, recent studies have also shown that blood Pb levels or Pb exposure
inhibits the activity of ferrochelatase, leading to increased RBC ZPP levels in children
and occupationally-exposed adults (Counter et al.. 2009. 2008; Mohammad et al.. 2008;
Counter etal.. 2007; Patil et al.. 2006b; Ademuviwa et al.. 2005b). Although the mean
4-581
-------
blood Pb levels in most of the studies investigating these effects in adults and children
were >20 (ig/dL, two studies did observe negative effects in populations with mean blood
Pb levels <10 (ig/dL: increased ALAD reactivation index in exposed painters
(5.4 (ig/dL), and statistically significant, positive associations between ALAD and blood
Pb level in children and the elderly (7.1 and 6.4 (ig/dL, respectively).
The cross-sectional nature of the above epidemiologic studies in adults and children, and
the measurement of concurrent blood Pb, introduces some uncertainty regarding the
temporality of effects and the timing and duration of exposure associations with altered
heme synthesis. Although most studies did not control for potential confounders, a few
studies investigating effects in children, and one study investigating effects in adults, did
adjust for confounders such as age, sex, urban/rural residence, height, weight, BMI,
smoking status, and alcohol use, and still observed negative effects on heme synthesis.
However, no studies controlled for nutritional status, includingFe intake. Further, while
the epidemiologic database may be limited for the above reasons, the findings in these
studies demonstrated coherence with findings from multiple toxicological and
ecotoxicological studies.
The causal relationship between Pb exposure and altered heme synthesis is further
supported by cross-sectional studies observing decreased Fib (measured as total Fib,
MCH, or MCHC) in occupationally-exposed adults (Ukaejiofo et al.. 2009; Khan et al..
2008; Patil et al., 2006b; Karita et al., 2005) and in children in association with blood Pb
levels (Oueiroloetal.. 2010: Shahet al.. 2010: Olivero-Verbel et al.. 2007: Riddell et al..
2007). Several recent toxicological studies also observed decreased Fib levels in
laboratory animals exposed to Pb (Sharma et al.. 2010b: Baranowska-Bosiacka et al..
2009: Simsek et al.. 2009: Marques et al.. 2006: Lee et al.. 2005). Decreased Hb levels
are a direct indicator of decreased heme synthesis due to Pb exposure.
In summary, altered heme synthesis is demonstrated by a small, but coherent, body of
studies in adult animals reporting that Pb exposures via drinking water and gavage
(resulting in blood Pb levels relevant to this ISA) for 15 days to 9 months decreased
ALAD and ferrochelatase activities. Supporting this toxicological evidence is a larger
body of ecotoxicological studies that demonstrate decreased ALAD activity across a wide
range of taxa exposed to Pb. Epidemiologic studies demonstrate evidence in both adults
and children that concurrent blood Pb levels are associated with decreased ALAD and
ferrochelatase activities. However, the majority of these studies are limited by the lack of
rigorous methodology and consideration for potential confounding. Although there are
limitations in the epidemiologic evidence, some studies in children did control for or
considered potential confounding and effects in adults and children are coherent with
effects observed in exposed animals. Because epidemiologic evidence is limited to
4-582
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associations with concurrent blood Pb levels, there is uncertainty regarding the timing,
duration, magnitude, and frequency of Pb exposure associated with decreased RBC
survival and function. Evidence for altered heme synthesis is also provided by a large
body of toxicological and epidemiologic studies that report decreased Hb concentrations
in association Pb exposure or blood Pb levels. Collectively, the strong evidence from
toxicological and ecotoxicological studies, which are supported by findings from
epidemiologic studies, are sufficient to conclude that there is a causal relationship
between Pb exposures and altered heme synthesis.
4-583
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Table 4-35 Summary of evidence supporting RBC survival and heme synthesis causal determinations.
Attribute in Causal
Framework3
Key Evidence
Recent References
Pb Exposure or Blood Pb
Levels Associated with
Effects0
Decreased RBC Survival and Function: Causal
Consistent evidence
in animals with
relevant exposures
to rule out chance,
bias, and
confounding with
reasonable
confidence
Large body of studies with consistent findings for
decreased RBC survival and function (decreased
Hb, Hct, PVC, increased eryptosis, decreased
hematopoiesis, increased oxidative stress) in
rodents with relevant concentrations of Pb and
routes of exposure
Baranowska-Bosiacka et al. (2009),
Lee et al. (2005),
Sharma et al. (201 Ob).
Simsek et al. (2009).
Marques et al. (2006).
Molina et al. (2011).
Jang et al (2011),
Celiketal. (2005),
Alghazal et al. (2008b),
Kharoubi et al. (2008b),
Gautam and Flora (2010)
See Sections 4.7.2.2,
4.7.2.3, and 4.7.2.7
Rodents:
Blood Pb level:
1.7-7.1 ug/dL
Exposures:
Drinking water 50-2,000 ppm,
21-270 days as adults
Oral gavage
25-500 mg/kg,
28-70 days
Support from
multiple, consistent
epidemiologic
studies with relevant
blood Pb levels
Some consider
potential
confounding.
Cross-sectional studies that considered potential
confounding factors found blood Pb-associated
decreases in Hb, increases in anemia prevalence,
increased oxidative stress in children ages 6 mo-5
yr
Association with Hb found in children with
concurrent blood Pb levels with consideration for
potential confounding by age, sex, mouthing
behavior, anemia, dairy product consumption,
maternal age, education, employment, marital
status, family structure, SES-related variables
Other studies of Hb, oxidative stress adjusted for
factors such as age, sex, birthweight,
breastfeeding history, urban/rural residence
Riddell et al. (2007),
Queirolo et al. (2010).
Ahamed et al. (2007).
Ahamed et al. (2008)
See Sections 4.7.2.2,
and 4.7.2.7
Queirolo et al. (2010)
See Section 4.7.2.2
Riddell et al. (2007).
Ahamed et al. (2008).
Ahamed et al. (2007)
See Sections 4.7.2.2, and 4.7.2.7
Children: majority of concurrent
blood Pb levels <15 ug/dL
4-584
-------
Table 4-35 (Continued): Summary of evidence supporting RBC survival and heme synthesis causal determinations.
Attribute in Causal
Framework3
Key Evidence
Recent References
Pb Exposure or Blood Pb
Levels Associated with
Effects0
Support from
multiple, consistent
epidemiologic
studies with relevant
blood Pb levels
(continued)
Consistent evidence in large body of cross-
sectional studies without consideration for
potential confounding in occupationally-exposed
adults and in children of associations of blood Pb
levels with decreases RBC survival, interferes with
hematopoiesis, and increases oxidative stress
Karita et al. (2005). Khan et al. (2008). Patil et
al. (2006a). Patil et al. (2006b). Ukaejiofo et al.
(2009), Conterato et al. (2013), Cabaravdic et
al. (2010), Ergurhan-llhan et al. (2008),
Mohammad et al. (2008), Quintanar-Escorza et
al. (2007), Sakata et al. (2007), Riddell et al.
(2007). Queirolo et al. (2010). Olivero-Verbel et
al. (2007). Ahamed et al. (2006). Ahamed et al.
(2007). Ahamed et al. (2008). Turgut et al.
(2007). Huo et al. (2007). Shah et al. (2010).
Rondo et al. (2006), Jin et al. (2006)
See Sections 4.7.2.2.4.7.2.3. and 4.7.2.7
Adults (occupational exposures):
majority of blood Pb levels
>20 ug/dL, some studies
observed effects in the range of
5-7 ug/dl_
Support from
evidence that clearly
describes mode of
action
Altered RBC
membrane ion
transport:
Phosphatidylserine
(PS) expression:
Evidence of increased [Ca2+]i and decreased
Ca2+/Mg2+ATPase activity in the RBCs of exposed
workers. [Ca2+]i levels highly correlated with blood
Pb even among unexposed controls.
[Ca2+]i levels increased in RBCs from healthy
volunteers when exposed in vitro to Pb
[Ca2+]i associated with increased RBC fragility and
alterations in RBC morphology
Consistent evidence from in vivo and in vitro
studies that Pb exposure increases PS expression
on RBC membranes via modulation of [Ca2+]i
concentrations. Increased PS expression leads to
eryptosis and phagocytosis by macrophages
See Section 4.7.2.2
See Section 4.7.2.2
4-585
-------
Table 4-35 (Continued): Summary of evidence supporting RBC survival and heme synthesis causal determinations.
Attribute in Causal
Framework3
Key Evidence
Recent References
Pb Exposure or Blood Pb
Levels Associated with
Effects0
Altered Heme Synthesis: Causal
Consistent evidence
in animals with
relevant exposures
to rule out chance,
bias, and
confounding with
reasonable
confidence
A small, but coherent toxicology database
indicates decreased heme synthesis in rodents
with relevant Pb concentrations and routes of
exposure
Consistent eco-toxicological evidence for
Pb-induced decreased ALAD activity observed
across many taxa (bivalves, fish, amphibians,
birds, and mammals) in multiple studies
Rendon-Ramirez et al. (2007), Lee et al.
(2005),
Gautam and Flora (2010)
See Section 4.7.3.1
Birds:
Berglund et al. (2010),
Gomez-Ramirez et al. (2011). Hansen et al.
(2011 a).
Martinez-Haro et al. (2011)
See Section 6.3.4.3
Freshwater Invertebrates:
Aisemberg et al. (2005)
See Section 6.4.5.2
Fish:
Schmitt et al. (2005).
Schmitt et al. (2007b).
Heier et al. (2009).
See Section 6.4.5.3
Bivalves:
Kalman et al. (2008).
Company et al. (2011).
See Section 6.4.15.2
Blood Pb levels 6.5 ug/dL
Exposures: 500-5,000 ppm
Drinking water, 15-30 days as
adults
Birds:
6->100ug/dL
Freshwater Invertebrates
(48-h exposure in aquaria)
0.2-300 ug/g wet tissue
Fish:
6-14 ug/g
(gill or liver concentrations)
Bivalves:
0.38-3.50 ug/g dry weight
4-586
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Table 4-35 (Continued): Summary of evidence supporting RBC survival and heme synthesis causal determinations.
Attribute in Causal
Framework3
Key Evidence
Recent References
Pb Exposure or Blood Pb
Levels Associated with
Effects0
Support from
multiple, consistent
epidemiologic
studies with relevant
blood Pb levels
Cross-sectional studies that considered potential
confounding by age, sex, urban/rural residence,
height, weight, BMI found associations with lower
ALAD and ferrochelatase activities in children.
Concurrent blood Pb level associated with lower
ALAD and higher ZPP in adults with consideration
for potential confounding by age, sex, smoking
status, and alcohol use.
Associations found in several studies, mostly in
occupationally-exposed adults, that did not
consider potential confounding
Ahamed et al. (2006).
Ahamed et al. (2007)
See Section 4.7.3.1. Section 4.7.4.1.
and Section 4.7.4.2
Wang et al. (2010q)
See Sections 4.7.3.1, and 4.7.3.2
Children:
Ahamed et al. (2005)
Occupational:
Ademuyiwa et al. (2005b), Mohammad et al.
(2008),
Patil et al. (2006a). (2006b).
Quintanar-Escorza et al. (2007). Conterato et
al. (2013)
See Sections 4.7.3.1,
and 4.7.3.2
Adults (occupational exposure)
and children: Majority of
concurrent blood Pb levels
>20 ug/dl_,
Two studies observed
associations in adults and
children in the range of
concurrent blood Pb levels 6-
7 ug/dl_.
Support from
evidence for
decreases in Hb, a
direct marker of
decreased heme
synthesis
Consistent evidence in animals with relevant Pb
exposures for decreases in Hb.
Association found in children between decreased
Hb and concurrent blood Pb levels with
consideration for potential confounding by age,
sex, mouthing behavior, anemia, dairy product
consumption, maternal age, education,
employment, marital status, family structure, SES-
related variables
Other studies in children had limited or no
consideration for potential confounding.
Animals:
Baranowska-Bosiacka et al. (2009),
Lee et al. (2005),
Marques et al. (2006),
Sharma et al. (201 Ob),
Simsek et al. (2009)
See Section 4.7.2.2
Queirolo et al. (2010)
See Section 4.7.2.2
Shah et al. (2010),
Olivero-Verbel et al. (2007), Riddell et al.
(2007)
See Section 4.7.2.2
Adult animals: Blood Pb levels
1.7-7.1 ug/dLafter15day-9
month Pb exposure
Children: Majority of concurrent
blood Pb<15ug/dL
4-587
-------
Table 4-35 (Continued): Summary of evidence supporting RBC survival and heme synthesis causal determinations.
Attribute in Causal
Framework3
Key Evidence
Recent References
Pb Exposure or Blood Pb
Levels Associated with
Effects0
Support from
evidence for
decreases in Hb, a
direct marker of
decreased heme
synthesis
(continued)
Associations found in adults and, as well as
coherent findings in animal toxicological studies,
for decreased Hb.
Adults: Karita et al. (2005).
Khan et al. (2008).
Patil et al. (2006b).
Ukaejiofo et al. (2009)
See Section 4.7.2.2
Adults (occupational exposure):
Majority of blood Pb >20 ug/dL
Support from
evidence that
describes mode of
action
Altered Ion Status: Evidence that Pb competitively
inhibits the binding of Zn ions necessary for ALAD
activity. Pb also inhibits the incorporation of Fe
into protoporphyrin IX by ferrochelatase, resulting
in Zn-protoporphyrin production
See Sections 4.2.2.3. 4.2.2.4,
and 4.2.4.1
Described in detail in Table II of the Preamble.
""Describes the key evidence and references contributing most heavily to causal determination and where applicable to uncertainties or inconsistencies. References to earlier sections
indicate where the full body of the evidence is described.
°Describes the blood Pb levels in humans with which the evidence is substantiated and blood Pb levels in animals most relevant to this ISA.
4-588
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4.8 Reproductive and Developmental Effects
The effect of Pb on reproductive and developmental outcomes has been of interest for
years, starting in cohorts of occupationally-exposed individuals. More recently,
researchers have begun to focus on reproductive and developmental effects in
populations without occupational exposures, with more environmentally-relevant levels
of Pb exposure. The toxicological and epidemiologic literature on reproductive effects of
Pb includes research on female and male reproductive function such as hormone levels,
fertility, spontaneous abortions, effects on sperm, estrus, and effects on reproductive
organs. Evaluation of effects on development includes effects on puberty onset, postnatal
growth, and effects on the development of the teeth, sensory organs, and other systems.
Research on birth outcomes includes birth defects, infant mortality, preterm birth, and
low birth weight. A few studies of pregnancy-induced hypertension and eclampsia have
been conducted and are reported on in the section on hypertension (Section 4.4.2.1).
Briefly, the relatively small number of studies found consistently positive associations
between blood Pb levels and pregnancy-induced hypertension. Biomarkers of Pb
exposure, including blood Pb and bone Pb, are used in the epidemiologic studies
reviewed in this section. Bone Pb typically indicates cumulative exposure to Pb, whereas,
blood Pb may indicate more recent exposure. However, Pb can also be remobilized from
the bone during times of active bone remodeling, such as pregnancy or lactation.
Therefore, blood Pb also may reflect cumulative Pb exposure. Toxicological studies
typically report exposure using blood Pb. More detailed discussion of these measures and
Pb transfer via umbilical cord blood Pb across the placenta, and via lactation is given in
Section 3.2.2.4 on Pb Toxicokinetics.
Overall, the recent literature on reproductive effects of Pb exposure continues to support
associations reported in earlier Pb AQCDs between Pb exposure and effects on various
parameters of sperm (function, motility, count, integrity, histology). The toxicological
and epidemiologic literature of developmental effects of Pb exposure also indicates that
Pb exposure is associated with delayed onset of puberty in both males and females.
Associations between Pb exposure and other reproductive and developmental effects
have less consistent findings. The recent information from epidemiologic and
toxicological studies is integrated with conclusions from previous Pb AQCDs.
4.8.1 Effects on Development
The 2006 Pb AQCD (U.S. EPA. 2006b) reported Pb-associated developmental effects on
teeth, sensory organs, the GI system, the liver, and postnatal growth as well as delayed
4-589
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puberty onset (U.S. EPA. 2006b). There was recognition that Pb is transferred across the
placenta and through the breast milk, contributing to exposure during development. The
2006 Pb AQCD reported delayed puberty onset in both male and female populations in
animal toxicology studies showing associations with dam blood Pb levels of-40 (ig/dL
and pup blood Pb levels of 26 (ig/dL. The research reported in this ISA continues to find
delayed puberty onset with Pb exposure at even lower Pb doses in animal toxicology
studies as is detailed below. Mechanistic understanding of delayed puberty onset is also
reported in this ISA. Lower dose Pb exposure studies in animal toxicology are also
reported in studies of retinal function and postnatal growth in this ISA. Studies included
in this ISA expand upon evidence reported in previous Pb AQCDs for the
aforementioned systems sensitive to developmental effects with recent studies showing
effects at lower doses of Pb. This section does not cover associations between Pb and
neurodevelopmental impacts, which are discussed in detail in Section 4.3. The studies
presented in the following text and tables are grouped by study design and
methodological strength.
4.8.1.1 Effects on Puberty among Females
Recent toxicological studies of rodents have examined the effects of Pb on pubertal and
reproductive organ development and on biomarkers of pubertal development among
females. There have also been recent epidemiologic studies examining associations
between blood Pb levels and onset of puberty among girls, which are summarized in
Table 4-36 and in the text below. All of the epidemiologic studies examined concurrently
measured blood Pb and puberty and are reported below. Additionally, while there was a
longitudinal investigation by Naicker et al. (2010). who followed girls to determine their
age of menarche, blood Pb levels were measured once at 13 years of age.
4-590
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Table 4-36 Summary of epidemiologic studies of associations between blood Pb levels and puberty for females.
Study Population
Study Location
Reference3 Outcome Study Years
Selevan et Tanner Girls ages 8-18 yr
a I. (2003) staging and from the NHANES
age at III study
menarche N NHWhite=600
or\c
NH Black-oUO
N Mexican-
American" ' "1
U.S.A.
1988-1994
Methodological
Details
Cross-sectional
study using
ordinal logistic
regression and
Cox proportional
hazards
Blood Pb Data
in ug/dL
Geometric mean
Non-Hispanic (NH) Whites:
1.4
NH Blacks: 2.1
Mexican-Americans: 1.7
Blood Pb levels>5ug/dl_:
NH Whites: 2.7%
NH Blacks: 11.6%
Mexican-Americans: 12.8%
Blood Pb levels >10 ug/dL:
NH Whites: 0.3%
NH Blacks: 1.6%
Mexican-Americans: 2.3%
Adjusted Effect Estimates
OR (95% Cl)
Breast development
MUl \A/(->i+^c-
INH vvnites
1 ug/dL: lOO(Ref)
3ug/dL: 0.82(0.47, 1.42)
NH Blacks:
1 ug/dl_: lOO(Ref)
3 ug/dl_: 0.64 (0.42, 0.97)
Mexican Americans:
1 ug/dL: lOO(Ref)
3ug/dL: 0.76(0.63, 0.91)
Pubic hair development
NH Whites:
"I i in/Hi • "I 00 fRpf\
Potential
Confounders
Adjusted for in
Analysis
For breast
development: Age,
age2, height, BMI,
family income, ever
smoked>100
cigarettes, dietary Fe,
dietary vitamin C,
dietary Ca2+.
For pubic hair
development: Age,
age2, height, family
income, ever
smoked>100
cigarettes, anemia,
dietary Fe, dietary
vitamin C
3ug/dL: 0.75(0.37, 1.51)
NH Blacks:
1 ug/dl_: lOO(Ref)
3ug/dL: 0.62(0.41, 0.96)
Mexican Americans:
1 ug/dL: 1.00(Ref)
3ug/dL: 0.70(0.54, 0.91)
Considered in all
models: age, smoking,
dietary Ca , dietary
Fe, dietary vitamin C,
dietary total fat,
anemia, urban
residence, family
income
4-591
-------
Table 4-36 (Continued): Summary of epidemiologic studies of associations between blood Pb levels and puberty for females.
Reference3 Outcome
Selevan et Tanner
al. (2003) staging and
(Continued) age at
menarche
Study Population
Study Location
Study Years
Girls ages 8-18 yr
from the NHANES
III study
N NH White =600
N NH Black=805
N Mexican-
American" ' o1
U.S.A.
-1 QQQ -IQQyl
i yoo- 1 yy^t
Methodological
Details
Cross-sectional
study using
ordinal logistic
regression and
Cox proportional
hazards
Blood Pb Data
in ug/dL
Geometric mean
Non-Hispanic (NH) Whites:
1.4
NH Blacks: 2.1
Mexican-Americans: 1.7
Blood Pb levels>5ug/dl_:
NH Whites: 2.7%
NH Blacks: 11.6%
Mexican-Americans: 12.8%
Blood Pb levels >10 ug/dL:
NH Whites: 0.3%
NH Blacks: 1.6%
Mexican-Americans: 2.3%
Adjusted Effect Estimates
HR (95% Cl)
"included only girls ages 8-16 yr
Age at menarche
NH Whites:
1 ug/dL: 1.00(Ref)
3ug/dL: 0.74(0.55, 1.002)
NH Blacks:
1 ug/dl_: 1.00(Ref)
3 ug/dl_: 0.78 (0.63, 0.98)
Mexican Americans:
1 ug/dL: 1.00(Ref)
3ug/dL: 0.90(0.73, 1.11)
Potential
Confounders
Adjusted for in
Analysis
For age at menarche:
Height, BMI, family
income, anemia,
dietary Ca2+.
Considered in all
models: age, smoking,
dietary Ca +, dietary
Fe, dietary vitamin C,
dietary total fat,
anemia, urban
residence, family
income
4-592
-------
Table 4-36 (Continued): Summary of epidemiologic studies of associations between blood Pb levels and puberty for females.
Reference3 Outcome
Study Population
Study Location
Study Years
Methodological
Details
Blood Pb Data
in ug/dL
Adjusted Effect Estimates
Potential
Confounders
Adjusted for in
Analysis
Wu et al.
(2003b)
Tanner
staging and
age at
menarche
Girls ages 8-16 yr
from the NHANES
III study
N=1706
U.S.A.
1988-1994
Cross-sectional
study using
logistic
regression with
weighting
Mean (SD): 2.5 (2.2)
Weighted proportion of the
sample with blood Pb
5.0-21.7: 5.9%
OR (95% Cl)
Breast development
0.7-2.0 ug/dl_: 1.00(Ref)
2.1-4.9 ug/dl_: 1.51 (0.90,2.53)
5.0-21.7 ug/dl_: 1.20(0.51,
2.85)
Pubic hair development
0.7-2.0 ug/dl_: 1.00(Ref)
2.1-4.9 ug/dl_: 0.48(0.25, 0.92)
5.0-21.7 ug/dl_: 0.27(0.08,
0.93)
Race/ethnicity, age,
family size, residence,
poverty income, ratio,
BMI
Menarche
0.7-2.0 ug/dl_: 1.00(Ref)
2.1-4.9ug/dL: 0.42(0.18, 0.97)
5.0-21.7 ug/dl_: 0.19(0.08,
0.43)
4-593
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Table 4-36 (Continued): Summary of epidemiologic studies of associations between blood Pb levels and puberty for females.
Reference3
Gollenberg
et al. (2010)
Outcome
Luteinizing
hormone
(LH)and
inhibin B
Study Population
Study Location
Study Years
Girls ages 6-11 yr
from the NHANES
III study
N=705
Methodological
Details
Cross-sectional
study using
survey logistic
regression
Blood Pb Data
in ug/dL
Median 2.5 (range 0.07-
29.4)
Blood Pb>10ug/dL: 5%
Adjusted Effect Estimates
OR (95% Cl)
Exceeding pubertal inhibin B
cutoff (>35pg/mL)
<1 ug/dl_: 1.00(Ref)
1-4.9 ug/dl_: 0.38(0.12, 1.15)
>5ug/dL: 0.26(0.11, 0.60)
Potential
Confounders
Adjusted for in
Analysis
Age, race/ethnicity,
BMI, census region,
poverty-income ratio
U.S.A.
1988-1994
Exceeding pubertal LH cutoff
(>0.4 mlU/mL)
<1 ug/dL: 1.00(Ref)
1-4.9 ug/dl_: 0.98(0.48, 1.99)
>5ug/dL: 0.83(0.37, 1.87)
*Note: a sensitivity analysis
including only those with blood
Pb <10 ug/dL had similar results
but ORs were slightly
attenuated
4-594
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Table 4-36 (Continued): Summary of epidemiologic studies of associations between blood Pb levels and puberty for females.
Reference3 Outcome
Study Population
Study Location
Study Years
Methodological
Details
Blood Pb Data
in ug/dL
Adjusted Effect Estimates
Potential
Confounders
Adjusted for in
Analysis
Denham et Age at
al. (2005) menarche
10-to 16.9-yr-old
girls in the
Akwesasne
community
N=138
Akwesasne
Mohawk Nation
(boundaries of
New York,
Ontario, and
Quebec
Years not reported
Cross-sectional
study using
probit and logistic
regression
Mean (SD): 0.49 (0.905)
Median: 1.2
Coefficients for binary logistic
regression predicting menarche
with Pb centered at the mean:
log blood Pb
-1.29(p-value0.01)
log blood Pb -squared:
-1.01 (p-value 0.08)
Non-linear relationship
observed and Pb below the
mean did not appear to affect
the odds of menarche.
Increasing blood Pb from 0.49
to 0.98 ug/dL decreased the
odds of menarche attainment by
72%
Age, SES, BMI
Naicker et
al. (2010)
Self-
reported
Tanner
staging at
age 13 and
age at
menarche
Girls of blacker
mixed ancestry
who were enrolled
in the Birth to
Twenty (Bt20)
cohort (born in
1990) that lived in
Johannesburg/
Soweto for at least
6 mo after birth
N=682
Johannesburg/
Soweto, South
Africa
Born in 1990
Cross-sectional
and longitudinal
study using
logistic
regression
Mean (SD): 4.9 (1.9)
Blood Pb levels > 10 ug/dL:
1%
OR (95% Cl)
Delay in breast development at
age 13
<5ug/dL: 1.00(Ref)
>5ug/dL:2.34(1.45, 3.79)
Delay in pubic hair development
at age 13
<5ug/dL: 1.00(Ref)
>5ug/dL: 1.81 (1.15,2.84)
Delay in attainment of
menarche at age 13
<5ug/dL: 1.00(Ref)
>5ug/dL:2.01 (1.38,2.94)
BMI
4-595
-------
Table 4-36 (Continued): Summary of epidemiologic studies of associations between blood Pb levels and puberty for females.
Reference3
Den Hond
et al. (2011)
Outcome
Tanner
staging, age
at
menarche,
regular
menses
Study Population
Study Location
Study Years
Girls ages 14 and
15 yr, in their 3rd
year of secondary
education and
living in the same
study areas for at
least 5 years
Methodological
Details
Cross-sectional
study using
logistic
regression
Blood Pb Data
in ug/dL
Median: 1.81
10th percentile: 0.88
90th percentile: 3.81
Adjusted Effect Estimates
OR (95% Cl)
Pubic hair development with
doubling of exposure:
0.65(0.45, 0.93)
*Association was no longer
e-to+ic'+i^'^ ll\/ c- i« n if i^o nt in/ki^n
Potential
Confounders
Adjusted for in
Analysis
Age, BMI, smoking,
oral contraceptive use
Considered but did not
include: food intake
and lifestyle
parameters
N=792
Flanders, Belgium
2003-2004
PCS marker included in the
model
No association between Pb and
breast development (results not
given)
4-596
-------
Table 4-36 (Continued): Summary of epidemiologic studies of associations between blood Pb levels and puberty for females.
Reference3 Outcome
Study Population
Study Location
Study Years
Methodological
Details
Blood Pb Data
in ug/dL
Adjusted Effect Estimates
Potential
Confounders
Adjusted for in
Analysis
Tomoum et
al. (2010)
Hormones
and pubertal
development
using
Tanner
staging
Healthy children
aged 10-1 Syr;
seeking treatment
for minor health
problems and
living in one of two
designated areas
(one with high-risk
forPb
contamination and
one with no Pb
source)
N=20
Cross-sectional
study using Chi-
square
Blood Pb
Not stated for girls only
(combined with boys in the
study the mean was 9.46
[3.08])
Breast Development
<10 ug/dL:
Stage 2: 36.4%
Stage 3: 63.6%
> 10 ug/dL:
Stage 2: 100%
Stage 3: 0%
Chi-square p-value<0.01
None
Cairo, Egypt
2007
Pubic Hair Development
<10 ug/dL:
Stage 2: 36.4%
Stage 3: 63.6%
> 10 ug/dL:
Stage 2: 77.8%
Stage 3: 22.2%
Chi-square p-value>0.05
4-597
-------
Table 4-36 (Continued): Summary of epidemiologic studies of associations between blood Pb levels and puberty for females.
Reference3 Outcome
Study Population
Study Location
Study Years
Methodological
Details
Blood Pb Data
in ug/dL
Adjusted Effect Estimates
Potential
Confounders
Adjusted for in
Analysis
Tomoum et
al. (2010)
(Continued)
Hormones
and pubertal
development
using
Tanner
staging
Healthy children
aged 10-1 Syr;
seeking treatment
for minor health
problems and
living in one of two
designated areas
(one with high-risk
forPb
contamination and
one with no Pb
source)
N=20
Cross-sectional
study using Chi-
square
Blood Pb
Not stated for girls only
(combined with boys in the
study the mean was 9.46
[3.08])
Median FSH level
<10 ug/dL:
7.3 (IQR 7.9) mlU/mL
> 10 ug/dL:
3.17 (IQR 2.6) mlU/mL
Mann-Whitney U-test
p-value <0.05
Median LH level
<10 ug/dL:
8.9(IQR6.1)mIU/mL
> 10 ug/dL:
1.23 (IQR 2.5) mlU/mL
Mann-Whitney U-test
None
Wolff etal.
(2008);
Wolf and
Daley
(2QQ7)
Pubertal
stages
defined
using
standard
drawings
Cairo, Egypt
2007
9-yr old girls from
the study hospital
and nearby
pediatric offices
N=192
New York City, NY
1996-1997
Cross-sectional Median: 2.4
study using
Poisson
multivariate
regression with
robust error
variance
p-value<0.05
Median Estradiol levels were
similar among girls with blood
Pb levels of <10 ug/dL and
> 10 ug/dL (quantitative results
not given).
PR (95% Cl) (unit not given,
assume results are per 1 ug/dL)
Breast stage:
1.01 (0.79, 1.30)
Pubic hair stage:
1.25(0.83, 1.88)
For breast
development: Age,
BMI, race
For hair stage: Height,
private clinic, race
aStudies are presented in order of first appearance in the text of this section, which is based on study design and methodological strength.
4-598
-------
Multiple studies have been performed examining blood Pb levels and puberty using
NHANES III data (Gollenberg etal.. 2010: Selevan et al. 2003: Wu et al.. 2003b). A
study included girls 8-18 years of age and reported the results stratified by race (Selevan
et al.. 2003). This study also included many important potential confounders, such as
nutritional factors. Higher blood Pb levels were associated with lower Tanner stage of
breast and pubic hair development and later age at menarche among African Americans
and with lower stage of breast and pubic hair development among Mexican Americans.
For whites, the associations were in the same directions, but none reached statistical
significance. Another NHANES III study included girls aged 8-16 years and did not
stratify the results by race/ethnicity (Wu et al., 2003b). The study, which also controlled
for a variety of potential confounders (although not nutritional factors), reported an
association between increased blood Pb and delayed attainment of menarche and pubic
hair development but not breast development. The associations were observed even at
blood Pb levels of 2.1-4.9 (ig/dL compared to girls with blood Pb levels <2.1 (ig/dL. In a
study of girls aged 6-11 years old from NHANES III data, higher blood Pb levels were
associated with lower inhibin B, a protein that inhibits follicle stimulating hormone
(FSH) production, but no association was observed for luteinizing hormone (LH).
(Gollenberg et al., 2010). The inverse association between blood Pb and inhibin B was
greater among girls with Fe deficiency compared to those with high Pb but sufficient
Fe levels. Inhibin B and LH were chosen for this study because, as the authors indicated,
these hormones are, "believed to be relevant for younger girls... near the onset of puberty
and... serve as markers for hypothalamic-pituitary-gonadal functioning."
A study of girls aged 10-16.9 years of age in the Akwesasne Mohawk Nation reported a
nonlinear association between higher blood Pb and greater age at menarche (Denham et
al., 2005). No association was observed below blood Pb of 0.49 (ig/dL in a nonlinear
model of the Pb-menarche relationship. A study conducted in South Africa reported an
association between increased blood Pb levels and older age at first menarche and
pubertal development (Naicker et al.. 2010). Another study reporting on girls with low
blood Pb concentrations observed an association between higher blood Pb and less pubic
hair development but not breast development (Den Hond et al.. 2011). The association
was no longer statistically significant when a marker for polychlorinated biphenyl
exposure was included in the model. A study among girls aged 10-13 years (median:
12 years) reported lower levels of FSH and LH levels in the group with blood Pb of at
least 10 (ig/dL compared to the group with blood Pb less than 10 (ig/dL (Tomoum et al..
2010). In addition, there were some indications of lower Tanner stages of breast
development associated with Pb levels of at least 10 (ig/dL, but this relationship was not
present for stages of pubic hair development and there was no control for potential
confounders. A study performed in NYC among 9-year old girls reported no association
4-599
-------
between Pb levels and pubertal development (Wolff etal.. 2008). but this age group may
be too young to study when investigating delayed puberty as the outcome.
4-600
-------
o Highest Concentration
» Lowest Cone, with Response
A Highest Cone, with No Response
o Lowest Concentration
Blood Lead Level (iig/dl)
Puberty; Neonate/adult; Mouse;
Female; lavicoli etal. (2006)
Neurotransmitter ; Adult; Mouse;
Both; Leasure et al. (2008)
Physical development; Adult; Mouse;
Male; Leasure et al. (2008)
Eye; Adult; Rat; Both; Fox et al. (2008)
Redox-oxidative stress; Adult; Rat;
Male; Nava-Hernandez et al. (2009)
Sperm; Adult; Rabbit; Male; Moorman
etal. (1998)
Neurobehavioral; Adult; Mouse; Male;
Leasure etal. (2008)
Hematological parameters; Adult; Rat;
Both;Teijon etal. (2006)
Histology; Adult; Rat; Both; Teijon et
al. (2006)
Biomarkers; Adult; Rat; Both; Teijon et
al. (2006)
Physical development; Adult; Rat;
Both;Teijon etal. (2006)
Note: This figure illustrates reproductive and developmental effects associated with Pb exposure in studies that examined multiple
exposure concentrations. Dosimetric representation reported by blood Pb level. (Studies are described in Table 4-37).
Figure 4-36 Toxicological concentration-response array for reproductive and
developmental effects of Pb.
4-601
-------
Table 4-37 Toxicological concentration-response array summary for
reproductive and developmental effects of Pb presented in
Figure 4-36.
Reference
Exposure
Blood Pb level with
Effect (|jg/dL) Altered Outcome
lavicoli et al. (2006a)
Lifetime Pb exposure starting in
utero. Dam exposure during
gestation (0.2- 40 ppm daily
dietary Pb); offspring
continuous dietary exposure
until the termination of the
experiment at puberty
1.3&13
Delayed onset female puberty
Leasure et al. (2008)
Gestational and Lactational 10&42
Exposure to one of 3 Pb
acetate doses (27, 55, or
109 ppm) in dam drinking 10, 24 & 42
water, 2 weeks before mating,
through gestation out to
PND10. 10&42
Neurotransmitter, Dopamine
homeostasis
Physical Development, Adult
obesity (males)
Aberrant response to
amphetamine
Fox et al. (2008)
Dam exposure 2 weeks before
mating, throughout gestation,
and until PND10 (27, 55, or
109 ppm Pb acetate in drinking
water)
12
Retinal aberrations
Nava-Hernandez et al.
(2009)
Adult male rats, 10 mg/kg BW
Pb chloride i.p. once daily for
8 weeks
19.5
Sperm affected via redox
imbalance
Moorman et al. (1998)
Adult male rabbits, chronic
exposure: subcutaneous
injection, loading dose of
0.2-3.85 mg/kg BW Monday
(M), Wednesday (W) and
Fridays (F) weeks 6-10,
followed by maintenance dose
of 0.13-2.0 mg/kg BW
Pb acetate MWF over weeks
11-20 of the study
25-130
Semen quality affected
Teijon et al. (2006)
Wistar rats, 200 ppm or
400 ppm Pb acetate in drinking
water to dams from GD1
through lactation & 1 and
3 months postweaning to pups
40 & 100
40 & 100
40 & 100
100
Hematology
Histology:
Offspring renal & hepatic
Biomarker:
Offspring renal function
Physical development:
Birth weight
4-602
-------
Table 4-37 (Continued): Toxicological concentration-response array summary for
reproductive and developmental effects of Pb presented in
Figure 4-36.
Reference
Exposure
Blood Pb level with
Effect (|jg/dL)
Altered Outcome
Fox et al. (2008)
Dam exposure 2 weeks before
mating, throughout gestation,
and until PND10(27, 55, or
109 ppm Pb acetate in drinking
water)
12
Retinal aberrations
Nava-Hernandez et al.
(2009)
Male adult albino rats, 10
mg/kg BW Pb chloride i.p. once
daily for 8 weeks
19.5
Sperm affected via redox
imbalance
Moorman et al. (1998)
Fox et al. (2008)
Adult male rabbits, chronic
exposure: subcutaneous
injection, loading dose of
0.2-3.85 mg/kg BW Monday
(M), Wednesday (W) and
Fridays (F) weeks 6-10,
followed by maintenance dose
of 0.13-2.0 mg/kg BW
Pb acetate MWF over weeks
11-20 of the study
25-130
Semen quality affected
Teijon et al. (2006) Wistar rats, 200 ppm or
400 ppm Pb acetate in drinking
water to dams from GD1
through lactation & 1 and
3 months postweaning to pups
40 & 100
40 & 100
40 & 1 00
100
Hematology
Histology:
Offspring renal & hepatic
Biomarker:
Offspring renal function
Physical development:
Birth weight
Dam exposure 2 weeks before
mating, throughout gestation,
and until PND10 (27, 55, or
109 ppm Pb acetate in drinking
water)
12
Retinal aberrations
Earlier studies showed that prenatal and lactational exposures to Pb can cause a delay in
the onset of female puberty in rodents. Recent studies corroborate these findings and
show that puberty onset is one of the more sensitive markers of reproductive or
developmental effects of Pb exposure as is demonstrated in the exposure response array
(Figure 4-36 and Table 4-37; including outcomes described in sections that follow).
Dumitrescu et al. (2008b) exposed adult Wistar female rats to varying doses of Pb acetate
(50-150 ppb) in drinking water for 3 months before mating and during pregnancy.
Vaginal opening, an indicator of sexual maturation, was statistically significantly delayed
in pups from all Pb treated groups when compared to pups from non-treated dams. The
age at vaginal opening in female pups from the Pb treated groups increased, in a
concentration-dependent manner, from 39 days to 43-47 days. The authors also observed
4-603
-------
a correlation between body weight and age at vaginal opening meaning that as body
weight decreased the age at vaginal opening increased. This effect also exhibited a
concentration-dependent relationship.
In recent studies by one lab, lavicoli et al. (lavicoli et al., 2006a; 2004) reported a
statistically significant delay in several indicators of sexual maturity in female offspring
(Swiss mice, Fl generation) born to dams that ingested 0.02- 40 ppm Pb in their daily
diet during pregnancy; offspring had continuous dietary exposure until the termination of
the experiment at puberty. Maternal ingestion of Pb at the various doses resulted in
female pup blood Pb levels of 0.7, 1.3, 1.6, 1.9, 3.5, 3.9, 8 and 13 (ig/dL. There was a
dose-dependent delay in age at vaginal opening, age of first estrus, age of vaginal plug
formation, and age of first parturition . Vaginal opening was significantly different
between the following groups at blood Pb levels of 0.7, 1.3, 1.6, 1.9-3.9, 8 and 13 (ig/dL.
Estrus cyclicity, age of first vaginal plug detection and age of first parturition were
significantly different between the groups with blood Pb levels of 0.7, 1.3-1.6, 1.9-3.9, 8
and 13 (ig/dL. After in utero Pb exposure, sexual development was delayed in a
concentration-dependent fashion in Pb-exposed female offspring.
In addition, Pb-induced shifts in sexual maturity were observed in the subsequent
generation (F2 generation) across the same dose range (lavicoli et al.. 2006a). These F2
animals continued to be exposed to the aforementioned concentrations of Pb over
multiple generations through the diet. Results in the F2 generation including blood Pb
levels and markers of delayed sexual maturity in female offspring closely resembled
those of the FI generation.
In the 2006 Pb AQCD (U.S. EPA. 2006b). it was reported that a statistically significant
reduction in the circulating levels of insulin-like growth factor 1 (IGF-1), LH, and
estradiol (E2) was associated with Pb-induced delayed puberty onset in Fischer 344 pups,
providing evidence for the effect of Pb on a mode of action for delayed puberty onset.
Subsequently, Pine et al. (2006) evaluated whether IGF-1 replacement could reverse the
effects of Pb on delayed female puberty onset, providing more evidence for the effect on
a mode of action associated with Pb on delayed puberty onset. The authors reported that
offspring from dams exposed to Pb during gestation and lactation (daily oral gavage of
dam with 1.0 mL solution of Pb acetate 12 mg/mL; mean maternal blood Pb level
40 (ig/dL) exhibited a marked increase in LH and luteinizing hormone releasing hormone
(LHRH) secretion after IGF-1 administration (200 ng3/(iL i.p. injection twice daily from
PND23 until the appearance of vaginal opening which appears in control animals at
-PND40) resulting in restored timing of vaginal opening to that of control animals. It
should be noted that, IGF-1 replacement in Pb-exposed animals did not cause advanced
puberty onset over non-Pb-exposed controls. The results of this study provide support to
4-604
-------
the theory that Pb-induced delayed onset of puberty may be due to disruption of pulsatile
release of sex hormones (U.S. EPA. 2006b) and not necessarily due to a direct toxic
effect on the hypothalamic-pituitary-gonadal axis (Salawu et al., 2009), and IGF-1 may
play a prominent role in the process.
In summary, epidemiologic studies consistently show an association between higher
concurrent blood Pb and delayed pubertal development in girls. This association is
apparent even at low blood Pb levels (mean and/or median concurrent blood Pb levels of
1.2 to 9.5 (ig/dL). Most of the studies had good sample sizes and controlled for some
potential confounders. Nutritional information was rarely controlled for although this
could be important, especially in populations where malnutrition is prevalent. These
epidemiologic studies are cross-sectional, which does not allow for the study of
temporality between Pb exposure and pubertal onset nor does it consider the influence of
past blood Pb levels. Recent evidence from the toxicology literature continues to indicate
Pb-induced delays in the onset of puberty. Further, the biological plausibility of delayed
puberty is expanded with the toxicological literature that shows this pathway is mediated
by IGF-1.
4.8.1.2 Effects on Puberty among Males
Recent epidemiologic studies examining the association between blood Pb and onset of
puberty in males are summarized in Table 4-38. The majority of studies used concurrent
measures of blood Pb and puberty (Den Hond et al.. 2011; Tomoum et al.. 2010; Hauser
et al., 2008). but Williams et al. (2010) performed a longitudinal analysis of blood Pb
levels measured at ages 8-9 years and pubertal onset, following the participants for 3
years. Little epidemiologic information was available regarding pubertal onset in the
2006 Pb AQCD (U.S. EPA. 2006b).
4-605
-------
Table 4-38 Summary of recent epidemiologic studies of associations between blood Pb levels and puberty for
males.
Reference3
Mauser et al.
(2008)
Outcome
Pubertal
stages
defined
using
standard
drawings
Study Population
Study Location
Study Years
Healthy boys aged
8-9 yr
N=489
Methodological
Details
Cross-sectional
study using
multivariable
logistic regression
Blood Pb
in ug/dL
Median: 3
Data
(IQR 2-5)
Blood Pb>10ug/dL: 3%
Adjusted Effect
Estimates
OR (95% Cl)
Pubertal onset based on
testicular volume
<5 ug/dL:
1.00(Ref)
Potential
Confounders
Adjusted for in
Analysis
Gestational age,
height, BMI, age at
exam
Considered but did not
in^l i trie*- r\*^ r^n+'^l
Chapaevsk, Russia
2003-2005
> 5 ug/dL:
0.83(0.43, 1.59)
*after adjustment for
macronutrients, the OR
(95% Cl) became 0.66
(0.44, 1.00)
education, household
income
Genital development
<5 ug/dL:
1.00(Ref)
> 5 ug/dL:
0.57 (0.34, 0.95)
*after adjustment for
macronutrients, the OR
(95% Cl) became 0.52
(0.31, 0.88)
Pubic hair development
<5ug/dL: 1.00(Ref)
>5ug/dL:0.74(0.34,
1.60)
4-606
-------
Table 4-38 (Continued): Summary of recent epidemiologic studies of associations between blood Pb levels and puberty for
males.
Study Population
Study Location Methodological Blood Pb Data
Reference3 Outcome Study Years Details in ug/dL
Williams et al. Pubertal Healthy boys aged Longitudinal cohort Blood Pb at ages 8-9 yr
(2010) stages 8-9 yr at enrollment using Cox Median' 3 (IQR 2-5)
defined who had 3 annual proportional
using follow-up hazards
standard evaluations Blood Pb level >
drawings 10ug/dL:3%
A Q A
-481
Chapaevsk, Russia
2003-2008
Adjusted Effect
Estimates
HR (95% Cl)
Pubertal onset based on
testicular volume
<5 ug/dL:
lOO(Ref)
> 5 ug/dL:
0.73 (0.55, 0.97)
Genital development
<5 ug/dL:
lOO(Ref)
> 5 ug/dL:
0.76 (0.59, 0.98)
Pubic hair development
<5 ug/dL:
lOO(Ref)
> 5 ug/dL:
0.69(0.44, 1.07)
Potential
Confounders
Adjusted for in
Analysis
Birthweight,
gestational age,
energy intake,
proportion of fat
consumption,
proportion of protein
consumption, maternal
alcohol consumption
during pregnancy,
height at study entry,
BMI at study entry,
household income,
parental education
NOTE: exclusion of
BMI and height, in
case they were part of
the causal pathway,
resulted in very similar
estimates
Considered but not
included: parity,
maternal or household
smoking during
pregnancy, maternal
age at birth
4-607
-------
Table 4-38 (Continued): Summary of recent epidemiologic studies of associations between blood Pb levels and puberty for
males.
Reference3
Tomoum et al.
(2010)
Outcome
Hormones
and pubertal
development
using Tanner
staging
Study Population
Study Location
Study Years
Healthy children
aged 10-13 yr
seeking treatment
for minor health
problems and living
in one of two
designated areas
(one with high-risk
for Pb contamination
and one with no Pb
source)
Methodological
Details
Cross-sectional
study using
Chi-square
Blood Pb Data
in ug/dL
Not stated for boys only
(combined with girls in the
study the mean was 9.46
[3.08])
_ . ...
r Ot6Dtl3l
Confounders
Adjusted Effect Adjusted for in
Estimates Analysis
Testicular size None
<10 ug/dL:
O4-^i«_. A . r\r\/
Stage 1. 0%
Stage 2: 44.4%
Stage 3: 55.6%
> 10 ug/dL:
Stage 1: 33.3%
Stage 2: 66.7%
Stage 3: 0%
N=21
Cairo, Egypt
2007
Chi-square p-value<0.01
Pubic Hair Development
<10 ug/dL:
Stage 1: 0%
Stage 2: 55.6%
Stage 3: 44.4%
> 10 ug/dL:
Stage 1: 33.3%
Stage 2: 66.7%
Stage 3: 0%
Chi-square p-value<0.05
4-608
-------
Table 4-38 (Continued): Summary of recent epidemiologic studies of associations between blood Pb levels and puberty for
males.
Reference3
Tomoum et al.
(2010)
(Continued)
Outcome
Hormones
and pubertal
development
using Tanner
staging
Study Population
Study Location
Study Years
Healthy children
aged 10-13 yr
seeking treatment
for minor health
problems and living
in one of two
designated areas
(one with high-risk
for Pb contamination
and one with no Pb
source)
Methodological
Details
Cross-sectional
study using
Chi-square
Blood Pb Data
in ug/dL
Not stated for boys only
(combined with girls in the
study the mean was 9.46
[3.08])
_ . ...
r Ot6Dtl3l
Confounders
Adjusted Effect Adjusted for in
Estimates Analysis
Penile staging None
<10 ug/dL:
Stage 1: 11.1%
Stage 2: 44.4%
Stage 3: 44.4%
> 10 ug/dL:
Stage 1: 58.3%
Stage 2: 41. 7%
Stage 3: 0%
N=21
Cairo, Egypt
2007
Chi-square p-value<0.05
Mean testosterone level
<10 ug/dL:
4.72 (SD 1.52)ng/mL
> 10 ug/dL:
1.84(SD 1.04)ng/mL
Student t-test
p-value <0.05
4-609
-------
Table 4-38 (Continued): Summary of recent epidemiologic studies of associations between blood Pb levels and puberty for
males.
Reference3
Tomoum et al.
(2010)
(Continued)
Outcome
Hormones
and pubertal
development
using Tanner
staging
Study Population
Study Location
Study Years
Healthy children
aged 10-13 yr
seeking treatment
for minor health
problems and living
in one of two
designated areas
(one with high-risk
for Pb contamination
and one with no Pb
source)
Methodological
Details
Cross-sectional
study using
Chi-square
Blood Pb Data
in ug/dL
Not stated for boys only
(combined with girls in the
study the mean was 9.46
[3.08])
Potential
Confounders
Adjusted Effect Adjusted for in
Estimates Analysis
Median FSH level None
<10 ug/dL:
5.6 (IQR 7.6) mlU/mL
>>l f\ 1 I'M Ift 1 •
_ 10 ug/dL.
1.88 (IQR 1.4)mIU/mL
Mann-Whitney U-test
p-value <0.05
Median LH level
N=21
<10 ug/dL:
6.5 (IQR 5.8) mlU/mL
> 10 ug/dL:
0.79 (IQR 1.0)mIU/mL
Den Hond et
al. (2011)
Tanner
staging and
gyneco-
mastia
Cairo, Egypt
2007
Boys ages 14 and
15 yr, in their 3rd
year of secondary
education and living
in the same study
areas for at least 5
years
Cross-sectional
study using logistic
regression
Blood Pb
Median: 2.50
10th percentile: 1.20
90th percentile: 5.12
Mann-Whitney U-test
p-value <0.05
OR (95% Cl) for
Gynecomastia with
doubling of exposure
1.84 (1.11, 3.05)
No association between
Pb and pubic hair or
Parental education,
age, BMI, smoking
status
Considered but not
included: food intake,
lifestyle parameters
N=887
Flanders, Belgium
2003-2004
genital development
(results not given)
NOTE: results were
the same when
hexachlorobenzene
was included in the
model
"Studies are presented in order of first appearance in the text of this section, which is based on study design and methodological strength.
4-610
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Studies were performed among a cohort of Russian boys enrolled between ages 8-9 years
("Williams et al.. 2010; Hauser et al.. 2008). The area where these studies were performed
had various environmental contaminants that may be associated with delayed puberty
onset, such as dioxin, polychlorinated biphenyls, and other metals, present but these were
not included in the analyses (although preliminary analyses found no correlation between
blood Pb levels and serum dioxin levels). Both the cross-sectional study (Hauser et al..
2008) and the prospective study with annual follow-ups (Williams et al.. 2010)
demonstrated an association; higher blood Pb levels at 8-9 years of age was associated
with later onset of puberty. In a study of boys in Egypt, boys with higher blood Pb had
delayed pubertal onset compared to those with lower levels (Tomoum et al., 2010). In
addition, compared to the low blood Pb group, those boys with higher blood Pb had
lower testosterone, FSH, and LH levels but there was no control for potential
confounding. A study in Flanders reported no associations between blood Pb
concentration and pubertal development among 14- and 15-year old boys (Den Hond et
al.. 2011). However, higher blood Pb levels were associated with an increased odds of
gynecomastia.
No recent toxicological studies address Pb-induced male sexual maturation and
development, but earlier studies do provide support to findings in epidemiologic cohorts.
Pb exposure resulted in delayed sexual maturity as measured by prostate weight in male
Sprague-Dawley pups at PND35. These pups were exposed chronically to 1,500 or
4,500 ppm Pb acetate in dam or their own drinking water from GD5 until PND85 and had
blood Pb ranges from low to high of 88-196 and 120-379 ug/dL, respectively (Ronis et
al.. 1998b). Cynomolgus monkeys exposed to Pb over a lifetime (an oral capsule of
1,500 (ig/kg body weight/day for 10 years, blood Pb levels ranging from 30-60 ug/dL)
had altered pituitary and Sertoli cell function along with decreases in inhibin/FSH ratio
and reduced gonadotropin-releasing hormone (GnRH) stimulation of LH release in
adulthood (Foster et al.. 1993). all indicators that are important in proper sexual
maturation. Further mechanistic understanding of the effect of Pb can be gleaned from
studies in adult male Wistar rats exposed to Pb for 1 month (starting at PND56, 1,000 or
3,000 ppm Pb acetate in drinking water, respective blood Pb levels of 34 or 60 ug/dL)
that showed significant decreases in FSH, ventral prostate weight and serum testosterone
but no change in serum LH (Sokol etal.. 1985). These Pb-exposed adult male rats
(3,000 ppm Pb acetate in drinking water starting at PND56 for 30 days) demonstrated an
impaired pituitary release of LH in response to challenge of the hypothalamic-pituitary-
adrenal (HPA) axis with the opiate antagonist naloxone, an enhanced release of LH from
the pituitary in response to direct stimulation of the pituitary with luteinizing hormone-
releasing hormone (LHRH), an enhanced response to human chorionic gonadotropin
(hCG) by the testes, increased pituitary LH stores, and increased GnRH mRNA levels in
4-611
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the hypothalamus (Klein etal. 1994; Sokol. 1987). Thus, these results indicate that Pb
likely interferes with the male HPA axis, contributing to its reproductive toxicity.
In summary, recent epidemiologic studies have demonstrated an inverse relationship of
Pb on pubertal development among boys at low blood Pb levels (mean and/or median
blood Pb levels of 3.0 to 9.5 (ig/dL). These studies were mostly cross-sectional, but
associations were observed between blood Pb levels and delayed puberty in a
longitudinal study as well ("Williams et al.. 2010). The larger studies controlled for some
potential confounders, with a few studies at least considering the inclusion of dietary
factors, which may be an important confounder, especially in populations with high
prevalence of malnutrition. Some populations, such as the Russian boys cohorts, had
potential exposures to dioxins and polychlorinated biphenyls, but these were not
considered in the analyses with blood Pb levels. No recent toxicological studies were
found that addressed the effect of Pb on male sexual development and maturation;
however, the 2006 Pb AQCD (U.S. EPA. 2006b) supported earlier findings that Pb
exposure may result in delayed onset of male puberty and altered reproductive function
later in life in experimental animals.
4.8.1.3 Effects on Postnatal Stature and Body Weight
Findings from previous toxicological studies of rodents and primates have demonstrated
Pb induced impairment of postnatal growth (U.S. EPA. 2006b). Little epidemiologic
evidence was available in the 2006 Pb AQCD on postnatal growth. Several recent
epidemiologic studies examining the association of various biomarkers of Pb exposure
with stature and body weight have been conducted and the evidence reported is mixed.
One new toxicology study reports effects on body weight.
4-612
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Table 4-39 Summary of recent epidemiologic studies of associations between Pb biomarker levels and
postnatal growth.
Reference3
Study Population
Study Location
Study Years
Methodological
Details
Pb Biomarker Data
Mean (SD)
Adjusted Effect Estimates
Potential Confounders
Adjusted for in Analysis
Afeiche et al.
(2011)
n=523 boys
n=477 girls
Mexico City,
Mexico
Children born
between 1994 and
2005
Longitudinal
cohort using
varying
coefficient
models with
random effects
Maternal bone Pb 1 month
postpartum
Patella: 10.4 (11.8) ug/g
Change in weight at 5 years of
age (g) per 1 SD increase in
maternal bone Pb (95% Cl):
Girls:-171.6 (-275.2,-68.0)
Boys:-35.0 (-132.4, 62.3)
Cohort, maternal age, calf
circumference, height,
education, number of
pregnancies, breast
feeding for 6 months,
Ca2+ treatment, child's
gestational age at birth,
height, repeated
measures of concurrent
child blood Pb
4-613
-------
Table 4-39 (Continued): Summary of recent epidemiologic studies of associations between Pb biomarker levels and postnatal
growth.
Reference
Study Population
Study Location
Study Years
Methodological
Details
Pb Biomarker Data
Mean (SD)
Adjusted Effect Estimates
Potential Confounders
Adjusted for in Analysis
Schell et al.
(2009)
N=244
Albany, New York
1986-1992, and
1992-1998
Longitudinal
cohort study
using multivariate
regression
Maternal blood Pb
During second trimester: 2.8
(2.6) ug/dL
During third trimester: 2.6 (2.2) ug/dL
At delivery: 2.8 (2.4) ug/dL
Infant blood Pb
At delivery: 2.3 (2.7) ug/dl_
At 6 months: 3.2 (3.3) ug/dl_
At 12 months: 6.3 (4.8) ug/dL
(3 (p-value) for maternal second
trimester Pb
Length for age (z-score):
6 month: 0.149(0.05)
12 month: 0.073(0.38)
Weight for age (z-score):
6 month: 0.013(0.89)
12 month: 0.124(0.25)
Weight for length (z-score):
6 month: -0.158(0.16)
12 month: 0.084(0.45)
Head circumference for age
(z-score):
6 month: -0.242(0.01)
12 month:-0.220 (0.05)
Upper arm circumference for age
(z-score):
12 month:-0.132 (0.25)
Note: When examining second
trimester maternal Pb > 3 ug/dL,
associations were observed for
6 mo weight for age, 6 mo weight
for length, 6 and 12 mo head
circumference, and 12 mo upper
arm circumference for age
Infant sex, infant birth
weight, infant nutrition,
maternal age, marital
status, employment, race,
height, parity, second
trimester smoking, and
education.
4-614
-------
Table 4-39 (Continued): Summary of recent epidemiologic studies of associations between Pb biomarker levels and postnatal
growth.
Reference
Study Population
Study Location
Study Years
Methodological
Details
Pb Biomarker Data
Mean (SD)
Adjusted Effect Estimates
Potential Confounders
Adjusted for in Analysis
Lamb et al.
(2008)
N=309 mother
child pairs
Kosovo,
Yugoslavia
1985-1986
Longitudinal
cohort study
using linear
regression
Maternal blood Pb measured mid-
pregnancy
Pristina: 5.60 (1.99) ug/dL
Mitrovica: 20.56 (7.38) ug/dL
Regression coefficients (95% Cl)
relating maternal blood Pb:
To Height (cm):
Pristina
1 yr:-0.61 (-2.24, 1.03)
4yr: 0.79 (-1.71, 3.29)
6.5 yr: 0.15 (-2.43, 2.74)
10yr:-0.09 (-3.69, 3.52)
Mitrovica
1 yr:-0.30 (-2.55, 1.96)
4 yr:-0.72 (-3.26, 1.82)
6.5 yr:-1.87 (-4.38, 0.64)
10yr:-2.87 (-6.21, 0.47)
Infant sex, ethnicity,
parity, maternal height or
maternal BMI, maternal
education, gestational age
at delivery, gestational
age at blood sample,
HOME score
To BMI (kg/m2):
Pristina
1 yr: 0.61 (-0.28, 1.50)
4yr: 0.17 (-0.67, 1.00)
6.5 yr: 0.61 (-0.09, 1.30)
10yr:-0.49 (-1.45, 0.46)
Mitrovica
1 yr: 0.23 (-0.84, 1.30)
4yr: 0.16 (-0.66, 0.98)
6.5 yr:-0.12 (-0.90, 0.66)
10 yr: 1.31 (-0.95, 3.57)
4-615
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Table 4-39 (Continued): Summary of recent epidemiologic studies of associations between Pb biomarker levels and postnatal
growth.
Reference
Study Population
Study Location
Study Years
Methodological
Details
Pb Biomarker Data
Mean (SD)
Adjusted Effect Estimates
Potential Confounders
Adjusted for in Analysis
Ignasiak et al. schoolchildren
(2006) ages 7-1 Syr
n=463 boys
n= 436 girls
South-western
Poland
1995
(Industrial area
with Cu smelters
and refineries)
Cross-sectional
study using
stepwise multiple
regression
analysis
Concurrent blood Pb
7.7 (3.5) ug/dl_
Estimated decrement per
10 ug/dL increase in blood Pb (p-
value)
Weight:
Boys: 2.8 kg (0.002)
Girls: 3.5 kg (0.007)
Height:
Boys: 3.2 cm (0.10)
Girls: 4.0 cm (0.001)
Trunk length:
Boys: 1.2 cm (0.02)
Girls: 1.1 cm (0.0001)
Leg length:
Boys: 2.1 cm (0.002)
Girls: 2.9 cm (0.0001)
Arm length:
Boys: 1.8 cm (0.0001)
Girls: 1.9 cm (0.008)
Age, age , education
Mauser et al.
(2008)
N=489 boys ages
8-9 yr
Chapaevsk,
Russia
May 2003 - May
2005
Cross-sectional
study using
multiple linear
regression
Concurrent blood Pb
Median (25th-75th percentile)
3 (2-5) ug/dl_
Regression coefficient (95% Cl)
Height (cm):
-1.439 (-2.25,-0.63)
Weight (kg):
-0.761 (-1.54, 0.02)
BMI (kg/m2):
-0.107 (-0.44, 0.23)
Birth weight, gestational
age, age at exam
4-616
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Table 4-39 (Continued): Summary of recent epidemiologic studies of associations between Pb biomarker levels and postnatal
growth.
Reference3
Little et al.
(2009)
Min et al.
(2008b)
Study Population
Study Location
Study Years
n=196 (1980s)
n-169 (2002)
ages 2-12 yr
Dallas, Texas
1980-1989 and
2002
n=62 boys
n= 46 girls
ages 5-13 yr
Seoul, South
Korea
Date(s) not
specified
Methodological Pb Biomarker Data
Details Mean (SD)
Cross-sectional Concurrent blood Pb
study using 1980s: 23.6 (1.3 SE) ug/dl_
SoOA,and 2002: 1.6 (0.2 SE)ug/dL
regression
models
Cross-sectional Concurrent blood Pb
study using 2.4 (0.7) ug/dl_
multiple linear
regression
Adjusted Effect Estimates
Changes in mean scaled
measure per 10 ug/dL Pb
increase (95%CI):
Height (cm):
-2.1 (-1.9, -2.3)
Weight (kg):
-1.9 (-1.7, -2.1)
BMI (kg/m2):
-0.5 (-0.4, -0.7)
Linear model estimate (SE; p-
value)
Height (cm):
-1.449(0.639; p=0.026)
Total arm length (cm):
-1.804(0.702; p=0.012)
Body weight (kg):
-0.646(0.718; p=0.370)
BMI (kg/m2):
-0.006 (0.272; p=0.982)
Potential Confounders
Adjusted for in Analysis
Age, age2, sex and cohort
effect
Age, sex, and father's
education
4-617
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Table 4-39 (Continued): Summary of recent epidemiologic studies of associations between Pb biomarker levels and postnatal
growth.
Reference
Study Population
Study Location
Study Years
Methodological
Details
Pb Biomarker Data
Mean (SD)
Adjusted Effect Estimates
Potential Confounders
Adjusted for in Analysis
Sanna and
Vallascas
(2011)
N = 825 children
ages 11-14 yr
Sardinia, Italy
Data collected in
1998, 2002 and
2007
Cross-sectional
study using
multiple
regression
analysis
Pb in hair
1998: 5.84 (6.56) ug/g
2002: 1.49(1.72) ug/g
2007: 0.78 (0.93) ug/g
(3 log Pb (p-value)
Height (mm)
1998:-0.121 (0.0021)
2002:-0.115 (0.0349)
2007: 0.011 (0.8665)
Sitting Height (mm)
1998:-0.117 (0.0017)
2002:-0.036 (0.5149)
2007: 0.028 (0.6633)
Age, sex
Estimated Lower Limb Length
(mm)
1998:-0.103 (0.0209)
2002:-0.164 (0.0057)
2007: -0.008 (0.9058)
Zailina et al.
(2008)
N = 269 children
ages 6.5-8.5 yr
n=169 urban
n=100 industrial
Kuala Lumpur,
Malaysia
Cross-sectional
study using
correlations
Concurrent blood Pb
Industrial: 3.75 ug/dL
Urban: 3.56 ug/dL
Correlation with blood Pb (p-
value):
Height for age:
Urban: -0.095 (0.219)
Industrial: -0.037 (0.716)
Weight for age:
None
Urban: 0.019(0.806)
Industrial: -0.063 (0.535)
Weight for height:
Urban: 0.136(0.079)
Industrial: -0.069 (0.493)
Left arm circumference:
Urban: 0.041 (0.595)
Industrial: -0.055 (0.587)
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Table 4-39 (Continued): Summary of recent epidemiologic studies of associations between Pb biomarker levels and postnatal
growth.
Reference
Study Population
Study Location
Study Years
Methodological
Details
Pb Biomarker Data
Mean (SD)
Adjusted Effect Estimates
Potential Confounders
Adjusted for in Analysis
Tomoum et al. n=41 boys and
(2010) girls ages
10-1 Syr
Cairo, Egypt
Jan-Jun 2007
Cross-sectional
study using t-test
or Mann-Whitney
U-test
Concurrent blood Pb
9.46 (3.08) ug/dl_
Percentage of the median (SD):
Pb<10ug/dL
Weight:
Boys: 127.56(16.26)
Girls: 114.8(10.8)
Height:
Boys: 98.06(3.19)
Girls: 96.75(2.91)
None
Pb>10ug/dL
Weight:
Boys: 122.0(16.71)
Girls: 123.11 (12.52)
Height:
Boys: 99.5 (5.04)
Girls: 100.33(4.53)
p-value for all comparisons >0.05
Olivero-Verbel
et al. (2007)
N=189 children
ages 5-9 yr
Cartagena,
Columbia
Jun-Aug 2004
Cross-sectional
study using
Spearman
correlations
Concurrent blood Pb
5.49 (0.23) ug/dl_
Spearman correlation coefficient
(p-value) between blood Pb and
Body size (cm):
-0.224 (0.002)
Weight (kg):
-0.126(0.087)
*no significance in partial
correlation between blood Pb
and size when controlled for age:
-0.096(0.189)
None
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Table 4-39 (Continued): Summary of recent epidemiologic studies of associations between Pb biomarker levels and postnatal
growth.
Reference
Study Population
Study Location
Study Years
Methodological
Details
Pb Biomarker Data
Mean (SD)
Adjusted Effect Estimates
Potential Confounders
Adjusted for in Analysis
Liu et al.
(2011 b)
N=303
ages 3-7 yr
Guiyu, China
Chendian, China
Jan-Feb 2008
Cross-sectional
study using
sample t-tests
Concurrent blood Pb
Guiyu: 13.2 (4.0-48.5) ug/dl_
Chendian: 8.2 (0-21.3) ug/dl_
Median (range)
Mean chest circumference
among girls (cm):
<10ug/dL: 50.31 +/-3.22 cm
> 10 ug/dL: 49.03+/-2.27 cm
(p-value <0.05)
Mean chest circumference
among children >6 years old
(cm):
<10 ug/dL: 51.70 +/- 3.35 cm
> 10 ug/dL: 52.87+7-2.49 cm
(p-value <0.05)
None
Mean head circumference
among children >6 years old
(cm):
<10 ug/dL: 48.71 +/-1.66cm
> 10 ug/dL: 50.04+/-1.29 cm
(p-value <0.01)
Mahram et al.
(2007)
n=42 boys
n= 39 girls
n=45 exposed
n=36 controls
ages 7-11 yr
Zanjan province,
Iran
Cross-sectional
study using t-
tests
Concurrent blood Pb
Area with Pb mining: 37.0
(24.7) ug/dL
Area without Pb mining: 15.6
(13.4)ug/dL
Comparison of children living in
areas with a Pb mine to children
living in areas without a Pb mine:
Height (cm),
standardized for age:
p-value 0.52
Weight (kg),
standardized for age:
p-value 0.8
None
Date(s) not
specified
"Studies are presented in order of first appearance in the text of this section, which is based on study design and methodological strength.
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Results from recent epidemiologic studies of postnatal growth are summarized in Table
4-39. Longitudinal epidemiologic studies have had inconsistent findings regarding the
association between Pb levels and post-natal growth. Afeiche et al. (2011) conducted a
longitudinal study of children in Mexico City, born between 1994 and 2005. Maternal
bone Pb during pregnancy was associated with a statistically significant decrease in
weight at age 5 years in girls but not in boys. The findings were robust to additional
adjustment for child's concurrent blood Pb level. A study in New York reported an
inverse association between maternal blood Pb during the second trimester of pregnancy
and various measures of growth, especially among those mothers with blood Pb levels of
at least 3 (ig/dL (Schell et al., 2009). These associations did not persist for those with
maternal blood Pb levels less than 3 (ig/dL. Among infants, age 6 month blood Pb levels
were not associated with measures of growth at age 12 months. In comparisons of
changes in blood Pb levels over time, high maternal blood Pb combined with low
12 month blood Pb among infants (indicating a decrease in blood Pb over time) resulted
in the greatest growth, even compared to those with both low, or both high, maternal and
infant blood Pb measures. In a prospective study of 309 mother-child pairs from
Yugoslavia, the relationship between maternal blood Pb measured mid-pregnancy and
attained height in children was investigated in those living in a highly exposed town with
a smelter and battery plant and those living in a relatively lower exposed town (Lamb et
al., 2008). In multivariate adjusted regression models, neither attained height (at birth, 1,
4, 6.6, or 10 years of age) nor rate of height change per month (at birth-1 year, 1-4 years,
4-6.5 years, 6.5-10 years of age) was associated in a consistent direction with maternal
pregnancy blood Pb levels in either the industrial or less exposed town. Weight was also
not associated with maternal blood Pb in this study.
Multiple cross-sectional studies reported an association between Pb levels and impaired
growth. Ignasiak et al. (2006) studied school children aged 7-15 years living close to
Cu smelters and refineries in Poland to assess the impact of Pb exposure on their growth
status. There was a statistically significant linear relationship between concurrent blood
Pb and reduced weight, height, trunk, leg and arm lengths. This decrease in height was
more influenced by decreases in leg length than trunk length. These results also indicated
that there was attenuation in osteoblast activity associated with higher blood Pb levels,
consistent with animal toxicological studies (Long etal. 1990). Hauser et al. (2008)
investigated the relationship between blood Pb and height in boys living in Chapaevsk,
Russia, an area contaminated with multiple pollutants including dioxins and metals. In a
multivariate adjusted regression analysis, height significantly decreased with increasing
blood Pb. Statistically nonsignificant decreases in weight and BMI were also observed.
The association of blood Pb with height, weight, and BMI was examined among two
cohorts of children living near Pb smelters in Texas (Little et al.. 2009). The first cohort
included children 2-12 years old in 1980 and the second cohort included children of the
4-621
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same age in 2002 when blood Pb levels were substantially lower. Decreases in height,
weight, and BMI with increasing blood Pb levels were observed among children in both
cohorts and increases in height and weight were observed comparing children from the
2002 cohort to those from the 1980 cohort. In a study with Korean children, Min et al.
(2008b) observed that height and total arm length decreased significantly with increasing
blood Pb in multivariate adjusted regression models. A statistically nonsignificant
decrease in body weight was observed with increasing blood Pb while no effect on BMI
was reported. In a study of children in Sardinia Italy, Sanna and Vallascas (2011)
measured Pb in hair at three points in time (1997, 2002, and 2007) and reported cross-
sectional results from regression analyses for each of these time periods. Pb in hair
decreased over time and significant associations of Pb in hair with height were observed
only in earlier time periods when hair Pb levels were relatively high. However, Pb in hair
samples is not a well-characterized biomarker (see Chapter 3 and Section 3.3.4.2).
Contrary to the results summarized above, several cross-sectional studies do not observe
associations between blood Pb levels and impaired growth. In a study with a similar
design, Zailina et al. (2008) studied the relationship of blood Pb and height in 7 year-old
Malaysian school children comparing those attending two schools in an urban setting to
those attending a school near an industrial area. After adjustment for age, no statistically
significant associations between concurrent blood Pb and physical development were
observed. Tomoum et al. (2010) investigated the association between blood Pb and height
in pubertal children in Cairo, Egypt. Neither boys nor girls with concurrent blood Pb
levels >10 (ig/dL differed significantly in height or weight when compared to those with
blood Pb <10 (ig/dL. In a simple correlation analysis of children aged 5-9 years in
Colombia, Olivero-Verbel et al. (2007) reported that concurrent blood Pb levels were
negatively associated with body size (r = -0.224, p <0.002). However, when a partial
correlation analysis was performed controlling for age, the association between blood Pb
and body size was no longer statistically significant. In a study of school children in
China, chest and head circumference were found to differ between high (>10(ig/dL) and
low concurrent blood Pb level groups; however, the direction of the difference was not
consistent (Liu et al.. 20 lib). Among girls, in comparison of those with high and low
blood Pb levels, a reduction in head circumference was observed. Among children greater
than 6 years of age, those with higher blood Pb levels were reported to have greater head
and chest circumferences. In a study of children aged 7-11 years and living in an area of
Iran with or without Pb smelters, age-standardized weight and height did not vary by
study area (Mahram et al.. 2007).
Evidence from previous toxicological studies has shown an association between
gestational Pb exposure and impaired postnatal growth (U.S. EPA. 2006b). Recent
toxicological studies report significant changes in postnatal or adult body weight after Pb
4-622
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exposure during different developmental windows. Masso-Gonzalez and Antonio-Garcia
(2009) found Pb-induced decreased body weights at weaning (PND21) in rat pups from
dams exposed to Pb during pregnancy and lactation (drinking water, 300 mg/L). Blood
Pb level in the control group was 1.43 ug/dL; in the Pb group it was 22.8 ug/dL. Dong et
al. (2009) reported decreased body weight in adult Kunming mice after exposure to
6,000 ppm Pb acetate in drinking water for 8 weeks. In contrast, Leasure et al. (2008)
reported a statistically significant inverse relationship between Pb dose and body weight
for male mice exposed to lower (27 ppm), moderate (55 ppm) and higher levels
(109 ppm) levels of Pb during gestation and lactation (dam drinking water, 2 weeks
before mating, through gestation and to PND10) with those exposed to the lowest dose
having the highest adult body weight among the overweight Pb-exposed animals. Male
mice exposed to 27 and 109 ppm Pb concentrations during gestation were 26% and 13%
heavier than were controls at 1 year of age, respectively. In this study, dams were
administered 27 ppm (lower), 55 ppm (moderate), and 109 ppm (higher) Pb in drinking
water beginning 2 weeks prior to mating and going through gestation and lactation to
PND10, which resulted in respective blood Pb levels from 10 ug/dL or less in the
lower-exposure offspring to 42 ug/dL in the higher-exposure offspring at PND10.
Leasure et al. (2008) also exposed a separate group of mice to Pb only during the
postnatal period (PNDO-PND21, lactation only exposure), and mice exposed to the same
aforementioned lower or higher dose of Pb did not exhibit a difference in body weight
when compared to control offspring. Wang et al. (2009e) observed a statistically
significant decrease in fetal body weight and body length of Wistar rats at GD20 after
maternal exposure to 250 ppm Pb acetate during gestation days 1-10, 11-20, or 1-20.
Also, associations were reported between elevated maternal blood Pb levels (0.6, 1.3, or
1.74 (iM, respectively or -12.4, 26.9, or 36.0 (ig/dL, respectively) compared to control
(0.04 (iM or -0.83 (ig/dL) and decreased pup body length and placental weight in Wistar
rats at GD20. The greatest decrease in fetal body weight and length was observed in the
group exposed to Pb during gestation days 1-20 followed by the group exposed to Pb
during gestation days 11-20. Teijon et al. (2006) observed reductions in birth weight of
litters administered 200 ppm or 400 ppm Pb acetate in drinking water (Wistar rats, Pb to
dams from GDI through lactation, and to 1 and 3 months postweaning to pups), but
found that this effect did not persist in the postnatal growth of the rats.
Notably, previous toxicological studies observed reductions in postnatal weight as well as
birth weight after exposure to Pb, albeit often at higher concentrations of Pb exposure.
Ronis et al. (2001; 1998a; 1998b; 1996) have published a series of papers reporting on
experiments exposing rats to Pb over different developmental windows, showing
associations between Pb exposure and deficits in growth. Sprague-Dawley rats with
lifetime Pb exposure to 6,000 ppm Pb acetate in drinking water (gestational-termination
of experiment Pb exposure, maximum blood Pb of 316 (ig/dL in males and 264 (ig/dL in
4-623
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females) had sex-independent pre-pubertal growth suppression, male-specific suppression
of pubertal growth and loss of growth effects postnatally but still maintained an overall
decreased body size out to PND60 due to earlier deficits. In a follow up study using the
same exposure duration with a dose of 4,500 ppm Pb acetate (resulting in blood Pb of
263 (ig/dL at PND85) yielded the same results (Ronis etal. 1996) with mechanistic
insight showing decrements in insulin-like growth factor 1 (IGF1) accompanying the
decreases in growth rates.
In summary, the body of toxicological literature on postnatal growth with Pb exposure
indicates that Pb exposure can induce decrements in both height/body length and BW that
may be persistent and differ by sex. However, findings from epidemiologic studies of
postnatal growth are not consistent. Many of these studies were limited by their cross-
sectional design. A few studies used longitudinal cohorts and controlled for multiple
potential confounders, such as age and parity, but the results of these studies are
inconsistent. Animal toxicology studies give insight to mechanistic changes that may
contribute to this Pb-induced decrement and to the windows of exposure that may
contribute greatest to these decrements.
4.8.1.4 Toxicological Studies of Other Developmental Effects
Developmental Effects on Blood and Liver
The 1986 and 2006 Pb AQCDs [(U.S. EPA. 1986b) and (U.S. EPA. 2006b)1 reported
studies that suggest Pb may alter hematopoietic and hepatic function during development.
Some recent studies provide evidence that support these findings; however recent results
are not consistent among the studies.
Masso et al. (2007) reported a decrease in liver weights of pups born to dams that
consumed 300 mg/L Pb in drinking water during gestation and lactation. They also
reported an increase in the number of erythrocytes; however the erythrocyte size was
diminished by 62%. Pb produced microcytic anemia as evidenced by decreased
hemoglobin content and hematocrit values without changes in mean corpuscular
hemoglobin (MCH) concentration. Alkaline phosphatase (ALP) activity, CAT activity, or
thiobarbituric acid reactive substances (TEARS) production did not change in pups at
PND 0, but increased statistically significantly by PND21 indicating reactive oxygen
generation. No change in acid phosphatase (ACP) activity was observed in the livers of
pupsatPNDOorPND21.
4-624
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Masso-Gonzalez and Antonia-Garcia (2009) reported normochromic and microcytic
anemia and a significant decrease in hematocrit values and blood 5-aminolevulinic acid
dehydratase (ALAD) activity (90% reduction) in pups from dams administered 300 mg/L
Pb acetate in drinking water during gestation. The authors also reported that erythrocyte
osmotic fragility was four times greater in Pb-exposed pups than in control pups.
Masso-Gonzalez and Antonia-Garcia (2009) reported increases in TEARS and CAT
activity in the liver after Pb exposure. Intoxication with Pb also resulted in decreased
liver protein concentrations and Mn-dependent SOD activity. Abnormalities in liver
function were further exemplified by increases in liver concentrations of ALP and ACP.
Teijon et al. (2006) observed that gestational exposure to Pb caused a decrease in
erythrocytes, hemoglobin, and MCH at weaning; however, by 1 and 3 months
postweaning, these parameters had returned to normal values. The authors observed a
slight increase in serum ALP, alanine aminotransferase (ALT), and aspartate
aminotransferase (AST) levels after Pb exposure in the absence of liver histological
changes.
Pb-induced effects on SOD activity in the liver of fetuses after Pb intoxication was
supported by a study by Uzbekov et al. (2007). The authors reported an initial increase in
SOD activity in livers of pups exposed to 0.3 mg/L and 3.0 mg/L Pb nitrate in drinking
water during gestation for 1 month (mean daily consumption 27 (ig/kg). In contrast, long-
term exposure (5 months) to the same concentrations of Pb nitrate concentration during
gestation resulted in decreased hepatic SOD activity.
Effects on hepatic Phase I and Phase II enzymes after early developmental exposure of
offspring to Pb during gestation and lactation was evaluated by Pillai et al. (2009). In the
study, pregnant Charles Foster rats were administered 0.05 mg/kg body weight Pb
subcutaneously throughout gestation until PND21. Pups were evaluated on PND56.
Results of the study show that Phase I xenobiotic-metabolizing enzymes (NADPH- and
NADH cytochrome c reductase) and Phase II xenobiotic- and steroid-metabolizing
enzymes (5-glutamyl transpeptidase, UDPGT, glutathione-s-transferase, and 17(3-
hydroxysteroid oxidoreductase) were reduced in both male and female pups by PND56.
Only inhibition in glutathione-s-transferase and 17(3-hydroxysteroid oxidoreductase
activities demonstrated a sex-specific pattern (glutathione-s-transferase inhibition in
males; 17|3-hydroxysteroid oxidoreductase inhibition greater in females). Observed
Pb-induced histological changes included massive fatty degeneration in hepatocytes,
large vacuoles in cytoplasm, appearance of pyknotic nuclei, and infiltration of
lymphocytes in the liver. Activities of antioxidant enzymes (SOD, CAT, glutathione
peroxidase, and glutathione reductase) were also reduced after Pb intoxication.
Alterations in biochemical parameters included decreased DNA, RNA, and cholesterol
4-625
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content, although it was not clear whether these changes were related to genetic
expression of xenobiotic-metabolizing enzymes or changes in steroid hormone
homeostasis.
Developmental Effects on Skin
The 2006 Pb AQCD (U.S. EPA. 2006b) reported a study that demonstrated Pb-induced
abnormalities in skin development. No current studies were identified that addressed
Pb-induced skin alterations.
Developmental Effects on the Retina
The 2006 Pb AQCD (U.S. EPA. 2006b) concluded that Pb exposure during early
postnatal development (resulting in blood Pb levels ~20 (ig/dL) impaired retinal
development in female Long-Evans hooded rats. A more recent study (Fox et al., 2008)
exposed female Long-Evans hooded rats to lower (27 ppm), moderate (55 ppm), and
higher (109 ppm) levels of Pb acetate in drinking water beginning 2 weeks before mating,
throughout gestation, and until PND10. Blood Pb levels measured in these pups on
postnatal days 0-10 were 10-12 (ig/dL (lower), 21-24 (ig/dL (moderate), and 40-46 (ig/dL
(higher). Results of the study demonstrated supernormal persistent rod photoreceptor-
mediated (scotopic) electroretinograms (ERGs) [(Fox et al., 2008). and Table 4-131 in
adult rats similar to ERG findings in male and female children in association with
maternal first trimester blood Pb levels 10.5-32 (ig/dL [(Rothenberg et al., 2002b). and
Table 4-131. In rats, lower and moderate levels of Pb increased neurogenesis of rod
photoreceptors and rod bipolar cells without affecting Miiller glial cells and statistically
significantly increased the number of rods in central and peripheral retina. Higher-level
Pb exposure (109 ppm) statistically significantly decreased the number of rods in central
and peripheral retina. Pb exposure induced concentration-dependent decreases in adult rat
retinal dopamine synthesis and utilization/release.
Developmental Effects on Teeth
Pb has been associated with multiple health effects including dental caries; however,
there is very limited information available on the temporal and spatial incorporation of Pb
in dental tissue (Arora et al. 2005). Arora et al. (2005) demonstrated that Wistar rat pups
exposed to Pb during gestation and lactation (40 mg/L of Pb nitrate in drinking water of
pregnant dams) had higher concentrations of Pb on the surface of enamel and in the
dentine immediately adjacent to the pulp.
4-626
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4.8.2 Effects on Birth Outcomes
The 2006 Pb AQCD reported on multiple studies of adverse birth outcomes such as, fetal
mortality, birth defects, preterm birth, and low birth weight/fetal growth (U.S. EPA.
2006b). The toxicological studies reviewed in the 2006 Pb AQCD concluded that Pb
exposure can increase fetal mortality and produce sublethal effects, smaller litters, and
fewer implantation sites. Epidemiologic studies using occupational histories reported the
possibility of small associations between increased Pb exposure and birth defects, and
toxicological studies demonstrated associations between exposure to high doses of Pb
and increased incidences of teratogenic effects in experimental animals. Epidemiologic
studies on preterm birth and low birth weight/fetal growth included in the
2006 Pb AQCD reported inconsistent findings. Evidence from previous toxicological
studies has shown an association between gestational Pb exposure and reduced birth
weight and decreased litter size or number of pups. Continued research on adverse birth
outcomes is described in the sections that follow. The studies presented in the following
text and tables are grouped by study design and methodological strength.
4.8.2.1 Infant Mortality and Embryogenesis
No recent epidemiologic or toxicological studies have reported on the relationship
between Pb levels and infant mortality. The 2006 Pb AQCD (U.S. EPA. 2006b)
concluded that Pb exposure can increase fetal mortality and produce sublethal effects
(disrupt growth and development) in offspring of Pb exposed dams at concentrations that
do not result in clinical toxicity to the dams by disrupting implantation and pregnancy,
particularly at the blastocyst stage of development. In rodent studies gestational exposure
to Pb (blood Pb levels 32 to >70 ug/dL) resulted in smaller litters and fewer implantation
sites and in non-human primates pre- and perinatal mortality was reported in squirrel
monkeys exposed to Pb (mean dam blood Pb level of 54 ug/dL) in the last two-thirds of
gestation (U.S. EPA. 2006b). There is substantial evidence to show that there is no
apparent maternal-fetal barrier to Pb and it can easily cross the placenta and accumulate
in fetal tissue during gestation (Pillai et al., 2009; Wang et al., 2009e; Uzbekov et al.,
2007V
4.8.2.2 Birth Defects
The 2006 Pb AQCD (U.S. EPA. 2006b) reported the possibility of small associations
between high Pb exposure and birth defects with many of the epidemiologic studies using
occupational histories. Among the studies included in the 2006 Pb AQCD, a couple
4-627
-------
reported associations between parental exposure to Pb and neural tube defects (Irgens et
al.. 1998; Bound et al.. 1997). Recent studies also examined indicators of Pb exposure
and neural tube defects (Table 4-40). No other recent epidemiologic studies of Pb
exposure and birth defects were identified in the literature. No recent toxicological
studies were found that investigated Pb-induced changes in morphology, teratology
effects, or skeletal malformations of developing fetuses as a result of maternal Pb
exposure; however, in the 2006 Pb AQCD toxicological studies demonstrated
associations between exposure to high doses of Pb and increased incidences of
teratogenic effect in experimental animals.
4-628
-------
Table 4-40 Summary of recent epidemiologic studies of associations between
Pb exposure indicators and neural tube defects.
Reference3
Zeyrek et
al. (2009)
Brender et
al. (2006)
Huang et
al. (2011 b)
Study
Population
Study Location
Study Years
Infants with
gestational age
of at least
20 weeks
NNTD=74
Ncontrols=70
Turkey
Years not
specified
Infants of
Mexican-
American
women
NNTD=184
NControls=225
Texas
1995-2000
Live and still
births of women
living in the
study area
(villages in the
Lvliang region of
Shanxi province)
Methodological
Details
Case-control
study using
Student's t-test
and Mann-
Whitney U-test
Case-control
study using
logistic
regression
Ecologic
Data for Pb Biomarker
or Exposure
Measures
Mean (SD)
Maternal and umbilical
cord blood Pb taken
0. Shatter birth
Cases:
Maternal:
15.5(1 5.0) ug/dl_
Umbilical cord:
18.2(1 7.8) ug/dl_
Controls:
Maternal: 12.5(12.7)
Umbilical cord:
16.5(16.1)
Maternal blood Pb
taken 5-6 weeks post-
partum
Cases: 2.4 (1.9) ug/dL
Controls:
2.5(1.6)ug/dL
2 soil samples from
each village
56.14ug/g(11.43ug/g)
Adjusted Effect
Estimates
P-values for
differences of
Student's t-test or
Mann-Whitney
U-test (dependent
on distribution) were
0.35 for maternal
blood Pb and 0.63
for umbilical cord
hlood Ph
UIUUU i U
OR (95% Cl):
Blood
Pb<6.0 ug/dl_:
1.0 (Ref)
Blood
Pb> 6.0 ug/dl_:
1.5(0.6,4.3)
N/A
Potential
Confounders
Adjusted for
in Analysis
None
Inclusion of
breast
feeding in the
model
changed the
OR(95%CI)
to 3.8 (0.8,
19.5)
N=112 villages
China
2002-2004
aStudies are presented in order of first appearance in the text of this section, which is based on study design and methodological
strength.
Among the recent epidemiologic studies (described in Table 4-40). a study of women in
Turkey detected no difference between the blood Pb of mothers or the umbilical cord
blood Pb of the newborns for healthy infants compared with infants with neural tube
defects (cases of spina bifida occulta were excluded, but other forms of spina bifida were
4-629
-------
included) (Zeyrek et al.. 2009). Brender et al. (2006) performed a study of Mexican-
American women living in Texas. Measurements were taken 5-6 weeks postpartum,
which is a limitation of this study because the blood Pb levels may be different from
those during the developmental period of gestation. The OR comparing women with at
least 6 (ig/dL blood Pb to those with less than 6 (ig/dL blood Pb was 1.5 (95% CI: 0.6,
4.3). This increased after adjusting for breast feeding, although this variable was not a
confounder because it cannot be associated with neural tube defects. For these women,
neither occupational exposure to Pb nor proximity of residence to a facility with Pb air
emissions at the time of conception was associated with increased odds of neural tube
defects. A study with an ecologic design was performed in China and did not use
individual-level biomarkers to determine Pb levels (Huang et al.. 201 Ib). A positive
association between Pb levels in soil samples and neural tube defects was reported.
Exposure to multiple other trace elements also demonstrated a positive association but no
control for co-exposures was included in the models for Pb.
In summary, previous studies included in the 2006 Pb AQCD observed associations
between Pb and neural tube defects but were based on Pb was measured in drinking water
(Bound etal.. 1997) and estimated from occupational reports (Irgens etal.. 1998). A
recent ecologic study reported an association between Pb in the soil and neural tube
defects but was also limited by its lack of biological samples, as well as a lack of
individual-level data and the prevalence of several other metals (Huang et al.. 201 Ib).
Other recent epidemiologic studies of maternal blood Pb levels and neural tube defects
observed no statistically significant associations (Zeyrek et al.. 2009; Brender et al..
2006). These studies also have limitations, including the timing of Pb measurements and
lack of control for potential confounders.
4.8.2.3 Preterm Birth
Epidemiologic studies on preterm birth included in the 2006 Pb AQCD (U.S. EPA.
2006b) reported inconsistent findings regarding the relationship between Pb and
gestational age. Recent studies examined this potential association, and again mixed
results were reported (Table 4-41). Of these studies, the ones that categorized births as
preterm or term all defined preterm birth as less than 37 weeks of gestation. One
limitation to note for these studies is that if Pb affects spontaneous abortion and length of
gestation via a similar pathway, then the studies that only collect data at delivery and not
at earlier stages of pregnancy would be biased toward the null.
4-630
-------
Table 4-41 Summary of recent epidemiologic studies of associations between Pb exposure indicators and
preterm birth.
Reference3 Outcome
Jelliffe- Preterm
Pawlowski birth
et al. (< 37
(2006) completed
week)
Study Population
Study Location
Years
Singleton births to
non-smoking
mothers with blood
Pb measures
during pregnancy
from either the
California
Childhood Lead
Poisoning
Prevention Branch
or the California
Occupational Lead
Poisoning
Prevention
Program
N preterm birth =30
N term birth =232
California
1995-2002
Data for Pb Biomarker
Methodological or Exposure
Details Indicators Adjusted Effect Estimates
Longitudinal Maximum maternal Odd Ratios:
cohort study blood Pb during < 5 Mg/dL. 1 Oo (Ref)
re^esSf ^^ 6-9 Mg/dL: 0.8 (0.1, 6.4)
10-19ug/dL: 1.1 (0.2, 5.2)
, 10 ug/dL 30.9o/o 20.39 pg/dL: 4.5 (1.8, 10.9)
>40ug/dL:4.7(1.1, 19.9)
<10ug/dL: 1.00 (Ref)
>10ug/dL: 3.2(1.2, 7.4)
Potential
Confounders
Adjusted for in
Analysis
In the 10ug/dL
dichotomized
model: race,
insurance,
maternal age,
parity, infant sex,
low birthweight
4-631
-------
Table 4-41 (Continued): Summary of recent epidemiologic studies of associations between Pb exposure indicators and
preterm birth.
Reference3
Outcome
Study Population
Study Location
Years
Methodological
Details
Data for Pb Biomarker
or Exposure
Indicators
Adjusted Effect Estimates
Potential
Confounders
Adjusted for in
Analysis
Vigeh et al.
(2011)
Preterm
birth
(20-37 wee
Singleton births
from non-smoking,
non-obese mothers
aged 16-35 yr and
referred for
prenatal care
during the
8th-12th week of
gestation
N preterm birth =44
N term birth =304
Tehran, Iran
2006
Longitudinal
cohort study
using logistic
regression
Maternal blood Pb at
8-12 weeks gestation
Mean (SD)
3.8(2.0)
Preterm birth: 4.52
(1.63)
Term birth: 3.72 (2.03)
p-value for difference:
<0.05
OR(95%CI): 1.41 (1.08, 1.84)
(unit not given, assume per 1 ug/dL)
Age, infant sex,
education, passive
smoking exposure,
pregnancy weight
gain, parity,
hematocrit, type of
delivery
4-632
-------
Table 4-41 (Continued): Summary of recent epidemiologic studies of associations between Pb exposure indicators and
preterm birth.
Study Population
Study Location
Reference3 Outcome Years
Cantonwine Preterm Births to mothers
et al. birth with at least 1
(201 Oa) (<37week), blood Pb
Gestational measurement
age during pregnancy
and no chronic
diseases requiring
medication
N preterm birth =22
N term birth =2 13
Mexico City
1997-1999
Data for Pb Biomarker
Methodological or Exposure
Details Indicators
Longitudinal Maternal blood Pb
cohort study during pregnancy
using linear Mean (SD)
regression „, , „,
a Blood Pb
Visit at <20 weeks
pregnant: 7.2 (5.2)
Visit at 20-28 weeks
pregnant: 6.3 (4.3)
Visit at >28 weeks
pregnant: 6.8 (4.5)
Plasma Pb
Visit at <20 weeks
pregnant: 0.17 (0.16)
Visit at 20-28 weeks
pregnant: 0.13 (0.10)
Visit at >28 weeks
pregnant: 0.16 (0.26)
Adjusted Effect Estimates
(3 (95% Cl) per SD increase in
centered log-Pb concentration
Blood Pb
Visit at <20 weeks: -2.76 (-5.21 ,
OO *1 \
.31)
Visit at 20-28 weeks: -1 .77 (-3.39,
-0.15)
Visit at >28 weeks: -0.47 (-1 .78, 0.84)
Average: -1.49 (-3.63, 0.64)
Plasma Pb
Visit at <20weeks: -2.38 (-4.97, 0.21)
Visit at 20-28 weeks: -1.34 (-2.98,
0.29)
Visit at >28 weeks: -1.28 (-2.63, 0.06)
Average: -0.28 (-2.81, 2.25)
Potential
Confounders
Adjusted for in
Analysis
Infant sex,
maternal age,
maternal
education, history
of adverse birth
outcomes,
cigarette smoking,
parity
Plasma-to-blood Pb ratio
Visit at <20 weeks: -3.23 (-6.01,
-0.44)
Visit at 20-28 weeks: -1.41 (-3.10,
0.29)
Visit at >28 weeks: -1.30 (-2.67, 0.07)
Average: -1.27 (-3.89, 1.35)
Cord blood Pb -0.68 (-2.37, 1.00)
4-633
-------
Table 4-41 (Continued): Summary of recent epidemiologic studies of associations between Pb exposure indicators and
preterm birth.
Reference3
Outcome
Study Population
Study Location
Years
Methodological
Details
Data for Pb Biomarker
or Exposure
Indicators
Adjusted Effect Estimates
Potential
Confounders
Adjusted for in
Analysis
Zhu et al.
(2010)
Preterm
birth (<37
completed
week)
Singleton births to
mothers aged
15-49 yr with blood
Pb measures
before or on the
date of delivery and
blood Pb
measuring
<10ug/dL
N preterm birth=351 9
N term birth =39,769
New York
2003-2005
Retrospective
cohort study
using logistic
regression with
fractional
polynomials
Maternal blood Pb
Mean: 2.1
Odd Ratios:
<1.0ug/dL: 1.00(Ref)
1.1-2.0 ug/dl_: 1.03(0.93, 1.13)
2.1-3.0 ug/dl_: 1.01 (0.92, 1.10)
3.1-9.9 ug/dl_: 1.04(0.89, 1.22)
Timing of Pb test,
maternal age,
race, Hispanic
ethnicity, smoking
status, drug
abuse, marital
status, special
financial program
participation,
parity, infant sex
4-634
-------
Table 4-41 (Continued): Summary of recent epidemiologic studies of associations between Pb exposure indicators and
preterm birth.
Reference3 Outcome
Chen et al. Preterm
(2006a) birth
(<37 week)
Patel and Gestational
Prabhu age
(2009)
Study Population
Study Location
Years
Infants born to at
least one parent
who was part of the
Program to Reduce
Exposure by
Surveillance
System - Blood
Lead Levels cohort
that monitored
workers
occupationally
exposed to Pb
N preterm birth =74
N term birth = 1537
*738 births had
maternal Pb
information and
967 had paternal
Pb information
Taiwan
1993-1997
Consecutive births
at the study
hospital
Data for Pb Biomarker
Methodological or Exposure
Details Indicators
Occupational Maternal blood Pb
cohort study during pregnancy (or if
using regression that wasn't available,
models the 1 year prior to
fertilization) and/or
paternal blood Pb
during spermatogenesis
(the 64 days before
fertilization, or if that
wasn't available, the 1
year prior to
spermatogenesis)
Mean (SD)
Maternal blood Pb:
10.1 (10.4)
Paternal blood Pb:
12.9(13.8)
Cross-sectional Umbilical cord blood Pb
study using linear Mean (SD): 4 7 (1Z1)
regression
Potential
Confounders
Adjusted for in
Adjusted Effect Estimates Analysis
Risk Ratios Parental age,
parental
„„ x , , , , r,, education, parity,
Maternal blood Pb infant sex
<10ug/dL: 1.00(Ref)
10-19ug/dL: 1.97(0.92, 3.86)
>20ug/dL: 1.86(0.68,4.28)
Paternal blood Pb
<10ug/dL: 1.00(Ref)
10-19ug/dL: 1.17(0.53,2.32)
>20ug/dL: 0.55(0.19, 1.28)
>5 ug/dL: mean gestational age Not specified
38 weeks
< 5 ug/dL: mean gestational age
N=205 (mean
gestational age 39
+/-2 weeks)
Nagpur, India
Years not specified
39 weeks
Linear regression: gestational age
decreased 1 week with every 1 ug/dL
increase in umbilical cord blood Pb
(exact values and 95% Cl: not given)
4-635
-------
Table 4-41 (Continued): Summary of recent epidemiologic studies of associations between Pb exposure indicators and
preterm birth.
Reference3
Jones et al.
(2010)
Wells et al.
(2011 a)
Outcome
Gestational
Age:
preterm
(<37week),
term
(37-40 wee
k), post-
term
(>40 week)
Gestational
age
Study Population
Study Location
Years
Singleton births
> 27 week
gestation from
mothers aged
16-45 yr living in
the Shelby County
area for at least
5 mo during
pregnancy
N preterm birth ~ 1 U
N term birth =81
N postterm birth ~1 1
Tennessee
2006
Singleton births
from the Baltimore
Tracking Health
Related to
Environmental
Exposures
(THREE) study
N preterm birth =39
N term birth =261
Baltimore, MD
2004-2005
Data for Pb Biomarker
Methodological or Exposure
Details Indicators Adjusted Effect Estimates
Cross-sectional Umbilical cord blood Pb Geometric Mean:
study comparing Mean (SD): 2.4 (4.3) Preterm birth: 1 .4
across geometric _ . . „„ _,_ ,..,._
means (test not Geometric mean: 1.3 Term birth: 1 .2
specified) Post-term birth: 1.3
p-value for difference: >0.10
Cross-sectional Umbilical cord Pb Ratio for Pb concentration per
study using Mean' 10 days of gestation: 0.99 (0.93,
multivariable g 1.06)
linear regression „„!!,
0.9b)
> 5 ug/dl_: 0.7%
Potential
Confounders
Adjusted for in
Analysis
None
Maternal age,
race, insurance,
pre-pregnancy
BMI, smoking
status, gestational
age, birthweight,
average year of
neighborhood
home construction
4-636
-------
Table 4-41 (Continued): Summary of recent epidemiologic studies of associations between Pb exposure indicators and
preterm birth.
Reference3 Outcome
Berkowitz Preterm
et al. birth
(2006) (<37week)
Study Population
Study Location
Years
Singleton births
with 28-45 week
gestation
-7 Q A O
preterm birth"' ,o4O
N term birth = 162, 035
Shoshone County,
ID
1970-1981
Methodological
Details
Cohort study
using logistic
regression
Data for Pb Biomarker
or Exposure
Indicators
Three time periods of
two locations
(unexposed and
exposed/near smelter):
pre-fire, "high-exposure
period" (when a fire
happened at the smelter
and resulted in
damages leading to
high air Pb
concentrations for 6
mo), and "post-fire"
Adjusted Effect Estimates
OR (90% Cl) (unexposed location is
referent group):
Pre-fire: 0.93(0.67, 1.28)
High exposure: 0.68 (0.34, 1.35)
Post-fire: 1.17(0.95, 1.45)
Potential
Confounders
Adjusted for in
Analysis
Maternal age,
infant sex, first
birth, previous
miscarriage or
abortion
During the time of the
fire, estimates of Pb in
ambient air were as
high as 30 ug/m3
Orun et al.
(2011)
Preterm
birth
(<37 week)
Births to mothers
not occupationally
exposed to toxic
metals and living in
a suburban but
non-industrial area
Cohort study
using Mann-
Whitney U-test
Breast milk 2 months
post-partum
Median: 20.6 ug/L
Median Pb(IQR)
>37 week: 20.6 (1 1 .2, 29.2) ug/L
< 37 week: 20.4 (14.4, 27.9) ug/L
None
N preterm birth" 17
N term birth = 127
>WHO limit (5 ug/L):
87%
p-value for Mann-Whitney U test:
>0.05
Turkey
Years not specified
"Studies are presented in order of first appearance in the text of this section, which is based on study design and methodological strength.
4-637
-------
In a study taking place in California, women with information on blood Pb levels during
pregnancy based on their participation in a surveillance program (reason for participation
in the surveillance program was unknown but the authors speculate it was likely because
of potential Pb exposure due to occupational or environmental exposures or a family
member was identified as exposed to Pb) were matched with the birth certificates of their
infants (Jelliffe-Pawlowski et al.. 2006). Almost 70% of women had maximum blood Pb
measurements <10 (ig/dL with just over 60%being <5 (ig/dL. Preterm birth was
associated with higher blood Pb when comparing women with maximum pregnancy
blood Pb levels > 10 (ig/dL to women with blood Pb levels <10 (ig/dL in adjusted
analyses. In analyses of maximum Pb levels further refined into additional categories, the
odds of preterm birth were elevated among women with maximum blood Pb
measurement > 20 (ig/dL compared with women with maximum blood Pb levels
< 5 (ig/dL. A study in Iran also reported higher maternal blood Pb for preterm births than
for term births (Vigeh et al., 2011). The women in this study had lower blood Pb levels
than did those observed in the Jelliffe-Pawlowski et al. (2006) study. Higher maternal
blood Pb level was associated with higher odds of preterm birth. Another study
examining blood Pb and gestational age reported an inverse association between maternal
blood Pb concentration and gestational age, especially for blood Pb levels early in
pregnancy (Cantonwine et al.. 2010a). However, in a study conducted in New York
among women with blood Pb levels less than 10 (ig/dL, no association was observed
between blood Pb levels and preterm birth (Zhu et al.. 2010). Similarly, a study of
maternal and paternal blood Pb concentrations reported no association between maternal
or paternal blood Pb levels and preterm birth (Chen et al.. 2006a).
In another study, measurements of umbilical cord blood were taken after birth at a
hospital in Nagpur, India (Patel and Prabhu. 2009). A sample of women had their blood
Pb measured and among this sample, maternal blood Pb was correlated with the umbilical
cord Pb levels. Mean gestational age differed between infants with >5 (ig/dL cord blood
Pb and infants with < 5 (ig/dL cord blood Pb. In a linear regression model, gestational
age was found to decrease with increasing umbilical cord Pb levels. A study of women in
Tennessee consisted primarily of African American women living in an urban setting
(Jones et al., 2010). The mean level of umbilical cord blood Pb was slightly higher
among infants born preterm but the difference was not statistically significant. Using
umbilical cord blood Pb measures, a study reported no association between cord blood Pb
levels and gestational age. The overwhelming majority (99.3%) of cord blood Pb
concentrations among study participants was < 5 (ig/dL (Wells et al., 201 la).
4-638
-------
A study of preterm birth included women living in two different residential areas over
three different time periods (Berkowitz et al.. 2006). One residential area had consistently
lower exposures but the other had a 6-month period of high Pb emissions due to damage
to emissions controls at a local Pb smelter (Pb measured in ambient air was up to
30 (ig/m3). Preterm birth rates were examined during three time periods: before, during,
and after the time of higher Pb exposure. No association with preterm birth was observed
between women living in the high exposure area compared to those in the low exposure
area during any of the exposure time periods, but the number of preterm infants born
during the period of higher exposure was small.
A study of breast milk in the second month postpartum reported no difference in breast
milk Pb levels for those infants born preterm or term; however, a limitation of this study
is that Pb levels were not measured until two months after the birth and may not reflect
maternal blood Pb levels during pregnancy (Orimet al.. 2011).
In summary, as in the 2006 Pb AQCD, (U.S. EPA. 2006b) recent epidemiologic studies
report inconsistent findings for a relationship between indicators of Pb exposure and
preterm birth. No patterns were apparent within type of exposure measurement or Pb
level. Many of these studies are limited by the small number of preterm births and their
cross-sectional design (i.e., studies of umbilical cord blood may not adequately
characterized blood Pb levels earlier in pregnancy).A few studies utilized a longitudinal
cohort design (Vigeh et al.. 2011; Cantonwine et al., 2010a; Chen et al., 2006a; Jelliffe-
Pawlowski et al.. 2006). and although results among these studies were mixed some did
report an association between maternal blood Pb during pregnancy and preterm birth.
Most studies controlled for potentially important confounders, such as maternal age and
smoking.
4.8.2.4 Low Birth Weight/Fetal Growth
The 2006 Pb AQCD reported inconsistent epidemiologic study results for the
associations between Pb and birth weight/fetal growth but concluded that there could be a
small effect of Pb exposure on birth weight and fetal growth (U.S. EPA. 2006b). Since
then, multiple epidemiologic studies on the relationship between Pb exposure and birth
weight and fetal growth have been published using various measures of exposure, such as
air levels, umbilical cord blood, and maternal blood and bone. These studies are
summarized in Table 4-42 (organized in the text and table by the type of Pb measurement
and then by study design). Additionally, there have been a few recent toxicological
studies evaluating the effect of Pb exposure during gestation on birth weight.
4-639
-------
Table 4-42 Summary of recent epidemiologic studies of associations between Pb exposure indicators and low
birth weight and fetal growth.
Reference3 Outcome
Gundacker Birth length,
et al. birth weight,
(2010) head
circumference
Study
Population
Study Location
Years
Infants of women
recruited during
their second
trimester
N=53
Vienna, Austria
2005
Data for Pb
Biomarkers or
Methodological other Exposure
Details Indicators
Cohort study Median (IQR):
using categorical Maternal blood Pb
regression between week 34-38
of gestation
2.5(1.8, 3.5)ug/dL
Whole placentas
25.8(21.0, 36.8)
ug/kg
Umbilical cord Pb
shortly after birth
1.3(0.8, 2.4) ug/dL
Meconium samples
in first five days after
birth
15.5(9.8,27.9)
ug/kg
Adjusted Effect Estimates
Regression coefficients (units not given,
assume results are per 10 ug/dL or
1 ug/kg)
Birth length:
Placenta Pb: 0.599 (SE: 0.154, p-value
<0.001)
Meconium Pb: -0.385 (SE 0.157, p-value
OI"M O\
.012)
Birth weight:
Placenta Pb: 0.658 (SE 0.136, p-value
<0.001)
Maternal blood Pb: -0.262 (SE 0.131, p-
value 0.058)
Potential
Confounders
Adjusted for in
Analysis
Model for birth
length: placenta
Pb, gestational
age, meconium
Pb
Model for birth
weight:
gestational age,
placenta Pb,
maternal blood
Pb
4-640
-------
Table 4-42 (Continued): Summary of recent epidemiologic studies of associations between Pb exposure indicators and low
birth weight and fetal growth.
Reference3 Outcome
Zhu et al. Birth weight,
(2010) small for
gestational age
(birth weight for
gestational age
<10th percentile
based on
national birth
weight by
gestational week
from weeks
25-42)
Study
Population
Study Location
Years
Singleton births to
mothers aged
1 5-49 yr with
blood Pb
measures before
or on the date of
delivery and
blood Pb
measuring
<10ug/dL
NLBw=2,744
N normal BW=40, 544
NSGA=4092
NnoSGA=39,084
New York
2003-2005
Data for Pb
Biomarkers or
Methodological other Exposure
Details Indicators
Retrospective Maternal blood Pb
cohort using before or at delivery
linear regression Mean: 2.1 ug/dl_
with factional
polynomials for
birth weight and
logistic
regression with
fractional
polynomials for
SGA
Adjusted Effect Estimates
Difference in birthweight in grams:
0 ug/dl_: Ref
1 ug/dl_: -27.4 (-37.8, -17.1)
2 ug/dl_: -38.8 (-53.4, -24.1)
3 ug/dl_: -47.5 (-65.4, -29.6)
4 ug/dl_: -54.8 (-75.5, -34.2)
5ug/dL:-61.3(-84.4, -38.2)
6 ug/dl_: -67.2 (-92.5, -41.8)
7 ug/dl_: -72.5 (-99.9, -45.2)
8 ug/dl_: -77.6 (-106.8, -48.3)
9ug/dL:-82.3(-113.3, -51.2)
10ug/dL: -86.7 (-11 9.4, -54.0)
After exclusion of blood Pb <1 ug/dL, a
1 ug/dL increase in blood Pb was
associated with a 7.0 g decrease in
birthweight
Odd Ratios for small for gestational age:
<1.0ug/dL: 1.00 (Ref)
1. 1-2.0 ug/dl_: 1.07(0.98, 1.17)
2.1-3.0 ug/dl_: 1.06(0.98, 1.16)
3.1-9.9 ug/dl_: 1.07(0.93, 1.23)
Potential
Confounders
Adjusted for in
Analysis
Model for birth
weight: Timing of
Pb test, maternal
age, race,
Hispanic
ethnicity,
education,
smoking status,
alcohol use,
drug abuse,
marital status,
financial
assistance
program
participation,
parity, infant sex
Model for SGA:
Timing of Pb
test, maternal
age, race,
education,
smoking status,
drug abuse,
marital status,
financial
assistance
program
participation,
parity, infant sex
4-641
-------
Table 4-42 (Continued): Summary of recent epidemiologic studies of associations between Pb exposure indicators and low
birth weight and fetal growth.
Reference3 Outcome
Chen et al. Low birth weight
(2006a) (<2,500 g),
small for
gestational age
(birth weight
< 10th percentile
of sex- and
gestational week
weights for
singletons in
1993-1996)
Study
Population
Study Location Methodological
Years Details
Infants born to at Occupational
least one parent cohort study
who was part of using regression
the Program to models
Reduce Exposure
by Surveillance
System - Blood
Lead Levels
cohort that
monitored
workers
occupationally
exposed to Pb
NLBw=72
N normal BW= 1,539
NSGA=135
NnoSGA = 1,476
*738 births had
maternal Pb
information and
967 had paternal
Pb information
Taiwan
1993-1997
Data for Pb
Biomarkers or
other Exposure
Indicators
Mean (SD)
Maternal blood Pb
during pregnancy (or
if that wasn't
available, the 1 year
prior to fertilization)
10.1 (10.4)
Paternal blood Pb
during
spermatogenesis
(the 64 days before
fertilization, or if that
wasn't available, the
1 year prior to
spermatogenesis)
12.9 (13.8)
Adjusted Effect
Risk Ratios
Low birth weight
Maternal blood Pb
<10ug/dL: lOO(Ref)
10-19 ug/dL: 2.22 (1.06,
>20ug/dL: 1.83(0.67,4
Paternal blood Pb
<10ug/dL: 1.00(Ref)
10-19 ug/dL: 0.83(0.34,
>20ug/dL: 0.42(0.12, 1.
O/^~> A
SGA
Maternal blood Pb
<10ug/dL: 1.00(Ref)
10-19 ug/dL: 1.62(0.91,
>20ug/dL:2.15(1.15, 3.
Paternal blood Pb
<10ug/dL: 1.00(Ref)
10-19 ug/dL: 0.94(0.49,
>20ug/dL: 0.94(0.51, 1
Estimates
4.26)
.20)
1.75)
06)
2.75)
83)
1.66)
.62)
Potential
Confounders
Adjusted for in
Analysis
Low birth weight
models: parental
age, parental
education, infant
sex, parity
SGA models:
parental age,
parental
education
4-642
-------
Table 4-42 (Continued): Summary of recent epidemiologic studies of associations between Pb exposure indicators and low
birth weight and fetal growth.
Reference3
Jelliffe-
Pawlowski
etal.
(2006)
Outcome
Low birth weight
(<2,500g)
Small for
gestational age
(birth weight for
gestational age
<10th percentile
of race- and
gender- specific
norms
Study Data for Pb
Population Biomarkers or
Study Location Methodological other Exposure
Years Details Indicators
Singleton births to Longitudinal Maximum maternal
non-smoking cohort study blood Pb during
mothers with using logistic pregnancy
blood Pb regression
preana'rtcv from9 ~ 10 V9/d\-: 30-9%
either the
California
Childhood Lead
Poisoning
Prevention
Branch or the
California
Occupational
Lead Poisoning
Prpvpntinn
1 1 C VGI 1 LIUI 1
Program and
matched to birth
records
NLBw=9
N normal BW=253
NSGA=17
NnoSGA=245
California
1995-2002
Adjusted Effect Estimates
Odd Ratios:
Low birth weight
<5ug/dL: 1.00(Ref)
6-9 ug/dL: not calculated (n=0)
10-19 ug/dL: 2.7 (0.5, 14.8)
20-39 ug/dL: 1.5(0.3, 7.7)
> 40 ug/dL: not calculated (n=0)
<10ug/dL: 1.00(Ref)
>10ug/dL: 3.6(0.3,40.0)
Small for gestational age
<5ug/dL: 1.00(Ref)
6-9 ug/dL: not calculated (n=0)
10-19 |jg/dL: 2.3 (0.6, 9.2)
20-39 ug/dL: 2.1 (0.7,6.7)
> 40 ug/dL: not calculated (n=0)
<10ug/dL: 1.00(Ref)
>10|jg/dL:4.2(1.3, 13.9)
Potential
Confounders
Adjusted for in
Analysis
10 ug/dL model
for birth weight:
preterm birth,
race, insurance,
parity, maternal
age, infant sex
10 ug/dL model
for SGA:
insurance,
parity, maternal
age
4-643
-------
Table 4-42 (Continued): Summary of recent epidemiologic studies of associations between Pb exposure indicators and low
birth weight and fetal growth.
Reference3 Outcome
Lamb et al. Height and BMI
(2008) at birth
Iranpouret Low birth weight
al. (2007) (<2,500g,
>37weeks)
Study
Population
Study Location
Years
Participants of the
Yugoslavia Study
of Environmental
Lead Exposure,
Pregnancy
Outcomes, and
Childhood
Development
N=292
Mitrovica and
Pristina,
Yugoslavia
1985-1986
Full-term infants
born at a hospital
affiliated with
Isfahan University
K I OO
Ni_Bw32
N normal BW=34
Isfahan, Iran
2005
Methodological
Details
Population-
based
prospective
cohort study
using linear
regression
Cross-sectional
study using t-
tests and
Spearman's
correlations
Data for Pb
Biomarkers or
other Exposure
Indicators
Mean (SD)
Mid-pregnancy
blood Pb
Mitrovica: 20.56
(7.38)
Pristina: 5.60(1.99)
Mean (SD)
Maternal blood Pb
within 12 h of
delivery
Cases: 12.5 (2.0)
[JCj/ClL
Controls: 13.5(2.7)
Mg/dL
Umbilical cord
Cases: 10.7(1.7)
Adjusted Effect Estimates
Regression Coefficients (95% Cl) for
1 Mg/dL increase in Pb:
BMI (kg/m2)
Mitrovica: -0.1 8 (-0.69, 0.33)
Pristina: -0.1 4 (-0.69, 0.42)
Height (cm)
Mitrovica: 0.43 (-0.83, 1.69)
Pristina: 0.35 (-0.64, 1.34)
P-values for t-tests:
Maternal blood Pb: 0.07
Umbilical cord blood Pb: 0.20
P-values for correlations:
Maternal blood Pb and Birth weight:
Low birth weight: 0.17
Normal birth weight: 0.3
Umbilical cord blood Pb and birth weight:
Potential
Confounders
Adjusted for in
Analysis
Infant sex,
ethnicity, parity,
maternal height
or BMI, maternal
education,
gestational age
at blood sample,
gestational age
at birth, quality
of home
environment
None
Controls: 11.3(1.9)
Low birth weight: 0.84
Normal birth weight: 0.26
4-644
-------
Table 4-42 (Continued): Summary of recent epidemiologic studies of associations between Pb exposure indicators and low
birth weight and fetal growth.
Reference3
Kordas et
al. (2009)
Outcome
Head
circumference,
birth weight,
birth length
Study
Population
Study Location Methodological
Years Details
Infants of mothers Cross-sectional
receiving study using
antenatal care at linear regression
hospitals serving
low-to-middle
income
populations
(cross-sectional
study of baseline
info from Ca
supplementation
trial)
Data for Pb
Biomarkers or
other Exposure
Indicators
Maternal tibia Pb
1 month post-partum
Mean (SD): 9.9
(9.8) ug/g
Maternal blood Pb
within 12 h of
delivery
>10ug/dL:27%
Umbilical cord
Adjusted Effect Estimates
Regression coefficients (SE) for each
1 ug/g increase in tibia Pb:
Birth weight: -4.9 (1.8)
Birth length: -0.02 (0.01)
Head circumference: -0.01 (0.01; p-value
<0.05)
Women in 4th quartile tibia Pb
(15.6-76.5 ug/g) delivered infants 140 g
less than women with tibia Pb in the lowest
quartile
Potential
Confounders
Adjusted for in
Analysis
Maternal age,
pre-pregnancy
BMI, maternal
height,
education,
parity, marital
status, ever
smoker,
postpartum calf
circumference,
gestational age,
infant sex
N=474
>10ug/dL: 13.7%
Mexico City,
Mexico
1994-1995
Afeiche et Birth weight Term, singleton Cross-sectional
al. (2011) births, at least study using
2,500 grams varying
enrolled in one of coefficient
three birth cohorts models with
recruited for other random effects
longitudinal
studies
N=1 000
IN — I , \J\J\J
Mexico City
1994-2005
Mean (SD)
measured 1 month
postpartum
Maternal patella Pb
10.4 (11. 8) ug/g
Maternal tibia Pb
8.7 (9.7) ug/g
(3 (95% Cl) for 1 SD increase in maternal
patella Pb
Girls: -45.7 (-131. 7, 40.2)
Boys: 72.3 (-9.8, 154.4)
No association for birth weight and tibia Pb
among girls. A positive association was
observed for tibia Pb and birth weight
among boys (quantitative results not given).
Birth cohort,
maternal age,
maternal calf
circumference,
maternal height,
education,
parity, breast
feeding,
Ca2+ treatment
group
assignment,
gestational age,
height at birth,
repeated
concurrent child
blood Pb
measures
4-645
-------
Table 4-42 (Continued): Summary of recent epidemiologic studies of associations between Pb exposure indicators and low
birth weight and fetal growth.
Reference3 Outcome
Cantonwine Birth weight
etal.
(2010b)
Wells et al. Birth weight
(2011 a)
Study
Population
Study Location
Years
Infants who were
part of a clinical
trial to assess
maternal Ca2+-
supplementation
on bone Pb
mobilization
during lactation
N=533
Mexico City
1994-1995
Singleton births
from the
Baltimore
Tracking Health
Related to
Environmental
Exposures
(THREE) study
K 1 OO
NiBW-33
N normal BW=267
Baltimore, MD
2004-2005
Data for Pb
Biomarkers or
Methodological other Exposure
Details Indicators
Cross-sectional Umbilical cord blood
study using Pb
linear regression Means varied from
6.3 to 6.9
(categories created
based on maternal
HFE genotype)
>10ug/dL: 12.6%
Maternal tibia and
patella Pb
one month after
delivery
Cross-sectional Umbilical cord Pb
study using Mean (95% Cl):
multivariable „„„,„-,„ „ ™>
linear regression 0.84(0.72,0.96)
ug/dL
> 5 ug/dL: 0.7%
Adjusted Effect Estimates
(3 (95% Cl), in grams
Umbilical cord blood Pb:
-31.1 (-105.4, 43.3) perl ug/dL
Maternal tibia Pb
Overall: -4.4 (-7.9, -0.9) per 1 ug/g
<1-4.1 ug/g: Ref
4.1-9.2 ug/g: 17.2 (-75.6, 110.1)
9.2-15.4 ug/g: -19.1 (-112.1, 73.9)
15.4-43.2 ug/g: -95.4 (-189.9, -0.8)
Ratio for Pb concentration per 100g birth
weight: 1.01 (0.99, 1.02)
Potential
Confounders
Adjusted for in
Analysis
Maternal age,
education, infant
sex, maternal
arm
circumference,
gestational age,
smoking status
during
pregnancy,
marital status,
maternal
hemoglobin
first month
postpartum,
parity
Maternal age,
race, insurance,
pre-pregnancy
weight, smoking
status,
gestational
length, birth
weight, average
year of
neighborhood
home
construction
4-646
-------
Table 4-42 (Continued): Summary of recent epidemiologic studies of associations between Pb exposure indicators and low
birth weight and fetal growth.
Study
Population
Study Location Methodological
Reference3 Outcome Years Details
Al-Saleh et Head Infants with a Cross-sectional
al. (2008b) circumference gestational age of study using
at least 34 weeks linear regression
born to healthy
mothers aged
17-46 yr and non-
occupationally
exposed to Pb
N=653
Saudi Arabia
2004
Atabek et Birth weight, Term, singleton Cross-sectional
al. (2007) birth length, infants born to study using
head healthy mothers linear regression
circumference, living in urban
mid-arm areas and
circumference assumed to have
high Pb
concentrations
N=54
Turkey
NS
Data for Pb
Biomarkers or
other Exposure
Indicators
Umbilical cord blood
Pb
Mean (SD):
2.21 0(1. 691 )ug/dl_
75th percentile:
2.475 ug/dl_
>10ug/dL: 1.23%
Umbilical cord blood
Pb
Mean (SD):
14.4(8.9)
>10ug/dL: 53.7%
> 25 ug/dl_: 9.2%
Adjusted Effect Estimates
Regression models for those above the
75th percentile of cord blood Pb levels
(3 (SE) per unit of log-transformed Pb
-0.158(0.718), p-value: 0.036
(3 (p-value)
Birth weight: -0.81 (0.01)
Birth length: 0.41 (0.05)
Mid-arm circumference: 0.30 (0.05)
Potential
Confounders
Adjusted for in
Analysis
BMI, gestational
age
Considered but
not included:
prenatal
supplements,
location
of residence
Age, sex
Note: inclusion
of SES did not
change the
results
4-647
-------
Table 4-42 (Continued): Summary of recent epidemiologic studies of associations between Pb exposure indicators and low
birth weight and fetal growth.
Reference3
Llanos and
Ronco
(2009)
Zentner et
al. (2006)
Janjua et
al. (2009)
Outcome
Fetal growth
restriction
(1,000-2,500g)
*note normal
birth weights
were >3,000g
Birth weight and
length
Low birth weight
(<2,500g)
Study
Population
Study Location
Years
Term births
(37-40 weeks)
from non-smoking
mothers
N growth restricted =20
N normal BW=20
Santiago, Chile
NS
Singleton births
with maternal
residence within
5 km of Pb
smelter
N=55
Santo Amaro,
Brazil
2002
Infants of
randomly
selected women
who planned to
deliver between
37-42 weeks
NLBw=100
N normal BW=440
Karachi, Pakistan
2005
Methodological
Details
Cross-sectional
study using
Mann-Whitney
U-test
Cross-sectional
study using
linear regression
Cross-sectional
study using
binomial
regression
Data for Pb
Biomarkers or
other Exposure
Indicators
Placenta Pb
Mean (SD)
Fetal growth
restricted: 0.21
(0.04) ug/g
Controls:
0.04 (0.009) ug/g
Umbilical cord blood
Pb from delivery
Mean(SD): 3.9(3.6)
Umbilical cord blood
Pb
Mean(SD): 10.8
(0.2)
Potential
Confounders
Adjusted for in
Adjusted Effect Estimates Analysis
P-value for Mann-Whitney U-test <0.01 None
Linear regression coefficient (p-value) with No other
umbilical cord blood Pb as the dependent variables
variable in model with only length and besides length
weight (units not given for any and weight were
measurements, assume per 1 ug/dL for included in the
cord blood Pb): model
Length: -0.46 (0.003)
Weight: -0.275 (0.048)
(i.e., in this study, Pb is assessed as the
outcome)
Prevalence ratio (95% Cl): None
<10ug/dL: 1.00(Ref)
>10ug/dL: 0.82(0.57, 1.17)
4-648
-------
Table 4-42 (Continued): Summary of recent epidemiologic studies of associations between Pb exposure indicators and low
birth weight and fetal growth.
Reference3 Outcome
Jones et al. Low birth weight
(2010) (<2,500g)
Orun et al. Birth weight and
(2011) head
circumference
Study
Population
Study Location
Years
Singleton births
> 27 weeks
gestation from
mothers aged
16-45 yr living in
the Shelby
County area for at
least 5 mo during
pregnancy
NLBw=10
N normal BW=92
Tennessee
2006
Births to mothers
not occupationally
exposed to toxic
metals and living
Methodological
Details
Cross-sectional
study comparing
across geometric
means (test not
specified)
Cohort study
using Pearson
correlation
coefficients
Data for Pb
Biomarkers or
other Exposure
Indicators
Umbilical cord blood
Pb
Mean (SD): 2.4 (4.3)
Geometric mean:
1.3
Breast milk
2 months post-
partum
Median: 20.6 ua/L
Potential
Confounders
Adjusted for in
Adjusted Effect Estimates Analysis
Geometric Mean: None
Low birth weight: 1.2
Normal birth weight: 1.3
p-value for difference: >0.10
Correlations for breast milk Pb and z- None
scores of head circumference
Girls: 0.087
in a suburban but
non-industrial
area
NLBw=9
N normal BW=135
Turkey
NS
>WHO limit (5 ug/L):
87%
Median (IQR)
<2,500g:
20.4(8.5, 27.1) ug/L
>2,500g:
20.6(11.8,
29.5) ug/L
Boys: 0.029
Correlations for breast milk Pb and z-
scores of birth weight
Girls: 0.097
Boys: 0.045
*AII p-values for correlations>0.05
4-649
-------
Table 4-42 (Continued): Summary of recent epidemiologic studies of associations between Pb exposure indicators and low
birth weight and fetal growth.
Study Data for Pb
Population Biomarkers or
Study Location Methodological other Exposure
Reference3 Outcome Years Details Indicators
Williams et Birth weight Infants from Longitudinal Air Pb levels during
a 1. (2007) singleton births or cohort study first trimester of
the firstborn infant using pregnancy
in a set of hierarchical Mean (SDV
multiples linear models „ „_ , 3
0.12ug/m
(0.04 ug/m3)
N=not specified
Tennessee
2002
Potential
Confounders
Adjusted for in
Adjusted Effect Estimates Analysis
p-value for multilevel regression of Pb with Previous
birth weight: 0.002 preterm birth,
Increase of air Pb from 0 to 0.04 relates to previous
a 38g decrease in birth weight birth>4000g,
pregnancy-
Increase of Pb from 0 to 0.13 (maximum) induced
ug/m3 relates to a 124g decrease in birth hypertension
Wej9ht chronic
hypertension,
oligohydramnios,
other maternal
risk factors,
education,
cigarettes/day,
black race,
Hispanic
ethnicity, other
race/ethnicity,
plurality, infant
sex, first
trimester SC>2,
within 5km of an
air monitor,
poverty,
interaction of
poverty and
other maternal
risk factors,
percentage of
previous
pregnancies that
resulted in non-
live births
4-650
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Table 4-42 (Continued): Summary of recent epidemiologic studies of associations between Pb exposure indicators and low
birth weight and fetal growth.
Reference3
Berkowitz
etal.
(2006)
Outcome
Low birth weight
(<2,500 g and
> 37 weeks)
Small for
gestational age
(birth weight
< 5th percentile
of sex- and
gestational week
weights for
singletons in
Idaho)
Study
Population
Study Location
Years
Singleton infants
with 28-45 weeks
gestation
NLBw=4,297
• v normal
Bw=162,035
NSGA=7,020
NnoSGA = 162,035
Idaho
1970-1981
Data for Pb
Biomarkers or
Methodological other Exposure
Details Indicators
Cohort study Three time periods
using logistic of two locations
regression (unexposed and
exposed/near
smelter): pre-fire,
"high-exposure
period" (when a fire
happened at the
smelter and resulted
in damages leading
to high air Pb
concentrations for 6
mo), and "post-fire"
During the time of
Adjusted Effect Estimates
Term Low birth weight:
OR (90% Cl) (unexposed location is
referent group):
Pre-fire: 0.81 (0.55, 1.20)
High exposure: 2.39 (1.57, 3.64)
Post-fire: 1.28(0.95, 1.74)
Small for gestational age:
OR (90% Cl) (unexposed location is
referent group):
Pre-fire: 0.98(0.73, 1.32)
High exposure: 1.92 (1.33, 2.76)
Post-fire: 1.32(1.05, 1.67)
Potential
Confounders
Adjusted for in
Analysis
Maternal age,
infant sex, first
birth, previous
miscarriage or
abortion
Pb in ambient air
were as high as
30 ug/m3
"Studies are presented in order of first appearance in the text of this section, which is based on study design and methodological strength.
4-651
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Multiple studies were conducted that examined the association between maternal blood
Pb and birth weight/fetal growth. A study in Vienna, Austria reported an inverse
association between maternal blood Pb levels and birth weight but no associations for
birth length or head circumference (Gundacker et al.. 2010). Similarly, increased
maternal blood Pb was associated with decreased birth weight among infants in a study
performed in New York (Zhu et al.. 2010). No association was observed between
maternal blood Pb levels and having an infant who was small for his/her gestational age
(SGA). A study in Taiwan examined both maternal and paternal blood Pb levels among
those occupationally exposed to Pb and their associations with birth weight and SGA
(Chen et al.. 2006a). Paternal blood Pb levels were not associated with increased risk of
low birth weight or SGA. Higher maternal blood Pb concentration was associated with
higher risk of low birth weight and SGA, although not all of the associations were
statistically significant. There were small numbers of infants with low birth weight or
SGA, especially at the highest blood Pb levels (> 20 (ig/dL). In California, blood Pb
measurements of women during pregnancy were matched with the corresponding birth
certificates (Jelliffe-Pawlowski et al.. 2006). The adjusted OR for low birth weight that
compared women with blood Pb levels > 10 (ig/dL to women with levels <10 (ig/dL was
elevated. However, it was difficult to draw conclusions about the relationship between
blood Pb and birth weight due to small numbers (n = 9 for low birth weight) and the
consequently wide 95% CI. An association was detected for increased blood Pb and
SGA. Women residing in two different towns in Yugoslavia (one with a Pb smelter and
one without a Pb smelter) were recruited during their first prenatal visit (Lamb et al..
2008) (study based on previous work by Factor-Litvak et al. (1991)). The mid-pregnancy
blood Pb levels were greater in women from the town with a Pb smelter. No association
was reported between maternal blood Pb and height or BMI at birth for the infants of
these women despite the differences in maternal blood Pb between the two towns. A
study of term births in Iran reported no difference in blood Pb levels of women giving
birth to a normal weight infant and women giving birth to an infant with low birth weight
(Iranpour et al.. 2007).
A study examining the association between Pb biomarker levels and birth weight used
tibia bone measurements one month post-partum from mothers living in Mexico City
(Kordas et al.. 2009). Tibia Pb levels were inversely associated with birth weight but not
with birth length. This association between Pb and birth weight was not modified by
maternal folate consumption or maternal or infant MTHFR genotype, although the
association between tibia Pb levels and birth weight was greater in magnitude among
women with certain MTHFR SNPs (statistical tests not reported). Another study in
Mexico City reported no association between maternal tibia Pb levels and birth weight
among girls but reported a positive association for boys (Afeiche et al.. 2011). No
associations were observed with maternal patella Pb concentration, although among boys,
4-652
-------
the relationship was positive but not statistically significant. One of the cohorts used by
Afeiche et al. (2011) was also evaluated in another study (Cantonwine et al.. 2010b). An
inverse association was observed between maternal tibia Pb and birth weight, especially
at higher levels (over 15.4 (ig/dL). This association was stronger among those mothers
with variants of the hemochromatosis Fe gene (HFE).
Multiple studies examined the relationship between Pb level and birth weight using Pb
measured from the placenta or umbilical cord. A study performed in Baltimore, MD
reported no association between umbilical cord blood Pb concentration and birth weight
(Wells etal.. 201 la). This study had low blood Pb levels, with only 0.7% of participants
having umbilical cord blood Pb measuring >5 (ig/dL. In Saudi Arabia, a study was
conducted among non-occupationally exposed women (Al-Saleh et al., 2008b). Umbilical
cord blood Pb concentrations were low and an association was observed between
umbilical cord Pb levels above the 75th percentile (2.475 (ig/dL) and head circumference.
A study with high Pb concentrations in umbilical cord blood reported an inverse
association between Pb levels and birth weight (Atabek et al., 2007). However, no
correlation was detected in an analysis restricted to umbilical cord Pb less than 10 (ig/dL.
No associations with other measures of growth, such as birth length and mid-arm
circumference, were detected. Researchers in Chile collected the placentas from term
births and compared the Pb levels for those born with normal birth weights to those with
low birth weights (Llanos and Ronco. 2009). Pb levels were greater in the placentas of
infants with low birth weights. In addition, the authors note that 3 low birth weight
infants had extremely high Pb levels in the placentas (>1.5 (ig/g) and were excluded from
these analyses. A study in Brazil examined Pb levels in umbilical cord blood from term
births of women residing within 5 km of a Pb smelter (Zentner et al.. 2006). The cord
blood Pb level was found to be inversely correlated with length and weight of the infants.
Another study recruited women in Pakistan (Janjua et al.. 2009). Umbilical cord blood Pb
levels were not associated with low birth weight. The study by Iranpour et al. (2007)
discussed above investigated the association with umbilical cord blood Pb levels in
addition to their examination of maternal whole blood Pb. They again report no
difference in cord blood Pb levels between term infants of normal and low birth weight.
A study comparing geometric mean umbilical cord blood Pb levels reported no difference
in the levels for normal and low birth weight infants born to women living primarily in
urban areas of Memphis, TN (Jones etal.. 2010). A study previously mentioned that
observed an inverse association between maternal tibia Pb and birth weight in Mexico
City reported no association between umbilical cord blood Pb concentration and birth
weight (Cantonwine et al.. 201 Ob). Finally, a study in Vienna measured Pb in the placenta
(Gundacker et al.. 2010). A positive correlation was observed between placenta Pb and
birth length and weight; however, in the same study, maternal blood Pb was inversely
related to birth weight.
4-653
-------
A study performed in Turkey examined the relationship between Pb levels in breast milk
two months postpartum and size at birth (Orim et al.. 2011). No association was observed
between breast milk Pb concentration and birth weight or head circumference.
A few studies examined air exposures and reported inverse associations between air Pb
concentrations and birth weight. However, a limitation of these studies is the difficulty in
assessing if the measured concentrations represent population exposures (see
Section 3.5.3). Williams et al. (2007) examined Pb concentrations in the air during the
first trimester. The purpose of their study was to demonstrate the use of hierarchical
linear models and they used the example of air pollution and birth weight in Tennessee.
The model results showed an association between ambient Pb concentration and birth
weight, with an estimated decrease in birth weight of 38 grams for every 0.04 (ig/m3
(i.e., one standard deviation) increase in Pb concentration. Another study of air Pb levels
was conducted in Idaho and included two areas over three time periods. One study area
was affected by damage to pollution control facilities for a large Pb smelter that led to
high Pb emissions during one of the time periods under study (Berkowitz et al.. 2006).
During the 6 month period when emission controls were affected by the fire damage,
estimates of Pb in ambient air were as high as 30 (ig/m3. Mean birth weight for term
births was decreased among infants born to women living in the high exposure area
during the period of high exposure compared to those living in the lower exposure area.
The difference in birth weight of term births remained, but was reduced, between the two
areas during the time period after the exposure ended. During the period of higher
exposure, the odds of low birth weight among term births was increased among those
living in the higher exposed area compared to those in the lower exposed area, but the
odds were not different between the two study areas during the time periods before or
after the high level of exposure. An increase in SGA infants (defined as infants with
weights less than or equal to the lowest 5th percentile of birth weight for their sex and
age) was also associated with living in the higher exposed area during the time period of
higher exposure. The odds of SGA infants decreased during the time period after the
exposure but the odds were still elevated compared to those residing in the lower exposed
area.
Evidence from previous toxicological studies has shown an association between
gestational Pb exposure and reduced birth weight (U.S. EPA. 2006b). More recent studies
have reported conflicting results. Wang et al. (2009e) demonstrated a statistically
significant decrease in fetal body weight and body length of Wistar rats after maternal
exposure to 250 ppm Pb acetate during gestation days 1-10, 11-20, or 1-20. The greatest
decrease in fetal body weight and length was observed in the group exposed to Pb during
gestation days 1-20 followed by the group exposed to Pb during gestation days 11-20.
Teijon et al. (2006) observed that when pregnant dams were administered 200 ppm or
4-654
-------
400 ppm Pb acetate in drinking water, litter weight was significantly decreased (400 ppm
Pb only) versus controls due to significant decrements in female pup birth weight; male
birth weight was unaffected. The results of these studies indicate that as Pb exposure
increases, the body weight of exposed offspring decreases. Masso-Gonzalez and
Antonia-Garcia (2009) also observed an 8-20% decrease in body weight of pups from rat
dams given 300 mg/L Pb acetate in drinking water (exposure during gestation and
lactation resulting in mean blood Pb level of 22.8 (ig/dL), but no changes in body length
were reported.
In summary, associations were observed between Pb and low birth weight in
epidemiologic studies of maternal bone Pb and studies of Pb air exposures and birth
weight. The associations were less consistent when using maternal blood Pb or umbilical
cord and placenta Pb as the exposure measurement although some studies did
demonstrate associations. Epidemiologic studies of Pb and fetal growth face multiple
limitations. One limitation is the cross-sectional nature of many studies. These do not
allow an understanding of the temporality for Pb and fetal growth. In addition, some
studies suffer from small sample size. The studies of air Pb levels and birth weight, one
of which involved air Pb concentrations as high as 30 (ig/m3, demonstrate positive
associations but are limited in that exposure levels are unknown and certain size fractions
have less certain relationships with population exposure levels (see Chapter 3). Also,
many of the studies controlled for important confounders, such as parity and gestational
age, but adjustment in some studies was lacking. Toxicological studies observed an
association between gestational Pb exposure and reduced birth weight with moderate
(22.8 (ig/dL blood Pb level) to higher dose Pb (400 ppm Pb acetate in drinking water).
4.8.3 Effects on Male Reproductive Function
The 2006 Pb AQCD (U.S. EPA. 2006b) reported on male Pb exposure or biomarker
levels and reproductive functions in males as measured by sperm
count/motility/morphology, time to pregnancy, reproductive history, and chromosomal
aberrations. Despite limitations, most of the studies found slight associations between
high blood Pb levels (i.e., > 45 (ig/dL) and reduced male fecundity or fertility (U.S. EPA,
2006b). Evidence reviewed in the 1986 Pb AQCD (U.S. EPA. 1986a) also demonstrated
that Pb exposure affects male reproductive function in humans and experimental animals.
Recently published research has continued to support an association between Pb and
sperm/semen production, quality, and function. Studies of Pb and male reproductive
function are described in the sections below. The studies presented in the following text
and tables are grouped by study design and methodological strength.
4-655
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4.8.3.1 Effects on Sperm/Semen Production, Quality, and
Function
Multiple epidemiologic and toxicological studies have examined the relationship between
Pb and sperm and semen production, quality, and function. These studies are summarized
in the text below. In addition, recent epidemiologic studies are included in Table 4-43.
All epidemiologic studies were cross-sectional with concurrent measurements of Pb
levels in biological samples and sperm-related outcomes.
4-656
-------
Table 4-43 Summary of recent epidemiologic studies of associations between Pb biomarker levels and effects
on sperm and semen.
Reference3
Hsuetal.
(2009b)
Study Population
Study Location
Years
Men working at a
battery plant
N=80
Taiwan
NS
Methodological
Details
Occupational
cohort study
(cross-sectional)
using ANOVA
and linear
regression
Pb Biomarker Data
Mean (SD) in ug/dL
Blood Pb: 42
Categorized into 3 groups:
<25 ug/dl_,
25-45 ug/dl_,
>45 ug/dl_
Adjusted Effect Estimates
p-values for difference across the three groups
were <0.05 for: sperm head abnormalities,
sperm neck abnormalities, sperm chromatin
structure assay (aT, COMPaT)
p-values for difference across the three groups
were >0.05 for: semen volume, sperm count,
Potential
Confounders
Adjusted for
in Analysis
Smoking status
immaturity, computer-assisted semen analysis,
% sperm with ROS production
Coefficients for regression analysis with blood
Pb:
Morphologic abnormality: 0.271 (p-value
<0.0001)
Head abnormality: 0.237 (p-value 0.0002)
aT: 1.468 (p-value 0.011)
COMPaT: 0.233 (p-value 0.21)
4-657
-------
Table 4-43 (Continued): Summary of recent epidemiologic studies of associations between Pb biomarker levels and effects on
sperm and semen.
Study Population
Study Location
Reference3 Years
Methodological
Details
Pb Biomarker Data
Mean (SD) in ug/dL
Adjusted Effect Estimates
Potential
Confounders
Adjusted for
in Analysis
Kasperczyk
et al. (2008)
Healthy, non-
smoking, fertile
men who worked
at the Zn and Pb
Metalworks
NControls = 14
N low exposure"^U
N high exposure"^"
Poland
NS
Occupational
cohort study
(cross-sectional)
using Kruskal-
Wallis ANOVA
and Spearman's
coefficient for
non-parametric
correlation
Blood Pb
Categorized as
High exposure workers : 53.1
(2.05)
Range: 40-81 ug/dL
Low exposed workers: 34.7 (0.83)
Range: 25-40 ug/dL
Controls (office workers with no
history of occupational Pb
exposure):
8.47 (0.54)
Seminal plasma Pb
High exposure workers: 2.02 (0.23)
Low exposure workers: 2.06 (0.40)
Controls: 1.73(0.16)
Mean (SE)
Sperm volume (ml_)
Controls: 2.94 (0.32)
Low exposure: 2.89 (0.22)
High exposure: 2.98 (0.22)
(p-value for ANOVA: 0.993)
Sperm cell count (mln/mL)
Controls: 43.1 (7.0)
Low exposure: 44.6 (10.1)
High exposure: 42.2 (5.86)
(p-value for ANOVA: 0.400)
Normal morphology (%)
Controls: 63.3 (2.7)
Low exposure: 57.3 (2.5)
High exposure: 58.4 (2.1)
(p-value for ANOVA: 0.266)
Progressively motile sperm after 1 h (%)
Controls: 16.4(3.2)
Low exposure: 14.8(2.6)
High exposure: 10.5 (1.9)
(p-value for ANOVA: 0.217)
Motile sperm after 24 h (%)
Controls: 4.4(1.8)
Low exposure: 7.3 (1.7)
High exposure: 3.1 (0.8)
(p-value for ANOVA: 0.188)
p-value for correlation between blood Pb and
sperm cell motility after 1 h: 0.011
None
4-658
-------
Table 4-43 (Continued): Summary of recent epidemiologic studies of associations between Pb biomarker levels and effects on
sperm and semen.
Reference3
Naha and
Manna
(2QQ7)
Study Population
Study Location
Years
Non-occupationally
exposed controls
and occupationally
exposed workers
NControls=50
N low exposure — OU
N high exposure —^\J
Methodological
Details
Occupational
cohort study
using ANOVA,
Student's t-test,
and Scheffe's F-
test
Pb Biomarker Data
Mean (SD) in ug/dL
Categorized by work history as
controls, low exposure (7-10 yr of
exposure for 8 h/day) and high
exposure (>1 0 yr of exposure for 8
h/day)
Blood Pb measurement
Controls: 10.25(2.26)
Adjusted Effect Estimates
p-values for difference across the three groups
for mean values of semen profiles were <0.01
for: liquefaction time, seminal volume, sperm
count, sperm DNA hyploidy, sperm
morphological abnormality, sperm motility,
sperm ATPase activity, seminal plasma
fructose, seminal plasma total protein, seminal
plasma free amino acid, seminal plasma
cholesterol
Potential
Confounders
Adjusted for
in Analysis
None
Bangalore, India
NS
Low exposure : 50.29 (3.45)
High exposure: 68.26 (2.49)
Semen Pb measurement
Controls: 2.99 (0.76)
Low exposure: 15.85(1.95)
High exposure: 25.30 (2.28)
Naha and
Chowdhury
(2006)
Men aged 31-45
that were non-
occupationally
exposed controls
and occupationally
exposed workers)
NControls=50
N low exposure"^U
N high exposure —bO
Kolkata, India
NS
Occupational
cohort study
using ANOVA,
Student's t-test,
and Scheffe's F-
test
Categorized by work history as
controls, low exposure (7-10 yr of
exposure for 8 h/day) and high
exposure (>10 yr of exposure for 8
h/day)
Blood Pb measurement
Controls 13.62 (2.45)
Low exposure 48.29 (4.91)
High exposure 77.22 (1.25)
Semen Pb measurement
Controls 3.99 (1.36)
Low exposure 10.85 (0.75)
High exposure 18.30 (2.08)
p-values for difference across the three groups
for mean values of semen profiles were <0.01
for: sperm count, sperm protein, sperm DNA
hyploidy, sperm DNA, sperm RNA, sperm
viability, sperm membrane lipid peroxidation,
seminal plasma total ascorbate, seminal plasma
DHAA, sperm ATPase activity, sperm motility,
sperm velocity, seminal plasma fructose
None
4-659
-------
Table 4-43 (Continued): Summary of recent epidemiologic studies of associations between Pb biomarker levels and effects on
sperm and semen.
Reference3
Telisman et
al. (2007)
Study Population
Study Location
Years
Men aged 19-55
yr, never
occupationally
exposed to metals
and going to a
clinic for infertility
examination or for
semen donation to
be used for
artificial
insemination
Methodological Pb Biomarker Data
Details Mean (SD) in ug/dL
Cross-sectional Blood Pb
study using
linear multiple Median: 4.92 (range 1.13-14.91)
regression
Adjusted Effect Estimates
Standardized regression coefficients for log
blood Pb (units not given)
Immature sperm: 0.13 (p-value <0.07)
Pathologic sperm: 0.31 (p-value <0.0002)
Wide sperm: 0.32 (p-value <0.0001 )
Round sperm: 0.16 (p-value <0.03)
Coefficients and p-values not given if not
statistically significant: semen volume, sperm
concentration, slow sperm, short sperm, thin
sperm, amorph sperm
Potential
Confounders
Adjusted for
in Analysis
Cd, Cu, Zn, Se,
age, smoking
status, alcohol
use
N=240
Croatia
2002-2005
4-660
-------
Table 4-43 (Continued): Summary of recent epidemiologic studies of associations between Pb biomarker levels and effects on
sperm and semen.
Reference3
Study Population
Study Location
Years
Methodological
Details
Pb Biomarker Data
Mean (SD) in ug/dL
Adjusted Effect Estimates
Potential
Confounders
Adjusted for
in Analysis
Meeker et
al. (2008)
Men aged 18-55 yr
going to infertility
clinics (distinction
not made between
clinic visits for
male or female
fertility issues)
N=219
Michigan
NS
Cross-sectional
study using
multiple logistic
regression
Blood Pb
Median: 1.50 (IQR 1.10, 2.00)
OR (95% Cl) for having below reference-level
semen parameters
Concentration
1st quartile: 1.00 (ref)
2nd quartile: 0.88 (0.32, 2.44)
3rd quartile: 2.58(0.86, 7.73)
4th quartile: 1.16(0.37, 3.60)
Motility
1st quartile: 1.00 (ref)
2nd quartile: 1.04(0.43, 2.53)
3rd quartile: 1.95(0.70, 5.46)
4th quartile: 1.66(0.64,4.29)
Age, smoking
status
Models with
multiple metals
included:
smoking status,
Mo, Mn, Cd,
and Hg
Considered but
did not include:
BMI, race
Morphology
1st quartile: 1.00 (ref)
2nd quartile: 0.83(0.37, 1.87)
3rd quartile: 1.41 (0.54, 3.67)
4th quartile: 1.18(0.50,2.79)
Models with adjustment for multiple metals
Concentration
1st quartile: 1.00 (ref)
2nd quartile: 0.89(1.57,2.89)
3rd quartile: 3.94(1.15, 13.6)
4th quartile: 2.48(0.59, 10.4)
4-661
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Table 4-43 (Continued): Summary of recent epidemiologic studies of associations between Pb biomarker levels and effects on
sperm and semen.
Reference3
Slivkova et
al. (2009)
Mendiola et
al. (2011)
Study Population
Study Location
Years
Men aged 22-48 yr
undergoing semen
analysis at an
infertility clinic
N=47
NS
Men attending
infertility clinics
and classified as
either normal
sperm (controls) or
oligo-astheno-
teratozoospermia
(cases) based on
WHO semen
quality criteria
Methodological
Details
Cross-sectional
study using
correlation
Case-control
study using
multiple linear
regression
Pb Biomarker Data
Mean (SD) in ug/dL
Semen Pb
1.49 (0.40) mg/kg
Seminal plasma: 2.90 (IQR 2.70,
3.20)
Whole blood: 9.50 (IQR 7.50,
11.90)
Blood plasma: 2.90 (IQR 2.70,
3 10)
/
Cases:
Seminal plasma: 3.0 (0.3)
Adjusted Effect Estimates
Correlation between Pb and flagellum ball :
-0.39 (p-value not given)
"correlations not given for any other sperm
pathological changes (therefore assume not
statistically significant): broken flagellum,
separated flagellum, separated flagellum, small
heads, retention of cytoplasmic drop, other
pathological spermatozoa, large heads,
acrosomal changes, and knob twisted flagellum
(3(95%CI)
Sperm concentration
Seminal plasma: -1.0 (-3.1, 2.3)
Whole blood: -0.2 (-1.7, 1.6)
Blood plasma: 0.08 (-4.1, 5.2)
% Immotile sperm
Seminal plasma: 1.5 (0.37, 1.9)
Whole blood: 0 05 (-0 32 0 43)
Potential
Confounders
Adjusted for
in Analysis
None
Age, BMI,
number of
cigarettes/day
NControls~31
Ncases=30
Spain
2005-2007
Whole blood: 9.8 (2.3)
Blood plasma: 2.9 (0.2)
Controls:
Seminal plasma: 2.9 (0.3)
Whole blood: 9.7 (2.3)
Blood plasma: 2.9 (0.3)
Blood plasma: -0.49 (-1.8, 0.62)
% Morphologically normal sperm
Seminal plasma: -0.54 (-3.1, 2.0)
Whole blood:-0.31 (-1.5, 0.89)
Blood plasma: -0.08 (-3.5, 3.4)
*Units not given (assume 1 ug/dL)
Note: No correlation in Pb levels among bloods
or seminal plasma. There was correlation
between Pb and other metals (Cd and Hg)
within each body fluid. Other metals were not
controlled for in models
aStudies are presented in order of first appearance in the text of this section, which is based on study design and methodological strength.
4-662
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International epidemiologic studies of men occupationally exposed to Pb have reported
on associations between Pb exposure or biomarker levels and sperm count and quality
and semen quality. In most of these occupational studies, mean blood Pb levels over
40 (ig/dL have been reported for individuals occupationally exposed to Pb. In addition,
they did not control for other potential occupational exposures. A study performed in
Taiwan among men with high levels of blood Pb reported that men with blood Pb levels
of 25-45 (ig/dL and higher had increased sperm head abnormalities, increased sperm
DNA denaturation, and increased sensitivity to denaturation compared to men with lower
blood Pb levels (Hsu et al., 2009b). No difference was detected between three Pb
exposure groups in semen volume, sperm count, motility, velocity, and reactive oxygen
species (ROS) production. A similar study in Poland included employees exposed to Pb
and compared them with a group of male office workers (Kasperczyk et al., 2008). Pb
levels measured in seminal fluid were slightly higher among those in the exposed groups
although they were not statistically different from the levels in the control group. No
difference was observed for semen volume, sperm count, or sperm morphology among
the groups. Sperm motility was lower in the highest exposure group (40-81 (ig/dL)
compared to both the control (no history of exposure) and moderate exposure (25-40
(ig/dL) groups. Lipid peroxidation, which can induce tissue damage in sperm via ROS,
was greater in the highest exposure group compared to the controls. Studies performed in
India (Naha and Manna. 2007; Naha and Chowdhury. 2006) reported that men in the
highest exposure group (men working in battery or paint manufacturing plants for
10-15 years for 8 hours/day) had mean blood Pb levels of77.22 (ig/dL (Naha and
Chowdhury. 2006) and 68.26 (ig/dL (Naha and Manna. 2007). Control groups in these
studies (those without occupational Pb exposure) had mean blood Pb levels below
15 (ig/dL. Increases in levels of Pb in semen were also noted across exposure groups.
Both studies report decreases in sperm count and in sperm velocity and motility with
increasing Pb exposure. Higher Pb exposure was also associated with greater hyploidy of
sperm DNA and morphologic abnormalities (Naha and Manna. 2007; Naha and
Chowdhury. 2006). Decreased viability and increased lipid peroxidation were detected
(Naha and Chowdhury. 2006).
A few studies examined blood or seminal plasma Pb levels and semen quality of men at
infertility clinics (Mendiola et al.. 2011; Slivkovaet al.. 2009; Meeker etal.. 2008;
Telisman et al.. 2007). In general, these men had lower levels of Pb biomarkers than men
who were occupationally exposed, but the studies are limited by the strong possibility of
selection bias related to the recruitment of men attending infertility clinics. A study
performed in Croatia recruited men attending a clinic for infertility examination or to
donate semen for use in artificial inseminations, who had never been occupationally
exposed to metals and therefore had lower blood Pb levels than the occupational studies
(although leaded gasoline was still sold during the study period) (Telisman et al., 2007).
4-663
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Increased blood Pb was associated with increased percentages of pathologic sperm, wide
sperm, and round sperm. There was also a slight increase in immature sperm although it
was not statistically significant. Similar results were seen when other biomarkers for Pb
(erythrocyte protoporphyrin and 5-aminolevulinic acid dehydratase [ALAD]) were used
instead. This study controlled for multiple potential confounders, including other metals.
Meeker et al. (2008) detected no associations between higher blood Pb and semen
concentration, morphology, or motility (although a slight positive trend was observed
between higher Pb levels and motility in unadjusted models). In models that included
multiple metals, blood Pb was associated with being below the WHO limit of sperm
concentration levels (less than 20 million sperm/mL), although the 95% CI was wide for
the 4th quartile of Pb levels and included the null. The precision of estimates in this study
was extremely low. Slivkova et al. (2009) reported a negative correlation between semen
Pb and pathological changes in sperm (specifically, flagellum ball), but no correlations
were observed for other alterations in the sperm. Another study reported a positive
association between seminal plasma Pb concentration and percentage of immotile sperm,
but this analysis did not adjust for exposure to other metals reported to be correlated with
Pb concentration in the seminal plasma (Mendiola et al.. 2011). No association was
observed for seminal plasma Pb concentration and sperm concentration or percentage of
morphologically normal sperm. Additionally, neither Pb levels in whole blood nor
plasma were associated with sperm concentration, percentage of immotile sperm, or
percentage of morphologically normal sperm.
Extensive evidence in the toxicological literature demonstrates that Pb exposure is
detrimental to the quality and overall health of testicular germ cells, affecting sperm and
semen quality and production. Earlier Pb AQCDs contained studies of Pb-induced
decreased sperm counts, decreased sperm production rate, and dose-dependent
suppression of spermatogenesis in adult rodents with 30 day drinking water Pb exposure
[(Sokol and Berman. 1991). blood Pb level 35 and 37 (ig/dL; (Sokoletal.. 1985). blood
Pb level 34 (ig/dL; (Sokol. 1989). blood Pb levels <43 (ig/dL]. Chronic Pb treatment
(15 weeks) of adult male rabbits, resulting in a blood Pb of 24 (ig/dL, induced statistically
significant decrements in semen quality and greater testicular pathology (Moorman et al..
1998) with dosing by subcutaneous injection, loading dose of 0.2-3.85 mg/kg BW
Monday (M), Wednesday (W) and Fridays (F) weeks 6-10, followed by maintenance
dose of 0.13-2.0 mg/kg BW Pb acetate MWF over weeks 11-20 of the study. The
2006 Pb AQCD also cited studies in which sperm from Pb exposed rats yielded lower
rates of fertilization when used for in vitro fertilization of eggs harvested from unexposed
females [(Sokoletal.. 1994). blood Pb level 33-46 (ig/dL].
Recent studies corroborate earlier findings that Pb alters sperm parameters such as sperm
count, viability, motility, and morphology. Anjum et al. (2010) exposed 50 day old male
4-664
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albino Wistar/NIN rats to Pb acetate (273 or 819 mg/L in drinking water, 500 or
1,500 ppm, respectively) for 45 days. Affected endpoints included reduced epididymal
sperm count, motile sperm, and viable sperm, indicating decreased sperm production and
quality. Anjum et al. (2010) did not report blood Pb values. Wistar/NIN rats (1,500 ppm
Pb acetate in drinking water for 70 days) supplemented with the herb Centella asiatica
(Sainath et al.. 2011) had significant attenuation of the Pb-induced changes observed by
Anjum et al. (2010). Pillai et al. (2012) found gestational and lactational treatment with
Pb acetate in Charles Foster rats (subcutaneous injection of 0.05 mg/kg BW/day) induced
effects on sperm in adults (PND65) including significant decreases in testicular sperm
count, epididymal sperm count, and sperm motility. These findings are consistent with
Pb-associated effects on sperm and male reproductive organs in wildlife from the
ecological literature including deer, Asian earthworms, rainbow trout, marine worms, and
the fathead minnow (see Sections 6.3.12.1 and 6.4.12.1 from the ecology terrestrial and
aquatic reproduction sperm sections).
Pb exposure has been shown to affect the male reproductive organs, as is seen with
histological or morphological changes. Studies included in previous Pb AQCDs showed
that histological and ultrastructural damage to the testes or seminiferous tubules was seen
in non-human primates with chronic oral Pb exposure (daily Pb exposure, gelatin
capsule; control plus 3 treatment groups) during three different time periods: (1) infancy
(PNDO-PND400), resulting in a maximum blood Pb level of 36 ug/dL, (2) post-infancy
(PND300 up to 10 years of age), resulting in a maximum blood Pb level of 33 ug/dL, and
(3) lifetime (PNDO up to 10 years of age), resulting in a maximum blood Pb level of
32 ug/dL (Foster et al., 1998; Singh et al.. 1993a). Rodent studies using i.p. injections of
Pb also showed ultrastructural damage to structures involved in spermatogenesis (blood
Pb level after i.p. injection treatment for 16 days: 7.4 ug/dL) (Murthy et al.. 1995). More
recently, Salawu et al. (2009) observed a decrease in absolute testicular weight after Pb
exposure (adult male SD rats, 10,000 ppm Pb acetate in drinking water for 8 weeks).
Anjum et al. (2010) reported decreased testicular and epididymal weights of male rats
exposed to Pb acetate (500 or 1,500 ppm Pb acetate in drinking water for 45 days) which
were significantly attenuated with Pb co-exposure to the herb Centella asiatica (Sainath
et al.. 2011). Pb induced morphological abnormalities in sperm in a concentration-
dependent manner (Allouche et al.. 2009; Oliveira et al.. 2009; Salawu et al.. 2009; Shan
et al.. 2009; Tapisso et al.. 2009; Massanyi et al.. 2007; Wang et al.. 2006a). Sperm
abnormalities reported after Pb exposures were amorphous sperm head, abnormal tail,
and abnormal neck. Dong et al. (2009) reported decreased epididymis and body weights
in mice after an 8-week exposure to 6,000 ppm Pb acetate in drinking water (adult male
Kunming mice, 8-week exposure). Oliveira et al. (2009) observed a negative correlation
between Pb dose and intact acrosomes (8 week old ICR-CD1 mice, subcutaneous
injection of 74 and 100 mg PbCl2/kg body weight for four consecutive days). Rubio et al.
4-665
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(2006) (adult mice treated with i.p. injections of 8, 16 or 24 mg/kg of Pb acetate for
35 days) and Biswas and Ghosh (2006) (adult male Wistar rats, i.p. injection of 8 mg/kg
body weight Pb acetate for 21 days) also observed a Pb-induced decrease in seminal
vesicle and ventral prostate weights. Rubio et al. (2006) reported that Pb acetate, in a
concentration-dependent manner (8-24 mg/kg body weight), reduced the length of certain
stages of the spermatogenic cycle of rat seminiferous tubules and thus affected
spermatogenesis. Oral Pb acetate exposure (25 mg/kg bw in drinking water for 3 months,
resulting in blood Pb level of 5.3 (ig/dL) to adult male albino rats produced significant
histological seminiferous tubule damage (epithelium, spermatocytes, acrosomes) that was
attenuated with ascorbic acid treatment (Pb exposure +100 mg/kg bw/day ascorbic acid,
resulting in blood Pb level of 4.7 (ig/dL) (El Shafai et al.. 2011). However, the majority
of studies did not observe a statistically significant difference in body weight or
reproductive organ weights after Pb exposure at the doses used in the studies. Not all of
the aforementioned studies observed changes in every parameter. This may be due to the
use of different strains or species, chemical form of the Pb compound administered,
dosage schedule, duration of exposure, and age of animals at the time of the study
(Oliveira et al.. 2009).Data from recent studies suggested that the generation of ROS in
the male reproductive tissues, which can then affect antioxidant defense systems of cells
(Pandya et al.. 2010) (adult male Charles Foster rats, Pb acetate 0.025 mg/kg body
weight/day i.p. for 8 weeks) contributes to the mode of action of Pb damage to the male
reproductive organs and sperm or semen. Salawu et al. (2009) observed a statistically
significant increase in malondialdehyde (MDA, oxidative stress marker) and a significant
decrease in the activity of antioxidant enzymes superoxide dismutase (SOD) and catalase
(CAT) in plasma and testes of adult male Sprague Dawley rats after administration of
10,000 ppm Pb acetate in drinking water for 8 weeks. Supplementation with tomato paste
(used as a source of antioxidants) reduced Pb-induced ROS production and prevented the
Pb-induced increase in MDA formation and decrease in SOD and CAT activity.
Furthermore, co-treatment of Pb with substances that are known to have antioxidant
properties [i.e., tomato paste, Maca (Lepidium meyenii), and ascorbic acid] prevented the
Pb-induced reduction in sperm count, sperm motility, and sperm viability (Salawu et al.,
2009: Shan et al.. 2009: Madhavi et al.. 2007: Rubio et al.. 2006: Wang et al.. 2006a).
Recent studies continue to demonstrate that Pb may be directly toxic to mature
spermatozoa (adult Algerian mice, Pb acetate 21.5 mg/kg BW/every other day i.p. for 11
or 21 days and adult NMRI mice, 6,000 ppm Pb chloride in drinking water for 16 weeks,
respectively) (Tapisso et al.. 2009: Hernandez-Ochoa et al.. 2006) as well as primary
spermatocytes (adult male Wistar rats treated with drinking water containing 250mg/L
Pb acetate and adult albino rats, 10 mg/kg BW Pb chloride i.p. once daily for 8 weeks)
(Nava-Hernandez et al.. 2009: Rafique et al.. 2009). Nava-Hernandez et al. (2009)
exposed two groups of adult male rodents to Pb via drinking water (LI and L2, 250mg/L
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or 500 mg/L Pb acetate starting at PND60 for 90 days). They found significant increases
in spermatid DNA damage with Pb exposure. In their study, all Pb-treated animals had
blood Pb levels statistically significantly higher than controls (LI: 19.54 (ig/dL and
L2:21.90 (ig/dL); no statistically significant difference in blood Pb levels existed between
the two Pb exposure groups likely because the L2 group drank less water than did the LI
group. Piao et al. (2007) reported that Pb exposure caused DNA damage to sperm; the Pb
exposed group had a blood Pb of 6.7 (ig/dL. Piao et al. (2007) also examined the effect of
Zn supplementation on Pb-induced sperm aberrations and found that the proportion of
abnormal sperm was statistically significantly higher in the Pb group and the Pb+Zn
group (25 mg/kg Pb acetate i.p., 4 mg/kg Zn acetate i.p., both Pb acetate and Zn acetate,
once every two days, for 2 weeks) than in controls. However, the proportion of abnormal
sperm in Pb+Zn group was statistically significantly lower than in Pb alone group.
Hernandez-Ochoa et al. (2006) reported that Pb reaches the sperm nucleus in the
epididymis of mice chronically exposed (16 weeks in adult animals) to Pb (resulting in
mean blood Pb of 75.6 (ig/dL) by binding to nuclear sulfhydryl groups from the
DNA-protamine complex, increasing sperm chromatin condensation, and thereby
interfering with the sperm maturation process without altering sperm quality parameters.
Tapisso et al. (2009) observed a statistically significant increase in the number of
micronuclei and frequency of sister chromatid exchange with increasing treatment
duration in adult male Algerian mice administered 21.5 mg/kg body weight Pb acetate by
i.p. injection (every other day for 11 or 21 days). Nava-Hernandez (2009) reported a
concentration-dependent increase in DNA damage in rat primary spermatocytes after a
13-week exposure period to Pb acetate in drinking water (resulting in mean blood Pb
levels between 19.5 and 21.9 (ig/dL). Rafique et al. (2009) reported degenerative changes
from pyknosis to apoptosis in primary spermatocytes (adult albino rats, 10 mg/kg BW
Pb chloride i.p. once daily for 8 weeks). Hepatic expression of spermatogenic genes was
transiently down-regulated in 8 week old male Wistar-Kyoto (WKY) rats in response to
Pb nitrate (100 (imol single i.v. injection) 3 hours after injection and recovered to
baseline by 12 hours (Nemoto et al.. 2011): this effect was not seen in the stroke-prone
spontaneously hypertensive rats, which are from a WKY background, or in Sprague-
Dawley rats, demonstrating strain specificity.
Pb-induced apoptosis in germ cells within the seminiferous tubules is another suggested
mechanism by which Pb exerts its toxic effects on sperm production and function (Wang
et al.. 2006a) (Kunming mice, 2,000 ppm Pb acetate in drinking water for 14-42 days).
Dong et al. (2009) reported a exposure concentration-related increase in apoptosis in
spermatogonia and spermatocytes of Kunming mice after exposure to 1,500-6,000 ppm
Pb acetate in drinking water. Pb-induced testicular germ cell apoptosis was associated
with up-regulation of genes involved in the signal pathway of MAPK and death receptor
signaling pathway of Fas. For instance, up-regulation of K-ras and Fas expressions was
4-667
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concomitant with activation of c-fos and active caspase-3 proteins. Wang et al. (2006a)
observed a Pb exposure concentration-dependent increase in the expression of apoptotic
markers TGF(31 and caspase-3 in spermatogenic cells, Sertoli cells, and Leydig cells.
Shan et al. (2009) (20 mg/kg BW intragastric Pb acetate for 6 weeks) also reported a
statistically significant increase in mRNA expression and protein levels of Fas, Fas-L and
caspase-3 after Pb treatment. Supplementation with ascorbic acid inhibited or reduced the
Pb-induced apoptosis in germ cells and protected testicular structure and function (El
Shafai etal.. 2011; Shan et al.. 2009; Wang et al.. 2006a) suggesting ROS generation is a
major contributing factor in decreased male fertility observed after chronic Pb exposure.
Similar to the results summarized in previous Pb AQCDs, recent epidemiologic and
toxicological studies indicate that Pb exposure has effects on sperm, semen, and male
reproductive organs. Consistent toxicological evidence from multiple labs with multiple
species of animals showed decrements in sperm or semen quality with Pb exposure
including decreased sperm counts, decreased sperm production rate, and a dose-
dependent suppression of spermatogenesis. Histological damage to rodent sperm and
ultrastructural damage to rodent and non-human primate seminiferous tubules has been
reported with Pb exposure. Sperm from Pb-exposed male rodents used for in vitro
fertilization of eggs from unexposed females yielded a lower rate of fertilization. Also,
direct effects of Pb on rodent sperm DNA have been reported in rodents with drinking
water exposure. The toxicological findings cross species and are seen in wildlife
including deer, earthworms, rainbow trout, marine worms, and fathead minnow.
In studies of men exposed to Pb in occupational settings, associations were observed
between blood Pb levels of at least 25 (ig/dL and sperm count and quality. Multiple
epidemiologic studies of occupational cohorts included control populations with high
blood Pb levels (close to or greater than 10 (ig/dL), which makes identification of effects
at lower levels difficult. Occupational studies had limited consideration for potential
confounding factors, such as other workplace exposures. An epidemiologic study of men
attending a clinic for purposes of infertility exam or semen donation demonstrated an
inverse relationship between Pb levels and sperm and semen quality (Telisman et al..
2007). This study also controlled for other metals in the analyses. Other studies of men at
infertility clinics had greater imprecision in their estimates, less control for confounding
(such as other metals), and/or small sample sizes. Additionally, studies limited to men at
infertility clinics may suffer from selection bias and are not generalizable.
4.8.3.2 Effects on Hormone Levels
The 2006 Pb AQCD (U.S. EPA. 2006b) provided evidence that Pb acts as an endocrine
disrupter in males at various points along the hypothalamic-pituitary-gonadal axis. The
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2006 Pb AQCD also reported inconsistencies in the effects of Pb exposure on circulating
testosterone levels. Recent epidemiologic and toxicological studies are reported below.
Epidemiologic studies are summarized in Table 4-44. Epidemiologic studies were cross-
sectional; biological samples used for the measurement of Pb were measured
concurrently with hormone levels. One study estimated cumulative blood Pb.
4-669
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Table 4-44 Summary of recent epidemiologic studies of associations between Pb exposure indicators and
hormones for males.
Study
Population
Study Location Methodological Data for Pb Biomarker or
Reference3 Outcome Years Details other Exposure Indicator
Telisman et FSH, LH, Men aged 19-55 Cross-sectional Blood Pb (ug/dl_)
al. (2007) testosterone, yr, never study using
estradiol, occupationally linear multiple „ ,. . „„ , „ „_ „„ „„,
prolactin exposed to regression Median: 4.92 (range 1.13-14.91)
metals and going
to a clinic for
infertility
examination or
f n r Q P m P n
IUI oCI I Id I
donation to be
used for artificial
insemination
N=240
Croatia
2002-2005
Adjusted Effect Estimates
Standardized regression coefficients
for log blood Pb (units not given)
Testosterone: 0.21 (p-value <0.003)
Estradiol: 0.22 (p-value <0.0008)
Prolactin: - 0.18 (p-value <0.007)
Note: Coefficients and p-values not
given if not statistically significant
(LH, FSH)
Potential
Confounders
Adjusted for
in Analysis
Age, smoking
status,
alcohol use,
Cd, Cu, Zn,
Se
An interaction
between Pb
and Cd was
included in
models for
testosterone
4-670
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Table 4-44 (Continued): Summary of recent epidemiologic studies of associations between Pb exposure indicators and
hormones for males.
Reference3 Outcome
Study
Population
Study Location
Years
Methodological
Details
Data for Pb Biomarker or
other Exposure Indicator
Adjusted Effect Estimates
Potential
Confounders
Adjusted for
in Analysis
Meeker et
al. (2010)
FSH, LH,
inhibin B,
testosterone,
SHBG, FAI,
testosterone/LH
Men aged
18-55 yr going to
infertility clinics
(distinction not
made between
clinic visits for
male or female
fertility issues)
N=219
Michigan
NS
Cross-sectional
study using
multiple linear
regression
Blood Pb (ug/dl_)
Median: 1.50 (IQR 1.10, 2.00)
Regression coefficients (95% Cl)
FSH
1st quartile: 0 (ref)
2ndquartile: 0.13 (-0.10, 0.37)
3rd quartile: 0.10 (-0.15, 0.35)
4th quartile: 0.07 (-0.18, 0.31)
LH
1st quartile: 0 (ref)
2nd quartile: 0.004 (-0.20, 0.21)
3rd quartile: 0.13 (-0.09, 0.35)
4th quartile: 0.88 (-0.14, 0.29)
Age, BMI,
current
smoking
Considered
but did not
include: race,
income,
season
Inhibin B
1st quartile: 0 (ref)
2nd quartile:-6.45 (-27.2, 14.3)
3rd quartile:-4.62 (-26.6, 17.4)
4th quartile: -7.79 (-29.0, 13.4)
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Table 4-44 (Continued): Summary of recent epidemiologic studies of associations between Pb exposure indicators and
hormones for males.
Reference3 Outcome
Study
Population
Study Location
Years
Methodological
Details
Data for Pb Biomarker or
other Exposure Indicator
Adjusted Effect Estimates
Potential
Confounders
Adjusted for
in Analysis
Meeker et
al. (2010)
(Continued)
FSH, LH,
inhibin B,
testosterone,
SHBG, FAI,
testosterone/LH
Men aged
18-55 yr going to
infertility clinics
(distinction not
made between
clinic visits for
male or female
fertility issues)
N=219
Michigan
NS
Cross-sectional
study using
multiple linear
regression
Blood Pb (ug/dl_)
Median: 1.50 (IQR 1.10, 2.00)
Testosterone
1st quartile: 0 (ref)
2nd quartile: 28.6 (-6.82, 64.1)
3rd quartile: 15.8 (-21.8, 53.3)
4th quartile: 39.9 (3.32, 76.4)
SHBG
1st quartile: 0 (ref)
2nd quartile: -0.01 (-0.16, 0.15)
3rd quartile: 0.04 (-0.12, 0.21)
4th quartile: 0.07 (-0.10, 0.23)
FAI
1st quartile: 0 (ref)
2nd quartile: 0.8 (-0.04, 0.20)
3rd quartile: 0.03 (-0.10, 0.17)
4th quartile: 0.08 (-0.05, 0.21)
Testosterone /LH
1st quartile: 1.00 (ref)
2nd quartile: 0.07 (-0.16, 0.30)
3rd quartile: -0.05 (-0.29, 0.19)
4th quartile: 0.07 (-0.17, 0.31)
Age, BMI,
current
smoking
Considered
but did not
include: race,
income,
season
4-672
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Table 4-44 (Continued): Summary of recent epidemiologic studies of associations between Pb exposure indicators and
hormones for males.
Reference3 Outcome
Study
Population
Study Location
Years
Methodological
Details
Data for Pb Biomarker or
other Exposure Indicator
Adjusted Effect Estimates
Potential
Confounders
Adjusted for
in Analysis
Mendiola et FSH, LH,
al. (2011) testosterone
Men attending
infertility clinics
and classified as
either normal
sperm (controls)
or oligo-astheno-
teratozoospermia
(cases)based on
WHO semen
quality criteria
NCases=30
NControls=31
Spain
2005-2007
Case-control
study using
multiple linear
regression
Mean (IQR) (ug/dl_)
Seminal plasma: 2.90 (2.70,
3.20)
Whole blood: 9.50(7.50, 11.90)
Blood plasma: 2.90 (2.70, 3.10)
Mean (SD)
Cases:
Seminal plasma: 3.0 (0.3)
Whole blood: 9.8 (2.3)
Blood plasma: 2.9 (0.2)
Controls:
Seminal plasma: 2.9 (0.3)
Whole blood: 9.7 (2.3)
Blood plasma: 2.9 (0.3)
Linear regression (3 (95% Cl)
FSH
Seminal plasma: 0.05 (-0.24, 0.39)
Whole blood: 0.04 (-0.03, 0.04)
Blood plasma: -0.20 (-0.64, 0.25)
LH
Seminal plasma: 0.14 (-0.13, 0.41)
Whole blood: 0.05 (-0.05, 0.07)
Blood plasma: -0.07 (-0.49, 0.31)
Testosterone
Seminal plasma: 0.11 (-0.10, 0.31)
Whole blood: 0.01 (-0.05, 0.02)
Blood plasma: -0.12 (-0.40, 0.14)
*Units not given (assume 1 ug/dL)
Age, BMI,
number of
cigarettes/day
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Table 4-44 (Continued): Summary of recent epidemiologic studies of associations between Pb exposure indicators and
hormones for males.
Reference3 Outcome
Hsiehetal. FSH, LH,
(2009a) testosterone,
inhibin B
Study
Population
Study Location
Years
Workers at a
Pb-acid battery
factory with
annual blood Pb
measures
N=181
Methodological Data for Pb Biomarker or
Details
Longitudinal
occupational
cohort study
using
multivariate
linear
regression
other Exposure Indicator
Concurrent blood Pb:
<10ug/dL: 11.6%
>40ug/dL: 17.1%
Cumulative blood Pb: Not
reported
Time-weighted cumulative blood
Pb:
Not reported
Adjusted Effect Estimates
(3 from linear regression
Inhibin B
Concurrent blood Pb: 0.40 (p-value
OA f\\
.40)
Cumulative blood Pb: 0.05 (p-value
0.02)
Time-weighted cumulative blood Pb:
1.33 (p-value 0.007)
Potential
Confounders
Adjusted for
in Analysis
LH, FSH,
testosterone,
age, smoking
status,
alcohol use,
BMI
1991-NS
Pearson's correlations analysis
detected no correlations between
concurrent blood Pb levels and FSH,
LH, or testosterone. Cumulative
blood Pb levels were correlated with
FSH and LH, but not testosterone.
Time-weighted cumulative blood Pb
levels were correlated with LH, but
not FSH or testosterone.
4-674
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Table 4-44 (Continued): Summary of recent epidemiologic studies of associations between Pb exposure indicators and
hormones for males.
Reference3 Outcome
Study
Population
Study Location
Years
Methodological
Details
Data for Pb Biomarker or
other Exposure Indicator
Adjusted Effect Estimates
Potential
Confounders
Adjusted for
in Analysis
Naha and
Manna
(2007)
FSH, LH,
testosterone
Non-
occupationally
exposed controls
and
occupationally
exposed workers
NControls=50
N low exposure"^U
Nhigh exposure~^0
Bangalore, India
NS
Occupational
cohort study
using ANOVA,
Student's t-test,
and Scheffe's F-
test
Categorized by work history as
controls, low exposure (7-10 yr
of exposure for 8 h/day) and
high exposure (>10 yr of
exposure for 8 h/day)
Mean (SD)
Blood Pb measurement
Controls 10.25 (2.26)
Low exposure 50.29 (3.45)
High exposure 68.26 (2.49)
Semen Pb measurement
Controls 2.99 (0.76)
Low exposure 15.85 (1.95)
High exposure 25.30 (2.28)
Mean FSH (SD)
Control: 2.69 (1.22)
Low exposure: 2.58 (1.94)
High exposure: 2.16 (0.99)
p-values for difference >0.05
Mean LH (SD)
Control: 5.14(2.35)
Low exposure: 4.27 (2.52)
High exposure: 3.9 (1.69)
p-values for difference >0.05
Mean Testosterone (SD)
Control: 5.24 (2.40)
Low exposure:4.83 (1.21)
High exposure: 4.59 (1.27)
p-values for difference >0.05
None
aStudies are presented in order of first appearance in the text of this section, which is based on study design and methodological strength.
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Hormone levels were measured in a few recent epidemiologic studies. In a study of men
non-occupationally exposed to Pb in Croatia, increased blood Pb level was associated
with increasing serum testosterone and estradiol but decreasing serum prolactin
(Telisman et al.. 2007). In addition, the analysis of an interaction term for blood Pb and
blood Cd levels demonstrated a synergistic effect on increasing serum testosterone levels.
No association was observed between blood Pb and FSH or LH. This study controlled for
multiple potential confounders, including other metals. Among men recruited from
infertility clinics in Michigan, median blood Pb levels were much lower than those
observed in the other studies of Pb and hormone levels among men (Meeker et al., 2010).
No association was detected between blood Pb and levels of FSH, LH, inhibin B, sex
hormone-binding globulin (SHBG), free androgen index (FAI), or a measure of Leydig
cell function (testosterone/LH). A positive association between the highest quartile of
blood Pb and testosterone was present, but this association did not persist when other
metals were included in the model. Similarly, another study of men recruited from
infertility clinics observed no association between Pb concentrations from seminal
plasma, whole blood, or blood plasma and FSH, LH, or testosterone (Mendiola et al..
2011).
A study of occupationally-exposed men in Taiwan reported an association between
measures of cumulative blood Pb levels and inhibin B levels, but no association was
detected with concurrent blood Pb levels (Hsieh et al.. 2009a). A correlation between
cumulative blood Pb measures and LH levels was detected but correlations were not
present with FSH or testosterone levels. No correlations were apparent between FSH,
LH, or testosterone and current blood Pb levels. Another occupational study of men with
high blood Pb levels reported no difference in serum FSH, LH, and testosterone among
the three groups (controls: mean blood Pb 10.25 (ig/dL, low exposure: mean blood Pb
50.29 (ig/dL, high exposure: mean blood Pb 68.26 (ig/dL) (Naha and Manna. 2007). This
study did not assess any potential confounding factors.
In a recent toxicological study, Rubio et al. (2006) observed a decrease in testosterone
levels in Pb acetate-treated rats in an exposure concentration-related fashion (8-24 mg/kg
body weight), and this decrease correlated with reduced lengths of spermatogenic cycle
stages VII-VIII (spermiation) and IX-XI (onset of spermatogenesis). Anjum et al. (2010).
who dosed 50 day old male rats with 273 or 819 mg/L Pb acetate in drinking water (500
or 1,500 ppm, respectively; blood Pb not reported), found significant decreases in serum
testosterone and testicular 3(3-HSD and 17(3-HSD levels in Pb-exposed animals versus
controls. Pandya et al. (2010) reported altered hepatic steroidogenic enzyme activity.
Pillai et al. (2012) found gestational and lactational treatment with Pb acetate in Charles
Foster rats (subcutaneous injection of 0.05 mg/kg BW/day, blood Pb not reported)
induced significant decreases in testicular 17(3-HSD and serum testosterone. Biswas and
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Ghosh (2006) reported a Pb-induced decrease in serum testosterone and gonadotropins
(FSH, LH) with inhibition of spermatogenesis, however, there was a statistically
significant increase in the adrenal steroidogenic enzyme, A5-3(3-HSD, activity and serum
corticosterone levels indicating disruption of the adrenocortical process. Exposure
concentration-dependent decreases in serum testosterone were reported in Pb-exposed
male rats (Anjum et al.. 2010). In contrast, Salawu et al. (2009) did not observe a
decrease in serum testosterone between control animals and animals administered
10,000 ppm Pb acetate in drinking water for 8 weeks. Allouche et al. (2009) not only did
not observe any statistically significant changes in serum FSH or LH, but reported an
increase in serum testosterone levels after 500-3,000 ppm Pb acetate treatment in
drinking water (only statistically significant in animals administered 500 ppm Pb acetate).
The results of these recent studies further support the theory that compensatory
mechanisms in the hypothalamic-pituitary-gonadal axis may allow for the adaptation of
exposed animals to the toxic endocrine effects of Pb (Rubio et al.. 2006; U.S. EPA.
2006b).
Overall, recent epidemiologic and toxicological studies report mixed findings regarding
hormone aberrations in males associated with Pb exposure or Pb biomarker levels. These
results are similar to those from the 2006 Pb AQCD (U.S. EPA. 2006b) on the effects of
Pb exposure on circulating testosterone levels. Epidemiologic studies are limited by their
sample populations, often occupational cohorts or men at infertility clinics, which may
not be generalizable. Occupational cohorts may have other exposures that confound the
associations, and studies at infertility clinics are subject to selection bias. A few of the
recent epidemiologic studies include important confounding factors, such as smoking, but
other factors, such as exposure to other metals, were often absent. Additionally, most
studies examine concurrent Pb biomarker and hormone levels which may not reflect
changes resulting from long-term exposures, as demonstrated by the longitudinal
occupational cohort study. Further, in cross-sectional studies, the temporality of effects
cannot be established.
4.8.3.3 Effects on Fertility
Epidemiologic studies have been performed comparing Pb and infertility in men. The
SMART study is a longitudinal study that examined the success of in vitro fertilization
(IVF) treatment for women and their partners starting their first round of treatment
(Bloom et al.. 20 lib; Bloom et al.. 2010). A small number of the male partners
participated (n=16). Their mean (SD) blood Pb level was 1.50 (0.80) (ig/dL. Higher blood
Pb levels, measured at the time of oocyte retrieval in female partners, were associated
with greater oocyte fertilization (OR 1.08 [95% CI: 0.97, 1.21] per 1 (ig/dL increase in
4-677
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blood Pb when adjusted for Cd, Hg, age, cigarette smoking, race/ethnicity, and
creatinine), which is not the expected direction (Bloom et al.. 2010). However, higher
blood Pb was associated with lower embryo cell number, a predictor of IVF success, and
with higher embryo fragmentation score, an inverse predictor of IVF success (OR for
embryo cell number: 0.58 [95% CI: 0.37, 0.91]; OR for embryo fragmentation score: 1.47
[95% CI: 1.11, 1.94] per 1 (ig/dL, controlled for age, race/ethnicity, cigarette smoking,
creatinine, Cd, and Hg, plus day of embryo transfer for embryo cell number) (Bloom et
al.. 20lib). A case-control study conducted in Turkey reported that concurrent blood and
seminal plasma Pb levels were different in fertile (n=45; mean [SD] blood Pb: 23.16
[5.59] (ig/dL) and infertile men (n=50; mean [SD] blood Pb: 36.82 [12.30] (ig/dL)
(p <0.001, ANOVA) (Kiziler et al.. 2007). There was no control for potential
confounding factors although the relationship persisted when limited to non-smokers.
Another case-control study examined occupational Pb exposure (determined by self-
report of occupational exposure in the past month) and detected no difference in reported
exposure for infertile (n=650) versus fertile men (n=698) (unadjusted OR 0.95 [95% CI:
0.6, 1.6]) (Gracia et al.. 2005). Blood Pb was not measured but approximately 5.0% of
infertile men and 5.3% fertile men reported occupational exposure to Pb. A limitation
present in these studies is that the cases included are men who are seeking help at fertility
clinics; the study populations are not a sample of the general population regarding
fertility. The results could be biased due to the recruitment of individuals going to an
infertility clinic, who may be different than individuals suffering from infertility without
knowing it or without going to a clinic.
Recent animal toxicology studies assessed paternal-mediated reproductive fitness by
examining the reproductive success of Pb-exposed males with non-exposed control
females. Anjum et al. (2010) found that adult male rats that were exposed to 273 or
819 mg/L Pb acetate in drinking water (500 or 1,500 ppm, respectively; blood Pb not
reported) spent a significantly longer time copulating than did their control littermates.
The Pb-exposed males were less successful copulators with only 73% of the 500 ppm
Pb acetate exposed males, and 53% of the 1,500 ppm exposed males generating
copulatory plugs in the unexposed female mates. While the number of pregnant females
did not significantly differ from controls, Pb exposed males contributed to the formation
of significantly fewer implantations/dam, and significantly fewer fetuses/dam.
Pb-exposed males were able to sire offspring, but produced fewer offspring per litter. In a
group of males rats with co-exposure to Pb and the herb Centella asiatica, these
reproductive decrements were attenuated relative to rats exposed to Pb alone (adult albino
male rats, 1,500 ppm Pb acetate in drinking water for 70 days) (Sainath et al.. 2011).
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Overall, large, well-conducted epidemiologic studies of Pb exposure and fertility in males
are lacking. The few available studies reported inconsistent findings. Toxicological
studies demonstrated paternal associated subfecundity (fewer pups sired per pregnancy)
with altered mating behavior (longer time spent copulating), albeit in studies with no
blood Pb levels reported. Supplementation with antioxidants in a separate study showed
restoration of this subfecundity, possibly contributing mode of action support to this Pb-
induced decreased fertility in male rodents.
4.8.3.4 Effects on Morphology and Histology of Male Sex Organs
Recent toxicological studies further support historical findings that showed an association
between Pb exposure and changes in the sex organs as well as germ cells. Histological
changes in testes of Pb nitrate-treated adult animals (a single i.p. dose of 12.5, 25, or
50 mg/kg of BW and sacrifice 48 hours later) included seminiferous tubule atrophy,
Sertoli cell and Leydig cell shrinkage with pyknotic nuclei (Shan et al.. 2009; Wang et
al.. 2006a). dilatation of blood capillaries in the interstitium, undulation of basal
membrane, and occurrence of empty spaces in seminiferous epithelium (adult male
Wistar rats, single i.p. dose of 12, 25 or 50 mg/kg BW Pb acetate) (Massanyi et al..
2007). Pillai et al. (2010) found gestational and lactational treatment with Pb acetate in
Charles Foster rats (subcutaneous injection of 0.05 mg/kg BW/day) induced significant
decreases in absolute organ weight (testes and epididymis) and significant decreases in
relative epididymal weight. Anjum et al. (2010). who exposed 50 day old male albino
Wistar/NIN rats to Pb acetate (273 or 819 mg/L in drinking water, 500 or 1,500 ppm,
respectively, blood Pb levels not reported) for 45 days, reported significant decreases in
relative reproductive organ weight (epididymis, testis, vas deferens, and seminal vesicle)
in Pb-exposed animals.
4.8.3.5 Summary of Effects on Male Reproductive Function
Evidence of associations between Pb exposure and male reproductive function varies by
outcome. The strongest evidence of an association is the relationship observed between
Pb and negative effects on sperm and semen in both recent epidemiologic and
toxicological studies and studies reviewed in previous Pb AQCDs. Decrements in sperm
count, sperm production rate and semen quality were reported in animal toxicological
studies in rodents with drinking water Pb exposure [(Sokol and Berman. 1991; Sokol et
al.. 1985). blood Pb level 34-37 (ig/dL] and rabbits treated subcutaneously with Pb
[(Moorman et al.. 1998). blood Pb levels of 25 (ig/dLJ. Rodents exposed to Pb had direct
effects of Pb on sperm DNA, i.e., elevated levels of DNA damage [(Nava-Hernandez et
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al.. 2009). blood Pb levels 19 and 22 (ig/dL]. Histological or ultrastructural damage to the
male reproductive organs was reported in studies of rodents [(El Shafai et al.. 2011).
blood Pb level 5.1 (ig/dL] and non-human primates [(Singh et al., 1993a). blood Pb level
43 (ig/dL]. Subfecundity has been reported in unexposed females mated to Pb exposed
males (decreased number of pups born/litter). Also, sperm from Pb-exposed rats (blood
Pb 33 to 46 ug/dL) used for in vitro fertilization of eggs harvested from unexposed
females yielded lower rates of fertilization (Sokol et al., 1994). Many of the
epidemiologic studies included occupational cohorts, which had high blood Pb levels
(> 25(ig/dL), or men attending infertility clinics, which have potential selection bias.
Additionally, control for potential confounding factors, such as other workplace
exposures, was not often performed. A study of men (who were attending a clinic for an
infertility exam or to donate semen for use in artificial insemination) did control for
multiple factors, including other metals and smoking status, and reported an association
between blood Pb levels and some indicators of poor sperm quality (Telisman et al..
2007). Recent toxicological studies also reported an association between Pb exposure and
decreases in reproductive organ weight, organ histological changes in the testes and germ
cells. Male rats exposed to Pb also showed subfecundity, in that they produced smaller
litters when mated with unexposed females (Anjum et al.. 2010). Further coherence for
these findings in laboratory animal models is found with findings in the ecological
literature for the effects of Pb exposure on reduced fecundity in terrestrial and aquatic
animal species (Sections 6.4.5.2. 6.3.4.2. and 6.4.5.3). Similar to the 2006 Pb AQCD
(U.S. EPA. 2006b). recent epidemiologic and toxicological studies reported inconsistent
results regarding hormone aberrations associated with Pb exposure. Mixed findings were
also apparent among epidemiologic studies of fertility among men.
4.8.4 Effects on Female Reproductive Function
The epidemiologic studies on Pb and female reproductive function presented in the
2006 Pb AQCD (U.S. EPA. 2006b) provided little evidence for an association between
Pb biomarkers and effects on female reproduction and fertility. However, the 1986 and
2006 Pb AQCDs (U.S. EPA. 2006b. 1986a) reported toxicological findings that Pb
exposure was associated with effects on female reproductive function that can be
classified as alterations in female sexual maturation, effects on fertility and menstrual
cycle, endocrine disruption, and changes in morphology or histology of female
reproductive organs including the placenta. Since the 2006 Pb AQCD, many
epidemiologic studies have been published regarding Pb biomarker levels in women and
reproductive effects. The studies presented in the following text and tables are grouped
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by study design and methodological strength. In addition, recent toxicological studies add
further knowledge of Pb-related effects on the female reproductive system.
4.8.4.1 Effects on Female Sex Endocrine System and Estrus
Cycle
Multiple epidemiologic studies have examined the association between blood Pb levels
and hormone levels and the estrus cycle. Recent epidemiologic studies (characterized in
Table 4-45) were cross-sectional in design, analyzing measures of Pb and hormones that
were collected either concurrently or close in time. These studies support the
toxicological findings, which are the major body of evidence on endocrine effects of Pb.
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Table 4-45 Summary of recent epidemiologic studies of associations between blood Pb levels and hormones
for females.
Reference3
Outcome
Study Population
Study Location
Years
Methodological
Details
Blood Pb
Mean
(SD)
in ug/dL
Adjusted Effect Estimates
Potential Confounders
Adjusted for in
Analysis
Krieg (2007) FSH, LH
Women aged 35-60 yr
from the NHANES III
study
N=3,375
U.S.
1988-1994
Cross-sectional
study using
linear regression
2.8 Linear regression slope (95% Cl)
for log-transformed Pb
FSH:
Post-menopausal
22.2(13.5, 30.8)
Pregnant
0.1 (-0.1, 0.3)
Menstruating at time of exam
2.1 (-2.1,6.3)
Both ovaries removed
32.6(10.1, 55.1)
Birth control pills being used
-6.3 (-10.0,-2.5)
Pre-menopausal
8.3(3.8, 12.7)
LH:
Post-menopausal
6.2(3.0, 9.5)
Pregnant
-0.8 (-1.9, 0.4)
Menstruating at time of exam
-0.3 (-1.8, 1.3)
Both ovaries removed
10.0(1.1, 18.9)
Birth control pills being used:
-0.6 (-2.9, 1.6)
Pre-menopausal
1.7 (-0.6, 4.1)
Age, total bone mineral
density, serum cotinine,
alcohol use, current
breast feeding,
hysterectomy, one ovary
removed, Norplant use,
radiation or
chemotherapy, hormone
pill use, vaginal cream
use, hormone patch use
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Table 4-45 (Continued): Summary of recent epidemiologic studies of associations between blood Pb levels and hormones for
females.
Reference3
Pollack et al.
(2011)
Outcome
FSH, estradiol,
LH,
progesterone,
and cycle length
Study Population
Study Location
Years
Healthy,
premenopausal women
aged 18-44 yrwith
menstrual cycle length
of 21-35 days, BMI of
18-35 kg/m , not
recently using birth
control, not planning to
become pregnant, and
not breast feeding
N=252
Blood Pb
Mean
Methodological (SD)
Details in ug/dL
Longitudinal 0.93
cohort using |QR: 0.68,
nonlinear mixed -| 20
models with
harmonic terms
and weighted
linear mixed
models
Adjusted Effect Estimates
Mean % Estradiol
0.30-0.72 ug/dl_: Ref
0.73-1. 10ug/dL: 8.2 (-1.2, 18.6)
1.1 1-6.20 ug/dl_: 4.7 (-4.7, 15.2)
Amplitude Estradiol
0.30-0.72 ug/dl_: Ref
0.73-1.10 ug/dl_: -0.01 (-0.06, 0.04)
1.11-6.20 ug/dl_: -0.02 (-0.7, 0.03)
Potential Confounders
Adjusted for in
Analysis
Age, BMI, race
Also examined, but did
not include: smoking,
income, education,
physical activity, parity,
dietary Fe, fish
consumption, shellfish
consumption, vegetable
consumption, total
calories
Buffalo, NY
2005-2007
Phase Shift Estradiol
0.30-0.72 ug/dl_: Ref
0.73-1.10 ug/dl_: -0.09 (-0.24, 0.05)
1.11 -6.20 ug/dl_: 0.14 (-0.01, 0.29)
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Table 4-45 (Continued): Summary of recent epidemiologic studies of associations between blood Pb levels and hormones for
females.
Reference3
Pollack et al.
(2011)
(Continued)
Outcome
FSH, estradiol,
LH,
progesterone,
and cycle length
Study Population
Study Location
Years
Healthy,
premenopausal women
aged 18-44 yrwith
menstrual cycle length
of 21 -35 days, BMI of
18-35 kg/m , not
recently using birth
control, not planning to
become pregnant, and
not breast feeding
N=252
Blood Pb
Mean
Methodological (SD)
Details in ug/dL
Longitudinal 0.93
cohort using |QR: 0.68,
nonlinear mixed -| 20
models with
harmonic terms
and weighted
linear mixed
models
Adjusted Effect Estimates
Mean % FSH
0.30-0.72 ug/dl_: Ref
0.73-1. 10ug/dL: 8.0 (-0.9, 17.7)
1.11 -6.20 ug/dl_: 3.6 (-5.3, 13.3)
Amplitude FSH
0.30-0.72 ug/dl_: Ref
0.73-1.10 ug/dl_: -0.01 (-0.03, 0.02)
1.1 1-6.20 ug/dl_: -0.02 (-0.04, 0.01)
Potential Confounders
Adjusted for in
Analysis
Age, BMI, race
Also examined, but did
not include: smoking,
income, education,
physical activity, parity,
dietary Fe, fish
consumption, shellfish
consumption, vegetable
consumption, total
calories
Buffalo, NY
2005-2007
Phase Shift FSH
0.30-0.72 ug/dl_: Ref
0.73-1.10 ug/dl_: -0.06 (-0.25, 0.12)
1.11-6.20 ug/dl_: -0.02 (-0.21, 0.18)
Mean % LH
0.30-0.72 ug/dl_: Ref
0.73-1.10ug/dL: 5.1 (-5.1, 16.4)
1.11-6.20 ug/dl_: -0.5 (-10.5, 10.7)
Amplitude LH
0.30-0.72 ug/dl_: Ref
0.73-1.10 ug/dl_: -0.01 (-0.03, 0.02)
1.11-6.20 ug/dl_: -0.02 (-0.04, 0.01)
Phase Shift LH
0.30-0.72 ug/dl_: Ref
0.73-1.10 ug/dl_: -0.16 (-0.36, 0.03)
1.11-6.20 ug/dl_: -0.11 (-0.32, 0.10)
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Table 4-45 (Continued): Summary of recent epidemiologic studies of associations between blood Pb levels and hormones for
females.
Reference3
Pollack et al.
(2011)
(Continued)
Outcome
FSH, estradiol,
LH,
progesterone,
and cycle length
Study Population
Study Location
Years
Healthy,
premenopausal women
aged 18-44 yrwith
menstrual cycle length
of 21 -35 days, BMI of
18-35 kg/m , not
recently using birth
control, not planning to
become pregnant, and
not breast feeding
N=252
Blood Pb
Mean
Methodological (SD)
Details in ug/dL
Longitudinal 0.93
cohort using |QR: 0.68,
nonlinear mixed -| 20
models with
harmonic terms
and weighted
linear mixed
models
Adjusted Effect Estimates
Mean % Progesterone
0.30-0.72 ug/dl_: Ref
0.73-1. 10ug/dL: 7.5(0.1, 15.4)
1.1 1-6.20 ug/dl_: 6.8 (-0.8, 14.9)
Amplitude Progesterone
0.30-0.72 ug/dl_: Ref
0.73-1.10 ug/dl_: 0.07 (0.01, 0.15)
1. 11-6.20 ug/dl_: -0.06 (-0.13, 0.01)
Potential Confounders
Adjusted for in
Analysis
Age, BMI, race
Also examined, but did
not include: smoking,
income, education,
physical activity, parity,
dietary Fe, fish
consumption, shellfish
consumption, vegetable
consumption, total
calories
Buffalo, NY
2005-2007
Phase Shift Progesterone
0.30-0.72 ug/dl_: Ref
0.73-1.10 ug/dl_: 0.04 (-0.06, 0.15)
1.11 -6.20 ug/dl_: 0.15(0.05, 0.26)
Linear models (3 (95% Cl)
Estradiol: 0.03 (-0.05, 0.11)
FSH: -0.01 (-0.07, 0.06)
LH: 0.02 (-0.06, 0.10)
Progesterone: 0.06 (-0.04, 0.17)
OR (95% Cl) for anovulation per
1 ug/dL: 1.20(0.62,2.34)
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Table 4-45 (Continued): Summary of recent epidemiologic studies of associations between blood Pb levels and hormones for
females.
Reference3
Jackson et al.
(2011)
Chang et al.
(2006)
Study Population
Study Location
Outcome Years
FSH, estradiol, Healthy, pre-
LH, menopausal women
progesterone, aged 18-44 yr with
and cycle length menstrual cycle length
of 21-35 days, BMI of
18-35 kg/m , not
recently using birth
control, not planning to
become pregnant, and
not breast feeding
N=252
Buffalo, NY
2005-2007
Estradiol Women receiving care
at a infertility clinic in
2000-2001 or delivering
a normal infant at a
nearby medical center
in 1999
N=147
Methodological
Details
Longitudinal
cohort study
using linear
regression and
logistic
regression
Case-control
study using
mult ivari ate
linear regression
Blood Pb
Mean
(SD)
in ug/dL
Median:
0.87
IQR:
0.68, 1.20
3.12(0.19)
Potential Confounders
Adjusted for in
Adjusted Effect Estimates Analysis
Adjusted percent change (95% Cl) Cd, Hg, age, race
in serum hormone level for change /ethnicity
in blood Pb
FSH: -2.5 (-11. 2, 7.0)
Estradiol: 4.9 (-5.0, 15.9)
LH: 2.5 (-12.3, 19.9)
Progesterone: 4.6 (-12.2, 24.6)
Cycle length: 0.2 (-2.8, 3.3)
OR(95%CI)perunitPb
<25 day vs. 25-35 day cycle length:
0.9 (0.4, 2.3)
>35 day vs. 25-35 day cycle length:
0.5(0.1, 1.9)
Linear regression (3 (SE) per 1 ug/dL Not specified
Pb
1.18(0.60)
p-value: 0.049
Kaohsiung City, Taiwan
1999, 2000-2001
aStudies are presented in order of first appearance in the text of this section, which is based on study design and methodological strength.
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An epidemiologic study using the NHANES III data and including women aged
35-60 years old examined the relationship between blood Pb levels (mean 2.8 (ig/dL) and
serum FSH and LH (Krieg. 2007). Deviation from normal FSH and LH levels may
indicate endocrine disruption related to ovary functioning. Researchers found that higher
blood Pb levels were associated with higher levels of serum FSH and LH among both
postmenopausal women and women with both ovaries removed. There was also a trend
of increasing serum FSH with blood Pb levels for pre-menopausal women who were not
menstruating at the time of the exam or pregnant, although the association was not
statistically significant for LH. A limitation of this portion of the study is that FSH and
LH were measured without attention to day of a woman's menstrual cycle and LH and
FSH are known to vary throughout the cycle of non-menopausal, cycling women who are
not taking birth control pills. Higher blood Pb levels were associated with lower levels of
serum FSH among women taking birth control pills. The inverse association was also
present for LH, but it was not statistically significant. No associations between blood Pb
and FSH or LH were apparent for women who were menstruating at the time of the exam
or were pregnant. Further analysis indicated that the lowest level of blood Pb for which a
statistically significant association between blood Pb and FSH could be observed was
1.7 (ig/dL among women with their ovaries removed. For LH, the lowest level of blood
Pb for which a statistically significant association between blood Pb and LH could be
observed was 2.8 (ig/dL among postmenopausal women. Associations between hormones
and blood Pb level were also investigated using the BioCycle study cohort (Jackson et al..
2011; Pollack et al., 2011). These women were premenopausal with normal cycles and
not on birth control. Blood Pb was measured at enrollment and hormones were measured
multiple times throughout the menstrual cycle. Neither study detected an association
between unit change in blood Pb and hormone levels. However, in examination of tertiles
of Pb, women in the highest tertile blood Pb (1.11-6.20 (ig/dL) had higher mean
progesterone and longer length of a phase shift compared to women in the lowest tertile
(0.30-0.72 (ig/dL) (Pollack et al.. 2011). Other associations were observed but were not
statistically significant (Pollack et al.. 2011). No associations were detected for
anovulation (Pollack et al., 2011) or for cycle length (Jackson et al., 2011). Another
epidemiologic study was performed in Kaohsiung City, Taiwan among two groups of
women aged 23-44 years: those who were seeking help at a fertility clinic after one year
of trying to conceive, and those who had previously delivered an infant and were
identified from medical records of a postpartum care unit (Chang et al.. 2006). The mean
(SD) concurrent blood Pb level in this study was 3.12 (0.19) (ig/dL. The study reported a
positive association between blood Pb levels and serum estradiol concentrations during
the early follicular phase, which reflects ovary activity.
The effect of Pb exposure on the female endocrine system was demonstrated in
toxicological studies reviewed in the 1986 and 2006 Pb AQCDs (U.S. EPA. 2006b.
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1986a). However, the mechanism by which Pb affects the endocrine system has not been
fully elucidated. Several recent articles continue to demonstrate that Pb alters the
concentration of circulating hormones in female experimental animals. As mentioned in
the 2006 Pb AQCD, Pine et al. (2006) observed that maternal Pb exposure (during
gestation and lactation) caused a decrease in basal LH levels in pre-pubertal female
Fischer 344 rat pups as compared to control, non-Pb-exposed pups. Dumitrescu et al.
(2008a) observed an alteration of hormone levels in adult female Wistar rats that ingested
Pb acetate (50, 100, 150 ppb) in drinking water for 6 months; measurements were made
during the pro-estrous stage of the estrous cycle to allow for consistent timing and control
for cyclic hormonal variation. The authors reported decreases in FSH, estradiol, and
progesterone levels with increases in LH and testosterone levels. Nampoothiri and Gupta
(2008) administered Pb acetate to Charles Foster female rats 5 days before mating and
during the gestational period at a concentration that did not affect reproductive
performance, implantation or pregnancy outcome (0.05 mg/kg body weight). They
observed a decrease in steroidogenic enzymes, 3(3- hydroxysteroid dehydrogenase (HSD)
and 17(3-HSD, activity in reproductive organs, as well as a decrease in steroid hormones
(progesterone and estradiol), suggesting that chronic exposure to low levels of Pb may
affect reproductive function of mothers and their offspring. Similarly, Pillai et al. (2010)
reported impaired ovarian steroidogenesis in Charles Foster adult female rats (PND56)
from dams treated gestationally and lactationally to Pb acetate (subcutaneous daily
injections of 0.05 (ig/kg BW). Pillai observed a decrease in steroidogenic enzymes, 3(3-
HSD and 17|3-HSD, but saw no changes in ovarian steroidogenic acute regulatory protein
(StAR) or CYP11 mRNA levels indicating Pb-induced inhibition of ovarian
steroidogenesis.
Kolesarova et al. (2010) conducted an in vitro study to examine the secretory activity of
porcine ovarian granulose cells after Pb administration for 18 hours. The results of the
study showed that Pb acetate concentrations of 0.046 mg/mL and 0.063 mg/mL
statistically significantly inhibited insulin-like growth factor-1 (IGF-1) release, but
concentrations of 0.25 mg/mL and 0.5 mg/mL Pb did not influence IGF-1 release.
Progesterone release was not affected by Pb treatment; however, Pb caused a reduction in
LH and FSH binding in granulose cells and increased apoptosis as evidenced by
increased expression of caspase-3 and cyclin Bl, suggesting a Pb-induced alteration in
the pathways of proliferation and apoptosis of porcine ovarian granulose cells. Decreased
gonadotropin binding was also observed in rats after Pb subcutaneously administration
(0.05 mg/kg body weight daily) before mating and during pregnancy, resulting in a blood
Pb of 2.49 (ig/mL (Nampoothiri and Gupta. 2006).
No recent toxicological studies were found that examined Pb-induced effects on the
estrus cycle.
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Overall, toxicological studies report alterations in hormone levels related to blood Pb
concentration. Similarly, epidemiologic studies reported associations between
concurrent/closely timed blood Pb levels and hormone levels in female adults. Although
Pb-associated changes in hormone levels are observed, there are discrepancies and the
hormones examined vary by study. One explanation for the inconsistent findings is that
changes could vary based on current hormonal and reproductive status of the participants.
Also, the covariates included in statistical models as potential confounders varied among
studies, which could contribute to between study heterogeneity. Another limitation of the
epidemiologic studies is that not all of the studies investigated important confounders,
such as other metal exposures or smoking. Additionally, the cross-sectional design of
these studies leaves uncertainty regarding Pb exposure magnitude, timing, duration, and
frequency that contributed to the observed associations.
4.8.4.2 Effects on Fertility
Previous studies indicated that Pb exposure does not produce total sterility, but it can
disrupt female fertility (U.S. EPA. 2006b). Recent epidemiologic studies and studies in
experimental animals have inconsistent results. The recent epidemiologic studies are
summarized in Table 4-46. Most of these studies examined biological measures of Pb
collected at or during the period of possible fertilization or start of fertility treatment,
although Bloom et al. (2011 a) measured blood Pb at baseline and followed women for at
least 12 menstrual cycles (or until pregnancy).
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Table 4-46 Summary of recent epidemiologic studies of associations between Pb biomarker
levels and fertility
for females.
Reference3 Outcome
Bloom et al. Achieving
(2011 a) pregnancy
Chang et al. Infertility
(2006)
Study Population
Study Location
Years
Women who were
aged 18-34 yr, were
previously part of a
study about fish
consumption, and
were not currently
pregnant and were
followed for 12
menstrual cycles or
until pregnant
N=80
New York
1996-1997
Women receiving
care at a infertility
clinic in 2000-2001 or
delivering a normal
infant at a nearby
medical center in
1999
Cases: N=64
Controls: N=83
Kaohsiung City,
Taiwan
1999, 2000-2001
Methodological Pb Biomarker
Details Mean (SD) in ug/dL Adjusted Effect Estimates
Longitudinal Blood Pb at baseline Probability (95% Cl) for
cohort using achieving pregnancy per
Cox 0.6 ug/dL blood Pb:
proportional No Posltlve n 031 M 066 1 004)
hazards pregnancy test: 1.55 u.uoi^ i.uoo, i.uut;
(0.16)
Positive pregnancy
test: 1.54(0.12)
Case-control Blood Pb OR (95% Cl) for Infertility
study using < 2.5 ug/dL: 1 .00 (Ref)
unconditional 3.12(0.19) >2.5 ug/dl_: 2.94 (1.18, 7.34)
IOQ ISTIC
regression
Potential Confounders
Adjusted for in Analysis
Baseline As, baseline Cd,
baseline Mg, baseline Ni,
baseline Se, baseline Zn,
total serum lipids, age,
parity, frequency of
intercourse during fertility
window, alcohol use,
cigarette use
Age, BMI, active smoking,
use of Western medicine
Considered but did not
include: irregular
menstruation, age at first
menses, marital status,
passive smoking,
contraceptive drugs
4-690
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Table 4-46 (Continued): Summary of recent epidemiologic studies of associations between Pb biomarker levels and fertility for
females.
Reference3
Bloom et al.
(2010)
Bloom et al.
(2011 b)
Outcome
Oocyte
maturity,
oocyte
fertilization
Embryo cell
number,
embryo
fragmentation
score
Study Population
Study Location
Years
Women who were
part of the Study of
Metals and Assisted
Reproductive
Technologies
(SMART): women
referred to the Center
for Reproductive
Health of UCSF for
infertility treatment
and their first IVF
procedure
N=15
California
2007-2008
Women who were
part of the Study of
Metals and Assisted
Reproductive
Technologies
Methodological
Details
Longitudinal
cohort using
multivariable
log-binomial
regression
Longitudinal
cohort using
ordinal logistic
regression
Pb Biomarker
Mean (SD) in ug/dL
Blood Pb at the time
of oocyte retrieval
0.82 (0.32)
Blood Pb at the time
of oocyte retrieval
0.83 (0.30)
Adjusted Effect Estimates
RR per 1 ug/dL
Oocyte maturity (determined
by Metaphase II arrest):
0.54(0.31, 0.93)
0.25(0.03,2.50)*
Oocyte fertilization:
0.97(0.66, 1.43)
1.09(0.72, 1.65)*
OR per 1 ug/dL
Higher embryo cell number:
0.25(0.07, 0.86)
Potential Confounders
Adjusted for in Analysis
Age, cigarette smoking,
race/ethnicity
*Also, adjusted for Cd.
Age, race/ethnicity,
cigarette smoking, urine
creatinine, Hg, Cd
(SMART) (women
referred to the Center
for Reproductive
Health of UCSF for
infertility treatment
and their first IVF
procedure) and who
generated embryos
N=24
California
2007-2008
Higher embryo fragmentation
score:
1.71 (0.45,6.56)
Additionally included for
embryo cell number: day
of embryo transfer
4-691
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Table 4-46 (Continued): Summary of recent epidemiologic studies of associations between Pb biomarker levels and fertility for
females.
Reference3 Outcome
Al-Saleh et Achieving
al. (2008a) pregnancy
and/or
fertilization
Study Population
Study Location
Years
Women aged 19-50
yr undergoing IVF
Pregnancy: N=203
No pregnancy: N=321
Fertility: N=556
No fertility: N=63
Riyadh, Saudi Arabia
2002-2003
Methodological
Details
Longitudinal
cohort using
logistic
regression
Pb Biomarker
Mean (SD) in ug/dL
Blood Pb
Follicular fluid Pb
Blood Pb:
3.34 (2.24)
Blood Pb levels
>10ug/dL:
1.7%
Follicular fluid:
0.68(1.82)
Adjusted Effect Estimates
OR (95% Cl) (unit not given,
assume results are per
1 Mg/dL)
Pregnancy
Blood Pb: 0.55(0.23, 1.31)
Follicular fluid Pb: 1.36(0.91,
2.02)
Fertilization
Blood Pb: 0.30(0.08, 1.03)
Follicular fluid Pb: 1.45(0.69,
•a no\
Potential Confounders
Adjusted for in Analysis
Age, husband's age, BMI,
location and duration at
that location, previous
location and duration at
that location, age at first
menses, number of days
of menstrual cycle,
education, work status,
husband's education,
family income, husband's
smoking status, blood and
follicular Cd and Hg,
follicular cotinine
Note: In a reduced adjusted
model for fertilization, the OR
for blood Pb was 0.38 (0.14,
0.99)
Also included for
pregnancy as outcome:
coffee consumption, tea
consumption, caffeinated
soft drink consumption,
Silberstein et
al. (2006)
Achieving
pregnancy
Women undergoing
IVF at the study
hospital
Pregnancy: N=4
No pregnancy: N=5
Providence, Rl
NS
Longitudinal
cohort study
using Mann-
Whitney U-test
Follicular fluid Pb
Not given
quantitatively
Estimated from a
figure in the paper:
Median Pb in
follicular fluid of
pregnant women:
-1.3
Median Pb in
follicular fluid of non-
pregnant women:
-2.2
P-value for difference in
medians by Mann-Whitney U
test: 0.0059
*Note, This study only included
9 women
None
aStudies are presented in order of first appearance in the text of this section, which is based on study design and methodological strength.
4-692
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A prospective cohort study followed women who previously participated in a study of
fish consumption for a length of up to 12 menstrual cycles and investigated the
relationship between blood Pb levels at baseline and having a positive pregnancy test at
some point during the next 12 menstrual cycles (Bloom etal.. 201 la). No association was
observed between blood Pb and achieving pregnancy.
Among women aged 23-44 years, a difference in blood Pb was reported between women
who were seeking help at a fertility clinic after one year of trying to conceive and women
who had previously delivered an infant and were identified from medical records of a
postpartum care unit at a medical center (Chang et al.. 2006). Higher odds of infertility
were observed when comparing women with blood Pb levels >2.5 (ig/dL to those with
blood Pb levels < 2.5 (ig/dL although this study is limited by its case-control design.
Epidemiologic studies have also examined women having difficulty conceiving by
performing studies among patients of fertility clinics or undergoing IVF. The Study of
Metals and Assisted Reproductive Technologies (SMART) enrolled women undergoing
their first round of IVF and investigated multiple steps before pregnancy as the outcomes
(Bloom et al.. 20lib; Bloom et al.. 2010). Higher blood Pb levels were associated with
lower oocyte maturity although the wide 95% CIs made interpretation of models
controlling for Cd difficult. No association was observed between blood Pb and oocyte
fertilization (Bloom et al.. 2010). In the examination of markers of IVF success,
inconsistent results were observed. Embryo cell number was lower in association with
higher blood Pb levels but no association was observed for embryo fragmentation score
(Bloom etal., 20lib). Another study examining fertility reported on women in Saudi
Arabia aged 19-50 years who were undergoing IVF treatment (Al-Saleh et al.. 2008a).
Women were categorized as having achieved a pregnancy versus not having achieved a
pregnancy and achieved fertilization versus not achieving fertilization. The majority of
women had follicular Pb levels that were below the limit of detection; whereas less than
2% of women had blood Pb levels below the limit of detection. In addition, less than 2%
of women had blood Pb levels that were above 10 (ig/dL. Follicular Pb levels were not
correlated with the blood Pb. No association was observed between blood or follicular Pb
and pregnancy outcomes in either crude or adjusted models. An association was not
detected between follicular Pb and fertilization, but higher blood Pb was associated with
lower rates of fertilization. Finally, a study that included nine women undergoing IVF
treatment in Rhode Island (Silberstein et al.. 2006) found that median follicular Pb levels
in women who achieved pregnancy were lower than the follicular Pb levels among
nonpregnant women.
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Overall, these epidemiologic studies examine a variety of fertility-related endpoints and
although some studies demonstrate an association between higher Pb biomarker levels
and fertility/pregnancy, as a whole the results are inconsistent across studies. One
limitation present in most of these studies is that the participants are women who are
seeking help for fertility problems. The participants are not samples of the general
population and therefore cannot be generalized to all women of childbearing age. This
may also have introduced substantial selection bias into the study.
Animal toxicology studies following female fertility looked at various outcomes. Several
studies observed a decrease in litter size when females were exposed to Pb before mating
or during pregnancy (Dumitrescu et al.. 2008b: lavicoli et al.. 2006a: Teijon et al.. 2006).
Pups in a study by Teijon et al. (2006) receiving 400 ppm Pb acetate in drinking water
had blood Pb of 97 (ig Pb/dL blood at 1 week post-weaning and 18.2 (ig Pb/dL blood at
2 week post-weaning. Dumitrescu et al. (2008b) observed a modification in sex ratio of
pups born to dams exposed to Pb before mating and during the entirety of pregnancy. As
the dose of Pb increased, the number of females per litter also increased (i.e., 1 male to
0.8 female in non-Pb-exposed group; 1 male to 0.66 female in 50 ppb Pb acetate group; 1
male to 2.25 females in 100 ppb group; and 1 male to 2.5 females in 150 ppb group).
These results are not consistent with earlier results of Ronis et al. (1998b). who did not
observe differences in sex ratio dams and offspring were exposed only during pregnancy.
Thus, Pb exposure in animal studies during or before pregnancy have shown effects on
litter size and mixed effects on sex ratio.
Nandi et al. (2010) demonstrated a concentration-dependent decline in viability rate,
maturation, fertilization, and cleavage rates of buffalo oocytes cultured in medium
containing 1-10 (ig/mL Pb acetate (24 hour culture). Karaca and §im§ek (2007) observed
an increase in the number of mast cells in ovary tissue after Pb exposure (2,000 (ig/mL in
drinking water for 6 weeks prior to estrous monitoring then for 1 additional month during
which estrous cyclicity was monitored) suggesting that Pb may stimulate an
inflammatory response in the ovaries which may contribute to Pb-induced female
infertility.
In contrast, Nampoothiri and Gupta (2008) did not observe any statistically significant
change in fertility rate or litter size in female rats subcutaneously administered Pb
(0.05 mg/kg body weight daily before mating and during pregnancy) with a resulting
blood Pb of 2.49 (ig/mL. Although reproductive performance was not affected in this
study, the authors did report an alteration in implantation enzymes. Cathepsin-D activity
decreased and alkaline phosphatase activity increased after Pb exposure.
4-694
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In summary, recent epidemiologic and toxicological studies on the effect of Pb on
fertility outcomes have generated inconsistent results. Most of the epidemiologic studies
are limited by their small sample sizes, potential selection bias, and lack of
generalizability due to a focus on women seeking help for infertility. Most of the studies
control for multiple potential confounders, such as smoking status and age. The studies
from the toxicological literature show that Pb exposure to female experimental animals
affects litter size (decreased litter size), sex ratio (ratio of male to female offspring in a
litter) and ovarian viability, albeit often at higher doses of Pb than relevant to this ISA.
However, the bulk of the toxicological evidence including the current and historical Pb
literature (U.S. EPA. 2006b) indicates that increased Pb exposure may decrease female
fertility.
4.8.4.3 Ovaries, Embryo Development, Placental function, and
Spontaneous Abortions
The 2006 Pb AQCD included studies of Pb exposure among men and women and their
associations with spontaneous abortions. The 2006 Pb AQCD concluded that overall
there was little evidence to support an association between Pb exposure among women
and spontaneous abortion (U.S. EPA. 2006b). Most of the studies examined in the
2006 Pb AQCD assigned exposure based on living near a smelter or working in
occupations that often result in Pb exposure and the results of these studies were
inconsistent. Little evidence was available in the 2006 Pb AQCD to suggest an
association with paternal Pb biomarker levels, and no recent studies have been performed
to examine paternal Pb biomarker levels and spontaneous abortion. Since the
2006 Pb AQCD, multiple epidemiologic studies have been published that examine Pb
biomarker levels in women and their possible association with spontaneous abortion.
Table 4-47 provides information on these longitudinal and cross-sectional studies.
Additionally, toxicological studies have studied the effects of Pb on fetal loss and the
contribution of the ovaries and placenta to fetal loss.
4-695
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Table 4-47 Summary of recent epidemiologic studies of associations between Pb biomarker levels and
spontaneous abortions.
Reference3
Outcome
Study Population
Study Location
Years
Methodological
Details
Pb Biomarker
Mean (SD) in ug/dL
Adjusted Effect Estimates
Potential
Confounders
Adjusted for
in Analysis
Vigeh et al.
(2010)
Pregnancy
ended before
20 weeks of
gestation
Women who were non-
smokers, non-obese,
had no chronic health
conditions, had their last
menstrual period less
than 12 weeks prior,
and were pregnant with
a singleton infant
> 1 spontaneous
abortions: N=15
No spontaneous
abortions: N=336
Longitudinal
cohort study
using t-test and
logistic
regression
Maternal blood Pb
during weeks 8-12 of
pregnancy
3.8(2.0)
Spontaneous abortion:
3.51 (1.42)
Non-spontaneous
abortion: 3.83 (1.99)
T-test for difference in mean values:
p-value 0.41
OR:
0.331 (95% Cl: 0.011, 10.096) for an
increase in log-transformed blood Pb i
assume per 1 ug/dL)
Age, parity,
hematocrit,
passive
cigarette
smoking
exposure
Tehran, Iran
2006-2008
4-696
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Table 4-47 (Continued): Summary of recent epidemiologic studies of associations between Pb biomarker levels and
spontaneous abortions.
Reference3
Outcome
Study Population
Study Location
Years
Methodological
Details
Pb Biomarker
Mean (SD) in ug/dL
Adjusted Effect Estimates
Potential
Confounders
Adjusted for
in Analysis
Yin et al.
(2008)
Anembryonic
pregnancy
Women age 25-35 yr
and at 8-12 weeks of
gestation at study entry;
cases were
anembryonic
pregnancies and
controls were normal
pregnancies that ended
in a live birth between
37-42 weeks
Case-control
study using t-
test
Maternal blood Pb
after miscarriage for
cases and at study
enrollment for controls
Cases:
5.3(95%CI: 5.2, 5.9)
Controls:
4.5(95%CI: 3.7, 5.0)
Comparisons between log-
transformed blood Pb levels of cases
and controls performed via Student's
t-test had a p-value of 0.03
None
Cases: N=40
Controls: N=40
Shanxi Province, China
2004-2006
Lamadrid-
Figueroa et al.
(2007)
Previous
miscarriage
Women who had a
previous pregnancy and
were currently pregnant
with gestational age of
< 14 weeks
Cross-sectional
study using
Poisson
regression
Maternal and umbilical
cord blood Pb,
maternal bone Pb
Overall:
IRR (95%CI) Categorized Plasma
Blood Pb ratio:
1st fertile: 1.00(Ref)
2nd fertile: 1.16 (p-value 0.61)
3rd fertile: 1.90 (p-value 0.015)
Age,
education
> 1 previous
miscarriages: N=71
No previous
miscarriages: N=136
Mexico City, Mexico
1997-1999,2001-2004
Blood Pb: 6.2 (4.5)
Plasma Pb: 0.014
(0.013)
Cases:
Blood Pb: 5.8 (3.4)
Plasma Pb: 0.014
(0.013)
Controls:
Blood Pb: 6.5 (4.9)
Plasma Pb: 0.013
(0.013)
IRR (95%CI) Per 1 SD increase:
Plasma Pb: 1.12 (p-value 0.22)
Blood Pb: 0.93 (p-value 0.56)
Plasma/Blood Pb ratio: 1.18 (p-value
0.02)
Patella Pb: 1.15 (p-value 0.39)
Tibia Pb: 1.07 (p-value 0.56)
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Table 4-47 (Continued): Summary of recent epidemiologic studies of associations between Pb biomarker levels and
spontaneous abortions.
Reference3
Gundacker et
al. (2010)
Study Population
Study Location
Outcome Years
Previous Women recruited during
miscarriage the second trimester of
pregnancy
> 1 previous
miscarriages: N=8
No previous
miscarriages: N=22
Methodological
Details
Cross-sectional
study using non-
parametric tests
Pb Biomarker
Mean (SD) in ug/dL
Whole placentas
shortly after birth
Median (IQR):
25.8(21.0, 36.8)
Adjusted Effect Estimates
Median Placenta Pb:
Women who had not previously
miscarried:
27 ug/kg
Women who had previously
miscarried:
39 ug/kg (p-value for difference:
0.039)
Potential
Confounders
Adjusted for
in Analysis
None
Vienna, Austria
2005
"Studies are presented in order of first appearance in the text of this section, which is based on study design and methodological strength.
4-698
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A longitudinal study examining spontaneous abortions occurring early in the pregnancy
was conducted in Iran (Vigeh et al.. 2010). Mean blood Pb concentrations, measured at
8-12 weeks of pregnancy, were similar in women who did and did not have spontaneous
abortions. Higher blood Pb levels were not associated with greater odds of spontaneous
abortions before 20 weeks of pregnancy. Yin et al. (2008) performed a study in the
Shanxi Province of China to examine if plasma Pb levels were associated with
anembryonic pregnancies (spontaneous abortions during the first trimester, which
account for 15% of all spontaneous abortions). Women were enrolled at 8-12 weeks of
gestation. Women who delivered a term pregnancy had mean plasma Pb levels that were
lower than those of women who had an anembryonic pregnancy (plasma Pb measured at
the time of miscarriage for cases and at 8-12 weeks for controls). Of note, among cases
plasma Pb level was inversely correlated with folate and vitamin B12, but this correlation
was not observed among those who delivered at term; no models examining plasma Pb
levels were adjusted for nutrient status. A study in Mexico City examined a group of
pregnant women (maximum gestational period at enrollment was 14 weeks) who had
previously been pregnant and either given birth or had a spontaneous abortion (Lamadrid-
Figueroa et al.. 2007). Women in the highest tertile of plasma/blood Pb ratio had a higher
rate of previous spontaneous abortions than did women in the lowest tertile. The authors
state that the plasma/whole blood ratio represents bioavailable Pb, which is capable of
crossing the placental barrier for a given blood concentration. No association was
observed for the relationship between Pb and spontaneous abortions using whole blood,
plasma, or bone Pb alone. Similarly, a study of placental Pb levels among pregnant
women in Austria observed higher placenta Pb levels among women who had miscarried
a previous pregnancy compared to women who had not miscarried a previous pregnancy
(Gundacker et al.. 2010). It is important to note that the number of women included in the
study was small (only 8 women reported previously having a miscarriage)
In toxicological studies, isolated bovine embryo cultures are often used to understand the
mechanisms responsible for aberrant embryo development as it may contribute to
teratogenesis, fetal loss or negative postnatal pup outcomes. Nandi et al. (2010)
demonstrated an exposure concentration-dependent decline in embryo development of
fertilized buffalo oocytes cultured for 24 hours in medium containing 0.05-10 (ig/mL
Pb acetate as evidenced by reduced morula/blastocyst yield and increased four-to eight-
cell arrest, embryo degeneration, and asynchronous division. This study provides
evidence of the negative effect of Pb on embryo development and contributes
mechanistic understanding to Pb-induced pregnancy loss.
A possible explanation for reduced fertility and impaired female reproductive success as
a result of Pb exposure is changes in morphology or histology in female sex organs and
the placenta (Dumitrescu et al.. 2007; U.S. EPA. 2006b). Wang et al. (2009e) observed
4-699
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that elevated maternal blood Pb (0.6-1.74 (iM, -12.4-36.0 (ig/dL) compared to control
(0.04 (iM, -0.83 (ig/dL) was associated with decreased fetal body weight, pup body
length, and placental weight in Wistar rats. The authors reported that placentae from
Pb-exposed groups showed concentration-dependent increasing pathology of
cytoarchitecture and cytoplasmic organelles. The authors also reported positive
expression of NF-KB, a transcription factor that controls the expression of genes involved
in immune responses, apoptosis, and cell cycle, in the cytotrophoblasts, decidual cells,
and small vascular endothelial cells in rat placenta under a low-level Pb exposure
condition which correlated with low blood Pb levels.
Pb-exposed (273 mg/L or 819 mg/L in drinking water, 500 or 1,500 ppm Pb acetate,
respectively) male rats from Anjum et al. (2010) that had an exposure concentration-
dependent decrease in serum testosterone, decreased male reproductive organ weight and
decreased sperm were mated to untreated females. These untreated dams bred to the Pb
exposed males concentration-dependent decreased implantation rate and higher pre- and
post-implantation loss, indicating paternally mediated fetal loss. The magnitude of these
effects in dams was dependent on the concentration of Pb exposure in their male mating
partners.
As observed in sperm cells, Pb stimulates changes in antioxidant enzyme activity in rat
ovaries indicating that oxidative stress may be a contributing factor in Pb-induced ovarian
dysfunction. Nampoothiri et al. (2007) observed a reduction in SOD activity and an
increase in CAT activity along with a decrease in glutathione content and an increase in
lipid peroxidation in rat granulosa cells after 15 days of Pb treatment (subcutaneously
administered Pb, 0.05 mg/kg body weight daily before mating and during pregnancy with
a resulting blood Pb of 2.49 (ig/mL).
Previous studies demonstrated that Pb can be found in the ovaries and causes histological
changes, thus potentially contributing to Pb-induced effects on female fertility (U.S.
EPA. 2006b). In support of historical studies, recent studies demonstrate Pb-induced
histological changes in ovarian cells of pigs (Kolesarova et al.. 2010) and rats
(Nampoothiri et al.. 2007; Nampoothiri and Gupta, 2006). Kolesarova et al. (2010)
observed a reduction of the monolayer of granulosa cells after Pb addition (0.5 mg/mL,
18 hours culture). Nampoothiri and Gupta (2006) reported that Pb treatment caused a
decrease in cholesterol and total phospholipid content in the membranes of granulosa
cells which resulted in increased membrane fluidity (subcutaneously administered Pb,
0.05 mg/kg body weight daily before mating and during pregnancy with a resulting blood
Pbof2.49(ig/mL).
Overall, the recent studies support the conclusions of the 2006 Pb AQCD (U.S. EPA.
2006b) that there is mixed evidence among epidemiologic studies to suggest an
4-700
-------
association between Pb and spontaneous abortions. It is important to note that studies of
spontaneous abortions are difficult to conduct. The majority of spontaneous abortions
occur during the first trimester, which makes them difficult to capture. Women may
miscarry before being enrolled in a study and many women may not have known they
were pregnant when they miscarried. This limits the ability to detect subtle effects,
especially if higher Pb exposures do lead to increased risk of early spontaneous abortions.
In addition, some studies are limited by their retrospective examination of current Pb
biomarker levels in relation to previous miscarriages. Sample size is another limitation of
the available epidemiologic studies. The epidemiologic studies also had little control for
potential confounding factors, with some studies including no potential confounders in
their analyses. Toxicological data provide a potential mechanistic explanation of the
contribution of Pb exposure to spontaneous abortions. These laboratory data show that Pb
exposure impaired placental function, induced oxidative stress and histological changes
in the ovaries, and affected embryo development. The toxicological and epidemiologic
data provide inconsistent findings for the role of Pb in spontaneous abortions.
4.8.4.4 Effects on Breast Milk
Experiments in laboratory animals have shown that dietary manipulation of maternal
fatty acid (FA) levels in diet can worsen Pb-related behavioral effects of offspring after
lactational Pb exposure (Lim et al.. 2005). To determine if components of dam milk
contributed to this change, dam milk fatty acids were altered via diet. Diets deficient in
(n-3) fatty acids (FA[s]) can lead to a deficiency of DHA, which is essential for proper
nervous system development. Lim et al. (2005) found that dam Pb exposure (Long-Evans
rats, 2,000 ppm Pb acetate trihydrate/BW) during lactation (PNDO-PND21) led to a
decrement in non-essential fatty acids in the maternal organs at PND25 but with blood Pb
levels much higher than those relevant to this ISA (mean [SD] blood Pb levels in dams:
308 [56] (ig/dL). In animals with a diet deficient in (n-3) FAs, there was a Pb-diet
interaction with a specific size PUFA (i.e., a 20-carbon [n-6] PUFA). In general, Pb
exposure caused a decrement in shorter chain monounsaturated and saturated FAs in
maternal organs.
Supplementation of the maternal diet with calcium can be an especially important
modulator of Pb mobilization into maternal blood during periods of high calcium demand
including pregnancy/lactation. For example, mothers with elevated blood Pb levels given
calcium phosphate and ascorbic acid supplementation during pregnancy and lactation had
a 90% decrease in placental Pb content and a 15% decrease in the concentration of Pb in
breast milk (Altmann et al.. 1981) versus the control group that did not receive dietary
treatment. Another study (Gulson et al.. 2004a) has shown that calcium supplementation
4-701
-------
during lactation is less beneficial in modulating maternal blood Pb levels (mean blood Pb
at first sampling was 2.4 (ig/dL); the Gulson cohort (Gulson et al.. 2004a) was limited by
power (n=10 women). In a cohort of women from Mexico City, daily calcium
supplementation during lactation reduced maternal blood Pb by 15-20% and Pb in breast
milk by 5-10% (Ettinger et al.. 2004a). Another study by the same investigators showed
that using calcium supplements daily during pregnancy also reduced blood Pb levels
during pregnancy (Ettinger et al., 2009) with the effect strongest in women with higher
biomarkers of Pb exposure (elevated baseline bone Pb or >5 (ig/dL blood Pb) or in
women with higher Pb exposure (self-reported use of Pb-glazed ceramics). Thus, dietary
modulation with calcium supplementation during pregnancy and lactation may decrease
the amount of Pb to which the developing fetus of infant is exposed. The evidence for
this seems especially strong for protection during pregnancy and more mixed for
protective effects of calcium during lactation.
4.8.4.5 Summary of Effects on Female Reproductive Function
In summary, Pb exposure was found to affect female reproductive function as
demonstrated by both epidemiologic and toxicological studies. Some evidence is also
available regarding associations between blood Pb levels and altered hormone levels in
adults, but varied among studies. The differences may have been due to the different
hormones examined and the different timing in the menstrual and life cycles of the
women. Studies reported inconsistent findings for the association between Pb and
fertility. Adjustment for potential confounders varies from study to study, with some
potentially important confounders, such as BMI, not included in all studies. Also, many
epidemiologic studies are limited by small samples sizes, and are generally comprised of
women attending infertility clinics; which presents the possibility of selection bias and
lack of generalizability. Toxicological studies found effects on female reproductive
function after prenatal or early postnatal exposures. Further coherence for these findings
in laboratory animal models is found in with findings in the ecological literature for the
effects of Pb exposure on reduced fecundity in terrestrial and aquatic animal species
(Sections 6.4.5.2. 6.3.4.2. 6.3.4.3. and 6.4.5.3). Although epidemiologic and toxicological
studies provide information on different exposure periods, both types of studies support
the conclusion that Pb affects at least some aspects of female reproductive function.
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4.8.5 Summary and Causal Determination
Many epidemiologic and toxicological studies of the effects of Pb on reproductive and
developmental outcomes have been conducted since the 2006 Pb AQCD. The evaluation
of causal relationships with Pb exposure focuses on four areas: developmental effects,
birth outcomes, reproductive function among males, and reproductive function among
females. The sections that follow describe the evaluation of evidence for these outcomes
with respect to causal relationships with Pb exposure using the framework described in
Table II of the Preamble. The application of the key evidence to the causal framework is
summarized in Table 4-48.
4.8.5.1 Effects on Development
The 2006 Pb AQCD (U.S. EPA. 2006b) reported Pb-associated effects on development in
toxicological studies. Multiple epidemiologic studies of Pb and puberty have shown
associations between concurrent blood Pb levels and delayed pubertal onset for girls and
boys. Delayed puberty has been linked to decreased peak bone mass and increased risk of
osteoporotic fractures (Gilsanz etal., 2011; Naves et al., 2005). In cross-sectional
epidemiologic studies of girls (ages 6-18 years) with mean and/or median concurrent
blood Pb levels from 1.2 to 9.5 (ig/dL, consistent associations with delayed pubertal onset
(measured by age at menarche, pubic hair development, and breast development) were
observed. In boys (ages 8-15 years), fewer epidemiologic studies were conducted but
associations between blood Pb levels and delayed puberty were observed, including
associations among boys in a longitudinal study. These associations are consistently
observed in populations with mean or median blood Pb levels of 3.0 to 9.5 (ig/dL.
Potential confounders considered in the epidemiologic studies of both boys and girls that
performed regression analyses varied. Most studies controlled for age and BMI. Other
variables, such as measures of diet, SES, and race/ethnicity, were included in some of the
studies. Adjustment for nutritional factors was done less often and this could be an
important confounder. A study using a cohort of girls from NHANES controlled for
various dietary factors as well as other potential confounders and reported an association
between increased concurrent blood Pb levels and delayed pubertal onset (Selevan et al..
2003). A limitation across most of the epidemiologic studies of blood Pb levels and
delayed puberty is their cross-sectional design, which does not allow for an
understanding of temporality. There is uncertainty with regard to the exposure frequency,
timing, duration, and level that contributed to the associations observed in these studies.
Recent toxicological studies indicate that delayed pubertal onset may be one of the more
sensitive developmental effects of Pb exposure with effects observed after gestational
4-703
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exposures leading to blood Pb levels in the female pup of 1.3-13 (ig/dL (lavicoli et al..
2006a: lavicoli et al.. 2004). Toxicological studies have reported delayed male sexual
maturity as measured with sex organ weight, among other outcomes, seeing significant
decrements at blood Pb levels of 34 (ig/dL (Sokol etal.. 1985). Thus, data from the
toxicological literature and from epidemiologic studies demonstrate that puberty onset in
both males and females is delayed with Pb exposure.
Findings from epidemiologic studies of the effect of Pb on postnatal growth are
inconsistent and findings from the toxicological literature of the effect of Pb exposure are
mixed with recent growth findings showing adult onset male obesity after gestational and
lactational Pb exposure (see Section 4.8.1.3). Toxicological studies demonstrated that
other effects of Pb exposure during development include impairment of retinal
development, effects on the lens of the eye, and alterations in the developing
hematopoietic, hepatic systems and teeth.
The collective body of evidence integrated across epidemiologic and toxicological
studies, based on the findings of delayed pubertal onset among males and females, is
sufficient to conclude that there is a causal relationship between Pb exposure and
developmental effects.
4.8.5.2 Effects on Birth Outcomes
Overall, results of pregnancy outcomes were similar to those of the 2006 Pb AQCD;
(U.S. EPA. 2006b) inconsistent epidemiologic evidence of a relationship with Pb was
available for preterm birth. The 2006 Pb AQCD reported potential associations between
Pb and neural tube defects, but a recent epidemiologic study found no association. Some
associations were observed between Pb and low birth weight when epidemiologic studies
used measures of postpartum maternal bone Pb or air exposures. The associations were
less consistent for maternal blood Pb measured during pregnancy or at delivery or
umbilical cord and placenta Pb (maternal blood Pb or umbilical cord and placenta Pb
were the biomarkers most commonly used in studies of low birth weight) but some
associations between increased Pb biomarker levels and decreased low birth weight/fetal
growth were observed. The effects of Pb exposure during gestation in animal
toxicological studies included mixed findings with some studies showing reduction in
litter size, implantation, and birth weight, and some showing no effect. Based on the mix
of inconsistent results of studies on various birth outcomes but some associations
observed in well-conducted epidemiologic studies of preterm birth and low birth
weight/fetal growth, the evidence is suggestive of a causal relationship between Pb
exposure and birth outcomes.
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4.8.5.3 Effects on Male Reproductive Function
Toxicological evidence and supporting epidemiologic evidence indicate that a causal
relationship exists between Pb exposure and effects on male reproductive function. Key
evidence is provided by toxicological studies in rodents, non-human primates, and rabbits
showing detrimental effects on semen quality, sperm and fecundity/fertility with
supporting evidence in epidemiologic studies of associations between Pb exposure and
detrimental effects on sperm. This is consistent with evidence reported in the
2006 Pb AQCD (U.S. EPA. 2006b).
Toxicological studies with relevant Pb exposure routes reported effects on rodent sperm
quality and sperm production rate [(Sokol and Berman. 1991; Sokol etal.. 1985). blood
Pb level 34-37 (ig/dL], sperm DNA damage lYNava-Hernandez et al.. 2009). blood Pb
levels 19 and 22 jig/dL], and histological or ultrastructural damage to the male
reproductive organs in studies from rodents [(El Shafai et al.. 2011). blood Pb level
5.1 (ig/dL] and non-human primates [(Cullen etal. 1993). blood Pb level 43 (ig/dL].
These effects were found in animals exposed to Pb during peripuberty or adults for 1
week to 3 months. The toxicological studies reported an association between Pb exposure
and decreases in reproductive organ weight and organ histological changes in the testes
and germ cells. Subfecundity (decreased number of pups born/litter) was reported in
unexposed females mated to Pb exposed males. Also, sperm from Pb-exposed rats (blood
Pb level: 33 to 46 ug/dL) used for in vitro fertilization of eggs harvested from unexposed
females yielded lower rates of fertilization. (Sokol etal.. 1994). Supporting evidence was
provided by decrements in sperm quality from rabbits administered Pb subcutaneously
(blood Pb levels of 25 (ig/dL) (Moorman et al.. 1998).
Detrimental effects of Pb on sperm were observed in epidemiologic studies with
concurrent blood Pb levels of 25 (ig/dL and greater among men occupationally exposed
(Hsu et al.. 2009b: Kasperczyk et al.. 2008; Naha and Manna. 2007; Nahaand
Chowdhury. 2006). The epidemiologic studies were limited due to these high exposure
levels among the occupational cohorts and the lack of consideration for potential
confounding factors, including occupational exposures other than Pb. Studies among men
with lower Pb levels were limited to infertility clinic studies, which may produce a biased
sample and findings that lack generalizability. However, a well-conducted epidemiologic
study that enrolled men going to a clinic for either infertility issues or to make a semen
donation and controlled for other metals as well as smoking reported a positive
association with various detrimental effects in sperm (Telisman et al.. 2007). The median
concurrent blood Pb level in this study was 4.92 (ig/dL (range: 1.13-14.91). A similar
study (Meeker et al.. 2008) also reported possible associations between concurrent blood
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Pb and various semen parameters, but the results were extremely imprecise, making it
difficult to draw conclusions.
Similar to the 2006 Pb AQCD (U.S. EPA. 2006b). recent epidemiologic and toxicological
studies reported inconsistent results regarding hormone aberrations associated with Pb
exposure. Due to the complexity of the reproductive system, uncertainty exists as to
whether Pb exerts its toxic effects on the reproductive system by affecting the
responsiveness of the hypothalamic-pituitary-gonad axis, by suppressing circulating
hormone levels or by some other pathway. Mixed findings were also apparent among
epidemiologic studies of fertility among men.
More recent toxicological studies suggest that oxidative stress is a major contributor to
the effects of Pb exposure on the male reproductive system, providing mode of action
support. The effects of ROS may involve interference with cellular defense systems
leading to increased lipid peroxidation and free radical attack on lipids, proteins, and
DNA. Several recent studies showed that Pb induced an increased generation of ROS
within the male sex organs, and germ cell injury, as evidenced by aberrant germ cell
structure and function. Co-administration of Pb with various antioxidant compounds
either eliminated Pb-induced injury or greatly attenuated its effects. In addition, many
studies that observed increased oxidative stress also observed increased apoptosis which
is likely a critical underlying mechanism in Pb-induced germ cell DNA damage and
dysfunction.
Based on the consistency and coherence of findings for the detrimental effects of Pb
exposure on sperm and semen in the toxicological literature, the support from
epidemiologic studies, and biological plausibility provided by mode of action evidence,
the evidence is sufficient to conclude that there is a causal relationship between Pb
exposures and male reproductive function.
4.8.5.4 Effects on Female Reproductive Function
Epidemiologic and toxicological studies of reproductive function among females
investigated whether Pb biomarker levels were associated with hormone levels, fertility,
estrus cycle changes, and morphology or histology of female reproductive organs
including the placenta. Toxicological studies reported in the 2006 Pb AQCD reported
associations between Pb exposure and female reproductive function, although little
evidence was provided by epidemiologic studies (U.S. EPA. 2006b). Some epidemiologic
studies have shown associations with concurrent blood Pb levels and altered hormone
levels in adults, but results varied among studies, possibly due to the different hormones
examined and the different timing in the menstrual and life cycles. There is some
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evidence of a potential relationship between Pb exposure and female fertility, but
findings are mixed. The majority of the epidemiologic studies are cross-sectional, and
adjustment for potential confounders varies from study to study, with some potentially
important confounders, such as BMI, not included in all studies. Also, most of the studies
have small samples sizes and are generally of women attending infertility clinics.
Toxicological study design often employs prenatal or early postnatal Pb exposures at
relevant Pb levels, with Pb contributing to placental pathology and inflammation,
decreased ovarian antioxidant capacity, altered ovarian steriodogenesis and aberrant
gestational hormone levels. Although epidemiologic and toxicological studies provide
information on different exposure periods, both types of studies, including some high-
quality epidemiologic and toxicological studies, support the conclusion that Pb possibly
affects at least some aspects of female reproductive function. Overall, the relationship
observed with female reproductive outcomes, such as fertility, placental pathology, and
hormone levels in some epidemiologic and toxicological studies, is sufficient to conclude
that evidence is suggestive of a causal relationship between Pb exposure and female
reproductive function.
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Table 4-48 Summary of evidence supporting reproductive and developmental
causal determinations.
Attribute in Causal
Framework3
Key Evidence
References
Pb Exposure or
Blood Pb Levels
Associated
with Effects0
Effects on Development - Causal
Delayed Puberty Onset
Consistent
associations with
relevant blood Pb
levels in high-
quality
epidemiologic
studies that help
rule out chance,
bias, and
confounding with
reasonable
confidence
Consistent evidence in multiple large
cross-sectional epidemiologic studies
for females and males plus a
longitudinal study for males. Most of
these studies have large sample
sizes and controlled for potential
confounding by covariates such as
age, BMI, and/or education/SES.
There is uncertainty related to
exposure patterns resulting in likely
higher past Pb exposure
A large study using females aged
8-18 years from the NHANES III
study also controlled for various
dietary factors and reported
associations between blood Pb levels
and delayed puberty onset
Tomoum et al. (2010),
Mauser et al. (2008),
Williams et al. (2010),
Denham et al .(2005).
Naicker et al. (2010).
Wu et al. (2003b).
Gollenberg et al.
(2010)
Sections
4.8.1.1 and 4.8.1.2
Selevan et al. (2003)
Concurrent blood
Pb levels:
1.2-9.5 ug/dl_
Consistent
toxicological
evidence with
relevant Pb
exposures to rule
out change, bias,
and confounding
with reasonable
confidence.
Consistent toxicological evidence
from multiple laboratories of delayed
male and female puberty onset with
Pb exposure via diet or oral gavage in
rodents
Dumitrescu et al.
(2008b).
lavicoli et al. (2006a)
Pine et al. (2006)
Sections
4.8.1.1 and 4.8.1.2
Blood Pb level after
dietary exposure, or
gavage, from gestation to
estrus:
1.3-13ug/dL
Evidence clearly
describes mode of
action.
Toxicological evidence describes
HPG axis dysfunction and. Mode of
action changes in IGF-1 contributing
to Pb-induced delay in puberty onset.
Section 4.8.1.1
Postnatal Growth
Available
toxicological
evidence is mixed
There are mixed findings in the
toxicological literature on Pb
exposure and postnatal growth with
some studies showing stunted growth
in animals exposed to Pb and some
showing no effects.
Recent toxicological evidence of
effect of Pb on postnatal growth:
obesity in adult male offspring after
gestational + lactational Pb exposure
22.8 ug/dl_
Section 4.8.1.3
10 ug/dL
42 ug/dl_
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Table 4-48 (Continued): Summary of evidence supporting reproductive and developmental
causal determinations.
Attribute in Causal
Framework3
Key Evidence
References
Pb Exposure or
Blood Pb Levels
Associated
with Effects0
Available Multiple studies, mostly cross-
epidemiologic sectional, for children of varying ages
evidence is have reported inconsistent results for
inconsistent the association between blood Pb
levels and various measures of
growth. There is uncertainty related to
exposure patterns resulting in likely
higher past Pb exposure.
Section 4.8.1.3
Impaired Organ Systems
Consistent
toxicological
evidence for effects
on various organ
systems but not
always with
relevant Pb
exposures.
Relevant gestational and lactational
Pb exposure of rats resulted in retinal
ERG aberrations and increased
retinal cell layer thickness.
Effects also found on bone, teeth,
and Gl system.
Section 4.8.1.4
Blood Pb level after
gestational-lactational
exposure:
10-12ug/dL
Effects on Birth Outcomes - Suggestive
A few high-quality
epidemiologic
studies show
associations with
relevant blood Pb
levels but findings
are overall
inconsistent.
Inconsistent findings for studies for
birth defects, preterm birth, and low
birth weight/fetal growth.
A few well-conducted epidemiologic
studies of preterm birth and low birth
weight/fetal growth using measures
of maternal blood Pb at the time of
pregnancy reported associations.
See Section 4.8.3 (and
all subsections):
4.8.3.1,4.8.3.2,4.8.3.3
and 4.8.3.4
Jelliffe-Pawlowski et al.
(2006),
Viqehetal.(2011),
Zhu et al. (2010),
Chen et al. (2006a),
Gundacker et al.
(2010)
Maternal pregnancy blood
Pb levels:
>10ug/dL
Inconsistent The toxicological literature reported
toxicological mixed findings with some studies
evidence showing smaller litter size (fewer
pups born) or decreased birth weight
with Pb exposure and some studies
showing no effect.
See Sections 4.8.2.4
and 4.8.2.1
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Table 4-48 (Continued): Summary of evidence supporting reproductive and developmental
causal determinations.
Attribute in Causal
Framework3
Key Evidence
References
Pb Exposure or
Blood Pb Levels
Associated
with Effects0
Effects on Male Reproductive Function - Causal
Sperm/semen production, quality, and function
High-quality and
consistent
toxicological
evidence with
relevant Pb
exposures to rule
out chance, bias,
and confounding
with reasonable
confidence.
Decreased sperm counts, decreased
sperm production rate, dose-
dependent suppression of
spermatogenesis in rodents with
drinking water Pb exposure.
Ultrastructural and histological
All results:
Section 4.i
Sokol et al
Sokol and
(1991)
3.3.1
.(1985)
Berman
Blood Pb level after adult
drinking water exposure
for 30 days: 34 ug/dL
Blood Pb level after
peripubertal or adult
drinking water exposure
for 30 days:
35 and 37 ug/dL
Maximum blood Pb levels
damage to non-human primates
testis and seminiferous tubules.
Histologic damage to rodent
seminiferous tubules including
spermatids and developing sperm.
Ultrastructural abnormalities to rat
spermatogenesis.
Direct effects on rodent sperm DMA
after drinking water Pb exposure.
Sperm from Pb exposed rats used
for in vitro fertilization of eggs
harvested from unexposed females
yielded lower rates of fertilization.
Semen and sperm quality in rabbits
with subcutaneous Pb treatment;
Ultrastructural damage to spermatids
with i.p. injection of Pb.
Findings of detrimental effects of Pb
exposure on sperm from multiple
species (Deer, Asian earthworm,
rainbow trout, marine worm,
H. elegans, Fathead minnow)
Singh et al. (1993a)
Foster et al. (1998)
El Shafai et al. (2011)
Murthy et al.(1995)
Nava-Hernandez et al.
(2009)
Sokol etal.CI994)
Moorman et al.(1998),
See Ecological Effects;
Sections 6.4.12.1 and
6.4.21.1)
after daily oral Pb
exposure (gelatin capsule)
during infancy, post
infancy, or over a lifetime
(up to 10 yr):
32 to 36 ug/dL
Blood Pb level after adult
exposure (oral gavage) for
3 months: 5.31 ug/dL
Blood Pb level after i.p.
injection for 16 days:
7.4 ug/dl_
Blood Pb level after adult
exposure for 13 weeks:
19and22ug/dL
Blood Pb level after adult
exposure for 14-60 days:
33-46 ug/dl_
Blood Pb level after adult
exposure for 15 weeks:
16-24 ug/dl_
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Table 4-48 (Continued): Summary of evidence supporting reproductive and developmental
causal determinations.
Attribute in Causal
Framework3
Key Evidence
References
Pb Exposure or
Blood Pb Levels
Associated
with Effects0
lexicological
evidence is
supported by
consistent findings
in epidemiologic
studies that cannot
rule out uncertainty
Consistent evidence from studies of
occupational cohorts with high blood
Pb levels. Results from occupational
cohorts may have been confounded
by other workplace exposures, which
were not adjusted for in the
epidemiologic studies. Potential
confounding by smoking was
considered in one study. There is
uncertainty related to exposure
patterns resulting in likely higher past
Pb exposure.
Results less consistent at lower
blood Pb level. A well-conducted
epidemiologic study at an infertility
clinic reported associations between
detrimental effects in sperm and
blood Pb levels after controlling for
smoking and other metal exposure.
A similar study also reported some
elevated effect estimates but the
results were too imprecise to draw
definitive conclusions. There is
uncertainty related to exposure
patterns resulting in likely higher past
Pb exposure.
Concurrent blood Pb
levels: > 25 ug/dL
Naha and Manna
(2007). Naha and
Chowdhury (2006),
Hsu et al. (2009b),
Kasperczyk et al.
(2008)
Groups with concurrent
blood Pb level:
<15ug/dL
Telisman et al. (2007),
Meeker et al. (2008)
Evidence describes
mode of action.
Consistent evidence in reproductive
organs of Pb-exposed male animals
(of increased apoptosis of germ cells
or spermatocytes), decreased
antioxidant activity (SOD and
CAT),and increased oxidative stress
(MDA).
Sections 4.8.3.1 and
4.8.3.2
Hormone levels
Inconsistent
findings in
epidemiologic
studies; few
studies available
There are a small number of studies
examining hormone levels and the
results are inconsistent. There is
uncertainty related to exposure
patterns resulting in likely higher past
Pb exposure.
Telisman et al. (2007).
Naha and Manna
(2007).
Hsieh et al. (2009a).
Meeker et al. (2010),
Mendiola et al. (2011)
Section 4.8.3.2
Fertility
Lack of large, well-
conducted
epidemiologic
studies but overall
inconsistent
evidence
The few epidemiologic studies
examining this outcome generally
have small samples sizes and are
drawn from men attending infertility
clinics. There is uncertainty related to
exposure patterns resulting in likely
higher past Pb exposure.
Kizileretal. (2007),
Bloom etal. (201 1b),
Bloom et al. (2010).
Gracia et al. (2005)
Section 4.8.3.3
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Table 4-48 (Continued): Summary of evidence supporting reproductive and developmental
causal determinations.
Attribute in Causal
Framework3
Limited
toxicological
evidence
Effects on Female
Key Evidence13
Paternal Pb exposure resulted in
less successful copulation, fewer
implantations, and longer periods of
time copulating for successful
matings. Unexposed females with
Pb-exposed male partners did not
have fewer pregnancies, but did
produce smaller litters.
References'3
Anjum et al. (2010)
Sainatha et al. (2011)
Pace et al. (2005)
Section 4.8.3.3
Pb Exposure or
Blood Pb Levels
Associated
with Effects0
45-day exposure of adult
male rats to 500 or
1,500 ppm Pb acetate
exposure in drinking
water, followed by
behavioral mating studies
with unexposed females.
Reproductive Function - Suggestive
A few high-quality
epidemiologic
studies of Pb levels
and hormones
demonstrate
associations but
are inconsistent
overall
Evidence in some high-quality cross-
sectional epidemiologic studies
demonstrates associations with
hormone levels but results are mixed
and vary by hormone examined and
timing in a woman's menstrual and
life cycles. In addition, the potential
confounders vary from study to
study, with some potentially
important confounders, such as BMI,
not included in all studies. There is
uncertainty related to exposure
patterns resulting in likely higher past
Pb exposure.
Concurrent mean blood
Pb levels:
<5 ug/dL
Jackson et al. (2011).
Pollack et al. (2011),
Chang et al. (2006),
Krieg (2007)
Section 4.8.4.1
Lack of large, well-
conducted
epidemiologic
studies examining
associations
between Pb levels
and fertility, but
overall inconsistent
evidence
Epidemiologic studies of this
association are limited by the small
sample sizes included in those
studies. In addition, most of the
study populations were drawn from
women undergoing IVF and/or
attending infertility clinics. There is
uncertainty related to exposure
patterns resulting in likely higher past
Pb exposure.
Section 4.8.4.2
Toxicological
studies of Pb and
effects on female
reproduction
demonstrate
effects in some
studies.
Evidence in the toxicological
literature of Pb contributing to
placental pathology and
inflammation, decreased ovarian
antioxidant capacity, altered ovarian
steriodogenesis and aberrant
gestational hormone levels.
Section 4.8.4.1 and
4.8.4.3
Described in detail in Table II of the Preamble.
bDescribes the key evidence and references contributing most heavily to causal determination and where applicable to uncertainties
or inconsistencies. References to earlier sections indicate where full body of evidence is described.
°Describes the blood Pb level in humans with which evidence is substantiated or the blood Pb levels or Pb exposure concentrations
in animals relevant to this ISA.
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4.9 Effects on Other Organ Systems
The 2006 Pb AQCD described limited evidence for the effects of Pb exposure on various
organ systems including the liver, GI tract, endocrine system, bone and teeth, eyes, and
respiratory tract. These lines of evidence largely are supported by recent toxicological
and epidemiologic studies, although the collective evidence remains relatively limited in
terms of the quantity and design of studies and/or the populations examined.
4.9.1 Effects on the Hepatic System
Hepatotoxic effects of Pb exposure (indicated in various animal models and/or human
populations) include alterations in hepatic metabolism, hepatic cell proliferation, changes
in cholesterol metabolism, as well as oxidative stress-related injury.
4.9.1.1 Summary of Key Findings of the Effects on the Hepatic
System from the 2006 Pb AQCD
The 2006 Pb AQCD (U.S. EPA. 2006b) stated that the experimental animal database
indicated hepatotoxic effects, including liver hyperplasia, at very high dose Pb exposures.
Other effects noted in the liver following exposure to Pb included altered cholesterol
synthesis, DNA synthesis, and glucose-6-phosphotase dehydrogenase (G6DP) activity.
The 2006 Pb AQCD reported that cytochrome P450 (CYP) levels decreased following
single doses of Pb nitrate (20-100 (imol/kg). Induced and constitutive expression of
microsomal CYP1A1 and CYP1A2 was inhibited by Pb exposure. Inhibition of these
(Phase I) xenobiotic metabolizing enzymes was accompanied by an increase in Phase II
enzymes following exposure to Pb nitrate and other Pb compounds, suggesting that Pb is
capable of inducing a biochemical phenotype similar to hepatic nodules. Studies
investigating Pb-induced hepatic hyperplasia suggested alterations in the gluconeogenic
mechanism, DNA hypomethylation, changes in proto-oncogene expression, as well as
altered cholesterol synthesis. Cholesterol metabolism changes following exposure to Pb
were reportedly mediated by induction of several enzymes related to cholesterol
metabolism as well as a decrease in the cholesterol catabolizing enzyme,
7 a-hydroxylase. Tumor necrosis factor alpha (TNF-a) was reported to be one of the
major mitogenic signals that mediated Pb nitrate-induced hepatic hyperplasia, based on
findings showing that inhibitors blocking TNF-a activity also blocked Pb-induced
hyperplasia. Other Pb-related effects presented in the 2006 Pb AQCD included liver cell
apoptosis and Pb-induced oxidative stress in in vitro cell cultures. The 2006 Pb AQCD
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further suggested that alterations in liver heme metabolism may involve changes in
5-aminolevulinic acid dehydrogenase (ALAD) activity, porphyrin metabolism, transferrin
gene expression and changes in Fe metabolism.
With regard to human studies, the 2006 Pb AQCD stated that increases in serum liver
enzymes suggest that Pb exposure (> 50 (ig/dL) results in nonspecific liver injury in
occupationally-exposed adults. However, studies did not adjust for potential confounding
factors, including other occupational exposures or establish explicit associations between
Pb exposure and hepatic injury (i.e., observation of histopathological effects). In addition,
similar effects were noted in animal studies, and decreased CYP activity was associated
with higher blood Pb levels in a few studies of children or adults (drawn from the general
population). The 2006 Pb AQCD reported that hepatic effects in humans were associated
only with high blood Pb levels, i.e., >30 (ig/dL.
4.9.1.2 Recent Epidemiologic Studies
A few recent epidemiologic studies examined antioxidant status and oxidative stress
effects, as measured by liver biochemical parameters, associated with occupational
exposure to Pb. However, all of these occupationally-exposed cohorts represented
populations highly exposed to Pb, resulting in mean or median blood Pb levels ranging
from 29 to 53 (ig/dL. Although the hepatic effects observed within these occupational
cohorts do not reflect the general population as a whole, they are useful in demonstrating
consistent effects on a number of liver outcomes, including altered liver function
(i.e., changes in the level of liver function enzymes), oxidative stress, and antioxidant
status (Can et al.. 2008; Khan et al.. 2008; Patil et al.. 2007). However, these studies were
cross-sectional in design with concurrent blood Pb measurement. Thus, there is
uncertainty regarding the directionality of effects and in the magnitude, timing,
frequency, and duration of Pb exposure that contributed to the observed associations.
Further, these cross-sectional analyses did not consider potential confounding by factors
such as age, diet, BMI, smoking, or other occupational exposures.
Patil et al. (2007) reported that concurrent (SD) blood Pb levels (mean [SD]: 22.32
[8.87] (ig/dL) were significantly higher (p <0.001, t-test) in 30 spray painters from
Kolhapur City in western Maharashtra, India (exposed to Pb for >6 hours/day for 2 to 20
years) than blood Pb levels (mean [SD]: 12.52 [4.08] (ig/dL) in the 35 concurrent
controls who had no history of Pb exposure and lived in rural areas. Levels of liver
function enzymes, including the two serum transaminase enzymes SGOT (also known as
AST; serum glutamic oxaloacetic transaminase/aspartate aminotransferase) and SGPT
(also known as ALT; serum glutamic pyruvic transaminase/alanine aminotransferase),
4-714
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were increased in spray painters compared to controls, whereas total serum protein levels
were decreased compared to controls (p <0.01, t-test). In another occupational study,
Conterato et al. (2013) investigated liver function parameters in automotive industry
painters and in battery manufacturing workers, who were exposed to Pb in Brazil. Mean
(SD) concurrent blood Pb levels were 5.4 (0.4) (ig/dL in the 50 Pb-exposed painters and
1.5 (0.1) (ig/dL in the 36 unexposed controls. The mean (SD) duration of exposure to Pb
in automotive industry painters was 133.9 (14.5) months (~11 years). In the Pb-exposed
paint workers, the levels of AST (but not y-glutamyltransferase), were increased nearly
2-fold compared to AST levels in controls (p <0.05). The activity of AST was positively
correlated with blood Pb levels (Spearman R = 0.26, p <0.05). The authors suggested that
confounding exposures to other toxic constituents of the paints regularly used by painters,
and not Pb, may be the etiological cause of decrements in AST function as these effects
were not also seen in battery workers with much higher blood Pb levels (49.8 (ig/dL)
(Conterato et al., 2013). Co-exposure to other environmental contaminants may also
explain the effects that were previously reported in occupationally-exposed spray-
painters in Patil et al. (2007).
4.9.1.3 Recent lexicological Studies
Hepatic Metabolism
As stated in the 2006 Pb AQCD (U.S. EPA. 2006b). acute (e.g., single dose, 5 to 10
mmol Pb/kg body wt) treatment of rodents with Pb nitrate and other Pb compounds was
found to result in a decrease in Phase I enzymes and a simultaneous increase in Phase II
enzymes. The conclusions presented in the 2006 Pb AQCD were also reviewed by
Mudipalli (2007).
Recent studies found changes in biochemical parameters, suggestive of liver damage, in
animals exposed to Pb; however, the relevance to humans is often uncertain because of
the high blood Pb levels examined in animal studies and/or the gavage or injection routes
of Pb administration used in some of the animal studies. Undernourished male Wistar rats
(fed low-protein diet without mineral supplements) exposed to 500 ppm Pb acetate in
drinking water over a 10 month period had decreases in serum protein and albumin levels
as well as increases in AST, ALT, serum alkaline phosphatase (ALP), and gamma
glutamyl transpeptidase (GGT) levels (Herman et al., 2009). In the Pb-treated animals,
the blood Pb levels steadily increased throughout the initial portion of the study period,
reaching a maximum of approximately 30 (ig/dL after 2 months. After this time, blood Pb
levels rapidly increased to approximately 110 (ig/dL by six months time, and remained at
this level until the termination of exposure at 10 months. Similar biochemical changes
4-715
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were not observed in animals treated with Pb acetate maintained on protein-adequate,
mineral rich diet.
Similarly, mice gavaged with Pb nitrate (50 mg/kg for 40 days) also demonstrated
increased activities of AST, ALT, ALP, and acid phosphatase (ACP) compared to
controls (Sharma et al.. 2010a). Upadhyay et al. (2009) reported that treatment of
Sprague-Dawley rats with Pb acetate (35 mg/kg via i.p. injection for 3 days, resulting
blood Pb level not reported) significantly increased the activities of ALT, AST, serum
ALP, and acid phosphatase over those in controls but decreased liver ALP activity.
Concomitant treatment with Zn and varying levels of vitamin C were observed to
ameliorate these toxic effects of Pb. The serum activities of glutamic pyruvic
transaminase (GPT) and lactate dehydrogenase (LDH) were similarly significantly
increased over those in controls in mice subcutaneously (s.c.) injected with Pb acetate
(50 mg/kg daily for 15 days, resulting blood Pb levels not reported) (Wang etal., 2010h).
Swarup et al. (2007) investigated serum biochemical changes in cows living in
Pb-contaminated environments. Serum levels of ALT, AST, ALP, total protein, albumin,
globulin, and A/G ratio were significantly altered in cows living near Pb-Zn smelters
(mean [SD] blood Pb level: 86 [6] (ig/dL) compared to control cows (mean [SD] blood
Pb level: 7 [1] (ig/dL). Significant positive correlations were found between blood Pb
level and ALT and AST levels, whereas a negative correlation was observed between
blood Pb level and total lipids, protein, and albumin.
Pillai et al. (2009) investigated hepatic Phase I and II enzymes in male and female rats
born to dams that were treated with Pb acetate (50 (ig/kg, via s.c. injection daily
throughout gestation and continuing until PND21). Thus, the offspring of treated dams
were exposed to Pb via placental and lactational transfer. The female and male pups were
then allowed to reach sexual maturity (PND55-PND56), in order to assess continuing
exposure to bioaccumulated Pb. The activities of hepatic Phase I enzymes NADPH- and
NADH-cytochrome c reductase were significantly reduced in Pb-exposed male and
female rats on PND56 (resulting blood Pb levels not reported), compared to controls. In
rats treated with Pb and Cd (25 (ig/kg), the effect on Phase I enzymes was increased. Pb
treatment additionally decreased the activities of Phase II enzymes uridine diphosphate-
glucoronyl transferase and GST in males and females, but no effect was observed on
GGT or 17(3-hydroxysteroid oxidoreductase. Additionally, no effect was observed on
serum glutamate pyruvate dehydrogenase or ALP activities in the Pb-treated male or
female rats. Histological observations demonstrated fatty degeneration of the liver,
vacuolization, and pycnotic nuclei, indicating general hepatotoxicity following Pb
treatment in both male and female rats.
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In a similar study, Teijon et al. (2006) exposed Wistar rats to Pb acetate (200 or 400 ppm
drinking water) throughout gestation, lactation, and continuing through 1 and 3 months
postweaning. Compared to the controls, the concentrations of Pb in the liver of the male
pups were elevated in the 200- and 400-ppm groups at both 1 and 3 months postweaning
(only male pups were followed after birth). Liver concentrations of Pb were greater in the
200 ppm animals compared to the 400 ppm animals at one month postweaning (mean
[SE]: 1.19 [0.30] (ig Pb/g tissue versus 0.76 [0.06] \ig Pb/g tissue, respectively), but were
similar between the 2 dosing regimens (200 ppm versus 400 ppm) by 3 months
postweaning (mean [SE]: 0.54 [0.06] versus 0.55 [0.07] ng Pb/g tissue, respectively).
Serum ALP activity was significantly increased over controls only at 2 weeks
postweaning, whereas ALT activity was significantly decreased from controls only at 2
months and at 3 months postweaning. Cheng et al. (2006) studied the mechanism of Pb
effects on bacterial lipopolysaccharide (LPS)-induced TNF-a expression. A/J mice were
injected with Pb acetate (100 (imol/kg via i.p.), with or without LPS (5 mg/kg). Pb alone
did not affect AST or ALT activity or the level of TNF-a in the serum of the mice. In
comparison, treatment of mice with low doses of Pb and LPS together caused a
statistically significant increase in TNF-a induction as well as enhanced liver injury,
suggesting that Pb potentiated LPS-induced inflammation. In a complementary in vitro
experiment, co-exposure of Pb and LPS stimulated the phosphorylation of p42/44
mitogen-activated protein kinase (MAPK) and increased TNF-a expression in mouse
whole blood cells, peritoneal macrophages, and RAW264.7 cells (a macrophage cell
line). These results indicated that Pb increased LPS-induced TNF-a levels via the protein
kinase C (PKC)/MAPK pathway in monocytes/macrophages rather than hepatocytes.
Similarly, Pb chloride potentiated bovine serum albumin (BSA)-induced inflammation in
the livers of mice subcutaneously injected with Pb (Saet al.. 2012).
Lipid Metabolism
Several recent toxicological studies indicated Pb-induced impaired lipid metabolism, as
evidenced by increases in liver cholesterol. There was some evidence in animals exposed
to Pb in diet, albeit at relatively high exposure concentrations or measured blood Pb
levels. Ademuyiwa et al. (2009) reported that male albino Sprague Dawley rats exposed
to 200, 300 and 400 ppm Pb acetate in drinking water had mean (SD) blood Pb levels of
40.63 (9.21), 61.44 (4.63), and 39.00 (7.90) (ig/dL, respectively. Animals exposed to
200 ppm Pb had mean (SD) liver Pb concentrations of 10.04 (1.14) (ig/g, compared to
3.24 (1.19) (ig/g and 2.41 (0.31) (ig/g in animals exposed to 300 or 400 ppm Pb,
respectively. The animals exposed to Pb exhibited increased hepatic cholesterogenesis at
all doses tested compared to controls. Additionally, a decrease in triglyceride levels was
observed at 300 and 400 ppm Pb; a decrease in phospholipid levels was observed at
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400 ppm Pb. The authors also reported positive correlations between tissue cholesterol
and phospholipids, and Pb accumulation in liver across all doses. In contrast, the
association between tissue triglyceride levels and Pb accumulation was negative. In
related studies, Khotimchenko and Kolenchenko (2007) reported that adult male albino
rats treated with Pb acetate (100 mg/kg for 14 days, resulting blood Pb levels not
reported) exhibited disorders in lipid metabolism that were accompanied by increased
levels of total cholesterol and triglycerides in the liver tissue. Sharma et al. (2010a)
reported increased liver cholesterol in mice gavaged with Pb nitrate (50 mg/kg for
40 days). Pillai et al. (2009) observed decreases in total liver cholesterol in PND56 male
and female rats that had been treated with Pb acetate (via s.c. injection, 50 (ig/kg,
continuously throughout gestation and lactation). These results suggest that Pb induction
of cholesterogenesis and phospholipidosis in the liver may cause subtle effects at the
cellular level that may lead to hepatotoxicity.
Kojima and Degawa (2006) examined sex-related differences in Pb-induced gene
expression of a rate limiting hepatic cholesterol biosynthesis enzyme, 3-hydroxy-3-
methylglutaryl-CoA reductase (HMGR) and its transcription factor, sterol regulatory
element binding protein-2 (SREBP-2). Male and female Sprague Dawley rats were
injected with Pb nitrate (100 (imol/kg body weight, intravenously [i.v.], resulting blood
Pb levels not reported). SREBP-2 expression was significantly increased in males and
females with this increase occurring earlier in male rats (6-12 hours, compared to 24-36
hours in females). In contrast, expression of HMGR was significantly increased in both
Pb-exposed males and females at earlier time frames and greater range of onset (3-48
hours in males; 12-48 hours in females) compared to that of SREBP-2. Significant
increases in total liver cholesterol were also observed in Pb-exposed males and females at
3-48 and 24-48 hours, respectively. These results suggest that the SREBP-2 and HMGR
gene expression increases and increase in total cholesterol levels in the liver, in response
to Pb exposure, all occur earlier in males compared to females; and also suggest that the
increase in HMGR gene expression and increase in total cholesterol levels in the liver
occur before an increase in the SREBP-2 gene expression in both sexes.
Hepatic Oxidative Stress
A number of studies demonstrated increased hepatic oxidative stress as a result of
exposure to various Pb compounds, measured by increases in reactive oxygen species
(ROS) or decreases in antioxidant levels or enzyme activity. ROS can potentially result in
damage to hepatic function and structure. Several of these observations were made in
animals exposed to Pb in drinking water that produced blood Pb levels relevant to this
ISA. In a study examining the effects of Pb exposure to fetuses, Masso et al. (2007)
exposed pregnant Wistar rats to 300 ppm Pb in drinking water from GDI to parturition
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(PNDO), or to weaning (PND21). Blood Pb levels were higher at parturition (mean [SD]:
31.5 [0.80] (ig/dL) than at weaning (mean [SD]: 22.8 [0.50] (ig/dL). Pups exhibited liver
damage that was accompanied by an increased production of thiobarbituric acid-reactive
species (TEARS, an indicator of lipid peroxidation) and increased CAT activity
compared to controls. In addition, increased ALP and acid phosphatase activity was
observed. Uzbekov et al. (2007) showed differential effects by duration of maternal
exposure before mating. Female Wistar rats were exposed to 0.3 and 3.0 ppm Pb nitrate
in drinking water, for 1 month or 5 months prior to pregnancy, and also continuing during
pregnancy; then the livers from both groups of GD20 fetuses were examined for hepatic
SOD activity. The pregnant control rats had a mean (SD) blood Pb level of 16.1
(0.63) (ig/dL, whereas the dams exposed to 0.3 and 3.0 ppm Pb had mean blood Pb levels
of 20.4 (ig/dL and 24.4 (ig/dL, respectively. In the GD20 fetuses from dams exposed for
1 month prior to pregnancy, a concentration-dependent increase in liver SOD activity was
observed, whereas SOD activity was decreased in the GD20 fetuses from dams exposed
for 5 months prior to pregnancy. The increase in SOD activity in the livers of fetuses
from dams exposed to 0.3 or 3.0 ppm Pb nitrate for one month suggests an initial
activation of SOD in response to increased free radical production, while the decrease in
SOD production in fetal livers from dams exposed to the same concentrations for
5 months suggests that longer durations of Pb exposure impairs the antioxidant defense
mechanism.
Increased hepatic oxidative stress was also found in animals with postnatal Pb exposure
in drinking water. Jurczuk et al. (2007) reported that adult male Wistar rats (PND56)
treated for 12 weeks with 500 ppm Pb in drinking water (resulting blood Pb level not
reported) exhibited decreases in liver vitamin E and GSH levels along with an increase in
lipid peroxidation. The correlation between vitamin E and lipid peroxidation suggested
that vitamin E is involved in the mechanism of peroxidative action of Pb in the liver. In a
study examining the role of low molecular weight thiols on peroxidative mechanisms,
Jurczuk et al. (2006) found that adult male Wistar rats (PND56) treated for 12 weeks with
500 ppm Pb acetate in drinking water exhibited a decrease in blood ALAD as well as
decreases in GSH and nonprotein sulfhydryl levels in the liver. Metallothionein levels
were also reported to be higher in the liver following exposure to Pb. Yu et al. (2008)
reported concentration-dependent increases in lipid peroxide levels and decreases in GSH
levels and CAT, SOD and GPx activities in livers from castrated male pigs that received
a diet mixed with 0, 5, 10, or 20 mg/kg Pb nitrate exposure, during ages 62-107 days. The
level of hepatic CuZnSOD mRNA was also reduced in Pb-treated animals. The study
authors suggested that this decrease in SOD mRNA expression and activity of antioxidant
enzymes may lead to a reduction in free radical scavenging capability and increased lipid
peroxidation.
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Studies administering Pb by bolus doses had similar findings. Adegbesan and Adenuga
(2007) reported that lipid peroxidation was increased and SOD activity was decreased in
protein undernourished male Wistar rats compared to well-fed control rats, and that these
effects were further exacerbated in protein undernourished rats receiving a single i.v.
injection of Pb nitrate (100 (imol/kg, resulting blood Pb levels not reported). Protein
undernourishment also decreased GSH levels and CAT activity compared to normal diet;
however, co-treatment with Pb mitigated the severity of these effects. GSH levels and
CAT activity were still lower in undernourished rats with Pb exposure compared to well-
fed rats, but greater than undernourished rats with no Pb exposure. The results indicated
Pb treatment exacerbated the effects of malnutrition on liver lipid peroxidation and
altered the involvement of free radicals. Adult male Charles-Foster rats treated for 15
days with Pb acetate (0.025 mg/kg per day [via i.p. injection], resulting blood Pb levels
not reported) also exhibited statistically significant increases in lipid peroxidation levels
and decreases in SOD, CAT, and G6PD levels in liver mitochondrial and
postmitochondrial fractions (Pandyaet al.. 2010). Statistically nonsignificant decreases
were observed in GSH levels, and in GPx and GR activities in these Pb-treated rats. In
mice gavaged with Pb nitrate (50 mg/kg for 40 days), lipid peroxidation was increased,
and SOD, CAT, and GSH were decreased compared to controls (Sharma et al., 2010a).
Additionally, Pb nitrate treatment resulted in histopathological changes in the structure of
the liver: hepatocytes were damaged and were marked by cytoplasmic vacuolization and
pycnotic nuclei.
Khotimchenko and Kolinchenko (2007) also reported an increase in lipid peroxidation
and development of hepatitis in adult male albino rat liver parenchyma following
treatment with Pb acetate (100 mg/kg per day, via gavage, for 14 days [resulting blood Pb
levels not reported]). The increased lipid peroxidation was measured by increases in
malondialdehyde (MDA) levels along with decreases in GSH and thiol groups; indicating
injury in the liver antioxidant system. In another study, levels of hepatic lipid
peroxidation were observed to be significantly increased in rats injected (i.p.) with
Pb acetate (35 mg/kg per day for 3 days, resulting blood Pb levels not reported), whereas
hepatic GSH was significantly decreased (Upadhyav et al.. 2009). A study examining
male and female rat pups that were continuously exposed to Pb during gestation and
lactation (pregnant dams were injected [s.c.] with Pb acetate, 50 (ig/kg per day [GDO to
PND21], resulting blood Pb levels not reported), did not find effects on GSH or MDA
levels at PND56 (Pillai et al.. 2009). In vitro exposure of cells from a hepatic human
embryonic epithelial cell line (WRL-68) to 5 (iM Pb acetate for 30 days resulted in
increased production of ROS throughout the incubation period (Hernandez-Franco et al..
2011). Concurrent with this increase in ROS generation, the activities of SOD and the
levels of membrane lipid peroxidative damage increased throughout the first 24 days of
in vitro exposure but returned to normal levels by exposure day 30.
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Hepatic Apoptosis
Fan et al. (2009b) reported that a single i.v. injection of Pb nitrate (200 (imol/kg in 0.5
mL) in adult male Sprague-Dawley rats resulted in an increase in the percentage of
apoptotic hepatocytes (mean: 2.5% [SD: 1.4] of total hepatocytes) compared with
controls (mean: 0.31% [SD: 0.31]). Expression of ferritin light-chain (FTL) protein also
increased (mean [SD]: 3.5 [1.0]-fold increase) over that in controls.
Immunohistochemical analysis revealed that hepatocytes around the hepatic lobules'
central veins were heavily stained by anti-FTL antibodies, as were nonparenchymal cells
identified as Kupffer macrophage cells. The authors hypothesized that the expression of
FTL in Kupffer cells may have resulted from the phagocytic engulfment of the apoptotic
cells by the activated Kupffer cells. Treatment of rats with clofibrate, a lipid lowering
agent, also increased FTL protein in hepatocytes around the central veins, but did not
increase FTL expression in Kupffer cells; and induced hepatocellular proliferation, but
not apoptosis.
4.9.1.4 Summary of Effects on the Hepatic System
While explicit associations between hepatic injury (i.e., histopathological effects) and Pb
exposure have not been established, evidence from epidemiologic and toxicological
studies have indicated that exposure to Pb can result in altered liver function and hepatic
oxidative stress. A few studies have reported associations between higher blood Pb levels
and decreased CYP enzymes (Phase I xenobiotic metabolism) in children and
nonoccupationally-exposed adults. However, most occupational evidence indicates
decreases in serum protein and albumin levels and increased AST, ALT, ALP, and GGT
activities (indicators of decreased liver function), increased oxidative stress, and
decreased antioxidant status in Pb-exposed workers with blood Pb levels >29 (ig/dL
(Conterato et al.. 2013: Can et al.. 2008: Khan et al.. 2008: Patil et al.. 2007). The
implications of the occupational epidemiologic evidence are limited because of the cross-
sectional design of the studies, the high blood Pb levels examined, and the lack of
consideration for potential confounding by factors such as age, diet, BMI, smoking, or
other occupational exposures.
Similar changes in liver function enzymes have been found in mature animals exposed to
high levels of Pb during adulthood (Sharma et al., 2010a; Wang et al.. 2010h; Herman et
al.. 2009; Cheng et al.. 2006). and animals exposed during gestation and lactation (Pillai
et al.. 2009; Teijon et al.. 2006). Pb exposure has been shown to impair lipid metabolism
in animals, as evidenced by increased hepatic cholesterogenesis, and altered triglyceride
and phospholipid levels (Ademuviwa et al.. 2009; Khotimchenko and Kolenchenko.
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2007). Multiple studies in humans and animals have observed Pb-associated increases in
hepatic oxidative stress, generally indicated by an increase in lipid peroxidation along
with a decrease in GSH levels and CAT, SOD, and GPx activities (Pandva et al., 2010;
Sharmaetal.. 2010a: Khan et al.. 2008: Yu etal. 2008: Adegbesan and Adenuga. 2007:
Jurczuk et al.. 2007: Khotimchenko and Kolenchenko. 2007: Jurczuk et al., 2006).
Indices of increased oxidative stress were also observed in the livers of rat fetuses
exposed to Pb throughout gestation (Masso et al.. 2007). The relevance of the
toxicological evidence is uncertain, as many studies administered Pb as bolus doses.
Additionally, few toxicological studies reported resulting blood Pb levels. Of those
studies that did report blood Pb levels, most were so high (>30 (ig/dL) as to limit their
relevance to humans. Because of the insufficient quality of studies, the evidence is
inadequate to determine if there is a causal relationship between Pb exposure and hepatic
effects.
4.9.2 Effects on the Gastrointestinal System
Gastrointestinal effects examined in relation to Pb exposure include abdominal pain,
constipation, and internal paralysis in humans and degeneration of the intestinal epithelial
mucosa and a decrease in duodenal motility in animals.
4.9.2.1 Summary of Key Findings on the Effects on the
Gastrointestinal System from the 2006 Pb AQCD
The 2006 Pb AQCD (U.S. EPA. 2006b) stated that a number of factors influence the
gastrointestinal absorption of Pb; including the chemical and physical form of Pb, the
developmental age at Pb intake, as well as various nutritional factors.
Rats exposed to Pb acetate in drinking water (0.1%) had degeneration of the intestinal
epithelial mucosa, potentially leading to malabsorption of nutrients. Uptake of Pb via oral
exposure was shown to be dependent on source of Pb. Pb administered in bovine or rat
milk, or infant milk formula is associated with casein micelles, and was observed to
accumulate in ileal tissues, whereas Pb administered in drinking water as a soluble salt
was observed to accumulate in duodenal tissues. This difference in Pb uptake suggests
that Pb exposure via drinking water may be more toxic than Pb exposure via milk. In a
different study reviewed in the 2006 Pb AQCD, decreases in duodenal motility and the
amplitude of contractility in the intestinal tract were observed in rats following Pb acetate
exposure (gavage, 44 mg/kg per day for 4 weeks) . Nutritional studies using chicks
examining different dietary levels of Pb, Ca2+, and vitamin D in rats indicated
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competition in absorption between Pb and Ca2+. Dietary supplementation with vitamin D
led to an increase in intestinal absorption of Pb and Ca2+. In instances where severe Ca2+
deficiency was noted, ingestion of Pb caused a clear decrease in 1,25-dihydroxy vitamin
D (1,25-(OH)2D3) levels. Overall, the 2006 Pb AQCD stated that studies in rat intestine
have shown that the largest amount of Pb absorption occurs in the duodenum with the
mechanisms of absorption involving active transport and diffusion via the intestinal
epithelial cells. Absorption has been reported to occur, through both saturable and
nonsaturable pathways, based on results from various animal studies.
The 2006 Pb AQCD also reported on human evidence that symptoms associated with
gastrointestinal colic (abdominal pain, constipation, intestinal paralysis) were more
prevalent in occupationally-exposed adults with blood Pb levels > 50 (ig/dL.
4.9.2.2 Recent Epidemiologic Studies
Consistent with previous occupational findings, Kuruvilla et al. (2006) reported
gastrointestinal effects in men, including stomach pain and gastritis along with other non-
gastrointestinal effects in 53 Pb-exposed paint workers (mean [SD] blood Pb: 8.04
[5.04] (ig/dL) in India compared with 50 controls (mean [SD] blood Pb level: 5.76
[4.45] (ig/dL) matched by sex, age, education, income, smoking, and alcohol
consumption. Fifty-two male battery workers were also examined. Prevalence of
gastrointestinal symptoms in painters did not differ from that in battery workers with
higher blood Pb levels (mean [SD] blood Pb level of 42.40 ([25.53] (ig/dL). Despite the
consistent evidence with earlier occupational studies, the implications of gastrointestinal
findings in Pb-exposed workers are limited by the cross-sectional study designs, high
blood Pb levels associated with effects (mostly > 50 (ig/dL), and limited consideration for
potential confounding by factors such as age, smoking, alcohol use, nutrition, or other
occupational exposures.
4.9.2.3 Recent lexicological Studies
A few recent studies pertaining to gastrointestinal effects of Pb exposure were identified
that provide evidence for additional mechanisms underlying gastrointestinal damage and
impaired function. Santos et al. (2006) examined the impact of Pb exposure on
relaxations to nonadrenergic noncholinergic (NANC) nerve stimulation in rat gastric
fundus. Male Wistar rats treated with 80 ppm Pb acetate via drinking water for 15, 30,
and 120 days (resulting blood Pb levels not reported) exhibited a significant difference in
NANC relaxations in the gastric fundus following electrical field stimulus. While
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frequency-dependent (0.5 - 8 Hz) relaxations were observed in all groups, including the
control group, the relaxations were significantly inhibited in rats treated with the 80 ppm
Pb acetate for all three durations (15,30, and 120 days). When gastric fundus strips from
rats were incubated with L-nitroarginine, a nitric oxide (NO) synthase inhibiter, no
additional inhibition in relaxations was observed. In contrast, incubation with sodium
nitroprusside and 8-Br-GMPc (a cyclic guanosine monophosphate [cGMP] analog)
resulted in a concentration-dependent relaxation in strips from the control group and from
the group that had been exposed to the 80 ppm Pb acetate for 120 days. The results
suggested that long-term exposure to Pb causes inhibition in NANC relaxation, probably
due to the modulated release of NO from the NANC nerves or due to interaction with the
intracellular transducer mechanism in the rat gastric fundus.
In a study examining Pb-induced oxidative stress in the gastric mucosa, Olaleye et al.
(2007) treated young adult (80-90 grams) male albino Wistar rats with 100 or 5,000 ppm
Pb acetate in drinking water for 15 weeks (resulting blood Pb levels not reported). After
the 15 week Pb acetate treatment, gastric ulcers were induced by a 4-hour exposure to
acidified ethanol. The 15-week pre-exposure to Pb acetate led to a significant increase in
gastric mucosal damage induced by the acidified ethanol, compared to the damage in the
control group. Gastric ulcers were not induced in all of the animals. After the 15 week
exposure to Pb acetate (but before any acidified ethanol treatment), stomach gastric acid
secretion and biochemical analyses were performed in some of the rats. While the basal
gastric acid secretory rate was not altered, stomach response to histamine was
significantly higher in animals treated with the Pb acetate compared to that in the
controls. Additionally, there was a significant increase in gastric lipid peroxidation at
both the 100 and 5,000 ppm dosage levels. In contrast, CAT, and SOD activities and
nitrite levels were significantly decreased in the gastric mucosa. These results indicated
that Pb-induced gastric damage may be mediated via increases in oxidative stress.
4.9.2.4 Summary of Gastrointestinal Effects
Relatively few human studies have been conducted on the gastrointestinal toxicity of Pb.
The evidence points to more prevalent symptoms, such as stomach pain, gastritis,
constipation, and intestinal paralysis, in Pb-exposed workers (Kuruvilla et al.. 2006).
However, the implications of gastrointestinal findings in Pb-exposed workers are limited
by the cross-sectional study designs, high blood Pb levels associated with effects (mostly
> 40 (ig/dL), and limited consideration of potential confounding by factors such as age,
smoking, alcohol use, nutrition, or other occupational exposures. Toxicological evidence
indicates that Pb is absorbed primarily in the duodenum by active transport and diffusion,
although variability is observed by Pb compound, age of intake, and nutritional factors.
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There is some coherence between the evidence in Pb-exposed workers and observations
in animals that Pb induces damage to the intestinal mucosal epithelium, decreases
duodenum contractility and motility, reduces absorption of Ca2+, inhibits NANC
relaxations in the gastric fundus, and induces oxidative stress (lipid peroxidation,
decreased SOD and CAT) in the gastric mucosa (Olaleve et al., 2007; Santos et al..
2006). The observation of oxidative stress was accompanied by gastric mucosal damage.
Because of the insufficient quantity and quality of studies, the evidence is inadequate to
determine if there is a causal relationship between Pb exposure and gastrointestinal
effects.
4.9.3 Effects on the Endocrine System
A summary of key findings pertaining to reproductive hormones in males and females is
presented in the section on Reproductive and Developmental Effects (Sections 4.8.3 and
4.8.4). Collective epidemiologic and toxicological evidence is inconsistent in
demonstrating the effects of Pb exposure on male and female sex hormone levels. Other
endocrine processes that are most commonly found to be impacted by Pb exposure
include changes in thyroid hormones, including thyroid stimulating hormone (TSH),
triiodothyronine (T3), and thyroxine (T4). A few studies have examined Ca2+ and
cortisol.
4.9.3.1 Summary of Key Findings of the Effects on the Endocrine
System from the 2006 Pb AQCD
The 2006 Pb AQCD (U.S. EPA, 2006b) reported that adult human endocrine processes
impacted by occupational Pb exposure include thyroid hormone levels, changes in male
sex hormone levels, as well as changes in the production of vitamin D (1,25-(OH)2D3).
However, these effects were observed only with blood Pb levels exceeding 30-40 (ig/dL
and in cross-sectional studies with little or no consideration for potential confounding by
factors such as sex, SES, nutritional status, BMI, smoking, comorbid conditions, and
other occupational exposures. In addition, alterations in calcitropic hormones were found
in children with blood Pb levels ranging from 10-120 (ig/dL and in an opposite direction
than that in Pb-exposed workers.
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4.9.3.2 Recent Epidemiologic Studies
Recent epidemiologic studies have reported associations between indicators of exposure
to Pb and thyroid hormone levels, in populations of children and in adults (with and
without occupational Pb exposure); although results have not been consistent for a
particular hormone. Further, the implications of these findings are limited because of the
use of cross-sectional study design, lack of rigorous statistical analysis, and limited
consideration for potential confounding factors. Inconsistent associations were found in a
study that did consider potential confounding factors. Abdelouahab et al. (2008)
examined an adult Canadian population characterized by high consumption of freshwater
fish contaminated with Pb and other environmental pollutants. The median concurrent
blood Pb level was 3.1 (ig/dL for men and 1.7 (ig/dL for women. The median blood Pb
level for women was lower than the limit of detection (2.1 (ig/dL), resulting in
measurement error of blood Pb level and greater uncertainty in the results. In an analysis
stratified by sex, TSH levels were negatively correlated with blood Pb in women with
adjustment for age, smoking status, estro-progestative intake, total plasma lipids, and
total blood selenium (Se). No associations with T3 and T4 levels were found in women.
TSH, T3 and T4 levels were not correlated with blood Pb level in males, after adjustment
for the same covariates (excluding hormone intake) plus pesticide exposure, corticoid
medication, concurrent alcohol consumption, and occupational exposure to metals.
Overall, the inconsistent associations and potential influence of other exposures did not
strongly demonstrate an effect of Pb exposure.
Studies with less rigorous methods also did not clearly indicate an association between
blood Pb level and a particular thyroid hormone. In a Kosovo, Yugoslavia population of
pregnant women, higher pregnancy blood Pb levels were associated with lower free T4
levels during pregnancy among the 156 women living in a highly exposed town with a
smelter and battery plant, but not among the 153 women living in a relatively unexposed
nearby town (Lamb et al., 2008). The mid-pregnancy blood Pb levels were highly
elevated in the industrial town compared to the unexposed town (mean [SD]: 20.56 [7.38]
versus 5.60 [1.99] (ig/dL). In 24 newborns delivered in Tokyo, Japan, neither TSH nor
free T4 (sampled 4-6 days postpartum) was correlated with cord blood Pb level (mean:
0.67 (ig/dL) (lijima et al., 2007). Neither of these two studies considered potential
confounding. Croes et al. (2009) examined the hormone levels in 1,679 adolescent girls
and boys residing in nine study areas in the Flanders Region of Belgium, with varying
exposures to multiple industrial pollutants including Pb. The median concurrent blood Pb
level of the participants from the nine different study areas ranged from 1.6 to 2.8 (ig/dL.
Analyses only indicated differences in free T3, at the region or neighborhood level with
adjustment for age, sex, recent disease, and BMI. No direct associations with blood Pb
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level were analyzed, thus the results could be attributed to other factors that varied by
location.
Contrasting results were found for free T4 in Pb-exposed workers. Dundar et al. (2006)
examined associations between blood Pb levels and thyroid function in 42 male
adolescent auto repair apprentice workers, with no history of prior disease and exposed
long term to Pb (15-17 year-olds in the auto repair apprenticeship at least 1 year). Mean
blood Pb level was higher in the auto repair workers compared to the 55 healthy
unexposed control subjects (mean [SD]: 7.3 [2.92] versus 2.08 [1.24] (ig/dL). Free T4
levels were significantly lower in the auto workers compared to the control group, which
had no abnormal free T4 levels reported. In contrast, free T3 and TSH levels were
comparable between auto workers and controls. Blood Pb level was negatively correlated
with free T4 levels. In contrast, another study (Pekcici et al.. 2010) found higher free T4
and TSH in adult auto mechanic or battery factory workers who were highly exposed to
Pb (mean blood Pb: 71.1 (ig/dL) compared to controls (mean blood Pb level: 0.2 (ig/dL).
Free T3 levels were similar between the two groups. The results from this study are likely
not generalizable to the general public due to the high blood Pb levels of the exposed
workers.
Previous findings for blood Pb-associated changes in serum vitamin D (1,25-(OH)2D3)
in children were mixed. A recent study in New Jersey examined winter (December to
March) and summer (July to September) seasonal changes in the associations between
blood Pb level and serum 1,25-(OH)2D3 status, in 142 young, U.S. urban African-
American or Hispanic children (ages 1-8 years, grouped by age [1-3 year-olds and 4-8
year-olds] and race/ethnicity) using a repeated measures design (Kemp et al.. 2007). The
percentage of 1-3 year-old African-American children (n = 49) with blood Pb levels
> 10 (ig/dL increased from 12.2% in winter to 22.5% in summer. This large seasonal
increase in blood Pb levels in these 1-3 year-old children was not accompanied by a
significant increase in serum 1,25-(OH)2D3 concentrations. There was also a larger
seasonal increase in blood Pb levels in 1-3 year-old children from both races combined
(n = 78) (mean [SE]: 4.94 [0.45] (ig/dL winter, 6.54 [0.82] (ig/dL summer) than in 4-8
year-old children from both races combined (n = 64) (mean [SE]: 3.68 [0.31] (ig/dL
winter, 4.16 [0.36] (ig/dL summer). However, no difference in seasonal 1,25-(OH)2D3
was observed in the 1-3 year-old children from both races combined. A larger winter to
summer increase in blood Pb level was correlated with a larger seasonal increase in
serum 1,25-(OH)2D3 in the 4-8 year-old children from both races combined and in the
4-8 year-old African American children (n = 42). In the 4-8 year-old children from both
races combined, there was a winter to summer increase in 1,25-(OH)2D3 (mean [SE]:
25.3 [1.2] (ig/L in winter versus 33.8 [1.1] (ig/L in summer), which may account for the
results in this older age group. Based on these results, the study authors concluded that
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higher summertime increase in serum 1,25-(OH)2D3 levels in children between ages 4
and 8 years is most likely due to increased sunlight-induced vitamin D synthesis and may
be a contributing factor to seasonal changes in blood Pb levels via changes in
gastrointestinal absorption or release of Pb from bone.
Hypothalamic-pituitary-adrenal (HPA) axis function was examined in a prospective
analysis of associations between prenatal maternal blood Pb levels (indicated by cord
blood collected at delivery) or postnatal blood Pb levels (at mean [SD] age: 2.62 [1.2]
years, data obtained from family physicians or state records), and saliva cortisol levels
during a stress protocol in the children at age 9.5 years (Gump et al.. 2008). For prenatal
blood Pb, the children were divided into the following quartiles: [< 1], [1.1-1.4],
[1.5-1.9], and [2.0-6.3] (ig/dL. For postnatal blood Pb, the quartiles were: [1.5-2.8],
[2.9-4.1], [4.2-5.4], and [5.5-13.1] (ig/dL. With adjustment for potential confounding by
SES-related factors, HOME score, pregnancy health, maternal substance abuse, blood Pb
level was not associated with initial salivary cortisol levels in the children when they
were older (9.5 years old), before the stress challenge protocol commenced. However,
following an acute stressor, which comprised of submerging the dominant arm of each
9 year-old child for a minute in a gallon of one part ice to one part water, increasing
prenatal and postnatal blood Pb levels were associated with statistically significant
increases in salivary cortisol responses. Children in the 2nd, 3rd, and 4th prenatal blood
Pb quartiles and in the 4th postnatal quartile had increased salivary cortisol responses
compared to children in the 1st quartile. When blood Pb was treated as a continuous
variable, regression analysis showed that both prenatal and postnatal blood Pb levels
were associated salivary cortisol reactivity. While associations were found in children
with blood Pb levels below 10 (ig/dL, they could have been attributed to higher earlier
childhood blood Pb levels of these children who were born in 1980-1990s.
4.9.3.3 Recent lexicological Studies
Pb-associated changes in thyroid hormones also were found in animal studies. In a study
examining the effects of Pb and Cd in adult cows reared in a polluted environment in
India, Swarup et al. (2007) found significantly higher mean plasma T3 and T4 levels in
cows living near Pb/Zn smelters (mean [SD] blood Pb: 86 [6] (ig/dL) and near closed
Pb/operational Zn smelters (mean [SD] blood Pb: 51 [9] (ig/dL) when compared to cows
in unpolluted areas (mean [SD] blood Pb: 7 [1] (ig/dL). Regression analyses of the 269
cows showed a significant positive correlation between blood Pb levels and plasma T3
and T4 levels, whereas the correlation between blood Pb levels and plasma cortisol was
not statistically significant. Mean plasma estradiol level was significantly higher in cows
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near the closed Pb/operational Zn smelters compared to the control group of cows.
Because of the Pb-Zn co-exposure, the effects cannot be attributed specifically to Pb.
Biswas and Ghosh (2006) investigated the effect of Pb treatment on adrenal and male
gonadal functions in Wistar rats treated with Pb acetate (8.0 mg/kg via i.p. injection for
21 days, resulting blood Pb levels not reported). Pb treatment significantly increased
adrenal steroidogenic enzyme activity and serum corticosterone levels. Accessory sex
organ (prostate and seminal vesicle) weights were decreased in Pb-treated animals,
whereas adrenal weights were increased. These effects were accompanied by a decrease
in spermatogenesis and serum concentrations of testosterone, FSH, and LH and by an
increase in the percent of spermatid degeneration. Supplementation with testosterone
during the last 14 days of Pb treatment was observed to ameliorate these effects.
4.9.3.4 Summary of Endocrine Effects
Collective epidemiologic and toxicological evidence is inconsistent in demonstrating the
effects of Pb exposure on male and female sex hormone levels (Sections 4.8.3 and 4.8.4)
and vitamin D levels. Several epidemiologic studies have reported associations between
indicators of Pb exposure and thyroid hormone levels in populations of children and
adults without (Lamb et al., 2008) and with occupational Pb exposure (Dundar et al.,
2006). although results have not been consistent for a particular hormone. Further, the
implications of these findings are limited because of the cross-sectional study design,
lack of rigorous statistical analysis, and limited consideration for potential confounding
factors. Blood Pb level was positively correlated with plasma T3 and T4 levels in adult
cows living near Pb-Zn smelters; however, the effects could not be attributed specifically
to Pb exposure (Swarup et al., 2007).
In a prospective study of children in New York, who were challenged with an acute
stressor, higher cord blood levels (as a reflection of prenatal maternal Pb blood level), or
2-year-old blood Pb levels, were associated with significant higher salivary cortisol in
response to a stress challenge at age 9 years (Gump et al. 2008). While these associations
were found with blood Pb levels <10 (ig/dL, they could have been attributed to higher
earlier childhood blood Pb levels of these children who were born in the 1980s and
1990s. Biswas and Ghosh (2006) found a Pb-induced increase in corticosterone in rats,
albeit by i.p. Pb treatment. Cortisol and corticosterone are the major glucocorticoids in
humans and rodent, respectively.
In conclusion, epidemiologic and toxicological evidence indicates Pb-associated
endocrine effects such as thyroid hormones, cortisol, and vitamin D, although results are
not consistent. Because of the lack of insufficient quantity and quality of studies, the
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evidence is inadequate to determine if there is a causal relationship between Pb exposure
and endocrine effects related to thyroid hormones, cortisol, and vitamin D.
4.9.4 Effects on Bone and Teeth
Primary effects on bone associated with Pb exposure or biomarker levels have included
an increase in osteoporosis, increased frequencies of falls and fractures, changes in bone
cell function as a result of replacement of bone calcium with Pb, and depression in early
bone growth. Other effects include tooth loss and periodontitis. Mechanistic evidence
from toxicological studies includes effects on cell proliferation, procollagen type I
production, intracellular protein, and osteocalcin in human dental pulp cell cultures.
4.9.4.1 Summary of Key Findings of the Effects on Bone and
Teeth from the 2006 Pb AQCD
The 2006 Pb AQCD reported many effects on bone and some in teeth in animals
following Pb exposure. Exposure of animals to Pb during gestation and the immediate
postnatal period was reported to significantly depress early bone growth with the effects
showing concentration-dependent trends. In mature animals, long-term Pb exposure (up
to one year), along with poor nutrition (low calcium) reduced bone growth as well as
bone density. Systemic effects of Pb exposure included disruption in bone mineralization
during growth, alteration in bone cell differentiation and function due to alterations in
plasma levels of growth hormones and calcitropic hormones such as 1,25-[OH]2D3 and
impact on Ca2+- binding proteins and increases in Ca2+ and phosphorus concentrations in
the bloodstream. Bone cell cultures exposed to Pb had altered vitamin D-stimulated
production of osteocalcin accompanied by inhibited secretion of bone-related proteins
such as osteonectin and collagen. In addition, Pb exposure caused suppression in bone
cell proliferation most likely due to interference from factors such as growth hormone
(GH), epidermal growth factor (EOF), transforming growth factor-beta 1 (TGF-(31), and
parathyroid hormone-related protein (PTHrP).
As in bone, Pb exposure was found to easily substitute for Ca2+ in the teeth and was taken
up and incorporated into developing teeth in experimental animals. Since teeth do not
undergo remodeling like bone does during growth, most of the Pb in the teeth remains in
a state of permanent storage. High dose Pb exposure to animals (30 mg/kg body weight)
was found to induce the formation of a "Pb line" that is visible in both the enamel and
dentin and is localized in areas of recently formed tooth structure. Areas of mineralization
were easily evident in the enamel and the dentin within these "Pb lines." Pb has also been
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shown to decrease cell proliferation, procollagen type I production, intracellular protein,
and osteocalcin in human dental pulp cell cultures. Adult rats exposed to Pb have
exhibited an inhibition of the posteruptive enamel proteinases, delayed teeth eruption
times, as well as a decrease in microhardness of surface enamel. Pb was reported to be
widely dispersed and incorporated into developing apatite crystal during enamel
formation process; however, post formation, Pb was reported to be capable of entering
and concentrating in specific enamel areas which were Ca2+-deficient. The
2006 Pb AQCD (U.S. EPA. 2006b) also reported that a number of animal studies and a
few epidemiologic studies each suggested that Pb is a caries-promoting element. The
strongest epidemiologic evidence comprised associations between concurrent blood Pb
level and dental caries in an NHANES analysis of thousands of children that adjusted for
age, sex, race/ethnicity, poverty to income ratio, exposure to cigarette smoke, geographic
region, head of household education, carbohydrate and calcium intake, and frequency of
dental visits. Other effects found in humans included bone disease (e.g., Paget's disease);
however, the evidence was provided by occupational or case-control studies.
4.9.4.2 Recent lexicological and Epidemiologic Studies
Consistent with evidence reported in the 2006 Pb AQCD, recent studies have found
associations between Pb exposure or biomarker levels and effects in bones of humans and
animals. The association between blood Pb levels and lower bone mineral density was
examined in several epidemiologic studies. Prospective evidence was provided by Khalil
et al. (2008) in 533 older women aged 65-87 years with a mean (SD) blood Pb levels of
5.3 (2.3) (ig/dL. Base line bone mineral density was measured in 1986-1988 (calcaneus),
measured again in 1988-1990 (total hip and femoral neck), and again in 1993-1994
(calcaneus, total hip and femoral neck), while blood Pb levels were measured in between
the second and last bone analyses (during 1990-1991; and categorized as low [n = 122,
< 3 ng/dL], medium [n = 332, 4-7 ng/dL], and high [n = 79, > 8 (ig/dL] Pb blood levels
[range: 1-21 ng/dL]). Information on falls and fractures was collected every 4 months,
starting after the initial enrollment (1986-1988) and continuing for more than 10 years.
The bone mineral density at the last measurement (1993-1994) was 7% lower in the total
hip (p <0.02) and 5% lower in the femoral neck (p <0.03) in the high blood Pb level
group (> 8 (ig/dL) compared to the low blood Pb level group (< 3 (ig/dL). A
concentration-dependent relationship was found for total hip and femoral neck bone
mineral density across the three blood Pb level groups. In addition, total hip, femoral
neck, and calcaneus bone loss was observed to be greater in the medium (blood Pb level:
4-7 (ig/dL) and high blood Pb level groups compared to the low Pb level group, with a
statistically significant trend found for calcaneus bone loss. Compared to the low blood
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Pb level group, women in the high blood Pb level group had an increased risk of
non-spine fracture (10.5 year interview follow-up), and women with medium or high
blood Pb levels had a higher risk of falls (4 year follow-up) with adjustment for age,
clinic, BMI, weight change between visits, smoking, chair stands (were able to stand up
five times from a chair, without using the arms of the chair), fracture history, estrogen
use, and baseline (1986-1988) bone mineral density. Nutritional factors were not
considered. The increased risk of lower bone density and falls leading to osteoporosis-
related fractures associated with blood Pb levels (>4 (ig/dL) are likely influenced by
higher past Pb exposures of these women.
Supporting evidence was provided by cross-sectional epidemiologic studies, although the
direction of the association and the magnitude, timing, frequency, and duration of Pb
exposure that contributed to the observed associations are uncertain. Further, most studies
did not consider potential confounding by nutritional factors. A large NHANES II
analysis of 8,654 adults > 50 years of age (Campbell and Auinger. 2007) was stratified by
non-Hispanic white men (mean blood Pb level: 4.9 [range: 0.7 to 48.1] (ig/dL),
non-Hispanic white women (mean blood Pb level: 3.6 [range: 0.7 to 28.7] (ig/dL),
African-American men (mean blood Pb level: 7.7 [range: 0.7 to 52.9] (ig/dL), and
African-American women (mean blood Pb level: 4.5 [range: 0.7 to 23.3] (ig/dL).
Analyses of covariance considered potential confounding by age, race, sex, BMI,
menopausal status, tobacco use, alcohol use, physical activity, Ca2+ intake, chronic
medical conditions, certain medication use, and SES. Non-Hispanic white men
(n = 1,693, p <0.05) and women (n = 1,754, p <0.10) in the highest tertile of concurrent
blood Pb level had lower mean total hip bone mineral density than non-Hispanic white
men and women in the lowest tertile of blood Pb levels (actual concentrations not
reported). Smaller differences were observed in African-American men and women
(possibly due to the smaller sample sizes (n = 613, and 629, respectively). No association
was observed between blood Pb levels and osteoporotic fractures in either sex or any
race/ethnicity group.
Similar observations were made by Sun et al. (2008a) in 155 males and 37 females in
China who were occupationally-exposed to Pb (mean blood Pb: 20.22 and 15.5 (ig/dL,
respectively). In analyses (including all workers, [plus 36 male and 21 female unexposed
controls] stratified into groups according to blood Pb and urinary Pb levels), groups with
urinary Pb levels > 5 (ig Pb/g of creatinine had lower (p <0.01) bone mineral density
compared to groups with lower urinary Pb in each sex. Prevalence of osteoporosis
increased with increasing blood Pb in a linear manner. In contrast, a significant
association was observed between blood Pb level and bone mineral density, but only in
men with blood Pb levels >30 (ig/dL. Prevalence of osteoporosis increased significantly
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with increasing blood Pb in a linear manner. Results were not adjusted for potential
confounding factors, including other occupational exposures.
Cross-sectional epidemiologic studies also found associations between concurrent blood
Pb level and biological markers of bone turnover. Among 329 male (mean age: 65 years,
median blood Pb level: 2.2 (ig/dL) and 342 female (mean age: 62 years, median blood Pb
level: 1.9 (ig/dL) adults in North Carolina, Nelson et al. (2011) found in women that
higher blood Pb level was associated with biochemical biomarkers of joint tissue
metabolism and turnover, specifically higher urinary N-terminal cross-linked telopeptide
of type I collagen (uNTX-I, a marker of bone resorption/turnover) and urinary C-terminal
cross-linked telopeptide fragments of type II collagen (uCTX-II, a marker associated with
the progression of radiographic knee and hip osteoarthritis) after adjusting for age, BMI,
race, and smoking status. In adjusted analyses of men, higher blood Pb level was
associated with higher uCTX-II, COMP (cartilage oligomeric matrix protein), and type II
collagen cleavage neoepitope / C-propeptide of type II procollagen (C2C:CPII) ratio (an
indication of the balance between cartilage collagen degradation and synthesis). In
women, a weaker association was found for COMP, a cartilage biomarker related to
osteroarthritis. The results indicated that blood Pb level is associated with bone turnover
and mineralized cartilage turnover in women, and with non-mineralized cartilage
turnover in men.
Similarly, Machida et al. (2009) investigated bone matrix turnover in Japanese women
farmers, and how it is related to age-related menopause status and blood Pb level.
Perimenopausal women (n = 319 [age range: 49 to 55 years]) had higher geometric mean
blood Pb level (2.0 (ig/dL) than the other 3 groups did: premenopausal women (n = 261
[age range: 35 to 48 years], blood Pb level: 1.6 (ig/dL), younger postmenopausal women
(n = 397 [age range: 56 to 65 years], blood Pb level: 1.8 (ig/dL), or older postmenopausal
women (n = 248 [age range: 66 to 75 years], blood Pb level: 1.7 (ig/dL). In a model that
simultaneously included bone-mineral density, uNTX-I, osteocalcin, and age, higher
blood Pb levels were positively associated with bone mineral density, uNTX-I, and
osteocalcin (all p <0.01). In perimenopausal women, higher blood Pb level was predicted
most strongly by higher osteocalcin levels. Age was positively associated with higher
blood Pb levels in perimenopausal women only. Associations also were reported for
bone-specific ALP in unadjusted analyses.
To characterize mechanisms underlying the effects of Pb on bone, Jang et al. (2008)
studied the effect of Pb exposure on Ca2+-release activated Ca2+-influx (CRACI) using
cultures of human fetal osteoblast-like (hFOB 1.19) cells (OLCs) in vitro. When cells
were incubated with 1,000 or 3,000 (iM Pb in the culture medium, a concentration-
dependent decrease in CRACI was observed, as was a concentration-dependent increase
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in the influx of Pb into the human OLCs. These results suggest that Pb inhibits the
measurable influx of Ca2+ upon re-addition of Ca2+, which in turn, results in an influx of
Pb into the OLCs.
Studies have found inconsistent associations between higher blood Pb levels and reduced
growth in children; however, Zuscik et al. (2007) hypothesized that Pb may alter growth
by altering chondrogenic commitment of mesenchymal cells and by affecting various
signaling pathways. Exposure of stage El 1.5 murine limb bud mesenchymal cells
(MSCs) to 1 (iM Pb in vitro caused increased basal and TGF-(3/BMP induction of
chondrogenesis, which was accompanied by nodule formation and upregulation of Sox-9,
type 2 collagen, and aggrecan, which are all key markers of chondrogenesis. Enhanced
chondrogenesis during induced ectopic bone formation also was found in mice that had
been pre-exposed to Pb acetate for six weeks via drinking water (55 or 233 ppm,
[previously shown to correspond to 14 or 40 (ig/dL blood Pb level, respectively]). MSCs
exposed to Pb in vitro exhibited an increase in TGF-|3, but BMP-2 signaling was
inhibited. Pb also induced NF-KB and inhibited AP-1 signaling. These results suggested
that the chondrogenesis induced by Pb exposure most likely involved modulation and
integration of multiple signaling pathways, including TGF-(3, BMP, AP-1, and NF-KB.
Effects of Pb exposure on teeth were examined in a few recent cross-sectional
epidemiologic studies. A subset of the U.S. NHANES III (1988-1994) population was
selected for a large periodontitis versus Pb blood level study of both men (n = 2,500) and
women (n = 2,399), 30-55 years-old, that considered potential confounding by a large set
of factors, including nutritional status (Saraiva et al.. 2007). Compared to individuals
with a concurrent blood Pb level of <3 (ig/dL, the prevalence ratios of periodontitis were
1.70 (95% CI: 1.02, 2.85) for men with concurrent blood Pb level of >7 (ig/dL and 3.80
(95% CI: 1.66, 8.73) for women with concurrent blood Pb level >7 (ig/dL. These results
were adjusted for age, NHANES III phase, cotinine levels, poverty to income ratio,
race/ethnicity, education, bone mineral density, diabetes, Ca2+ intake, dental visits, and
menopause status in women.
Arora et al. (2009) examined the association between blood and bone Pb level and the
loss of natural teeth, in 333 men (age range: 50 to 94 years) from a subset of the Veterans
Affairs Normative Aging Study (NAS). Tooth loss was ascertained as the number of teeth
present during a dental assessment, and was categorized into three groups: 0 missing
teeth (n = 44), 1-8 missing teeth (n = 164), or > 9 missing teeth (n = 125). Men with > 9
teeth missing had significantly higher tibia and patella Pb concentrations (measured
within 3 years of dental assessment) compared to those with no tooth loss. Men with the
highest tibia Pb concentrations (>23 (ig/g) had higher odds of tooth loss (OR: 3.03 [95%
CI: 1.60, 5.75]) compared to men with tibia Pb levels < 15 (ig/g. Men with the highest
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patella Pb levels (>36 (ig/g) also had higher odds of tooth loss (> 9 missing teeth versus
0-8 missing teeth; or > 1 missing teeth versus 0 missing teeth: OR: 2.41 [95% CI: 1.30,
4.49]) compared to men with patella Pb levels < 22.0 (ig/g. Men with tibia Pb levels
16-23 (ig/g, and men with patella Pb levels 23-36 (ig/g also had elevated odds of tooth
loss. Results were adjusted for age, education, smoking status, pack-years of smoking,
and diabetes, but nutritional factors were not considered. Tooth loss was not associated
with higher blood Pb levels which were measured within 3 years of dental assessment
[see Hu et al. Q996b) for details of the blood Pb analysis]. This indicates that long-term
cumulative exposure to Pb is associated with increased odds of tooth loss.-However,
because the timing of tooth loss was not ascertained, and concurrent bone Pb levels may
represent exposures after tooth loss occurred, the directionality of effects is uncertain.
4.9.4.3 Summary of Effects on Bone and Teeth
A few studies have indicated associations of Pb exposure or Pb biomarker levels with
bone disease (e.g., Paget's disease); however, the implications are limited by examination
of Pb-exposed workers or recruitment of individuals based on bone disease status
(i.e., case-control). Numerous epidemiologic studies indicated an association between
higher Pb biomarker levels and lower bone density in adults. Prospective evidence was
provided by a study of elderly women (65-87 years-old), in which higher blood Pb levels
(> 4 (ig/dL vs. < 3 (ig/dL) were associated with lower bone density measured after 2-4
years and greater risk of falls and osteoporosis-related fractures (Khalil et al.. 2008).
Cross-sectional epidemiologic associations between higher blood Pb levels and lower
bone mineral density were found in adults without (Campbell and Auinger. 2007) and
with occupational Pb exposure (Sun et al.. 2008a). Cross-sectional studies also indicated
associations between higher blood Pb levels and biochemical biomarkers of higher bone
turnover in elderly populations (Nelson etal.. 2011; Machida et al.. 2009). In the cross-
sectional epidemiologic evidence, it is difficult to determine whether an increase in blood
Pb level results from lower bone density or from higher bone turnover, and whether
either of these effects lead to a greater release of Pb from bone into the bloodstream.
Except for Sun et al. (2008a), studies adjusted for several potentially important
confounding factors, including age, BMI, and smoking. However, studies did not
consider nutritional status, which could affect the release of Pb from bone to blood. To
support the direction and independent effects of Pb on bone development, toxicological
studies have found Pb-induced (gestational and postnatal) decreases in bone growth in
juvenile animals. Further, these toxicological studies have characterized potential modes
of action, by showing Pb-induced decreases in bone mineralization and bone cell
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differentiation, inhibition of CRACI, and alterations in signaling pathways involved in
skeletal development (Jang et al.. 2008; Zuscik et al.. 2007).
Epidemiologic studies have found associations between blood Pb levels and effects on
teeth. Large NHANES analyses adjusted for several potentially important confounding
factors including age, SES-related factors, and nutritional factors, and found associations
between concurrent blood Pb level and dental caries in children (Moss et al.. 1999) and
periodontitis in adults (Saraiva et al.. 2007). Higher patella and tibia Pb levels were
associated with tooth loss in NAS men (Arora et al., 2009). The results for blood Pb and
bone Pb levels in adults indicate that long-term, cumulative exposure to Pb exposure is
associated with negative effects on teeth. This epidemiologic evidence was based on
cross-sectional study design analyses, which precludes conclusions about the
directionality of effects. However, these findings are supported by toxicological evidence
in animals for Pb-induced increases in Pb uptake into teeth; and decreases in cell
proliferation, procollagen type I production, intracellular protein, and osteocalcin in cells
exposed to Pb in vitro. Despite evidence for associations between Pb exposure and effects
in bone and teeth at relatively low concurrent blood Pb levels, these outcomes were most
often examined in older adults that have been exposed to higher levels of Pb earlier in
life. Therefore, uncertainty still remains concerning the Pb exposure level, timing,
frequency, and duration that contribute to the observed associations.
However, the small body of epidemiologic evidence showing associations between Pb
biomarker levels and various bone and teeth effects after adjusting for potential
confounding by age, SES-related factors, and nutritional factors, plus the supporting
toxicological evidence, is sufficient to conclude that a causal relationship is likely to exist
between Pb exposure and effects on bone and teeth.
4.9.5 Effects on Ocular Health
Ocular effects most commonly associated with exposure to Pb include formation of
cataracts, impaired vision, retinal edema, and retinal stippling.
4.9.5.1 Summary of Key Findings of the Effects on Ocular Health
from the 2006 Pb AQCD
The 2006 Pb AQCD stated that various changes in the visual system were observed with
Pb poisoning including retinal stippling and edema, cataracts, ocular muscle paralysis and
impaired vision. Maternal prenatal blood Pb levels in the range of 10.5 to 32.5 ug/dL
during the first trimester were associated with supernormal retinal electroretinograms
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(ERGs) in children at age 7-10 years (Section 4.3.6.2). Cataracts were noted in middle-
aged men with tibia bone Pb levels of 31-126 ug/g.
4.9.5.2 Recent lexicological and Epidemiologic Studies
The recent cross-sectional epidemiologic studies of ocular effects in adults did not
produce clear evidence, and each was limited by the lack of rigorous statistical analysis
and lack of consideration for potential confounding. Erie et al. (2009) measured Pb and
Cd in retinal tissue from 25 eye donors with age-related macular degeneration (cases) and
36 normal control donors. Pb, but not Cd, concentration was significantly elevated in the
neural retina tissue of the 25 donors with macular degeneration (50 eyes; median [IQR]:
12.0 [8-18] ng/g Pb) versus normal control donors (72 eyes; median [IQR]: 8.0 [0-11]
ng/g Pb). Neither of these heavy metals was significantly elevated in the retinal pigment
epithelium (RPE)/choroid complex in donors with macular degeneration over normal
controls. Mosad et al. (2010) compared Pb, Cd, vitamin C, vitamin E, and beta carotene
blood levels between 45 middle-aged male smokers and nonsmokers with cataracts.
Blood Pb levels were elevated (p <0.0001) in 15 light (mean [SD]: 14.5 [0.41] (ig/dL), 15
moderate (14.5 [0.41] (ig/dL), and 15 heavy smokers (18.7 [1.24] (ig/dL) compared to 15
nonsmokers (12.2 [0.21] (ig/dL). Similar associations were observed for Cd blood levels
and lens concentrations. There was no direct analysis of the association between Pb blood
level or lens Pb concentration and the severity of cataracts.
Recent animal studies have observed Pb-induced retinal progenitor cell proliferation and
neurogenesis (Section 4.3.6.2). An in vitro study found increased opacity of rat (age
4-6 weeks) lens exposed to 1 (iM Pb nitrate with or without secondary oxidative
challenge after 5-8 days but not after 3 days (Neal etal. 2010b). Thus, short-term Pb
exposure did not induce osmotic swelling or lens shrinkage. Following a 5-day exposure,
approximately 30% of the Pb-exposed lenses displayed "definite cataracts" compared to
only 5% of control lenses. By culture day 8, 100% of the Pb-exposed lenses were
described either as having point opacities or definite cataracts, while only approximately
14% of control lenses displayed these characteristics, indicating that prolonged exposure
of lenses to Pb induced an accelerated formation of opacity/cataract compared to
unexposed lenses. Pb-exposed lenses cleared the media of hydrogen peroxide more
rapidly than did control lenses, potentially due to increased CAT activity. Exposure to
hydrogen peroxide resulted in total (100%) opacity in Pb-exposed lenses at culture day 7,
compared -20% in control cells. Exposure to Pb additionally altered epithelial nutrient
transport and lens histology, relative to that in controls.
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In summary, a prospective epidemiologic study indicated associations between prenatal
maternal blood Pb levels of 10.5-32.5 (ig/dL during the first trimester and supernormal
retinal ERGs in children ages 7-10 years [Section 4.3.6.2. Rothenberg et al. (2002b)].
However, the biological relevance of supernormal ERGs is uncertain. Evidence in adults
for associations between eye tissue Pb levels and macular degeneration (Mosad et al..
2010) and cataracts in adults (Erie et al.. 2009) is limited by weak statistical methods and
lack of consideration for potential confounding to warrant conclusions. Toxicology
studies have reported Pb-induced retinal progenitor cell proliferation, retinal ERGs, and
lens opacity (Section 4.3.6.2). Because of the insufficient quantity and quality of studies
in the cumulative body of evidence, the evidence is inadequate to determine a causal
relationship between Pb exposure and ocular effects.
4.9.6 Effects on the Respiratory System
Blood Pb level has been associated with asthma and allergy in children in prospective and
cross-sectional epidemiologic studies (Section 4.6.5.2). As described in Section 4.2.4.
Pb exposure has been shown to induce the generation of ROS. ROS are implicated in
mediating increases in bronchial responsiveness and activating neural reflexes, leading to
decrements in lung function. Collectively, studies investigating these airway responses in
asthma-free populations are limited in number, lack rigorous statistical analysis, and
collectively do not provide strong evidence of an association with blood Pb level
(Section 4.6.5.3). Collectively, panel and time-series epidemiologic studies demonstrate
associations between short-term increases in ambient air Pb (measured in PM2 5 or PMi0
air samples), and decreases in lung function and increases in respiratory symptoms, and
asthma hospitalizations in children but not adults (Section 4.6.5.3). Toxicological studies
have found pulmonary inflammation induced by concentrated ambient air particles
(CAPs) in which Pb was one of the numerous components (Wei et al.. 2011; Duvall et al..
2008; Godleski et al.. 2002; Saldiva et al.. 2002). Despite this evidence for respiratory
effects related to air-Pb concentrations, the limitations of air-Pb studies, including the
limited data on the size distribution of Pb-PM (Section 2.5.3). the uncertain relationships
of Pb-PMio and Pb-PM25 with blood Pb levels, and the lack of adjustment for other
correlated PM chemical components preclude firm conclusions about air Pb-associated
respiratory effects. Because of the insufficient quantity and quality of studies in the
cumulative body of evidence, the evidence is inadequate to determine a causal
relationship between Pb exposure and respiratory effects in populations without asthma.
4-738
-------
4.10 Cancer
Previous Pb AQCDs have demonstrated that Pb is a well-established animal carcinogen.
Oral Pb acetate exposure to male and female rodents has consistently been shown to be a
kidney carcinogen in multiple separate studies, inducing adenocarcinomas and adenomas
after chronic exposure. Developmental Pb acetate exposure also induced kidney tumors
in offspring whose dams received Pb acetate in drinking water during pregnancy and
lactation. Gliomas of the brain have also been reported after oral Pb exposure. These
rodent toxicological studies have been conducted at high doses of Pb and have shown that
Pb is an animal carcinogen. Because of this strong body of historical data, the
2006 Pb AQCD, in describing the evidence reviewed therein, states, "limited
tumorigenesis studies have been conducted in animal models and the focus has been
more on the mechanism of neoplasia...and possible immunomodulatory effects of Pb in
the promotion of cancer." More recent studies have focused on co-administration of Pb
with known carcinogens or modifiers such that lifestage, diet, and mechanism of action
can be better understood.
The previous epidemiologic studies included in the 2006 Pb AQCD (U.S. EPA. 2006b)
"provide [d] only very limited evidence suggestive of Pb exposure associations with
carcinogenic or genotoxic effects in humans," and the studies were summarized as
follows:
"The epidemiologic data ... suggest a relationship between Pb exposure and cancers of the
lung and the stomach... Studies of genotoxicity consistently link Pb-exposed populations
with DNA damage and micronuclei formation, although less consistently with
chromosomal aberrations."
The International Agency for Research on Cancer (IARC) classified inorganic Pb
compounds as probable human carcinogens (Group 2A of IARC classification) based on
sufficient evidence in animal studies (evidence in human studies was limited), and
organic Pb compounds as not classifiable (Group 3 of IARC classifications) (IARC,
2006a: Rousseau et al.. 2005). Additionally, the National Toxicology Program (NTP)
listed Pb and Pb compounds as "reasonably anticipated to be human carcinogens" (NTP.
2011). The typical cancer bioassays employed by IARC or NTP as evidence of
Pb-induced carcinogenicity used rodents that were continuously exposed to Pb acetate in
chow or drinking water for 18 months to two years in duration. These two year cancer
bioassays and the doses administered are typical of cancer bioassays used with other
chemicals.
In the following sections, recent epidemiologic and toxicological studies, published since
the 2006 Pb AQCD, regarding Pb and cancer mortality and incidence are examined. In
addition, recent studies of Pb exposure associated with DNA and cellular damage, as well
4-739
-------
as epigenetic effects, are summarized. When the information is available, the form of the
Pb compound under study (e.g., inorganic, organic) is indicated. In epidemiologic
studies, various biological indicators of Pb exposure are used including Pb measured in
blood and bone. The biological indicators of Pb associated with cancer-related endpoints
are considered in drawing conclusions about potentially important levels and timing of Pb
exposure. Bone Pb is indicative of cumulative Pb exposure. Blood Pb can represent more
recent exposure, but because it can also represent remobilized Pb occurring during times
of bone remodeling, blood Pb level may also be an indicator of long-term Pb exposure in
adults. More detailed discussion of these measures is given in Section 3.3.5.
Toxicological studies report exposure or administered dose. Details of the recent
epidemiologic and toxicological studies follow.
4.10.1 Cancer Incidence and Mortality
Recent studies have included epidemiologic evaluations of the associations between Pb
exposure and both specific cancers (such as lung cancer and brain cancer), and overall
cancer (cancer of any type). Table 4-49 provides an overview of the study characteristics
and results for the epidemiologic studies that reported effect estimates. The studies in the
text and table are grouped according to study design and methodological strength. This
section also evaluates toxicological evidence on the potential carcinogenicity of Pb.
4-740
-------
Table 4-49 Summary of recent epidemiologic
Reference
(In order of
appearance in Study Cancer
text) Location Outcome
Study
Population
studies of cancer incidence and overall cancer mortality.
Measure of Pb
Exposure/
Methodological Concentration
Details Data Adjusted Effect Estimates
Potential
confounders
adjusted for in
analysis
Overall Cancer Mortality:
Menke et al. Multiple U.S. Overall
(2006) locations cancer
mortality
NHANES III
cohort with
Blood Pb
measures in
1988-1994
Follow-up: At
least 12 years
Blood Pb
<10ug/dL
N=13,946
N for cancer
mortality = 411
Cohort study Blood Pb at HR (95% Cl):
using Cox baseline
regression and Ge0metric mean TertNe 1 . 1 00
other 2 58 uq/dl_
techniques _ M9 Tertile 2: 0.72 (0.46, 1.12)
.cig's'ug/dL Tertile 3: 1.10 (0.82, 1.47)
Tertile 2:
1. 94-3.62 ug/dl_
Tertile 3:
> 3.63 ug/dL
Age,
race/ethnicity,
sex, diabetes
mellitus, body
mass index,
current or
former smoking,
alcohol
consumption,
physical activity,
low income,
CRP, total
cholesterol, high
school
education,
urban
residence,
postmenopausal
status,
hypertension,
and level of
kidney function
4-741
-------
Table 4-49 (Continued): Summary of recent epidemiologic studies of cancer incidence and overall cancer mortality.
Reference
(In order of
appearance in
text)
Study
Location
Cancer
Outcome
Study
Population
Methodological
Details
Measure of Pb
Exposure/
Concentration
Data
Adjusted Effect Estimates
Potential
confounders
adjusted for in
analysis
Schober et al.
(2006)
Multiple U.S.
locations
Overall
cancer
mortality
NHANES III
cohort
At least 40
years of age
Median follow-
up period for
this study: 8.6
yr
N= 9,757
N for cancer
mortality = 543
Cohort study
using Cox
proportional
hazard
regression
analysis and
other
techniques
Blood Pb at
baseline
Blood Pb<5 ug/dL:
67.7%
Blood Pb 5-9 ug/dl_
: 26.0%
Blood Pb>10ug/dL:
6.3%
RR (95% Cl):
Blood Pb<5ug/dL: 1.00
Blood Pb 5-9 ug/dl_: 1.44(1.12,
1.86)
Blood Pb>10ug/dL: 1.69(1.14,
2.52)
Note: Modification by age
assessed and associations
varied slightly
Sex,
race/ethnicity,
education, and
smoking status
Age used as
time-scale in
models
Additional
covariates
considered but
not included:
Census region
and urban
status of
residence,
alcohol intake
4-742
-------
Table 4-49 (Continued): Summary of recent epidemiologic studies of cancer incidence and overall cancer mortality.
Reference
(In order of
appearance in Study Cancer
text) Location Outcome
Weisskopf et al. Boston, MA Overall
(2009) area cancer
mortality
Khalil et al. Baltimore, Overall
(2009b) MD, and cancer
Monongahela mortality
Valley, PA
Study
Population
Normative
Aging Study
(MAS)
Included men
only (mostly
white)
Mean follow-
up period for
this study: 8.9
yr
Blood Pb
measures
available
N=1038
N for cancer
mortality=85
Bone Pb
measures
available
N=727
N for cancer
mortality=57
Subgroup of
the Study of
Osteoporotic
Fractures
cohort
Included white
women aged
65-87;
Follow-up: 12
yr (+/- 3 yr)
N=533
N for cancer
mortality=38
Methodological
Details
Cohort study
using
Cox
proportional
hazards
Cohort study
using
Cox
proportional
hazards
regression
analysis and
other
techniques
Measure of Pb
Exposure/
Concentration
Data
Blood Pb at
baseline (ug/dL)
Mean (SD): 5.6
(3.4)
Tertile 1 of Blood
Pb: <4
Tertile 2 of Blood
Pb: 4-6
Tertile 3 of Blood
Pb: >6
Patella Pb at
baseline (ug/g)
Tertile 1 of patella
Pb: <22
Tertile 2 of patella
Pb: 22-35
Tertile 3 of patella
Pb: >35
Blood Pb at
baseline
Mean (SD) 5.3 (2.3)
ug/dl_
Adjusted Effect Estimates
HR (95% Cl):
Blood Pb Tertile 1: 1.00
Blood Pb Tertile 2: 1.03(0.42,
2CC\
.DO)
Blood Pb Tertile 3: 0.53(0.20,
1.39)
Patella Pb Tertile 1: 1.00
Patella Pb Tertile 2: 0.82 (0.26,
2.59)
Patella Pb Tertile 3: 0.32(0.08,
1.35)
HR (95% Cl):
Blood Pb<8ug/dL: 1.00
Blood Pb> 8 ug/dl_: 1.64(0.73,
3.71)
Potential
confounders
adjusted for in
analysis
Age, smoking,
and education
Additional
covariates
considered but
not included:
alcohol intake,
physical activity,
body mass
index, total
cholesterol,
serum high-
density
lipoprotein,
diabetes
mellitus, race,
and
hypertension
Age, clinic, body
mass index,
education,
smoking,
alcohol intake,
estrogen use,
hypertension,
walking for
exercise,
diabetes, and
total hip bone
mineral density
4-743
-------
Table 4-49 (Continued): Summary of recent epidemiologic studies of cancer incidence and overall cancer mortality.
Reference
(In order of
appearance in
text)
Overall Cancer
Absalon and
Slesak (2010)
Obhodas et al.
(2007)
Mendy et al.
(2012)
Study
Location
Incidence:
Silesia
province,
Poland
Island of Krk,
Croatia
Multiple U.S.
locations
Cancer
Outcome
Overall
cancer
incidence
Incidence
rates for
neoplasms
Incidence of
cancer or
"malignancy
of any kind"
Study
Population
Children living
in this province
at least five
years
N = not
specified
Individuals
living in the
Island of Krk
from
1997-2001
N= 1,940
2007-2008
NHANES
cohort - at
least 20 years
of age
N= 1,857
Methodological
Details
Ecologic
analysis using
correlations.
Cross-sectional
study using
correlations and
linear
regression
Cross-sectional
study using
logistic
regression
Measure of Pb
Exposure/
Concentration
Data
Pb-related air
pollution measures
Concentrations: NA
Soil and vegetation
samples,
household potable
water samples,
children's hair
samples
Concentrations: NA
Concurrently
measured
creatinine-adjusted
urinary Pb
Geometric mean for
creatinine-adjusted
urinary Pb marker:
0.59 ug/g
(95% Cl: 0.57,
OC *1 \
.61)
Adjusted Effect Estimates
Reported correlations between
changes in Pb and cancer
incidence - no/low correlations
observed (correlation
coefficients between -0.3 and
0.2)
No association observed
between Pb in the samples and
incidence of neoplasm
(numerical results not provided)
OR (95% Cl): Greater than log-
transformed mean creatinine-
adjusted urinary Pb level
compared to less than log-
transformed mean creatinine-
adjusted urinary Pb level: 0.76
(0.44, 1.33)
Potential
confounders
adjusted for in
analysis
None specified
Examined
correlation by
sex, no
difference
reported
None specified
Age, sex,
race/ethnicity,
education level,
ratio family
income to
poverty, alcohol
consumption,
cigarette
smoking, and
other heavy
metals
4-744
-------
Table 4-49 (Continued): Summary of recent epidemiologic studies of cancer incidence and overall cancer mortality.
Reference
(In order of
appearance in
text)
Study
Location
Cancer
Outcome
Study
Population
Methodological
Details
Measure of Pb
Exposure/
Concentration
Data
Potential
confounders
adjusted
Adjusted Effect Estimates analysis
for in
Lung Cancer:
Lundstrom et al. Sweden
(2006)
Lung
cancer
(incidence
and
mortality)
Male Pb
smelter
workers first
employed for
> 3 months
between 1928
and 1979
Followed up
for mortality
from 1955
-1987
N=187
N lung
cancer=46
Nested case-
referent study
using
conditional
logistic
regression
Median peak blood
Pb level
Cases: 49.7 ug/dL
Controls:
55.9 ug/dl_
Median number of
years with at least
one blood sample
obtained
Cases: 4.5 yr,
Controls: 6.0 yr
OR (95% Cl):
Median peak blood Pb Level:
1.00(0.71, 1.42) per 10 ug/dL
Median number of years with at
least one blood sample
obtained: 0.96 (0.91, 1.02) per
year
Median cumulative Blood Pb
index: 1.00 (0.98, 1.01) per
10 ug/dL
Matched by age
Adjusted for
smoking and As
exposure
Median cumulative
blood Pb index
(sum of annual
blood Pb Level)
Cases: 186 ug/dL
Controls: 246 ug/dL
Note: similar results were
observed when restricted to
smokers only
4-745
-------
Table 4-49 (Continued): Summary of recent epidemiologic studies of cancer incidence and overall cancer mortality.
Reference
(In order of
appearance in
text)
Jones et al.
(2QQ7)
Study
Location
Humberside,
U.K.
Cancer
Outcome
Lung
cancer
mortality
Study
Population
Male tin
smelter
employees
1X1=1, 462
Methodological
Details
Cohort study
using Poisson
regression
Measure of Pb
Exposure/
Concentration
Data
Personnel record
cards and air
sampling
conducted from
1972-1991
Adjusted Effect Estimates
RR (90% Cl) for Pb exposure
weighted by age and time since
exposure:
1.54(1.14,2.08)
Potential
confounders
adjusted for in
analysis
Not specified
Three exposure
scenarios
determined for
working lifetime
cumulative
exposure - all have
similar medians of
approximately
2 mg/m3 per yr
Mean
-2.0 mg/m3 per yr
Note: Similar results for other
exposure determination
scenarios.
4-746
-------
Table 4-49 (Continued): Summary of recent epidemiologic studies of cancer incidence and overall cancer mortality.
Reference
(In order of
appearance in Study
text) Location
Rousseau et al. Montreal,
(2007) Canada
Cancer Study
Outcome Population
Lung Men aged
cancer 35-79
incidence N population
based
controls=
ranged from
271 to 471
depending on
exposure of
interest
N controls with
other cancers=
ranged from
737 to 1203
depending on
exposure of
interest
N lung
cancer=ranged
from 433 to
751 depending
on exposure of
interest
Methodological
Details
Population-
based case-
control study
using
unconditional
logistic
regression
Measure of Pb
Exposure/
Concentration
Data
Interview of job
history and
exposure matrix
Ever exposed to:
Organic Pb 3.0%
Inorganic Pb 17.0%
Pb in gasoline
emissions 38.6%
Adjusted Effect Estimates
OR (95% Cl) for any versus no
exposure
Organic Pb exposure:
1.3(0.5, 3.1)
Inorganic Pb exposure:
1.1 (0.7, 1.7)
Pb in gasoline emissions
exposure:
0.8(0.6, 1.1)
Note: results are for
comparisons using population-
based controls; results for
controls with other types of
cancers were similar
Potential
confounders
adjusted for in
analysis
Age, family
income, cultural
origin, proxy
status, ever
exposure to
asbestos, silica,
As, Cd, and
chromium (VI)
4-747
-------
Table 4-49 (Continued): Summary of recent epidemiologic studies of cancer incidence and overall cancer mortality.
Reference
(In order of
appearance in
text)
Study
Location
Cancer
Outcome
Study
Population
Methodological
Details
Measure of Pb
Exposure/
Concentration
Data
Potential
confounders
adjusted
Adjusted Effect Estimates analysis
for in
Brain Cancer:
van Wijngaarden Multiple U.S Brain
and Dosemeci locations cancer
(2006) mortality
National
Longitudinal
Mortality Study
Included
individuals
with
occupational
information
Included
follow-up from
1970-1989
N= 317,968
Cohort study
using
proportional
hazards,
Poisson
regression
techniques, and
standardized
mortality ratios
(SMR)
Interview about
current or most
recent job within the
past 5 years and a
job exposure matrix
Concentrations: NA
HR (95% Cl)
For any Pb exposure compared
to no exposure: 1.56 (1.00,
2.43)
RR (95% Cl) from Poisson
regression of any Pb exposure
compared to no exposure: 1.42
/n 01 o on\
^u.y i , Ł..Ł.\j )
SMR (95% Cl):
Not exposed: 0.87(0.70, 1.06)
Any exposure: 1.11 (0.74, 1.59)
Note: Effect estimates were
greatest among those with high
probabilities of exposure and
medium/high exposure intensity
Sex, age, race,
living in an
urban area,
marital status
and educational
level
Additional
covariate
considered but
not included:
Family income
(not used due to
large % missing;
additional
analysis
including it gave
similar results)
4-748
-------
Table 4-49 (Continued): Summary of recent epidemiologic studies of cancer incidence and overall cancer mortality.
Reference
(In order of
appearance in Study Cancer
text) Location Outcome
Rajaraman et al. Phoenix, AZ, Brain
(2006) Boston, MA, cancer
and incidence
Pittsburgh,
PA
Study
Population
NCI Brain
Tumor Study
Included
individuals >18
yr diagnosed
with brain
cancer less
than 8 week
before
hospitalization;
frequency-
matched
controls were
individuals
admitted to the
same hospitals
for non-
neoplastic
conditions
N controls
=799
N glioma
= 489
N meningioma
=197
Measure of Pb
Exposure/
Methodological Concentration
Details Data Adjusted Effect Estimates
Case-control Interviews of OR (95% Cl) for ever versus no
study using lifetime work history exposure
unconditional and exposure Meningioma-
logistic databases 0 8 (0 5 13)
regression Concentrations: NA Glioma:'
0.8(0.6, 1.1)
Note: positive associations
between Pb exposure and
meningioma incidence was
observed among individuals
with ALAD2 genotypes, but not
individuals with ALAD1
homozygotes genotypes; these
associations were not observed
for glioma incidence
Potential
confounders
adjusted for in
analysis
Age, sex, race /
ethnicity,
hospital, and
residential
proximity to
hospital
4-749
-------
Table 4-49 (Continued): Summary of recent epidemiologic studies of cancer incidence and overall cancer mortality.
Reference
(In order of
appearance in
text)
Bhatti et al.
(2009)
Study Cancer Study
Location Outcome Population
Phoenix, AZ, Brain NCI Brain
Boston, MA, cancer Tumor Study
and incidence included non-
Pittsburgh, Hispanic
PA
whites > 18 yr
diagnosed with
brain cancer
less than
8 week before
hospitalization;
frequency-
matched
controls were
individuals
admitted to the
same hospitals
for non-
neoplastic
conditions
N controls
=494
N glioma
= 362
N meningioma
=134
Measure of Pb
Exposure/
Methodological Concentration
Details Data
Case-control Interviews of
study using lifetime work history
unconditional and exposure
logistic databases
regression Mean (SD)
Glioma:
70.5 ug/m3 per yr
(193.8 ug/m3 peryr)
Glioblastoma
multiform:
97.5 ug/m3 per yr
(233.9 ug/m3 peryr)
Meningioma:
101.1 ug/m3 peryr
(408.7 ug/m3 peryr)
Controls:
69.7 ug/m3 per yr
(248.8 ug/m3 peryr)
Adjusted Effect Estimates
OR (95% Cl) per
1 00 ug/m3 per yr increase in
cumulative Pb exposure
Glioma: 1.0(0.9, 1.1)
Glioblastoma multiform: 1.0
(0.9, 1.1)
Meningioma: 1.1 (1.0, 1.2)
Note: modification by SNPs was
analyzed and associations
varied by SNP
Potential
confounders
adjusted for in
analysis
Age, sex,
hospital, and
residential
proximity to the
hospital
4-750
-------
Table 4-49 (Continued): Summary of recent epidemiologic studies of cancer incidence and overall cancer mortality.
Reference
(In order of
appearance in Study
text) Location
Cancer Study
Outcome Population
Measure of Pb
Exposure/
Methodological Concentration
Details Data
Adjusted Effect Estimates
Potential
confounders
adjusted for in
analysis
Breast Cancer:
Pan et al. (2011) Canada
Breast National
cancer Enhanced
incidence Cancer
Surveillance
System
(NECSS).
Population-
based sample
of cancer
cases and
controls with
information
collected from
1994-1997
N controls
=2,467
N cases=
2Oy1O
,0^0
Population- Self-reported
based case- previous addresses
control study and their proximity
using to Pb smelters
unconditional (determined using
logistic Environmental
regression Quality Database
[EQDB])
Concentration: NA
OR(95%CI):
Residing >3.2 km from Pb
smelter or no nearby smelter:
1.00
Residing 0.8-3.2 km from Pb
smelter:
0.41 (0.11, 1.51)
Residing <0.8 km from Pb
smelter'
0.61 (0.11, 3.42)
Age, province of
residence,
education,
smoking pack
years, alcohol
consumption,
body mass
index,
recreational
physical activity,
number of live
births, age at
menarche,
menopausal
status, total
energy intake,
and
employment in
the industry
under
consideration
Additional
covariate
considered but
not included:
Family income
4-751
-------
Table 4-49 (Continued): Summary of recent epidemiologic studies of cancer incidence and overall cancer mortality.
Reference
(In order of
appearance in Study
text) Location
Cancer Study
Outcome Population
Methodological
Details
Measure of Pb
Exposure/
Concentration
Data
Adjusted Effect Estimates
Potential
confounders
adjusted for in
analysis
Other Cancers:
Rousseau et al. Montreal,
(2007) Canada
Various Men aged
cancer 35-79 yr
incidences N population
based
controls=
ranged from
271 to 471
depending on
the cancer and
exposure of
interest
N controls with
other cancers=
ranged from
697 to 2,250
depending on
exposure of
interest
N cancer
=ranged from
60 to 442
depending on
the cancer and
exposure of
Population-
based case-
control study
using
unconditional
logistic
regression
Interview of job
history and
exposure matrix
Ever exposed to:
o ™/n'C
Inorganic Pb
17.0%
Pb in gasoline
emissions
38.6%
OR (95% Cl):
Never exposed is referent group
Organic Pb:
Esophageal 1.7(0.5, 6.4)
Stomach 3.0 (1.2, 7.3)
Colon 1.5(0.7, 3.6)
Rectum 3. 0(1. 2, 7.5)
Pancreas 0.9 (0.1, 5.2)
Prostate 1.9(0.8, 4.6)
Bladder 1.7 (0.7, 4.2)
Kidney 2.3 (0.8, 6.7)
Non-Hodgkin's lymphoma
0.4(0.1,2.2)
Note: results are for
comparisons using population-
based controls; results for
controls with other types of
ranrarc \A/ara cimilar avrant nn
Age, family
income, cultural
origin or
birthplace, and
proxy status; all
models except
those for
melanoma and
non-Hodgkin's
lymphoma were
adjusted for
smoking
association was present
between organic Pb and rectal
cancer
4-752
-------
Table 4-49 (Continued): Summary of recent epidemiologic studies of cancer incidence and overall cancer mortality.
Reference
(In order of
appearance in Study
text) Location
Rousseau et al. Montreal,
(2007) Continued Canada
Cancer Study
Outcome Population
Various Men aged
cancer 35-79 yr
incidences N population
based
controls=
ranged from
271 to 471
depending on
the cancer and
exposure of
interest
N controls with
other cancers=
ranged from
697 to 2,250
depending on
exposure of
interest
N cancer
=ranged from
60 to 442
depending on
the cancer and
exposure of
interest
Methodological
Details
Population-
based case-
control study
using
unconditional
logistic
regression
Measure of Pb
Exposure/
Concentration
Data
Interview of job
history and
exposure matrix
Ever exposed to:
Organic Pb
3.0%
Inorganic Pb
17.0%
Pb in gasoline
emissions
38.6%
Adjusted Effect Estimates
OR (95% Cl):
Never exposed is referent group
Inorganic Pb:
Esophageal 0.6 (0.3, 1.2)
Stomach 0.9 (0.6, 1.5)
Colon 0.8 (0.5, 1.1)
Rectum 0.8 (0.5, 1.3)
Pancreas 0.9 (0.4, 1.8)
Prostate 1.1 (0.7, 1.6)
Bladder 1.1 (0.7, 1.5)
Kidney 1.0(0.6, 1.7)
Melanoma 0.4 (0.2, 1.0)
Non-Hodgkin's lymphoma
0.7(0.4, 1.2)
Note: results are for
comparisons using population-
based controls; results for
controls with other types of
Potential
confounders
adjusted for in
analysis
Age, family
income, cultural
origin or
birthplace, and
proxy status; all
models except
those for
melanoma and
non-Hodgkin's
lymphoma were
adjusted for
smoking
cancers were similar except no
association was present
between organic Pb and rectal
cancer
4-753
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Table 4-49 (Continued): Summary of recent epidemiologic studies of cancer incidence and overall cancer mortality.
Reference
(In order of
appearance in Study
text) Location
Rousseau et al. Montreal,
(2007) Continued Canada
Cancer Study
Outcome Population
Various Men aged
cancer 35-79 yr
incidences N population
based
controls=
ranged from
271 to 471
depending on
the cancer and
exposure of
interest
N controls with
other cancers=
ranged from
697 to 2,250
depending on
exposure of
interest
N cancer
=ranged from
60 to 442
depending on
the cancer and
exposure of
interest
Methodological
Details
Population-
based case-
control study
using
unconditional
logistic
regression
Measure of Pb
Exposure/
Concentration
Data
nterviewof job
history and
exposure matrix
Ever exposed to:
Organic Pb
3.0%
Inorganic Pb
17.0%
Pb in gasoline
emissions
38.6%
Adjusted Effect Estimates
OR (95% Cl):
Never exposed is referent group
Pb in gasoline emissions:
Esophageal 0.6 (0.4, 1.1)
Stomach 1.0(0.7, 1.4)
Colon 0.8 (0.6, 1.1)
Rectum 1.0 (0.7, 1.4)
Pancreas 0.9 (0.5, 1.4)
Prostate 0.9 (0.7, 1.2)
Bladder 0.8 (0.6, 1.1)
Kidney 1.0(0.7, 1.5)
Melanoma 0.8 (0.5, 1.4)
Non-Hodgkin's lymphoma
0.7(0.5, 1.0)
Note: results are for
comparisons using population-
based controls; results for
controls with other types of
Potential
confounders
adjusted for in
analysis
Age, family
income, cultural
origin or
birthplace, and
proxy status; all
models except
those for
melanoma and
non-Hodgkin's
lymphoma were
adjusted for
smoking
cancers were similar except no
association was present
between organic Pb and rectal
cancer
4-754
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Table 4-49 (Continued): Summary of recent epidemiologic studies of cancer incidence and overall cancer mortality.
Reference
(In order of
appearance in Study
text) Location
Rousseau et al. Montreal,
(2007) Continued Canada
Cancer Study
Outcome Population
Various Men aged
cancer 35-79 yr
incidences N population
based
controls=
ranged from
271 to 471
depending on
the cancer and
exposure of
interest
N controls with
other cancers=
ranged from
697 to 2,250
depending on
exposure of
interest
N cancer
=ranged from
60 to 442
depending on
the cancer and
exposure of
interest
Methodological
Details
Population-
based case-
control study
using
unconditional
logistic
regression
Measure of Pb
Exposure/
Concentration
Data
nterviewof job
history and
exposure matrix
Ever exposed to:
Organic Pb
3.0%
Inorganic Pb
17.0%
Pb in gasoline
emissions
38.6%
Adjusted Effect Estimates
OR (95% Cl):
Never exposed is referent group
Pb in gasoline emissions:
Esophageal 0.6 (0.4, 1.1)
Stomach 1.0(0.7, 1.4)
Colon 0.8 (0.6, 1.1)
Rectum 1.0 (0.7, 1.4)
Pancreas 0.9 (0.5, 1.4)
Prostate 0.9 (0.7, 1.2)
Bladder 0.8 (0.6, 1.1)
Kidney 1.0(0.7, 1.5)
Melanoma 0.8 (0.5, 1.4)
Non-Hodgkin's lymphoma
0.7(0.5, 1.0)
Note: results are for
comparisons using population-
based controls; results for
controls with other types of
Potential
confounders
adjusted for in
analysis
Age, family
income, cultural
origin or
birthplace, and
proxy status; all
models except
those for
melanoma and
non-Hodgkin's
lymphoma were
adjusted for
smoking
cancers were similar except no
association was present
between organic Pb and rectal
cancer
4-755
-------
Table 4-49 (Continued): Summary of recent epidemiologic studies of cancer incidence and overall cancer mortality.
Reference
(In order of
appearance in Study
text) Location
Santibanez et al. Valencia
(2008) and
Alicante,
Spain
Cancer Study
Outcome Population
Esophageal PANESOES
cancer study
incidence included 30-80
yr old men
hospitalized in
any of the
participating
study hospitals
N controls
=285
N cancer =185
(147
squamous cell,
QQ aH&nn
GO duel IU-
carcinoma)
Methodological
Details
Case-control
study using
unconditional
logistic
regression
Measure of Pb
Exposure/
Concentration
Data
Interviews to
determine
occupational history
and a job exposure
matrix
Concentration: NA
Adjusted Effect Estimates
OR (95% C I):
All esophageal cancers:
Unexposed: 1.00
Low workplace Pb exposure
(<4.9ug/dL): 0.79(0.43, 1.46)
High workplace Pb exposure
(>4.9 ug/dL): 1.69(0.57, 5.03)
Esophageal squamous cell
carcinoma:
Unexposed: 1.00
Low workplace Pb exposure
(<4.9ug/dL): 0.70(0.34, 1.43)
High workplace Pb exposure
(>4.9 ug/dL): 0.91 (0.22, 3.75)
Potential
confounders
adjusted for in
analysis
Age, hospital
location,
educational
level, smoking
and alcohol use
Adenocarcinoma:
Unexposed: 1.00
Low workplace Pb exposure
(<4.9ug/dL): 0.95(0.32,2.82)
High workplace Pb exposure
(>4.9 ug/dL): 5.30(1.39,20.22)
*Note: associations not changed
or slightly increased when
restricted to occupational
exposures > 15yr
4-756
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4.10.1.1 Overall Cancer Mortality
Several recent cohort studies examined the association between Pb concentrations and
cancer mortality, including multiple analyses of the NHANES III population. In one
NHANES III analysis, the cohort of 13,946 (N for cancer mortality=411) was followed
for 12 years and individuals with blood Pb levels greater than 10 (ig/dL were excluded
from the study (mean baseline blood Pb level was 2.58 (ig/dL). No association was
observed between blood Pb and cancer mortality (HR of highest tertile [> 3.63 (ig/dL]
compared to lowest tertile [<1.93 (ig/dL]: 1.10 [95% CI: 0.82, 1.47]) (Menke et al..
2006). Another analysis of the NHANES III population, which was restricted to
individuals 40 years and older at the time of blood Pb collection and included 9,757 (N
for cancer mortality=543) individuals with all blood Pb levels (including those greater
than 10 (ig/dL), reported associations between blood Pb and cancer mortality (Schober et
al., 2006). In this study, median follow-up time was 8.6 years. The RRs were 1.69 (95%
CI: 1.14, 2.52) for individuals with blood Pb levels of at least 10 (ig/dL and 1.44 (95%
CI: 1.12, 1.86) for blood Pb levels of 5-9 (ig/dL compared to individuals with blood Pb
levels less than 5 (ig/dL. When stratified by age, point estimates comparing blood Pb
levels of 5-9 versus less than 5 (ig/dL were similar across all age groups but only
statistically significant among 75-84 year olds. The risks of mortality associated with
blood Pb levels > 10 (ig/dL in the groups aged 40-74 years and 85 years and older were
elevated.
A study of men (primarily white) from the greater Boston, MA area enrolled in the
Normative Aging Study (NAS) found no association between blood or bone Pb and
cancer mortality in adjusted analyses after an average of 8.9 years of follow-up (N=l,038,
N for cancer mortality=85 when using blood Pb measures; N=727, N for cancer
mortality=57 when using bone Pb measures). At baseline, the mean (SD) blood Pb level
for this population was 5.6 (3.4) (ig/dL, and blood Pb was poorly correlated with
measured bone Pb (Weisskopf et al.. 2009). As part of the Study of Osteoporotic
Fractures, 533 white women aged 65-87 (N for cancer mortality=38) were included in a
sub-study of blood Pb level and cancer mortality and were followed for approximately 12
years (Khalil et al.. 2009b). The mean (SD) blood Pb level at baseline was 5.3
(2.3) (ig/dL, and no association was detected between blood Pb and cancer mortality in
the study population.
Overall, recent epidemiologic studies of blood Pb levels and cancer mortality reported
inconsistent results. An epidemiologic study using NHANES III data demonstrated the
strongest association between blood Pb and increased cancer mortality; however, other
4-757
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studies reported weak or no associations. These cohort studies were well-conducted
longitudinal studies with control for potential confounders, such as age, smoking, and
education (see list of the potential confounders addressed in each study in Table 4-49).
One limitation is that the studies in populations other than NHANES cohorts each had a
small number of cancer mortality cases.
4.10.1.2 Overall Cancer Incidence
Studies of overall cancer incidence have also been performed (Table 4-49). An ecologic
analysis compared Pb-related air pollution over 5 year increments from 1990 to 2005
with incidence rates of cancer during this time period (cancer sites not specified) among
children (N not specified) (Absalon and Slesak. 2010). The highest Pb air pollution levels
were measured in 1990 when over 50% of the study area exceeded the limit of
1 (ig/m3 per year. No correlation was observed both overall and in sex-specific analyses.
Another study (N= 1,940) examined correlations between Pb concentrations in soil, water,
vegetation, and hair samples with incidence of neoplasms (Obhodas et al.. 2007). The Pb
concentrations were not correlated with incidence of neoplasms. A recent study using the
2007-2008 NHANES cohort reported no association between higher creatinine-adjusted
urine Pb levels and elevated odds of having ever had cancer or a malignancy (N=l,857)
(Mendy etal., 2012). The timing of cancer diagnosis in relation to the urine sample
collection was not identified.
Overall, recent epidemiologic studies reported no positive associations between various
measures of Pb exposure and overall cancer incidence. These studies are limited by their
ecologic and cross-sectional designs. Absalon and Slesak (2010) and Obhodas et al.
(2007) did not collect biological measurements, and no control for potential confounding
was mentioned.
4.10.1.3 Lung Cancer
Most of the recent evidence regarding lung cancer incidence is provided by a few studies
of occupationally-exposed adults. These are described in Table 4-49. Some studies in the
2006 Pb AQCD (U.S. EPA. 2006b) reported associations between Pb exposure and lung
cancer in occupational cohorts, although the studies were limited due to possible
confounding by smoking or other workplace exposures. In a more recently published
study of smelter workers (N=187, N for lung cancer=46), no association was observed
between several metrics of Pb exposure (peak blood Pb values, number of years Pb
samples were obtained, and cumulative blood Pb index) and lung cancer incidence and
4-758
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mortality combined (Lundstrom et al.. 2006). The median follow-up in the study was
about 30 years, and the median peak blood Pb values during employment were
49.7 (ig/dL for lung cancer cases and 55.9 (ig/dL for controls. In a study of 1,462 tin
smelter workers, no association was observed between Pb exposure and lung cancer
mortality in unweighted analyses, but when the analyses were weighted by age and time
since exposure, positive associations were apparent (Jones et al.. 2007). In this study,
cumulative Pb exposure was calculated by combining historical air sampling data and
personnel record cards, which specified work histories. The median cumulative Pb
exposure was estimated to be approximately 2 mg/m3 per year. It is important to note that
the tin smelter workers were exposed to other metals as well, such as As and antimony,
and the study did not specify if additional potential confounders were evaluated (beyond
the weighting for age and time since exposure). A population-based case-control study
performed among men in Montreal, Canada in the 1980s assessed occupational Pb
exposure via interviews regarding job histories and determined the likely Pb exposures
associated with the job activities (Rousseau et al.. 2007). No association was apparent
between organic Pb, inorganic Pb, or Pb from gasoline emissions and lung cancer (N
ranged from 271 to 1,203 depending on the exposure of interest).
Studies were also conducted that compared lung tissue Pb measurements for individuals
with lung cancer to those without lung cancer. The controls for these studies were
individuals with metastases in the lung from other primary cancers (De Palmaetal.
2008) and individuals with non-cancerous lung diseases (De Palma et al., 2008; Kuo et
al.. 2006). Limitations in these studies include their cross-sectional design, the
measurement of Pb in cancerous tissue, which may have altered Pb distribution, and the
use of controls with other cancers and lung diseases. Findings are mixed among the
studies. De Palma et al. (2008) reported higher Pb concentrations in the cancerous and
non-cancerous lung tissue of individuals with non-small cell lung cancer compared to
control groups, although the authors report these results may be confounded by smoking.
Kuo et al. (2006) found no statistical difference in Pb levels for lung tissue of individuals
with lung cancer compared to controls.
In summary, some studies in the 2006 Pb AQCD reported associations between
indicators of Pb exposure and lung cancer among occupational cohorts; however, recent
epidemiologic studies of lung cancer reported no associations. Overall, these recent
epidemiologic studies included only men, limiting the generalizability. The studies by
Jones et al. (2007) and Rousseau et al. (2007) also have the disadvantage of not obtaining
actual measures of Pb exposure or biomarker levels. In addition, these studies, as well as
those in the 2006 Pb AQCD, are of occupational cohorts, and the relationships with Pb
exposures may be confounded by other workplace exposures and covariates that were not
considered, such as smoking.
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4.10.1.4 Brain Cancer
A few studies of brain cancer examined the association between cancer and occupational
Pb exposure using exposures determined via exposure databases and patient interviews
about past jobs and known exposures Table 4-49). The National Longitudinal Mortality
Study, a study that included a national sample of the U.S. population (N=317,968),
estimated occupational Pb exposure based on current/most recent employment among
individuals (van Wijngaarden and Dosemeci. 2006). Although not all estimates using
various statistical techniques and measures of Pb exposure/intensity are statistically
significant, a pattern of increased associations between Pb exposure and brain cancer
mortality was observed in the study population. In a case-control study of brain tumors
(N for controls=799, N for glioma=489, N for meningioma =197), glioma was reported to
have no association with any Pb exposure metric. However, positive associations were
observed between high cumulative occupational Pb exposure and meningioma among
individuals withALAD2 genotypes (OR 2.4 [95% CI: 0.7, 8.8] comparing individuals
ever exposed to Pb with those not exposed to Pb; OR 12.8 [95% CI: 1.4, 120.8]
comparing individuals with cumulative Pb exposure > 100 (ig/m3 per year to those not
exposed to Pb) (Rajaraman et al.. 2006). This association was not present among
individuals with the ALAD1 homozygous genotypes (OR 0.5 [95% CI: 0.3, 1.0]
comparing individuals ever exposed to Pb with those not exposed to Pb; OR 0.7 [95% CI:
0.2, 1.8] comparing individuals with cumulative Pb exposure > 100 (ig/m3 per year to
those not exposed to Pb). Another study of the association between occupational Pb
exposure (measured using self-reported occupational exposure history) and brain tumors
reported none or slight overall associations with types of brain tumors; however, positive
associations were observed among individuals with certain genetic single nucleotide
polymorphisms (SNPs) (n for controls=494, n for glioma=362, n for meningioma =134)
(Bhatti et al., 2009). After control for multiple comparisons, individuals with the
rs 105 0450 GPX1 variant had positive associations between cumulative Pb exposure and
glioblastoma multiforme and meningioma. Individuals without the rs2239774^4C2
variant showed a positive association between Pb and glioblastoma multiforme. Also,
individuals without the rs75 74920 XDH variant displayed a positive association between
Pb and meningioma.
Overall, associations between occupational Pb exposure and brain cancer incidence and
mortality were found to vary according to several genetic variants. Studies of the
association between Pb exposure and brain cancer were not reported in the
2006 Pb AQCD. These studies were limited in the implications of their results because
they do not have individual level biological or exposure Pb measurements and the
potential for confounding by other workplace exposures exists.
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4.10.1.5 Breast Cancer
The association between proximity to a Pb smelter and breast cancer was evaluated in a
non-occupational cohort. A population-based case-control study in Canada (N for
controls=2,467, N for cases=2,343) examined the proximity to a Pb smelter based on
residential addresses (Pan etal. 2011) (Table 4-49). No association was reported
between proximity of a Pb smelter and breast cancer incidence, but the study was limited
by the small number of women who resided near a Pb smelter (n=13 lived < 3.2 km from
Pb smelter). No Pb biomarker or exposure data were available.
A few case-control studies examined Pb levels in biological samples among individuals
with and without breast tumor and/or cancer. A study of newly diagnosed breast cancer
patients and controls examined Pb levels in blood and hair samples and reported higher
Pb levels in both for cancer cases, although the difference in the Pb content in hair
samples was not statistically significant (Alatise and Schrauzer. 2010). Siddiqui et al.
(2006) observed higher blood Pb levels in women with benign and malignant tumors
compared to controls. Additionally, although blood Pb levels were higher among those
with malignant breast tumors compared to those with benign tumors, both had similar
levels of Pb detected in breast tissues. Another study of Pb levels present in breast tissue
reported no statistical difference in Pb levels (Pasha et al.. 2008b). However, a study of
breast tissue did observe a statistically significant difference between Pb levels in the
breast tissue of cancer cases and controls (lonescu et al.. 2007). Finally, a study of Pb
levels in urine reported a positive association between urine Pb and breast cancer, but this
association became null when women taking nonsteroidal aromatase inhibitors but not
taking bisphosphonates (a combination responsible for bone loss) were excluded from the
analysis (McElroy et al.. 2008).
The 2006 Pb AQCD (U.S. EPA. 2006b) did not report any studies examining Pb levels
and breast cancer. Overall, recent studies suggest that women with breast cancer may
have higher blood Pb levels than those without breast cancer. However, results are mixed
in studies that compared breast tissue Pb concentrations between breast tumor and control
samples. These studies are limited by their design. The samples are taken after cancer is
already present in the cases, thus, the directionality between tissue or blood Pb levels and
cancer development cannot be established. Additionally, the sample sizes are often small,
and the studies may be underpowered [most of the studies had less than 25 cases (Alatise
and Schrauzer. 2010; lonescu et al.. 2007; Siddiqui et al.. 2006)1. A case-control study,
also limited by its method of exposure measurement, reported no association between
living near a Pb smelter and breast cancer (Pan et al.. 2011).
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4.10.1.6 Other Cancers
There have been a few studies of cancer types other than those listed above. The
2006 Pb AQCD reported evidence of an association between Pb exposure and stomach
cancer in several occupational cohorts. A study performed among men in Montreal,
Canada in the 1980s evaluated multiple cancer outcomes and estimated occupational
exposures to organic Pb, inorganic Pb, and Pb from gasoline emissions via interviews
regarding job histories and subsequent exposure approximations by chemists and
hygienists (N for cases and controls varied from 60 to 2,250 based on the cancer type and
exposure) (Rousseau et al.. 2007). Adults with occupational exposure to organic Pb
exposure had greater odds of having stomach cancer compared to adults without
occupational exposure to organic Pb. A positive association was also observed for rectal
cancer when population-based controls were used but was null when the control
population was limited to individuals with other types of cancers. No association was
detected for cancers of the esophagus, colon, pancreas, prostate, bladder, kidney,
melanoma, or non-Hodgkin's lymphoma. None of the cancers were associated with
occupational exposure to inorganic Pb. When occupational exposure to Pb in gasoline
emissions was categorized as "unexposed," "nonsubstantial level," or "substantial level,"
a positive association with stomach cancer was observed when cancer controls were used
as the comparison group; however, the association was not present when population
controls were utilized as the control group). Another case-control study using participant
interviews and a job exposure matrix, including only men, (N for controls=285, N for
cancer=185) reported no association between occupational Pb exposure and esophageal
squamous cell carcinomas, but an association was present between high occupational Pb
exposure and adenocarcinoma of the esophagus (Santibanez et al.. 2008). However,
neither of these studies quantified Pb levels in biological or exposure samples.
Several studies compared Pb levels in blood, tissue, and urine of individuals who have
cancer with Pb levels in individuals who are cancer-free. Compared to control groups,
higher Pb levels were observed in the blood and bladder tissue of individuals with
bladder cancer (Golabek et al.. 2009). the kidney tissue of individuals with renal cell
carcinoma (with highest levels among those with the highest stage tumors) (Calvo et al..
2009). the tissue (but not serum) of individuals with laryngeal cancer (Olszewski et al..
2006). the blood of individuals with gastric cancer (Khorasani et al.. 2008). the plasma
and hair of individuals with gastrointestinal cancer (Pasha etal.. 2010). the blood and hair
of individuals with non-specified types of cancer (Pasha et al.. 2008c: Pasha et al.. 2007).
and the hair of individuals with benign tumors (Pasha et al.. 2008a). No statistical
difference in Pb levels was reported for colon tissue of individuals with colorectal polyps
(Alimonti et al.. 2008) or urine of individuals with bladder cancer (Lin et al.. 2009)
compared to control groups. A study examining Pb levels in kidney tissue reported the
4-762
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highest levels of Pb in normal kidney tissue samples that were adjacent to neoplastic
tumors. The Pb levels reported in the kidney tissue of neoplastic tumors were elevated
compared to those detected in corpses without neoplastic tumors of the kidney (Cerulli et
al.. 2006). All of these comparison studies are limited by the inability to determine
temporality as Pb biomarkers were measured after the cancer diagnosis; the level of Pb
may be due to changes that result from having cancer, not changes that result in cancer.
Many of these studies attempted to control for this by including only cases that have not
undergone certain treatments. Additionally, studies are limited by their small sample
sizes and the selection of the control populations. Control populations are supposed to
represent the general population from which the cases are drawn; some of the control
subjects in these studies are individuals with diseases/conditions warranting tissue
resections, which are not prevalent in the general population.
In summary, recent epidemiologic studies examining the potential for associations of Pb
exposure with the incidence of specific cancers reported varying associations with
occupational Pb exposure. Associations were null for occupational Pb exposure and most
cancer sites examined. However, a positive association was observed between
occupational Pb exposure and adenocarcinoma of the esophagus as well as exposure to
occupational organic Pb and stomach cancer, which is supported by evidence of a
relationship between Pb exposure and stomach cancer in occupational cohorts reported in
the 2006 Pb AQCD. Associations between occupational organic Pb exposure and rectal
cancer and occupational exposure to Pb in gasoline emissions and stomach cancer were
inconsistent. These studies of various cancer sites have limited generalizability due to the
study populations comprising only men. In addition, there are no personal biological or
exposure samples used in the epidemiologic analyses and confounding by other
occupational exposures is possible. In other studies, biological samples were used in
biomarker comparisons of cancer and cancer-free individuals but as stated above, these
studies have multiple other limitations.
4.10.1.7 Animal Models of Carcinogenicity
Previous AQCDs have established that Pb has been shown to act as a carcinogen in
animal toxicology models, albeit at relatively high concentrations. Chronic oral
Pb acetate exposure to male and female rodents has consistently been shown to be a
kidney carcinogen in multiple separate studies, inducing adenocarcinomas and adenomas
after chronic exposure. Gliomas of the brain have also been reported after oral Pb
exposure. The kidneys are the most common target of Pb-induced carcinogenicity
(Kasprzak et al.. 1985; Roller etal.. 1985; Azaretal.. 1973; Van Esch and Kroes. 1969)
but the testes, brain, adrenals, prostate, pituitary, and mammary gland have also been
4-763
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affected (IARC. 2006a). The typical cancer bioassays used by IARC or NTP as evidence
of Pb-induced carcinogenicity were designed using rodents, typically males but
sometimes animals of both sexes, that were continuously exposed to Pb acetate in chow
(i.e., 1,000 or 10,000 ppm Pb acetate) or drinking water (i.e., 26 or 2,600 ppm Pb acetate)
for 18 months to two years in duration, the typical life span of a rodent (Kasprzak et al.
1985: Roller etal.. 1985: Azaretal. 1973: Van Esch and Kroes. 1969). These two-year
cancer bioassays and the doses employed are typical of cancer bioassays employed for
other chemicals, albeit at doses that are higher than Pb doses cited in other toxicological
sections of the ISA. In cancer bioassays, to obtain statistically valid data from small
groups of animals, doses are selected such that any dose-related effects will occur
frequently enough to be detected. The 2006 Pb AQCD (U.S. EPA. 2006b) pointed out
that because Pb is a "well-established animal carcinogen...., focus has been more on the
mechanism of neoplasia and possible immunomodulatory effects of Pb in the promotion
of cancer." This focus continues to date. More recent studies have focused on
administration of Pb with known carcinogens or antioxidants such that lifestage, diet, and
mode of action can be better understood. Developmental Pb acetate exposure also
induced kidney tumors in offspring whose dams received Pb acetate in drinking water
during pregnancy and lactation.
Recognition of the importance of windows of exposure in Pb-induced cancer bioassays is
a focus of more recent studies. In one study, gestational and lactational exposure of
laboratory rodents to inorganic Pb (500, 750 or 1,000 ppm Pb acetate in drinking water)
induced carcinogenicity in adult offspring (Waalkes etal.. 1995). Another recent study
considered Pb-induced carcinogenesis in laboratory animals with early life Pb exposure
(gestation, lactation and continued until 8 weeks of age) in which Tokar et al. (2010)
examined tumorigenesis in homozygous metallothionein I/II knockout mice and their
corresponding wild type controls (groups often mice each). The dams/mothers were
exposed by drinking water to 2,000 or 4,000 ppm Pb acetate during gestation and
lactation and compared to untreated controls. Study animals were exposed in utero,
through birth and lactation, and then postnatally to drinking water until 8 weeks old. The
Pb-exposed metallothionein I/II knockout mice had increased testicular teratomas and
renal and urinary bladder preneoplasia. The tumor burdens of Pb-exposed wild-type mice
were not statistically significantly different than controls. The data suggest that
metallothionein can protect against Pb-induced tumorigenesis. Concerns with the study
are that the doses are at levels of Pb to which humans would not likely be exposed and
there is no metallothionein null condition in humans, though there is variability in the
expression of metallothionein. The data do not address whether this variability would
have any impact on Pb-induced carcinogenesis in humans. Collectively, the animal
toxicology data demonstrate that Pb is a well-established animal carcinogen in studies
employing high-dose Pb exposure over a continuous extended duration of exposure
4-764
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(i.e., 2 years), which is typical of cancer bioassays. Newer studies are showing early-life
maternal Pb exposure can contribute to carcinogenicity in offspring and suggest that
metallothionein is protective against cancer in this pathway.
4.10.2 Cancer Biomarkers
Studies of Pb exposure and prostate specific antigen (PSA), a biomarker for prostate
cancer, were not reported in the 2006 Pb AQCD (U.S. EPA. 2006b). A recent cross-
sectional study of men aged 21-40 years without occupational history of exposure to
metals examined PSA (N=57). This study reported a positive association between Pb
levels and PSA levels (measured in the same blood samples) in regression models
adjusted for the following potential confounders: age, smoking, alcohol consumption, and
other metals (Cd, Zn, Se, and Cu) (Pizent et al., 2009). The median concurrent blood Pb
level was 2.6 (ig/dL (range 1.0-10.8 (ig/dL). The authors note that the study population
was young and at lower risk of prostate cancer than are older men.
4.10.3 Modes of Action for Pb-induced Carcinogenicity
Most evidence clearly supports the mode of action of Pb inducing carcinogenicity in
animal models, but the exact chain of events supporting this segue has not been
completely characterized. Specifically, it is unclear whether the mode of action of Pb is
best understood within the framework of multistage carcinogenesis, genomic instability
or epigenetic modification. For example, multistage carcinogenesis involves a series of
cellular and molecular changes that result from the progressive accumulation of
mutations that induce alterations in cancer-related genes. Pb does not appear to follow
this paradigm, and the literature suggests it is weakly mutagenic. Pb does appear to have
some ability to induce DNA damage (Section 4.10.3.2). However, the ability of Pb to
alter gene expression through epigenetic mechanism (Section 4.10.3.3) and to interact
with proteins may be a means by which Pb induces carcinogenicity. It is known that Pb
can replace Zn in Zn-binding (Zn-fmger) proteins (Section 4.2). which include hormone
receptors, cell-cycle regulatory proteins, the Ah receptor, estrogen receptor, p53, DNA
repair proteins, protamines, and histones. These Zn-fmger proteins all bind to specific
recognition elements in DNA. Thus, Pb may act at a post-translational stage to alter
protein structure of Zn-fmger proteins, which can in turn alter gene expression, DNA
repair and other cellular functions. To recapitulate, cancer develops from one or a
combination of multiple mechanisms including modification of DNA via epigenetics or
enzyme dysfunction and genetic instability or mutation. These modifications then provide
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the cancer cells with a selective growth advantage. In this schematic, Pb may contribute
to epigenetic changes and chromosomal aberrations.
The genomic instability paradigm requires a cascade of genome-wide changes caused by
impaired DNA repair, kinetochore assembly, cellular checkpoints, centrosome
duplication, microtubule dynamics or a number of cell maintenance processes. These
processes have been rarely studied for Pb, thus there are few data that suggest Pb may
interfere with some of these processes. Furthermore, the bulk of the literature in this area
involves Pb chromate and it is unclear if the effects are due to Pb or chromate. Epigenetic
modifications may lead to cancer by altering cellular functions without altering the DNA
sequence. The most commonly studied epigenetic change is methylation alterations. A
small number of studies show that Pb can induce epigenetic changes (Section 4.10.3.3).
but studies are still missing to clearly tie these effects to Pb-induced carcinogenesis and
genotoxicity. Thus, either genomic instability or epigenetic modification paradigms or
some combination of the two may underlie Pb-induced carcinogenicity.
Exposure to mixtures can also contribute to understanding of modes of action. No recent
studies of the protective role of Ca2+ or Zn in Pb-induced carcinogenesis or genotoxicity
were found. Pb can displace these and other divalent cations, affecting physiological
processes. There were some data suggesting that metallothionein (Section 4.10.4). which
sequesters Pb and makes it less bioavailable, protects rodents from Pb-induced cancers.
Boron, melatonin, N-acetylcysteine, turmeric and myrrh protected cells against
Pb-induced genotoxicity (Section 4.10.3.2) and affected antioxidant status, especially the
glutathione pathway. There were some data suggesting that Pb mimics or antagonizes the
essential micronutrient Se in rodents. These data are discussed in more detail below
(Section 4.10.4) and point to the relevance of mixtures in assessing toxicity.
4.10.3.1 Neoplastic Transformation Studies, Human Cell Cultures
Carcinogenesis can be measured in cell culture systems through neoplastic transformation
models that monitor change by following morphological transformation of cells,
i.e., formation of a focus (or foci) of cell growth. Xie et al. (2007) treated BEP2D cells
(human papilloma virus-immortalized human bronchial cells) with 0, 1, 5, or 10 (ig/cm2
Pb chromate for 120 hours. PbCrO4 induced foci formation in a concentration-dependent
manner. Xie et al. (2008) treated BJhTERT cells (hTERT-immortalized human skin
fibroblasts) and ATLD-2 cells (hTERT-immortalized human skin fibroblasts deficient in
Mrel 1) with 0, 0.1, 0.5, and 1 (ig/cm2 Pb chromate for 120 hours. PbCrO4 induced foci
formation in a concentration-dependent manner in the Mrel 1 deficient cells. Mrel 1 is a
component of the MRN complex and plays a role in telomere maintenance and double-
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strand break repair. Mrel 1 was required to prevent Pb chromate-induced neoplastic
transformation.
Immune Modulation of Tumorigenesis by Pb
As described in the 2006 Pb AQCD (U.S. EPA. 2006b). Pb-induced immunotoxicity can
contribute to increased risk of cancer, primarily as the result of suppressed Thl responses
and misregulated inflammation. First, Pb-induced misregulation of inflammation
involving innate immune cells has been shown to result in chronic insult to tissues. These
insults, excessive lipid and DNA oxidation production by overproduction of ROS and
weakened antioxidant defenses, can increase the likelihood of mutagenesis, cellular
instability, and tumor cell formation. For example, results from Xu et al. (2008) support
the association with Pb exposure and DNA damage, and investigators concluded that it is
a possible route to increased Pb-induced tumorigenesis. The second component of
increased risk of cancer involves Pb-induced suppression of Thl-dependent anti-tumor
immunity as acquired immunity shifts statistically significantly toward Th2 responses.
With cytotoxic T lymphocytes and other cell-mediated defenses dramatically lessened,
the capacity to resist cancer may be compromised.
4.10.3.2 DNA and Cellular Damage
Multiple studies have been performed examining the relationship between Pb and DNA
and cellular damage. Details of the recent epidemiologic and toxicological studies follow.
Epidemiologic Evidence for DNA and Cellular Damage
Multiple studies examined the relationship between blood Pb and sister chromatid
exchange (SCE). SCEs are exchanges of homologous DNA material between chromatids
on a chromosome and are a test for mutagenicity or DNA damage. A study of male
policemen reported a mean blood Pb level for the study population of 43.5 (ig/dL
(Wiwanitkit et al., 2008). In analyses dichotomized as high or low blood Pb levels (cut-
off at 49.7 (ig/dL), the higher blood Pb group was observed to have higher mean SCE.
Another study of adult males compared the SCE of storage battery manufacturing
workers (mean blood Pb level of 40.14 (ig/dL) and office workers (mean blood Pb level
of 9.11 (ig/dL) (Duvdu et al., 2005). The exposed workers had higher SCE levels and also
a greater number of cells in which the SCEs per cell were higher than the 95th percentile
of the population. Finally, a study of children aged 5-14 years old (mean [SD] blood Pb
level of 7.69 [4.29] (ig/dL) reported no correlation between blood Pb levels and SCE
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(Mielzynska et al., 2006). However, the study did report a positive association between
blood Pb and micronuclei (MN) levels.
Other studies of DNA damage have reported mixed results. A study of children ages 6-11
years old and environmentally-exposed to Pb reported no association between blood Pb
and baseline DNA damage or repair ability after a peroxide challenge (children attending
a school far from a Pb smelter: median blood Pb level 4.6 (ig/dL; children attending a
school near a Pb smelter: median blood Pb level 28.6 (ig/dL) (Mendez-Gomez et al..
2008). Another study included adult participants aged 50-65 years and reported an
association between blood Pb and carcinoembryonic antigen (CEA) but not with DNA-
strand breaks, MN frequency, or oxidative DNA damage (median blood Pb level of the
study population: 3.92 (ig/dL) (De Coster et al., 2008). A study conducted among
workers exposed to Pb (mean blood Pb level: 30.3 (ig/dL) and unexposed controls (mean
blood Pb level: 3.2 (ig/dL) reported greater cytogenetic damage (measured by MN
frequency), chromosomal aberrations, and DNA damage in the Pb-exposed group
(although this was not statistically significant in linear regression models controlling for
age) (Grover et al.. 2010). A study of painters in India, where Pb concentrations in paint
are high, reported a mean (SD) blood Pb level of 21.56 (6.43) (ig/dL among painters who
reported painting houses for 8-9 hours/day for 5-10 years (Khan et al.. 2010b): the mean
(SD) blood Pb level was 2.84 (0.96) (ig/dL for healthy workers who had not been
occupationally exposed to Pb. Cytogenetic damage was higher among the painters
compared to the healthy controls. Another study compared the blood Pb of metal workers
and office workers and reported higher blood Pb levels (both current and 2 year average)
among the metal workers (blood Pb level > 20 (ig/dL) compared to the office workers
(blood Pb level <10 (ig/dL) (Olewinska et al.. 2010). Overall, the workers had increased
DNA strand breaks versus the office workers (this held true at various blood Pb levels).
Finally, a study of Pb battery workers with symptoms of Pb toxicity and a group of
controls were examined (Shaik and Jamil. 2009). Higher chromosomal aberrations, MN
frequency, and DNA damage were reported for the battery workers as compared to the
controls. These workplace studies are limited by the lack of consideration for potential
confounding factors, including other occupational exposures.
lexicological Evidence for DNA and Cellular Damage
Sister Chromatic! Exchanges
Pb has been shown to induce SCEs both in vivo and in vitro. Tapisso et al. (2009).
considered SCEs in adult Algerian mice (groups of six mice each) that were treated by
i.p. injection with 5 or 10 doses of 0.46 mg/kg Pb acetate. The SCE in bone marrow were
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elevated after Pb exposure alone and increased with time. Co-exposure with Cd or Zn
further increased SCE levels.
SCE was also followed in cultured human cells. Ustundag and Duydu (2007) considered
the ability of N-acetylcysteine and melatonin to reduce Pb nitrate-induced SCE in a
single human donor. Cells were treated with 0, 1, 5, 10, or 50 (iM Pb nitrate. SCE
statistically significantly increased at every Pb concentration in a concentration
dependent manner. Both 1 and 2 mM N-acetylcysteine and melatonin were able to
statistically significantly reduce SCE levels in Pb-exposed cells. In another study, Turkez
et al. (2011) considered the ability of boron compounds, essential micronutrients, to
prevent Pb chloride-induced SCE in human lymphocytes. Cells were obtained from 4
non-smoking donors. Both 3 and 5 ppm Pb chloride induced a statistically significant
increase in SCE levels over controls. Boron was able to statistically significantly
diminish these levels. For both studies, exposure times were not provided, and the full
interpretation of these data is limited by the limited number of donors and the absence of
an exposure time for the SCE assay.
Micronuclei Formation
The 2006 Pb AQCD stated "studies of genotoxicity consistently find associations of Pb
exposure with DNA damage and MN formation" and recent studies continue to report
these associations. Alghazal et al. (2008b) considered the ability of Pb acetate trihydrate
to induce MN in bone marrow of adult Wistar rats. Animals were given a daily dose of
100 mg/L in their drinking water for 125 days. The mean number of MN in male and
female rats was statistically significantly higher in Pb-exposed animals than in unexposed
controls. Tapisso et al. (2009) considered Pb-induced MN in rodents. Algerian mice were
treated by i.p. injection with 5 or 10 doses of 0.46 mg/kg Pb acetate and compared to
untreated controls. The MN in bone marrow were elevated after Pb exposure and
increased with time.
MN formation has also been followed in cultured human cells. Ustundag and Duydu
(2007) considered the ability of N-acetylcysteine and melatonin to reduce Pb nitrate-
induced MN in a single human donor. Cells were treated with 0, 1,5, 10, or 50 (iM
Pb nitrate. MN formation statistically significantly increased at the two highest Pb
concentrations in a concentration-dependent manner. Both 1 and 2 mM N-acetylcysteine
and melatonin were not able to statistically significantly reduce MN levels. In another
study, Turkez et al. considered the ability of boron compounds to prevent Pb chloride-
induced MN in human lymphocytes. Cells were obtained from 4 non-smoking donors.
Both 3 and 5 ppm Pb chloride induced a statistically significant increase in MN levels
over controls. Boron induced a statistically significant attenuation of these Pb-induced
levels. For both studies, exposure times were not provided, and the full interpretation of
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these data is limited by the limited number of donors and the absence of an exposure time
for the MN assay. Gastaldo et al. (2007) evaluated the ability of Pb to induce MN.
Human endothelial HMEC cell line was treated with 1-1,000 (iM Pb nitrate for 24 hours.
MN increased in a statistically significant, concentration-dependent manner.
Hypoxanthine-guanine phosphoribosyltransferase Mutations
The potential mutagenicity of Pb in human or animal cells has been evaluated by
monitoring mutations at the hypoxanthine-guanine phosphoribosytransferase (HPRT)
locus. Li et al. (2008a) evaluated Pb acetate-induced HPRT in the non-small-cell lung
carcinoma tumor cell line, CL3, and in normal human diploid fibroblasts (specific tissue
source not reported). All cells were exposed to 0, 100, 300, or 500 (iM Pb acetate for
24 hours in serum-free medium ± a 1-hour pretreatment with a MKK1/2 inhibitor or a
PKC-alpha inhibitor. Pb alone did not induce HPRT mutations. Inhibiting the ERK
pathway via either inhibitor statistically significantly increased Pb-induced mutagenesis.
Wang et al. (2008c), investigated Pb acetate-induced HPRT mutations in CL3 cells. All
cells were exposed to 0, 100, 300 or 500 (iM Pb acetate for 24 hours in serum-free
medium ± a 1-hour pretreatment with a PKC-alpha inhibitor or siRNA for PKC-alpha. Pb
alone did not induce HPRT mutations. Inhibiting PKC-alpha via either inhibitor
statistically significantly increased Pb-induced mutagenesis. McNeill et al. (2007)
examined Pb acetate-induced HPRT mutations in Chinese hamster ovary AA8 cells and
AA8 cells overexpressing human Ape 1. Cells were treated with 5 (iM Pb acetate for 6
hours. No increases in HPRT mutations were observed after Pb exposure in either cell
line but with specific pathway perturbations (PKC-alpha or ERK), Pb was able to induce
HPRT mutations.
Chromosomal Aberrations
Chromosomal aberrations, another indicator of cancer risk, were followed in Pb-exposed
rodents (El-Ashmawv et al.. 2006). Dietary exposure to Pb acetate administered as a
single dose of 5,000 ppm w/w to adult male Swiss albino mice caused statistically
significant increased levels of chromosomal aberrations in the Pb treatment alone group,
particularly with respect to fragments, deletions, ring chromosomes, gaps, and end-to-end
associations. In addition, the authors found turmeric and myrrh powders were protective
against the effects of Pb. Concerns with the study include the use of only a single dose of
Pb acetate along with the high levels of unusual aberrations such as ring chromosomes
and end-to-end associations. Typically, these aberrations are rare after metal exposure,
but were the most commonly observed aberration in this study raising questions about the
quality of the metaphase preparations. An additional concern was that only 50
metaphases per dose were analyzed instead of the more common 100 metaphases per
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dose. The authors did not explain why their spectrum of aberrations was so different, why
they only used one dose, or analyzed fewer metaphases per dose.
Multiple studies considered the ability of Pb to induce chromosomal aberrations in
cultured human cells. The ability of Pb nitrate to induce chromosomal aberrations was
examined in primary human peripheral blood lymphocytes obtained from healthy,
nonsmoking donors (Pasha Shaik et al.. 2006). Cells were treated with 0, 1,200 or 2,000
(iM Pb nitrate for 2 hours. No increase in chromosomal aberrations was reported. Some
aneuploidy was observed. Concerns with the study are that only a 2-hour exposure was
used, which may not be long enough for DNA damage to be expressed as a chromosomal
aberration. It also appears from the data presentation that only three subjects were used;
one for a control, one for the low dose and one for the high dose. Experiments were not
repeated, thus given the small number of subjects, this study may not have had sufficient
power to detect any effects. Holmes et al. (2006a), treated WHTBF-6 cells (hTERT-
immortalized human lung cells) with 0, 0.1, 0.5, or 1 (ig/cm2 Pb chromate for 24-120
hours or with 0, 0.1, 0.5, 1, 5 or 10 (ig/cm2 Pb oxide for 24 or 120 hours. Pb chromate
induced statistically significant, concentration-dependent increases in centrosome
abnormalities and aneuploidy; lead oxide did not induce significant changes. Wise et al.
(2006a) treated BEP2D cells with 0, 0.5, 1, 5, or 10 ug/cm2 Pb chromate for 24 hours.
V / 7 ? 5 ? r O
Pb chromate induced statistically significant concentration-dependent increases in
chromosomal aberrations. Holmes et al. (2006b), treated WHTBF-6 cells with 0, 0.1, 0.5,
or 1 (ig/cm2 Pb chromate for 24-72 hours. Pb chromate induced statistically significant,
concentration-dependent increases in chromosomal aberrations. The effects of the
chromate anion cannot be ruled out as causative in inducing these chromosomal
aberrations. Wise et al. (2006b), treated WHTBF-6 cells with 0, 0.1, 0.5, or 1 (ig/cm2
Pb chromate for 24-120 hours. Pb chromate induced statistically significant,
concentration-dependent increases in spindle assembly and checkpoint disruption, effects
of mitosis and aneuploidy. By contrast, chromate-free Pb oxide did not induce
centrosome amplification. The effects were likely attributable to the chromate anion. Xie
et al. (2007) treated BEP2D cells with 0, 1, 5, or 10 (ig/cm2 Pb chromate for 24 hours.
Pb chromate induced statistically significant, concentration-dependent increases in
chromosomal aberrations and aneuploidy. Wise et al. (2010) treated WHTBF-6 cells with
0, 0.1, 0.5, or 1 (ig/cm2 Pb chromate for 24 hours in a study comparing 4 chromate
compounds. Pb chromate induced statistically significant, concentration-dependent
increases in chromosomal aberrations.
Multiple investigators considered the ability of Pb chromate to induce chromosome
aberrations in rodent cell cultures. Grlickova-Duzevik et al. (2006) treated Chinese
hamster ovary (CHO) cells with 0, 0.1, 0.5, or 1 (ig/cm2 Pb chromate for 24 hours.
Specific CHO lines used included AA8 (wildtype) EM9 (XRCC1-deficient), and H9T3
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(EM9 complemented with human XRCC1 gene). Pb chromate induced statistically
significant, concentration-dependent increases in chromosomal aberrations that were
statistically significantly increased by deficiency of the DNA repair geneXRCCl.
Nestmann and Zhang (2007) treated Chinese hamster ovary cells (clone WB(L)) with 0,
0.1,0.5, 1,5, orlO (ig/cm2 Pb chromate (as pigment yellow) for 18 hours. No increases
in chromosomal aberrations were observed. Savery et al. (2007) treated CHO cells with
0, 0.1, 0.5, 1, or 5 (ig/cm2 Pb chromate for 24 hours. Specific CHO lines used included
AA8 (wildtype), KO40 (Fa«cg-deficient), and 40BP6 (Fa«cg-complemented). The
Fancg gene plays an important role in cellular resistance to DNA interstrand crosslinks,
protecting against genetic instability. Pb chromate induced statistically significant,
concentration-dependent increases in chromosomal aberrations that were increased by
^a«cg-deficiency. Camyre et al. (2007) treated CHO cells with 0, 0.1, 0.5, 1, 5, or
10 (ig/cm2 Pb chromate for 24 hours. Specific CHO lines used included CHO-K1
(parental), xrs-6 (Ku80 deficient), and 2E (xrs-6 complemented with Ku80 gene).
Pb chromate induced statistically significant, concentration-dependent increases in
chromosomal aberrations that were not affected by Ku80 deficiency. Ku80 is a gene
involved in nonhomologous end-joining repair and its absence can contribute to genetic
instability. Stackpole et al. (2007) treated CHO and Chinese hamster lung (CHL) cells
with 0, 0.1, 0.5, or 1 (ig/cm2 Pb chromate for 24 hours. Specific CHO lines used included
AA8 (wildtype), irslSF (XRCCS-deficient), and ISFwtS (XRCC3 complemented).
XRCC3 is DNA repair enzyme involved in homologous recombination. CHL lines used
included V79 (wildtype), irs3 (Rad51C deficient) and irs3#6 (Rad51C complemented).
Rad51C is a gene that encodes strand-transfer proteins that are thought to be involved in
recombinational repair of damaged DNA and in meiotic recombination. Pb chromate
induced statistically significant, concentration-dependent increases in chromosomal
aberrations that were statistically significantly increased by both XRCC3 and Rad51C
deficiency.
Multiple studies considered the ability of Pb chromate to induce chromosome aberrations
in marine mammal cell cultures. Li Chen et al. (2009) treated primary North Atlantic
right whale lung and skin fibroblasts with 0, 0.5, 1.0, 2.0, and 4.0 (ig/cm2 Pb chromate for
24 hours. Wise et al. (2009) treated primary Steller sea lion lung fibroblasts with 0, 0.1,
0.5,1 and 5 (ig/cm2 Pb chromate for 24 hours. Wise et al. (2011) treated primary sperm
whale skin fibroblasts with 0, 0.5, 1, 3, 5, and 10 (ig/cm2Pb chromate for 24 hours. In all
three studies, Pb chromate induced statistically significant, concentration-dependent
increases in chromosomal aberrations.
In summary, exposure of various cell models and an in vivo model to Pb (acetate,
chromate, or nitrate) induced significant increases in chromosomal aberration that often
responded in a concentration-dependent manner. The use of various cell lines deficient in
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specific DNA repair enzymes helped to elucidate which pathways may be most sensitive
to Pb-dependent chromosomal aberration. However, a number of studies used
Pb chromate exposures and the effects of the chromate anion cannot be ruled out as
causative in inducing these chromosomal aberrations.
COMETAssay
Multiple studies considered the ability of Pb to induce DNA single strand breaks in
laboratory animals and human and animal cells using the COMET assays. The COMET
assay measures DNA damage assessed by single cell electrophoresis of a lysed cell and
measurement of the fragmented DNA or tail length. Xu et al. (2008) examined DNA
damage in male ICR mice treated with Pb acetate. Animals (5 per group) were given
Pb acetate by gavage at doses of 0, 10, 50, or 100 mg/kg body weight every other day for
4 weeks. Pb exposure statistically significantly increased both tail length and tail moment
in a concentration-dependent manner. Nava-Hernandez et al. (2009) considered the
ability of Pb acetate to induce DNA damage in primary spermatocyte DNA of male
Wistar rats. Animals (3 per group) were treated for 13 weeks with 0, 250, or 500 mg/L Pb
in their drinking water. There was statistically significantly less DNA damage in the
controls compared to the two treatment groups. Narayana and Al-Bader (2011) examined
DNA damage in liver tissue of adult male Wistar rats exposed to Pb nitrate. Animals (8
per group) were treated for 60 days with doses of 0, 5,000, or 10,000 ppm Pb nitrate in
their drinking water. There were no statistical differences between treated groups and
controls. Drosophila melanogaster larvae (72 hours old) exposed to Pb nitrate (2,000,
4,000, and 8,000 (iM in culture media for 24 hours) yielded haemocytes that tested
positive in the COMET assay; Pb chloride (8,000 \\M) did not cause DNA damage with
the COMET assay (Carmona et al.. 2011).
Other studies used the COMET assay in cultured human cells. Pasha Shaik et al. (2006)
treated primary human peripheral blood lymphocytes obtained from healthy, nonsmoking
donors with 0, 2.1x, 2.4x, 2.7x, 3.Ox, and 3.3xl03 (iM Pb nitrate for 2 hours and found
concentration-dependent increases in COMET tail length. Concerns with the study are
that apparently no negative control was used. It also appears from the data presentation
that only five subjects were used; one for each dose. Experiments were not repeated.
Thus, given the small number of subjects and the absence of a negative control, this study
may only be detecting background levels of DNA damage. Xie et al. (2008) treated
BJhTERT cells (hTERT-immortalized human skin fibroblasts) and ATLD-2 cells
(hTERT-immortalized human skin fibroblasts deficient in Mrel 1) with 0, 0.1, 0.5, and
1 (ig/cm2 Pb chromate for 24 hours. Pb chromate induced a concentration-dependent
increase in DNA double strand breaks measured by the COMET assay. Pb chromate
exposure and the effects of the chromate anion cannot be ruled out as causative in
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inducing these aberrations. In another study, Pb nitrate exposure (30 (ig/mL) induced
statistically significant increased DNA damage in human liver HepG2 cells that was
attenuated with co-exposure with the antioxidant NAC (500 (iM) (Yedjou et al.. 2010).
Other studies used the COMET assay to examine Pb-induced DNA single strand breaks
in rodent cell cultures. Xu et al. (2006). treated PC12 cells with 0, 0.1, 1 or 10 (iM
Pb acetate. Both tail length and tail moment statistically significantly increased in a
concentration-dependent manner. Kermani et al. (2008) exposed mouse bone marrow-
mesenchymal stem cells to 60 (iM Pb acetate for 48 hours. There was an increase in
several COMET assay measurements including tail length.
The COMET assay showed multiple positive findings of DNA damage after Pb exposure
in rodents, flies, primary human cells, and cell lines. In vivo studies with rodents exposed
to Pb acetate yielded significant increases in tail length and moment via COMET assays
in separate studies that used lymphocytes and sperm. In Drosophila, Pb nitrate but not
Pb chloride produced significant increases with the COMET assay. Human cell culture
from primary cells (lymphocytes) and from cell lines (fibroblasts and liver) had increased
DNA damage as assessed using the COMET assays with separate Pb nitrate and
Pb chromate exposures. Thus, the COMET assay showed multiple positive findings of
DNA damage after in vitro and in vivo Pb exposure.
Other Indicators of DNA Damage
Other studies considered the ability of Pb to induce DNA double strand breaks by
measuring gamma-H2A.X foci formation in cultured human cells. Xie et al. (2008)
treated BJhTERT cells (hTERT-immortalized human skin fibroblasts) and ATLD-2 cells
(hTERT-immortalized human skin fibroblasts deficient in Mrel 1) with 0, 0.1, 0.5, and
1 (ig/cm2 Pb chromate for 24 hours. Pb chromate induced a concentration-dependent
increase in DNA double strand breaks measured by gamma-H2A.X foci formation.
Pb chromate exposure and the effects of the chromate anion cannot be ruled out as
contributory. Gastaldo et al. (2007) evaluated the ability of Pb to induce DNA double
strand breaks with both gamma-H2A.X foci formation and pulse-field gel electrophoresis
in cultured human cells. The human endothelial HMEC cell line was treated with 1 to
1,000 (iM Pb nitrate for 24 hours. DNA double strand breaks increased in a
concentration-dependent manner. Wise et al. (2010) treated WHTBF-6 cells with 0, 0.1,
0.5, or 1 (ig/cm2 Pb chromate for 24 hours in a study comparing four chromate
compounds. Pb chromate induced statistically significant, concentration-dependent
increases in DNA double strand breaks measured by gamma-H2A.X foci formation, at a
similar level to the three other compounds. A few studies demonstrated the ability of Pb
to destabilize DNA by forming DNA-histone cross links, which can lead to histone
aggregation (Rabbani-Chadegani et al., 2011; Rabbani-Chadegani et al.. 2009). In
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extracts of rat liver, Pb nitrate (<300 (iM) was shown to react with chromatin components
and induce chromatin aggregation via histone-DNA cross links.
Genotoxicity testing ofDrosophila melanogaster larvae (72 hours old) using the Wing
Spot test showed that neither Pb chloride nor Pb nitrate (at concentrations of 2,000, 4,000
and 8,000 (iM in culture media with exposure until pupation) was able to induce
significant increases in the frequency of wing spots (Carmona et al., 2011). The wing
spot test can detect mitotic recombination and multiple mutational events such as point
mutations, deletions, and certain types of chromosome aberrations (Graf and Wurgler.
1986). Further, wing spot assays employing Pb co-exposure with gamma radiation
showed no effect of Pb on gamma radiation induced spotting frequency.
Multiple studies examined the effects of Pb on DNA repair. Most were conducted in
cultured cells, and one was done in an animal model. El-Ghor et al. (2011) followed
microsatellite instability (MSI) in Pb acetate trihydrate exposed adult male rats. MSI
reflects impaired DNA mismatch repair and contributes to an increased risk of cancer.
DNA from leukocytes of adult male albino rats exposed to Pb acetate (acute: single oral
dose of 467 mg/kg BW or sub-chronic: 47 mg/kg BW six days/week for 4 week) showed
increased MSI at three microsatellite loci (D6mit3, D9mit2, and DISMghl). This study is
limited by its small sample size (n=2 to 3 rodents per treatment group). Li et al. (2008a)
evaluated Pb acetate-induced effects on nucleotide excision repair efficiency in CL3
cells. All cells were exposed to 0, 100, 300 or 500 (iM Pb acetate for 24 hours in serum-
free medium. Pb increased nucleotide excision repair efficiency. Gastaldo et al. (2007)
evaluated the ability of Pb to affect DNA repair in cultured human cells. The human
endothelial HMEC cell line was treated with 100 (iM Pb nitrate for 24 hours. Pb inhibited
non-homologous end joining repair, over activated MRE11-dependent repair, and
increased Rad51-related repair. Xie et al. (2008) treated BJhTERT cells (hTERT-
immortalized human skin fibroblasts) and ATLD-2 cells (hTERT-immortalized human
skin fibroblasts deficient in Mrel 1) with 0, 0.1, 0.5, and 1 (ig/cm2 Pb chromate for 24 or
120 hours. Mrel 1 was required to prevent Pb chromate-induced DNA double strand
breaks. In this finding, Pb chromate exposure and the effects of the chromate anion
cannot be ruled out as causative. McNeill et al, (2007) considered Pb acetate effects on
Apel. Chinese hamster ovary cells (AA8 subtype) were treated with 0, 0.5, 5, 50, or
500 (iM Pb acetate and then whole cell extracts were used to determine AP site incision
activity. The data show that Pb reduced AP endonuclease function. Finally, studies
considered Pb-induced cellular proliferation in laboratory animals. Kermani et al. (2008)
exposed mouse bone marrow-mesenchymal stem cells to 0-100 (iM Pb acetate for 48
hours. As measured by the MTT assay, Pb decreased cell proliferation at all
concentrations tested. An earlier study in rats showed Pb nitrate-induced increased
proliferation of liver cells after a partial hepatectomy, with more prominent effects found
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in males than females (sexual dimorphism) (Tessitore et al.. 1995). Recent studies
showed similar trends in males. Fortoul et al. (2005) exposed adult male CD1 mice (24
animals per group) to IxlO4 (iM Pb acetate, 0.006 M Cd chloride or a mixture of the two
chemicals for 1 hour twice a week for 4 weeks by inhalation. Electron microscopy
indicated Pb-induced cellular proliferation in the lungs.
4.10.3.3 Epigenetics
Air pollution exposure is being linked increasingly with epigenetic changes in humans
and toxicological models (Pavanello et al., 2010; Baccarelli and Bollati. 2009; Tarantini
et al.. 2009; Bollati et al.. 2007). Epigenetic changes are changes in gene expression that
occur without actual changes in the DNA sequence, and these changes may be heritable.
Epigenetic changes are mediated by histone modification, DNA methylation, miRNA
changes, or pathways that affect these three mediators. Differential epigenetic
modification has the possibility to contribute to disease. Epigenetic studies have been
conducted to examine the associations between Pb biomarker levels and global DNA
methylation markers [Alu and long interspersed nuclear element-1 (LINE-1)] in humans
(Wright etal. 2010; Pilsner et al.. 2009). Wright et al. (Wright. 2013; Wright et al..
2010) examined men from the Normative Aging Study (N=517) with mean (SD) Pb
levels of 20.5 (14.8) (ig/g in tibia, 27.4 (19.7) (ig/g in patella, and 4.1 (2.4) (ig/dL in
blood. In both crude and adjusted analyses, patella Pb levels were inversely associated
with LINE-1 methylation but not with Alu. The adjusted models all included age, body
mass index, percent lymphocytes, with some adjusted models also controlling for
education, smoking, and blood Pb levels. In examination of the relationship between
patella Pb and LINE-1 more closely, a non-linear trend was observed with a smaller
magnitude of effect estimated for higher patella Pb (> 40 (ig/g). No associations were
observed for tibia or blood Pb and either LINE-1 or Alu. Another study included
maternal-infant pairs from the Early Life Exposures in Mexico to Environmental
Toxicants study (N=103) and measured LINE-1 and Alu methylation in umbilical cord
blood samples (Pilsner et al.. 2009). In unadjusted models, maternal tibia Pb levels
one month postpartum (mean [SD]: 10.5 [8.4] (ig/g) were inversely associated with Alu
methylation in the cord blood. Maternal patella Pb levels one month postpartum (mean
[SD]: 12.9 [14.3] (ig/g) were inversely associated with LINE-1 methylation. The
associations persisted in models adjusted for maternal age, maternal education, sex of
infant, smoking during pregnancy, and umbilical cord blood Pb levels (the results were
no longer statistically significant when umbilical cord blood was removed from the
model). No association was detected between umbilical cord Pb levels and the DNA
methylation markers. Overall, the studies consistently demonstrate an association
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between higher patella Pb levels and lower LINE-1 methylation. Lower DNA
methylation is associated with increased gene expression; however, the link between
global DNA methylation and risk of disease, has not been established.
Toxicological studies have examined Pb-induced epigenetic changes and gene
expression, DNA repair, and mitogenesis. Glahn et al., (2008) performed a gene array
study in primary normal human bronchial epithelial cells from four donors after in vitro
treatment of the cells with 55 (ig/dL Pb chloride, 15 (ig/L Cd sulfate, 25 (ig/L Co chloride
or all three combined for 72 hours. The authors describe a pattern of RNA expression
changes indicating "...coordinated stress-response and cell-survival signaling,
deregulation of cell proliferation, increased steroid metabolism, and increased expression
of xenobiotic metabolizing enzymes." These are all known targets of possible epigenetic
changes, but attributing the results to epigenetic changes is complicated. In a recent
publication (Li etal.. 2011). exposure of HepG2 cells to a high dose of Pb (100 (iM
Pb acetate) resulted in ALAD gene promoter hypermethylation and decreased ALAD
transcription. This was in agreement with findings in battery plant workers who showed
ALAD hypermethylation (versus non-occupationally exposed controls) and an
association of this hypermethylation with elevated risk of Pb poisoning (Li etal. 2011).
These latter results have implications for Pb toxicokinetics or disposition of Pb as
modified by ALAD.
4.10.4 Effects of Pb within Mixtures
Several studies considered the impact of Pb as part of a mixture on genotoxicity and
mutagenesis. Mendez-Gomez et al. (2008) evaluated 65 children in Mexico with high
environmental exposures to both As and Pb. DNA damage and decreased DNA repair
were seen with increased proximity to a Pb smelter using the COMET assay and other
assays but did not correlate with blood Pb levels and no interaction between blood Pb and
urinary As levels was observed. Tapisso et al. (2009) examined Pb alone, Pb plus Zn and
Pb plus Cd-induced MN in rodents. Algerian mice (groups of six mice each) were treated
i.p. with 5 or 10 doses of 0.46 mg/kg Pb acetate and compared to untreated controls. The
MN in bone marrow were elevated after Pb treatment alone and increased with time.
Co-exposure with Cd or Zn did not further increase MN levels but did increase SCE
levels. Glahn et al. (2008) performed a gene array study in primary normal human
bronchial epithelial cells from four donors treated with 55 (ig/dL Pb chloride, 15 (ig/L Cd
sulfate, 25 (ig/L Co chloride or all three combined for 72 hours. There was a clear
interaction of all three metals impacting RNA expression.
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Studies in the 2006 Pb AQCD (U.S. EPA. 2006b) found a protective role for calcium in
genotoxic and mutagenic assays with Pb co-exposure. No recent studies of the protective
role of calcium in Pb-induced carcinogenesis or genotoxicity were found. There were
some data suggesting that boron, melatonin, N-acetylcysteine, turmeric and myrrh protect
cells against Pb-induced genotoxicity (Section 4.10.3.2).
A recent study details Pb and Se interactions in virus-dependent carcinogenesis in
laboratory animals. Schrauzer (2008) considered the impact of Se on carcinogenesis by
studying four groups of weanling virgin female C3H/St mice infected with murine
mammary tumor virus (groups of 20-30 mice), which induces mammary tumor
formation. One set of two groups were fed a diet containing 0.15 ppm Se and then were
exposed via drinking water to acetic acid (control group) or 0.5 ppm Pb acetate. The
second set of two groups were fed a diet containing 0.65 ppm Se and then similarly
exposed to acetic acid or 0.5 ppm Pb acetate. The study was primarily focused on the
general effects of a low Se diet. The data suggest that Se is anticarcinogenic as in the
groups without Pb exposure, the animals exposed to the higher Se levels had fewer
mammary tumors and these tumors had a delayed onset of appearance. Pb exposure with
low Se caused the same delayed onset as did the higher dose of Se and also caused some
reduction in the tumor frequency. Pb exposure with higher Se increased the tumor
frequency and the onset of the tumors. Pb also induced weight loss at 14 months in both
exposed groups. The data suggest that there may be interactions of Pb and Se, but they
suggest that Pb mimics or antagonizes Se. They do not suggest that Se is protective of
Pb-induced toxicity or carcinogenesis.
In summary, the new data on Pb exposure as part of a mixture are derived from studies
designed with co-exposure to metals or antioxidants. Children in Mexico living near a Pb
smelter with co-exposure to high levels of Pb and As showed elevated DNA damage and
impaired DNA repair but no conclusions were able to be made regarding whether this
was a result of exposure to individual metals or a combination of metals. Pb and Cd
co-exposure in mice elevated SCE levels but did not further exacerbate MN levels above
Pb exposure alone. Primary lung cells exposed to a metals mixture showed an interaction
at the mRNA level among the three metals tested. In other genotoxicity assays, various
antioxidants (melatonin, NAC, turmeric and myrrh) and metals (boron) were protective
against Pb-induced genotoxicity. In an animal model of breast cancer, Se modified the
onset and multiplicity of murine mammary tumor virus-induced tumorogenicity in
Pb-exposed animals. These data show that co-exposure of Pb with antioxidants or metals,
modifies the effect of Pb on DNA damage, DNA repair, mutagenicity, genotoxicity, or
tumorogenicity.
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4.10.5 Summary and Causal Determination
Toxicological and epidemiologic studies of the association between Pb exposure and
cancer and cancer-related outcomes have been reviewed in the preceding sections.
Evaluation of the relationship between Pb exposure and cancer with respect to causality
was based primarily on evidence for tumor incidence in experimental animals and
evidence describing potential modes of action including mutagenesis, DNA aberrations,
and epigenetic changes. Supporting evidence was provided from associations of Pb
exposure with cancer incidence and mortality in humans. The application of the key
supporting evidence from these studies to the causal framework is summarized in Table
4-50 and the following text.
The toxicological literature provides consistent evidence of the carcinogenic potential of
Pb and possible contributing modes of action, including genotoxic, mutagenic and
epigenetic effects. In laboratory studies, chronic Pb exposure for 18 months or two years
to high concentrations such as 10,000 ppm Pb acetate in diet or 2,600 ppm Pb acetate in
drinking water has been demonstrated to be an animal carcinogen. Chronic Pb exposure
to male and female rodents has consistently induced kidney and brain carcinogenesis in
multiple separate studies, inducing various tumors, (i.e., adenocarcinomas, adenomas,
and gliomas. Pb has also been shown to cause mammary gland, prostate, adrenal, and
testicular tumors in animals. Developmental Pb acetate exposure also induced tumors in
offspring whose dams received Pb acetate in drinking water during pregnancy and
lactation. Multiple toxicological studies showed neoplastic transformation in cultured
cells providing an additional potential mode of action, but most used Pb chromate, and it
is possible that the chromate ion contributed to these findings. The toxicological and
epidemiologic literature provides evidence for potential carcinogenic modes of action
from genotoxic, mutagenic and epigenetic assays.
Multiple longitudinal epidemiologic studies have been performed examining the
association between cancer incidence and mortality and Pb exposures, estimated with
biological measures and exposure databases. Mixed results have been reported for cancer
mortality studies; a large NHANES epidemiologic study demonstrated a positive
association between baseline blood Pb and cancer mortality with median 8.6 years of
follow up on subjects (Schober et al.. 2006). but the other studies reported null results
(Khalil et al.. 2009b: Weisskopf etal.. 2009; Menke et al.. 2006). These were well-
conducted epidemiologic studies with control for important potential confounders such as
age, smoking, and education. Although the 2006 Pb AQCD (U.S. EPA. 2006b) reported
some studies that found an association between Pb exposure indicators and lung cancer,
recent studies mostly included occupationally-exposed adults and observed no
associations. Most studies of Pb and brain cancer were null among the overall study
4-779
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population, but positive associations were observed among individuals with certain
genetic variants. However, the studies of Pb and brain cancer were all limited due to
possible confounding by several factors, including other workplace exposures. A limited
amount of research has been performed on other types of cancer. The 2006 Pb AQCD
reported evidence that suggested an association between Pb exposure and stomach
cancer, but in a recent study of stomach cancer the results were inconsistent, reporting a
positive association between organic Pb exposure and stomach cancer but null findings
for exposure to inorganic Pb or Pb from gasoline emissions and stomach cancer.
Among epidemiologic studies, high Pb levels (over 40 (ig/dL in adults) were associated
with SCEs among adults. This association was not observed among children (mean
concurrent blood Pb 7.69 (ig/dL). Other epidemiologic studies of DNA damage reported
inconsistent results. Consistent with previous toxicological findings, Pb does appear to
have genotoxic activity in animal and in vitro models, inducing SCE, MN and DNA
strand breaks. The majority of the chromosomal aberration studies with Pb-induced
significant findings used Pb chromate exposure and the aberrations are likely due to the
chromate. Pb does not appear to be strongly mutagenic as the HPRT assays were
typically negative unless a cell signaling pathway was disturbed.
Mechanistic understanding of the carcinogenicity of Pb in toxicological models is
expanding with work on the antioxidant Se and metallothionein, a protein that binds Pb
and reduces its bioavailability. Low Se diet affects tumorigenesis and tumor multiplicity
with Pb exposure. Metallothionein has been shown to be protective against the effect of
Pb on carcinogenicity. Pb is clastogenic and mutagenic in some but not all models.
Clastogenicity and mutagenicity may be possible mechanisms contributing to cancer but
are not absolutely associated with the induction of cancer. Because Pb has a higher
atomic weight than does Zn, Pb replaces Zn at many Zn binding sites or Zn finger
proteins. This substitution has the potential to induce effects that can indirectly contribute
to carcinogenicity via interactions with hormone receptors, cell-cycle regulatory proteins,
tumor suppressor genes like p53, DNA repair enzymes, histones, etc. These indirect
effects may act at a post-translational level to negatively alter protein structure and DNA
repair.
Epigenetic changes associated with Pb exposure or biological markers of Pb exposure are
beginning to appear in the literature. Epigenetic modifications may contribute to
carcinogenicity by altering DNA repair or changing the expression of a tumor suppressor
gene or oncogene. A small number of epidemiologic studies examining Pb and global
epigenetic changes demonstrated an inverse association between Pb concentrations and
LINE-1 or Alu methylation. Lower DNA methylation is associated with increased gene
expression, but epigenetic contributions to cancer are not yet fully characterized in this
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emerging area of research. Toxicological studies show that Pb can activate or interfere
with a number of signaling and repair pathways, though it is unclear whether these are
due to epigenetic responses or genotoxicity. Thus, an underlying mechanism is still
uncertain, but likely involves genomic instability, epigenetic modifications, or both.
In conclusion, the toxicological literature provides the strong evidence for long-term
exposure (i.e., 18 months or 2 years) to high concentrations of Pb (> 2,600 ppm) and
cancer. The consistent evidence indicating Pb-induced carcinogenicity in animal models
is substantiated by the mode of action findings from multiple high-quality toxicological
studies in animal and in vitro models from different laboratories. This is substantiated by
the findings of other agencies including IARC, which has classified inorganic Pb
compounds as a probable human carcinogen and the National Toxicology Program,
which has listed Pb and Pb compounds as "reasonably anticipated to be human
carcinogens." Strong evidence from animal toxicological studies demonstrates an
association between Pb and cancer, genotoxicity or epigenetic modification.
Carcinogenicity in animal toxicology studies with relevant routes of Pb exposure has
been reported in the kidneys, testes, brain, adrenals, prostate, pituitary, and mammary
gland, albeit at high doses of Pb. Based on such evidence, IARC has classified
inorganic Pb compounds as a probable human carcinogen and the National Toxicology
Program (NTP) has listed Pb and Pb compounds as "reasonably anticipated to be human
carcinogens." Epidemiologic studies of cancer incidence and mortality reported
inconsistent results; one strong epidemiologic study demonstrated an association between
blood Pb and increased cancer mortality, but the other studies reported weak or no
associations. In the 2006 Pb AQCD, indicators of Pb exposure were found to be
associated with stomach cancer, and a recent study on stomach cancer and occupational
Pb exposure reported mixed findings depending on the type of Pb exposure (organic Pb,
inorganic Pb, or Pb from gasoline emissions). Similarly, some studies in the
2006 Pb AQCD reported associations between Pb exposure indicators and lung cancer.
Recent epidemiologic studies of lung cancer focused on occupational exposures and
reported inconsistent associations. The majority of epidemiologic studies of brain cancer
had null results overall, but positive associations between Pb exposure indicators and
brain cancer were observed among individuals with certain genotypes. Overall, the
consistent and strong body of evidence from toxicological studies on tumor incidence and
potential modes of action but inconsistent epidemiologic evidence is sufficient to
conclude that a causal relationship is likely to exist between Pb exposure and cancer.
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Table 4-50 Summary of evidence supporting cancer and genotoxicity
causal determinations.
Attribute in
Causal
Framework3 Key Evidence13
Pb Exposure or
Blood Levels
Associated
References'3 with Effects
Cancer - Likely Causal
Consistent Consistent findings across multiple
evidence from toxicology studies using 18 month
multiple animal or two year cancer bioassays in
studies with rats wherein rodents are fed chow
chronic Pb or received drinking water
exposure enriched with Pb acetate and
show tumor development.
Gestational and lactational Pb
exposure induced carcinogenicity
in adult offspring.
Azar et al. (1973).
Kasprzak et al. (1985).
Koller et al. (1985).
Van Esch and Kroes (1969)
Waalkesetal., (1995)
Section 4.10.1.7
Chronic 10,000 ppm
Pb acetate diet or
2,600 ppm drinking
water Pb acetate, no
blood Pb measurement
available.
500, 750 and
1,000 ppm Pb in
drinking water, no
blood Pb measurement
available.
Most evidence Consistent toxicological evidence
clearly for mutagenicity, carcinogenicity,
supports mode and genotoxicity of Pb reported by
of action multiple laboratories in humans,
animals and in vitro models using
multiple assays (MN, SCE,
COMET).
Some evidence for epigenetic
changes. Bone Pb levels were
inversely associated with LINE-1
methylation in a study of adult
men.
Study showed inverse association
between maternal postpartum
bone Pb levels and Alu and
LINE-1 methylation in cord blood.
Occupational battery workers had
ALAD hypermethylation compared
with controls; cell culture study of
high dose Pb exposure caused
ALAD hypermethylation.
Toxicological Some toxicological studies employ
evidence of Pb chromate when investigating
clastogenic, the clastogenic, mutagenic, and
mutagenic, genotoxic effects of Pb. The effect
and genotoxic of the chromate ion in contributing
effects with Pb to these effects cannot be ruled
chromate out.
Epidemiology evidence of DMA
and cellular damage:
See Section 4.10.3.2
Toxicology evidence of DMA and
cellular damage:
See Section 4.10.3.2
See Section 4.10.3.3
See Section 4.10.3.2
4-782
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Table 4-50 (Continued): Summary of evidence supporting cancer and genotoxicity
causal determinations.
Attribute in
Causal
Framework3
Epidemiologic
evidence is
limited and
inconsistent
Key Evidence13
Epidemiologic studies of overall
cancer mortality have inconsistent
findings. These are high-quality,
longitudinal studies with an
average follow-up of at least 8.6
years and control for potential
confounders, such as age,
smoking, and education. There is
uncertainty related to exposure
patterns resulting in likely higher
past Pb exposure.
Epidemiologic studies of specific
References'3
Menke et al. (2006)
Schober et al. (2006)
Weisskopf et al. (2009),
Khalil (2009b)
Overall Cancer Mortality:
See Section 4.10.1.1
Pb Exposure or
Blood Levels
Associated
with Effects
In the mortality studies,
the majority of study
participants' baseline
blood Pb levels were
<10ug/dL (those that
reported means ranged
from 2.58 to 5.6 ug/dl_).
In studies of specific
cancer types, exposure
measurements varied
(e.g. report of ever
exposure, work history,
blood Pb levels).
studies were not consistent with
previous findings of possible
associations for lung and stomach
cancers reported in the
2006PbAQCD. Many of the
epidemiologic studies examining
specific cancer sites were case-
control studies and not all included
potentially important confounders,
such as smoking. There is
uncertainty related to exposure
patterns resulting in likely higher
past Pb exposure.
Specific Cancer:
See Sections:
(4.10.1.3. Lung);
(4.10.1.4. Brain);
(4.10.1.5 Breast);
(4.10.1.6 Other cancers)
"Described in detail in Table II of the Preamble.
bDescribes the key evidence and references contributing most heavily to causal determination and where applicable to uncertainties
or inconsistencies. References to earlier sections indicate where full body of evidence is described.
4-783
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CHAPTER 5 POTENTIALLY AT-RISK
POPULATIONS
Introduction
The NAAQS are intended to protect public health with an adequate margin of safety. In
so doing, protection is provided for both the population as a whole and those groups
potentially at increased risk for health effects from exposure to the air pollutant for which
each NAAQS is set (Preface to this ISA). To facilitate the identification of populations at
increased risk for Pb-related health effects, studies have evaluated various factors that
may contribute to susceptibility and/or vulnerability to Pb. The definitions of
susceptibility and vulnerability vary across studies, but in most instances "susceptibility"
refers to biological or intrinsic factors (e.g., age, sex) while "vulnerability" refers to
nonbiological or extrinsic factors (e.g., socioeconomic status [SES]) (U.S. EPA. 2010.
2009a). Additionally, in some cases, the terms "at-risk" and "sensitive" populations have
been used to encompass these concepts more generally. In this ISA, "at-risk" groups are
defined as those with characteristics that increase the risk of Pb-related health effects in a
population. These characteristics include various factors, such as genetic background,
race and ethnicity, sex, age, diet, pre-existing disease, SES, and characteristics that may
modify exposure or the response to Pb.
Individuals, and ultimately populations, could experience increased risk for air pollutant
induced health effects via multiple avenues. A group with intrinsically increased risk
would have one or more factors that increase risk for an effect through a biological
mechanism. In general, people in this category would have a steeper concentration-
risk relationship or would manifest effects at a lower exposure, compared to those not
in the category. Potential factors that are often considered intrinsic include genetic
background and sex. A group of people could also have extrinsically increased risk,
which would be through an external, non-biological factor. Examples of extrinsic
factors include SES and diet. Some groups are at risk of increased internal dose at a
given exposure concentration. In addition, some groups could have greater exposure
(concentration x time), regardless of the delivered dose. Finally, there are those who
might be placed at increased risk for experiencing a greater exposure by being
exposed at a higher concentration. An example of this is people living near Pb
smelters.
Some factors described above are multifaceted and may influence the risk of an air
pollutant related health effect through a combination of avenues. For example, SES may
5-1
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affect access to medical care, which itself may contribute to the presence of preexisting
diseases and conditions considered as intrinsic factors. Additionally, children's outdoor
activities can lead to more hand-to-mouth contact with contaminated soil than adults,
which leads to increased intake dose and exposure. At the same time, children have
biological (i.e., intrinsic) differences from adults that may influence their uptake,
metabolism, storage, and excretion.
The emphasis of this chapter is to identify and understand the factors that potentially
increase the risk of Pb-related health effects, regardless of whether the increased risk is
due to intrinsic factors, extrinsic factors, increased dose/exposure, or a combination due
to the often interconnectedness of factors. The following sections examine factors that
potentially lead to increased risk of Pb-related health effects and characterize the overall
weight of evidence for each factor.
Approach to Classifying Potential At-Risk Factors
To identify factors that potentially lead to some populations being at greater risk of
Pb-related health effects, the evidence across relevant scientific disciplines (i.e., exposure
sciences, dosimetry, toxicology, and epidemiology) was evaluated. In this systematic
approach, the collective evidence is used to examine coherence of effects across
disciplines and determine biological plausibility. The collective results across the
scientific disciplines comprise the overall weight of evidence that is used to determine
whether a specific factor results in a population being at increased risk of an air pollutant
related health effect. The first section of this chapter (Section 5.1) summarizes
physiological factors that influence Pb levels in the body. The second section of this
chapter (Section 5.2) summarizes information on factors potentially related to differential
Pb exposure. The studies presented in this section supplement the material provided in
Chapter 2 and Chapter 3 by examining how factors such as age, sex, race and ethnicity,
SES, proximity to Pb sources, and residential factors may affect Pb exposure or blood Pb
levels. The third section of this chapter (Section 5.3) discusses the epidemiologic and
toxicological studies evaluated in Chapter 4 that provide information on factors
potentially related to increased risk of Pb-induced health effects through the use of
analysis targeting specific factors. To examine whether Pb differentially affects certain
populations, the epidemiologic studies discussed in Section 5.3 conduct stratified
analyses to identify the presence or absence of effect measure modification. A thorough
evaluation of potential effect measure modifiers may help identify populations that are at
increased risk for Pb-related health effects. Highlighted studies include only those where
the population was stratified into subgroups (e.g., males versus females or smokers
versus nonsmokers) for comparative analysis. In the case of many biomarker studies and
5-2
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the epidemiologic studies considered, this approach allowed for a comparison between
populations exposed to similar Pb concentrations and within the same study design.
Toxicological studies also provide evidence of Pb effects and biological plausibility for
factors that may lead to increased risk for Pb-related health effects. Included
toxicological studies discussed in Section 5.3 may have categorized the study populations
by different factors, such as age, sex, diet/nutrition status, and genetics, or are those that
examined animal models of disease. These epidemiologic and toxicological studies
provide the scientific basis for an overall weight of the evidence evaluation for the
identification of specific populations potentially at risk of Pb-related health effects.
Details on the magnitude of effects for studies in this third section (Section 5.3) are
included in summaries of the studies presented in Chapter 4.
Numerous studies that focused on only one potentially at-risk population were described
in previous chapters (Chapter 4) but are not discussed in detail in this chapter because
they lacked stratified analysis with adequate comparison groups. For example, pregnancy
is a lifestage with potentially increased risk for mothers and fetuses, but because there are
no comparison groups for stratified analyses, these studies were presented in Chapter 4
but are not included here. Additionally, it is understood that some of the stratified
variables/factors discussed in this third section (Section 5.3) may not be effect measure
modifiers but instead may be mediators of Pb-related health effects. Mediators are factors
that fall on the causal pathway between Pb exposure and health outcomes, whereas effect
measure modifiers are factors that result in changes in the measured associations between
Pb exposure and health effects. Because mediators are caused by Pb exposure and are
also intermediates in the disease pathway that is studied, mediators are not correctly
termed "at-risk" factors. Some of the factors discussed in this third section could be
mediators and/or modifiers. These are noted in Table 5-5.
Building on the causal framework discussed in detail in the Preamble and used
throughout the ISA, conclusions are made regarding the strength of evidence for each
factor that may contribute to increased risk of a Pb-related health effect based on the
evaluation and synthesis of evidence across scientific disciplines. The conclusions drawn
considered the "Aspects to Aid in Judging Causality" discussed in Table I of the
Preamble. The categories considered for evaluating the potential increased risk of an air
pollutant-related health effect are "adequate evidence," "suggestive evidence,"
"inadequate evidence," and "evidence of no effect." They are described in more detail in
Table 5-1.
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Table 5-1 Classification of evidence for potential at-risk factors.
Health Effects
Adequate
evidence
Suggestive
evidence
Inadequate
evidence
Evidence of
no effect
There is substantial, consistent evidence within a discipline to conclude that a factor results in a
population or lifestage being at increased risk of air pollutant-related health effect(s) relative to
some reference population or lifestage. Where applicable this includes coherence across
disciplines. Evidence includes multiple high-quality studies.
The collective evidence suggests that a factor results in a population or lifestage being at
increased risk of an air pollutant-related health effect relative to some reference population or
lifestage, but the evidence is limited due to some inconsistency within a discipline or, where
applicable, a lack of coherence across disciplines.
The collective evidence is inadequate to determine if a factor results in a population or lifestage
being at increased risk of an air pollutant-related health effect relative to some reference
population or lifestage. The available studies are of insufficient quantity, quality, consistency,
and/or statistical power to permit a conclusion to be drawn.
There is substantial, consistent evidence within a discipline to conclude that a factor does not
result in a population or lifestage being at increased risk of air pollutant-related health effect(s)
relative to some reference population or lifestage. Where applicable this includes coherence
across disciplines. Evidence includes multiple high-quality studies.
5.1 Physiological Factors that Influence the Internal
Distribution of Pb
Blood and bone Pb measures are influenced to varying degrees by biokinetic processes
(e.g., absorption, distribution, metabolism, excretion), which are discussed in detail in
Chapter 3. These processes can be affected by multiple factors, such as age, genetics,
diet, and co-exposure with other metals and non-metals.
Age influences the biokinetic response to Pb within the body. Infants may be considered
an at-risk population because Pb easily crosses the placental barrier and accumulates in
fetal tissue during gestation (Pillai et al.. 2009; Wang et al.. 2009e: Uzbekov et al.. 2007).
This transfer of Pb from mother to fetus is partly due to the remobilization of the
mother's bone stores (O'Flahertv. 1998; Franklin et al.. 1997). This also results in
increased maternal blood Pb levels (Lamadrid-Figueroa et al.. 2006; Gulson et al.. 2004a;
Hertz-Picciotto et al.. 2000; Gulson et al.. 1997; Lagerkvist et al.. 1996; Schuhmacher et
al.. 1996; Rothenberg et al., 1994a). Infants also absorb Pb more efficiently from their
gastrointestinal tracks than older children and adults (Mushak. 1991; Ziegler et al.. 1978).
Bone growth rate is high during childhood. The majority of a child's Pb body burden is
not permanently incorporated in the bone, but some Pb does remain in the bone until
older age (McNeill et al.. 2000; O'Flaherty. 1995; Leggett 1993). Older adults are more
5-4
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likely to have age-related degeneration of bones and organ systems and a possible
redistribution of Pb stored in the bones into the blood stream (Popovic et al.. 2005;
Garrido Latorre et al., 2003; Gulson et al., 2002).
Various genes can also affect Pb biomarker concentrations. Genetic variants of the
vitamin D receptor (VDR) in humans have been associated with varied bone and plasma
Pb levels (Rezende et al.. 2008; Theppeang et al.. 2004; Schwartz et al.. 2000a). Multiple
studies have also examined the association between the aminolevulinate dehydratase
(ALAD) polymorphism and blood Pb levels and found that the ALAD-2 polymorphism
may be biologically related to varying Pb levels, although some studies report no
difference for ALAD alleles (see also Section 4.2.3) (Miyaki et al.. 2009; Shaik and
Jamil. 2009; Sobin et al.. 2009; Chen et al.. 2008c: Rabstein et al.. 2008; Scinicariello et
al.. 2007: Zhao et al.. 2007: Montenegro et al.. 2006: Wananukul et al.. 2006).
It is well established that diets sufficient in minerals such as calcium (Ca), iron (Fe), and
zinc (Zn) offer some protection from Pb exposure by preventing or competing with Pb for
absorption in the GI tract. A study in China reported that children who regularly
consumed breakfast had lower blood Pb levels than those children that did not eat
breakfast after controlling for confounders such as urban-rural location of the school,
maternal education level and occupation, and paternal occupation (LiuetaL 201 la).
Diets designed to limit or reduce caloric intake and induce weight loss have been
associated with increased blood Pb levels in adult rats (Han et al.. 1999). A toxicological
study reported negative effects of Pb on osmotic fragility, TEARS production, catalase
activity, and other oxidative parameters, but most of these effects were reduced to the
levels observed in the control group when the rats were given supplementation of zinc
and vitamins (Masso-Gonzalez and Antonio-Garcia. 2009). Toxicological studies by
Jamieson et al. (2008: 2006) also reported that a zinc-deficient diet increases bone and
renal Pb content and impairs skeletal growth and mineralization in rats. A zinc-
supplemented diet attenuated bone and renal Pb content. Toxicological studies among
rats and chicks have shown that dietary deficiency of calcium induces increased Pb
absorption and retention (Mykkanen and Wasserman. 1981: Six and Gover. 1970).
Increased calcium intake reduces accumulation of Pb in bone and mobilization of Pb
during pregnancy and lactation in rats (Bogden et al.. 1995). Additionally, human studies
have reported that Fe deficiencies may result in higher Pb absorption or altered
biokinetics (Schell et al.. 2004: Marcus and Schwartz. 1987: Mahaffev and Annest.
1986).
In summary, age, genetics, and diet affect the biokinetics of Pb, which in turn affects the
internal distribution of Pb. These factors were discussed in greater detail in Chapter 3
5-5
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where more information on overall biokinetic and physiological factors affecting Pb
distribution is provided.
5.2 Population Characteristics Potentially Related to Differential
Pb Exposure
Elevated or differential Pb exposure and related biomarker levels (such as blood Pb),
have been shown to be statistically related to several population characteristics, including
age, sex, race and ethnicity, SES, proximity to Pb sources, and residential factors (U.S.
EPA. 2006b). In most cases, exposure, absorption, and biokinetics of Pb are all
influenced to varying degrees by such characteristics. Additionally, the relative
importance of such population characteristics in affecting exposure, absorption, and
biokinetics varies among individuals in a population and is difficult to quantify. This
section presents recent studies demonstrating a relationship between each population
characteristic and exposure status. The studies presented in this section build upon the
current body of literature suggesting that population characteristics differentially
influence Pb exposure; the new literature does not alter previous understanding of the
differential influence of population characteristics on Pb exposure. Differential response
to given Pb exposures is discussed in Section 5.3.
5.2.1 Age
5.2.1.1 Early Childhood
Typically, children have higher exposure to Pb compared with adults because children's
behaviors and activities include hand-to-mouth contact, crawling, and poor hand-washing
that typically result in greater Pb ingestion compared with adults (U.S. EPA. 2006b).
Children can also have increased Pb exposure because outdoor activities can lead to
hand-to-mouth contact with contaminated soil. For example, Zahran et al. (2010)
observed that a 1% reduction in soil Pb concentration led to a 1.55 (ig/dL reduction in
median blood Pb levels (p <0.05) among New Orleans children.
Age of the children may influence blood Pb levels through a combination of behavioral
and biokinetic factors. The 2009-2010 NHANES data are presented in Table 5-2 by age
and sex. Among children, highest blood Pb levels occurred in the 1-5 year age group
(children under age 1 were not included), and within this subgroup (not shown on the
table), 1-year old children had the highest blood Pb levels (99th percentile: 9.47 ug/dL)
5-6
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(CDC. 20lib) (CDC. 2013). It is possible that high blood Pb levels among these young
children may also be related to in utero exposures resulting from maternal Pb
remobilization from bone stores from historic exposures (Miranda et al., 2010) or from
contemporaneous Pb exposures if the mothers had appreciable current Pb exposure.
Jones et al. (2009a) analyzed the NHANES datasets for the years 1999-2004 to study
trends in blood Pb among two different age groups of children over time (see Table 5-3).
They observed greater percentages of children aged 1-2 years having blood Pb levels of
2.5 to <5 (ig/dL, 5 to <7.5 (ig/dL, and > 10 (ig/dL, compared with 3-5 year-old children,
but no age difference was noted for the 7.5 to <10 (ig/dL bracket. At the same time, 1-2
year-old children had lower percentages of blood Pb levels <1 (ig/dL and 1 to <2.5 (ig/dL
compared with children ages 3-5 years old. This implies that there is a shift in the
distribution of blood Pb levels as children age, even during early childhood. These
distribution differences may be attributable to differences in exposure (including
behavioral influences, such as hand-to-mouth contact and crawling in proximity to
contaminated surfaces), residual contributions from the mother's Pb burden, age-
dependent variability in biokinetics or diet (e.g., milk versus solid diets). Yapici et al.
(2006) studied the relationship between blood Pb level and age among a cohort of
children between 6 and 73 months of age with elevated blood Pb levels (87.6% of study
group with blood Pb greater than 20 (ig/dL) living near a Turkish coal mine. They
observed a low but statistically significant negative correlation between blood Pb and age
(r = -0.38,p<0.001).
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Table 5-2 Blood Pb levels (|jg/dL) by
Age
1-5yr
6-1 1 yr
12-1 9 yr
20-59 yr
60+ yr
Overall
Sex
Total
Male
Female
Total
Male
Female
Total
Male
Female
Total
Male
Female
Total
Male
Female
Total
Male
Female
N
836
429
407
1009
521
488
1183
632
551
3856
1843
2013
1909
941
968
8793
4366
4427
Avg.
1.61
1.59
1.64
1.05
1.10
0.99
0.84
0.98
0.69
1.50
1.88
1.15
2.09
2.46
1.73
1.50
1.75
1.25
Std. Dev.
1.49
1.32
1.65
0.74
0.73
0.75
0.68
0.69
0.62
1.83
2.33
1.10
1.51
1.78
1.07
1.57
1.88
1.13
age and sex, 2009-2010 NHANES.
5%
0.53
0.51
0.54
0.42
0.45
0.38
0.33
0.40
0.30
0.44
0.56
0.40
0.72
0.87
0.65
0.43
0.50
0.39
25%
0.85
0.83
0.90
0.61
0.66
0.58
0.50
0.58
0.44
0.72
0.92
0.61
1.16
1.39
1.01
0.72
0.84
0.63
50%
1.21
1.22
1.20
0.83
0.88
0.79
0.69
0.80
0.57
1.08
1.37
0.89
1.69
1.99
1.43
1.10
1.29
0.96
75%
1.81
1.84
1.77
1.22
1.30
1.12
0.96
1.11
0.79
1.70
2.12
1.35
2.53
2.90
2.14
1.76
2.05
1.48
95%
4.00
4.09
3.69
2.36
2.37
2.35
1.82
2.09
1.31
3.53
4.49
2.63
4.79
5.56
3.75
3.66
4.31
2.97
99%
8.03
7.49
9.59
4.29
4.18
3.98
3.10
3.91
2.25
7.27
9.68
4.41
8.28
9.89
5.42
7.21
8.62
5.17
Source: (CDC, 2013).
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Table 5-3 Percentage of children within six categories/brackets of blood Pb
levels, 1999-2004 NHANES.
N
Overall 2,532
Geometric
mean3
1.9
(1.8-2.0)
Percentage (%) of children within categories/brackets (95% CIs)
<1
jjg/dL
14.0
(11.6-16.6)
1 to <2.5
jjg/dL
55.0
(52.1-57.9)
2.5 to <5
jjg/dL
23.6
(21.1-26.1)
5 to <7.5
jjg/dL
4.5
(3.3-5.9)
7.5 to <10
jjg/dL
1.5
(1.0-2.1)
>10
jjg/dL
1.4
(1.0-2.0)
Sex
Female 1,211
Male 1,321
1.9
(1.7-2.0)
1.9
(1.7-2.0)
14.1
(10.8-17.7)
14.0
(11.4-16.7)
54.5
(51.1-57.8)
55.5
(51.4-59.5)
23.9
(20.3-27.8)
23.2
(20.3-26.3)
4.5
(3.3-5.8)
4.6
(3.0-6.5)
1.4
(0.8-2.3)
1.5
(0.9-2.3)
1.7
(0.9-2.6)
1.3
(0.7-2.6)
Age
1-2yr 1,231
3-5 yr 1,301
2.1
(2.0-2.2)
1.7
(1.6-1.9)
10.6
(7.7-13.9)
16.2
(12.9-19.9)
51.0
(46.7-55.3)
57.6
(53.8-61.4)
27.9
(24.9-31.0)
20.7
(17.9-23.7)
6.7
(5.0-8.6)
3.1
(1.9-4.6)
1.4
(0.8-2.2)
1.5
(0.8-2.3)
2.4
(1.4-3.5)
0.9
(0.4-1.5)
Race/Ethnicity
Non-
Hispanic 755
Black
Mexican Q12
American
Non-
Hispanic 731
White
Poverty-Income Ratio
<1.3 1,302
>1.3 1,070
2.8
(2.5-3.0)
1.9
(1.7-2.0)
1.7
(1.6-1.8)
4.0 (2.5-5.7)
10.9
(8.6-13.4)
17.6
(14.0-21.5)
(PIR)
2.4
(2.2-2.5)
1.5
(1.4-1.6)
6.7 (4.6-9.2)
19.9
(16.3-23.8)
42.5
(37.8-47.2)
61.0
(56.9-65.1)
57.1
(52.4-61.7)
49.3
(44.9-53.7)
60.4
(56.9-63.8)
36.2
(33.1-39.3)
22.1
(18.0-26.5)
19.7
(16.1-23.5)
32.5
(28.6-36.4)
16.0
(12.9-19.3)
9.4
(6.9-12.2)
3.4
(2.2-5.0)
3.6
(1.9-5.8)
6.9
(2.2-8.8)
2.3
(1.2-3.7)
4.6
(3.0-6.5)
1.3
(0.6-2.2)
0.8
(0.3-1.6)
2.8
(1.7-4.1)
0.6
(0.1-1.4)
3.4
(1.8-5.5)
1.2
(0.4-2.6)
1.2
(0.6-2.0)
1.8
(1.1-2.7)
0.8
(0.3-1.6)
"Geometric mean Pb Units: |jg/dL (95% Cl)
Source: Reprinted with permission of the American Academy of Pediatrics; Jones et al. (2009a)
Fetal and child Pb biomarkers have been demonstrated to relate to maternal Pb
biomarkers as reported in the 2006 Pb AQCD (U.S. EPA. 2006b). Kordas et al. (2010)
observed that maternal hair Pb concentration was a statistically significant predictor of
child hair Pb concentration ((3 = 0.37 ± 0.07, p <0.01). Elevated blood Pb levels among
mothers present a potential exposure route to their children in utero or through breast
milk; see Miranda et al. (2010).
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5.2.1.2 Older Adulthood
Blood Pb levels tend to be higher in older adults compared with the general adult
population, as described in the 2006 Pb AQCD (U.S. EPA. 2006b). Table 5-2 presents
2009-2010 NHANES data broken down by age group and shows that blood Pb levels
were highest in the among participants 60 years old or older, in comparison with adults
aged 20-59 years and with adolescents. In a study of blood Pb and saliva Pb in a mostly
female population in Detroit, Nriagu et al. (2006) found that age was a statistically
significant positive predictor of blood Pb (p O.001). Average blood Pb levels among 14-
to 24-year-old subjects was 2.60 ± 0.16 ug/dL compared with 4.29 ± 0.56 ug/dL among
subjects aged 55 years or older. Higher average and median levels among older adults
could potentially be due to a shared experience of higher historical Pb exposures stored in
bone in conjunction with remobilization of stored Pb during bone loss (Section 3.2).
Theppeang et al. (2008b) studied Pb concentrations in the blood, tibia, and patella of
subjects age 50-70 as part of the Baltimore Memory Study. They found a statistically
significant relationship between age and tibia Pb ((3 = 0.37, p <0.01 in a model including
age, race/ethnicity, Yale energy index, and 2 diet variables; (3 = 0.57, p <0.01 in a model
including age, sex, and an interaction term for sex and age, which was also statistically
significant at p = 0.03). Theppeang et al. (2008b) also noted that patella Pb
concentrations were also positively associated with age, although the authors did not
present the data or significance levels. A statistically significant relationship was not
observed between the log-transform of blood Pb and age ((3 = 0.007, p = 0.11), although
the age range of subjects may not have been sufficient to discern a difference in blood Pb
level.
Miranda et al. (2010) observed that older pregnant women (ages 30-34 years and 35-39
years) had statistically significant higher odds of having greater blood Pb levels than
younger pregnant women (25- to 29-year-olds) in the reference age category. These
results could be related to a historical component to Pb exposure among mothers. These
findings were also consistent with observations that Pb storage in bones increased with
age before subsequent release with bone loss occurring during pregnancy, as described in
Section 3.2 and summarized in Section 5.1.
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5.2.2 Sex
The AQCD (U.S. EPA. 2006b) described several studies showing higher Pb biomarker
levels in male adults compared with female adults. The 2009-2010 NHANES showed
that overall, males have significantly higher blood Pb levels (average: 1.75 ug/dL) than
females (average: 1.25 ug/dL) (p <0.0005). Among adults aged 20-59 years, average
blood Pb levels were 64% higher for males compared with females (p <0.0005). Among
adults 60 years or older, average blood Pb levels were 30% higher for males compared
with females (p <0.0005) (NCHS. 2010). In their study of Pb burden among Baltimore
adults aged 50-70 years, Theppeang et al. (2008b) observed that average blood Pb levels
were statistically significantly higher (p <0.01) among men (4.4 (ig/dL) than women
(3.1 (ig/dL). For average tibia Pb levels, Theppeang et al. (2008b) noted no difference
(p = 0.12) between men (18.0 (ig/g) and women (19.4 (ig/g).
Among U.S. children, the 2009-2010 NHANES data showed that blood Pb levels were
higher among girls than boys for the 1- to 5-years age group (Table 5-2). Blood Pb levels
became slightly higher among boys for the 6- to 11-years age group, and levels were
substantially higher among adolescent males than females 12- to 19-years old. The
2009-2010 NHANES data suggest that sex-based differences in blood Pb levels are not
substantial until adolescence.
5.2.3 Race and Ethnicity
Higher blood Pb and bone Pb levels among African Americans have been well
documented (U.S. EPA. 2006b). Structural equation model results by Lanphear and
Roghmann (1997) demonstrated not just elevations in blood Pb among African
Americans but also significant associations between blood Pb and race. Recent studies
are consistent with those previous findings. For instance, Levin et al. (2008) and Jones et
al. (2009a) both analyzed NHANES survey data to examine trends in childhood blood Pb
levels. Data from the Jones et al. (2009a) study, using NHANES data (NCHS. 2009.
2008) from 1988-1991 and 1999-2004 are shown in Figure 5-1. The authors found that
differences among children from different racial/ethnic groups with regard to the
percentage with blood Pb levels > 2.5 ug/dL over the period 1999-2004, have decreased
since the period of 1988-1991. The non-Hispanic black group still had higher percentages
with blood Pb levels > 2.5 ug/dL compared with non-Hispanic whites and Mexican
Americans, with large observable differences for blood Pb levels between 2.5 and
<10 ug/dL. It is notable that the distributions of blood Pb levels among Mexican
American and non-Hispanic white children were nearly identical in the 1999-2004
dataset. Theppeang et al. (2008b) also explored the effect of race and ethnicity on several
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Pb biomarkers in a study of older adults living in Baltimore, MD. They observed a
statistically significant difference between African American (AA) and Caucasian (C)
subjects with respect to tibia Pb (AA: 21.8 (ig/g, C: 16.7 (ig/g, p <0.01) but not patella Pb
(AA: 7.1 ng/g, C: 7.1 ng/g, p = 0.46) or blood Pb levels (AA: 3.6 ng/dL, C: 3.6 ng/dL,
p = 0.69). Greater tibia (but lower patella) Pb levels may indicate greater historical
exposure among African Americans compared to Caucasians in the Baltimore population
studied by Theppeang et al. (2008b).
Differences in potential exposure among ethnic and racial groups have also been noted in
a study in the greater metropolitan New Orleans area. Campanella and Mielke (2008)
found that, in Census blocks where surface soil Pb levels were less than 20 mg/kg, the
population was 36% black, 55% white, 3.0% Asian, and 6.0% Hispanic, based on the
2000 Census, with the percentage based on the total number living in Census blocks with
the same soil Pb levels. In contrast, they found that for Census blocks in which soil Pb
levels were between 1,000 and 5,000 mg/kg, the population was 62% black, 34% white,
1% Asian, and 4% Hispanic (Figure 5-2), although the total population size generally
declined with soil Pb concentration, with the Census blocks with soil Pb of 1,000-5,000
mg/kg having less than half the population of that in the <20 mg/kg blocks. As described
in Section 5.2.4. the differences observed by Campanella and Mielke (2008) may also be
attributable to SES factors, or SES may be a confounding factor in the relationship
between Pb soil levels and race/ethnicity of nearby residents.
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70 -
60 -
= 50
OJ
5 40
M—
o
Ł 30
OJ
u
^
OJ
O. 20
10 -
70 i
1988-1991
l-<2.5 2.5-<5 5-<7.5 7.5 - <10 > 10
1999 - 2004
o
<1 l-<2.5 2.5-<5 5-<7.5 7.5 - <10 > 10
Blood Pb Level (ug/dL)
..^.. Non-Hispanic black^^^^Mexican American ~^~ Non-Hispanic white
Note: from the NHANES survey, 1988-1991 (top) and 1999-2004 (bottom).
Data used with permission of the American Academy of Pediatrics, Jones et al. (2009a)
Figure 5-1 Percent distribution of blood Pb levels by race/ethnicity among
U.S. children (1-5 years).
5-13
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• Black
D White
D Asian
D Hispanic
Soil Pb Concentration (mg/kg)
Note: By Census 2000 race/ethnicity demographic groups.
Source: Data used with permission of Springer Science; Campanella and Mielke (2008).
Figure 5-2 Soil Pb concentration exposure among the population of three
parishes within greater metropolitan New Orleans.
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5.2.4 Socioeconomic Status (SES)
Socioeconomic factors have sometimes been associated with Pb exposure biomarkers.
Previous results reported in the 2006 Pb AQCD found negative associations between
income or other SES metrics and blood Pb, although these relationships were not always
statistically significant (U.S. EPA. 2006b). Nriagu et al. (2006) performed a multiple
regression analysis of blood Pb and saliva Pb levels on various Socioeconomic,
demographic, and exposure variables among an adult population in Detroit, Michigan.
Blood and saliva Pb were both used as indicators of Pb in unbound plasma that is
available to organs. Nriagu et al. (2006) found that education (p <0.001), income
(p <0.001), and employment status (p = 0.04) were all statistically significant predictors
of blood Pb levels, with blood Pb decreasing with some scatter as education and income
level increased. Statistically significant relationships were also reported by Nriagu et al.
(2006) for saliva Pb level with respect to education (p <0.001), income (p <0.001), and
employment (p = 0.06). However, the highest educational attainment and income
categories had higher saliva Pb levels compared with other groups; Nriagu et al. (2006)
attributed these inconsistencies to small sample sizes among the high educational
attainment and income categories.
On a national level, the difference in blood Pb levels that have historically been seen to
exist between different income levels has been decreasing. For example, Levin et al.
(2008) cited 1991-1994 NHANES data [analyzed in Pirkle et al. (1994)1 that the
percentage of children aged 1-5 years with blood Pb levels > 10 ug/dL was 4.5% for the
lowest income group compared with 0.7% for the highest income group. Levin et al.
(2008) also analyzed data from the 1999-2002 NHANES and found no statistically
significant difference between the percent of children with blood Pb levels above
10 ug/dL for Medicaid-enrolled children (1.7%) compared with non-enrolled children
(1.3%). However, Medicaid-enrolled children did have higher median blood Pb levels
(2.6 ug/dL) compared to children not enrolled in Medicaid (1.7 ug/dL). Adding data for
2003-2004 to the analysis (i.e., for 1999-2004), widened the difference between Medicaid
enrolled and non-enrolled children with regard to percentage having blood Pb levels
> 10 ug/dL (1.9% versus 1.1%), but the difference was still not statistically significant (p
>0.05) and median blood Pb levels for the two groups did not change (Levin et al., 2008).
Likewise, Jones et al. (2009a) analyzed blood Pb levels with respect to poverty-income
ratio (PIR), which is the ratio of family income to the poverty threshold appropriate for a
given family size. They found statistically significant differences in median blood Pb for
PIR < 1.3 compared with PIR >1.3. The percentage of 1- to 5-year-old children having
blood Pb > 10 ug/dL was higher for PIR < 1.3 (1.8 versus 0.8); however, this difference
was not statistically significant. Additionally, in residential areas of metropolitan New
Orleans with soil concentrations below 20,000 mg/kg, Campanella and Mielke (2008)
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observed a linear increase in surface soil Pb concentration with decreasing median
household income, suggesting a relationship of potential exposure with household
income. The census block-averaged median household income in areas with soil Pb
between 2.5 and 20 mg/kg was $40,000 per year, while the corresponding median income
in areas with soil Pb between 5,000 and 20,000 mg/kg was $24,000 per year. The highest
soil concentrations (20,000 mg/kg and above) was associated with a median income of
$27,000.
5.2.5 Proximity to Pb Sources
Air and soil Pb concentrations are higher in some industrialized and urbanized areas, as
described in Sections 2.2. 2.3. 2.5. and 3.1. as a result of historical and contemporary Pb
sources. The highest air Pb concentrations measured using the Pb-TSP monitoring
network have been measured at monitors located near sources emitting Pb. Elevated soil
Pb concentrations have also been measured in urbanized areas compared with less
urbanized or rural locations (Tilippelli et al.. 2005). Air Pb concentrations exhibit high
spatial variability even at low concentrations (-0.01 ug/m3) (Martuzevicius et al.. 2004).
Proximity to an industrial source likely contributes to higher Pb exposures, as described
in the 2006 Pb AQCD (U.S. EPA. 2006b) for several studies of Superfund and other
industrial sites. This is consistent with the observation of higher air concentrations at
source oriented Pb monitoring sites compared with non-source oriented sites in the
2008-2010 data presented in Section 2.5.
Jones et al. (2010) found that neonates born near a Pb-contaminated hazardous waste site
had significantly higher umbilical cord blood Pb levels (median: 2.2 ug/dL [95% CI: 1.5,
3.3 ug/dL]) compared with a reference group of neonates not living near a potentially
contaminated site (median: 1.1 ug/dL [95% CI: 0.8, 1.3 ug/dL]), suggesting that
Pb-contaminated hazardous waste sites contribute to neonatal Pb levels. The population
studied in Jones et al. (2010) was 88% African American; 75% had a high school degree
or equivalent, while 20% had a college degree and 5% attended but did not graduate from
high school. However, Jones et al. (2010) did not analyze covariation between exposure
and maternal characteristics, so it cannot be determined if differences in characteristics
among the maternal groups (which did and did not report nearby hazardous waste sites)
confounded these results.
Studies have suggested that concentration of Pb in soil, a potential exposure media, is
related to land use type and historical sources, as described in Section 2.6.1. For instance,
Wu et al., (2010) observed that bioavailable Pb concentrations in Los Angeles surface
soil samples were significantly associated with traffic-related variables and parcel age
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(i.e., length of time since the parcel was first developed), with parcel age being a highly
significant predictor of bioavailable soil Pb in most models (p <0.0001). Zahran et al.
(2010) observed that surface soil Pb levels in 46 Census tracts of metropolitan New
Orleans dropped following Hurricanes Katrina and Rita, from 330 mg/kg to 200 mg/kg
(averages of median measurements across all Census tracts for 2000 and 2006) and
attributed this observation to coverage by relatively cleaner river sediments. Blood Pb
levels obtained from children (ages 0-6 years) also declined subsequent to the hurricanes;
statistical modeling of the changes in soil and blood Pb estimated the decline to be
1.55 (ig/dL for each 1% reduction in soil Pb (p < 0.05).
5.2.6 Residential Factors
Findings from a recent study of the association between blood Pb and housing factors by
Dixon et al. (2009). which analyzed data from the NHANES national survey for
1999-2004, are consistent with those from previous studies presented in the
2006 Pb AQCD that observed positive associations between increased blood Pb and
increased house dust Pb levels (U.S. EPA. 2006b; Lanphear et al.. 1998; Laxen et al..
1987). Dixon et al. (2009) used NHANES data from 1999-2004 to perform a linear
regression of blood Pb among children 12-60 months old on several factors including
year of home construction, floor surface condition, floor dust Pb level, windowsill dust
Pb level, and renovation in homes built before 1978. They found an inverse association
between blood Pb (log transformed) and residence in homes being built after 1950
(p = 0.014). Blood Pb was positively associated with windowsill Pb level (p = 0.002),
dust Pb level (p <0.001), and occurrence of renovation in pre-1978 homes (p = 0.045).
Detailed results of this regression are shown in Table 5-4. As part of the same study,
Gaitens et al. (2009) performed a regression analysis of floor dust Pb (PbD) and
windowsill dust Pb on several factors. Floor dust Pb (log transformed) was positively
associated with the following housing-related factors: not smooth/cleanable floor surface
(p <0.001), windowsill dust Pb (log transformed) (p <0.001), and older year of
construction (p <0.001). Windowsill dust Pb (log transformed) was positively associated
with the following housing-related factors: older year of construction (p <0.001), not
smooth/cleanable window surface (0.001) and deteriorated indoor paint (p = 0.028).
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Table 5-4 Regression of log-transformed blood Pb level of children 12-60
months old on various factors related to housing condition, from
1999-2004 NHANES dataset.
Variables
Intercept
Age (in years)
Year of construction
PIR
Race/ethnicity
Country of birth
Floor surface/condition *
log Floor PbD
Floor surface/condition *
(log Floor PbD)2
Floor surface/condition *
(log Floor PbD)3
Overall p-value Levels9
0.172
<0.001 Age
Age2
Age3
Age4
0.014 Intercept for missing
1990-present
1978-1989
1960-1977
1950-1959
1940-1949
Before 1940
<0.001 Intercept for missing
Slope
<0.001 Non-Hispanic white
Non-Hispanic black
Hispanic
Other
0.002 Missing
U.S.b
Mexico
Elsewhere
<0.001 Intercept for missing
Not smooth and cleanable
Smooth and cleanable or
carpeted
Not smooth and cleanable
Smooth and cleanable or
carpeted
Uncarpeted not smooth and
cleanable
Smooth and cleanable or
rarnptprl
Estimate (SE)
-0.517(0.373)
2.620 (0.628)
-1.353(0.354)
0.273 (0.083)
-0.019(0.007)
-0.121 (0.052)
-0.198(0.058)
-0.196(0.060)
-0.174(0.056)
-0.207 (0.065)
-0.012(0.072)
0.000
0.053 (0.065)
-0.053(0.012)
0.000
0.247 (0.035)
-0.035 (0.030)
0.128(0.070)
-0.077(0.219)
0.000
0.353(0.097)
0.154(0.121)
0.178(0.094)
0.386 (0.089)
0.205 (0.032)
0.023(0.015)
0.027 (0.008)
-0.020(0.014)
-0.009 (0.004)
p-Value
0.172
<0.001
<0.001
0.002
0.008
0.024
0.001
0.002
0.003
0.003
0.870
—
0.420
<0.001
—
<0.001
0.251
0.073
0.728
—
<0.001
0.209
0.065
<0.001
<0.001
0.124
0.001
0.159
0.012
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Table 5-4 (Continued): Regression of log-transformed blood Pb level of children 12-60
months old on various factors related to housing condition, from
1999-2004 NHANES dataset.
Variables Overall p-value
log Window-sill PbD 0.002
Home-apartment type <0.001
Anyone smoke inside the home 0.015
log Cotinine concentration 0.004
(ng/dL) in blood
Window cabinet or wall 0.045
renovation in a pre-1978 home
Levels3
Intercept for missing
Slope
Intercept for missing
Mobile home or trailer
One family house detached
One family house attached
Apartment (1-9 units)
Apartment (> 10 units)
Missing
Yes
No
Intercept for missing
Slope
Missing
Yes
No
Estimate (SE)
0.053 (0.040)
0.041 (0.011)
-0.064 (0.097)
0.127(0.067)
-0.025 (0.046)
0.000
0.069 (0.060)
-0.133(0.056)
0.138(0.140)
0.100(0.040)
0.000
-0.150(0.063)
0.039(0.012)
-0.008(0.061)
0.097 (0.047)
0.000
p-Value
0.186
<0.001
0.511
0.066
0.596
—
0.256
0.022
0.331
0.015
—
0.023
0.002
0.896
0.045
—
Children: N = 2,155 (age 10-60 months); R2 = 40%
""Includes the 50 states and the District of Columbia
Source: Dixon et al. (2009).
Renovation activities on older homes have been shown to produce excess Pb dust
concentrations. Gaitens et al. (2009) performed a regression analysis on dust Pb
concentrations from 1994-2004 NHANES on demographic and housing variables and
found that renovation of windows, cabinets, or walls in a pre-1950 home was positively
associated with floor dust Pb concentration (p <0.001). Paint scraping within the last
twelve months was positively associated with windowsill dust Pb concentration
(p = 0.053). Dixon et al. (2009) performed a regression analysis on log-transformed blood
Pb levels from NHANES (1999-2004) on several demographic and housing variables and
found that renovation of windows, cabinets, or walls in pre-1978 homes was positively
associated with blood Pb concentration (p = 0.045). A case study by Mielke et al. (2001)
reports on elevated indoor and outdoor dust Pb levels at two houses where exterior paint
has been either power sanded (without confinement of released material) or hand scraped
(with collection of released material) to prepare for repainting. The latter approach
appeared to yield lower dust Pb levels, although given the extremely limited dataset,
conclusions are uncertain. In an occupational study of men performing home renovations
in the U.K., window renovation and wood-stripping workers specializing in renovation of
old houses had significantly higher median blood Pb levels compared with all workers in
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similar occupations (wood strippers: 37 (ig/dL; window renovators: 32 (ig/dL; all
workers: 13.7 (ig/dL; p <0.001) (Mason et al.. 2005).
5.3 Factors Potentially Related to Increased Risk of Pb-lnduced
Health Effects
This section evaluates factors examined in recent studies as effect measure modifiers that
potentially increase the risk of various Pb-related health effects. There was limited
evidence from the 2006 Pb AQCD (U.S. EPA. 2006b) for many of potential at-risk
factors described below. Where available, information on conclusions regarding at-risk
populations from the 2006 Pb AQCD is included in the subsections.
5.3.1 Age
Below is information from epidemiologic and toxicological studies regarding studies of
increased risk for Pb-related health effects among children and older adults. Other age
groups, such as adolescents, have not been evaluated here, if they were not part of
stratified studies of lifestage.
5.3.1.1 Childhood
According to the 2010 Census, 24.0% of individuals living in the U.S. were under the age
of 18, with 6.5% aged 0-4 years and 17.5% aged 5-17 years (Howden and Meyer. 2011).
It is recognized that Pb can cross the placenta and affect the developing nervous system
of the fetus (Sections 3.2.2.4 and 4.3) and there is strong evidence of increased risk to the
neurocognitive effects of Pb exposure during several lifestages throughout gestation,
childhood, and into adolescence (for more detail, Section 4.3). However, most recent
studies among children do not have adequate comparison groups between children of
various age groups or between children and adults, and were therefore only presented in
Chapter 4. Overall, early childhood has been demonstrated to be a lifestage of increased
risk for Pb-related health effects.
A study including multiple U.S. locations examined associations of blood Pb levels with
various immune parameters among individuals living near Pb industrial sites and
matched controls (Sarasua et al.. 2000). For several of these endpoints, the association in
the youngest group (ages 6-35 months) and the oldest group (ages 16-75 years) were in
opposite directions. For example, among children ages 6-35 months, the associations
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between blood Pb levels and Immunoglobulin A (IgA), Immunoglobulin M (IgM), and
B-cell abundance were positive, whereas the associations among 16-75 year olds were
negative. The opposite associations were also present for T cell abundance. Ig antibodies,
which are produced by activated B cells, are important mediators of the humoral immune
response to antigens. T cells are important mediators of cell-mediated immune responses
that involve activation of other immune cells and cytokines. These findings by Sarasua et
al. (2000) indicate that very young children may be at increased risk for Pb-associated
activation of humoral immune responses and perturbations in cell-to-cell interactions that
underlie allergic, asthma, and inflammatory responses (for more information, see
Sections 4.6.2.1 and 4.6.3).
A study among Lebanese children examined the association between blood Pb levels and
transferrin saturation (TS) less than 12% (a measure of Fe-deficiency with or without
accompanying anemia) and Fe (iron)-deficiency anemia (IDA) (Muwakkit et al., 2008).
A positive association was detected for blood Pb levels > 10 (ig/dL and both TS less than
12% and IDA among children aged 11-23 months old; however, null associations were
observed among children 24-35 months old. Calculations were not performed for children
aged 36-75 months because there were no children in the highest Pb group (> 10 (ig/dL)
with either TS <12% or IDA. The authors noted that it is difficult to know whether the Pb
levels were "a cause or a result of IDA levels since previous studies linked Fe deficiency
with Pb toxicity.
Overall evidence indicates early childhood as a lifestage of increased risk for Pb-related
health effects. Both epidemiologic studies summarized above reported associations
among the youngest age groups, although different age cut-points were used with one
study including only infants 35 months of age and younger. Toxicological studies provide
support for increased health effects of Pb among younger age groups. Toxicological
studies have reported that younger animals, whose nervous systems are developing
(i.e., laying down and pruning neuronal circuits) and whose junctional barrier systems in
the brain (i.e., the blood brain barrier) and GI system (i.e., gut closure) are immature, are
more at risk from the effects Pb exposure (Rice and Barone. 2000; Landrigan et al..
1999). In sum, there are consistent findings, coherent across disciplines that adequate
evidence exists to conclude that children are an at-risk population. There are insufficient
data to identify specific critical windows of exposure for Pb-related effects.
5.3.1.2 Older Adulthood
The number of Americans over the age of 65 will be increasing in upcoming years
(estimated to increase from 12.4% of the U.S. population to 19.7% between 2000 to
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2030, which is approximately 35 million and 71.5 million individuals, respectively)
(SSDAN CensusScope. 2010: U.S. Census Bureau. 2010). As of the 2010 Census, 7.0%
of the U.S. population were ages 65-74, 4.2% were 75-84, and 1.7% were age 85 and
older (Werner. 2011). Overall, inconsistent results are reported for age-related
modification of the association between Pb and mortality. There is limited evidence for
the associations between Pb and other health effects among older adults.
A study using the NHANES III cohort examined blood Pb levels and mortality among
individuals less than 60 years old and individuals 60 years and older (Menke et al., 2006).
Positive hazard ratios were observed in both age groups but the hazard ratios were greater
in those less than 60 years old. The interaction terms were not statistically significant. A
similar study using the NHANES III cohort examined the relationship between blood Pb
levels and mortality from all-cause, cardiovascular disease, and cancer broken down into
more specific age groups (Schober et al.. 2006). Point estimates were elevated for the
association comparing blood Pb levels > 10 (ig/dL to blood Pb levels <5 (ig/dL and all-
cause mortality for all age groups (40-74, 75-84, and 85+ year olds), although the
association for 75-84 year olds did not reach statistical significance. The association was
also present when comparing blood Pb levels of 5-9 (ig/dL to blood Pb levels <5 (ig/dL
among 40-74 year olds and 75-84 year olds, but not among those 85 years and older.
None of the associations between blood Pb and cardiovascular disease-related mortality
reached statistical significance but the point estimates for cardiovascular disease-related
mortality comparing blood Pb levels > 10 (ig/dL to blood Pb levels <5 (ig/dL were
elevated among all age groups. Finally, the association between blood Pb levels
> 10 (ig/dL and cancer mortality was positive among those 40-74 years old and 85 years
and older but the association was null for those 75-84 years old. Among 75-84 year olds
the association was positive comparing blood Pb levels of 5-9 (ig/dL to <5 (ig/dL. The
other age groups had similar point estimates but the associations were not statistically
significant.
A study using the Normative Aging Study cohort reported an interaction between Pb and
age (Wright et al. 2003). The inverse association between age and cognitive function was
greater among those with high blood or patella Pb levels. Effect estimates were in the
same direction for tibia Pb but the interaction was not statistically significant.
Finally, a study of current and former Pb workers reported that an interaction term of Pb
and age (dichotomous cutpoint at 67th percentile but exact age not given) examined in
models of Pb (measured from blood and patella) and blood pressure was not statistically
significant (Weaver et al.. 2008). Thus, no modification by age was observed in this study
of Pb and blood pressure.
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Toxicological studies have demonstrated Pb-related health effects among older
populations. The kidneys of older animals appear to be more at-risk for Pb-related health
effects from the same dose of Pb (i.e., continuous 50 mg/L Pb acetate drinking water)
than younger animals (Berrahal et al.. 2011). Increased risk related to older age is also
observed for effects on the brain. Recent studies have demonstrated the importance of Pb
exposure during early development in promoting the emergence of Alzheimer's like
pathologies in aged animals. Development of pathologies of old age in brains of aged
animals that were exposed to Pb earlier in life has been documented in
psychopathological effects in adults (mice and monkeys), (for more details see Section
4.3.10.1). These pathologies include the development of neurofibrillary tangles and
increased amyloid precursor protein and its product beta-amyloid (Bashaetal.. 2005;
Zawia and Basha. 2005). Some of these findings were seen in animals that no longer had
elevated blood Pb levels.
In summary, results for age-related modification of the association between Pb and
mortality had mixed results. Limited evidence was available for the associations between
Pb and cognitive function or other health effects among older adults. Toxicological
studies have shown increases in Pb-related health effects by age that may be relevant in
humans. Future studies will be instrumental in understanding older age as a factor that
potentially affects the risk of Pb-related outcomes.
5.3.2 Sex
The distribution of males and females in the U.S. is similar. In 2010, 49.2% of the U.S.
population was male and 50.8% was female. The distribution of sex varied by age with a
greater prevalence of females in older age groups compared to males (Howden and
Meyer. 2011). The 2006 Pb AQCD reported that boys are often found to have higher
blood Pb levels than girls, but findings were "less clear" regarding differences in
Pb-related health effects between males and females (U.S. EPA. 2006b). Recent studies
examining effect measure modification by sex have also reported inconsistent findings.
Multiple epidemiologic studies have examined Pb-related effects on cognition stratified
by sex. In previous studies using the Cincinnati Lead Study cohort, Dietrich et al. (1987)
and Ris et al. (2004) observed interactions between blood Pb (prenatal and postnatal) and
sex; associations of prenatal and postnatal blood Pb and subsequent decrements in
memory, attention, and visuoconstruction were observed only among male adolescents.
More recently, Wright et al. (2008) examined early life blood Pb levels and criminal
arrests in adulthood. The risks attributable to Pb exposure were greater among males than
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females. Additionally, the association between childhood blood Pb levels and adult gray
matter volume loss was greater among males than females (Cecil et al.. 2008). In an
expanded analysis of the developmental trajectory of childhood blood Pb levels on adult
gray matter, researchers found that associations between yearly mean blood Pb levels and
volume of gray matter loss were more pronounced in the frontal lobes of males than
females (Brubakeretal.. 2010). Multiple studies were also conducted in Port Pirie,
Australia that examined blood Pb levels at various ages throughout childhood and
adolescence (Tong et al.. 2000; Baghurst et al.. 1992; McMichael et al.. 1992). These
studies observed Pb effects on cognition deficits were stronger in girls throughout
childhood and into early adolescence. A study in Poland also investigated the association
between umbilical cord blood Pb levels and cognitive deficits and reported a positive
association for boys at 36 months but not for girls (Jedrychowski et al.. 2009a). No
association was detected for boys or girls at 24 months.
An epidemiologic study examined the association between concurrent blood Pb levels
and kidney function among 12-20 year olds using the NHANES III study cohort
(Fadrowski et al.. 2010). The results were stratified by sex and no effect measure
modification was apparent.
Similarly, a study of current and former Pb workers examined an interaction term
between sex and Pb for the study of blood Pb and blood pressure (Weaver et al.. 2008).
No modification by sex was present.
Epidemiologic studies have also been performed to assess differences between males and
females for Pb-related effects on various biomarkers. A study comprised mostly of
females reported positive associations between blood Pb and total immunoglobulin E
(IgE) for women not taking hormone replacement therapy or oral contraceptives (Pizent
et al.. 2008). No association was reported in males, but other associations, such as
bronchial reactivity and reactive skin prick tests were observed in the opposite of the
expected direction, which questions the validity of the results among the male study
participants. Analysis of an NHANES dataset detected no association between blood Pb
levels and inflammatory markers (Songdej et al.. 2010). Although there was no clear
pattern, a few of the associations were positive between blood Pb and C-reactive protein
for males but not females. A study of children living at varying distances from a Pb
smelter in Mexico reported that blood Pb was associated with increased release of
superoxide anion from macrophages, which was greater among males than females
(Pineda-Zavaleta et al.. 2004).
Epidemiologic investigations of cancer have also examined the associations by sex. A
study of the association between occupational exposure to Pb and brain tumors reported
no sex-specific associations for gliomas, but a positive association for cumulative Pb
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exposure and meningiomas for males but not females (Rajaraman et al.. 2006). An
ecologic analysis of Pb pollution levels and cancer incidence among children reported
weak correlations overall and the weak correlations were more apparent among males,
whereas no correlation was observed among females (Absalon and Slesak. 2010).
A study of all-cause and cardiovascular mortality using the NHANES III cohort reported
no modification of the association between blood Pb and all-cause or cardiovascular
mortality by sex (Menke et al.. 2006). This did not differ among women when classified
as pre- or post-menopausal.
Toxicological studies have also reported sex differences in Pb-related effects to various
organ systems. Donald et al. (1986) reported a different time course of enhanced social
investigatory behavior between male and female mice exposed to Pb. In a subsequent
publication, Donald et al. (1987) showed that non-social behavior in mice decreased in
females and increased in males exposed to Pb. Males also had a shorter latency to
aggression with Pb treatment versus controls. Pb affected mood disorders differently for
males and females. Behavioral testing in rats showed males experienced emotional
changes and females depression-like changes with Pb exposure (de Souza Lisboa et al..
2005). In another study, gestational exposure to Pb impaired memory retrieval in male
rats at all 3 doses of Pb exposure; memory retrieval was only impaired in low-dose
female rats (Yang et al.. 2003). Sex-specific differences in mice were also observed for
gross motor skills; at the lowest Pb dose, balance and coordination were most affected
among males (Leasure et al.. 2008).
Pb and stress are co-occurring factors that act in a sex-divergent manner to affect
behavior, neurochemistry, and corticosterone levels. Pb and stress act synergistically to
affect fixed interval operant behavior and corticosterone in female rat offspring. Virgolini
et al. (2008a) found that effects on the offspring's central nervous system by
developmental Pb exposure (maternal exposure and transferred to the offspring through
lactation) were enhanced by combined maternal and offspring stress and females were
most at risk. Behavioral related outcomes after gestational and lactational Pb exposure
(with and without stress) exhibited sex-differences in exposed offspring (Virgolini et al.,
2008b). Pb-induced changes in brain neurochemistry, with or without concomitant stress
exposure, are complex with differences varying by brain region, neurotransmitter type,
and sex of the animal.
The brain is known to have a sexually dimorphic area in the hypothalamus, termed the
sexually dimorphic nucleus (SON). Lesions in this area affect sex-specific phenotypes
including behavior. Across species the SDN has a greater cell number and larger size in
males versus females. This sexually dichotomous area is especially vulnerable to
perturbation during fetal life and the early postnatal period. This may be one area of the
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brain that could explain some of the sexually dichotomous effects that are seen with Pb
exposure. One study supporting this line of thought showed that high-dose in utero Pb
exposure (pup blood Pb level 64 (ig/dL at birth) induced reductions in SDN volume in
35% of Pb-exposed male rats (McGivern et al. 1991). Interestingly, another chemical
that is known to cause a hypothalamic lesion in this area, monosodium glutamate, is
associated with adult onset obesity among mice (Olnev. 1969): adult onset obesity is seen
in the Pb literature.
Obesity in adult offspring exposed to low-dose Pb in utero was reported for male but not
female mice (Leasure et al.. 2008). Obesity was also found in male rat offspring exposed
in utero to high doses of Pb that persisted to 5 weeks of age/end of the study, but among
female rats, body weight remained elevated over controls only to 3 weeks of age (Yang et
al.. 2003). Additionally, low-dose Pb exposure induced retinal decrements in exposed
male mice offspring (Leasure et al.. 2008).
A toxicological study of Pb and antioxidant enzymes in heart and kidney tissue reported
that male and female rats had differing enzymatic responses, although the amount of Pb
in the heart tissue or the disposition of Pb also varied between males and females
(Sobekova et al.. 2009; Alghazal et al.. 2008a). The authors reported these results could
be due to greater deposition of Pb in female rats or greater clearance of Pb by males
(Sobekova et al.. 2009V
Multiple associations between Pb and various health endpoints have been examined for
effect measure modification by sex. Some studies reported differences in the risk of Pb-
related outcomes by sex, but did not consistently demonstrate whether male or female sex
was associated with a greater risk; studies on cognition from the Cincinnati Lead Study
cohort and a study in Poland reported males to be an at-risk population, whereas studies
from Australia pointed to females as an at-risk population. A difference in sex is
supported by toxicological studies. Further research is needed to confirm the presence or
absence of sex-specific associations between Pb and various health outcomes and to
determine in which sex the associations are greater.
5.3.3 Genetics
The 2006 Pb AQCD stated that, "genetic polymorphisms in certain genes have been
implicated as influencing the absorption, retention, and toxicokinetics of Pb in humans"
(U.S. EPA. 2006b). The majority of discussion there focused on the aminolevulinate
dehydratase (ALAD) and vitamin D receptor (VDR) polymorphisms. These two genes, as
well as additional genes examined in recent studies, are discussed below. The presence of
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ALAD variants is associated with an increase in Pb-related health outcomes, but there is
limited information for other genes.
5.3.3.1 Aminolevulinate Dehydratase
The aminolevulinate dehydratase (ALAD) gene encodes for an enzyme that catalyzes the
second step in the production of heme and is also the principal Pb-binding protein (U.S.
EPA. 2006b). Studies have examined whether ALAD variants altered associations
between Pb and various health effects.
Associations between Pb and brain tumors observed in an epidemiologic study varied by
ALAD genotype status (Raiaraman et al.. 2006). Positive associations between Pb
exposure (determined via interview about occupational exposures) and meningioma were
reported among ALAD2 individuals, but this association was not found among
individuals who had the ALAD1 allele. No associations were observed between Pb and
glioma regardless of ALAD genotype.
Studies investigating the association between Pb levels and cognitive function have also
examined modification by ALAD polymorphisms. The evidence is provided by an
NHANES analysis (Krieg et al.. 2009) as well as multiple analyses from the NAS cohort
examining different tests of cognitive function (Rajan et al.. 2008; Weuve et al.. 2006). In
the study using a cohort from NHANES III, for several indices of cognitive function,
associations with concurrent blood Pb levels were more pronounced in groups with CC
and CG ALAD genotypes (i.e., ALAD2 carriers) (Krieg et al.. 2009). In the NAS cohort
of men, Weuve et al. (2006) found that concurrent blood Pb level but not bone Pb level
was associated with a larger decrease in a test of general cognitive function among
ALAD2 carriers. Another NAS study examined functioning of specific cognitive
domains (e.g., vocabulary, memory, visuospatial skills) and found variable evidence for
effect modification by ALAD genotype across tests (Rajan et al.. 2008). For example,
among ALAD2 carriers, concurrent blood Pb level was associated with a more
pronounced decrease in vocabulary score but less pronounced decrease in a memory
index and no difference in the associations with other cognitive tests. For tibia and patella
Pb levels, ALAD genotype was found to modify associations with different tests, for
example, executive function and perceptual speed. It is not clear why the direction of
effect modification would vary among different cognitive domains. The limited number
of populations examined, and the different cognitive tests performed in each study, make
it difficult to conclusively summarize findings for effect modification by ALAD variants.
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However, in the limited available body of evidence, blood and bone Pb levels were
generally associated with lower cognitive function in ALAD2 carriers.
A study of current and former workers exposed to Pb examined the association between
blood Pb and blood pressure and reported no modification by ALAD genotype (Weaver
et al.. 2008). However, another study of blood Pb and blood pressure reported
interactions between blood Pb and ALAD, but this varied by race/ethnicity (non-Hispanic
white, non-Hispanic black, and Mexican American) (Scinicariello et al.. 2010). In this
study, non-Hispanic white carriers of the ALAD2 genetic variant in the highest blood Pb
quartile had a higher risk of hypertension compared with ALAD1 homozygous
individuals. The association between ALAD variants and hypertension did not vary
within any quartiles of blood Pb for non-Hispanic blacks or Mexican-Americans, but the
precision of estimates was low.
Individuals with ALAD2 variants had greater associations between Pb and kidney
effects; among those with the variant, higher Pb was associated with higher glomerular
filtration measures (Weaver et al.. 2006; Weaver et al.. 2005b: Weaver et al.. 2003b). A
study of workers at a battery plant storage facility in China reported workers with the
ALAD2 allele demonstrated greater associations between blood Pb levels and renal
injury (Gao et al.. 2010a). Another study of renal function among Pb workers in Asia also
reported greater associations between blood Pb concentrations and renal function by
ALAD, especially at high blood Pb levels (Chia et al.. 2006).
5.3.3.2 Vitamin D Receptor
The vitamin D receptor (VDR) is a regulator of calcium absorption and metabolism. A
recent study of the NHANES III population examined the association between blood Pb
levels and various neurocognitive tests with assessment of effect measure modification
by SNPs and haplotypes of VDR (Krieg et al.. 2010). The results were varied, even
among specific SNPs and haplotypes, with some variants being associated with greater
modification of the relationship between Pb and one type of neurocognitive test
compared to the modification of the relationship between Pb and other neurocognitive
tests. In an epidemiologic study of blood Pb levels and blood pressure among a group of
current and former Pb-exposed workers, no modification was reported by VDR (Weaver
et al.. 2008).
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5.3.3.3 Methylenetetrahydrofolate reductase
Methylenetetrahydrofolate reductase (MTHFR) catalyzes the conversion of
5,10-methylenetetrahydrofolate to 5-methyltetrahydrofolate, which in turn, is involved in
homocysteine remethylation to the amino acid methionine. A study in Mexico of the
association between Pb and Bayley's Mental Development Index (MDI) score at 24
months reported no effect measure modification by MTHFR 677T allele (Pilsner et al..
2010). Another study in Mexico examined the association between maternal Pb and birth
weight (Kordas et al.. 2009). No modification of the Pb-birth weight association by
MTHFR was observed.
5.3.3.4 Hemochromatosis
The hemochromatosis (HFE) gene encodes a protein believed to be involved in
Fe absorption. A difference was observed between the association of tibia Pb levels and
cognitive function for men with and without HFE allele variants (Wang et al.. 2007a). No
association between tibia Pb and cognitive function was present for men with HFE
wildtype, but a decline in function was associated with tibia Pb levels among men with
any HFE allele variant. A study of bone Pb levels and HFE reported no difference in
effect estimates for bone Pb and pulse pressure between different HFE variants and HFE
wild-type (Zhang et al.. 2010a). An interaction was observed between an HFE variant in
mothers and maternal tibia Pb in a study of maternal Pb and birth weight (Cantonwine et
al.. 2010b). The inverse association between maternal tibia Pb levels and birth weight
was stronger for those infants whose mothers had the HFE variant. The interaction was
not present between the HFE variants and maternal blood Pb or cord blood Pb
concentrations.
5.3.3.5 Other Genetic Polymorphisms
Some other genetic polymorphisms were also examined as to whether they modify
Pb-related health effects, but only limited data were available for these polymorphisms.
These include dopamine receptor D4 (DRD4), dopamine receptor D2 (DRD2), dopamine
transporter (DAT1), glutathione S-transferase Mu 1 (GSTM1), tumor necrosis factor-
alpha (TNF-a), endothelial nitric oxide synthase (eNOS), and various SNPS.
A prospective birth cohort reported that increasing blood Pb levels were associated with
poorer rule learning and reversal, spatial span, and planning in their study population
(Froehlich et al.. 2007). These inverse associations were exacerbated among those
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lacking DRD4-7. A study of prenatal and postnatal Pb levels in Mexico City reported no
modification of the associations between Pb levels and neurocognitive development by
DRD2 or DAT1 (Kordas et al.. 2011).
A study of university students in South Korea reported blood Pb levels to be associated
with biomarkers of inflammation among individuals with GSTM1 null genotype and not
among individuals with GSTM1 present (Kim et al.. 2007). This study of blood Pb levels
and inflammation also examined individuals with TNF-a GG, GA, or AA alleles. An
association was present for those with TNF-a GG but not for those with TNF-a GA or
AA.
A study of blood Pb and plasma NOX reported no overall association but did report an
inverse correlation among subjects with the eNOS TC+CC genotype (Barbosa et al..
2006c). No correlation was observed for subjects with the eNOS TT genotype; however
the number of subjects in this group was small, especially for those with high blood Pb
levels.
One study examined how the association between occupational Pb exposure and brain
tumors varied among multiple single nucleotide polymorphisms (SNPs) (Bhatti et al..
2009). No effect measure modification of the association between Pb and glioma was
observed for any of the SNPs. GPX1 (the gene encoding for glutathione peroxidase 1)
modified the association for glioblastoma multiforme and meningioma. The association
between Pb and glioblastoma multiforme was also modified by a RAC2 (the gene
encoding for Rac2) variant, and the association between Pb and meningioma was also
modified by XDH (the gene encoding for xanthine dehydrogenase) variant.
Overall, studies of ALAD observed increased Pb-related health effects associated with
certain gene variants. Other genes, such as VDR, HFE, DRD4, GSTM1, TNF-a, and
eNOS, may also affect the risk of Pb-related health effects but conclusions are limited
due to the small number of studies.
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5.3.4 Pre-existing Diseases/Conditions
Studies have also been performed to examine whether certain morbidities increase an
individual's risk of Pb-related effects on health. Recent studies have explored
relationships for autism, diabetes, and hypertension. In general, consistent associations
are observed indicating that hypertension increases the risk of Pb-related renal effects but
the evidence is limited for other pre-existing conditions. Although there are many
childhood conditions that could hypothetically increase the risk of Pb-related health
effects in children (e.g. prenatal exposure to alcohol or drugs, birth asphyxia, head
trauma, sickle cell anemia) epidemiologic studies to evaluate such hypotheses are
unavailable and therefore they are not discussed.
5.3.4.1 Autism
Rates of individuals with autism have increased in recent years. A study reported a
prevalence rate in 2006 of 9.0 per 1,000 individuals (95% CI: 8.6, 9.3) determined from a
monitoring network (Autism and Developmental Disabilities Monitoring Network) with
11 sites across the U.S. (CDC. 2009).
A cross-sectional study of children with and without autism examined the association
between blood Pb levels and various immune function and inflammation genes (Tian et
al., 2011). Blood Pb levels of children with and without autism were associated with
expression of the genes under study; however, the associations observed were in opposite
directions (for children with autism, increased blood Pb levels were associated with
increased expression, whereas for children without autism, increased blood Pb levels
were associated with decreased expression).
5.3.4.2 Diabetes
Approximately 8% of U.S. adults have diabetes (Pleis et al., 2009). A few studies have
been conducted to investigate the possibility of diabetes as a modifying factor for Pb and
various health outcomes.
Differences in the association between bone and blood Pb levels and renal function for
individuals with and without diabetes at baseline were examined using the Normative
Aging Study cohort (Tsaih et al. 2004). Tibia and blood Pb levels were positively
associated with measures of poor renal function among individuals with diabetes but not
among individuals without diabetes. However, this association was no longer statistically
significant after the exclusion of individuals who were hypertensive or who used diuretic
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medications. Another study with this cohort reported no associations between bone Pb
and heart rate variability, which did not differ among those with and without diabetes
(Park et al. 2006).
The NHANES III data were used to evaluate whether the association between blood Pb
and both all-cause and cardiovascular mortality varied among individuals with and
without diabetes (Menke et al., 2006). The 95% CIs among those with diabetes were
large and no difference was apparent among those with and without diabetes.
Overall, recent epidemiologic studies found that associations between Pb concentrations
and health outcomes did not differ for individuals with and without diabetes. However,
results from the 2006 Pb AQCD found that individuals with diabetes are at "increased
risk of Pb-associated declines in renal function" (U.S. EPA. 2006b). Future research
examining associations between Pb and renal function, as well as other health outcomes,
among individuals with and without diabetes will inform further on the potential for
increased risk among individuals with diabetes.
5.3.4.3 Hypertension
Hypertension affects approximately 24% of adults in the U.S. and the prevalence of
hypertension increases with age (61% of individuals > 75 years old have hypertension)
(Pleis et al.. 2009).
The Normative Aging Study mentioned above evaluating modification of the association
between Pb levels and renal function by diabetes also examined modification by
hypertensive status (Tsaih et al.. 2004). The association between tibia Pb and renal
function, measured by change in serum creatinine, was present among individuals with
hypertension but not among individuals that were normotensive. Models of the follow-up
serum creatinine levels demonstrated an association with blood Pb for individuals with
hypertension but not individuals without hypertension (this association was not present
when using tibia or patella Pb). Another study using this population examined
modification of the association between bone Pb and heart rate variability, measured by
low frequency power, high frequency power, and their ratio (Park et al.. 2006). Although
a statistically significant association between bone Pb and heart rate variability was not
observed among individuals with or without hypertension, the estimates were different,
with greater odds for individuals with hypertension (bone Pb levels were positively
related to low frequency power and the ratio of low frequency to high frequency power
and were inversely related to high frequency power).
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A study using the NHANES III cohort reported a positive association between blood Pb
levels and both all-cause and cardiovascular mortality for individuals with and without
hypertension but the associations did not differ based on hypertensive status (Menke et
al.. 2006).
The 2006 Pb AQCD reported that individuals with hypertension had increased risk of
Pb-related effects on renal function (U.S. EPA. 2006b). This is supported by recent
epidemiologic studies. As described above, studies of Pb-related effects on renal function
and heart rate variability have observed some differences among individuals with
hypertension, but the difference between adults with and without hypertension was not
observed for Pb-related mortality.
Overall, studies of Pb-related health effects related to pre-existing conditions have some
evidence of a potential increased risk of Pb-related health effects. The evidence is
consistent for Pb-related renal effects and hypertension but is limited for other pre-
existing conditions.
5.3.5 Smoking Status
The rate of smoking among adults 18 years and older in the U.S. is approximately 20%
and about 21% of individuals identify as former smokers (Pleis et al.. 2009). Studies of
Pb and various health effects have examined smoking as an effect measure modifier and
reported inconsistent results.
A study of blood Pb levels and all-cause and cardiovascular mortality reported no
modification of this association by smoking status, measured as current, former, or never
smokers (Menke et al.. 2006). The Normative Aging Study also examined the association
between blood and bone Pb levels and renal function and also reported no interaction
with smoking status (Tsaih et al.. 2004).
A study of Pb-exposed workers and controls reported similar levels of absolute neutrophil
counts (ANC) across Pb exposure categories among nonsmokers (Pi Lorenzo et al..
2006). However, among current smokers, higher Pb exposure was associated with higher
ANC. Additionally, a positive relationship was observed between higher blood Pb levels
and TNF-a and granulocyte colony-stimulating factor (G-CSF) among both smokers and
nonsmokers, but this association was greater among smokers (Di Lorenzo et al., 2007). A
recent study of fertile and infertile men examined blood and seminal plasma Pb levels for
smokers and nonsmokers (Kiziler et al., 2007). The blood and seminal plasma Pb levels
were higher for smokers of both fertile and infertile groups. Additionally, the Pb levels
were lowest among nonsmoking fertile men and highest among smoking infertile men.
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Maternal smoking during pregnancy was examined in a study of children's concurrent
blood Pb levels and prevalence of attention-deficit/hyperactivity disorder (ADHD)
among children aged 8-15 years. An interaction was observed between children's current
blood Pb levels and prenatal tobacco smoke exposure; those children with high Pb levels
and prenatal tobacco smoke exposure had the highest odds of ADHD (Froehlich et al.
2009).
Overall, the studies have inconsistent findings on whether smoking modifies the
relationship between Pb levels and health effects. Future studies of Pb-related health
effects and current, former, second-hand, and prenatal smoking exposures among various
health endpoints will aid in determining changes in risk by this factor.
5.3.6 Socioeconomic Status
According to the Current Population Survey Annual Social and Economic Supplements
conducted by the Census Bureau, in 2011 15.0% of Americans lived in poverty (average
poverty threshold for family of 4 was $23,021) (DeNavas-Walt et al., 2012). Few studies
have compared blood Pb level effect estimates among groups in different
sociodemographic strata. Larger blood Pb-associated decreases in cognitive function
were found with lower SES in some studies (Ris et al., 2004; Tong et al., 2000; Bellinger
et al.. 1990). In contrast, a meta-analysis of eight studies found a smaller decrement in
Full Scale Intelligence Quotient (FSIQ) for studies in disadvantaged populations than for
studies in advantaged populations (Schwartz. 1994). While results indicate that blood Pb
level is associated with FSIQ deficits in both higher and lower sociodemographic groups,
they do not clearly indicate whether groups with different socioeconomic status differ in
Pb-related changes for cognitive function.
5.3.7 Race/Ethnicity
Based on the 2010 Census, 63.7% of the U.S. population is comprised of non-Hispanic
whites. Approximately 12.2% of people reported their race/ethnicity as non-Hispanic
black and 16.3% reported being Hispanic (Humes et al., 2011). Studies of multiple
Pb-related health outcomes examined effect measure modification by race/ethnicity and
although there are a limited number of studies, there is evidence that certain Pb-related
outcomes may vary by race/ethnicity.
A study of adults from the NHANES III cohort examined the association between blood
Pb levels and all-cause and cardiovascular mortality (Menke et al., 2006). Stratified
analyses were conducted for non-Hispanic whites, non-Hispanic blacks, and Mexican
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Americans and no interaction for race/ethnicity was reported. Other studies have also
used NHANES cohorts to study blood Pb levels and hypertension (Scinicariello et al..
2010; Muntner et al., 2005). While no association was observed between blood Pb and
hypertension for non-Hispanic whites or Hispanics, a positive association was reported
for non-Hispanic blacks in a study using the NHANES III cohort (Scinicariello et al.,
2010). In another study, although none of the associations between blood Pb levels and
hypertension were statistically significant, increased odds were observed among
non-Hispanic blacks and Mexican Americans but not for non-Hispanic whites (Muntner
et al.. 2005).
A study of girls aged 8-18 years from the NHANES III cohort reported an inverse
association between blood Pb levels and pubertal development among blacks and
Mexican Americans (Selevan et al.. 2003). For non-Hispanic whites, the associations
were in the same direction but did not reach statistical significance. Of note, less than 3%
of non-Hispanic whites had blood Pb levels over 5 (ig/dL, whereas 11.6% and 12.8% of
blacks and Mexican Americans, respectively, had blood Pb levels greater than 5 (ig/dL.
A study linking educational testing data for 4th grade students in North Carolina reported
declines in reading and mathematics scores with increasing levels of blood Pb (Miranda
et al., 2007a). Although not quantitatively reported, a figure in the study depicted the
association stratified by race, and the slopes appeared to be similar for white and black
children.
Blood Pb and asthma incidence was examined for white and black children living in
Michigan (Joseph et al.. 2005). When utilizing separate referent groups for the two races,
the only association is an increase among whites (although not statistically significant),
but when restricting to the highest blood Pb levels, the association was no longer
apparent. Whites with low blood Pb levels were used as the referent group for both races
in additional analysis. Although the estimates were elevated for black children compared
to white children (including at the lowest blood Pb levels), the confidence intervals for
the associations overlapped indicating a lack of a difference by race.
The results of these recent epidemiologic studies provide some evidence that there may
be race/ethnicity-related increased risk with higher Pb levels for certain outcomes,
although the overall understanding of potential effect measure modification by
race/ethnicity is limited by the small number of studies. Additionally, these results may
be confounded by other factors, such as socioeconomic status or nutritional factors.
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5.3.8 Body Mass Index
In the U.S. self-reported rates of obesity were 26.7% in 2009, up from 19.8% in 2000
(Sherry etal.. 2010). The NHANES III cohort was utilized in a study of blood Pb levels
and all-cause and cardiovascular mortality, which included assessment of the associations
by obesity (Menke et al., 2006). Positive associations were observed among individuals
within both categories of body mass index (BMI; normal [<25 kg/m2] and
overweight/obese [> 25 kg/m2], determined using measured values of height and weight)
but there was no difference in the association between the two categories. Using the
Normative Aging Study data, an investigation of bone Pb levels and heart rate variability
was performed and reported slight changes in the association based on the presence of
metabolic syndrome; however, none of the changes resulted in associations that were
statistically significant (Park et al.. 2006). Overall, no modification by BMI or obesity
was observed among recent epidemiologic studies, but the available epidemiologic and
supporting toxicological studies are limited.
5.3.9 Alcohol Consumption
There are a limited number of studies examining alcohol as a factor affecting Pb-related
risk. A study using the Normative Aging Study cohort investigated whether the
association between blood and bone Pb levels and renal function would be modified by
an individual's alcohol consumption (Tsaih et al., 2004). No interaction with alcohol
consumption was observed. However, a toxicological study reported that ethanol
potentiated the effect of Pb exposure by decreasing renal total protein sulfhydryls
(endogenous antioxidants) in rats. Pb and ethanol also decreased other endogenous renal
antioxidants (glutathione and non-protein sulfhydryls) (Jurczuk et al.. 2006). Overall,
evidence to determine if alcohol consumption is a potential at-risk factor is of limited
quantity and consistency.
5.3.10 Nutritional Factors
Different components of diet may affect the association between Pb concentrations and
health outcomes. Recent epidemiologic and toxicological studies of specific mineral
intakes/dietary components are detailed below but evidence is limited.
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5.3.10.1 Calcium
Using the Normative Aging Study (NAS) cohort, researchers examined the association
between Pb levels and hypertension, modified by calcium intake over the year before Pb
measurement (determined by a food frequency questionnaire) (Elmarsafawv et al.. 2006).
The associations between Pb levels (measured and modeled separately for blood, patella,
and tibia) and hypertension did not differ based on dichotomized calcium intake
(800 mg/day).
5.3.10.2 Iron
The 2006 Pb AQCD included studies that indicated individuals with Fe deficiency and
malnourishment had greater inverse associations between Pb and cognition (U.S. EPA.
2006b). A recent epidemiologic study of pubertal development among girls observed
inverse associations between blood Pb and inhibin B. This association was modified by
Fe deficiency; girls with Fe deficiency had a stronger inverse association between Pb and
inhibin B than those who were Fe sufficient (Gollenberg et al.. 2010). Toxicological
studies also reported that Fe-deficient diets exacerbate or potentiate the effect of Pb. A
study of pregnant rats given an Fe-deficient diet and exposed to Pb through drinking
water over GD6-GD14, had decreased litter size, more pups with reduced fetal weight
and reduced crown-rump length, increased litter resorption, and a higher dam blood Pb
level in the highest exposure groups (Singh et al.. 1993b; Saxenaetal.. 1991). Thus, in
this model, Fe deficiency makes rat dams more at risk for Pb-dependent embryo and
fetotoxicity (Singh et al.. 1993b).
5.3.10.3 Folate
A study by Kordas et al. (2009) examined Pb levels and birth size among term births in
Mexico City. The authors reported no interaction between maternal tibia Pb and folate
levels.
5.3.10.4 Protein
No recent epidemiologic studies have evaluated protein intake as a factor affecting
Pb-related health effects. However, a toxicological study demonstrated that differences in
maternal protein intake levels could affect the extent of Pb-induced immunotoxicity
among offspring (Chen et al.. 2004).
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In sum, the evidence is limited for most dietary factors but evidence for Fe deficiency as
a factor that potentially increases risk of Pb-induced effects is present and coherent in
epidemiologic and toxicological studies.
5.3.11 Stress
Recent studies have examined modification of the association between Pb and various
health effects by stress-level. Evidence demonstrates that higher stress levels may
increase the risk of Pb-related health effects.
A study of bone (tibia and patella) Pb levels and hypertension reported modification of
the association by perceived stress levels (Peters et al.. 2007). Among individuals with
greater perceived stress levels, stronger associations between blood Pb levels and
hypertension were present. Among the same study population, higher perceived stress
was also reported to affect the association between blood Pb levels and cognitive
function; the higher stress group showed a greater inverse association between Pb and
cognitive function than those in the low stress group (Peters et al., 2008). In another
study, the inverse association between tibia Pb levels and some measures of cognitive
function were similarly strengthened by neighborhood psychosocial hazards (Glass et al..
2009).
Toxicological studies have demonstrated that early life exposure to Pb and maternal
stress can result in toxicity related to multiple systems (Rossi-George et al.. 2009; Cory-
Slechtaetal.. 2008; Virgolini et al.. 2008a: Virgolini et al.. 2008b). including
dysfunctional corticosterone responses (Rossi-George et al.. 2009; Virgolini et al..
2008b). Additionally, toxicological studies have demonstrated that stressors to the
immune system can also affect associations with Pb exposure. Chickens with low Pb
exposure in ovo, with additional viral stressors, had increased immune cell mobilization
and trafficking dysfunction (Lee et al.. 2002). Similarly, mice with neonatal Pb exposure,
and an additional immune challenge, had a sickness behavior phenotype, likely driven by
IL-6 production (Dyatlov and Lawrence. 2002).
Although examined in a small number of studies, which limited the power to provide
coherence, recent epidemiologic studies observed modification of the association
between Pb and various nervous system health effects by stress-level. Increased risk of
Pb-related health effects by stress is further supported by toxicological studies.
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5.3.12 Maternal Self-Esteem
Maternal self-esteem has been shown to modify associations between blood Pb levels and
health effects in children. Surkan et al. (2008) studied the association between children's
blood Pb levels and Bayley's MDI and Psychomotor Development Index (PDI) among
mother-child pairs. High maternal self-esteem was independently associated with higher
MDI score and also appeared to attenuate the negative effects of the child's increased
blood Pb levels on MDI and PDI scores. Greater decreases in MDI and PDI were
associated with increased blood Pb levels among children whose mothers were in the
lower quartiles of self-esteem.
5.3.13 Cognitive Reserve
Cognitive reserve has been defined as "the maintenance of cognitive performance in spite
of ongoing underlying brain pathology" (Bleecker et al.. 2007a). A study of Pb smelter
workers reported that an inverse association between lifetime weighted blood Pb levels
and cognitive function was present among workers with low cognitive reserve (measured
using a reading achievement test) but no association was present in workers with high
cognitive reserve (Bleecker et al.. 2007a). Inverse associations between lifetime-weighted
blood Pb levels and motor functions existed among all workers regardless of cognitive
reserve. No other recent epidemiologic studies were performed examining cognitive
reserve as a factor affecting risk of Pb-related health outcomes, thus providing limited
evidence to conclude that cognitive reserve is a potential at-risk factor.
5.3.14 Other Metal Exposure
The 2006 Pb AQCD reported that the majority of studies that examined other toxicants
did so as confounders and not as effect measure modifiers (U.S. EPA. 2006b). Recent
epidemiologic studies have begun to explore the possible interaction between Pb
exposure and co-exposures with other metals and reported increased risk of Pb-related
outcomes. These studies, as well as toxicological studies of these metals, are described
below.
5.3.14.1 Cadmium
In a study of girls in the NHANES III cohort, inverse associations were observed
between blood Pb and inhibin B concentrations (Gollenberg et al.. 2010). These inverse
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associations were stronger among girls with high cadmium (Cd) and high Pb compared to
those with high Pb and low Cd. Additionally, higher blood Pb and Cd levels together
were positively associated with albuminuria and reduced estimated glomerular filtration
rate, compared to those with the lowest levels of Pb and Cd (Navas-Acien et al. 2009).
Toxicological studies reported that in rats, the addition of Cd to Pb exposure reduced the
histological signs of renal toxicity from each element alone; however, urinary excretion
of porphyrins were increased, indicating that although measured tissue burdens of Pb
were reduced, the biologically available fraction of Pb was actually increased (Wang and
Fowler. 2008). In other studies, Cd synergistically exacerbated Pb-dependent renal
mitochondrial dysfunction (Wang et al.. 2009c).
Overall, epidemiologic and toxicological studies have reported increased risk of
Pb-related health effects among those with high Cd levels as well; however, the number
of studies examining both metals is small.
5.3.14.2 Manganese
Among children in South Korea taking part in a study of IQ, an interaction was reported
between Pb and manganese (Mn) blood levels (Kim et al.. 2009b). Children with high
blood Mn levels were observed to have reductions in full scale IQ and verbal IQ
associated with increased blood Pb levels, whereas no association between blood Pb
levels and full scale IQ and verbal IQ were noted among those children with low blood
Mn levels. No effect measure modification by Mn was observed for the association
between blood Pb levels and performance IQ. A study performed among children in
Mexico City observed greater decreases in neurodevelopment with increases in blood
levels of Pb and Mn at 12 months, compared to decreases in neurodevelopment observed
for increased Pb levels with low levels of Mn (Claus Henn et al.. 2012). No interaction
was observed between the two metals and neurodevelopment at 24 months.
Overall, studies have reported increased risk of various health effects with exposure to
other metals in addition to Pb; however, this is limited by the small number of studies.
Toxicological studies, when available, have provided support for these findings.
5.4 Summary
Table 5-5 provides an overview of the factors examined as potentially increasing the risk
of Pb-related health effects based on the recent evidence integrated across disciplines.
They are classified according to the criteria discussed in the introduction to this chapter.
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Table 5-5 Summary of evidence for factors that potentially increase the risk of
Pb-related health effects.
Factor Evaluated
Childhood (Sections 5.2.1, 5.3.1)
Older Adulthood (Sections 5.2.1, 5.3.1)
Sex (Sections 5.2.2, 5.3.2)
Genetics (Section 5.3.3)
Pre-existinq Disease3 (Section 5.3.4)
Smokinq Status (Section 5.3.5)
Socioeconomic Status (SES) (Sections 5.2.4, 5.3.6)
Race/Ethnicitv (Sections 5.2.3, 5.3.7)
Proximity to Pb Sources (Section 5.2.5)
Residential Factors (Section 5.2.6)
Body Mass Index (BMI) (Section 5.3.8)
Alcohol Consumption (Section 5.3.9)
Nutrition (Section 5.3.10)
Stress (Section 5.3.11)
Maternal Self-Esteem (Section 5.3.12)
Coqnitive Reserve3 (Section 5.3.13)
Other Metals (Section 5.3.14)
Classification
Adequate
Suggestive
Suggestive
Suggestive
Suggestive
Inadequate
Suggestive
Adequate
Adequate
Adequate
Inadequate
Inadequate
Adequate
Suggestive
Inadequate
Inadequate
Suggestive
aPossible mediator
There are consistent findings, coherent across disciplines that adequate evidence exists to
conclude that childhood is an at-risk lifestage. Among children, the youngest age groups
were observed to be most at risk of elevated blood Pb levels, with levels decreasing with
increasing age of the children. Children may have increased exposure to Pb compared
with adults because children's behaviors and activities (including increased hand-to-
mouth contact, crawling, and poor hand-washing), differences in diets, and biokinetic
factors.
For adults, elevated Pb biomarkers were associated with increasing age. It is generally
thought that these elevated levels are related to remobilization of stored Pb during bone
loss and/or higher historical Pb exposures. Studies of older adults had inconsistent
findings for effect measure modification of Pb-related mortality but no difference was
observed for other health effects. However, toxicological studies support the possibility
of age-related differences in Pb-related health effects. The overall evidence is suggestive
that older adults are a potential at-risk population based on limited epidemiologic
evidence but support from toxicological studies and differential exposure studies.
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Some studies suggest that males at some ages have higher blood Pb levels than
comparably aged females; this was supported by stratifying the total sample of NHANES
subjects. Sex-based differences appeared to be prominent among the adolescent and adult
age groups but were not observed among the youngest age groups (1-5 years and 6-11
years). Studies of effect measure modification of Pb and various health endpoints by sex
were inconsistent; although it appears that there are some differences in associations for
males and females. This is also observed in toxicological studies. Overall, there is
suggestive evidence to conclude that sex is a potential at-risk factor, limited due to
inconsistencies between whether males or females are at greater risk of certain outcomes.
Regarding race and ethnicity, recent data suggest that the difference in blood Pb levels
between black and white subjects is decreasing over time, but black subjects still tend to
have higher Pb body burden and Pb exposures than white subjects. Compared to whites,
non-white populations were observed to be more at risk of Pb-related health effects (i.e.
hypertension, delayed puberty onset); however, this could be related to confounding by
factors such as SES or differential exposure levels, which was noted in some of the
epidemiologic studies. Studies of race/ethnicity provide adequate evidence that
race/ethnicity is an at-risk factor based on the higher exposure observed among non-white
populations and some modification observed in studies of associations between Pb levels
and health effects.
Similar to race and ethnicity, the gap between SES groups with respect to Pb body burden
appears to be diminishing. Studies of SES and its relationship with Pb-related health
effects are limited and different studies demonstrate increased risk among higher or lower
SES groups, providing limited evidence to determine if SES is an at-risk factor for
Pb-related health effects. However, biomarkers of Pb exposure have been shown to be
higher among lower SES groups even in recent studies in which differences among SES
groups have lessened. Therefore, the evidence is suggestive to conclude that low SES is a
potential at-risk factor for Pb-related health effects.
There is evidence associating proximity to areas with Pb sources, including areas with
large industrial sources, with increased Pb body burden and risk of Pb exposure. High
concentrations of ambient air Pb have been measured near sources, compared with large
urban areas without sources. Additionally, high Pb exposures have been documented near
Superfund sites.
Studies utilizing the NHANES dataset have reported increased Pb biomarker measures
related to increase house dust Pb levels, homes built after 1950, and renovation of
pre-1978 homes. These findings were consistent with those of several high quality
studies. Thus, there is adequate evidence that residing in a residence with Pb exposures
will increase the risk of Pb-related health effects.
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There is suggestive evidence to conclude that various genes are potentially modifying the
associations between Pb and health effects. Epidemiologic and toxicological studies
reported that ALAD variants may increase the risk of Pb-related health effects. Other
genes examined that may also affect risk of Pb-related health effects were VDR, DRD4,
GSTM1, TNF-a, eNOS, and HFE, although the number of studies examining effect
measure modification by these genes was small.
Among nutritional factors, diets sufficient in minerals such as Ca2+, Fe, and Zn offer
some protection from Pb exposure by preventing or competing with Pb for absorption in
the GI tract. Additionally, those with Fe deficiencies were observed to be an at-risk
population for Pb-related health effects in both epidemiologic and toxicological studies.
Thus, there is adequate evidence across disciplines that some nutritional factors
contribute to a population being at increased risk. Other nutritional factors, such as Ca2+,
Zn, and protein intake, demonstrated the potential to modify associations between Pb and
health effects in toxicological studies.
There was suggestive evidence for several other factors as potentially increasing the risk
of Pb-related health effects: pre-existing diseases/conditions, stress, and co-exposure with
other metals. Pre-existing diseases/conditions have the potential to affect the risk of
Pb-related health effects. Recent epidemiologic studies did not support modification of
associations between Pb and health endpoints by the prevalence of diabetes; however,
past studies have found individuals with diabetes to be an at-risk population with regard
to renal function. Hypertension was observed to be a factor affecting risk in both past and
recent epidemiologic studies. Studies of Pb levels and both renal effects and heart rate
variability demonstrated greater odds of the associations among hypertensive individuals
compared to those that are normotensive. Epidemiologic studies also examined autism as
potential factors affecting Pb-related health effects; differences were observed but few
studies were available to examine this factor. Stress was evaluated as a factor that
potentially increases the risk of Pb-related health outcomes and although limited by the
small number of epidemiologic studies, increased stress was observed to impact the
association between Pb and health endpoints. Toxicological studies supported this
finding. Finally, interactions between Pb and co-exposure with other metals were
evaluated in recent epidemiologic and toxicological studies of health effects. High levels
of other metals, such as Cd and Mn, were observed to result in greater effects for the
associations between Pb and various health endpoints but evidence was limited due to the
small number of studies.
Finally, there was inadequate evidence to conclude that smoking, BMI, alcohol
consumption, maternal self-esteem, and cognitive reserve are potential at-risk factors due
to limited quantities of studies regarding their effect on Pb-related health outcomes.
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Epidemiologic studies examining smoking as a factor potentially affecting risk reported
mixed findings. It is possible that smoking modifies the effects of only some Pb-related
health outcomes. In the limited number of studies, modification of associations between
Pb and various health effects (mortality and heart rate variability) was not observed for
BMI/obesity. Also, no modification was observed in an epidemiologic study of renal
function examining alcohol consumption as a modifier, but a toxicological study
supported the potential of alcohol to affect risk. Maternal self-esteem was examined in an
epidemiologic study and individuals with mothers who had lower self-esteem had greater
Pb-related decreases in MDI and PDI. An epidemiologic study evaluated cognitive
reserve as a modifier of the associations between Pb and cognitive and motor functions.
Cognitive reserve was an effect measure modifier for the association between Pb and
cognitive function but not motor function.
Common metrics that would allow estimation of the magnitude of increased risk,
conferred by the at- risk factors considered in this chapter, are generally unavailable.
There is variation in the Pb measures used (e.g. blood or bone Pb) and in population
groups (e.g. different age groupings for children). In addition, available studies show
differences in magnitudes of association that vary within and across each outcome
examined for an at-risk factor. For example, studies of race report no effect measure
modification by race for some outcomes (mortality, asthma, and scores on reading and
math examinations) (Miranda et al.. 2007a: Menke et al.. 2006; Joseph et al.. 2005).
whereas other outcomes (delayed puberty onset and hypertension) demonstrated effect
measure modification (Scinicariello et al.. 2010; Muntner et al.. 2005; Selevan et al..
2003). In sum, there is adequate evidence that several factors - childhood, race/ethnicity,
nutrition, residential factors, and proximity to Pb sources - confer increased risk of Pb-
related health effects. However, the evidence is not conducive to quantifying the
differences in magnitude of increased risk related to these factors.
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CHAPTER 6 ECOLOGICAL EFFECTS OF LEAD
This chapter synthesizes and evaluates the most policy-relevant science to help form the
foundation for the review of the secondary (welfare-based) NAAQS for Pb. The Clean
Air Act definition of welfare effects includes, but is not limited to, effects on soils, water,
wildlife, vegetation, visibility, weather, and climate, as well as effects on materials,
economic values, and personal comfort and well-being. This chapter discusses the effects
of Pb on ecosystem components and processes and is organized into five sections. The
introduction (Section 6.1) presents the organizing principles of this chapter and several
important general ecology concepts. An overview of fate and transport of Pb in
ecosystems including measured concentrations of this metal in various environmental
media (i.e., soil, water, sediment) is presented in Section 6.2. Section 6.3 reviews the
effects of Pb on terrestrial ecosystems; how soil biogeochemistry affects Pb
bioavailability, biological effects of Pb exposure and subsequent vulnerability of
particular ecosystems. A similar discussion of the effects of Pb on freshwater and
saltwater ecosystems is presented in Section 6.4. including water-only exposures and
sediment-related effects. The terrestrial, freshwater and saltwater sections each conclude
with an integrative synthesis of new evidence for Pb effects and causal determinations,
based on the synthesis of new evidence and findings from previous Pb AQCDs.
Section 6.5 summarizes the causal determinations. Areas not addressed here include
literature related to ingestion of Pb shot or pellets and studies that examine human health-
related endpoints which are described in other chapters of this document.
6.1 Introduction to Ecological Concepts
Metals, including Pb, occur naturally in the environment at measurable concentrations in
soils, sediments, and water. Organisms have developed adaptive mechanisms for living
with metals, some of which are required micronutrients (but not Pb). However,
anthropogenic enrichment can result in concentrations that exceed the capacity of
organisms to regulate internal concentrations, causing a toxic response and potentially
death. Differences in environmental chemistry may enhance or inhibit uptake of metal
from the environment, thus creating a spatial patchwork of environments that are at
greater risk than other environments. Similarly, organisms vary in their degree of
adaptation to, or tolerance of, the presence of metals. These fundamental principles of
how metals interact with organisms and ecosystems are described in detail in EPA's
Framework for Metals Risk Assessment (U.S. EPA. 2007c). This section introduces
critical concepts for understanding how Pb from atmospheric deposition may affect
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organisms, communities, and ecosystems. The sections that follow provide more detail
for how aquatic and terrestrial ecosystems respond to Pb and how environmental
chemistry interacts with organisms to affect exposure and uptake.
6.1.1 Ecosystem Scale, Function, and Structure
For this assessment, an ecosystem is defined as the interactive system formed from all
living organisms (biota) and their abiotic (chemical and physical) environment within a
given area (IPCC. 2007).The boundaries of what could be called an ecosystem are
somewhat arbitrary, depending on the focus of interest or study. Thus, the extent of an
ecosystem may range from very small spatial scales to, ultimately, the entire Earth
(IPCC. 2007). Ecosystems cover a hierarchy of spatial scales and can comprise the entire
globe, biomes at the continental scale, or small, well-circumscribed systems such as a
small pond (U.S. EPA. 2008e). A pond may be a small but complex system with multiple
trophic levels ranging from phytoplankton to several feeding guilds of fish plus fish-
eating birds or mammals. A large lake, on the other hand, may be a very simple
ecosystem, such as the Great Salt Lake in Utah that covers approximately 1,700 square
miles but contains only bacteria, algae, diatoms, and two invertebrate species. All
ecosystems, regardless of size or complexity, share the commonality of multiple
interactions between biota and abiotic factors, and a reduction in entropy through energy
flow from photosynthetic organisms to top predators. This includes both structural
(e.g., soil type and food web trophic levels) and functional (e.g., energy flow,
decomposition, nitrification) attributes. Changes are often considered undesirable if
important structural or functional components of ecosystems are altered following
pollutant exposure (U.S. EPA. 1998V
Ecosystems are most often defined by their structure, and are based on the number and
type of species present. Structure may refer to a variety of measurements including the
species richness, abundance, community composition, and biodiversity as well as
landscape attributes. Individual organisms of the same species are similar in appearance
and genetics, and can interbreed and produce fertile offspring. Interbreeding groups of
individual organisms within the same species that occupy some defined geographic space
form populations, and populations of different species form communities (Barnthouse et
al.. 2008). The community composition may also define an ecosystem type, such as a
pine forest or a tall grass prairie. Pollutants can affect the ecosystem structure at any of
these levels of biological organization (Suter et al., 2005). Individual plants or animals
may exhibit changes in metabolism, enzyme activities, hormone function, or overall
growth rates or may suffer gross lesions, tumors, deformities, or other pathologies.
Effects on the nervous system of animals may cause behavioral changes that alter
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breeding behaviors or predator avoidance. However, only some organism-level endpoints
such as growth, survival, and reproductive output have been definitively linked to effects
at the population level and above. Examples of organism-level endpoints with direct links
to population level effects include mass mortality, gross anomalies, survival, fecundity
and growth (Suteretal.. 2004). Population level effects of pollutants include changes
over time in abundance or density (number of individuals in a defined area), age or sex
structure, and production or sustainable rates of harvest (Barnthouse et al., 2008).
Community level attributes affected by pollutants include species richness and abundance
(also known as biodiversity), dominance of one species over another, or size (area) of the
community. Pollutants may affect communities in ways that are not observable in
organisms or populations (Bartell 2007). including: (1) effects resulting from interactions
between species, such as altering predation rates or competitive advantage; (2) indirect
effects, such as reducing or removing one species from the assemblage and allowing
another to emerge (Petraitis and Latham. 1999); and (3) alterations in trophic structure.
Alternatively, ecosystems may be defined on a functional basis. "Function" refers to the
suite of processes and interactions among the ecosystem components and their
environment that involve nutrient and energy flow as well as other attributes including
water dynamics and the flux of trace gases such as rates of photosynthesis,
decomposition, nitrification, or carbon cycling. Pollutants may affect abiotic conditions
(e.g., soil chemistry), which indirectly influences biotic structure and function (Bartell
2007). Feedback loops or networks influence the stability of the system, and can be
mathematically described through simplistic or complex process, or energy flow, models
(Bartell. 2007). For example, the Comprehensive Aquatic Systems Model (CASM) is a
bioenergetics-based multi-compartment model that describes the daily production of
biomass (carbon) by populations of aquatic plants and animals over an annual cycle
(DeAngelis et al.. 1989). CASM, originally designed to examine theoretical relationships
between food web structure, nutrient cycling, and ecosystem stability, has since been
adapted for risk assessments and has been applied to numerous lakes with a variety of
pollutants (Bartell 2007). Likewise, other theoretical ecosystem models are being
modified for use in assessing ecological risks from pollutant exposures (Bartell. 2007).
Some ecosystems, and some aspects of particular ecosystems, are less vulnerable to long-
term consequences of pollutant exposure. Other ecosystems may be profoundly altered if
a single attribute is affected. Thus, spatial and temporal definitions of ecosystem structure
and function become an essential factor in defining impacted ecosystem services and
critical loads of particular pollutants, either as single pollutants or in combination with
other stressors. Both ecosystem services (Section 6.1.2) and critical loads (Section 6.1.3)
serve as benchmarks or measures of the impacts of pollutants on ecosystems.
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6.1.2 Ecosystem Services
Ecosystem structure and function may be translated into ecosystem services (Daily.
1997). Ecosystem services are the benefits people obtain from ecosystems (UNEP, 2003).
Ecosystem services are defined as the varied and numerous ways that ecosystems are
important to human welfare and how they provide many goods and services that are of
vital importance for the functioning of the biosphere. This concept has gained recent
interest and support because it recognizes that ecosystems are valuable to humans, and
are important in ways that are not generally appreciated (Daily. 1997). Ecosystem
services also provide a context for assessing the collective effects of human actions on a
broad range of the goods and services upon which humans rely.
In general, both ecosystem structure and function play essential roles in providing goods
and services. Ecosystem processes provide diverse benefits including absorption and
breakdown of pollutants, cycling of nutrients, binding of soil, degradation of organic
waste, maintenance of a balance of gases in the air, regulation of radiation balance and
climate, and fixation of solar energy fWRI. 2000: Daily. 1997: Westman. 1977). These
ecological benefits, in turn, provide economic benefits and values to society (Costanza et
al.. 1997: Pimentel etal.. 1997). Goods such as food crops, timber, livestock, fish and
clean drinking water have market value. The values of ecosystem services such as flood
control, wildlife habitat, cycling of nutrients and removal of air pollutants are more
difficult to measure (Goulder and Kennedy. 1997).
Particular concern has developed within the past decade regarding the consequences of
decreasing biological diversity (Tilman. 2000: Ayensuetal.. 1999: Wall. 1999: Chapin et
al.. 1998: Hooper and Vitousek. 1997). Human activities that decrease biodiversity also
alter the complexity and stability of ecosystems and change ecological processes. In
response, ecosystem structure and function can be affected (Daily and Ehrlich. 1999:
Wall. 1999: Chapin etal. 1998: Levin. 1998: Peterson et al.. 1998: Tilman. 1996: Tilman
and Downing. 1994: Pimm. 1984). Biodiversity is an important consideration at all levels
of biological organization, including species, communities, populations, and ecosystems.
Human-induced changes in biotic diversity and alterations in the structure and
functioning of ecosystems are two of the most dramatic ecological trends of the past
century (U.S. EPA. 2004: Vitousek et al.. 1997).
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Hassan et al. (2005) identified four broad categories of ecosystem services:
• Supporting services are necessary for the production of all other ecosystem
services. Some examples include biomass production, production of
atmospheric O2, soil formation and retention, nutrient cycling, water cycling
and provisioning of habitat. Biodiversity is a supporting service in that it is
increasingly recognized to sustain many of the goods and services that humans
enjoy from ecosystems. These supporting services provide a basis for an
additional three higher-level categories of services.
• Provisioning services such as products (Gitay et al., 2001) i.e., food (including
game meat, roots, seeds, nuts, and other fruit, spices, fodder), water, fiber
(including wood, textiles) and medicinal and cosmetic products.
• Regulating services that are of paramount importance for human society such as
(1) carbon sequestration, (2) climate and water regulation, (3) protection from
natural hazards such as floods, avalanches, or rock-fall, (4) water and air
purification, and (5) disease and pest regulation.
• Cultural services that satisfy human spiritual and aesthetic appreciation of
ecosystems and their components.
6.1.3 Critical Loads as an Organizing Principle for Ecological Effects of
Atmospheric Deposition
A critical load is defined as, "a quantitative estimate of an exposure to one or more
pollutants below which significant harmful effects on specified sensitive elements of the
environment do not occur according to present knowledge" (Nilsson and Grennfelt.
1988). Critical loads are a powerful organizing principle for information that links
atmospheric deposition with ecological impairment. They allow for heterogeneity in
ecosystem sensitivity and exposure which often results in critical load values that vary by
ecosystem (e.g., aquatic-water; aquatic-sediment; terrestrial), and differ by endpoint of
concern. It is important to consider that critical loads are often calculated assuming
steady state conditions (i.e., how much input is required to balance the rate of output),
and there may be time required to reach the critical load (i.e., the lag time between onset
of exposure and induction of measurable effects). The following types of information are
required to calculate a critical load, each of which is discussed in more detail in the
subsequent sections of this chapter:
• Ecosystem at risk;
• Receptors of concern (plants, animals, etc.);
• Endpoints of concern (organism, population or community responses, changes
in ecosystem services or functions);
• Dose (concentration) - response relationships and threshold levels of effects;
• Bioavailability and bioaccumulation rates;
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• Naturally occurring (background) Pb (or other metal) concentrations; and
• Biogeochemical modifiers of exposure.
There is no single "definitive" critical load for a pollutant, partly because critical load
estimates reflect the current state-of-knowledge and policy priorities, and also because of
local or regional differences among ecosystems (U.S. EPA, 2008e). Changes in scientific
understanding may include, for example, expanded information about dose-response
relationships, better understanding of bioavailability factors, and improved quantitative
models for effects predictions. Changes in policy may include new mandates for resource
protection, inclusion of perceived new threats that may exacerbate the effects of the
pollutant of concern (e.g., climate change), and a better understanding of the value of
ecosystem services.
6.2 Fate and Transport of Pb in Ecosystems
Fate and transport of Pb in ecosystems are difficult to assess because Pb detected in the
environment could have multiple sources and passes through various environmental
media within a watershed. These issues are described in detail in Section 2.3. Pb can be
emitted to air, soil, or water and then cycle through any or all of these media. In addition
to primary emission of particle-bound or gaseous Pb to the atmosphere, Pb can be
resuspended to the air from soil or dust (Section 6.2.2). Additionally, Pb-bearing PM can
be deposited from the air to soil or water through wet and dry deposition. The
complicated nature of Pb fate and transport in ecosystems is illustrated in Figure 6-1 in
which the Venn diagram depicts how Pb can move through multiple environmental media
that encompass both terrestrial and aquatic systems (see also Figure 2-9). The
"air/soil/water" arrows illustrate Pb exposures to plants and animals. Many of the studies
presented in the subsequent material focus on observations of Pb exposure via one
medium: air, soil, sediment, or water.
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Newly Emitted Pb
Historically Emitted Pb
NATURAL WATERS
AND SEDIMENTS
OUTDOOR SOIL
AND DUST
Non-air Pb
eleases
AIR
SOIL
SEDIMENT
WATER
AIR
SOIL
SEDIMENT
WATER
PLANT
EXPOSURE
ANIMAL
EXPOSURE
Figure 6-1
Fate of atmospheric Pb in ecosystems.
6.2.1 Fate and Transport
This section provides a brief overview of the fate and transport of Pb in ecosystems. Fate
of Pb is determined by the chemical and physical properties of the medium in wet
deposition, bodies of water, or soil (e.g., pH, salinity, oxidation status, flow rate and the
suspended sediment load and its constituents). Desorption, dissolution, precipitation,
sorption and complexation processes can all occur concurrently and continuously, leading
to transformations and redistribution of Pb within a watershed. A primary factor
controlling fate and transport and the subsequent bioavailability of Pb in both aquatic and
terrestrial systems is solubility. In terrestrial systems, Pb is typically bound to organic
matter and soil. The most important factors in determining Pb solubility in soils are pH
and CEC (Smolders et al.. 2009) (Section 6.3.2). while in aquatic systems, solubility
varies primarily with pH and water hardness as well as Ca2+ concentration, total
suspended solids and DOC (Section 6.4.2). Therefore, the pH of water is of primary
importance in determining the likely chemical fate of Pb in terms of solubility,
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precipitation, or organic complexation. For more detailed information about the fate and
transport of Pb, please see Section 2.3.
Soluble Pb in air is mostly removed by wet deposition, and most of the insoluble Pb is
removed by dry deposition. As a result, dry deposition is the major removal mechanism
for Pb in coarse PM (which is mainly insoluble) and wet deposition is the most important
removal mechanism for fine PM and Pb halides (which were more soluble)
(Section 2.3.1). Recent research provides considerable evidence that appreciable amounts
of Pb can accumulate on coarse PM during transport, and that the physical and chemical
characteristics of Pb can be altered by this process due to accompanying transformations
(Section 2.3.1.1). Atmospheric removal of metals by wet or dry deposition is largely
controlled by solubility of Pb in rain water. The relative importance of wet and dry
deposition is highly variable with respect to location and season, probably reflecting both
variations in Pb speciation and variations in external factors such as pH and rain water
composition (Section 2.3.1.2).
Pb deposited to terrestrial ecosystems may remain in soils or eventually be transported in
runoff to streams, lakes or rivers in the watershed. Pb has a relatively long retention time
in the organic soil horizon, although its movement through the soil column also suggests
potential for contamination of groundwater (Section 2.3.3). Pb deposition to soils has
decreased since the phase-out of leaded on-road gasoline (Section 2.3.3.1). Recent studies
of metal concentrations in leaf litter and organic roadside debris suggest that the litter can
act as a temporary sink for metals from the soil around and below leaves on the ground
(Section 2.3.3.2). Leaching has been consistently observed to be a slower process for Pb
than for other contaminants because Pb is only weakly soluble in pore water, but
anthropogenic Pb is more available for leaching than naturally occurring Pb in soil
(Section 2.3.3.3). Overall, recent research confirms the generally low mobility of Pb in
soil. This limited mobility is strongly dependent on colloid amount and composition, as
well as pH, and may be greater in some contaminated soils. Low mobility allows soils to
act as a sink for atmospheric Pb potentially for decades or longer. Hence, atmospheric Pb
concentrations that peaked several decades ago may still be present in soil in the absence
of remediation.
Sources of Pb to surface waters include direct atmospheric deposition and indirect
deposition via runoff and industrial discharge (Section 2.3.2). Because dispersal in
waterways is a relatively rapid process, concentrations in surface waters are highest near
sources of pollution before substantial Pb removal by flushing and sedimentation occurs.
Transport in surface water is largely controlled by exchange with sediments, and the
cycling of Pb between water and sediments is governed by chemical, biological, and
mechanical processes that are affected by many factors, including salinity, organic
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complexation, oxidation-reduction potential, and pH. Metals in waterways are transported
primarily as soluble chelates and ions, or adsorbed on colloidal surfaces, including
secondary clay minerals, Fe and Mn oxides or hydroxides, and organic matter, and
adsorption on organic or inorganic colloids is particularly important for Pb. The extent of
sorption strongly depends on particle size as smaller particles have larger collective
surface areas. Pb is relatively stable in sediments, with long residence times and limited
mobility (Section 2.3.2.1). As described in previous sections, Pb enters and is distributed
in bodies of water largely in PM form. In rivers, particle-bound metals can often account
for > 75% of the total load (Section 2.3.2.2). The flux of Pb in aquatic ecosystems is
therefore influenced by the dynamic physical and chemical interactions within a
watershed.
Particles associated with runoff are mostly PM, with a relatively small dissolved fraction,
and dissolution of carbonate and related compounds are important contributors to Pb
pollution in runoff waters. Pb release into runoff is dependent on storm intensity and
length of dry periods between rain events. A "first flush effect" occurs with highest
runoff concentrations observed at the beginning of a rain event. Most recent studies have
concluded that, during storm events, Pb is transported together with large PM. Some
studies, however, found that Pb was concentrated in the fine PM fraction and,
occasionally, Pb was found predominantly in the dissolved fraction. Since the
2006 Pb AQCD, snowmelt and rain-on-snow events are better understood, and it has
been observed that greater runoff occurs from snowmelt and in rain on snow events than
when snow is not present, and that metals, including Pb, are often associated with coarse
PM under these circumstances. Runoff in rural areas is strongly controlled by soil type
and the presence of vegetation, with less runoff and greater retention in mineral soils or
when grass is present, and more runoff for soils high in organic matter (OM).
Sediments can be either a source or a sink for metals in the aquatic environment
(Section 2.3.2). Release can be via re-suspension of the sediment bed via wind, wave, and
tidal action or by dissolution from sediment to the water column. Sediment resuspension
from marine environments is important, with disturbance of bed sediments by tidal action
in estuarine areas resulting in a general greater capacity for re-suspension of PM. Recent
research on Pb flux from sediments in natural waters has demonstrated that resuspended
Pb is largely associated with OM or Fe and Mn particles, but that anoxic or depleted
oxygen environments in sediments play an important role in Pb cycling. This newer
research indicated that resuspension and release from sediments largely occurs during
discrete events related to storms. It has also confirmed that resuspension is an important
process that strongly influences the lifetime of Pb in bodies of water. Finally, there have
been important advances in understanding and modeling of Pb partitioning in complex
aquatic environments.
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6.2.2 Ecosystem Exposure, Lag Time and Re-entrainment of Historically
Deposited Pb
Ecosystem exposure from atmospheric emissions of Pb depends upon the amount of Pb
deposited per unit time. Ecosystem response will also depend upon the form in which the
Pb is deposited, the areal extent of such deposition, and modifying factors that affect Pb
bioavailability in soil, sediments, and water (e.g., pH, organic matter) (Sections 6.3.2.
6.4.2 and 6.4.3). However, there is frequently a lag in time between when metals are
emitted and when an effect is seen, particularly in terrestrial ecosystems and, to a lesser
extent, in aquatic sediments. This is because the buffering capacity of soils and sediments
permits Pb to become sequestered into organic matter, reducing its availability for uptake
by organisms. The lag time from start of emissions to achieving a critical load can be
calculated as the time to reach steady state after Pb was initially added to the system.
Excluding erosion processes, the time required to achieve 95% of steady state is about 4
half-lives (ti/2): (Smolders et al.. 2007). Conversely, once emissions cease, the same
amount of time is required to reduce metal concentrations to background levels.
Time to steady state for metals in soils depends upon rates of erosion, uptake by plants,
and leaching or drainage from soils. Ignoring erosion, half-life of metals can be predicted
(Smolders et al.. 2007) for a soil as:
_ 0.69 xdx 10,000
PKd
Equation 6-1
where:
d is the soil depth in meters (m)
y is the annual crop yield (tons/ha per year)
TF is the ratio of the metal concentration in plant to that in soil
R is the net drainage loss from the soil depth of concern (m3/ha per year)
P is the bulk density of soil [kg(dry weight)/L]
Kd is the ratio of the metal concentration in soil to that in soil pore solution (L/kg)
' Time required to reduce the initial concentration by 50% if metal input is zero.
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Metals removed by crops (or plants in general) comprise a very small fraction of the total
soil metal and can be ignored for the purpose of estimating time to steady state. Thus,
Equation 6-1 is simplified to:
0.69 xdx 10,000
Equation 6-2
and becomes a function of soil depth, the amount of rainfall, soil density, and soil
properties that affect Kd. Pb has a relatively long time to steady state compared to other
metals, as shown in Table 6-1.
Table 6-1 Comparison among several metals: Time to achieve 95% of steady
state metal concentration in soil; example in a temperate system.
Metal
Se
Cu
Cd
Pb
Cr
Loading rate (g/ha-yr)
100
100
100
100
100
0.3
480a
690a
19,000a
16,700a
aMean Kd (ratio of total metal concentrations in soils to that in
concentration in soil. (49 Dutch soils) (de Groot et al.. 1998).
Kd (L/kg)
1.3
1,860a
2,670a
73,300a
64,400a
soil pore water); and Time to achieve 95%
Time (years)
of steady-state
Note: Based on a soil depth of 25 cm, a rain infiltration rate of 3,000 m /ha per yr, and the assumption that background was zero at
the start of loading.
Source: Reprinted with permission of CRC Press, Smolders et al. (2007)
In aquatic systems, ti/2 for Pb in the water column depends on the ratio of the magnitudes
of the fluxes coming from and going into the sediment, the ratio of the depths of the
water column and sediment, and the sediment ti/2. Sediment ti/2 is dependent upon the
particulate and dissolved fractions and is calculated as for soils (Equation 6-2).
Re-entrainment of Pb particles via windblown dust from surface soils or dry sediments
may occur. Amount and distance of re-entrained particles and deposition rates are
dependent upon wind velocity and frequency; size, density, shape, and roughness of the
particle; soil or sediment moisture; and terrain features including openness (including
amount of vegetation), aspect relative to wind direction, and surface roughness.
Resuspension is defined in terms of a resuspension factor, K, with units of (meters'1), or a
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resuspension rate (A), with units of seconds"1 (Equation 6-3). The resuspension rate, A, is
the fraction of a surface contaminant that is released per time and is defined by:
R
A=c
Equation 6-3
where:
R is the upward resuspension flux ((ig/m2 per second),
C is the soil (or dry sediment) Pb concentration ((ig/m2).
Such emissions may have local impacts, but are not likely to have long-range effects, as
particles generally remain low to the ground and are not lifted into the atmosphere.
Although re-entrainment may alter the particle size distribution in a local area, it
generally does not alter the bioavailable fraction, and deposited particles will be subject
to the same biogeochemical forces affecting bioavailability. Therefore, exposure via
re-entrainment should be considered additive to exposure from atmospheric particulate
deposition in terrestrial and aquatic ecosystems.
6.2.3 Concentrations in Non-Air Media
Pb from multiple sources moves through environmental media as described in
Section 6.2.1 and Figure 6-1 and has led to measurable Pb concentrations in soil, water,
sediment and biota in terrestrial and aquatic ecosystems (Table 6-2). The highest
concentrations of Pb in the environment are currently found near Pb sources, such as
metal smelters and industrial processing. After phase-out of Pb from on-road gasoline, Pb
concentrations have decreased considerably in rain, snowpack and surface waters.
Declining Pb concentrations in tree foliage, trunk sections, and grasses, as well as surface
sediments and soils in some locations, have also been observed (U.S. EPA. 2006b). In
contrast, Pb is retained in soils and sediments, where it may provide a historical record of
deposition and associated concentrations. In remote lakes, sediment profiles indicate
higher Pb concentrations in near surface sediment as compared to pre-industrial era
sediment from greater depth, with peak concentrations between 1960 and 1980 (when
leaded on-road gasoline was at peak use).
Atmospheric deposition has led to measurable Pb concentrations observed in rain,
snowpack, soil, surface waters, sediments, agricultural plants, livestock, and wildlife.
Concentrations of Pb in moss, lichens, peat, and aquatic bivalves have been used to
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understand spatial and temporal distribution patterns of air Pb concentrations. The
amount of Pb in ecosystems is influenced by numerous factors, however, and it is not
currently possible to determine the contribution of atmospherically-derived Pb from total
Pb. Food, drinking water, and inhalation are likely major routes of exposure for livestock
and terrestrial wildlife. Ingestion and water intake are the major routes of Pb exposure for
aquatic organisms. In these exposure pathways, the bioavailable Pb may be from multiple
sources. Information on ambient Pb concentrations in non-air media and biota is reported
in Section 2.6. and concentrations considered in the interpretation of the ecological
evidence are tabulated in Table 6-2.
The most extensive survey of background soil Pb concentration in the contiguous U.S.
was conducted between 1961 and 1976, and comprised 1,319 non-urban, undisturbed
sampling locations, where 250 cm3 of soil was collected at a depth of 20 cm (Shacklette
and Boerngen. 1984). The lower detection limit was 10 mg Pb/kg, and 14% of the 1,319
samples were below it. The mean Pb concentration was 19.3 mg Pb/kg, the median 15 mg
Pb/kg, and the 95th percentile 50 mg/kg. Sixteen locations had Pb concentrations
between 100 and 700 mg Pb/kg. These results were in agreement with 3 previous
surveys. When creating the Ecological Soil Screening Level (Eco-SSL) guidance
document, the U.S. EPA (2007d_, 2003b) augmented these data with observations from an
additional 13 studies conducted between 1982 and 1997, most of them limited to one
state. The resulting data were summarized using state means for each of the fifty states.
Those state means ranged between 5 and 38.6 mg Pb/kg, with an overall national mean of
18.9 mg Pb/kg. No new data on background concentrations of Pb in U.S. soils have been
published since 2005.
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Table 6-2 Pb concentrations in non-air media and biota considered for
ecological assessment.
Media
Soil (non-urban)
Freshwater
Sediment
Saltwater
Sediment
Fresh Surface
Water
(Dissolved Pb)b
Saltwater0
Vegetation
Pb Concentration
Contiguous U.S. Median: 15 mg Pb/kg (dry weight)
Contiguous U.S. 95th Percentile:
50 mg Pb/kg (dry weight)
National Average: 18.9 mg Pb/kg (dry weight)
Range of state averages:
5-38.6 mg Pb/kg (dry weight)
Median: 73 mg Pb/kg (dry weight)
Median: 28 mg Pb/kgb (dry weight)
Range: 0.6 to 1,050 mg Pb/kga
Median: 0.50 ug Pb/Lb;
Max: 30 ug Pb/L, 95th percentile 1.1 ug Pb/L
Range: 0.0003-0.075 ug Pb/L
(Set of National Parks in western U.S.)
Range: 0.01-27 ug Pb/L
Lichens: 0.3-5 mg Pb/kg (dry weight)
(Set of National Parks in western U.S.)
Grasses:
Geometric Mean: 0.31 kg Pb/kg (dry weight)
Years Data
Obtained
1961-1976
1961-1997
1996-2001
1991-2003
Dates not
available
1991-2003
2002-2007
Dates not
available
2002-2007
1980s-2000s
References
Shaklette (1984)
U.S. EPA
(2007d, 2006b,
2003b)
Mahler et al.
(2006)
U.S. EPA
(2006b)
Sadiq (1992)
U.S. EPA
(2006b)
Field and
Sherrell (2003),
U.S. National
Park Service
(2011)
Sadiq (1992)
U.S. National
Park Service
(2011)
Vandenhove et
al. (2009)
Vertebrates
Fish:
Geometric Mean: 0.59 mg Pb/kg (dry weight)
(whole fish)
Geometric Mean: 0.15 mg Pb/kg (dry weight) (liver)
Range: 0.08-22.6 mg Pb/kg (dry weight) (whole fish)
Range: 0.01-12.7 mg Pb/kg (dry weight) (liver)
1991-2003
Fish (from a set of national parks in western U.S.):
0.0033 (fillet) to 0.97 (liver) mg Pb/kg (dry weight)
Moosed'e: 0.008-0.029 mg Pb/kg (dry weight) (meat)
Moosed'e: 0.012-0.023 mg Pb/kg (dry weight) (liver)
2002-2007
U.S. EPA
(2006b)
U.S. National
Park Service
(2011)
aNo information available regarding wet or dry weight
"Based on synthesis of National Water-Quality Assessment (NAWQA) data reported in 2006 Pb AQCD (U.S. EPA, 2006b)
°Data from a combination of brackish and marine saltwater samples. In general, Pb in seawater is higher in coastal areas and
estuaries since these locations are closer to sources of Pb contamination and loading from terrestrial systems.
d The reference cited and its source citations show that observations date from studies published in 1977-1990, indicating that the
data were obtained no later than those years. Further, these measurements seem to be for non-U.S. locations, including the max,
which is well above other reported values in these refs.
eThree moose in one Alaskan park
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The 2006 Pb AQCD reported representative median and range of Pb concentrations in
surface waters (median 0.50 (ig Pb/L, range 0.04 to 30 (ig Pb/L) and sediments (median
28 mg Pb/kg dry weight, range 0.5 to 12,000 mg Pb/kg dry weight) in the U.S. based on a
synthesis of National Water Quality Assessment (NAWQA) data (U.S. EPA. 2006c). In
an additional study using data collected from 1996-2001 the median Pb concentration in
sediment was reported to be 73 mg Pb/kg dry weight (Mahler et al.. 2006). A range of
0.01 to 27 (ig Pb/L for saltwater was reported by Sadiq although the values are not
specific to the U.S. and include open sea areas as well as estuarine and coastal waters
(Sadiq. 1992). In general, Pb in seawater is higher in coastal areas and estuaries since
these locations are closer to sources of Pb contamination and loading from terrestrial
systems (Sadiq. 1992).
Measured concentrations of Pb in soils, sediment and water are not necessarily
representative of the amount of Pb that is bioavailable to plants, invertebrates and
vertebrates. Both bioaccessibility and bioavailability (Sections 6.3.3. 6.4.3. and 6.4.11) of
Pb are dependent upon the physical, chemical, and biological conditions under which an
organism is exposed at a particular geographic location. Experimental exposures may be
difficult to compare with exposures under natural field conditions in terrestrial and
aquatic systems where a variety of abiotic and biotic modifying factors affect Pb toxicity.
6.3 Terrestrial Ecosystem Effects
6.3.1 Introduction to Effects of Pb on Terrestrial Ecosystems
Numerous studies of the effects of Pb on components of terrestrial systems were
reviewed in the 1977 Pb AQCD, the 1986 Pb AQCD and the 2006 Pb AQCD. The focus
of the present document is on studies published since the last AQCD. Many of those
studies were conducted near stationary sources of atmospheric Pb such as metal
industries and mines, or using soil collected near those sources. Increasing proximity to
the source was often used to generate a gradient of increasing exposure. As may be
expected, concentrations found in close proximity to those sources are many times
greater than those found at most locations around the country (data on concentrations of
Pb in U.S. soils are reviewed in Section 6.2.3 and summarized in Table 6-2). and as
indicated in the present document's Preamble, concentrations within one to two orders of
magnitude of current conditions were considered. In addition, it is important to note that
in all studies where a gradient of multiple concentrations was used, effects increased with
increasing concentration. This is an important aspect in determining causality (see
Preamble), and therefore justifies inclusion of some studies with very high exposures.
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Inclusion of those studies also provides potential data for establishing dose-response
relationships, and predicting effects at all concentrations, including those found away
from stationary sources. Finally, some studies at very high concentrations were used to
provide mechanistic information on Pb toxicity, allow for comparison of Pb uptake
across taxa, or demonstrate the wide range of sensitivity among closely-related species.
Concentrations used in studies where Pb was added to soil experimentally are difficult to
relate to concentrations found in natural environments that have been exposed to Pb
pollution. As reviewed in the following sections, there is ample evidence that multiple
factors, many of them known but not quantified, interact with Pb concentration to
produce responses of widely varying magnitude for similar concentrations, or similar
responses for varying concentrations of Pb. Thus, experimental concentrations that
appear relatively low may be most comparable to relatively high concentrations in natural
soils, and vice-versa. The various factors that interact with Pb concentration, and the
evidence for those interactions, are discussed in the following sections. However, the
same justifications for inclusion apply to added-Pb experiments as they do to studies
where proximity to sources is used to vary exposure: gradients of Pb concentrations
create gradients of response, and they often provide information on underlying
mechanisms of toxicity even if the concentrations cannot be easily compared to natural
ones.
The literature on terrestrial ecosystem effects of Pb published since the 2006 Pb AQCD,
is considered with brief summaries from the 1977 Pb AQCD, the 1986 Pb AQCD and the
2006 Pb AQCD, where relevant. Section 6.3 is organized to consider uptake of Pb and
effects at the species level, followed by community and ecosystem level effects. Recent
evidence for Pb effects on reproduction, growth, and survival in terrestrial plants,
invertebrates, and vertebrates is summarized in Table 6-4. Alterations to reproduction,
growth, and survival of terrestrial organisms can lead to changes at the community and
ecosystem levels of biological organization such as decreased abundance, reduced taxa
richness, and shifts in species composition (Section 6.1). Soil biogeochemistry of Pb is
reviewed in Section 6.3.2. Section 6.3.3 considers the bioavailability and uptake of Pb by
plants, invertebrates, and vertebrates in terrestrial systems. Biological effects of Pb on
terrestrial ecosystem components including plants and lichen, invertebrates, and
vertebrates (Section 6.3.4) are followed by data on exposure and response of terrestrial
species (Section 6.3.5). Effects of Pb at the ecosystem level of biological organization are
discussed in Section 6.3.6. Section 6.3 concludes with a discussion of critical loads in
terrestrial systems (Section 6.3.7). soil screening levels (Section 6.3.8). characterization
of sensitivity and vulnerability of ecosystem components (Section 6.3.9). and effects on
ecosystem services (Section 6.3.10). Concentration of Pb in soil is expressed in mg Pb/kg
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soil, and concentration in solutions applied to soil or extracted from soil is expressed in
mg Pb/L solution.
6.3.2 Soil Biogeochemistry and its Influence on Bioavailability
According to data presented in the 2006 Pb AQCD (U.S. EPA. 2006b). the fraction of
soil metal that is directly available to plants is the fraction found in soil pore water, even
though the concentration of metals in pore water is generally small relative to bulk soil
concentration. At any given bulk soil concentration, the amount of Pb dissolved in soil
solution is controlled by at least six variables: (1) solubility equilibria; (2) adsorption-
desorption relationship of total Pb with inorganic compounds (e.g., oxides of Al, Fe, Si,
Mn; clay minerals); (3) adsorption-desorption reactions of dissolved Pb phases on soil
organic matter; (4) pH; (5) cation exchange capacity (CEC); and (6) aging. Adsorption-
desorption of Pb to soil solid phases is largely controlled by total metal loading.
Therefore, areas with high Pb deposition will exhibit a lower fraction of total Pb
partitioned to inorganic and organic matter. Decreasing soil pH, CEC, and organic matter
have been strongly correlated to increases in the concentration of dissolved Pb species.
Aging of metals in soils results in decreased amounts of labile metal as the Pb becomes
incorporated into the soil solid phase (McLaughlin et al.. 2010). Data from recent studies
have further defined the impact of pH, CEC, organic matter (OM), and aging on Pb
mobilization and subsequent bioavailability in soils.
6.3.2.1 pH, CEC and Salinity
Models of metal bioavailability calibrated from 500+ soil toxicity tests on plants,
invertebrates, and microbial communities indicated that soil pH and CEC are the most
important factors governing both metal solubility and toxicity (Smolders et al.. 2009).
The variability of derived EC50 values was most closely associated with CEC. Smolders
et al. (2007) determined that 12 to 18 months of artificial aging of soils amended with
metal decreased the soluble metal fraction by approximately one order of magnitude.
Relatedly, lower soil pH in forest environments relative to adjacent agricultural land
resulted in higher solubility, and the mobility of smelter-produced metals was found to be
greater in forest than in agricultural lands (Douay et al.. 2009). Further, decreasing the
soil pH via simulated acid rain events increased naturally occurring Pb bioavailability in
field tests (Hu et al.. 2009b). Miretzky et al. (2007) also showed that the concentration of
mobile Pb was increased in acidic soils, and discovered that Pb adsorption to sandy loam
clay was a function of weak electrostatic bonds with charged soil surfaces and was
influenced by Fe and Mn oxide. Dayton et al. (2006) and Bradham et al. (2006) used path
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analysis to help identify the main determinants of both organism Pb content and
responses from among multiple soil characteristics. In parallel studies with lettuce and
earthworms, they amended an array of 21 soils with varying characteristics with the same
amount of Pb (2,000 mg /kg as Pb nitrate), and found that in lettuce, the main
determinant of both accumulation and biological responses was OC, with contribution
from pH and Fe/Al oxides. These latter characteristics only influenced accumulation and
responses through their own impact on CEC. In earthworms, the main determinant of
accumulation was pH, with contribution from CEC, but only through its association with
other variables including OC, Fe/Al oxides, and pH. The main determinant of
reproductive effects in earthworms was Fe/Al oxides, while pH drove differential
mortality between the various soils.
Salinity can also alter Pb mobility and bioavailability in soils. Application of CaCl2,
MgCl, or NaCl salts to field-collected soils containing 31 to 2,764 mg Pb/kg increased
the proportion of mobile metal. As the strength of the salt application was increased from
0.006 to 0.3 M, the proportion of released Pb increased from less than 0.5% to over 2%
for CaCl2 and from less than 0.5% to over 1% for MgCl (Acostaetal.. 2011). However,
the majority of salinity-induced effects occurred in soils containing less than 500 mg
Pb/kg, and the proportion of released Pb decreased with increasing total soil Pb
concentrations. In addition, the authors noted that Pb release from soils under increasing
salinity was reduced at higher carbonate concentrations, indicating that the effect of soil
salinity on Pb release is dependent on still other soil factors. A sequential extraction
procedure was employed by Ettler et al. (2005) to determine the mobility of different Pb
fractions present in soils collected from a mining and smelting area in the Czech
Republic. Five Pb fraction categories were identified: (Fraction A) exchangeable,
(Fraction B) acid extractable (bound to carbonates), (Fraction C) reducible (bound to Fe
and Mn oxides), (Fraction D) oxidizable (complexed with organic carbon), and (Fraction
E) residual (silicates). Tilled agricultural soils were found to have decreased Pb, likely as
a result of repeated cultivation, with the majority of Pb represented as the reducible
Fraction C. Pb concentration in undisturbed forest soils, however, was largely present as
the exchangeable fraction (A), weakly bound to soil OM. However, the validity of
associating sequentially extracted fractions with discrete geochemical components has
not been definitively established, and as a consequence, the association between
fractionation and bioavailability remains uncertain.
6.3.2.2 Organic Matter
Organic matter decreases bioavailability of Pb, but as it is turned over and broken down,
pedogenic minerals become more important in Pb sequestration (Schroth et al.. 2008).
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Shaheen and Tsadilas (2009) noted that soils with higher clay content, organic matter,
total calcium carbonate equivalent, and total free sesquioxides also exhibited higher total
Pb concentration, indicating that less Pb had been taken up by resident plant species.
Huang et al. (2008) examined the re-mobilization potential of Pb in forest soils, and
determined that mobilization of total Pb was strongly associated with dissolved organic
matter (DOM). Groenenberg et al. (2010) used a non-ideal competitive adsorption
Donnan model to explain the variability of organic matter binding affinity and
uncertainties associated with metal speciation. They found that natural variations in fulvic
acid binding properties were the most important variable in predicting Pb speciation. Guo
et al. (2006b) determined that the -COOH and -OH groups associated with soil OM were
important factors in Pb sequestration in soil, and Pb sorption was increased as pH was
raised from 2 to 8. Because organic content increased the Pb sequestration efficiency of
soils, OM content had an inhibitory effect on Pb uptake by woodlouse species Oniscus
asellus and Porcellio scaber (Gal et al.. 2008). Vermeulen et al. (2009) demonstrated that
invertebrate bioaccumulation of Pb from contaminated soils was dependent on pH and
OM, but that other unidentified habitat-dependent factors also contributed. The
relationship of bioaccumulation and soil concentration was modified by pH and OM, and
also by habitat type. Kobler et al. (2010) showed that the migration of atmospherically
deposited Pb in soil matrices was strongly influenced by soil type, indicating that certain
soil types may retain Pb for longer periods of time than others. In soils characterized by
well-drained substrate and limestone bedrock, Pb concentration decreased over time,
likely as a result of water drainage and percolation. The authors contrasted this
observation with reports of prolonged residence time in humic soils, particularly at the
lower depths of the humus layer. They theorized that the most significant Pb migration
route was transportation of particulate-bound Pb along with precipitation-related flow
through large soil pores.
A number of recent laboratory studies have further defined the relationship of soil
biogeochemical characteristics and Pb uptake by plants. As noted above, Dayton et al.
(2006) found through path analysis that the main determinant of both accumulation and
biological responses in lettuce grown on amended soil was OC. As part of a metal
partitioning study, Kalis et al. (2007) determined that not only did metal concentration in
the soil solution decrease as pH increased, but pH-mediated metal adsorption at the root
surface ofLolium perenne determined root Pb concentration, with concentration in the
shoot correlated with root concentration. Interestingly, Kalis et al. (2007) and Lock et al.
(2006) also observed that the influx of Pb in the water-soluble fraction had an impact on
soil pH. In addition, 1 (iM humic acid decreased root Pb concentration in L. perenne
plants grown in 0.1 and 1 (iM Pb solution, likely as a result of Pb complexation and
sequestration with the added OM (Kalis et al.. 2006). Ma et al. (2010) also reported that
long-term agricultural cultivation can decrease the rate of Pb desorption in soil through a
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gradual OM-enrichment. Phosphorous soil amendments equivalent to 35 mg P/kg soil
were observed to reduce the quantity of DPTA-extractable Pb from an average of 19 and
24 mg Pb/kg in unamended soils to 12 to 15 mg Pb/kg in P-amended soils. As a result,
maize and soybean seedlings accumulated significantly less Pb: average concentrations in
soybean shoot and root ranged from 4.4 to 5.2 mg Pb/kg with P addition (versus 9.21 mg
Pb/kg without), while maize shoot concentrations average between 4.8 to 5.3 mg Pb/kg in
P-amended soils (as compared with 10.16 mg Pb/kg in controls) (Xie et al., 2011).
6.3.2.3 Aging
Smolders et al. (2007) defined aging as the process responsible for decreasing the
bioavailability of metals in soils independently of their persistence. Smolders et al. (2009)
reviewed the effects of aging of Pb in soils on the toxicity of Pb to plants and soil
invertebrates, with aging achieved in most studies primarily by leaching amended soil,
but also through natural binding and complexation. In nearly half of the Pb soil studies
reviewed, responses that were observed with freshly amended soil could no longer be
detected following soil leaching, indicating that aged soils likely contain less bioavailable
Pb. The authors concluded that competitive binding between soil ligands and biotic
ligands on plant roots or invertebrate guts can be used to model the relationship of
observed availability and toxicity of metals in soils. Because this concept is the basis of
the Biotic Ligand Model (BLM) (Section 6.3.3). the authors proposed a terrestrial BLM
approach to estimate the risk of metals to terrestrial organisms. However, Antunes et al.
(2006) noted that there were several key challenges involved in development of a
terrestrial BLM applicable to plants, particularly the reliable measurement of free ion
activities and ligand concentration in the rhizosphere, the identification of the organisms'
ligands associated with toxicity, and the possible need to incorporate kinetic dissolution
of metal-ligand complexes as sources of free ion. Further, Pb in aged field soils has been
observed to be less available for uptake into terrestrial organisms, likely as a result of
increased sequestration within the soil particles (Antunes et al.. 2006). Magrisso et al.
(2009) used a bioluminescent strain of the bacterium Cupriavidus metallidurans to detect
and quantify Pb bioavailability in soils collected adjacent to industrial and highway areas
in Jerusalem, Israel, and in artificial preparations corresponding to separate soil
components (carbonates, Fe-oxides, clays, organic matter, and quartz) freshly spiked with
varying concentrations of Pb. The bacterium was genetically engineered to give off the
bioluminescent reaction as a dose-dependent response, and was inoculated in soil slurries
for three hours prior to response evaluation. Spiked soil components induced the
bioluminescent response, and field-collected components did not. However, the
comparability of the simulated soils and their Pb concentration with the field-collected
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samples was not entirely clear. Lock et al. (2006) compared the Pb toxicity to springtails
(Folsomia Candida) from both laboratory-spiked soils and field-collected
Pb-contaminated soils of similar Pb concentrations. Total Pb concentrations of 3,877 mg
Pb/kg dry weight and higher always caused significant effects on F. Candida reproduction
in the spiked soils. In field soils, only the soil with the highest Pb concentration of
14,436 mg Pb/kg dry weight significantly affected reproduction. When expressed as soil
pore-water concentrations, reproduction was never significantly affected at Pb
concentrations of 0.5 mg Pb/L, whereas reproduction was always significantly affected at
Pb concentrations of 0.7 mg Pb/L and higher, independent of the soil treatment. Leaching
soils prior to use in bioassays had only a slight effect on Pb toxicity to resident
springtails, suggesting that among the processes that constitute aging of Pb in field soils,
leaching is not particularly important with respect to bioavailability.
Red-backed salamanders (Plethodon cinereus) exposed to Pb-amended soils (553 mg
Pb/kg, 1,700 mg Pb/kg, 4,700 mg Pb/kg, and 9,167 mg Pb/kg) exhibited lowered appetite
and decreased white blood cell counts at the two highest concentrations, as compared to
controls (Bazar etal. 2010). However, salamanders tolerated field-collected, aged soils
containing Pb concentration of up to 16,967 mg Pb/kg with no significant deleterious
effects.
In summary, studies published during the past 5 years continue to substantiate the
important role that soil geochemistry plays in sequestration or release of Pb. Soil pH and
CEC have long been known to be the primary controlling factors of the amount of
bioavailable Pb in soils, and a recent review of more than 500 studies corroborates these
findings (Smolders et al.. 2009). Fe and Mn oxides are now known to also play an
important role in Pb sequestration in soils. Pb binds to OM, although relatively weakly,
and as the OM is broken down the Pb may be released into soil solution. Leaching of
metal through soil pores may be the primary route for loss of bioavailable soil Pb; OM
may reduce leaching and thus appear to be associated with Pb sequestration. Aging of Pb
in soils (through incorporation of the metal into the particulate solid-phase of the soil)
results in long term binding of the metal and reduced bioavailability of Pb to plants and
soil organisms.
6.3.3 Bioavailability in Terrestrial Systems
Bioavailability was defined in the 2006 Pb AQCD as "the proportion of a toxin that
passes a physiological membrane (the plasma membrane in plants or the gut wall in
animals) and reaches a target receptor (cytosol or blood)" (U.S. EPA, 2006c). In 2007,
EPA took cases of bioactive adsorption into consideration and revised the definition of
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bioavailability as "the extent to which bioaccessible metals absorb onto, or into, and
across biological membranes of organisms, expressed as a fraction of the total amount of
metal the organism is proximately exposed to (at the sorption surface) during a given
time and under defined conditions" (U.S. EPA. 2007c). The bioavailability of metals
varies widely depending on the physical, chemical, and biological conditions under
which an organism is exposed (U.S. EPA. 2007c). Characteristics of the toxicant itself
that affect bioavailability are: (1) chemical form or species, (2) particle size, (3) lability,
and (4) source. The bioavailability of a metal is also dependent upon the fraction of metal
that is bioaccessible. As stated in the Framework for Metals Risk Assessment (U.S. EPA.
2007c_), the bioaccessible fraction of a metal is the portion (fraction or percentage) of
environmentally available metal that actually interacts at the organism's contact surface
and is potentially available for absorption or adsorption by the organism. The Framework
states that "the bioaccessibility, bioavailability, and bioaccumulation properties of
inorganic metals in soil, sediments, and aquatic systems are interrelated and abiotic
(e.g., organic carbon) and biotic (e.g., uptake and metabolism) modifying factors
determine the amount of an inorganic metal that interacts at biological surfaces (e.g., at
the gill, gut, or root tip epithelium) and that binds to and is absorbed across these
membranes. A major challenge is to consistently and accurately measure quantitative
differences in bioavailability between multiple forms of organic metals in the
environment." A conceptual diagram presented in the Framework for Metals Risk
Assessment (U.S. EPA. 2007c) summarizes metals bioavailability and bioaccumulation
in aquatic, sediment and soil media (Figure 6-2).
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Bioaccessible Fraction (BF) ":
Percent soluble metal ion
concentration relative to total
metal concentration (measured in
solution near biomembrane)
Relative Bioavai lability (RBA) b:
Percent adsorbed or absorbed
compared to reference material
(measure of membrane dynamics)
Absolute Bioavailability (ABA) c:
Percent of metal mass absorbed
internally corrpared to external
exposure (measures systerric
uptake/accumulation)
Bioaccessibility
X
io\vailability
Environmental availability \
Exposure ==
Bioaccumulation of metal
==============-» Efects
lembrane
uptake
\
Physiological
V^rnembrane
Total Metal Concentration
Predation
Foraging
Toxicoloical
accumulation
Detoxification
and Storage
Benign /
accumulation
Internal r
Transport
and
Distribution
aBF is most often measured using in vitro methods (e.g., artificial stomach), but it should be validated by in vivo methods.
bRBA is most often estimated as the relative absorption factor, compared to a reference metal salt (usually calculated on the basis of
dose and often used for human risk, but it can be based on concentrations).
°ABA is more difficult to measure and used less in human risk; it is often used in ecological risk when estimating bioaccumulation or
trophic transfer.
Source: ERG (2004) and U.S. EPA (2007c).
Figure 6-2 Conceptual diagram for evaluating bioavailability processes and
bioaccessibility for metals in soil, sediment, or aquatic systems.
The BLM attempts to integrate the principal physical and chemical variables that
influence Pb bioavailability. The model considers the reactions of Pb with biological
surfaces and membranes (the site of action) to predict the bioavailability and uptake of
the metal (Figure 6-3). and integrates the binding affinities of various natural ligands and
the biological uptake rates of organisms to predict both the bioaccessible and bioavailable
fraction of Pb in the environment, and to determine the site-specific toxicity of the
bioavailable fraction. In principle, the BLM can be used for determining toxicity in water,
sediment, and soil media, however, the parameter values that influence BLM are, in
general, characterized to a greater extent in aquatic systems than in terrestrial systems
(Section 6.4.4).
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Organic Matter
Complexation
Site of Action
Inorganic Ligand
Complexation
Source: Reprinted with permission of John Wiley and Sons; from Di Toro et al. (2001)
Figure 6-3 Schematic diagram of the biotic ligand model.
New information on sources of Pb in terrestrial ecosystems, and their influence on
subsequent bioavailability, was reviewed in Chapter 2. while new information on the
influence of soil biogeochemistry on speciation and chemical lability was presented in
Section 6.3.2. This section summarizes recent literature on uptake and subsequent
presence of Pb in tissues. The 2006 Pb AQCD (U.S. EPA. 2006b) extensively reviewed
the methods available for quantitative determination of the mobility, distribution, uptake,
and fluxes of atmospherically delivered Pb in ecosystems, and they are not reviewed in
this section. The 2006 Pb AQCD also reported bioaccumulation factors (BAF) and
bioconcentration factors (BCF). BAF is defined as the field measurement of metal
concentration in tissues, including dietary exposures, divided by metal concentration in
environmental media (Smolders et al.. 2007). BCF is defined as the same measurement
carried out in artificial media in the laboratory that does not include dietary exposure
(Smolders et al.. 2007). The EPA Framework for Metals Risk Assessment states that the
latest scientific data on bioaccumulation do not currently support the use of BCFs and
BAFs when applied as generic threshold criteria for the hazard potential of metals (U.S.
EPA. 2007c).
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6.3.3.1 Terrestrial Plants
At the time of the 1977 Pb AQCD, it was understood that Pb uptake in plants was
influenced by plant species and by the available Pb pool in the soils (U.S. EPA. 1977).
The role of humic substances in binding Pb was better characterized by the 1986 Pb
AQCD where it was stated that most plants cannot survive in soil containing
10,000 (ig Pb/g (mg Pb/kg) dry weight if the pH is below 4.5 and the organic matter
content is below 5% (U.S. EPA. 1986b). At the time of the 1986 AQCD, it was thought
that Pb can be absorbed across the leaf surface into internal plant tissues, but that the vast
majority of uptake is via roots (U.S. EPA. 1986b). The 2006 Pb AQCD (U.S. EPA.
2006b) noted that terrestrial plants accumulate atmospheric Pb primarily via two routes:
direct stomatal uptake into foliage, and incorporation of atmospherically deposited Pb
from soil into root tissue, followed by variable translocation to other tissues. Foliar Pb
may include both incorporated Pb (i.e., from atmospheric gases or particles) and surficial
particulate Pb deposition. Although the plant may eventually absorb the surficial
component, its main importance is its likely contribution to the exposure of plant
consumers. This section will first review recent studies on uptake of Pb by plants through
foliar and soil routes, and their relative contribution, followed by the consideration of
translocation of Pb from roots to shoots, including a discussion of variability in
translocation among species. Data on ambient Pb levels associated with vegetation are
summarized in Section 2.6.6.
Leaf and Root Uptake
Although Pb is not an essential metal, it is taken up from soils through the symplastic
route, the same active ion transport mechanism used by plants to take up water and
nutrients and move them across root cell membranes (U.S. EPA. 2006c). As with all
nutrients, only the proportion of a metal present in soil pore water is directly available for
uptake by plants. In addition, soil-to-plant transfer factors in soils enriched with Pb have
been found to better correlate with bioavailable Pb soil concentration, defined as DTPA-
extractable Pb, than with total Pb concentration (U.S. EPA. 2006c). Since the publication
of the 2006 Pb AQCD, suggestive evidence has become available that a substantial
proportion of Pb accumulated in shoots of some species of trees originates in direct leaf
uptake of atmospheric Pb. Evidence for such direct uptake is weaker in herbaceous
plants, and all data came from locations near stationary sources.
Field studies carried out in the vicinity of Pb smelters have determined the relative
importance of direct foliar uptake and root uptake of atmospheric Pb deposited in soils.
Hu and Ding (2009) analyzed ratios of Pb isotopes in the shoots of commonly grown
vegetables and in soil at three distances from a smelter (0.1, 0.2, 5.0 km). Pb isotope
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ratios in plants and soil were different at two of those locations, leading the authors to the
conclusion that airborne Pb was being assimilated via direct leaf uptake. Soil Pb
concentration in the rhizosphere at the three sites ranged between 287 and 379 mg Pb/kg
(Site I), 155 and 159 mg Pb/kg (Site II), and 58 and 79 mg Pb/kg (Site III, selected as the
control site). The median shoot and root Pb concentrations at each site were 36 and
47 mg Pb/kg, 176 and 97 mg Pb/kg, and 1.3 and 7 mg Pb/kg, respectively, resulting in
shootroot Pb ratios exceeding 1.0 in Site I (for Malabar spinach \Basella alba],
ratio = 1.6, and amaranth \Amaranthus spinosus], ratio = 1.1), and in Site II (for the
weeds Taraxacum mongolicum, ratio = 1.9, andRostellariaprocumbens, ratio = 1.7).
However, the two species studied at Site II were not studied at Site I or Site III. In the
control site (Site III), no plant was found with a Pb shootroot ratio greater than 1.0. Hu
and Ding (2009) concluded that metal accumulation was greater in shoot than in root
tissue, which suggested both high atmospheric Pb concentration and direct stomatal
uptake into the shoot tissue.
Cui et al. (2007) studied seven weed species growing in the vicinity of an old smelter
(average soil Pb concentration of 4,020 mg Pb/kg) in Liaoning, China, to measure Pb
accumulation rates in roots and shoots. Cutleaf groundcherry (Physalts angulata)
accumulated the most Pb, with root and shoot concentration of 527 and 331 mg Pb/kg,
respectively, and velvetleaf (Abutilon theophrasti) was the poorest absorber of Pb (root
and shoot concentration of 39 and 61 mg Pb/kg, respectively). In all cases, weed species
near the smelter accumulated more Pb than plants from non-polluted environments (5 mg
Pb/kg), indicating that aerially deposited Pb produced by smelting is bioavailable to
plants. However, the ratio of rootshoot Pb concentration varied by species, and the
authors presented no data to differentiate Pb taken up from soil from Pb incorporated via
foliar uptake. Angelova et al. (2010) examined Pb uptake by rapeseed plants (Brassica
napus) grown in heavy metal contaminated soils 0.5 km and 15 km from the Non-Ferrous
Metal Works, in Bulgaria. Average surface soil Pb concentration decreased with distance
from the plant (200.3 and 24.6 mg Pb/kg, respectively), as did average DTPA-extractable
Pb (69.7 and 4.9 mg Pb/kg, respectively). Pb content in stems and leaves in rapeseed
grown at 0.5 km from the plant averaged 1.73 and 8.69 mg Pb/kg ; average stem and leaf
Pb concentrations in rapeseed grown at the more distant location were reported as 0.72
and 1.42 mg Pb/kg, respectively (Angelova et al.. 2010).
Pb plant BAFs for plants grown in 70 actively cropped fields in California averaged
0.052 for vegetable crops and 0.084 for grains; the highest reported Pb BAF (0.577) was
found in onions. Authors compared the BAFs based on total Pb and Pb in solution and
determined that both were accurate predictors of plant uptake (Chen et al.. 2009b).
Likewise, Zhang et al. (20lib) compiled Pb uptake data for several crop species in
China, and reported an average BAF for grains (rice) of 0.009 (0.0009-0.03) and
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0.41(0.0007-0.17) for leafy vegetables, such as spinach, Chinese cabbage and celery
(Zhang et al.. 201 Ib). Chrastny et al. (2010) characterized the Pb contamination of an
agricultural soil in the vicinity of a shooting range. Pb was predominantly in the form of
PbO and PbCO3, and Pb was taken up by plants through both atmospheric deposition
onto the plant and by root uptake.
The Pb content of ripe date palm (Phoenix dactyliferd) fruit collected in Riyadh, Saudi
Arabia was determined to be indicative of areas of heavy industrialization and
urbanization; Pb concentrations in fruit flesh ranged from 0.34 to 8.87 mg Pb/kg dry
weight, with the highest Pb date concentrations detected near freeways and industrial
areas (Aldjain et al.. 2011). Likewise, Pb concentrations in rosemary (Rosmarinus
officinalis) flowers, stems, and leaves were significantly higher in the urban areas of Al-
Mafraq and Irbid, Jordan than in the smaller town of Ma'an, Jordan (53.6 to 86.5 mg
Pb/kg versus 16.2 to 16.7 mg Pb/kg). Authors noted a significant difference between Pb
concentrations in washed and unwashed rosemary samples, indicating that aerial
deposition and surface dust is likely a significant source of plant-associated Pb (El-Rjoob
et al.. 2008).
Bilberry (Vaccinium myrtillus), accumulated the highest amount of Pb out of four total
herbaceous species growing in Slovakian spruce ecosystems with variable soil Pb
concentrations, giving BAFs of 0.09 to 0.44, depending on location (Kuklova et al..
2010). Because of their long life spans, trees can provide essential information regarding
the sources of bioavailable Pb. A Scots pine forest in northern Sweden was found to
incorporate atmospherically derived Pb pollution directly from ambient air, accumulating
this Pb in bark, needles, and shoots (Klaminder et al.. 2005). Nearly 50% of total tree
uptake was estimated to be from direct adsorption from the atmosphere, as determined
using isotopic ratios and a binary mixing model. Further, Aznar et al. (2009a) found that
the Pb content of black spruce (Picea mariana) needles collected along a metal
contamination gradient emanating from a Canadian smelter in Murdochville, Quebec,
showed a significant decrease in Pb concentration with increasing distance from the
smelter. Interestingly, older needles were determined to accumulate larger quantities of
Pb than younger ones. Foliar damage and growth reduction were also observed in the
trees (Aznar et al.. 2009a). They were significantly correlated with Pb concentration in
the litter layer. In addition, there was no correlation between diminished tree growth and
Pb concentration in the deeper mineral soil layers, strongly suggesting that only current
atmospheric Pb was affecting trees (Aznar et al.. 2009b). Similarly, Kuang et al. (2007)
noted that the Pb concentration in the inner bark ofPinus massoniana trees growing
adjacent to a Pb-Zn smelter in the Guangdong province of China was much higher
(1.87 mg Pb/kg dry weight) than in reference-area trees. Because concentration in the
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inner bark was strongly correlated with concentration in the outer bark, they concluded
that the origin of the Pb was atmospheric.
Dendrochronology (tree ring analysis) has become an increasingly important tool for
measuring the response of trees to Pb exposure (Watmough, 1999). Tree ring studies
reviewed in the 1977 Pb AQCD showed that trees could be used as indicators of
increasing environmental Pb concentrations with time. Additional studies in the 1986 Pb
AQCD indicated that Pb could be translocated from roots to the upper portions of the
plant and that the amounts translocated are in proportion to concentrations of Pb in soil
(U.S. EPA. 1986b). The advent of laser ablation inductively coupled plasma mass
spectrometry has made measurement of Pb concentration in individual tree rings possible
(Witte et al., 2004; Watmough, 1999). This allows for close analysis of the timing of Pb
uptake relative to smelter activity and/or changes in soil chemistry. For example, Aznar
et al. (2008a) measured Pb concentration in black spruce tree rings to determine the
extent and timing of atmospheric deposition near the Murdochville smelter. Variability in
tree-ring Pb content seemed to indicate that trees accumulated and sequestered
atmospheric Pb in close correlation with the rates of smelter emission, but that
sequestration lagged about 15 years behind exposure. However, the ability to determine
time of uptake from the location in growth rings is weakened in species that transfer Pb
readily from outer bark to inner bark. Cutter and Guyette (1993) identified species with
minimal radial translocation from among a large number of tree species, and
recommended the following temperate zone North American species as suitable for metal
dendrochronology studies: white oak (Quercus alba), post oak (Q. stellata), eastern red
cedar (Juniperus virginiana), old-growth Douglas fir (Pseudotsuga menziesii), and big
sagebrush (Artemisia tridentatd). In addition, species such as bristlecone pine (Pinus
aristata), old-growth redwood (Sequoia sempervirens), and giant sequoia (S. giganted)
were deemed suitable for local purposes. Patrick and Farmer (2006) determined that
European sycamore (Acer pseudoplatanus) are not suitable for this type of
dendrochronological analysis because of the formation of multiple annual rings.
Pb in sapwood and heartwood is more likely a result of soil uptake than of direct
atmospheric exposure (Guyette et al.. 1991). Differentiation of geogenic soil Pb in tree
tissue from Pb that originated in the atmosphere requires measurement of stable Pb
isotope ratios (Patrick, 2006). Tree bark samples collected from several areas of the
Czech Republic were subjected to stable Pb isotope analysis to determine the source and
uptake of atmospheric Pb (Conkova and Kubiznakova. 2008). Results indicated that
beech bark is a more efficient accumulator of atmospheric Pb than spruce bark. A
decrease in the 206Pb/207Pb ratio was measured in bark and attributed to increased usage of
leaded gasoline between 1955 and 1990; an increased 206Pb/207Pb ratio was ascribed to
coal combustion (Conkova and Kubiznakova. 2008). Similarly, Savard et al. (2006)
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compared isotope ratios of 206Pb/207Pb and 208Pb/206Pb in tree rings from spruce trees
sampled at a control site near Hudson Bay, with those sampled near the Home smelter
active since 1928, in Rouyn-Noranda, Canada. The concentration of total Pb showed a
major increase in 1944 and a corresponding decrease of the 206Pb/207Pb ratios, suggesting
that the smelter was responsible for the increased Pb uptake (Savard et al.. 2006). The
authors suggested that the apparent delay of 14 years may have been attributable to the
residence time of metals in airborne particles the buffering effect of the soils and, to a
lesser extent, mobility of heavy metals in tree stems. Furthermore, through the use of the
two different isotope ratios, Savard et al. (2006) were able to differentiate three types of
Pb in tree rings: natural (derived from the mineral soil horizons), industrial (from coal
burning urban pollution), and mining (typical of the volcanogenic massive sulfide ore
deposits treated at the Home smelter).
Devall et al. (2006) measured Pb uptake by bald-cypress trees (Taxodium distichum)
growing in a swamp near a petroleum refinery and along a bank containing
Pb-contaminated dredge spoils. They measured Pb in tree cores and showed
greater uptake of Pb by trees in the swamp than by trees growing on the dredge spoil
bank, attributing the difference to exposure source (refinery versus dredge spoils) and
differences in soil chemistry between the swamp and the dredge spoil bank (Devall et al..
2006). Similarly, Gebologlu et al. (2005) found no correlation between proximity to
roadway and accumulated Pb in tomato and bean plants at sites adjacent to two state
roads in Turkey (average Pb concentration 5.4 and 6.0 mg Pb/kg), indicating that uptake
may be influenced by multiple factors, including wind direction, geography, and soil
chemistry. Average Pb levels in leaves were 0.6 and 0.5 mg Pb/kg for tomato and bean
plants, respectively, while fruit concentration averaged 0.4 mg Pb/kg for both species.
Conversely, if foliar contamination is due primarily to dust deposition, distance from a
source such as a road may be easily correlated with Pb concentration on the plants. For
example, Ai-Khlaifat and Al-Khashman (2007) collected unwashed date palm (Phoenix
dactyliferd) leaves at 3-meter trunk height from trees in Jordan to assess the extent of Pb
contamination from the city of Aqaba. Whereas relatively low levels of Pb were detected
in leaves collected at background sites (41 mg Pb/kg), leaves collected adjacent to
highway sites exhibited the highest levels of Pb (177 mg Pb/kg). The authors determined
that Pb levels in date palm leaves correlated with industrial and human activities
(e.g., traffic density) (Ai-Khlaifat and Al-Khashman. 2007). Likewise, Pb concentrations
were significantly enriched in tree bark samples and road dust collected in highly
urbanized areas of Buenos Aires, Argentina (approximate average enrichment factors of
30 and 15 versus reference samples) (Fujiwara et al.. 2011). However, decreases in tissue
Pb concentration with increasing distance from stationary sources can also follow from
decreasing Pb in soil. Bindler et al. (2008) used Pb isotopes to assess the relative
importance of pollutant Pb versus natural Pb for plant uptake and cycling in Swedish
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forested soils. The Pb isotopic composition of needles/leaves and stemwood of different
tree species and ground-cover plants indicated that the majority of Pb present in these
plant components was derived from the atmosphere, either through aerial interception or
actual uptake through the roots. For the ground-cover plants and the needles/leaves, the
206pb/207pb isotopic ratios (i 12 to 1.20) showed that the majority of Pb was of
anthropogenic origin. Stemwood and roots have higher 206Pb/207Pb ratio values (1.12 to
1.30) which showed the incorporation of some natural Pb as well as anthropogenic Pb.
For pine trees, the isotopic ratio decreased between the roots and the apical stemwood
suggesting that much of the uptake of Pb by trees is via aerial exposure. Overall, it was
estimated that 60-80% of the Pb in boreal forest vegetation originated from pollution; the
Pb concentrations were, however, quite low - not higher than 1 mg Pb/kg plant material,
and usually in the range of 0.01-0.1 mg Pb/kg plant material (while soils had a range of 5
to 10 mg Pb/kg in the mineral horizons and 50 to 150 mg Pb/kg in the O horizons).
Overall, the forest vegetation recycles very little of the Pb present in soils (and thus does
not play a direct role in the Pb biogeochemical cycle in boreal forest soils).
Fungal species, as represented by mushrooms, accumulate Pb from soils to varying
degrees. Based on the uptake of naturally occurring 210Pb, Guillen et al. (2009)
established that soil-associated Pb was bioavailable for uptake by mushrooms, and that
the highest 210Pb accumulation was observed in Fomes fomentarius mushrooms, followed
by Lycoperdon perlatum, Boletus aereus, andMacrolepiotaprocera, indicating some
species differences. Benbrahim et al. (2006) also showed species differences in uptake of
Pb by wild edible mushrooms, although they found no significant correlations between
Pb content of mushrooms and soil Pb concentration. Pb concentrations in mushroom
carpophores ranged from 0.4 to 2.7 mg Pb/kg from sites with soil concentrations ranging
from 3.6 and 7.6 mg Pb/kg dry soil. Likewise, Semreen and Aboul-Enein (2011).
reported the heavy metal uptake of wild edible mushrooms collected in various
mountainous regions of Jordan. Pb BCFs ranged between 0.05 (Russula delicd) and 0.33
(Bovistaplumbea) for six mushroom species. Pb BAFs for edible mushrooms collected
from quartzite acidic soils in central Spain (containing 19.2 mg Pb/kg) ranged from 0.07
(Macrolepiotaprocera) to 0.45 (Lepista nuda) (Campos and Tejera. 2011).
Translocation and Sequestration of Pb in Plants
In the 1977 Pb AQCD it was recognized that most Pb taken up from soil remains in the
roots and that distribution to other portions of the plant is variable among species (U.S.
EPA. 1977V The 2006 Pb AQCD (U.S. EPA. 2006b) stated that most of the Pb absorbed
from soil remains bound in plant root tissues either because (1) Pb may be deposited
within root cell wall material, or (2) Pb may be sequestered within root cell organelles.
More recent research largely confirms that Pb taken up from soil largely remains in roots,
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but suggests that some species translocate meaningful amounts into shoot tissue.
Sequestration of Pb may be a protective mechanism for the plant. Recent findings have
been consistent with this hypothesis: Han et al. (2008) observed Pb deposits in the cell
walls and cytoplasm of malformed cells of Iris lactea exposed to 0 to 10 mM Pb (0 to
2,072 mg Pb/L) solution in sand culture for 28 days. They hypothesized that preferential
sequestration of Pb in a few cells, which results in damage to those cells, helps in
maintaining normal overall plant activities through the sacrifice of a small number of
active cells. Similarly, macroscopic analysis of the roots of broad bean (Viciafabd)
cultivated in mine tailings (average Pb concentration of 1,112 mg Pb/kg) by Probst et al.
(2009) revealed dark ultrastructural abnormalities that were demonstrated to be metal-
rich particles located in or on root cell walls. It is unclear whether the presence of these
structures had any effect on overall plant health.
Clark et al. (2006) investigated Pb bioavailability in garden soils in Roxbury and
Dorchester, MA. The sources of Pb were considered to be Pb from paints and from
leaded gasoline additives, with 40 to 80% coming from paint. The average Pb
concentration in foliar tissue of bean plants was 14 ± 5 mg Pb/kg while the concentration
in the bean pod was only 20.6 mg Pb/kg. For mustard plants, there was a linear
relationship (R2=0.85) between Pb concentration in plant tissues and Pb concentration in
the soil (both for plants grown in situ and those grown under greenhouse conditions).
Murray et al. (2009) investigated the uptake and accumulation of Pb in several vegetable
species (carrot [Daucus carota], radish [Raphanus sativus], lettuce [Lactuca sativa],
soybean [Gfycine max], and wheat [Triticum aestivum]) from metal-contaminated soils,
containing 10 to 40 mg Pb/kg and demonstrated that most Pb remained in the roots. No
Pb was measured in the above-ground edible soybean and wheat tissues, while carrots,
the most efficient accumulator of Pb, contained a maximum Pb tissue concentration of
12 mg Pb/kg dry mass. Similarly, (Cho et al.. 2009) showed that green onion (Allium
fistulosuni) plants also take up little Pb when planted in soil spiked with Pb nitrate. No
plant tissues contained a Pb concentration greater than 24 mg Pb/kg when grown for
14 weeks in soils of up to 3,560 mg Pb/kg, and the majority of bioavailable Pb was
determined to be contained within the roots. Chinese spinach (Amaranthus dubius) also
translocates very little Pb to stem and leaf tissue, and uptake from Pb-containing soils (28
to 52 mg Pb/kg) is minimal (Mellem et al.. 2009). Wang et al. (20 lie) determined tissue-
specific BCFs for wheat grown in soils containing 93 to 1,548 mg Pb/kg. Although the
average calculated root BCF was 0.3, very little Pb was translocated to shoots (average
BCF=0.02), shells (0.006), and kernels (0.0007) (Wang et al.. 20lie). However, it is not
known how much of the Pb applied was present in pore water. Sonmez et al. (2008)
reported that Pb accumulated by three weed species (Avena sterilis, Isatis tinctoria,
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Xanthium strumarium) grown in Pb-spiked soils was largely concentrated in the root
tissues, and little was translocated to the shoots (Sonmez et al.. 2008).
The Pb BCFs for alfalfa (Medicago sativa) and crimson clover (Trifolium incarnatuni)
grown in mixtures of heavy metals (Pb concentrations of 10 to 500 (ig Pb/kg) were
reportedly low. For alfalfa, BCFs ranged from 0.02 to 0.12, while for crimson clover,
these values were between 0.04 and 0.06 (Comino et al.. 2011). The low shoot-root
translocation factors reported for alfalfa (0.17 to 0.43) indicated that plant Pb content was
largely contained in root tissue. Businelli et al. (2011) calculated whole-plant Pb BAFs
for lettuce, radish, tomato and Italian ryegrass using Pb-spiked soils (average values of
0.025, 0.021, 0.032, and 0.65, respectively). Again, the majority of accumulated Pb was
stored in root tissue, with comparatively little translocated to above-ground tissues
(Businelli etal.. 2011).
Recent research has shown that Pb translocation to stem and leaf tissues does occur at
significant rates in some species, including the legume Sesbania drummondii (Peralta-
Videa et al.. 2009) and buckwheat (Fagopyrum esculentum) (Tamura et al.. 2005). Wang
et al. (2006b) noted that Pb soil-to-plant transfer factors were higher for leafy vegetables
(Chinese cabbage, pak-choi, and water spinach) than for the non-leafy vegetables tested
(towel gourd, eggplant, and cowpea). Tamura et al. (2005) demonstrated that buckwheat
is an efficient translocator of Pb. Buckwheat grown in Pb-containing soils collected from
a shooting range site (average 1M HC1 extractable Pb= 6,643 mg Pb/kg) preferentially
accumulated Pb in leaves (8,000 mg Pb/kg) and shoots (4,200 mg Pb/kg), over root
tissues (3,300 mg Pb/kg). Although plant growth was unaffected, this level of leaf and
shoot accumulation is likely to have significant implications for exposure of herbivores.
Similarly, Shaheen and Tsadilas (2009) reported that vegetables (pepper, okra, and
eggplant) grown in soils containing 24 to 30 mg Pb/kg total Pb were more likely to
accumulate Pb in leaves (range: undetected to 25 mg Pb/kg) rather than in fruits (range:
undetected to 19 mg Pb/kg); however, no significant correlation between soil Pb
concentration and plant tissue Pb concentration could be established (Shaheen and
Tsadilas. 2009). Tobacco plants were also observed to take up significant amounts of Pb
into leaf tissue. Field-grown plants in soils containing an average of 19.8 mg Pb/kg
contained average lower, middle and upper leaf Pb concentrations of 11.9, 13.3, and
11.6 mg Pb/kg respectively (Zaprjanova et al.. 2010). Uptake by tobacco plants was
correlated with both total soil Pb concentrations and the mobile Pb fraction (average
3.8 mg Pb/kg soil).
There is broad variability in uptake and translocation among plant species, and
interspecies variability has been shown to interact with other factors such as soil type. By
studying multiple species in four Pb-Zn mining sites in Yunnan, China, Li et al. (2009d)
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demonstrated not only significant differences in uptake and translocation among the
species studied, but also modification of the effect on species by type of soil. Plants
sampled represented nine species from four families—Caryophyllaceae, Compositae,
Cruciferae, and Pteridaceae. Overall, soil Pb concentration averaged 3,772 mg Pb/kg dry
weight, with the highest site average measured at the Minbingying site (5,330 mg Pb/kg),
followed by Paomaping (2,409 mg Pb/kg), Jinding (1,786 mg Pb/kg), and Qilinkeng
(978 mg Pb/kg). The highest average shoot Pb concentration (3,142 mg Pb/kg) was
detected in Stellaria vestita (Caryophyllaceae) collected at Paomaping, while Sinopteris
grevilloides (Pteridaceae) collected from Minbingying exhibited the lowest shoot Pb
concentration (69 mg Pb/kg). A similar trend was detected in root tissues. S. vestita root
collected from the Paomaping area contained the maximum Pb concentration measured
(7,457 mg Pb/kg), while the minimum root Pb levels were measured in Picris
hieracioides (Pteridaceae) tissues collected from Jinding. These results indicate
significant interspecies differences in Pb uptake, as well as potential soil-specific
differences in Pb bioavailability. S. vestita, in particular, was observed to be an efficient
accumulator of Pb, with a maximum enrichment coefficient of 1.3. Significant
correlations between soil Pb concentration and average shoot and root Pb levels were also
established (Li et al., 2009d). Within plant species, the variability in uptake and
translocation of Pb may extend to the varietal level. Antonious and Kochhar (2009)
determined uptake of soil-associated Pb for 23 unique genotypes from four species of
pepper plants (Capsicum chinense, C.frutescens, C. baccatum, and C. annum). Soil Pb
concentration averaged approximately 0.6 mg Pb/kg dry soil. No Pb was detected in the
fruits of any of the 23 genotypes, except two out of seven genotypes of C. baccatum,
which had 0.9 and 0.8 mg Pb/kg dry weight Pb in fruit.
Recent studies substantiated findings from the 2006 Pb AQCD that plants store a large
portion of Pb in root tissue. Pb soil-to-plant transfer factors are higher for leafy
vegetables than for the non-leafy vegetables (Wang et al., 2006b) and buckwheat has
recently been shown to be an efficient translocator of Pb from soil to above-ground
shoots (Tamura et al., 2005).
Field studies carried out in the vicinity of Pb smelters (Hu et al., 2009b) show that Pb
may accumulate in shoot tissue through direct stomatal uptake rather than by soil-root-
shoot translocation. For instance, Hovmand and Johnsen (2009) determined that about
98% of Pb sequestered in Norway spruce needles and twigs was derived from
atmospheric sources, and that less than 2% of Pb was translocated from the roots
(Hovmand et al.. 2009). Dendrochronology has become more advanced in recent years
and is a useful tool for monitoring historical uptake of Pb into trees exposed to
atmospheric or soil Pb. Trees accumulate and sequester atmospheric Pb in close
correlation with the rate of smelter emissions, although one study indicated that
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sequestration can lag behind exposure from emissions by 15 years. Pb in the outer woody
portion of the tree is more likely the result of direct atmospheric exposure, while Pb in
sapwood is more likely a result of soil uptake. This difference provides an important tool
for analyzing source apportionment of Pb accumulation in plants (Guvette et al.. 1991).
6.3.3.2 Terrestrial Invertebrates
At the time of publication of the 2006 Pb AQCD (U.S. EPA. 2006b). little information
was available regarding the uptake of atmospheric Pb pollution (direct or deposited) by
terrestrial invertebrate species. Consequently, few conclusions could be drawn
concerning the Pb uptake rate of particular species although there was some evidence that
dietary or habitat preferences may influence exposure and uptake. Recent literature
indicates that invertebrates can accumulate Pb from consuming a Pb-contaminated diet
and from exposure via soil, and that uptake and bioaccumulation of Pb by invertebrates is
lower than that observed for other metals.
Snails
Pauget et al. (2011) reported that uptake of Pb from soil by the land snail (Cantareus
asperses) was most significantly influenced by soil pH and organic matter, as increases in
these variables were correlated to decreased Pb bioavailability. Cantareus asperses snails
exposed to dietary Pb at 3.3, 86, and 154 mg/kg of diet (spiked with Pb sulfate) for up to
64 days were found to assimilate a significant proportion of Pb, and feeding rates were
unaffected by the presence of the metal (Beebv and Richmond. 2010). While BCFs for
Cd were observed to increase over the 64-day study period, the rate of Pb assimilation
remained consistent over time and the authors inferred the absence of a regulatory
mechanism for uptake of Pb. The authors speculated that uptake is a function of growth
or cell turnover instead. Helix aspersa snails rapidly accumulated Pb from contaminated
soil (1,212 mg Pb/kg) and from eating contaminated lettuce (approximately 90 mg Pb/kg
after 16 weeks' growth on Pb-contaminated soil) during the first 2 weeks of exposure, at
which point snail body burdens reached a plateau (Scheifler et al., 2006b). There were no
observed effects of Pb exposure or accumulation on survival or growth in C. asperses or
H. aspersa. In another study (Ebenso and Ologhobo. 2009b). juvenile Achatina achatina
snails confined in cages on former Pb-battery waste dump sites were found to accumulate
Pb from both plant and soil sources. Soil Pb concentration averaged 20, 200, and
1,200 mg Pb/kg at the three main waste sites, while leaf tissues of radish (Raphanus
sativus) grown at these sites averaged 7, 30, and 68 mg Pb/kg dry weight, respectively.
Concentration of Pb in snail tissues rose with concentration in both soil and plants, and
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the authors found that for both sources, a log-log relationship could be estimated with a
very close fit (r2 =0.94 and 0.95, respectively). Pb concentration in snail tissues averaged
12, 91, and 468 mg Pb/kg, respectively, at the three sites, which the authors stipulated
were above the maximum permissible concentration of Pb for human consumption of
mollusks, mussels, and clams (1.5 mg Pb/kg tissue) as determined by the U.K. Food
Standards Agency. Pb concentration in snail tissues generally is much lower than that of
the soil substrates upon which they were reared, but higher than in other soil-dwelling
organisms. De Vaufleury et al. (2006) exposed Helix aspersa snails to standardized
(International Organization for Standardization methodology [ISO 11267:1999])
artificial-substrate soils amended with sewer sludge containing 13, 26, 39, or 52 mg
Pb/kg for 28 days without supplemental food. After the exposure period, snail foot tissue
contained increased levels of Pb—1.9, 1.7, and 1.5 mg Pb/kg dry weight versus
concentration averaging 0.4 mg Pb/kg in control organisms. Viscera also exhibited
increased Pb levels at the two highest exposures, with measured tissue concentration of
1.2 and 1.1 mg Pb/kg, respectively, as compared with control tissue Pb levels of 0.4 mg
Pb/kg. However, there was no significant increase in snail-tissue Pb concentration when
natural soil was used in place of ISO medium, and there was no relationship between soil
Pb concentration and snail tissue concentration, strongly suggesting the presence of soil
variables that modify bioavailability. Notten et al. (2008) investigated the origin of Pb
pollution in soil, plants, and snails by means of Pb isotope ratios. They found that a
substantial proportion of Pb in both plants and snails was from current atmospheric
exposure.
Finally, a study by Coeurdassier et al. (2007) found that the presence of snails was
associated with higher Pb content in earthworms, suggesting that snails themselves may
have an effect on bioavailability.
Earthworms
Accumulation studies conducted with Eisenia sp. earthworms documented the difficulty
of extrapolating accumulation kinetic constants from one soil type to another, and
showed that many soil physiochemical properties, including pH, organic matter, and
CEC, among others, affect metal bioavailability (Nahmani et al.. 2009). Source of Pb,
and proportion of soil:leaf litter also affect Pb bioavailability. Bradham et al. (2006)
examined the effect of soil chemical and physical properties on Pb bioavailability.
Eisenia andrei earthworms were exposed to 21 soils with varying chemical and physical
properties that were freshly spiked with Pb to give a standard concentration of 2,000 mg
Pb/kg dry weight. At equivalent Pb exposure, the main determinants of both internal
earthworm Pb concentration and mortality were pH first (with lower pH resulting in
higher concentration and mortality), then CEC. However, the apparent importance of
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CEC was due to its correlation with several other less important soil characteristics.
These data corroborate that Pb bioavailability and toxicity are increased in acidic soils
and in soils with a low CEC (Section 6.3.2). This finding was confirmed by Gandois et al.
(2010). who determined that the free-metal-ion fraction of total Pb concentration in field-
collected soils was largely predicted by pH and soil Fe content.
The role of soil profile and preferred depth was studied using eight species of earthworms
from 27 locations in Switzerland, representing three ecophysiological groups (Ernst et al..
2008): epigeic (surface-dwelling worms), endogeic (laterally burrowing worms that
inhabit the upper soil layers), and anecic (vertically burrowing worms that can reach
depths of 2 meters). For epigeic and anecic earthworms, the total concentration of Pb in
leaf litter and in soil, respectively, were the most important drivers of Pb body burdens.
By contrast, the level of Pb in endogeic earthworms was largely determined by soil pH
and CEC. As a result of these differences, the authors suggested that atmosphere-sourced
Pb may be more bioavailable to epigeic than endogeic species, because it is less
dependent on modifying factors. Suthar et al. (2008). on the other hand, found higher Pb
bioaccumulation in the endogeic earthworm Metaphire posthuma than in the anecic
earthworm species Lampito mauritii, and speculated that differences in Pb tissue level
arose from differing life-history strategies, such as feeding behaviors, niche preferences,
and burrowing patterns, all of which exposed the endogeic species to greater Pb
concentration. Garg et al. (2009) reported that the smaller native earthworm
Allolobophoraparva accumulated significantly greater Pb concentrations than E. fetida.
Subsequently, it was concluded that native earthworm species may exhibit a higher Pb
accumulation potential as a result of increased tolerance to the heavy metal (Garg et al..
2009).
Earthworm activity can alter Pb bioavailability and subsequent uptake by earthworms
themselves and other organisms. Sizmur and Hodson (2009) speculated that earthworms
affect Pb mobility by modifying the availability of cations or anions. The concentration
of water-soluble Pb was observed to increase following earthworm (Lumbricus terrestris)
feeding activity in field-collected soils containing 132.7, 814.9, and 821.4 mg total Pb/kg
(calculated BAFs of 0.27, 0.33, and 0.13, respectively) (Alonso-Azcarate et al.. 2011).
However, Coeurdassier et al. (2007) found that snails did not have a higher Pb content
when earthworms were present, and that unexpectedly, Pb was higher in earthworm
tissue when snails were present.
Despite significant Pb uptake by earthworms, Pb in earthworm tissue may not be
bioavailable to predators. Pb in the earthworm (Aporrectodea caliginosa) was determined
to be contained largely in the granular fraction (approximately 60% of total Pb), while the
remaining Pb body burden was in the tissue, cell membrane, and intact cell fractions
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(Vijver et al.. 2006). However, this may vary by species, as (Li et al.. 2008b) found that
more than half of the Pb accumulated by E. fetida was contained within earthworm tissue
and cell membranes. Regardless, Vijver et al. (2006) concluded that only a minority of
earthworm-absorbed Pb would be lexicologically available to cause effects in the
earthworms or in their predators.
Arthropods
Pb and other metals were analyzed in honeybees (Apis mellifera) foraging in sampling
sites that included both urban areas and wildlife reserves in central Italy. (Perugini et al..
2011). Pb in whole bees ranged from 0.28 to 0.52 mg Pb/kg with the highest
concentration in honeybees caught in hives near an airport and industrial area within the
Rome metropolitan area. Cicadas pupating in historically Pb-arsenate-treated soils
accumulated Pb at concentrations similar to those reported previously for earthworms
(Robinson et al.. 2007). Likewise, tissue Pb levels measured in Coleoptera specimens
collected from areas containing average soil concentration of 45 and 71 mg Pb/kg
exhibited a positive relationship with soil Pb content, although abundance was unaffected
(Schipper et al.. 2008). By contrast, two grasshopper species inhabiting Pb and Cd-
contaminated areas near Zn smelting facilities exhibited different Pb accumulation rates.
Locust (Locusta migratoria) collected from areas with an average Pb soil concentration
of 540mg Pb/kg contained 47 mg Pb/kg, while grasshoppers (Acrida chinensis)
inhabiting the same area accumulated 93.9 mg Pb/kg (Zhang etal.. 2012). This gives
respective BAFs of 0.09 and 0.17. Similarly, the Pb sequestration rates that were
observed in two woodlouse species, O. asellus and P. scaber, were species-dependent
(Gal et al.. 2008). Both species were field collected at Pb-contaminated sites (average
concentration, 245 mg Pb/kg dry weight; range, 21-638 mg Pb/kg dry weight), with
O. asellus Pb levels averaging 43 mg Pb/kg over all sites, while P. scaber contained no
detectable Pb residues. Pb concentration measured in granivorous rough harvester ants
(Pogonomyrmex rugosus), in the seeds of some plant species they consume, and in
surface soil, were all shown to decline with increasing distance from a former Pb smelter
near El Paso, Texas, where soil leachable Pb at the three sites of ant collection ranged
from 0.003 to 0.117 mg Pb/kg (Del Toro et al., 2010). Ants accumulated approximately
twice as much Pb as was measured in seeds, but the study did not separate the effects of
dietary exposure from those of direct contact with soil or respiratory intake.
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6.3.3.3 Terrestrial Vertebrates
At the time of the 1977 Pb AQCD few studies of Pb exposure and effects in wild animals
other than birds had been conducted. A limited number of rodent trapping studies near
roadsides indicated general trends of species differences in Pb uptake and higher
concentrations of Pb in habitats adjacent to high-traffic areas (U.S. EPA. 1977). In the
1986 Pb AQCD concentration of Pb in bone tissue was reported for selected herbivore,
omnivore, and carnivore species [Table 8-2 in (U.S. EPA, 1986b)1.
Tissue Pb residues in birds and mammals associated with adverse toxicological effects
were presented in the 2006 Pb AQCD. In general, avian blood, liver, and kidney Pb
concentrations of 0.2-3 (ig Pb/dL, 2-6 mg Pb/kg wet weight, and 2-20 mg Pb/kg wet
weight, respectively, were linked to adverse effects. A few additional studies of Pb
uptake and tissue residues in birds and mammals conducted since 2006 are reviewed
here.
In a study of blood Pb levels in wild Steller's eiders (Polysticta stelleri) and black scoters
(Melanitta nigrd) in Alaska, the authors compiled avian blood Pb data from available
literature to develop reference values for sea ducks (Brown et al., 2006). The background
exposure reference value of blood Pb was <20 (ig Pb/dL, with levels between 20 and
59 (ig Pb/dL as indicative of Pb exposure. Clinical toxicity was in the range of
60-99 (ig Pb/dL in birds while >100 (ig Pb/dL results in acute, severe toxicity. In
measurement of blood Pb with a portable blood Pb analyzer, only 3% of birds had values
indicating exposure and none of the birds had higher blood Pb levels or clinical signs of
toxicity. Tissue distribution of Pb in liver, kidney, ovary and testes of rain quail (Coturnix
coramandelica) following oral dosing of 0.5 mg Pb/kg, 1.25 mg Pb/kg or 2.5 mg Pb/kg
Pb acetate for 21 days indicated that Pb uptake was highest in liver and kidney and low in
ovary and testes (Mehrotra et al.. 2008). Resident feral pigeons (Columba livid) captured
in the urban and industrial areas of Korea exhibited increased lung Pb concentration,
ranging from 1.6 to 1.9 mg Pb/kg wet weight (Nam and Lee. 2006). However, tissue
concentration did not correlate with atmospheric Pb concentration, so the authors
concluded that ingestion of particulate Pb (paint chips, cement, etc.) in the urban and
industrial areas was responsible for the pigeons' body burden. Similarly, 70% of
American woodcock (Scolopax minor) chicks and 43% of American woodcock young-of-
year collected in Wisconsin, U.S., exhibited high bone Pb levels of 9.6-93 mg Pb/kg dry
weight and 1.5-220 mg Pb/kg, respectively, even though radiographs of birds'
gastrointestinal tracts revealed no evidence of shot ingestion (Strom et al., 2005). Authors
hypothesized that unidentified anthropogenic sources may have caused the observed
elevated Pb levels.
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In addition to birds, soil-dwelling mammals can also bioaccumulate atmospherically-
sourced Pb. Northern pocket gophers (Thomomys talpoides) trapped within the Anaconda
Smelter Superfund Site were shown to accumulate atmospherically deposited Pb. Gopher
liver and carcass Pb concentration averaged 0.3 and 0.4 mg Pb/kg wet weight on low Pb
soils (47 mg Pb/kg), 0.4 and 0.9 mg Pb/kg wet weight in medium Pb soils (95 mg Pb/kg)
and 1.6 and 3.8 mg Pb/kg wet weight in high Pb soils (776.5 mg Pb/kg) (Reynolds et al..
2006). Likewise, rats trapped in the vicinity of a Kabwe, Zambia Pb-Zn mine had
significantly elevated liver and kidney Pb concentrations. Soil Pb concentrations were
measured between 9 and 51,188 mg Pb/kg (approximate average of 200 mg Pb/kg dry
weight), while rat liver and kidney Pb concentrations ranged between 0.009 and 7.3 mg
Pb/kg dry weight and 0.3 and 22.1 mg Pb/kg dry weight, respectively. Consequently,
residence in the mining region was correlated to significantly increased Pb body burdens
for rats (Nakayama et al.. 2011). Angelova et al. (2010) reared rabbits on a fodder
mixture containing Pb-contaminated rapeseed (grown adjacent to a metal works plant).
Following a four-week exposure, Pb was most heavily concentrated in rabbit kidney
tissue (3.9 mg Pb/kg and 1.9 mg Pb/kg, for high and low diet respectively), bone (1.0 and
0.3 mg Pb/kg, respectively), and liver (0.6 and 0.4 mg Pb/kg, respectively). Yucatan
micropigs (Sus scrofa) and Sprague-Dawley rats (Rattus norvegicus) reared on
Pb-contaminated soil (5% of 1,000 mg Pb/kg soil as dietary component) consumed
significantly different amounts of Pb. Over a 30-day period, rats consumed an average of
19.4 mg Pb, while micropig intake averaged 948 mg Pb (Smith et al.. 2009a). This
resulted in significantly higher Pb accumulation in micropigs, based on liver, blood,
kidney and bone Pb concentrations (average concentrations of 1.2, 25, 0.9, and 9 mg
Pb/kg for micropigs, and 0.2, 7, 0.5, and 1.5 mg Pb/kg for rats, respectively).
Casteel et al. (2006) found that bioavailability of Pb from environmental soil samples in
swine (Sus domestica) depended on Pb form or type, with high absorption of Pb from
cerussite and Mn-Pb oxides and poor absorption of Pb from galena and anglesite.
Juvenile swine (approximately 5-6 weeks old and weighing 8-11 kg) were fed
Pb-contaminated soils collected from multiple sources for 15 days (concentration range
of 1,270 to 14,200 mg Pb/kg) to determine the relative bioavailability. While Pb
concentrations were roughly equivalent in blood, liver, kidney, and bone tissues,
individual swine exhibited different uptake abilities (Casteel et al.. 2006).
Consistent with observations in humans, dietary Ca2+ deficiency (0.45 mg Ca2+ daily
versus 4 mg under normal conditions) was linked to increased accumulation of Pb in
zebra finches (Taeniopygia guttata) that were provided with drinking water containing
20 mg Pb/L (Dauwe et al., 2006). Liver and bone Pb concentration were increased by an
approximate factor of three, while Pb concentration in kidney, muscle, and brain tissues
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were roughly doubled by a Ca2+-deficient diet. However, it is not known whether this
level of dietary Ca2+ deficiency is common in wild populations of birds.
6.3.3.4 Food Web
In addition to the organism-level factors reviewed above, understanding the
bioavailability of Pb along a simple food chain is essential for determining risk to
terrestrial animals. While the bioavailability of ingested soil or particles is relatively
simple to measure and model, the bioavailability to secondary consumers of Pb ingested
and sequestered by primary producers and primary consumers is more complex. Kaufman
et al. (2007) caution that the use of total Pb concentration in risk assessments can result in
overestimation of risk to ecological receptors, and they suggest that the bioaccessible
fraction may provide a more realistic approximation of receptor exposure and effects.
This section reviews recent literature that estimates the bioaccessible fraction of Pb in
dietary items of higher order consumers, and various studies suggesting that Pb may be
transferred through the food chain but that trophic transfer of Pb results in gradual
attenuation, i.e., lower concentration at each successive trophic level.
Earthworm and plant vegetative tissue collected from a rifle and pistol range that
contained average soil Pb concentration of 5,044 mg Pb/kg were analyzed for Pb content
and used to model secondary bioavailability to mammals (Kaufman et al.. 2007).
Earthworms were determined to contain an average of 727 mg Pb/kg, and the Pb content
of unwashed leaf tissues averaged 2,945 mg Pb/kg. Canonical correspondence analysis
detected no relationship between earthworm and soil Pb concentration, but did show
correlation between unwashed vegetation and soil concentration. The authors noted that
the relatively high Pb concentration of unwashed as opposed to washed vegetation
indicated the potential importance of aerial deposition (or dust resuspension) in
determining total vegetative Pb concentration. Based on the mammalian gastric model,
they noted that 50% of vegetation tissue Pb and 77% of earthworm tissue Pb was
expected to be bioavailable to consumers. The avian gizzard model indicated that 53% of
soil Pb and 73% of earthworm Pb was bioaccessible to birds; and, for both mammals and
birds, the bioaccessible fraction of Pb was a function of total Pb concentration.
The transfer of Pb from soils contaminated by a Pb-Zn mine was limited along a soil-
plant-insect-chicken food chain (Zhuang et al.. 2009). In soils averaging 991 mg Pb/kg,
plants of the fodder plant Rumex patientia X tianschanicus sequestered an average of
1.6 mg Pb/kg wet weight in the shoot tissue, while larvae of the leafworm Spodoptera
litura accumulated an average Pb concentration of 3.3 mg Pb/kg wet weight S. liturn-fed
chickens (Gallus gallus domesticus) accumulated 0.58 mg Pb/kg and 3.6 mg Pb/kg in
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muscle and liver tissue, respectively, but only liver Pb burden was increased significantly
relative to controls. A large proportion of ingested Pb was excreted with the feces.
Likewise, a wading bird species, the black-tailed godwit (Limosa limosd) was shown to
accumulate Pb from earthworms residing in Pb-contaminated soils (Roodbergen et al..
2008). Pb concentration in eggs and feathers was increased in areas with high soil and
earthworm Pb concentration (336 and 34 mg Pb/kg, respectively): egg Pb concentration
averaged 0.17 mg Pb/kg and feather concentration averaged 2.8 mg Pb/kg. This suggests
that despite a residence breeding time of only a few months, this bird species could
accumulate Pb when breeding areas are contaminated.
Rogival et al. (2007) showed significant positive correlations between soil Pb
concentration along a gradient (approximately 50 to 275 mg Pb/kg) at a metallurgical
plant, and Pb concentration in both acorns (from Quercus robuf) and earthworms
(primarily Dendrodrilus rubidus and Lumbricus rubellus) collected on site. Acorn and
earthworm Pb contents were, in turn, positively correlated with the Pb concentration in
the liver, kidney, and bone tissues of locally trapped wood mice (Apodemus sylvaticus).
The uptake and transfer of Pb from soil to native plants and to red deer (Cervus elaphus)
was investigated in mining areas of the Sierra Madrona Mountains in Spain (Reglero et
al.. 2008). The authors reported a clear pattern between plant Pb concentration and the Pb
content of red deertissues with attenuation (i.e., decreasing concentration) of Pb up the
food chain. Interestingly, soil geochemistry likely was affected by mining activity as
Holm oak (Quercus ilex), gum rockrose (Cistus ladanifer), elm leaf blackberry (Rubus
ulmifolius), and grass (Graminae) tissues collected from mining areas exhibited increased
Pb levels (up to 98 mg Pb/kg in grasses and 21 mg Pb/kg in oak) despite the fact that total
soil Pb concentrations were not significantly greater than those of the non-mining areas.
Positive relationships were observed between Cepaea nemoralis snail tissue Pb levels
and Pb concentration measured in Urtica dioica leaves in field-collected samples from
areas characterized by metal soil contamination (approximately 200 to 400 mg Pb/kg)
(Notten et al.. 2005). Inouye et al. (2007) found that several invertebrate prey offence
lizards, including Acheta domestica crickets, Tenebrio molitor beetles, and P. scaber
isopods, accumulate Pb from dietary exposures (10, 50, 100, 250, 500, 750, and 1,000 mg
Pb/kg) lasting between 44 and 72 days. By day 44, Pb body burdens of crickets were 31,
50 and 68 mg Pb/kg (wet weight) at the three highest dietary exposures, respectively.
Isopods and beetle larvae accumulated significantly less Pb, with average body burdens
of 10, 15, and 14 mg Pb/kg following 56 days of exposure; and 12, 14, and 31 mg Pb/kg
following 77 days of exposure, respectively. For all invertebrates tested, Pb was
sequestered partly in the exoskeleton, and partly in granules. Exoskeleton Pb may be
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available to predators, but returns to background level with each shedding, while granular
Pb is likely unavailable, at least to other invertebrates (Vijver et al.. 2004).
In a comparison of rural and urban blackbirds (Turdus merula), Sheifler et al.(2006a)
found that while Pb concentration in unwashed tail feathers was equivalent in both
populations, urban birds had higher tissue concentrations. Pb content of urban
earthworms was also higher than that of rural earthworms. Hypothesizing that tail feather
Pb reflected deposition from air and resuspended dust, the authors suggested that elevated
Pb in the urban birds was mostly dietary in origin.
Overall, studies of Pb transfer in food webs have established the existence of pervasive
trophic transfer of the metal, but no consistent evidence of trophic magnification. It
appears that on the contrary, attenuation is common as Pb is transferred to higher trophic
levels. However, many individual transfer steps, as from particular plants to particular
invertebrates, result in concentration, which may then be undone when stepping to the
next trophic level. It is possible that whether trophic transfer is magnifying or attenuating
depends on Pb concentration itself. Kaufman et al. (2007) determined that, at low
concentrations of soil Pb, risk to secondary consumers (birds and mammals) was driven
by the bioavailability of Pb in worm tissues, while at high soil concentrations,
bioavailability of soil-associated Pb was more critical. The authors concluded that
incorporation of bioavailability/bioaccessibility measurements in terrestrial risk
assessments could lead to more accurate estimates of critical Pb levels in soil and biota.
Finally, while trophic magnification does greatly increase exposure of organisms at the
higher levels of the food web, these studies establish that atmospherically deposited Pb
reaches species that have little direct exposure to it. For those species, detrimental effects
are not a function of whether they accumulate more Pb than the species they consume,
but of the absolute amount they are exposed to, and their sensitivity to it.
6.3.4 Biological Effects of Pb in Terrestrial Systems
Various effects can be observed in exposed terrestrial species following uptake and
accumulation of Pb. While many of the responses are specific to organism type, induction
of antioxidant activities in response to Pb exposure has been reported in plants,
invertebrates, and vertebrates. Observed biological effects caused by exposure to
atmosphere-derived Pb are discussed in this section, while the results of dose-response
experimentation are reported in Section 6.3.5. Because environmental releases of Pb
often include simultaneous release of other metals, it can be difficult to identify
Pb-specific effects in field studies, with the exception of effects from leaded gasoline and
some Pb smelter deposition. Many laboratory studies that expose organisms to natural
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soils (or to biosolids-amended soils) also include exposure to multiple metals. There is
some information about mechanisms of metal interactions, such as through competition
for binding locations on specific enzymes or on cellular receptors, but generally such
interactions (particularly of multiple metals) are not well understood (ATSDR, 2004).
Despite a few well-known examples of metal antagonism (e.g., Cu and Mo or Cd and
Zn), it is common practice to assume additivity of effects (Fairbrother et al.. 2007).
Because this review is focused on effects of Pb, studies reviewed for this section and the
following include only those for which Pb was the only (or primary) metal to which the
organism was exposed. All reported values are from exposures in which concentrations
of Pb were analytically verified unless nominal concentrations are stated.
6.3.4.1 Terrestrial Plants and Lichen
Pb exposure has been linked to decreased photosynthesis in affected plants, significant
induction of antioxidant activities, genetic abnormalities, and decreased growth.
Induction of antioxidant responses in plants has been shown to increase tolerance to
metal soil contamination, but at sufficiently high levels, antioxidant capacity is exceeded,
and metal exposure causes peroxidation of lipids and DNA damage, eventually leading to
accelerated senescence and potentially death (Stobrawa and Lorenc-Plucinska. 2008).
Effects on Photosystem and Chlorophyll
Photosynthesis and mitosis were recognized as targets of Pb toxicity in plants in the 1977
Pb AQCD and additional effects of Pb on these processes were reported in subsequent Pb
AQCDs (U.S. EPA. 2006c. 1986b. 1977). The effect of Pb exposure on the structure and
function of plant photosystem II was recently studied in giant duckweed, Spirodela
polyrrhiza (Ling and Hong. 2009). Although this is an aquatic plant, photosystem II is
present in all plants. This finding thus provides support for effects on photosystem II
being the cellular-level mechanism that leads to decreases photosynthesis observed in
other plants. The Pb concentration of extracted photosystem II particles was found to
increase with increasing environmental Pb concentration, and increased Pb concentration
was shown to decrease emission peak intensity at 340 nm, amino acid excitation peaks at
230 nm, tyrosine residues, and absorption intensities. This results in decreased efficiency
of visible light absorption by affected plants. The authors theorized that Pb2+ may replace
either Mg2+ or Ca2+ in chlorophyll or the oxygen-evolving center, inhibiting photosystem
II function through an alteration of chlorophyll structure. Consistently with these results,
Wu et al. (2008b) demonstrated that Pb exposure interfered with and decreased light
absorption by spinach (Spinacia oleraced) plants. Spinach seeds were soaked in 5, 12, or
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25 mM Pb chloride (1036, 2486, or 5180 mg Pb/L) for 48 hours prior to germination, and
following 42 days of growth, plants were sprayed with Pb chloride solutions. Chloroplast
absorption peak intensity, fluorescence quantum yield at 680 nm, and whole-chain
electron transport rate all decreased with Pb exposure, as did photosystem II
photoreduction and oxygen evolution. Similarly, the photosynthetic rate of maize (Zea
mays) seedlings decreased over 21 days exposure to Pb, and measured leaf Pb
concentrations in photosynthetically-depressed seedlings ranged from approximately 100
to 300 mg Pb/kg dry weight (Ahmad et al.. 2011). Liu et al. (2010a) observed that
chlorophyll a and b content in wheat grown in soils spiked with Pb nitrate rose with
length of exposure until 14 days, at which point chlorophyll decreased. At nominal
exposures of 0.1 and 0.5 mM Pb (20.72 and 103.6 mg Pb/L) in hydroponic solution for
50 days, concentration of chlorophyll a and b was decreased in radish (R. sativus)
(Kumar and Tripathi. 2008). Changes in chlorophyll content in response to Pb were also
observed in lichen and moss species following exposures intended to simulate
atmospheric deposition (Carreras and Pignata. 2007). Usnae ambfyoclada lichen was
exposed to aqueous Pb solutions of 0.5, 1, 5, and 10 mM Pb nitrate (103.6, 207.2, 1,036,
and 2,072 mg Pb/L); chlorophyll a concentration was shown to decrease with increasing
Pb exposure. However, the ratio of lichen dry weight to fresh weight increased following
Pb exposures. It should be noted that highly productive Sphagnum mosses accumulated
atmospheric Pb at the same rate as slower growing mosses, indicating that moss growth
allowed for further Pb uptake, rather than a "dilution" effect (Kempter et al.. 2010). As
compared to other metals, however, Pb caused less physiological damage, which the
authors attributed to the metal's high affinity for binding to and sequestration within cell
walls (Carreras and Pignata. 2007).
The effect of Pb exposure on chlorophyll content of the moss and liverwort species
Thuidium delicatulum, T. sparsifolium, and Ptychanthus striatus was investigated
following immersion in six solutions of Pb nitrate containing from 10"10 to 10"2 M Pb
(0.00002 to 2,072 mg Pb/L) (Shakva et al.. 2008). Both chlorophyll a and total
chlorophyll content of the mosses (T. delicatulum and T. sparsifolium) decreased with
increasing Pb exposure. For the liverwort, increasing Pb exposure resulted in decreases in
content of chlorophyll a, chlorophyll b, and total chlorophyll. Further, the total
chlorophyll content of Hypnumplumaeforme mosses was decreased by 5.8% following
exposure to the highest concentration, while lower exposures slightly elevated
chlorophyll content.
These studies suggest that exposure to Pb has an impact on photosynthetic pigments, but
the exposure methods (seed soaking, spraying of Pb chloride solutions, hydroponic
growth systems) make it difficult to compare these effects to what might occur under the
uncontrolled conditions encountered in natural environments. These experiments bring to
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light the presence of effects, and the underlying mechanisms, but strong uncertainties
remain regarding the natural concentrations at which these effects would be observed.
Response of Antioxidants
Increased antioxidant activity is a common response to Pb exposure, although this
endpoint may not necessarily be an indication of deleterious effects on plant vitality.
Increases in reactive oxygen species with increasing exposure to Pb from 20 mg Pb/kg
soil to 2,000 mg Pb/kg have been demonstrated in broad bean (Viciafaba) (Wang et al.,
2010c: Wang et al.. 2010a: Wang et al.. 2008b) and tomato (Lycopersicon esculentum)
(Wang et al.. 2008a). where they were accompanied up to approximately 500mg Pb/kg by
proportional increases in superoxide dismutase (SOD), glutathione, guaiacol peroxidase,
and lipid peroxidation, as well as decreases in catalase. Spinach seedlings grown in soil
containing six increasing concentrations of Pb from 20 to 520 mg Pb/kg exhibited higher
production of reactive oxygen species, increased rates of lipid peroxidation and increased
SOD concentrations. Many of these responses persisted for 50 days after germination and
growth in the Pb-contaminated soil (Wang et al.. 201 la). Similarly, the bryophyte mosses
Hypnum plumaeforme, Thuidium cymbifolium, and Brachythecium piligerum exposed to
Pb solutions of greater than 0.1 mM Pb for 48 hours exhibited increased production of
O2 radical and H2O2, although no single moss species could be identified as most
sensitive to Pb exposure (Sun et al.. 2011). Increased rates of lipid peroxidation were also
observed in Pb-exposed mosses; however, SOD and catalase activity was suppressed or
unaffected by Pb.
Reddy et al. (2005) found that horsegram (Macrotyloma uniflorum) and bengalgram
(Cicer arietinum) plants watered with Pb solutions containing 200, 500, and 800 mg Pb/L
exhibited increased antioxidant activity: at exposures of 800 mg Pb/L, root and shoot
SOD activity increased to 2-3 times that of controls, and induction was slightly higher in
M. uniflorum. Similarly, catalase, peroxidase, and glutathione-S-transferase activities
were elevated in Pb-stressed plants, but were again higher forM uniflorum. Antioxidant
activities were also markedly greater in the root tissues than the shoot tissues of the two
plants, and were very likely related to the higher Pb accumulation rate of the roots. The
effectiveness of the up-regulation of antioxidant systems in preventing damage from Pb
uptake was evidenced by lower malondialdehyde (MDA) (a chemical marker of lipid
peroxidation) concentration inM uniflorum versus C. arietinum, indicating a lower rate
of lipid peroxidation in response toM uniflorum's higher antioxidant activity.
Gupta et al. (2010) contrasted responses of two ecotypes ofSedum alfredii (an Asian
perennial herb), one an accumulator of Pb collected from a Pb and Zn mining area, and
the other not. Glutathione level was increased in both, and root and shoot lengths were
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decreased following long-term exposures to Pb up to 200 (iM (41.4 mg Pb/L) in
hydroponic solution. However, the accumulator plants exhibited greater SOD and
ascorbate peroxidase activity, likely as a result of greater Pb uptake and a concurrent
increased detoxification capacity. Similar results were reported by Islam et al. (2008):
following Pb exposures of 200 (iM (41.4 mg Pb/L), catalase, ascorbic acid, and
glutathione levels of another Chinese herb, Elsholtzia argyi, were increased, while SOD
and guaiacol peroxidase activities decreased. Microscopic analysis also showed that
affected plants exhibited abnormal chloroplast structures. The response of glutathione
was further confirmed in wheat (Liu et al., 2010a) grown in soils spiked with Pb nitrate.
Evidence of increasing lipid peroxidation (MDA accumulation) with increasing Pb
exposure was also found in mosses (Sun et al.. 2009) and lichens. Lichens field-collected
from the trunks of poplar (Populus tremuld) trees in eastern Slovakia were chemically
analyzed for metal concentration arising from exposure to smelter pollution (Dzubaj et
al., 2008). These concentrations (ranging from 13 to 1,523 mg Pb/kg dry weight) were
assessed in relation to physiological variables, including chlorophyll a and b, carotenoids,
photosystem II activity, CO2 gas exchange (respiration), and MDA content. Lichen Pb
levels were significantly correlated only with MDA content. Determination of plant
chitinase content following exposure to As, Cd and Pb indicated that while levels of these
defense proteins were elevated by As and Cd, chitinase levels were not increased
following exposure to Pb (Bekesiova et al., 2008). As in studies of effects on
photosynthesis, the methods used for exposure make it difficult to compare these effects
to what might occur under the uncontrolled conditions encountered in natural
environments.
Growth
Evidence of effects of Pb on higher growth processes in terrestrial plants was reported in
early NAAQS reviews. Impacts to growth can lead to effects at the population-level of
biological organization and higher (Section 6.1.1). Growth effects of Pb on plants in the
1977 Pb AQCD primarily included visible growth responses observed in laboratory
studies with plants grown in artificial nutrient culture (U.S. EPA. 1977). No Pb toxicity
was observed in plants growing under field conditions at the time of the 1977 Pb AQCD.
Indirect effects of Pb on plant growth (i.e., inhibition of uptake of other nutrients when
Pb is present in the plant) were also reported in the 1977 Pb AQCD. In the 1986 Pb
AQCD mechanisms of Pb effects on growth included reduction of photo synthetic rate,
inhibition of respiration, cell elongation, root development, or premature senescence
(U.S. EPA. 1986b). In 1986, EPA reported that all of these effects were observed to occur
in isolated cells or in plants grown hydroponically in solutions comparable to 1 to 2 mg
Pb/kg soil or in soils with 10,000 mg Pb/kg or greater (U.S. EPA. 1986b). Pb effects on
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other plant processes, especially maintenance, flowering, and hormone development had
not been studied at the time of the 1986 Pb AQCD and remain poorly characterized.
Recent evidence for growth effects in terrestrial plants available since the
2006 Pb AQCD is reviewed below and summarized in Table 6-4. Both growth and
carotenoid and chlorophyll content of Brassica juncea (mustard) plants were negatively
affected by Pb exposure (John et al. 2009). Nominal Pb treatments of 1,500 (iM (311 mg
Pb/L) as Pb acetate solution decreased root lengths and stem heights by 50% after 60
days. Exposure to 600 (iM Pb (124 mg Pb/L) and greater, decreased carotenoid content,
while chlorophyll a was decreased at Pb exposures of 450 (iM (93 mg Pb/L) and higher.
However, when smelter ash-spiked soils containing 1,466 mg Pb/kg (and 18.6 mg Cd/kg)
or 7,331 mg Pb/kg (98.0 mg/kg Cd) were used to grow maize (Zea mays), as well as other
metals in high concentrations, effects were seen in growth or chlorophyll production only
at the higher concentration (Komarek et al., 2009). Given the low solubility of smelter
ash, these observations are consistent with solubility being a key determinant of
bioavailability. Similarly, wheat seedling growth was unaffected when exposed to soil
leachate containing up to 0.7 mg Pb/L for six weeks. Lettuce seedling root growth was
negatively correlated to leachate Pb concentration, but this correlation was only
significant for week 3 and week 6 measurements. Authors concluded that although the
total concentrations of multiple metals in tested soils and leachates exceeded Canadian
Environmental Quality Guidelines, no toxic or only slightly toxic effects occurred
following exposure to the metal mixture (Chapman et al. 2010).
Chinese cabbage (Brassica pekinensis) exposed to Pb-containing soils exhibited
depressed nitrogen assimilation as measured by shoot nitrite content, nitrate reductase
activity, and free amino acid concentration (Xiong et al., 2006). The authors planted
germinated cabbage seeds in soils spiked with Pb acetate to give final soil concentrations
of 0.2 (control), 4, and 8 mmol Pb/kg dry weight total Pb (41, 829, and 1,658 mg Pb/kg)
and collected leaf samples for 11 days. At exposures of 4 and 8 mmol Pb/kg (828.8 and
1,657.6 mg Pb/kg), leaf nitrate content was decreased by 29% and 20% relative to
control. Free amino acid content in exposed plants was 81% and 82% of control levels,
respectively. B. pekinensis shoot biomass was observed to decrease with increasing Pb
exposures, with biomass at the two highest Pb exposures representing 91% and 84% of
control growth, respectively.
Nitrogen, potassium, and phosphorus concentrations in the shoot and root tissues of four
canola cultivars (Brassica napus) also decreased as spiked soil Pb concentrations
increased from 0 to 90 mg/kg. At the highest soil Pb concentration, nitrogen
concentrations were reduced 56% in roots and 58% in shoots versus control levels, while
phosphorous concentrations were reduced 37% and 45%, respectively, and potassium
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content decreased by 42% in both tissues (Ashraf et al.. 2011). Cultivation in Pb-spiked
soils was also linked to decreased shoot and root biomass (32% and 62%, respectively at
90 mg Pb/kg).
Genetic and Reproductive Effects
Exposure to Pb also resulted in genetic abnormalities, including bridges, condensed
bivalents, and laggards, in the meiotic cells of pea plants (Lathyrus sativus) (Kumar and
Tripathi, 2008). Seeds were germinated in soils amended with Pb nitrate at concentrations
of 25, 50, 100, 200, and 300 mg Pb/kg, and concentrations of 100 mg Pb/kg and greater
were found to be genotoxic or detrimental to pea viability. Cenkci et al. (2010) exposed
fodder turnip (B. rapa) to 0.5 to 5 mM of Pb nitrate (103.6 to 1036 mg Pb/L) for 6 days
and showed decreased genetic template stability (as quantified by random amplified
polymorphic DNA profiles) and decreased photosynthetic pigments.
Two genotypes of maize seedlings exhibited a significant and concentration-dependent
reduction in seed germination following 7 days of Pb treatment in nutrient solution of
0.01, 0.1 and 1 mg Pb/L as Pb sulfate (Ahmad et al.. 2011). Pb exposure also decreased
germination rate and growth, and increased pollen sterility in radish grown for 50 days in
hydroponic solutions containing 0.5 mM Pb (104 mg Pb/L) (Kumar and Tripathi. 2008).
Plants exposed to Pb exhibited decreased growth, curling and chlorosis of young leaves,
and decreased root growth. In addition, Gopal and Rizvi (2008) showed that Pb exposure
increased uptake of phosphorus (P) and iron (Fe) and decreased sulfur (S) concentration
in radish tops.
Interestingly, as in zebra finch (Section 6.3.3.3) Ca2+ was found to moderate the effects of
Pb in both monocotyledon and dicotyledon plant seedlings, with tomato (Lycopersicon
esculentum), rye (Lolium sp.), mustard, and maize plants exhibiting increased tolerance
to Pb exposures of 5, 10, and 20 mg Pb/L in the presence of Ca2+ concentration of 1.2
mM (249 mg Pb/L) and higher (Antosiewicz. 2005).
6.3.4.2 Terrestrial Invertebrates
Exposure to Pb also causes antioxidant effects, reductions in survival and growth, as well
as decreased fecundity in terrestrial invertebrates as summarized in the 2006 Pb AQCD
(U.S. EPA. 2006b). Alterations in reproduction, growth and survival at the species level
can lead to effects at the population-level of biological organization and higher
(Section 6.1.1). In addition to these endpoints, recent literature also indicates that Pb
exposure can cause significant neurobehavioral aberrations, and in some cases,
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endocrine-level impacts. Second-generation effects have been observed in some
invertebrate species.
The morphology of y-aminobutyric acid (GAB A) motor neurons in Caenorhabditis
elegans nematodes was affected following exposure to Pb nitrate for 24 hours (Du and
Wang. 2009). The authors determined that exposures as low as 2.5 (iM Pb nitrate (0.5 mg
Pb/L) could cause moderate axonal discontinuities, and observed a significant increase in
the number of formed gaps and ventral cord gaps at Pb nitrate exposures of 75 and
200 (iM (6 and 41 mg Pb/L). Younger C. elegans larvae were more likely to exhibit
neurobehavioral toxicity symptoms in response to Pb exposure at 2.5 (iM (0.5 mg Pb/L)
(Xing et al.. 2009b). Neural degeneration, as demonstrated by dorsal and ventral cord
gaps and neuronal loss was also more pronounced in young larval C. elegans than in
older larvae and adults (Xing et al.. 2009c). C. elegans nematodes exposed to Pb
concentration as low as 2.5 (iM (0.5 mg Pb/L) for 24 hours also exhibited significantly
altered behavior characterized by decreased head thrashes and body bends. Exposures of
50 (iM Pb (10 mg Pb/L) and greater decreased the number of nematode forward turns
(Wang and Xing. 2008). Chemotaxis toward NaCl, cAMP, and biotin was also decreased
in C. elegans nematodes exposed to Pb concentration greater than 2.5 (iM (0.5 mg Pb/L)
(Xing et al.. 2009a). This evidence suggests that Pb may exert neurotoxic action in
invertebrates as it does in vertebrates. However, it is unclear how these behavioral
aberrations would affect fitness or survival (Wang and Xing. 2008).
In a study of C. elegans exposed to 4 sub-lethal concentrations of Pb nitrate between 25
and 100 (iM (5 and 21 mg Pb/L), Vigneshkumar et al. (2013) observed upregulation of
both catalase and antimicrobial response -related genes. When challenged with addition of
a pathogenic strain of Pseudomonas aeruginosa, exposed C. elegans showed greater
resistance to microbial colonization than controls.
Younger individuals also appear to be more sensitive to the reproductive effects of Pb
exposure. Guo et al. (2009) showed that concentrations of 2.5, 50, and 100 (iM Pb (0.5,
10, and 21 mg Pb/L) had greater significant adverse effects on reproductive output when
early-stage larval C. elegans were exposed. Adult C. elegans exhibited decreased brood
size only when exposed to the highest Pb concentration.
The progeny of C. elegans nematodes exposed nominally to 2.5, 75, and 200
Pb nitrate (0.5, 16, and 41 mg Pb/L) exhibited significant indications of multi-
generational toxicity (Wang and Yang. 2007). Life spans of offspring were decreased by
increasing parental Pb exposure, and were comparable to the reductions in parental life-
spans. Similarly, diminished fecundity was observed in the progeny of C elegans
exposed to Pb (9%, 19%, and 31% reductions of control fecundity, respectively),
although the decrease was smaller than in the exposed parental generation (reductions of
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52%, 58%, and 65%, respectively). Significant behavioral defects affecting locomotion
were also observed in the offspring, but these impacts were not determined to be
concentration-dependent. Reproductive effects of Pb exposure were also observed in
springtails F. Candida following 10-day exposure to Pb-spiked soils. Egg hatch
significantly decreased at nominal concentrations of 1,600 mg Pb/kg dry soil and higher
and the EC50 for hatching was 2,361 mg Pb/kg dry soils (Xu et al.. 2009b).
E. andrei earthworms exposed to 21 different soils, each containing 2,000 mg Pb/kg
freshly added Pb, for 28 days exhibited highly variable mortality, ranging from 0% to
100%, (Bradham et al.. 2006). Pb body burden of exposed worms ranged from 29 to
782 mg Pb/kg. Internal Pb concentration was also negatively correlated to reproductive
output. CEC and pH were found to be the principal soil characteristics determining the
differences in those effects, although the apparent role of CEC may only have been due to
its correlation with other soil characteristics. Low soil Pb concentration (5 mg Pb/kg) also
decreased the protein content of E. fetida earthworms during a 7-day exposure (Li et al..
2009b). Higher Pb concentration had no effect on protein production. However, cellulase
activity was increased by the 7-day exposures to Pb at all exposure concentrations (31%,
13%, and 23% of control activity at exposures of 5, 50, and 500 mg Pb/kg, respectively),
which the authors reported as an indication of detrimental effects on worm metabolism.
By contrast, Svendsen et al. (2007) found thatZ. rubellus earthworms exposed for 42
days to field-collected smelter-polluted soils containing average Pb concentration of 106,
309, and 514 mg Pb/kg dry weight exhibited normal survival and cocoon production
rates, even though they accumulated more Pb with increased environmental
concentration. The much smaller effect may be explained by the increased aging time
undergone by field soil, causing a larger fraction of the total Pb to be complexed and
sequestered by organic and inorganic compounds. Similarly, earthworms (E. fetida)
exposed to field-collected soils with concentrations of Pb and As up to 390 mg/kg and
128 mg/kg, respectively, due to historical treatments of Pb-arsenate pesticides, exhibited
no change in survival, behavior or morphology (Delistratv and Yokel. 2011). Soil aging
(e.g., from of the time of Pb-arsenate applications in 1942 to soil collection in
approximately 2009) likely reduced Pb bioavailability to earthworms.
As in plants, induction of metal chelating proteins and antioxidant activity in
invertebrates is affected by exposure to Pb. Metallothionein production in earthworms
(Lampito mauritii) was significantly induced following exposure to Pb-contaminated soil.
Tissue metallothionein levels increased after a two-week exposure to 75 to 300 mg Pb/kg
soil, although by 28 days levels had begun to decrease, perhaps as a result of Pb toxicity
(Maity et al., 2011). Further, the induction of antioxidant activity was correlated to
standard toxicity measurements in Thebapisana snails (Radwan et al.. 2010). Topical
application of Pb solutions (estimated to be 500 to 2,000 (ig Pb per animal) to snails
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resulted in decreased survival, increased catalase and glutathione peroxidase activities,
and decreased glutathione concentration. The 48-hour LD50 concentration was
determined to be 653 (ig per snail, as measured in digestive gland tissue. Snail
glutathione content was decreased at exposures of 72.2% of the 48-hour LD50 value,
while Pb exposure at 40% of the 48-hour LD50 value induced catalase and glutathione
peroxidase activities.
Dietary exposure to Pb also affected T. pisana snail growth. After three weeks on
Pb-contaminated diet, snail feeding rates were depressed by all Pb exposures (nominal
concentration of 50 to 15,000 mg Pb/kg diet dry weight) (El-Gendv et al.. 2011). A five
week dietary exposure to 1,000 mg Pb/kg and greater resulted in reduced snail growth.
Food consumption, growth, and shell thickness were also observed to decrease with
increasing diet Pb in juvenile A. achatina snails (7 levels between 0 and 1,344 mg Pb/kg
feed, for 12 weeks) (Ebenso and Ologhobo. 2009a). A similar depression of growth was
observed in sentinel juvenile A. achatina snails deployed at Pb-polluted sites in the Niger
Delta region of Nigeria. Although snail mortality was not increased significantly by
exposure to soil Pb up to 1,200 mg Pb/kg, a concentration-dependent relationship was
established for growth, with significant reduction observed at 12-week exposures to
20 mg Pb/kg (Ebenso and Ologhobo. 2009b). However, consumption of field-collected
Pb-polluted U. dioica leaves containing 3 mg Pb/kg stopped all reproductive output in
C. nemoralis. Snails also exhibited diminished food consumption rates when offered
leaves with both low (1.5 mg Pb/kg) and high Pb content, but the mechanism of the
dietary aversion was not defined (Notten et al.. 2006).
Chronic dietary exposure to Pb was also examined in post-embryonic oribatid mites
(Archegozetes longisetosus) (Kohler et al., 2005). Both algae and bark samples were
soaked in 100 mg/L Pb as Pb nitrate and provided as diet and substrate, respectively, to
larval mites. In addition to elevated heat shock proteins (hsp70), 90.8% of the
protonymphs exhibited significant leg deformities, including abnormal claws, shortened
and thickened legs, and translocated setae. Although not specifically discussed, it is very
likely that these deformities would decrease mite mobility, prey capture, and reproductive
viability. While there is some evidence that oribatid mites exhibit Pb avoidance behavior,
this response may not significantly reduce Pb exposure and effects. Although soil-
inhabiting mites (Oppia nitens) were observed to avoid high Pb concentrations, the EC50
for this behavior was approximately five times higher than the chronic EC50 for
reproduction (8,317 and 1,678 mg Pb/kg, respectively) (Owojori et al., 2011).
Consequently, it is unlikely that oribatid mites will avoid soils containing toxic Pb
concentrations.
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Lock et al. (2006) compared the toxicity of both laboratory-spiked soils and field-
collected Pb-contaminated soils to springtails (F. Candida). The 28-day EC50 values
derived for F. Candida ranged from 2,060 to 3,210 mg Pb/kg in leached and unleached
Pb-spiked soils, respectively, whereas field-collected soils had no significant effect on
springtail reproduction up to (but not including) 14,436 mg Pb/kg (Lock et al., 2006).
Consequently, leaching soils prior to use in bioassays had only a slight effect on Pb
toxicity to resident springtails, and did not provide an appropriate model for field-
weathered, Pb-contaminated soils. This indicates that physiochemical factors other than
leaching may be more important determinants of Pb bioavailability. A 4-week exposure
to Pb-amended soils containing up to 3,200 mg Pb/kg (nominal concentration) had no
significant effect on Sinella curviseta springtail survival or reproduction (Xu et al..
2009a).
Carabid beetles (Pterostichus oblongopunctatus) inhabiting soils contaminated by
pollution from a Pb-Zn smelter (containing 136 to 2,635 mg Pb/kg) were field-collected
and then laboratory-reared for two generations (Lagisz and Laskowski. 2008). While
fecundity was positively correlated to soil metal concentration (e.g., more eggs were
produced by females collected from contaminated areas), the hatching rate of eggs
diminished with increasing soil metal contamination. For the Fl generation, females
produced by parents inhabiting highly polluted areas exhibited decreased body mass. The
authors stated that these results indicate that invertebrates inhabiting metal- (or Pb-)
contaminated soils could face "significantly altered life-history parameters." Similarly,
aphids (Brevicoryne brassicae) reared on cabbage and radish plants exposed to 0.068 mg
Pb daily exhibited altered development and reproduction when compared to those reared
on non-exposed plants. Development time was increased by approximately two days,
which led to a reduction in relative fecundity (Goriir. 2007). Although the authors noted
that study exposures were greater than what would be expected in naturally polluted
areas, Pb exposure under field conditions could alter invertebrate life history patterns.
Several studies suggest that Pb may disrupt hormonal homeostasis in invertebrates. Shu
et al. (2009) reported that vitellogenin production in both male and female S. litura moths
was disrupted following chronic dietary exposure to Pb. Adult females reared on diets
containing 25, 50, 100, or 200 mg Pb/kg exhibited decreased vitellogenin mRNA
induction, and vitellogenin levels decreased with increasing Pb exposure. In addition,
vitellogenin mRNA induction was detected in males exposed to 12 and 25 mg Pb/kg, and
low levels of vitellogenin were found at those lower Pb exposures, when males normally
do not produce any. In the Asian earthworm (Pheretima guillelmi), sperm morphology
was found to be altered significantly following 2-week exposure to soils containing
nominal concentration of 1,000, 1,400, 1,800, and 2,500 mg Pb/kg (Zheng and Li. 2009).
Common deformities were swollen head and head helices, while head bending was also
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recorded in some cases. These deformities were observed following exposures to
concentration below the 14-day LC50 (3,207 mg Pb/kg) and below the concentration at
which weight was diminished (2,800 mg Pb/kg). Experimentation with the model
organism Drosophila indicates that Pb exposure may increase time to pupation and
decrease pre-adult development, both of which are endocrine-regulated (Hirsch et al..
2010).
6.3.4.3 Terrestrial Vertebrates
Pb poisoning is one of the earliest recognized toxicoses of terrestrial vertebrates,
occurring primarily through the ingestion of spent shot by birds. While the focus of the
ISA is on more environmentally relevant exposures, studies of Pb poisoning provide
historical context for the review. The widespread nature of this toxicosis was first noticed
in American waterfowl around the turn of the last century [see Jones, (1939) for an
historical summary]. Wetmore (1919) demonstrated that Pb shot caused the observed
effects and described in detail the species affected, associated symptoms, and additional
factors involved. By 1959, the estimated annual loss of waterfowl to Pb poisoning was
2-3 percent of the fall population (Bellrose. 1959). Smaller numbers of shorebirds and
upland game birds were also found poisoned by Pb (Locke and Thomas. 1996).
The first reported Pb poisoning of a bald eagle (Haliaeetus leucocephalus) was described
by Mulhern et al. (1970). and subsequently several hundred bald eagle Pb poisonings
were diagnosed throughout the U.S. prior to the ban on use of Pb shot for waterfowl
hunting (Kramer and Redig. 1997). Eagles and other raptors are poisoned by consuming
Pb pellets imbedded in the flesh of ducks or upland prey species and may also be exposed
to other sources of Pb, such as fishing sinkers and weights (Kramer and Redig. 1997).
The use of Pb shot for waterfowl hunting was banned in 1991 due to the poisoning of
bald eagles, which had been previously added to the endangered species list and were
specially protected under the Bald Eagle Protection Act of 1940.
Anderson et al. (2000) reported that by 1997, mallard (Anas platyrhynchus) deaths from
Pb poisoning in the Mississippi flyway were reduced by 64 percent, and ingestion of
toxic pellets had declined by 78 percent. They estimated the ban prevented approximately
1.4 million duck deaths in the first 6-year period. However, Pb exposure remains
widespread in bald eagles, although blood Pb concentrations have significantly decreased
(Kramer and Redig. 1997). The endangered California condor (Gymnogyps
californianus) also continues to have significantly elevated blood Pb levels as well as
Pb-associated mortality resulting from exposure to ammunition fragments contained in
food items (Cade. 2007; Church et al.. 2006). Although there is a significant amount of
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information on Pb tissue residues of mammals, there are very few reports of Pb
poisoning; exceptions are reports of Pb poisoned bats in a cave in the southern U.S. and
small mammals in the vicinity of several smelters (Shore and Rattner. 2001).
At the time of the 1977 Pb AQCD few studies of the effects of exposure to Pb had been
conducted in wild animals other than birds, and the majority of those studies were of
direct poisoning (U.S. EPA. 1977). Several studies of domestic animals grazing near Pb
smelters indicated that horses are more susceptible than cattle to chronic Pb exposure
although the findings were not conclusive due to the presence of other metals. Delta-
aminolevulinic acid dehydratase (ALAD) was recognized as a sensitive indicator of Pb
exposure in rats and waterfowl. In the 1986 Pb AQCD, additional effects of Pb on small
mammals and birds were reported. According to the 2006 Pb AQCD (U.S. EPA. 2006b).
commonly observed effects of Pb on avian and mammalian wildlife include decreased
survival, reproduction, and growth, as well as effects on development and behavior. More
recent experimental data presented here expand and support these conclusions, and also
indicate that Pb can exert other effects on exposed terrestrial vertebrates, including
alteration of hormones and other biochemical variables.
Since the 2006 Pb AQCD, there is additional evidence for hematological effects of Pb
exposure in terrestrial vertebrates. Red-backed salamanders (Plethodon cinereus)
exposed to Pb-amended soils (553, 1,700, 4,700, and 9,167 mg Pb/kg) by Bazar et al.
(2010) exhibited lowered appetite and decreased white blood cell counts at the two
highest concentrations, but tolerated field-collected, aged soils containing Pb
concentrations of up to 16,967 mg Pb/kg with no significant deleterious effects. The
white blood cell count of adult South American toads, (Bufo arenarum) was also
decreased by weekly sublethal i.p. injections of Pb acetate at 50 mg Pb/kg body weight,
(Chiesaetal.. 2006). The toads also showed altered serum profiles and increased number
of circulating blast cells. Final toad blood Pb levels were determined to be 8.6 mg Pb/dL,
although it is unclear whether this is representative of Pb concentrations observed in field
B. arenarum populations exposed to Pb. The authors suggested that, based on these
findings, long-term environmental exposure to Pb could affect toad immune response. In
western fence lizards (S. occidentalis), sub-chronic (60-day) dietary exposure to 10 to
20 mg Pb/kg per day resulted in significant sublethal effects, including decreased cricket
consumption, decreased testis weight, decreased body fat, and abnormal posturing and
coloration (Salice et al.. 2009). Long-term dietary Pb exposures are thus likely to
decrease lizard fitness.
Even in cases of high environmental Pb exposures, however, linking Pb body burdens to
biological effects can be difficult. Pb concentration in the breast feathers, washed tail
feathers, and blood of field-collected blackbirds (Turdus merula) were determined to be
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3.2 mg Pb/kg, 4.9 mg Pb/kg, and 0.2 mg Pb/kg wet mass in urban birds, as opposed to
1.4 mg Pb/kg, 1 mg Pb/kg, and 0.05 mg Pb/kg in rural birds (Scheifler et al.. 2006a).
However, the elevated Pb tissue concentrations in urban birds were not significantly
correlated to any index of body condition.
The long-term effect of atmospheric Pb deposition on pied flycatcher (Ficedula
hypoleucd) nestlings was determined in native communities residing in the Laisvall
mining region of Sweden (Berglund et al.. 2010). Moss samples indicated that Pb
deposition in study areas ranged between 100 and 2,000 mg Pb/kg dry weight during
operations and 200 and 750 mg Pb/kg when operations ceased. A simultaneous slight
reduction was observed in pied flycatcher blood Pb levels, from 0.4 to 0.3 mg Pb/kg.
However, clutch size was decreased in pied flycatchers inhabiting the mining area both
during and after mining operations, and mean nestling mortality was 2.5 times higher in
the mining region than in reference areas during mining operations, and 1.7 higher five
years after cessation of mining operations. The authors noted that Pb deposition in the
mining region remained elevated even after mining operations ceased, and that stable Pb
isotope analysis suggested that smelter Pb remained available to pied flycatcher through
the transfer of historically deposited Pb in soil to prey items.
Berglund et al. (2010) also analyzed ALAD activity in pied flycatchers at the later period,
and found that it was 46% lower at the mine site. Beyer et al. (2004) observed that
elevated blood Pb levels in several types of birds inhabiting the Tri-State Mining District
(Oklahoma, Kansas, Missouri) were correlated with decreases in ALAD activity. Based
on reduction in ALAD activity, robins (Turdus migratorius) were most sensitive to Pb
exposure (35% reduction), followed by cardinals (Cardinalis cardinalis), waterfowl, and
bobwhite quail (Colinus virginianus) (40%, 41%, and 56% reductions, respectively).
Eagle owl (Bubo bubo) nestlings living in a historical mining area in Spain also exhibited
elevated blood Pb levels (average 8.61 (ig/dL as compared to an average reference area
value of 3.18 (ig/dL), and this was correlated to an approximate 60% reduction in ALAD
activity (Gomez-Ramirez et al.. 2011). Hansen et al. (2011 a) determined that ground-
feeding songbirds were frequently exposed to Pb within the Coeur d'Alene, ID mining
region. Robins, in particular, were significantly likely to exhibit blood Pb levels in the
clinical and severe clinical poisoning ranges (50 to 100 (ig/dL and >100 (ig/dL,
respectively). Ingested soil Pb accounted for almost all of the songbirds' exposure to Pb,
with Pb exposure correlated with estimated soil ingestion rates (20% for robins, 17% for
song sparrows, and 0.7% for Swainson's thrushes, Catharus ustulatus). More than half of
the robins and song sparrows from all contaminated sites and more than half of the
Swainson's thrushes from highly contaminated sites showed at least 50% inhibition of
ALAD. The highest hepatic Pb concentration of 61 mg/kg (dry weight) was detected in a
song sparrow (Hansen et al.. 201 la).
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Blood Pb was significantly elevated in waterfowl in the Lake Coeur d'Alene areas of
Blackwell Island and Harrison Slough (mean sediment concentrations of 679 and
3,507 mg Pb/kg dry weight, respectively). Twenty-seven percent of the waterfowl
sampled in the Blackwell Island region had blood Pb concentrations suggestive of severe
clinical poisoning (average concentration=0.17 mg Pb/kg); in the Harrison Slough, 60%
of sampled waterfowl had highly elevated blood Pb levels that exceeded the severe
clinical poisoning threshold (average concentration=2.2 mg Pb/kg) (Spears et al., 2007).
The level of corticosteroid hormones in field populations of white stork nestlings
(Ciconia ciconid) in a mining area affected by Pb and other metals was positively
correlated with blood Pb levels (Baos et al., 2006). The effect was more pronounced for
single nestlings than for multiple-chick broods. Surprisingly, average blood Pb levels in
chicks inhabiting reference areas was 910 (ig Pb/dL (± 51), which was higherthan blood
Pb levels from the mining area (440 ± 340 (ig Pb/dL). However, the correlation between
blood Pb levels and the corticosteroid stress response in white stork nestlings was
observed in both groups of birds. Burger and Gochfeld (2005) exposed herring gull
(Lams argentatus) chicks to Pb acetate via an i.p. injection of 100 mg Pb/kg body
weight, to produce feather Pb concentration approximately equivalent to those observed
in wild gulls. Pb-exposed gulls exhibited abnormal behaviors, including decreased
walking and food begging, erratic behavioral thermoregulation, and diminished
recognition of caretakers. Interestingly, subchronic exposure of Japanese quail (Coturnix
coturnixjaponica) to 5 and 50 mg Pb/L in drinking water caused an increase in their
immune response. Exposed quail exhibited significantly lower rates of death or health
effects (including septicemia, perihepatitis, and pericarditis among others) than control
animals following infection with Escherichia coli, and the incidence of infection-related
effects was dependent on Pb exposure (Nain and Smits. 2011). These observations
contrast with immunotoxicology results in mice reported in Section 4.6.5.1.
Again, dietary or other health deficiencies unrelated to Pb exposure are likely to
exacerbate the effects of Pb. Ca2+-deficient female zebra finches (T. guttatd) had a
suppressed secondary humoral immune response following 28-day exposures to 20 mg
Pb/L in drinking water (Snoeijs et al.. 2005). This response, however, was not observed
in birds fed sufficient Ca2+. Although a significant finding, these data are difficult to
interpret under field conditions where the overall health of avian wildlife may not be
easily determined.
Chronic Pb exposures were also demonstrated to affect several mammalian species.
Young adult rats reared on a diet containing 1,500 mg Pb/kg Pb acetate for 50 days
demonstrated less plasticity in learning than non-exposed rats (McGlothan et al., 2008).
indicating that Pb exposure caused significant alteration in neurological function. Yu et
al. (2005) showed that dietary Pb exposure affected both the growth and endocrine
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function of gilts (S. domesticd). Consumption of 10 mg Pb/kg diet resulted in lower body
weight and food intake after 120 days of dietary exposure; Pb exposure decreased final
weight by 8.2%, and average daily food intake of Pb-exposed pigs was decreased by
6.8% compared to control intake. Additionally, concentration of estradiol, luteinizing
hormone, and pituitary growth hormone were decreased (by 12%, 14%, and 27% versus
controls, respectively), while blood Pb level was increased by 44% to an average
2.1 (ig/dL. In cattle grazing near Pb-Zn smelters in India, blood Pb levels were positively
correlated with plasma levels of the thyroid hormones thyroxine (T4) and tri-
iodothyronine (T3) and the hepatic biomarkers alanine transaminase and aspartate
transaminase (Swarup et al., 2007). Total lipids, total protein and albumin levels were
decreased in the same animals. Rodriguez-Estival et al. (2011) determined that red deer
(Cervus elaphus) and wild boar (Sus scrofd) inhabiting a Pb-contaminated mining area in
Spain exhibited increased liver and bone Pb concentrations (geometric means of 0.35 and
0.46 mg Pb/kg for red deer, and 0.81 and 7.36 mg Pb/kg for wild boar, respectively).
These tissue concentrations were correlated to a significant decrease in red deer
glutathione production, but corresponded to an increase in wild boar glutathione
(Rodriguez-Estival et al.. 2011). Authors proposed that the different antioxidant
responses may be indicative of different Pb susceptibilities in the two species.
Following previous reports of in vivo follicle and oocyte damage in animals with
low-level Pb accumulation, Nandi et al. (2010) treated oocytes of buffalo (Bubalus
bubalis) in vitro with Pb at concentrations ranging from 0.005 to 10 mg/L in one-day
cultures indicated a significant decline in viability of oocytes at 1 mg/L. Dose-dependent
effects on oocyte viability, morphological abnormalities, cleavage, blastocyst yield and
blastocyst hatching were observed in Pb-treated oocytes with maturation significantly
reduced at 2.5 mg/L and 100% oocyte death at 32 mg/L. These results appear to confirm
previous reports, but the in vitro concentrations of Pb are difficult to relate to in vivo
exposures. On the other hand, the reproductive viability of wild red deer from the
Pb-contaminated mining area of Spain studied by Rodriguez-Estival et al. (2011) was
shown to be altered, with 11% and 15% reductions in spermatozoa and acrosome
integrity observed in male deer from the mining area compared with those residing in
reference areas (Reglero et al.. 2009a).
6.3.5 Exposure and Response of Terrestrial Species
Evidence regarding exposure-response relationships and potential thresholds for Pb
effects on terrestrial populations can inform determination of standard levels that are
protective of terrestrial ecosystems. Given that exposure to Pb may affect plants,
invertebrates, and vertebrates at the organism, population, or community level,
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determining the rate and concentration at which these effects occur is essential in
predicting the overall risk to terrestrial organisms. This section updates available
information derived since the 2006 Pb AQCD, summarizing several dose-response
studies with soil invertebrates. As shown in the studies summarized in Table 6-4 and
Figure 6-4. several experiments have been published that used multiple levels of Pb under
controlled conditions. However, none of them treated Pb concentration as a continuous
variable, i.e., none attempted to analyze results as a concentration-response relationship.
In addition, given the well-established presence of strong interactions with variables such
as pH, CEC, OC, or aging, applying exposure-response relationships from those
experiments to natural conditions with different values of those interacting variables
could be difficult.
Dose-dependent responses in antioxidant enzymes were observed in adult L. mauritii
earthworms exposed to soil-associated Pb contamination (75, 150, 300 mg Pb/kg) (Maity
et al.. 2008). By day seven of exposure, glutathione-S-transferase activity and glutathione
disulfide concentration were positively correlated with increasing Pb exposures, while
glutathione concentration exhibited a negative dose-response relationship with soil Pb
concentration. However, these trends had become insignificant by the end of the total
exposure period (28 days), as a result of normalization of antioxidant systems following
chronic exposure. This strongly suggests that changes to earthworm antioxidant activity
are an adaptive response to Pb exposures.
Both survival and reproductive success ofE.fetida earthworms showed concentration-
dependent relationships with soil Pb concentration during the course of standard 14- and
56-day toxicity tests (Jones et al.. 2009b). Five levels of Pb soil concentration were
prepared for the acute 14-day study via spiking with Pb nitrate—0, 300, 711, 1,687, and
2,249 mg Pb/kg, while soil concentration of 0, 355, 593, 989, and 1,650 mg Pb/kg were
used in chronic (56-day) earthworm bioassays. A 14-day acute LC50 of 2,490 mg Pb/kg
was determined from the dose-response relationship, while the approximate 56-day
NOEC (no observed effect concentration) and EC50 values were about 400 mg Pb/kg and
1,000 mg Pb/kg, respectively. Jones et al. (2009b) made use of continuous (regressional)
models to characterize the relationship between Pb soil concentration and Pb
accumulation in earthworms, but did not use continuous models for the relationship of
exposure and other responses. Currie et al. (2005) observed mortality of E. fetida after 7
and 14 days in spiked field soil at seven levels of Pb (0 to 10,000 mg Pb/kg). They
reported LC50 values of 2,662 mg Pb/kg at 7 days and 2,589 mg Pb/kg at 14 days or
2,827 mg Pb/kg at both 7 and 14 days, depending on the number of worms in the
experimental enclosure.
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Other studies have shown no correlation between Pb concentration in either earthworm
tissue or soil, and earthworm survival rate. Although the Pb content ofE.fetida held in
metal-contaminated soils containing between 9.7 and 8,600 mg Pb/kg was positively
correlated with Pb concentration of fully aged soil collected from disused mines, there
was no statistical relationship with earthworm survival during a 42-day exposure period
(Nahmani et al.. 2007). However, Pb concentration in soil leachate solution was
significantly correlated with decreased earthworm survival and growth (linear regression:
R2= 0.64, p<0.0001). The 42-day Pb EC50 for E. fetida growth was 6,670 mg Pb/kg.
Langdon et al. (2005) exposed three earthworm species (E. andrei, L. rubellus, and
A. caliginosa) to Pb nitrate-amended soils at concentrations of 1,000 to 10,000 mg Pb/kg
to determine species variability in uptake and sensitivity. Twenty-eight-day LC50 values
for the three species were 5,824 mg Pb/kg, 2,867 mg Pb/kg, and 2,747 mg Pb/kg,
respectively, indicating thatZ. rubellus and A. caliginosa are significantly more
vulnerable to Pb contamination than E. andrei, a common laboratory species. This is
comparable to previous findings by Spurgeon et al. (1994) who reported 14-day LC50 of
4,480 mg Pb/kg and 50-day LC50 of 3,760 mg Pb/kg forE.fetida, another standard
laboratory test species. In the more recent study of E. fetida sensitivity summarized
above, Jones et al. (2009b) reported LC50 values for E. fetida that are similar to those for
L. rubellus and A. caliginosa. It is likely that these apparent species differences are a
result of differential bioavailability of the Pb in test soils. However, the Pb body burden
of all three species in the study by Langdon et al. (2005) increased with increasing
environmental concentration, and there were no species differences in Pb tissue content.
When given a choice between treated and untreated soils, all worm species exhibited
significant avoidance of Pb-contaminated soils, and altering pH (and, consequently, Pb
bioavailability) had no impact on avoidance (Langdon et al.. 2005). Field earthworms
may thus be able to reduce their exposure to Pb through behavior.
Reproductive success of other soil invertebrates is impacted by Pb. The organismal and
population-level responses of the springtail Paronychiurus kimi to Pb were determined by
Son et al. (2007) using artificial soils, following the 1999 ISO methodology. The 7-day
Pb LC50 was determined to be 1,322 mg Pb/kg dry weight, while the 28-day reproduction
EC50 was established as 428 mg Pb/kg. The intrinsic rate of population increase was
lower at a Pb soil concentration of 1,312 mg Pb/kg, and the authors estimated that, at this
level, P. kimi populations would be extirpated. The authors noted that, in this case, the
reproductive endpoint overestimated the population-level risk for P. kimi springtails
exposed to Pb, and proposed that more specific measures of population-level endpoints
(such as the reduction in intrinsic rate of increase) be used to determine risk to
populations. Menta et al. (2006) showed that a nominal soil concentration of 1,000 mg
Pb/kg decreased the reproductive output of two collembolans, Sinetta coeca and
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F. Candida. Pb concentrations of 50, 100, and 500 mg Pb/kg slightly but significantly
depressed S. coeca adult survival, while F. Candida survival was statistically unaffected
by Pb exposure. The hatching success ofF. Candida eggs was diminished by 10-day
exposure to Pb-spiked soils; the 10-day EC50 for hatching success was reported as
2,361 mg/kg Pb (Xu et al.. 2009b). However, authors noted that egg development was
more sensitive to Cu and Zn exposure, and by comparison, was less susceptible to Pb.
In addition to species variability, physical and chemical factors affecting Pb
bioavailability were also demonstrated to significantly influence the toxicity of Pb to
terrestrial species. As noted previously in Section 6.3.2. laboratory-amended artificial
soils provide a poor model for predicting the toxicity of Pb-contaminated field soils,
because aging and leaching processes, along with variations in physiochemical properties
(pH, CEC, OM), influence metal bioavailability. Consequently, toxicity values derived
from exposure-response experimentation with laboratory-spiked soils probably
overestimate true environmental risk, with the possible exception of highly acidic sandy
soils. Because toxicity is influenced by bioavailability and soil biogeological and
chemical characteristics, extrapolation of toxic concentrations between different field-
collected soils will be difficult. Models that account for those modifiers of bioavailability,
such as the terrestrial BLM proposed by Smolders et al. (2009). have proven difficult to
develop due to active physiological properties of soil organisms affecting either uptake
(such as root phytochelatins) or sequestration of Pb (such as granule formation in root
tissues and earthworms, or substitution of Pb for calcium in bones).
6.3.6 Terrestrial Community and Ecosystem Effects
A study reviewed in the 1977 Pb AQCD provided evidence for Pb effects on forest-
nutrient cycling and shifts in community composition. Reduced arthropod density,
biomass and richness were observed in the vicinity of a smelting complex in southeastern
Missouri where Pb, Cd, Zn and Cu were measured in the litter layer and soil (U.S. EPA.
1977; Watson et al.. 1976). In the 1986 Pb AQCD it was reported that Pb at
environmental concentrations occasionally found near roadsides and smelters (10,000 to
40,000 mg Pb/kg dry weight) can eliminate populations of bacteria and fungi on leaf
surfaces and in soil. At soil concentrations of 500 to 1,000 mg Pb/kg or higher,
populations of plants, microorganisms, and invertebrates may shift toward Pb-tolerant
populations of the same or different species (U.S. EPA. 1986b).
According to the 2006 Pb AQCD (U.S. EPA. 2006b). natural terrestrial ecosystems near
significant Pb stationary sources (such as smelters and mines) exhibited a number of
ecosystem-level effects, including decreased species diversity, changes in floral and
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faunal community composition, and decreasing vigor of terrestrial vegetation. These
findings are summarized in Table AX7-2.5.2 of the Annex to the 2006 Pb AQCD (U.S.
EPA, 2006c). More recent literature explored the interconnected effects of Pb
contamination on soil bacterial and fungal community structure, earthworms, and plant
growth, in addition to impacts on soil microbial community function.
Inoculation of maize plants with Glomus intraradices arbuscular mycorrhizal fungi
isolates decreased Pb uptake from soil, resulting in lower shoot Pb concentration and
increased plant growth and biomass (Sudova and Vosatka. 2007). Similarly, Wong et al.
(2007) showed that the presence of arbuscular mycorrhizal fungi improved vetiver grass
(Vetiveria zizanioides) growth, and while Pb uptake was stimulated at low soil
concentration (10 mg Pb/kg), it was depressed at higher concentration (100 and 1,000 mg
Pb/kg). Bojarczuk and Kieliszewska-Rokicka (2010) found that the abundance of
ectomycorrhizal fungi was negatively correlated with the concentration of metals,
including Pb, in the leaves of silver birch seedlings. Arbuscular mycorrhizal fungi may
thus protect plants growing in Pb-contaminated soils. Microbes too may dampen Pb
uptake and ameliorate its deleterious effects: biomass of plants grown in metal-
contaminated soils (average Pb concentration 24,175 mg Pb/kg dry weight) increased
with increasing soil microbial biomass and enzymatic activity (Epelde et al.. 2010).
However, above certain Pb concentration, toxic effects on both plants and microbial
communities may prevent these ameliorating effects. Yang et al. (2008b) found that both
the mycorrhizal colonization and the growth ofSolidago canadensis were negatively
affected by soil Pb contamination. They suggested that, more generally, Pb-mediated
alterations in plant-fungal dynamics may be the cause of ecological instability in
terrestrial vegetative communities exposed to metals.
The presence of both earthworms and arbuscular mycorrhizal fungi decreased the
mobility of Pb in mining soils undergoing phytoremediation (Ma et al.. 2006).
Inoculation with both earthworms and fungi increased plant growth at sites contaminated
with mine tailings compared to that observed at sites with 75% less Pb contamination.
Most likely, this was a result of the decrease in bioavailable (DTPA-extractable and
ammonium acetate-extractable) Pb to 17% to 25% of levels in areas without the
earthworm and arbuscular mycorrhizal fungi amendments. The presence of earthworms
in metal-contaminated soils decreased the amount of water-soluble Pb (Sizmur and
Hodson. 2008). but despite this decrease, ryegrass accumulated more Pb from
earthworm-worked soils than soils without worms present. Sizmur and Hodson
speculated that increased root dry biomass may explain the increased uptake of Pb in the
presence of earthworms. However, Sizmur et al. (2011) found that the presence of anecic
(deep-burrowing) earthworms (L. terrestris) increased soil leachate Pb concentrations by
190%. The authors observed that worms promoted a faster breakdown of organic matter,
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which caused a decrease in soil pH and a concurrent increase in Pb solubility. As a result,
ryegrass (L. perenne) accumulated a greater amount of Pb in systems with earthworms
(Sizmur et al., 2011). Further, the presence of earthworms (Lumbricus terrestris) was
found to increase Pb concentrations in both maize and barley, although growth of these
species was unaffected (Ruiz etal. 2011). Authors noted that worm activity increased Pb
extraction yields by factors of 4.4 and 7.6, for barley and maize. By contrast,
Coeurdassier et al. (2007) found that Pb was higher in earthworm tissue when snails were
present, but that snails did not have a higher Pb content when earthworms were present.
Microbial communities of industrial soils containing Pb concentrations of 61, 456, 849,
1,086, and 1,267 mg Pb/kg dry weight were also improved via revegetation with native
plants, as indicated by increased abundances of fungi, actinomycetes, gram-negative
bacteria, and protozoa, as well as by enhanced fatty acid concentration (Zhang et al..
2006). Increased plant diversity ameliorated the effects of soil Pb contamination (300 and
600 mg Pb/kg) on the soil microbial community (Yang et al.. 2007).
The effect of Pb on microbial community function has been quantified previously using
functional endpoints such as respiration rates, fatty acid production, and soil acid
phosphatase and urease activities, which may provide an estimate of ecological impacts
separate from microbial diversity and abundance measurements. Most studies of metal-
induced changes in microbial communities have been conducted using mixtures of
metals. However, Akerblom et al. (2007) tested the effects of six metals (Cr, Zn, Mo, Ni,
Cd, and Pb) individually. All tested metals had a similar effect on the species
composition of the microbial community. Exposure to a high Pb concentration (52 mg
Pb/kg) also negatively affected respiration rates. Total phospholipid fatty acid content
was determined to negatively correlate with increasing Pb exposure, indicating alteration
of the microbial community. When Yang et al. (2006) compared the microbial properties
of metal-contaminated urban soils to those of rural soils, significant differences were
detected in basal community respiration rates and microbial abundance. The urban soils
studied contained multiple metal contaminants, but microbial biomass was the only
measured endpoint to be significantly and negatively correlated to Pb concentration.
Similarly, the fungal community in a naturally Pb-enriched forest in Norway exhibited
differences in community composition and abundance when compared with other, low Pb
sites. The number of colony-forming fungal units was diminished by soil Pb, and was
approximately 10 times lower in the highest Pb soil concentration (-4300 mg Pb/kg).
Further, only one fungus species was isolated from both high Pb and control soils,
indicating highly divergent communities; species diversity was also reduced by high soil
Pb concentrations (Baath et al.. 2005). These studies suggest that anthropogenic Pb
contamination may affect soil microbial communities, and alter their ecological function.
However, (Khanet al.. 2010c) reported that it is possible for indicators of microbial
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activity to recover after an initial period depression. (Khanet al., 2010c) found that
following a 2-week exposure to three levels of Pb (150, 300, and 500 mg Pb/kg), the
number of culturable bacteria at the highest exposure concentration tested was decreased.
Acid phosphatase and urease levels (measures of soil microbial activity) decreased
significantly, but they had recovered by the ninth week. Another study (Bamborough and
Cummings. 2009) reported that no changes in bacterial and actinobacterial diversity in
metallophytic soils containing 909 to 5,280 mg Pb/kg (43 to 147 mg Pb/kg bioavailable
Pb (as defined by the study authors)). Soil bacteria community structure and basal
respiration rates were examined in natural soils with pH values ranging from 3.7 to 6.8
(Lazzaro et al., 2006). Six soil types of differing pH were treated with Pb nitrate
concentrations of 0.5, 2, 8, and 32 mM (104, 414, 1,658, and 6,630 mg Pb/L). Basal
respiration was decreased in two soil types tested at the highest Pb treatment (32 mM,
=6,630 mg Pb/L), and in a third at the two highest Pb treatments (8 and 32 mM, =1658
and 6,630 mg Pb/L). Terminal Restriction Fragment Length Polymorphism analysis
indicated that bacterial community structure was only slightly altered by Pb treatments.
While pH was correlated with the amount of water-soluble Pb, these increases were
apparently not significant enough to affect bacterial communities, because there were no
consistent relationships between soil pH and respiration rate or microbial community
structure at equivalent soil Pb concentration. Pb contamination was also demonstrated to
reduce phenol oxidase activity in several types of soils; when treated with Pb as Pb
acetate in concentrations between 5 and 50 nM Pb (0.001 and 0.01 mg Pb/L), phenol
oxidase activity significantly decreased in all soils tested, while 400 nM (0.08 mg Pb/L)
and greater completely arrested phenol oxidase activity in one soil tested (a high pH
sandy loam) (Carine et al., 2009). Carine et al. (2009) suggested that the decreased soil
enzymatic activity resulted from changes in the microbial community following Pb
exposure. Pb concentrations between 50 and 500 mg Pb/kg significantly reduced
microbial abundance and diversity, and also resulted in lower soil phosphatase, urease,
and dehydrogenase activities (Gao etal. 2010b). Further, the weekly soil carbon dioxide
evolution rate was significantly reduced by concentrations of 5, 10, and 50 mg Pb/kg,
which also indicated decreased microbial respiration and adverse effects on the microbial
community (Nwachukwu and Pulford. 2011). Gai et al. (2011) examined the microbial
activity of three soils via microcalorimetric methods following Pb exposure. They noted
an increase in activity immediately following Pb application (giving 10, 20, 40, 80, and
160 mg Pb/kg), and theorized that this was a result of rapid mortality of sensitive
microbial species, followed by a concurrent proliferation of Pb-tolerant microorganisms.
As Pb concentrations increased, however, the calculated microbial growth rate constant
decreased, indicating a suppression of microbial activity (Gai et al.. 2011). Authors also
noted a strong correlation between microcalorimetry estimates and the number of colony
forming units isolated from soil samples.
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Pb exposure negatively affected the prey capture ability of certain fungal species.
Nematophagous fungi are important predators of soil-dwelling nematodes, collecting
their prey with sticky nets, branches, and rings. The densities of traps they constructed
decreased in soils treated with 0.15 mM Pb chloride (31 mg Pb/L) (Mo et al.. 2008). This
suppression caused a subsequent reduction in fungal nematode capturing capacity, and
could result in increased nematode abundance.
In a study of microbial communities and enzyme activity, Vaisvalavicius et al. (2006)
observed that high concentration of soil metals were linked to a significant reduction in
soil microorganism abundance and diversity. Soil columns spiked with Cu, Zn, and
Pb acetate (total Pb concentration of 278 to 838 mg Pb/kg, depending on depth) exhibited
a 10- to 100-fold decrease in microbial abundance, with specific microbe classes
(e.g., actinomycetes) seemingly more affected than others (Vaisvalavicius et al.. 2006).
Concurrently, decreases in soil enzymatic activity were also observed, with saccharase
activity decreased by 57-77%, dehydrogenase activity by 95-98%, and urease activity
65-97%. Although this suggests that Pb contamination may alter the nutrient cycling
capacity of affected soil communities, it is difficult to separate the impact of Pb in this
study from the contributions of Cu and Zn that were also added. In contrast, Zeng et al.
(2007) reported that soil concentrations of 300 mg Pb/kg and less stimulated soil
enzymatic activity. Both urease and dehydrogenase levels were increased and rice dry
weight was unaffected by concentrations of 100 and 300 mg Pb/kg. However, at 500 mg
Pb/kg, both rice and soil enzyme activities and microbial biomass were decreased
suggesting impacts at the community level for the soil-rice system. The authors proposed
that these concentrations could be considered the critical Pb concentration in rice paddy
systems (Zeng et al.. 2007).
The microbial communities of soils collected from a Pb-Zn mine and a Pb-Zn smelter
were significantly affected by Pb and other metals (e.g., Cd) (Hu et al., 2007b). At a mine
site, Pb concentration of 57 to 204 mg Pb/kg and Cd concentration of 2.4 to 227 mg
Cd/kg decreased the number of bacteria-forming colonies extracted from soils. Principal
component analysis of microbial community structure demonstrated that different
communities were associated with different metal soil concentration. Similarly, soil
microbial communities exposed to metal contamination from a smelter site (soil Pb
concentration ranging from 30 to 25,583 mg Pb/kg dry weight) showed decreased
bacterial functional diversity (although fungal functional diversity increased) and no
effects on soil respiration rates were observed (Stefanowicz et al., 2008). This led the
authors to conclude that bacterial diversity is a more sensitive endpoint and a better
indicator of metal exposure than fungal diversity or microorganism activity. In a similar
study, Kools et al. (2009) showed that soil ecosystem variables measured after a 6-month
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exposure to metal-contaminated soil indicated that Pb concentration (536 or 745 mg
Pb/kg) was an important driver of soil microbial species biomass and diversity.
Pb-resistant bacterial and fungal communities were extracted regularly from soil samples
at a shooting range site in southern Finland (Hui et al., 2009). While bioavailable Pb
concentration averaged 100 to 200 mg Pb/kg as determined by water extraction, the total
Pb concentrations measured on site were 30,000 to 40,000 mg Pb/kg. To determine Pb
tolerance, bacterial colonies extracted and cultured from shooting range and control soils
were grown on media containing either 0.4 or 1.8 mM Pb (83 or 373 mg Pb/L). While
bacteria isolated from control soil did not proliferate on high-Pb media, shooting-range
soil microbe isolates grew on high-Pb media and were deemed Pb tolerant. The authors
noted that bacterial species common in control samples were not detected among the
Pb-tolerant species isolated from shooting-range soils. They speculated that if long-term
exposure to minimally bioavailable Pb can alter the structure of soil decomposer
communities, decomposition rates could be altered. However, this would require that the
microbial ecosystem decomposing function be altered along with structure, and the
authors provided no evidence for alteration of function.
Microbial communities associated with habitats other than soils are also affected by
exposure to atmospherically deposited Pb. Alder (Alnus nepalensis) leaf microorganism
populations were greater in number at non-affected sites than at sites adjacent to a major
Indian highway with increased Pb pollution (Joshi. 2008). The density, species richness,
and biomass of testate amoebae communities grown on Sphagnum fallax mosses were
significantly decreased following moss incubation in Pb solutions of either 0.6 or 2.5 mg
Pb/L (Nguyen-Viet et al.. 2008). More importantly, species richness and density were
negatively correlated with Pb concentration accumulated within the moss tissue. The
structure of microbial communities associated with lichen surfaces was affected by lichen
trace-element accumulation, including Pb content. Lichens collected from industrial areas
had elevated Pb concentration (10 to 20 mg Pb/kg versus 5 to 7 mg Pb/kg in urban and
rural areas, respectively) and housed bacterial communities characterized by increased
cyanobacteria biomass (Meyer etal.. 2010).
Following a 28-day exposure to field-collected soils contaminated with metals (including
Pb at 426 mg Pb/kg), both population growth and individual growth of the earthworm
L. rubellus were diminished (Klok et al.. 2006). The authors proposed that, although
these reductions were unlikely to result in extirpation, avian predators such as the godwit
(Limosa limosd) that feed heavily on earthworms may be affected by a reduction of
available earthworm biomass.
During the past 5 years, there has been increasing interest in the effects of Pb and other
metals on the functional aspects of soil microbial communities. Most studies show that
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Pb decreases diversity and function of soil microorganisms. However, in an example of
ecological mutualism, plant-associated arbuscular mycorrhizal fungi were found to
protect the host plant from Pb uptake, while fungal viability is protected by the host
plants. Similarly, soil microbial communities (bacterial species as well as fungi) in
Pb-contaminated soils are improved by revegetation. A few studies have reported on
effects of Pb to populations of soil invertebrates. They demonstrated that Pb can decrease
earthworm population density, although not to levels that would result in local extinction.
There have been no recently reported studies on the potential effects of Pb on terrestrial
vertebrate populations or communities, or possible indirect effects through reduction of
prey items such as earthworms.
6.3.7 Critical Loads in Terrestrial Systems
The general concept and definition of critical loads is introduced in Section 6.1.3 of this
chapter [also see Section 7.3 of the 2006 Pb AQCD (U.S. EPA. 2006c)1. An international
workshop was conducted in 2005 on the development of critical loads for metals and
other trace elements (Lofts et al., 2007). Among the findings of the workshop it was
reported that soil transport and transformation processes are key in controlling the fate of
metals and trace elements, thus their importance in the input-output mass balance needs
to be considered. The degree to which these processes are understood and can be
quantified varies. Complexation, sorption, ion exchange and precipitation are well
understood under laboratory conditions, but to a lesser extent in the field (Lofts et al.,
2007). Slower processes of weathering and fixation are less well understood or studied
than leaching (Lofts et al., 2007).
As noted in previous sections, soil pH and organic matter influence Pb availability.
De Vries et al. (2007) demonstrate that critical limits, measured as critical reactive metal
content, can significantly vary between soil types that differ in pH and organic matter.
Critical limits of Pb increased from 30 to 64 (mg Pb/kg) over a pH range of 4-7 when soil
organic matter content was 5%, while these limits increased from 187 to 400 (mg Pb/kg)
over the same pH range when organic content was 80%. These implications suggest that
critical limits increase with increasing soil organic matter. This has important
consequences for forest soils because many are covered by an organic layer where roots,
fungi and other microorganisms are located. Baath (1989) evaluated the effects of organic
matter on critical limits for microorganisms, measured via enzyme synthesis, litter
decomposition and soil respiration. Results indicate critical limits are up to four times
higher in the organic (135 to 976 mg Pb/kg) than the mineral soil layer (32 to 690 mg
Pb/kg) at hazardous concentration ranging from 5-50% of species. In general, De Vries
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et al. (2007) found support that ecotoxicological critical limits in European soils for Pb
decrease with increasing pH.
Several methods are routinely used for Pb risk assessment of terrestrial animals. Buekers
et al. (2009) proposed the use of a Tissue Residue Approach as a risk estimation method
for terrestrial vertebrates that eliminates the need for quantitative estimation of food
intake or Pb species bioavailability. Blood Pb no observed effect concentration (NOEC)
and lowest observed effects concentration (LOEC) data derived from 25 studies
examining the effects of Pb exposure on growth, reproduction, and hematological
endpoints were used to construct a series of species sensitivity distributions for mammals
and birds. They also used the HC5 criterion (5th percentile of species NOEC values for
collection of species) proposed by Aldenberg and Slob (1993). For mammals, the HC5
values obtained ranged from 11 to 18 (ig Pb/dL blood; HC5 values for birds ranged from
65 to 71 (ig Pb/dL. The authors proposed the use of 18 and 71 (ig Pb/dL as critical
threshold values for mammals and birds respectively, which are below the lowest NOEC
for both data sets used, and are above typical background Pb values. It is difficult to
determine environmental Pb toxicity given the variation of physicochemical and soil
properties that alter bioavailability and toxicity. This variability makes it difficult to
extrapolate between areas. Furman et al. (2006) proposed the use of a physiologically
based extraction test to predict risks posed to waterfowl from environmental Pb
contamination. The extraction process was modeled after gastric and intestinal conditions
of waterfowl, and was used to gauge the bioavailability of Pb from freshly amended and
aged contaminated soils. The concentration of Pb extracted through the use of the
physiologically based extraction test was demonstrated to be significantly correlated to
Pb tissue concentration in waterfowl exposed via in vivo studies of the same soils.
No critical loads for Pb from terrestrial ecosystems in the U.S. were identified for review
in this ISA; however, work has been conducted in Europe. Given that local conditions
(including historic loading, soil transport and transformation processes) are key elements
to critical load calculation the utility of critical loads that are developed from other
countries for application to U.S. ecosystems is unclear. The most recent European
publications on Pb critical loads include assessments of the U.K., Netherlands and Italy.
Hall et al. (2006) used the critical load approach to conduct a national risk assessment of
atmospheric Pb deposition for the U.K. While specific regions were determined to have
low critical load values for Pb (central England [Peak district area], the Pennines [in
north-central England], in southern Wales, and scattered areas in other parts of England,
Wales, and Scotland), the authors noted that this approach can be significantly biased, as
available ecotoxicological data used in the modeling were from studies that were not
conducted in soils representative of all U.K. soils. De Vries and Groenenberg (2009)
similarly observed that the uncertainty inherent in a critical load approach to Pb risk
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assessment is influenced by the critical concentration of dissolved metal and the
absorption coefficients of exposed soils. However, this approach did indicate that for
forest soils in the Netherlands, 29% of the areas would be expected to exceed the critical
load, based on currently available toxicity data and Pb pollution data (de Vries and
Groenenberg. 2009). Similarly, although Pb soil concentrations in the Bologna Province
of Italy were far below concentrations harmful to soil organisms, current atmospheric Pb
deposition rates suggest that critical load exceedances are likely in the future, unless
annual Pb emissions are decreased (Morselli et al.. 2006).
Given the heterogeneity of ecosystems exposed to Pb, and the differences in expectations
for ecosystem services attached to different land uses, it is expected that critical load
values for Pb for soils within the U.S. would fall within a broad range. Refer to Section
7.3.5 of the 2006 Pb AQCD for additional discussion of critical loads of Pb in terrestrial
systems.
6.3.8 Soil Screening Levels
Developed by EPA, ecological soil screening levels (Eco-SSLs) are maximum
contaminant concentrations in soils that are predicted to result in little or no quantifiable
effect on terrestrial receptors. These conservative values were developed so that
contaminants that could potentially present an unacceptable hazard to terrestrial
ecological receptors are reviewed during the risk evaluation process while removing from
consideration those that are highly unlikely to cause significant effects. The studies
considered for the Eco-SSLs for Pb and detailed consideration of the criteria for
developing the Eco-SSLs are provided in the 2006 Pb AQCD (U.S. EPA. 2006c).
Preference is given to studies using the most bioavailable form of Pb, to derive
conservative values. Soil concentration protective of avian and mammalian diets are
calculated by first converting dietary concentration to dose (mg/kg body weight per day)
for the critical study, then using food (and soil) ingestion rates and conservatively derived
uptake factors to calculate soil concentration that would result in unacceptable dietary
doses. This frequently results in Eco-SSL values below the average background soil
concentration [19 mg Pb/kg dry weight (U.S. EPA. 2005b. 2003b)], as is the case with Pb
for birds. The Pb Eco-SSL was completed in March 2005 and has not been updated since.
Values for terrestrial birds, mammals, plants, and soil invertebrates are 11, 56, 120, and
1,700 mg Pb/kg soil (dry weight), respectively.
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6.3.9 Characterization of Sensitivity and Vulnerability
Research has long demonstrated that Pb affects survival, reproduction, growth,
metabolism, and development in a wide range of species. The varying severity of these
effects depends in part upon species differences in metabolism, sequestration, and
elimination rates. Dietary factors also influence species sensitivity to Pb. Because of
effects of soil aging and other bioavailability factors discussed above (Section 6.3.2). in
combination with differing species assemblages and biological accessibility within prey
items, ecosystems may also differ in their sensitivity and vulnerability to Pb. The
2006 Pb AQCD reviewed many of these factors which are updated herein by reference to
recent literature.
6.3.9.1 Species Sensitivity
There is wide variation in sensitivity of terrestrial species to Pb exposure, even among
closely related organisms. Langdon et al. (2005) showed a two-fold difference in LC50
values among three common earthworm species, with the standard laboratory species,
E. andrei, being the least sensitive. Mammalian NOEC values expressed as blood Pb
levels were shown to vary by a factor of 8, while avian blood NOECs varied by a factor
of 50 (Buekers et al., 2009). Age at exposure, in particular, may affect sensitivity to Pb.
For instance, earlier instar C. elegans were more likely than older individuals to exhibit
neurobehavioral toxicity following Pb exposure (Xing et al.. 2009b). and also
demonstrated more pronounced neural degeneration than older larvae and adults (Xing et
al.. 2009c).
6.3.9.2 Nutritional Factors
Dietary factors can exert significant influence on the uptake and toxicity of Pb in many
species of birds and mammals. The 2006 Pb AQCD (U.S. EPA. 2006b) describes how
Ca2+, Zn, Fe, vitamin E, Cu, thiamin, P, Mg, fat, protein, minerals, and ascorbic acid
dietary deficiencies increase Pb absorption and its toxicity. For example, vitamin E
content was demonstrated to protect against Pb-induced lipid peroxidation in mallard
ducks. Generally, Pb exposure is more likely to produce behavioral effects in conjunction
with a nutrient-deficient diet. As previously reported in the 2006 Pb AQCD, Ca2+
deficiencies may increase the susceptibility of different terrestrial species to Pb, including
plant (Antosiewicz. 2005). avian (Dauwe et al.. 2006; Snoeijs et al.. 2005) and
invertebrate species. Antosiewicz (2005) determined that, for plants, Ca2+ deficiency
decreased the sequestration capacity of several species (tomato, mustard, rye, and maize),
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and that this likely resulted in an increased proportion of Pb at sites of toxic action.
Because Pb ions can interact with plant Ca2+ channel pores, in the presence of low Ca2+
and high Pb concentration, a higher proportion of Pb can interact with these channels and
be taken up by plants. A similar phenomenon has been observed in invertebrates, where
the metabolic pathway of metals mimics the metabolic pathway of Ca2+ [Simkiss et al.
(1982). as cited in Jordaens et al. (2006)1. Hence, in environments with
disproportionately high Pb versus Ca2+ concentration, accumulation of Pb may be
accelerated, as in plants. Ca2+ deficiency in birds was demonstrated to stimulate the
production of Ca2+-binding proteins in the intestinal tract, which extract more Ca2+ from
available diet; however, this response also enhances the uptake and accumulation of Pb
from diet and drinking water [Fullmer (1997). as cited in Dauwe et al. (2006)].
6.3.9.3 Soil Aging and Site-Specific Bioavailability
Total soil Pb concentration is a poor predictor of hazards to avian or mammalian wildlife,
because site-specific biogeochemical and physical properties (e.g., pH, OM, metal oxide
concentration) can affect the sequestration capacity of soils. Additionally, soil aging
processes have been demonstrated to decrease the bioavailable Pb fraction; as such,
laboratory toxicity data derived from spiked soils often overestimate the environmental
risk of Pb. Smolders et al. (2009) compared the toxicity of freshly Pb-spiked soils to
experimentally aged spiked soils and field-collected Pb-contaminated soils. Experimental
leaching and aging was demonstrated to increase invertebrate Pb EC50 values by factors
of 0.4 to greater than 8. In approximately half the cases, experimental aging of freshly
spiked soils through leaching removed the proportional relationship between Pb content
and toxicity. The leaching/aging factor for Pb, or ratio of EDi0 values derived in
experimentally aged soils to freshly spiked soils, was determined to be 4.2 (factors
greater than one indicate decreased toxicity in aged field soils relative to laboratory
spiked soils). These results draw attention to the difficulty of extrapolating results from
experiments with spiked soil to natural soils, but nonetheless suggest that the sensitivity
of terrestrial invertebrates to environmental Pb exposures is likely heavily dependent on
the relative rate of aging and site-specific bioavailability.
6.3.9.4 Ecosystem Vulnerability
Relative vulnerability of different terrestrial ecosystems to effects of Pb can be inferred
from the information discussed above on species sensitivity and how soil geochemistry
influences the bioavailability and toxicity of Pb. Soil ecosystems with low pH,
particularly those with sandy soils, are likely to be the most sensitive to the effects of Pb.
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Examples of such systems are forest soils, including oak, beech, and conifer forests.
The Pine Barrens in southern New Jersey (also known as the Pinelands) is an example of
a highly vulnerable ecosystem: it is a dense coniferous (pine) forest with acidic, sandy,
nutrient poor soil. As agricultural areas are taken out of production and revert to old
fields and eventually forests, their vulnerability to Pb is likely to increase as a result of
decreasing OM and acidification of soils (from discontinuation of fertilizing and liming).
On the other hand, increasing density of native or invasive plants with associated
arbuscular mycorrhizal fungi will likely act to ameliorate some of the effects of Pb (see
previous discussion of studies by Sudova and Vostka (2007) and Wong et al. (2007). It is,
however, difficult to categorically state that certain plant or soil invertebrate communities
are more vulnerable to Pb than others, as the available toxicity data have not yet been
standardized for differences in bioavailability (because of use of different Pb salts,
different soil properties, and different lengths of aging of soil prior to testing), nutritional
state, or organism age, or other interacting factors. Data from field studies are
complicated by the co-occurrence of other metals and alterations of pH, such as
acidification from SO2 in smelter emissions, which are almost universal at sites of high
Pb exposure, especially at mine or smelter sites. However, because plants primarily
sequester Pb in the roots, uptake by soil invertebrates is the most likely pathway for Pb
exposure of higher trophic level organisms. Invertivores are likely at higher risk than
herbivores. In fact, estimations of Pb risk at a former Pb smelter in northern France
indicated that area Pb concentration presented the greatest threat to insectivorous bird and
mammal species, but only minimal risk to soil invertebrate and herbivorous mammals
(Fritsch et al.. 2010). By extension, birds and mammals in ecosystems with a richer
biodiversity of soil invertebrates may be more vulnerable to Pb than those in ecosystems
with fewer invertebrates (e.g., arid locations). Regardless, the primary determinant of
terrestrial ecosystem vulnerability is soil geochemistry, notably pH, CEC, and amount of
OM.
6.3.10 Ecosystem Services Associated with Terrestrial Systems
Pb deposited on the surface of, or taken up by organisms has the potential to alter the
services provided by terrestrial biota to humans. There are no publications at this time
that specifically focus on the ecosystem services affected by Pb in terrestrial systems and
the directionality of impacts is not always clear. For example, terrestrial soils provide a
service to aquatic ecosystems by sequestering Pb through sorption and precipitation. At
the same time, the sequestration of Pb by soils may result in a degredation in the quality
of soil and may result in decreased crop productivity. The evidence reviewed in the
present document illustrates that Pb can cause ecological effects in each of the four main
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categories of ecosystem services (Section 6.1.2) as defined by Hassan et al. (2005). These
effects are sorted into ecosystem services categories and summarized here:
• Supporting: altered nutrient cycling, decreased biodiversity, decline of
productivity, food production for higher trophic levels
• Provisioning: plant yields
• Regulating: decline in soil quality, detritus production
• Cultural: ecotourism and cultural heritage values related to ecosystem integrity
and biodiversity, impacts to terrestrial vertebrates.
A few studies since the 2006 Pb AQCD, consider the impact of metals in general on
ecosystem services. Honeybees are important for provisioning services such as
pollination and production of honey. They can be exposed to atmospheric Pb by direct
deposition or through Pb associated with plants, water or soil. In a study of heavy metals
in honeybees in central Italy, there was a statistically significant difference in Pb between
bees collected in wildlife reserves compared to bees collected in urban areas with the
highest concentration of Pb detected from bees caught in hives near an airport (Perugini
et al., 2011). In a review of the effects of metals on insect behavior, ecosystem services
provided by insects such as detritus reduction and food production for higher trophic
levels were evaluated by considering changes in ingestion behavior and taxis (Mogren
and Trumble. 2010). Pb was shown in a limited number of studies to affect ingestion by
insects. Crickets (Chorthippus spp) in heavily contaminated sites reduced their
consumption of leaves in the presence of increasing Cd and Pb concentrations (Migula
and Binkowska. 1993). Decreased feeding activity in larval and adult Colorado potato
beetle (Leptinotarsa decemlineata) were observed as a result of dietary exposures of Pb
and Cu (Kwartirnikov et al.. 1999). while no effects were found in ingestion studies of Pb
with willow leaf beetle, Lochmaea caprae (Rokytova et al.. 2004) mottled water hyacinth
weevil, Neochetina eichhorniae (Kay and Haller. 1986) and hairy springtail, Orchesella
cincta (van Capelleveen et al.. 1986J.
Soil health for agricultural production and other soil-associated ecosystem services is
dependent upon the maintenance of four major functions: carbon transformations,
nutrient cycles, soil structure maintenance, and the regulation of diseases and pests and
these parameters may be altered by metal deposition (Kibblewhite et al., 2008). Pb
impacts to terrestrial systems reviewed in the previous sections provide evidence for
impacts to supporting, provisioning, and regulating ecosystem services provided by soils.
For example, earthworms were shown to impact soil metal mobility and availability,
which in turn resulted in changes to microbial populations (biodiversity), pH, dissolved
organic carbon, and metal speciation (Sizmur and Hodson. 2009). all of which may
directly affect soil fertility.
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Pb is bioaccumulated in plants, invertebrates and vertebrates inhabiting terrestrial and
aquatic systems that receive Pb from atmospheric deposition. This represents a potential
route for Pb mobilization into the food web or into food products. For example, Pb
bioaccumulation in leaves and roots of an edible plant may represent an adverse impact to
the provisioning of food, an essential ecosystem service. Although there is no consistent
evidence of trophic magnification there is substantial evidence of trophic transfer. It is
through consumption of Pb-exposed prey or Pb-contaminated food that atmospherically
deposited Pb reaches species that may have very little direct exposure to it.
There is limited evidence of Pb impacts to plant productivity. Productivity of gray birch
(Betula populifolia) was impaired in soils with elevated As, Cr, Pb, Zn and V (Gallagher
et al., 2008). Tree growth measured in both individuals and at the assemblage level using
satellite imagery and field spectrometry was significantly decreased with increasing metal
load in soil.
6.3.11 Synthesis of New Evidence for Pb Effects in Terrestrial Systems
This synthesis of the effects of Pb on terrestrial ecosystems covers information from the
publication of the 2006 Pb AQCD to present. It is followed in Section 6.5 by
determinations of causality that take into account evidence dating back to the 1977 Pb
AQCD.
High concentrations of Pb
The state-level mean concentration of Pb in U.S. soils ranges from 5 to 39 mg Pb/kg.
Studies of the effects of Pb always include much higher concentrations, whether they use
soils that have been exposed to Pb pollution, or experimental amendment with salts of Pb
(Table 6-4). Studies that use soils that have been exposed to Pb pollution were conducted
in situ, or with soil collected from natural environments in highly contaminated areas
near stationary sources. All of them took advantage of gradients of exposure produced by
distance from the source to vary the levels of Pb. In some cases, there were only two
levels-control and elevated-created by using two sampling locations rather than a larger
set of levels representative of the entire gradient. In most studies that used soils exposed
to Pb pollution, the highest concentration of Pb in the study is very high relative to those
found anywhere except at heavily exposed sites.
In amendment experiments, variation in Pb was generated by addition of Pb salts to either
natural or artificial soils. These experiments often included concentrations that were even
higher than those found in heavily exposed natural environments. In either type of study,
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however, effects gradually increased with increasing exposure. This exposure-response
dependency is an important component of causality determinations (Preamble). In
addition, until the presence of a discontinuity or breakpoint in the exposure-response
relationship can be shown, evidence of a monotonic rise in response suggest that effects
are present at all concentrations, even though the magnitude of those effects may not be
known in the lower part of the range.
Using concentration-response models where concentration is taken as a continuous
variable to analyze data with multiple values of Pb concentration would allow better
estimation of the size of effects at any value of exposure, including low ones, and also
better estimation of uncertainties around the size of effects. However, none of studies
with multiple Pb concentrations used a continuous model to characterize the relationship
between concentration and effects. In the analyses of data where multiple levels of Pb
were recorded, published studies only used discrete models of exposure, where increasing
concentrations of Pb are treated as separate, categorical values.
Comparability of effect concentrations
Strong interactions of Pb concentration and several other soil variables, including pH,
CEC, OC, and Fe/Al oxides have been amply demonstrated with respect to various
biological responses. For example, Dayton et al.(2006) and Bradham et al.(2006) tested
an array of different soils to which the same amount of Pb was added, using lettuce and
earthworms, respectively. They found differences in biological effects that were as large
as 27, 35, or even 72-fold between soils.
In studies where Pb was introduced through amendment, those interacting variables can
be changed experimentally in a controlled way, or held constant. In studies where natural
soils were used in which Pb originated from pollution, they are left to vary freely. In
either case, the presence and magnitude of those interactions make calculations of
expected responses under other sets of conditions particularly difficult, as well as
comparisons between studies conducted under different conditions.
In addition, the amount of Pb dissolved in soil pore water determines the impact of soil
Pb on terrestrial ecosystems to a much greater extent than the total amount present. It has
long been established that the amount of Pb dissolved in soil solution is controlled by at
least six variables: (1) solubility equilibria; (2) adsorption-desorption relationship of total
Pb with inorganic compounds; (3) adsorption-desorption reactions of dissolved Pb phases
on soil OM; (4) pH; (5) CEC; and (6) aging. Since 2006, further details have been
contributed to the understanding of the role of pH, CEC, OM, and aging. Smolders et al.
(2009) demonstrated that the two most important determinants of solubility (and also
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toxicity) in soils are pH and CEC. However, they had previously shown that aging,
primarily in the form of initial leaching following deposition, decreases soluble metal
fraction by approximately one order of magnitude (Smolders et al., 2007). Since 2006,
OM has been confirmed as an important influence on Pb sequestration, leading to longer-
term retention in soils with higher OM content, and also creating the potential for later
release of deposited Pb. Aging, both under natural conditions and simulated through
leaching, was shown to substantially decrease bioavailability to plants, microbes, and
vertebrates. However, most studies report some measure of total extracted Pb, or total
added Pb, rather than pore water or soluble fraction.
Plants
Recent studies with herbaceous species growing at various distances from smelters added
to the existing strong evidence that atmospherically transported Pb is taken up by plants.
These studies did not establish the relative proportion that originated from atmospheric
Pb deposited in the soil, as opposed to that taken up directly from the atmosphere through
the leaves. Studies found that in trees, Pb that is taken up through the roots is then
generally translocated from the roots to other parts. However, multiple recent studies
showed that in trees, the proportion of Pb that is taken up through the leaves is likely to
be very substantial. One study attempted to quantify it, and suggested that 50% of the Pb
contained in Scots Pine in Sweden is taken up directly from the atmosphere
(Section 6.3.3.1). Studies with herbaceous plants found that in most species tested, soil
Pb taken up by the roots is not translocated into the stem and leaves, but when growth
and survival were reported, growth of the whole plant decreased with increasing Pb, and
mortality increased (Table 6-4). Experimental studies have added to the existing evidence
of photosynthesis impairment in plants exposed to Pb, and have found damage to
photosystem II due to alteration of chlorophyll structure, as well as decreases in
chlorophyll content in diverse taxa, including lichens and mosses. A substantial amount
of evidence of oxidative stress in response to Pb exposure has also been produced.
Reactive oxygen species were found to increase in broad bean and tomato plants exposed
to increasing concentrations of soil Pb, and a concomitant increase in superoxide
dismutase, glutathione, peroxidases, and lipid peroxidation, as well as decreases in
catalase were observed in the same plants. Monocot, dicot, and bryophytic taxa grown in
Pb-contaminated soil or in experimentally spiked soil all responded to increasing
exposure with increased antioxidant activity. In addition, genotoxicity, decreased
germination, and pollen sterility were observed in some experiments. All effects were
small outside of contaminated areas (Section 6.3.4.1).
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Invertebrates
Since the 2006 Pb AQCD, various species of terrestrial snails have been found to
accumulate Pb from both diet and soil, although effects on growth, survival and
reproduction are inconsistent. Recent studies with earthworms have found that both
internal concentration of Pb and mortality increase with decreasing soil pH and CEC. In
addition, tissue concentration differences have been found in species of earthworms that
burrow in different soil layers. The rate of accumulation in each of these species could
result from layer differences in interacting factors such as pH and CEC (Section 6.3.3.2).
Because earthworms often sequester Pb in granules, some authors have suggested that
earthworm Pb is not bioavailable to their predators. There is some evidence that
earthworm activity increases Pb availability in soil, but it is inconsistent. In arthropods
collected at contaminated sites, recent studies found gradients in accumulated Pb that
corresponded to gradients in soil with increasing distance from stationary sources.
Recently published studies have shown neuronal damage in nematodes exposed to low
concentrations of Pb (2.5 uM = 0.5 mg Pb/L), accompanied by behavioral abnormalities.
Reproductive effects were found at lower exposure in younger nematodes, and effects on
longevity and fecundity were shown to persist for several generations. Increased
mortality was found in earthworms, and was strongly dependent on soil characteristics
including pH, CEC, and aging. Snails exposed to Pb through either topical application or
through consumption of Pb-exposed plants had increased antioxidant activity and
decreased food consumption, but effects on growth and survival were inconsistent.
Effects on arthropods exposed through soil or diet varied with species and exposure
conditions, and included diminished growth and fecundity in springtails, endocrine and
reproductive anomalies, and body deformities. Increasing concentration of Pb in the
exposure medium generally resulted in increased effects within each study, but the
relationship between concentration and effects varied between studies, even when the
same medium, e.g., soil, was used (Section 6.3.4.2). Evidence suggested that aging and
pH are important modifiers.
Vertebrates
There were few recent studies of Pb bioavailability and uptake in vertebrates since the
2006 Pb AQCD. A study of two species of sea ducks in Alaska found that 3% of the birds
had tissue levels of Pb that indicated exposure above background. Urban pigeons in
Korea were found to accumulate 1.6 to 1.9 mg Pb/kg wet weight Pb in the lungs, while in
Wisconsin 70% of American woodcock chicks and 43% of young-of-year had elevated
bone Pb (9.6 to 93 mg Pb/kg dry weight in chicks, 1.5 to 220 mg Pb/kg dry weight in
young-of-year). None of the locations for these studies was in proximity to stationary
sources of heavy contamination, and none was able to identify the origin of the Pb
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(Section 6.3.3.3). Effects on amphibians and reptiles included decreased white blood cell
counts, decreased testis weight, and behavioral anomalies. However, large differences in
effects were observed at the same concentration of Pb in soil, depending on whether the
soil was freshly amended or field-collected from contaminated areas. As in most studies
where the comparison was made, effects were smaller when field-collected soils were
used. A study at the Anaconda Smelter Superfund site found increasing Pb accumulation
in gophers with increasing soil Pb around the location of capture. Effects of dietary
exposure were studied in several mammalian species, and cognitive, endocrine,
immunological, and growth effects were observed. Pigs fed various Pb-contaminated
soils showed that the form of Pb determined accumulation, and another study showed
lower feed efficiency and weight in pigs with 2.08 versus 1.44 ug Pb/dL in blood,
originating in Pb-sulfate feed supplement. In some birds, maternal elevated blood Pb
level was associated in recent studies with decreased hatching success, smaller clutch
size, high corticosteroid level, and abnormal behavior. Some species show little or no
effect of elevated blood Pb level. A study of Japanese quail found that Pb added to the
diet could improve survival and incidence of several pathologies, and a long term study
of pied flycatchers at a mine site produced mixed evidence for the effects of Pb
(Section 6.3.4.3).
Food web
Recent studies were able to measure Pb in the components of various food chains that
included soil, plants, invertebrates, arthropods and vertebrates. They confirmed that
trophic transfer of Pb is pervasive, but no consistent evidence of trophic magnification
was found (Section 6.3.3.4).
Community and Ecosystem Effects
New evidence of effects of Pb at the community and ecosystem levels of biological
organization include several studies of the ameliorative effects of mycorrhizal fungi on
plant growth, attributed to decreased uptake of Pb by plants, although both mycorrhizal
fungus and plant were negatively affected. The presence of both earthworms and
mycorrhizal fungi decreased solubility and mobility of Pb in soil in one study, but the
presence of earthworms was associated with higher uptake of Pb by plants in another.
The presence of snails increased uptake of Pb by earthworms, but not vice-versa. Most
recently published research on community and ecosystem effects of Pb has focused on
soil microbial communities, which have been shown to be impacted in both composition
and activity. Many recent studies have been conducted using mixtures of metals, but have
tried to separate the effects of individual metals when possible. One study compared the
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effects of 6 metals individually (Akerblom et al.. 2007). and found that their effects on
community composition were similar. In studies that included only Pb, or where effects
of Pb could be separated, soil microbial activity was generally diminished, but in some
cases recovered overtime. Species and genotype composition were consistently altered,
and those changes were long-lasting or permanent (Section 6.3.6).
Exposure-Response
Several studies with various organisms have included gradients of Pb exposure. None has
characterized the exposure-relationship using a continuous model of exposure-response.
However, evidence indicates clearly that increased exposure to Pb is associated with
increases in observed effects in terrestrial ecosystems. Evidence also demonstrates that
many factors, including species and various soil physiochemical properties, interact
strongly with Pb concentration to modify those effects. In terrestrial ecosystems, where
soil is generally the main component of the exposure route, Pb aging is a particularly
important factor, and one that may be difficult to reproduce experimentally. Without
adequate quantification of those interactions, characterizations of exposure-response
relationships may be difficult to transfer outside of experimental settings.
6.3.12 Causal Determinations for Pb in Terrestrial Systems
In the following sections, organism-level effects on reproduction and development,
growth and survival are considered first since these endpoints can lead to effects at the
population level or above and are important in ecological risk assessment.
Neurobehavioral effects are considered next followed by sub-organismal responses
(hematological effects, physiological stress) for which Pb has been shown to have an
impact in multiple species and across taxa, including humans. Causal determinations for
terrestrial, freshwater and saltwater ecological effects are summarized in Table 6-3.
6.3.12.1 Reproductive and Developmental Effects-Terrestrial Biota
In terrestrial invertebrates and vertebrates, evidence assessed for the present document
and in Pb AQCDs indicates an association between reproductive effects and Pb exposure.
Impaired fecundity at the organism level of biological organization can result in a decline
in abundance and/or extirpation of populations, decreased taxa richness, and decreased
relative or absolute abundance at the community level (Suter et al.. 2005; U.S. EPA.
2003a). Evaluation of the literature on Pb effects in terrestrial species indicates that
exposure to Pb is associated with reproductive effects. Various endpoints have been
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measured in various taxa of terrestrial organisms to assess the effect of Pb on fecundity,
development, and hormone homeostasis. Although reproductive effects were
demonstrated, no single endpoint in a single taxon has been extensively studied. Recent
evidence available since the 2006 Pb AQCD for effects of Pb on reproductive endpoints
in terrestrial species is summarized in Table 6-4.
In terrestrial plants, few studies were available to the 2006 Pb AQCD (U.S. EPA. 2006b).
and few are available more recently that specifically address reproductive effects of Pb
exposure. Two genotypes of maize seedlings exhibited a significant and concentration-
dependent reduction in seed germination following 7 days of Pb treatment in nutrient
solution of nominal concentration of 0, 0.007, 0.7 and 7 mg Pb/L as Pb sulfate (Ahmad et
al., 2011). Germination inhibition and chromosomal abnormalities also increased in a
concentration-dependent manner in Grass pea grown in soil irrigated with solutions
containing nominal concentration of 0 to 188 mg Pb/L (Kumar and Tripathi. 2008).
However, germination increased in a broad sample of soils when amended with 2,000 mg
Pb/kg (Dayton et al.. 2006).
In terrestrial invertebrates, Pb can alter developmental timing, hatching success, sperm
morphology and hormone homeostasis. The number of species studied has been small,
but reproductive effects consistently increase with increasing exposure. The
2006 Pb AQCD reported effects on reproduction in collembolans and earthworms, with
LOECs and NOECs typically well above Pb soil concentrations observed away from
stationary sources of contamination, more recently, an increase in development time
(approximately two days) and a reduction in relative fecundity were observed in aphids
feeding on plants exposed to high concentrations of Pb (Goriir. 2007). Hatching success
of the collembolan F. Candida was decreased following 10 day exposure to Pb-spiked
soils (EC50 2,361 mg Pb/kg dry soils) (Xu et al.. 2009b). Sperm morphology in Asian
earthworms was significantly altered following 2-week exposures to soils containing
nominal concentration of 1,000, 1,400, 1,800, and 2,500 mg Pb/kg soil (Zheng and Li.
2009). Pb may also disrupt hormonal homeostasis in invertebrates as studies with moths
have suggested (Shu et al.. 2009). Adult female moths reared on diets containing 25, 50,
100, or 200 mg Pb/kg exhibited decreased vitellogenin mRNA induction, and
vitellogenin levels were demonstrated to decrease with increasing Pb exposure. Evidence
of multi-generational toxicity effects of Pb is also present in terrestrial invertebrates,
specifically springtails, mosquitoes, carabid beetles, and nematodes where decreased
fecundity in progeny of Pb-exposed individuals was observed. The magnitude of effects
is variable, but they are present in multiple phyla, and increase with increasing exposure
within studies. Reproductive effects in terrestrial invertebrates are also coherent with
similar effects observed in aquatic invertebrates.
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In terrestrial vertebrates, there is evidence for reproductive effects associated with Pb
exposure in recent evidence and Pb AQCDs. The 2006 Pb AQCD (U.S. EPA. 2006c)
concluded that exposure to affects reproduction and development in terrestrial
vertebrates. Effects reported in that document included declines in clutch size, number of
young hatched, number of young fledged, decreased fertility, and decreased eggshell
thickness observed in birds near areas of Pb contamination and in birds with elevated Pb
tissue concentration regardless of location. More recently, decreased testis weight was
observed in lizards administered a sublethal dose of 10 or 20 mg Pb/kg day by oral
gavage for 60 days (Salice et al. 2009). Few studies in the field have addressed
reproductive effects of Pb specifically in mammals, due to most available data in wild or
grazing animals being from near smelters, where animals are co-exposed to other metals.
For example, the reproductive viability of red deer (C. elaphus) inhabiting a
Pb-contaminated mining area of Spain was shown to be altered, with 11% and 15%
reductions in spermatozoa and acrosome integrity observed in male deer from the mining
area compared with those residing in reference areas (Reglero et al.. 2009aX but multiple
other metals were present at high concentrations. Evidence from AQCDs and the present
document for terrestrial vertebrates is coherent with evidence from freshwater
amphibians, and fish (Section 6.4.12.1). However, experimental evidence obtained using
mammals in the context of human health research demonstrates a consistency of adverse
effects of Pb on sperm (Section 4.8.3.1) and the onset of puberty in males and females
(Sections 4.8.1.1 [females] and 4.8.1.2 [males]) with strong evidence from both
toxicology and epidemiology studies. Other reproductive endpoints including
spontaneous abortions, pre-term birth, embryo development, placental development, low
birth weight, subfecundity, hormonal changes, and teratology were also affected, but less
consistently (Section 4.8).
For reproductive and developmental effects in terrestrial ecosystems, the current body of
evidence is inadequate to conclude that exposure to Pb is causal in plants, and is
sufficient to conclude that there is a causal relationship in invertebrates and vertebrates.
6.3.12.2 Growth Effects-Terrestrial Biota
Alterations in growth at the organism level of biological organization can have impacts at
the population, community, and ecosystem levels. In terrestrial ecosystems, evidence for
effects of Pb on growth is strongest in terrestrial plants, although these effects are
typically observed in laboratory studies with high exposure concentrations or in field
studies near stationary sources. In terrestrial plants, there is evidence over several decades
of research that Pb inhibits photosynthesis and respiration, all of which can reduce the
growth of the plant (U.S. EPA. 2006c. 1986a. 1977). Decreases in chlorophyll a and b
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content have been observed in various algal and plant species. Many laboratory toxicity
studies report effects on the growth of plants in synthetic growing media, but observed
effects typically occur at concentrations higher than the average background
concentrations in U.S. soils (19 mg Pb/kg dry weight) (U.S. EPA. 2005b) and there are
few field studies. Effects on plant growth can result in reduced productivity and
decreased biomass. The 2006 Pb AQCD relied principally on evidence assembled in the
Ecological Soil Screening Levels for Lead document (U.S. EPA. 2005b), which
concluded that growth (biomass) was the most sensitive and ecologically relevant
endpoint for plants.
Evidence for growth effects in terrestrial fauna is sparse. In the 1986 Pb AQCD, a study
was reviewed in which the Fl and F2 generations of the springtail Onychiurus armatus
fed a diet of Pb-exposed fungi (0.008 to 3.1 mg Pb/g) experienced a delay in achieving
maximum length (Bengtsson et al., 1983). The authors suggested that the reduced growth
may be accompanied by delayed sexual maturity. The 2006 Pb AQCD (U.S. EPA.
2006b) reported that growth effects observed in both terrestrial invertebrates and
vertebrates were more pronounced in juvenile organisms, underscoring the importance of
lifestage to overall Pb susceptibility. Recent evidence available since the 2006 Pb AQCD
for effects of Pb on growth endpoints in terrestrial species is summarized in Table 6-4:
reduced growth of garden snail T. pisana, increasing with increasing exposure, was
observed following a five week dietary exposure to eight nominal concentrations of Pb
(El-Gendy et al., 2011). Studies also show concentration-dependent inhibition of growth
in earthworms raised in Pb-amended soil (Zheng and Li. 2009; Currie et al.. 2005;
Langdon et al., 2005). In AQCDs, growth effects of Pb have been reported in birds
(changes in juvenile weight gain), at concentrations typically higher than currently found
in the environment away from heavily exposed sites. The current body of evidence is
sufficient to conclude that there is a causal relationship between Pb exposures and growth
effects in terrestrial plants, and that a causal relationship is likely to exist between Pb
exposure and growth effects in terrestrial invertebrates. Evidence is inadequate to
establish causal relationship between Pb exposures and growth effects in terrestrial
vertebrates.
6.3.12.3 Survival-Terrestrial Biota
The relationship between Pb exposure and survival has been well demonstrated in
terrestrial species as presented in the Pb AQCDs and in Section 6.3.5 of the present
document. Exposure can be either lethal, or produce sublethal effects that diminish
survival probabilities. In the 1977 Pb AQCD, deaths from Pb poisoning in domestic
animals caused by emissions from stationary sources were reported (U.S. EPA. 1977).
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Additional studies in the 1986 and 2006 Pb AQCDs and current ISA provide evidence for
a concentration-dependent response of mortality in terrestrial biota. Recent evidence
available since the 2006 Pb AQCD for effects of Pb on survival in terrestrial species is
summarized in Table 6-4.
Survival is a biologically important response that can have direct impact on population
size. Survival is often quantified using LC50 (the concentration of toxicant where 50%
mortality is observed or modeled), which may be a better metric for acute toxicity than
for typical environmental exposure, which is more often comparatively low, and
cumulative or chronic. From the LC50 data on Pb in this review and previous Pb AQCDs,
a wide range of sensitivity to Pb is evident across taxa and within genera. As expected,
reported LC50 are usually much higher than current environmental levels of Pb in the U.S
away from heavily exposed sites, even though physiological dysfunction that adversely
impacts the fitness of an organism often occurs well below concentrations that result in
mortality. When available, LCio, NOEC, or LOEC have been reported in the present
document.
Pb is generally not phytotoxic to plants at concentrations found in the environment away
from heavily exposed sites, probably due to the fact that plants often sequester large
amounts of Pb in roots, and that translocation to other parts of the plant is limited. No
data have become available to change this assessment since the 2006 Pb AQCD.
Survival of soil-associated organisms is adversely affected by Pb exposure. In the 1986
Pb AQCD it was reported that Pb at the high extreme of concentrations found near
roadsides and smelters at the time (10,000 to 40,000 mg Pb/kg dry weight) can eliminate
populations of bacteria and fungi on leaf surfaces and in soil. Severe impairment of
decomposition has long been accepted to be one of the most apparent results of soil
contamination with Pb and other metals. In nematodes, the 2006 Pb AQCD reported LC50
values varying from 10 to 1,550 mg Pb/kg dry weight dependent upon soil OM content
and soil pH (U.S. EPA. 2006c). In earthworms, 14 and 28 day LC50 values typically fell
in the range of 2,400-5,800 mg Pb/kg depending upon the species tested. More recent
evidence has been consistent with these values, and also showed concentration-dependent
decreases in survival in collembolans and earthworms under various experimental
conditions. The evidence in terrestrial invertebrates is coherent with evidence in
freshwater invertebrates.
In terrestrial avian and mammalian species, toxicity is observed in laboratory studies over
a wide range of doses (<1 to > 1,000 mg Pb/kg body weight per day) as reviewed for the
development of Eco-SSLs (U.S. EPA. 2005b). and subsequently reported in the
2006 Pb AQCD. The NOAELs for survival ranged from 3.5 to 3,200 mg Pb/kg per day.
Surprisingly, the only study to have reported survival data following exposure to Pb in an
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avion species since the 2006 Pb AQCD, found that survival was greater than in controls
in quail exposed to 50 mg Pb/L in drinking water for 7 weeks (Nain and Smits. 2011).
Evidence for association of Pb exposure with mortality in terrestrial vertebrates is
coherent with observations in freshwater vertebrates (Section 6.4.12.3). Therefore, the
evidence is sufficient to conclude that a causal relationship is likely to exist between Pb
exposures and survival in terrestrial vertebrates and that there is a causal relationship
between Pb exposures and survival in terrestrial invertebrates. The evidence is inadequate
to conclude that there is a causal relationship between Pb exposures and survival in
terrestrial plants.
6.3.12.4 Neurobehavioral Effects-Terrestrial Biota
The central nervous system of animals was recognized as a target of Pb toxicity in the
1977 Pb AQCD (U.S. EPA. 1977). and subsequent Pb reviews have provided additional
supporting evidence of Pb as a neurotoxicant in terrestrial invertebrates and vertebrates.
Effects of Pb on neurological endpoints in terrestrial animal taxa include changes in
behaviors that may decrease the overall fitness of the organism such as food
consumption, prey capture ability and avoidance behaviors.
Some organisms exhibit behavioral avoidance while others do not seem to detect the
presence of Pb (U.S. EPA. 2006c). Decreased food consumption of Pb-contaminated diet
has been demonstrated in some invertebrates (snails) and vertebrates (lizards, pigs).
Decreased food consumption was observed in juvenile A. achatina snails exposed to
Pb-contaminated (concentration greater than 134 mg Pb/kg) diet for 12 weeks (Ebenso
and Ologhobo. 2009a). Similarly, feeding rate in T. pisana snails was depressed in 3
week dietary exposures of 50 to 15,000 mg Pb/kg (El-Gendy et al.. 2011). while other
snails exposed to Pb at similar concentrations have shown no effects on feeding rate
(Beeby and Richmond. 2010). Consumption of 10 mg/Pb kg diet resulted in lower food
intake after 120 days of dietary exposure in pigs (S. domestied) (Yu et al.. 2005).
In limited studies available on nematodes there is evidence that Pb may affect the ability
to escape or avoid predation (Wang and Xing. 2008). Additional new evidence of
changes in the morphology of GABA motor neurons was also found in nematodes
(C. elegans) (Du and Wang. 2009).
Gull chicks experimentally exposed to Pb exhibited abnormal behaviors such as
decreased walking, learning deficits, erratic behavioral thermoregulation, and food
begging that could make them more vulnerable in the wild (Burger and Gochfeld. 2005).
Pb was administered via injection to reach a Pb concentration in feathers equivalent to Pb
levels in feathers of wild gull populations. Lizards exposed to Pb through diet of 10 to
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20 mg Pb/kg per day for 60 days in the laboratory exhibited abnormal coloration and
posturing behaviors.
These findings in terrestrial invertebrates and vertebrates are coherent with findings from
studies in aquatic biota that showed neurobehavioral alterations in various species offish,
and also in some aquatic invertebrates (Section 6.4.12.4). They are also coherent with
findings in laboratory animals that show that Pb induces changes in learning and memory
(Section 4.3.2.3). New behaviors induced by exposure to Pb reviewed in Chapter 4 that
are relevant to effects of Pb observed in terrestrial systems include hyperactivity and
mood disorders, effects on visual and auditory sensory systems, changes in structure and
function of neurons and supporting cells in the brain, and effects on the blood brain
barrier. Mechanisms that include the displacement of physiological cations, oxidative
stress and changes in neurotransmitters and receptors are also reviewed. Data from
ecological studies are highly coherent with these data from animal experiments,
especially neurobehavioral findings and evidence of structural changes. Overall, the
evidence from aquatic and terrestrial systems is sufficient to conclude that a causal
relationship is likely to exist between Pb exposures and neurobehavioral effects in
invertebrates and vertebrates in terrestrial ecosystems.
6.3.12.5 Hematological Effects-Terrestrial Biota
Hematological responses are commonly reported effects of Pb exposure in vertebrates in
terrestrial systems. In the 1977 Pb AQCD, ALAD was recognized as the most sensitive
indicator of Pb exposure in rats (U.S. EPA, 1977). Furthermore, inhibition of ALAD was
associated with death of waterfowl following ingestion of Pb shot. In the 1986 Pb AQCD,
decreases in red blood cell ALAD activity were documented in birds and mammals near a
smelter (Beyer et al.. 1985). Additional evidence for effects on blood parameters and
their applicability as biomarkers of Pb exposure in terrestrial birds and mammals were
presented in the 2005 Ecological Soil Screening Levels for Lead, the 2006 Pb AQCD and
the current ISA (U.S. EPA. 2006c. 2005b). Field studies available since the
2006 Pb AQCD, include evidence for elevated blood Pb levels correlated with decreased
ALAD activity in songbirds and owls living in historical mining areas (Gomez-Ramirez
et al.. 2011: Hansen et al.. 201 la).
This evidence is strongly coherent with evidence from freshwater invertebrates and
vertebrates (Section 6.4.12.5) and observations from human epidemiologic and animal
toxicology studies showing that exposure to Pb induces effects on hematological
endpoints, including altered heme synthesis mediated through decreased ALAD and
ferrochelatase activities, decreased red blood cell survival and function, and increased red
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blood cell oxidative stress. Taken together, the overall weight of human epidemiologic
and animal toxicological evidence is sufficient to conclude that a causal relationship
exists between Pb exposure and decreased RBC survival and function, and altered heme
synthesis in humans and in laboratory animals (Section 4.7). Based on observations in
terrestrial birds and mammals and additionally supported by observations in aquatic
organisms, and toxicological and epidemiological findings in laboratory animals and
humans evidence is sufficient to conclude that there is a causal relationship between Pb
exposures and hematological effects in terrestrial vertebrates. The evidence is inadequate
to conclude that there is a causal relationship between Pb exposures and hematological
effects in terrestrial invertebrates.
6.3.12.6 Physiological Stress-Terrestrial Biota
Induction of enzymes associated with oxidative stress response in terrestrial plants,
invertebrates and vertebrates is a recognized effect of Pb exposure (U.S. EPA, 2006c).
Several studies from the 2006 Pb AQCD in birds and plants provide evidence that Pb
induces lipid peroxidation, however, exposures in these studies were higher than would
be found generally in the environment (U.S. EPA. 2006c). Building on the body of
evidence presented in the 2006 Pb AQCD, recent studies provide evidence of
upregulation of antioxidant enzymes and increased lipid peroxidation associated with Pb
exposure in many species of plants, invertebrates and vertebrates. In plants, increases of
antioxidant enzymes with Pb exposure occur in some terrestrial species at concentrations
approaching the average Pb concentrations in U.S. soils (18.9 mg Pb/kg). For example, in
a series of studies Wang et al. observed increases in reactive oxygen species with
increasing exposure to Pb from 20 mg Pb/kg soil to 2,000 mg Pb/kg in broad bean (V.
faba) (Wangetal.. 2010c; Wang etal.. 201 Oa; Wang et al.. 2008b) and tomato (L.
esculentum) (Wang et al.. 2008a). where they were accompanied up to approximately
500mg Pb/kg by proportional increases in SOD, glutathione, guaiacol peroxidase, and
lipid peroxidation, as well as decreases in catalase. Spinach seedlings grown in soil
containing six increasing concentrations of Pb from 20 to 520 mg Pb/kg exhibited higher
production of reactive oxygen species, increased rates of lipid peroxidation and increased
SOD concentrations (Wang etal.. 201 la). Markers of oxidative damage are also observed
in terrestrial invertebrates, including snails and earthworms, and in terrestrial mammals.
Across these biota, there are differences in the induction of antioxidant enzymes that
appear to be species-dependent.
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The oxidative stress responses associated with Pb exposure in terrestrial plants,
invertebrates and vertebrates are consistent with responses in freshwater
(Section 6.4.12.6) and saltwater organisms (Section 6.4.21.6), and in humans (Section
4.2.4). This oxidative stress is characterized by increased presence of reactive oxygen
species and membrane and lipid peroxidation that can promote tissue damage,
cytotoxicity, and dysfunction. Increases in reactive oxygen species are often followed by
a compensatory and protective upregulation in antioxidant enzymes, such that this
upregulation is itself indicative of oxidative stress conditions. Continuous production of
reactive oxygen species may overwhelm this defensive process, leading to further
oxidative stress and injury.
Upregulation of antioxidant enzymes and increased lipid peroxidation are considered
reliable biomarkers of stress, and provide evidence that Pb exposure induces a stress
response in those organisms which may itself increase susceptibility to other stressors and
reduce individual fitness. Evidence is sufficient to conclude that there is a causal
relationship between Pb exposures and physiological stress in terrestrial plants, and that a
causal relationship is likely to exist between Pb exposure and physiological stress in
terrestrial invertebrates and vertebrates.
6.3.12.7 Community and Ecosystem Level Effects-Terrestrial Biota
Most direct evidence of community and ecosystem level effects is from near stationary
sources where Pb concentrations are higher than typically observed environmental
concentrations for this metal. Impacts of Pb on terrestrial ecosystems near smelters,
mines, and other industrial sources have been studied for several decades. Emissions of
Pb from smelting and other industrial activities are accompanied by other trace metals
(e.g., Zn, Cu, Cd) and SO2 that may cause toxic effects independently or in concert with
Pb. Those impacts include decreases in species diversity and changes in floral and faunal
community composition. Ecosystem-level field studies are complicated by the
confounding of Pb exposure with other factors such as the presence of trace metals and
acidic deposition and the inherent variability in natural systems. In natural systems, Pb is
often found co-existing with other stressors, and observed effects may be due to
cumulative toxicity.
In laboratory and microcosm studies where it is possible to isolate the effect of Pb, this
metal has been shown to alter competitive behavior of species, predator-prey interactions
and contaminant avoidance. These dynamics may change species abundance and
community structure at higher levels of ecological organization. Uptake of Pb into
aquatic and terrestrial organisms and subsequent effects on mortality, growth,
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physiological stress, blood, neurobehavioral and developmental and reproductive
endpoints at the organism level are expected to have ecosystem-level consequences, and
thus provide consistency and plausibility for causality in ecosystem-level effects.
In the 1977 Pb AQCD the potential for Pb to interfere with ecosystem level processes
was explored in a detailed review of a study on the effects of Pb on relationships between
arthropods and leaf litter decomposition (U.S. EPA. 1977). Reduced arthropod density,
biomass and richness were observed in the vicinity of a Pb smelting complex in Missouri.
There were also several studies correlating feeding habits, habitat, and Pb concentrations
in body tissues reported in the 1977 Pb AQCD, specifically in insects and small
mammals indicating that species differences in Pb concentrations are determined in part
by trophic position and habitat preference.
In the 1986 Pb AQCD it was reported that Pb at environmental concentrations
occasionally found near roadsides and smelters (10,000 to 40,000 mg Pb/kg dry weight
[mg Pb/kg]) can eliminate populations of bacteria and fungi on leaf surfaces and in soil
(U.S. EPA. 1986b). Some key populations of soil microorganisms and invertebrates die
off at 1,000 mg Pb/kg soil interrupting the flow of energy through decomposition
processes and altering community structure. At soil concentrations of 500 to 1,000 mg
Pb/kg or higher, populations of plants, microorganisms, and invertebrates may shift
toward Pb-tolerant populations of the same or different species (U.S. EPA. 1986b).
The 2006 Pb AQCD reported that decreased species diversity, changes in floral and
faunal community composition and decreased vigor of terrestrial vegetation were
observed in ecosystems surrounding former smelters including the Anaconda smelter in
southwestern Montana (U.S. EPA. 2006c). Several studies in the 2006 Pb AQCD
documented reduced organic matter decomposition rates and decreased microbial
biomass in areas heavily polluted by metals. Lower abundance and reduced biodiversity
of soil invertebrate communities were observed in field surveys in proximity to Pb
stationary sources.
Recent evidence published since the 2006 Pb AQCD (summarized in Table 6-4) supports
previous findings of a link between high concentration of soil metals and substantial
changes in soil microorganism community composition, as well as decreased abundance
and diversity. In a naturally Pb-enriched forest in Norway, The number of fungal colony
forming units was approximately 10 times lower in the highest Pb soil concentration
(~4.5 mg Pb/g dry weight) than in control soils (Baath et al.. 2005). The composition of
the fungal community was drastically altered, with only one species common to both
soils, and the number of species present was substantially lower.
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The effect of Pb on microbial community function has been quantified previously using
functional endpoints such as respiration rates, fatty acid production, and soil acid
phosphatase and urease activities. These may provide estimates of ecological impacts that
emphasize functionality irrespective of microbial diversity or abundance measurements.
Studies available since the 2006 Pb AQCD provide further evidence of Pb effects on
microbial processes. Pb contamination reduced phenol oxidase activity in several types of
soils; concentrations between 5 and 50 nM Pb (0.001 and 0.01 mg Pb/L) significantly
decreased phenol oxidase activity in all soils tested, while 400 nM (0.08 mg Pb/L) and
greater completely arrested phenol oxidase activity in one soil tested (a high pH sandy
loam) (Carine et al., 2009). Pb concentrations between 50 and 500 mg Pb/kg significantly
reduced microbial abundance and diversity, and also resulted in lower soil phosphatase,
urease, and dehydrogenase activities (Gao et al., 2010b). When the microbial properties
of metal-contaminated urban soils were compared to those of rural soils, significant
differences (Sudova and Vosatka. 2007) were detected in basal community respiration
rates and microbial abundance (Yang et al.. 2006). Gai et al. (2011) examined the
microbial activity of three soils via microcalorimetric methods following Pb exposure.
They noted an increase in activity immediately following Pb application (giving 10, 20,
40, 80, and 160 mg Pb/kg), and theorized that this was a result of rapid mortality of
sensitive microbial species, followed by a concurrent proliferation of Pb-tolerant
microorganisms. As Pb concentrations increased, however, the calculated microbial
growth rate constant decreased, indicating a suppression of microbial activity (Gai et al..
2011). Akerblom et al. (2007) tested the effects of six metals (Cr, Zn, Mo, Ni, Cd, and
Pb) individually. All tested metals had a similar effect on the species composition of the
microbial community. Exposure to a high Pb concentration (52 mg Pb/kg) negatively
affected respiration rates.
In addition to microbial communities, there is new evidence for effects of Pb on other
terrestrial ecosystem components. Increased plant diversity was shown to ameliorate
effects of Pb contamination on a microbial community (Yang et al.. 2007). The presence
of arbuscular mycorrhizal fungi may protect plants growing in Pb-contaminated soils
(Bojarczuk and Kieliszewska-Rokicka. 2010; Sudova and Vosatka. 2007). Invertebrates
affected by Pb in terrestrial systems may be altering community structure. Recent
evidence since the 2006 Pb AQCD, indicates that some species of worms avoid
Pb-contaminated soils (Langdon et al., 2005). Reductions in microbial and detritivorous
populations can affect the success of their predators (U.S. EPA. 2006c). Following a
28-day exposure to field-collected soils contaminated with metals (including Pb at
426 mg Pb/kg), both population growth and individual growth of the earthworm
L. rubellus were diminished (Klok et al.. 2006). The authors proposed that, although
these reductions were unlikely to result in extirpation, avian predators such as the godwit
(Limosa limosa) that feed heavily on earthworms may be affected by a reduction of
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available earthworm biomass. Furthermore, the presence of earthworms increased Pb
uptake by plants (Ruiz et al.. 2011; Sizmur et al.. 2011).
In terrestrial ecosystems, most studies show decreases in microorganism abundance,
diversity, and function with increasing soil Pb concentrations in areas near point-sources.
Specifically, shifts in nematode communities, bacterial species, and fungal diversity have
been observed. Most evidence for Pb toxicity to terrestrial plants, invertebrates and
vertebrates is from single-species assays in laboratory studies. Although the evidence is
strong for effects of Pb on growth (Section 6.3.12.2). reproduction (Section 6.3.12.1) and
survival (Section 6.3.12.3) in certain species, considerable uncertainties exist in
generalizing effects observed under small-scale, particular conditions up to predicted
effects at the ecosystem level of biological organization. In many cases it is difficult to
characterize the nature and magnitude of effects and to quantify relationships between
ambient concentrations of Pb and ecosystem response due to existence of multiple
stressors, variability in field conditions, and to differences in Pb bioavailability at that
level of organization. However, the cumulative evidence for Pb effects at higher levels of
ecological organization is sufficient to conclude that a causal relationship is likely to exist
between Pb exposures and the alteration of species richness, species composition and
biodiversity in terrestrial ecosystems.
6.4 Aquatic Ecosystem Effects
6.4.1 Introduction to Effects of Pb on Aquatic Ecosystems
This section of the Pb ISA reviews the recent literature published since the
2006 Pb AQCD (U.S. EPA. 2006c). on the effects of Pb on freshwater and saltwater
ecosystems. Freshwater and marine/estuarine systems are considered separately due to
differences in Pb speciation, bioavailability of Pb, salinity, and physiological adaptations
of organisms in freshwater versus saltwater environments, as modifying factors for Pb
toxicity. The focus is on the effects of Pb to aquatic organisms including algae, aquatic
plants, invertebrates, vertebrates, and other biota with an aquatic lifestage
(e.g., amphibians). In the freshwater and saltwater sections, aqueous concentrations of Pb
are reported as (ig Pb/L and sediment concentrations are in mg Pb/kg.
In the present document, studies in some freshwater and saltwater organisms are included
where responses are observed at very high Pb concentrations that might not be expected
in most environmental scenarios or where the relevance of the exposure method to
atmospherically-deposited Pb is unknown. These studies can provide mechanistic
information on Pb toxicity, allow for comparison of Pb uptake across taxa, or
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demonstrate the wide range of sensitivity among closely-related species. Furthermore,
although exposure to Pb in natural systems is likely characterized as a chronic, low dose
exposure, it is not always feasible to conduct long-term experiments under natural
conditions. Observations from short-term experiments in which high concentrations are
used can help to elucidate the shape of concentration-response relationships and provide
evidence for a gradient of response to Pb exposure but the extent to which effects would
be observed at concentrations of Pb typically found in the environment is uncertain.
There are a few studies in the following sections for which effects are reported at very
low concentrations of Pb that appear to be below analytical detection limits. These
studies are included to the extent that they provide information on responses to Pb.
However, the difficulty in maintaining low concentrations of Pb and the potential for
contamination limits the interpretation of the reported observations and consideration of
the observed effects in the absence of analytical verification. In these cases, less weight is
placed on study findings in drawing conclusions regarding the effects of Pb exposure.
In the following sections, the literature on aquatic ecosystem effects of Pb, published
since the 2006 Pb AQCD, is considered with brief summaries from the 1977 Pb AQCD,
the 1986 Pb AQCD and the 2006 Pb AQCD where relevant. Biogeochemistry of Pb in
aquatic systems is reviewed in Section 6.4.2. Sections 6.4.3 and 6.4.4 consider the
bioavailability and uptake of Pb by freshwater plants, invertebrates, and vertebrates.
Biological effects of Pb on freshwater ecosystem components (plants, invertebrates, and
vertebrates) are discussed in Section 6.4.5. In this section, effects are generally presented
from sub-organismal responses (i.e., enzymatic activities, changes in blood parameters)
to endpoints relevant to the population-level and higher (growth, reproduction and
survival; summarized in Table 6-5). Biological effects are followed by data on exposure
and response of freshwater species (Section 6.4.6). Effects of Pb at the freshwater
ecosystem level of biological organization are discussed in Section 6.4.7. Section 6.4
includes a discussion of critical loads in freshwater systems (Section 6.4.8).
characterization of sensitivity and vulnerability of freshwater ecosystem components
(Section 6.4.9) and a discussion of Pb effects on ecosystem services (Section 6.4.10). A
synthesis of the new evidence for Pb effects on freshwater organisms (Section 6.4.11) is
followed by causal determinations based on evidence dating back to the 1977 Pb AQCD
(Section 6.4.12). Corresponding sections on saltwater systems introduced in
Section 6.4.13 include bioavailability of Pb in saltwater (Section 6.4.14). biological
effects of Pb in saltwater (Section 6.4.15). exposure and response of saltwater species
(Section 6.4.16). community and ecosystem level effects (Section 6.4.17) and
characterization of sensitivity and vulnerability in saltwater species (Section 6.4.18 ) and
ecosystem services (Section 6.4.19). The saltwater ecosystem section concludes with a
synthesis of new evidence for Pb effects in marine/estuarine systems (Section 6.4.20) and
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causal determinations based on evidence dating back to earlier AQCDs when available
(Section 6.4.21).
6.4.2 Biogeochemistry and Chemical Effects of Pb in Freshwater and
Saltwater Systems
Quantifying Pb speciation in aquatic environments is critical for determining the toxicity
of the metal to aquatic organisms. As reviewed in the 2006 Pb AQCD (U.S. EPA. 2006b)
and discussed in detail in Sections 2.3 and 6.2 of this assessment (Fate and Transport),
the speciation process is controlled by many environmental factors. Although aerially
deposited Pb largely consists of the labile Pb fraction, once the atmospherically-derived
Pb enters surface waters its fate and bioavailability are influenced by Ca2+ concentration,
pH, alkalinity, temperature, total suspended solids, and dissolved organic carbon (DOC),
including humic acids. In sediments, Pb is further influenced by the presence of sulfides
and Fe and Mn oxides. For instance, in neutral to acidic aquatic environments, Pb is
typically present as PbSO4, PbCl4, Pb2+, cationic forms of Pb hydroxide, and ordinary
hydroxide [Pb(OH)2]. In near neutral pH the solubility of Pb is low and Pb may be
present as Pb(OH)2, cerussite (PbCO3), or hydrocerussite (Pb3(OH)2(CO3)), while in
alkaline waters, common forms of Pb include Pb carbonates [Pb(CO3)] and hydroxides
[Pb(OH)2]. In addition to these inorganic forms, Pb humate is present in the solid phase
and Pb fulvate is present in solution. In freshwater systems, Pb complexes with inorganic
OH" and CO32 and forms weak complexes with Cl~; conversely, Pb speciation in
seawater is a function of chloride concentration and the primary species are PbCl3,
PbCO3,PbCl2, andPbCl+.
In many, but not all aquatic organisms, Pb dissolved in water can be the primary
exposure route to gills or other biotic ligands. The toxicity associated with Pb in the water
column or sediment pore waters is directly affected by the competitive binding of Pb to
the anions listed above. In aqueous exposures, Pb is typically added as either the chloride
or the nitrate salt. Both dissociate completely to release lead ions, Pb2+. The Pb ions can
react with ligands in the solution to form soluble complexes and insoluble precipitates.
For solutions of a near neutral pH and lower the solubility of Pb is low. Recent research
has suggested that due to the low solubility of Pb in water, dietary Pb (i.e., lead adsorbed
to sediment, particulate matter, and food) may contribute substantially to exposure and
toxicity in aquatic biota (Sections 6.4.4.2 and 6.4.4.3). Factors determining solubility of
Pb such as distribution of absorbed, dissolved, and solid Pb species can be estimated with
aquatic speciation models discussed below.
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Currently, national and state ambient water quality criteria for Pb attempt to adjust
measured concentrations to better represent the bioavailable free ions, and express the
criteria value as a function of the hardness (i.e., amount of Ca2+ and Mg ions) of the water
in a specific aquatic system. Models such as the BLM (Figure 6-3) (Paquin et al. 2002;
Pi Toro etal.. 2001) include an aquatic speciation model (WHAM V; see below)
combined with a model of competitive binding to gill surfaces, and provides a more
comprehensive method for expressing Pb concentrations at specific locations in terms of
the bioavailable metal. Sediment quality criteria have not been established, although the
EPA has developed methods based on equilibrium partitioning theory to estimate
sediment benchmarks for Pb and a few other metals (U.S. EPA. 2005d). The approach is
based on the ratio of the sum of simultaneously extracted metals and amount of AVS,
adjusted for the fraction of organic carbon present in the sediments, and is reviewed in
detail in the 2006 Pb AQCD (U.S. EPA. 2006c). It is important to note that this method
cannot accurately predict which sediments are toxic or which metal is the primary risk
driver.
A more detailed understanding of the biogeochemistry of Pb in aquatic systems (both the
water column and sediments) is critical to accurately predicting toxic effects of Pb to
aquatic organisms. It should be recognized, however, that in addition to exposure via
sediment and water, chronic exposures to Pb also include dietary uptake, even though the
toxicokinetics of this exposure pathway are not yet well understood in aquatic organisms
and the influence of the bioavailability factors described above is unknown. Furthermore,
changes in environmental factors that reduce the bioaccessible Pb fraction can result in
either sequestration in sediments or subsequent release as mobile, bioaccessible forms.
This section provides updated information about the influence of chemical parameters
that affect Pb bioaccessibility in the aquatic environment (in sediments and the water
column).
Several models are available for estimating the speciation of dissolved Pb. These models
were tested by Balistrieri and Blank (2008) by comparing the speciation of dissolved Pb
in aquatic systems affected by historical mining activities with that predicted by several
models, including Windermere humic aqueous model (WHAM VI), non-ideal
competitive absorption Donnan-type model (NICA-Donnan), and Stockholm humic
model (SHM). Accurate prediction of labile Pb concentrations was achieved only with
SHM, although other metal concentrations were better described by the WHAM model.
Whereas both WHAM VI and NICA-Donnan predicted that the bulk of Pb contamination
would be complexed with Fe, SHM predicted Pb speciation predominantly characterized
by Fe and inorganic Pb complexes. Predicted dynamic Pb concentrations developed with
the WHAM VI and NICA-Donnan methods overestimated Pb concentrations measured
using diffusive gradients in thin-films in Lake Greifen (Switzerland), but underestimated
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concentrations in Furbach stream (located in both the Coeur d'Alene and Spokane River
Basins in Idaho), indicating that such models may not be able to accurately describe
metal speciation under all environmental conditions (Balistrieri and Blank. 2008).
Another model that can be used to calculate the equilibrium composition of dilute
aqueous solutions in the laboratory or natural aqueous systems is MINTEQA2 which
considers dissolved species, absorbed species, and multiple solid phases under a variety
of conditions (U.S. EPA. 2006e).
Quantification of different sediment metal-binding phases, including sulfide, organic
carbon (OC), Fe, and Mn phases, is important to fully understand the bioaccessible
fraction of Pb and the toxicity to benthic organisms (Simpson and Batley. 2007).
However, physical disturbance, pH change, and even the biota themselves also alter
sediment binding or release of Pb. Atkinson et al. (2007) studied the effects of pH on
sequestration or release of Pb from sediments. Although high and circumneutral water pH
(8.1 and 7.2) did not affect the release of sequestered Pb from sediments, lowering the pH
to 6 increased the concentration of Pb in overlying waters from less than 100 (ig Pb/L to
200-300 (ig Pb/L. Physical sediment disturbance also increased the amount of sediment-
bound Pb released into the aqueous phase. When Pb-contaminated sediment was
physically disturbed, the dissolved oxygen content of the overlying water was observed
to significantly impact Pb mobilization, with greater Pb mobilization at lower dissolved
oxygen levels (3 to 9 mg/L O2) (Atkinson et al.. 2007). In addition, although Pb
concentrations in the sediments of a mine-impacted wetland in Hezhang, China, were
determined to be strongly associated with organic/sulfide and residual fractions (e.g., 34
to 82% of total Pb), the presence of aquatic macrophytes altered the Pb speciation,
increasing the fraction of Pb bound to Fe-Mn oxides (42% to 47% of total Pb) (Bi et al..
2007). This phenomenon was investigated in greater depth by Sundby et al. (2005). who
determined that release of oxygen from macrophyte roots resulted in the oxidation of
sediment-bound Pb, leading to the release of bioaccessible Pb fractions (Sundby et al..
2005V
6.4.2.1 Other Metals
Multiple metals are present simultaneously in many aquatic environments and may
interact with one another influencing Pb uptake and toxicity. Interactions of Pb with other
metals were reviewed in the 2006 Pb AQCD, and more recent evidence supports previous
findings of altered bioavailability associated with metal mixtures. Komjarova and Blust
(2008) looked at the effect of the presence of Cd2+ on the uptake of Pb by the freshwater
cladoceran Daphnia magna. While Pb uptake rates were not affected by Cu, Ni or Zn,
enhanced Pb accumulation was observed in the presence of 0.2 (iM Cd. The highest Pb
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concentration, 0.25 (iM (51.8 (ig Pb/L) in turn facilitated Cu uptake. Area-specific and
whole organism Pb transport rates were greatest in the mid-intestine. It was concluded
that Pb-induced disruptions of ion homeostasis and metal absorption processes might be a
possible explanation of stimulated Pb uptake in the presence of Cd, as well as the
increase in Cu uptake rates provoked by presence of Pb at its highest studied
concentration. Komjarova and Blust (2009b) then considered the effect of Na, Ca2+ and
pH on simultaneous uptake of Cd, Cu, Ni, Pb and Zn. Cd and Pb showed increased
uptake rates at high Na concentration. It was thought that increased Na uptake rates
promoted Pb entrance to the cell. With respect to the effect of pH, reduced proton
competition begins to influence Pb uptake in waters with high pH. A clear suppression of
Cd, Ni, Pb and Zn uptake was observed in the presence of Ca2+ (2.5 mM). Ca2+ has been
reported to have a protective effect in other studies (involving other organisms). The
presence of other metals may also affect the uptake of Pb by fish. At low concentrations,
Cd in a Pb-Cd mixture out-competed Pb at gill tissue binding sites in rainbow trout
(Oncorhynchus mykiss), resulting in a less-than additive toxicity when fish were exposed
to both metals in tandem (Birceanu et al., 2008). Evidence for the presence of Pb
influencing the uptake of other metals was observed in the marine bivalvesMacomona
hliana and Austrovenus stutchburyi. Significantly, more Zn bioaccumulated in the
presence of Pb in these mussels than with Zn alone following a 10-day exposure to
spiked sediments (Fukunaga and Anderson. 2011).
6.4.2.2 Biofilm
Farag et al. (2007) measured Pb concentrations in various media (water, colloids,
sediment, biofilm) as well as invertebrates and fish collected within the Boulder River
watershed, MT, U.S. They concluded that the fraction of Pb associated with Fe-oxides
was most frequently transferred to biofilms and the other biological components of the
sampled systems (Farag et al.. 2007). Consequently, an increase in the Pb Fe-oxide
fraction could signify a potential increase in the bioaccessible pool of Pb. The authors
also noted that this fraction may promote downstream transport of Pb contamination.
Ancion et al. (2010) investigated whether urban runoff metal contaminants could modify
biofilm bacterial community structure and diversity and therefore potentially alter the
function of biofilms in stream ecosystems. They found that accumulation rates for metals
in biofilm were maximal during the first day of exposure and then decreased with time.
Equilibrium between metal concentrations in the water and in the biofilm was reached for
all metals after 7-14 days of exposure. The affinity of the biofilm for Pb was, however,
much greater than for Cu and Zn. With respect to recovery, the release of metals was
slow and after 14 days in clean water 35% of Pb remained in the biofilm. By retaining
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and releasing such metal pollutants, biofilms may play a key role in determining both the
concentration of the dissolved metals in the water column and the transfer of the metals
to invertebrates and fish grazing on them. An enrichment factor of 6,000:1 for Pb
between the biofilm and the water was measured after 21 days exposure to synthetic
urban runoff. The relatively slow release of such metal may greatly influence the transfer
of Pb to organisms feeding on the biofilms. This may be of particular importance during
storm events when large amounts of Pb are present in the urban runoff. It was suggested
that biofilms constitute an integrative indicator of metal exposure over a period of days to
weeks.
6.4.2.3 Carbonate
An investigation of heavy metal concentrations in an industrially impacted French canal
(Deule canal) indicated that total extractable Pb in sediments ranged from 27 to
10,079 mg Pb/kg, with 52.3% present in Fe-Mn oxide fractions, 26.9% as organic sulfide
fraction, 10.7% in carbonates, and 10.1% in the residual fraction (Boughriet et al.. 2007).
The relatively high fraction of Pb associated with carbonates was not observed at other
sites, as sediments in these areas contained low proportions of carbonates. Hence,
addition of carbonates (either from anthropogenic or natural sources) can significantly
impact Pb speciation in sediments, and potential bioavailability to resident organisms. In
addition, increased surface water carbonate concentrations also reduced the bioaccessible
Pb fraction as measured by chronic Pb accumulation in the fathead minnow, (Pimephales
promelas) (Mager et al.. 2010). and by Pb toxicity to fathead minnow and the cladoceran
(Ceriodaphnia dubia) (Mager et al.. 201 Ib).
6.4.2.4 Dissolved Organic Matter (DOM)
Uptake of Pb by water-column organisms is affected by the concentration of DOM
(Mager etal.. 201 la: Mager etal.. 2010). In a 7-day chronic study with C. dubia, DOM
protected against toxicity while water hardness was not protective (Mager et al., 201 la).
The specific composition of DOM has been shown to affect the bioaccessibility of
environmental Pb. Humic acid-rich DOM resulted in decreased free Pb ion concentration
when compared to systems containing DOM with high concentrations of polysaccharides
(Lamelas and Slavevkova, 2008). When the sequestering abilities of various components
of DOM were compared, humic acid again was shown to be most efficient at reducing the
Pb free ion concentration, followed by fulvic acid, alginic acid, polygalacturonic acid,
succinoglycan, and xanthan (Lamelas et al.. 2005). Lamelas et al. (2009) considered the
effect of humic acid on Pb(II) uptake by freshwater algae taking account of kinetics and
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cell wall speciation. The uptake flux was described by a Michaelis-Menten type equation.
Comparison of Cu(II), Cd(II) and Pb(II) uptake by green freshwater algae, (Chlorella
kessleri), in the presence of either citric acid or humic acid was made. The uptake fluxes,
percentage adsorbed and percentage internalized for Cu and Cd were identical in the
presence of either citric or humic acid. In contrast, however, there was a ten-fold increase
in the respective values for Pb. The increase in adsorbed Pb was attributed to the increase
in adsorption sites from the adsorbed humic acid on the surface of the algae. Two
hypotheses were considered to explain the increase in internalized Pb and the
internalization flux: (1) direct interaction of Pb-humic acid complexes with the
internalization sites, and (2) uptake of Pb(II) after dissociation from the Pb-humic acid
complex. The authors favor the former hypothesis but no evidence is presented for the
proposed ternary Pb-humic acid-internalized site complexes, nor is there an explanation
as to why this behavior is not observed for Cd or Cu.
There is evidence, however, that DOC/DOM does not have the same effect on free Pb ion
concentration in marine systems as in freshwater systems. No correlation was observed
between DOM concentration or composition and Pb toxicity when examined using the
sea urchin (Paracentrotus lividus) embryo-larval bioassay (Sanchez-Marin et al., 2010a).
For marine invertebrates, the presence of humic acid increased both the uptake and
toxicity of Pb, despite the fact that a larger fraction of Pb is complexed with humic acid
(25 to 75%). Although the authors could not provide a precise explanation for this, they
theorized that in marine environments, addition of humic acid could induce and enhance
uptake of Pb via membrane Ca2+ channels (Sanchez-Marin et al.. 2010b). This
mechanism was observed in the marine diatom (Thalassiosira weissflogii), in that humic
acids absorbed to cell surfaces increased metal uptake; however, water column Pb-humic
acid associations did appear to reduce free Pb ion concentrations (Sanchez-Marin et al..
201 Ob). Formation of a ternary complex that is better absorbed by biological membranes
was another proposed mechanism that could describe the increased bioaccessibility to
marine invertebrates of Pb bound to humic acid (Sanchez-Marin et al.. 2007).
Sanchez-Marin and Beiras (2011) subsequently have shown that different components of
DOM have different effects on Pb bioavailability in marine systems. Their initial research
using Aldrich humic acid found that increasing humic acid concentrations increased Pb
uptake by mussel gills and increased toxicity to sea urchin larvae in marine environments
(Sanchez-Marin et al.. 2007). In contrast, a subsequent investigation found that fulvic
acid reduced Pb bioavailability in marine water (Sanchez-Marin et al., 2011). The
contradictory effects of different components of DOM on marine bioavailability likely
reflect their distinct physico-chemical characteristics. More hydrophobic than fulvic acid,
humic acid may adsorb directly to cell membranes and enhance Pb uptake through some
(still unidentified) mechanism (Sanchez-Marin et al., 2011).
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Humic acid may increase the bioaccessible Pb fraction for green algae through formation
of a ternary complex that promotes algal uptake of the metal. Lamelas and Slaveykova
(2007) found that aqueous Pb formed complexes with humic acid, which in turn would
become adsorbed to C. kesslerii algal surfaces, and that the presence of Pb sorbed to
humic acid did not interfere with humic acid-algae complexation. The authors concluded
that humic acids bound to algae acted as additional binding sites for Pb, thus increasing
the concentrations associated with the algal fraction (Lamelas and Slaveykova. 2007).
Based on the above, the recent literature indicates the existence of a number of deviations
from current models used to predict bioaccessibility of Pb. In marine aquatic systems, for
instance, surface water DOM was found to increase (rather than decrease) uptake of Pb
by fish gill structures, potentially through the alteration of membrane Ca2+-channel
permeability. This phenomenon would not be accurately predicted by a BLM developed
using data from freshwater organisms. Further, in both freshwater and marine
environments, algal biosorption of labile Pb fraction was also increased by humic acid
and DOM, likely through the formation of ternary complexes that increase Pb binding
sites on the algal surface. Although it is unclear whether Pb in this form is available for
toxic action on algae, it is likely to comprise a significant source of dietary Pb for
primary consumers. Moreover, in an attempted field verification of freshwater
bioaccessibility models conducted at sites upstream, within and downstream of a mining
district only one model (SHM) adequately predicted Pb bioaccessibility (Balistrieri and
Blank. 2008).
6.4.2.5 Sulfides
In sediments, Pb bioavailability is further influenced by sulfides. In the presence of
sulfides, most of the reactive metal in sediments will form insoluble metal sulfide that is
not bioavailable for uptake by benthic organisms. Acid volatile sulfide (AVS) has been
used to predict the toxicity of Pb and other metals in sediments (Ankley etal. 1996; Di
Toro et al. 1992) and in the development of sediment quality criteria (Section 6.4.3). The
role of sulfides in the flux of Pb from sediments is discussed further in Section 2.3.2.3.
6.4.3 Introduction to Bioavailability and Biological Effects of Pb in
Freshwater Ecosystems
Freshwater ecosystems across the U.S. encompass many habitats including ponds,
streams, rivers, wetlands and lakes. Concentrations of Pb available for fresh surface-
water and freshwater sediments are reported in Section 6.2.3 and Table 6-2 and are
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summarized here. Representative median and range of Pb concentrations in surface
waters (median 0.50 (ig Pb/L, range 0.04 to 30 (ig Pb/L), sediments (median 28 mg Pb/kg
dry weight, range 0.5 to 12,000 mg Pb/kg dry weight) and fish tissues (geometric mean
0.54 mg Pb/kg dry weight, range 0.08 to 23 mg Pb/kg dry weight [whole body]) in the
U.S. based on a synthesis of NAWQA data reported in the previous 2006 Pb AQCD
(U.S. EPA. 2006c). Additional information on ambient Pb levels in waters, sediments and
biota is presented in Section 2.6.5 and Table 6-2 including new data from the Western
Airborne Contaminants Assessment Project (WACAP) on Pb in environmental media and
biota from remote ecosystems in the western U.S. WACAP assessed concentrations of
semi-volatile organic compounds and metals in up to seven ecosystem components (air,
snow, water, sediment, lichen, conifer needles and fish) in watersheds of eight core
national parks during a multi-year project conducted from 2002-2007 (Landers et al.,
2008). The goals of the study were to assess where these contaminants were
accumulating in remote ecosystems in the western U.S., identify ecological receptors for
the pollutants, and to determine the source of the air masses most likely to have
transported the contaminants to the parks.
The 2006 Pb AQCD (U.S. EPA. 2006b) provided an overview of regulatory
considerations for water and sediments in addition to consideration of biological effects
and major environmental factors that modify the response of aquatic organisms to Pb
exposure. Regulatory guidelines for Pb in water and sediments have not changed since
the 2006 Pb AQCD, and are summarized below with consideration of limited new
information on these criteria since the last review. This section is followed by new
information on biogeochemistry, bioavailability and biological effects of Pb since the
2006 Pb AQCD.
The most recent ambient water quality criteria for Pb in freshwater were released in 1985
(U.S. EPA. 1985) by the EPA Office of Water which employed empirical regressions
between observed toxicity and water hardness to develop hardness-dependent equations
for acute and chronic criterion. These criteria are published pursuant to Section 304(a) of
the Clean Water Act and provide guidance to states and tribes to use in adopting water
quality standards for the protection of aquatic life and human health in surface water. The
ambient water quality criteria for Pb are expressed as a criteria maximum concentration
(CMC) for acute toxicity and criterion continuous concentration (CCC) for chronic
toxicity (U.S. EPA. 2009R In freshwater, the CMC is 65 ng Pb/L and the CCC is
2.5 (ig Pb/L at a hardness of 100 mg/L.
The 2006 Pb AQCD summarized two approaches for establishing sediment criteria for Pb
based on either bulk sediment or equilibrium partitioning (Section 6.2.1 and
Section AX7.2.1.4). The first approach is based on empirical correlations between metal
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concentrations in bulk sediment and associated biological effects to derive threshold
effect concentrations (TEC) and probable effects concentrations (PEC) (MacDonald et
al., 2000). The TEC/PEC approach derives numeric guidelines to compare against bulk
sediment concentrations of Pb. The other approach in the 2006 Pb AQCD was the
equilibrium partitioning procedure published by the EPA for developing sediment criteria
for metals (U.S. EPA. 2005d). The equilibrium partitioning approach considers
bioavailability by relating sediment toxicity to pore water concentration of metals. The
amount of simultaneously extracted metal (SEM) is compared with the metals extracted
via AVS since metals that bind to AVS (such as Pb) should not be toxic in sediments
where AVS occurs in greater quantities than SEM.
Since the 2006 Pb AQCD, both of these methods for estimating sediment criteria for
metals, have continued to be used and refined. The SEM approach was further refined in
the development of the sediment BLM (Di Toro et al., 2005). The BLM is discussed
further in Sections 6.3.3 and 6.4.4 . Comparison of empirical approaches with AVS-SEM
in metal contaminated field sediments shows that samples where either method predicted
there should be no toxicity due to metals, no toxicity was observed in chronic amphipod
exposures (Besser et al., 2009; MacDonald et al., 2009). However, when the relationship
between invertebrate habitat (epibenthic and benthic) and environmental Pb
bioaccumulation was investigated, De Jonge et al. (2010) determined that different
environmental fractions of Pb were responsible for invertebrate uptake and exposure. Pb
uptake by benthic invertebrate taxa was not significantly correlated to AVS Pb levels, but
rather to total sediment concentrations (De Jonge et al.. 2009). Conversely, epibenthic
invertebrate Pb body burdens were better correlated to AVS concentrations, rather than
total Pb sediment concentrations (De Jonge et al.. 2010).
In the following sections, recent information since the 2006 Pb AQCD on Pb in
freshwater ecosystems will be presented. Throughout the sections, brief summaries of
conclusions from the 1977 Pb AQCD, the 1986 Pb AQCD and 2006 Pb AQCD are
included where appropriate. The sections are organized to consider uptake of Pb and
effects at the species level, followed by community and ecosystem level effects. New
research on the bioavailability and uptake of Pb into freshwater organisms including
plants, invertebrates, and vertebrates is presented in Section 6.4.4. Effects of Pb on the
physiology of freshwater flora and fauna (Section 6.4.5) are followed with data on
exposure and response of freshwater organisms (Section 6.4.6). Responses at the
community and ecosystem levels of biological organization are reviewed in Section 6.4.7
followed by a brief consideration of critical loads in freshwater systems (Section 6.4.8).
characterization of sensitivity and vulnerability of ecosystem components (Section 6.4.9)
and a discussion of ecosystem services (Section 6.4.10). The freshwater ecosystem
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section concludes with a synthesis of new evidence (Section 6.4.11) and causal
determinations based on evidence dating back to the 1977 Pb AQCD (Section 6.4.12).
6.4.4 Bioavailability in Freshwater Systems
Bioavailability was defined in the 2006 Pb AQCD as "the proportion of a toxin that
passes a physiological membrane (the plasma membrane in plants or the gut wall in
animals) and reaches a target receptor (cytosol or blood)." In 2007, EPA took cases of
bioactive adsorption into consideration and revised the definition of bioavailability as
"the extent to which bioaccessible metals absorb onto, or into, and across biological
membranes of organisms, expressed as a fraction of the total amount of metal the
organism is proximately exposed to (at the sorption surface) during a given time and
under defined conditions" (U.S. EPA. 2007c). See Section 6.3.3 for additional discussion
of bioavailability.
The bioavailability of metals varies widely depending on the physical, chemical, and
biological conditions under which an organism is exposed (U.S. EPA. 2007c). The
bioavailability of a metal is also dependent upon the bioaccessible fraction of metal. The
bioaccessible fraction of a metal is the portion (fraction or percentage) of
environmentally available metal that actually interacts at the organism's contact surface
and is potentially available for absorption or adsorption by the organism (U.S. EPA.
2007c). The processes for evaluating bioavailability and bioaccessibility are presented in
Figure 6-2 and in Section 6.3.3. In brief, trace metals, and their complexes, must first
diffuse from the external medium to the surface of the organism (mass transport). Metal
complexes may dissociate and re-associate in the time that it takes to diffuse to the
biological surface. These processes are considered further in Chapter 2. To have an effect
on the organism, metals must then react with a sensitive site on the biological membrane
(adsorption/desorption processes), often but not necessarily followed by biological
transport (internalization). Any of these processes may be the rate limiting step for the
overall biouptake process. Internalization is, however, the key step in the overall
biouptake process. Although the transport sites often have a high affinity for required
metals they do not always have high selectivity and so a toxic metal may bind to the site
of an essential metal with a similar ionic radius or co-ordination geometry, e.g., Pb2+,
Cd2+ and Zn2+ are similar to Ca2+. At the molecular level, there are three major classes of
transition metal transporter: P-type ATPases, Zn regulated transporter/Fe-regulated
transporter, and natural resistance associated macrophage proteins (Worms et al., 2006).
Of these, natural resistance associated macrophage proteins have been shown to promote
the uptake of various metals including Pb. This type of trace metal transport can be
described by Michaelis-Menten uptake kinetics and equilibrium considerations.
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Routes of Exposure
According to the 2006 Pb AQCD (U.S. EPA. 2006b). Pb adsorption, complexation,
chelation, etc., are processes that alter its bioavailability to different aquatic species, and
it was suggested that multiple exposure routes may be important in determining overall
bioavailability of Pb. Given its low solubility in water, bioaccumulation of Pb by aquatic
organisms may preferentially occur via exposure routes other than direct absorption from
the water column, including ingestion of contaminated food and water, uptake from
sediment pore waters, or incidental ingestion of sediment. If uptake and accumulation are
sufficiently faster than depuration and excretion, Pb tissue levels may become sufficiently
high to result in physiological effects (Luoma and Rainbow. 2005). Pb accumulation rates
are controlled, in part, by metabolic rate. Other factors that influence bioavailability of Pb
to organisms in aquatic systems are reviewed in Section 6.4.2. As summarized in the
2006 Pb AQCD, organisms exhibit three Pb accumulation strategies: (1) accumulation of
significant Pb concentrations with low rate of loss resulting in substantial accumulation;
(2) balance between excretion and bioavailable metal in the environment; and (3) very
low metal uptake rate without significant excretion, resulting in weak net accumulation
(Rainbow. 1996). Uptake experiments with aquatic plants, invertebrates, and vertebrates
reviewed in the 2006 Pb AQCD showed increases in Pb uptake with increasing Pb in
solution. The 2006 Pb AQCD findings included consideration of bioaccumulation in
different trophic levels. Pb concentrations were found to be typically higher in algae and
benthic organisms and lower in higher trophic-level consumers.
In this section:
1) Recent information on bioavailability and uptake in algae, plants,
invertebrates and vertebrates from freshwater systems are reviewed with
summary material from the 2006 Pb AQCD and earlier Pb AQCDs where
appropriate.
2) An overview of the BLM is presented as the most widely used method for
predicting both the bioaccessible and bioavailable fractions of Pb in the
aquatic environment.
3) Bioavailability in algae, plants, invertebrates and vertebrates is discussed. As
reviewed by Wang and Rainbow (2008), aquatic organisms exhibit distinct
patterns of metal bioaccumulation. The authors suggest that the observed
differences in accumulation, body burden, and elimination between species
are due to metal biogeochemistry and physiological and biological responses
of the organism. The studies presented below generally support the
observations of Wang and Rainbow (2008) that closely related species can
vary greatly in bioaccumulation of Pb and other non-essential metals.
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The bioaccumulation and toxicity of Pb to aquatic organisms are closely linked to the
environmental fate of the metal under variable environmental conditions (Section 2.3) as
they are highly dependent upon the relative proportion of free metal ions in the water
column. However, information is lacking on the uptake of Pb through ingestion of
Pb-sorbed particles or dietary exposure to biologically-incorporated Pb. Such routes of
exposure are not included in models such as the BLM that predict toxicity as a function
of Pb concentration in the water column. This uncertainty may be greater for Pb than for
other more soluble metals (such as Cu) as a greater proportion of the total mass of Pb in
an aquatic ecosystem is likely to be bound to particulate matter. Therefore, estimating
chronic toxicity of Pb to aquatic receptors may have greater uncertainty than predicting
acute effects.
BLM Models
In addition to the biogeochemical effects that govern the environmental pool of
accessible Pb, reactions of Pb with biological surfaces and membranes determines the
bioavailability and uptake of the metal by aquatic organisms. The BLM (Figure 6-3)
predicts both the bioaccessible and bioavailable fraction of Pb in the aquatic
environment, and can be used to estimate the importance of environmental variables such
as DOC in limiting uptake by aquatic organisms (Alonso-Castro et al., 2009). The BLM
integrates the binding affinities of various natural ligands in surface waters and the
biological uptake rates of aquatic organisms to determine the site-specific toxicity of the
bioavailable fraction.
In the 2006 Pb AQCD, limitations of the use of BLM in developing air quality criteria
were recognized including the focus of this model on acute endpoints and the absence of
consideration of dietary uptake as a route of exposure. Atmospheric deposition of Pb to
aquatic systems and subsequent effects on ecosystem receptors is likely characterized as a
chronic, cumulative exposure rather than an acute exposure. Recommendations from the
2006 Pb AQCD included developing both chronic toxicity BLMs and BLMs that
consider the dietary route of Pb uptake. The EPA recently incorporated the BLM into the
Framework for Metal Risk Assessment (U.S. EPA, 2007c) and has published an ambient
freshwater criteria document for Cu based on the BLM model (U.S. EPA. 2007a). This
section reviews the literature from the past 5 years on applications of the BLM to
predicting bioavailability of Pb to aquatic organisms. However, the primary focus of
initial BLMs has been acute toxicity endpoints for fish and invertebrates following gill or
cuticular uptake of metals.
Di Toro et al. (2005) constructed BLMs for metals exposure in sediments, surface water,
and sediment pore water to determine how to most accurately predict the toxicity of
metals-contaminated sediments. Results from models were compared with literature-
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derived acute toxicity values for benthic and epibenthic invertebrates to establish the
accuracy of the developed models. Although the models tended to overestimate the
toxicity of aqueous and sediment-bound Pb in freshwater environments, it was
determined that the model significantly underestimated Pb toxicity to marine
invertebrates (Pi Toro et al.. 2005). This may be because pore water metal concentrations
were not modeled. Consequently, these results may suggest that either 1) mobilization of
Pb concentrations from sediments into pore water is greater in marine environments, or 2)
marine invertebrates are significantly more sensitive to Pb exposures than are freshwater
species.
A number of deviations from results predicted by Pb exposure models (such as the BLM)
were documented by Ahlf et al. (2009). They highlighted that uptake of metals by
sediment-dwelling bivalves was significantly greater than predicted, because bivalves
accumulate Pb from multiple sources not included in the model, such as ingestion of
algae, bacteria, and colloidal matter. Species-specific dietary assimilation of ingested
particulate-bound metals is also likely to play a role in the toxicity of Pb to aquatic
organisms, yet insufficient data are available to permit modeling of this additional factor
(Ahlf etal. 2009). The authors outlined the need for additional data in developing
bioavailability models for chronic metal exposures. As recent evidence suggests that the
hydrophobic DOC fraction (e.g., humic and fulvic acids) sequesters the greatest fraction
of Pb in aquatic systems (Pernet-Coudrier et al.. 2011). understanding the influence of
this adsorption on Pb toxicity is critical for the prediction of chronic aquatic Pb toxicity.
For instance, although the presence of humic acid is considered to reduce the bioavailable
fraction of metals in surface water, green algae uptake and biosorption of metals,
including Pb, was actually increased by humic acid. The authors determined that humic
acid bound to algal surfaces served to increase the total number of metal binding sites
over those afforded solely by the algal surface (Lamelas and Slavevkova. 2007). This
highlights the complexity of modeling chronic metals bioavailability through multiple
exposure routes, as humic acid would decrease gill or cuticular uptake of metals from the
water column, but could potentially enhance dietary exposure by increasing algal metal
content. Slaveykova and Wilkinson (2005) also noted that humic acid is likely to interact
with other biological membranes and alter their permeability to metals, especially in
acidic environments. Further, they observed that increased surface water temperatures not
only increase membrane permeability but also change metabolic rates, both of which can
enhance metals uptake and assimilation; however, this factor is not included in
bioavailability models such as the BLM (Slavevkova and Wilkinson. 2005). Despite this,
the authors noted that, in most cases, the BLM could predict acute metals toxicity with a
reasonable degree of accuracy.
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6.4.4.1 Freshwater Plants and Algae
In the 1977 Pb AQCD, the root system of plants was recognized as the major route of
uptake for Pb (U.S. EPA. 1977). Uptake and translocation studies of Pb in plants and
algae reviewed in the 1977 Pb AQCD and the 2006 Pb AQCD indicated that plants tend
to sequester larger amounts of Pb in their roots than in their shoots. Recent studies on
bioavailability of Pb to plants support the findings of the previous Pb AQCDs and
provide additional evidence for species-dependent differences in responses to Pb in water
and sediments.
Most biouptake studies in aquatic plants and algae available since the 2006 Pb AQCD,
were typically conducted at very high concentrations of Pb that are not representative of
current levels of Pb typically encountered in freshwater. However, most of these
exposures included a series of increasing concentrations of Pb and generally, Pb was
accumulated in a dose-dependent manner. Studies in which high concentrations of Pb are
used and an exposure-response relationship is observed may imply effects at lower
concentrations but uncertainty remains to the extent to which effects would be observed
at concentrations of Pb typically found in the environment. The role of modifying factors,
such as the presence of other metals, on uptake rates as well as species differences in Pb
uptake rates can be determined from experimental Pb concentrations that are higher than
measured Pb in the environment. Plants that are hyperaccumulators of Pb and other
metals may be used for phytoremediation at highly contaminated sites and there is a large
body of literature on uptake of very high concentrations of metals by different species.
This chapter focuses on environmentally relevant concentrations of Pb and also those
studies with doses or exposures in the range of one or two orders of magnitude above
current or ambient conditions, as described in the Preamble. In freshwater ecosystems in
the U.S., the average Pb concentration in surface water is 0.5 (ig Pb/L (Table 6-2).
however, total Pb in water has been measured as high as 2,000 (ig Pb/L where mining
operations have affected streams (Table 2-11).Studies with freshwater algae available
since the 2006 Pb AQCD, are primarily limited to nominal media exposures at high
concentrations of Pb with metal quantified in tissues. For example, the microalgae
Spirulina platensis was demonstrated to accumulate Pb from Zarrouk culture medium in
a concentration-dependent manner with nominal initial concentrations of 5,000, 10,000,
30,000, 50,000 and 100,000 (ig Pb/L (Pb in medium was measured every two days
thereafter), following a 10-day incubation period (Arunakumara et al., 2008). Pb
concentrations accumulated by algae appeared to decrease when culture time increased
from 2 to 10 days. This may have occurred as a result of a gradual recovery of growth
and an addition of biomass that would have reduced the concentration of Pb in algal
tissue. An aquatic moss, Fontinalis antipyretica, accumulated up to an average of 622 mg
Pb/kg dry weight over a 7-day nominal exposure to 20,700 (ig Pb/L (Rau et al.. 2007).
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Interestingly, experimentation with concurrent Cu and Pb exposure indicated that the
presence of Cu increased the uptake of Pb by the green algae Chlamydomonas reinhardtii
(Chen et al., 2010c). The authors noted that, in the case of Cu-Pb binary exposures,
uptake rates of Pb exhibited complex non-linear dynamics in other aquatic organisms as
well.
Additional uptake studies conducted since the 2006 Pb AQCD include new information
for freshwater macrophytes. When exposed to nominal water concentrations of up to
20,700 (ig Pb/L, floating (non-rooted) coontail plants (Ceratophyllum demersum)
accumulated an average Pb concentration of 1,748 mg Pb/kg after 7 days, although this
was not significantly higher than levels accumulated in the first day of exposure (Mishra
et al., 2006b). Induction of the antioxidant system improved the tolerance of the aquatic
plant Najas indica for bioaccumulated Pb, allowing for increased biomass and the
potential to accumulate additional Pb mass. High Pb accumulation (3,554 mg Pb/kg dry
weight tissue following a 7-day exposure to 20,720 (ig Pb/L) was considered to be a
function of plant morphology; as a submerged, floating plant, N. indica provides a large
surface area for the absorption of Pb (Singh etal.. 2010).
Given that atmospherically-derived Pb is likely to become sequestered in sediments
(Section 6.2). uptake by aquatic macrophytes is a significant route of Pb removal from
sediments, and a potential route for Pb mobilization into the aquatic food web. The rooted
aquatic macrophyte Eleocharis acicularis was determined to be a hyperaccumulator of
Pb in an 11-month bioaccumulation experiment with mine tailings. When grown in
sediments containing 1,930 mg Pb/kg, the maximum concentration of Pb in E. acicularis
was determined to be 1,120 mg Pb/kg dry weight. However, calculated BCFs for Pb were
all less than one, indicating that Pb uptake, although high, was less efficient than for other
metals present (Ha et al.. 2009).
Aquatic plants inhabiting a wetland containing an average sediment Pb concentration of
99 mg Pb/kg exhibited variable Pb tissue concentrations, but these do not appear to be
related to macrophyte type (e.g., submerged, floating, emergent, etc.). Consequently, the
authors concluded that uptake of Pb by aquatic plants appears to be dependent on species,
at the exclusion of habitat or type. For instance, among the submerged plant species,
Ceratophyllum demersum accumulated the greatest amount of Pb (22 mg Pb/kg dry
weight), while Potamogeton malainus tissue contained the least amount of Pb, 2.4 mg
Pb/kg dry weight (Bi et al., 2007). Tissues of the floating plants Azolla imbricata and
Spirogyra communis were found to contain 12 and 20 mg Pb/kg dry weight, respectively,
while emergent macrophytes Scirpus triqueter an&Alternanthera philoxeroides
accumulated 1.4 and 10 mg Pb/kg dry weight. Fritioff and Greger (2006) determined that
anywhere from 24-59% of the total Pb taken up by Potamogeton natans aquatic plants
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was sequestered in the cell wall fraction, depending on plant tissue and environmental Pb
concentration. More importantly, no translocation of Pb was observed when plant tissues
(leaf, stem, root) were exposed to Pb solutions separately (Fritioff and Greger. 2006).
Dwivedi et al. (2008) reared nine different species of aquatic plants in a fly-ash
contaminated medium containing approximately 7 mg Pb/kg dry weight. Not only did
species exhibit different Pb accumulation efficiencies but they also compartmentalized
sequestered Pb differently. The submerged macrophyte Hydrilla verticillata accumulated
the greatest amount of Pb (approximately 180 mg Pb/kg dry weight tissue), but Pb was
sequestered solely in the shoot tissue. In contrast, other plant species accumulated
between 15 and 100 mg Pb/kg dry weight (Ranunculus scloralus andMarsilea
quadrifolia) with the majority compartmentalizing the metal in root tissue, except for
C. demersum and M. quadrifolia, which also utilized shoot tissue for Pb storage (Dwivedi
et al.. 2008).
Pb concentrations in the root, leaf, and stem tissues of three aquatic plant species were
found to correlate most closely with the concentration of the exchangeable Pb fraction
(e.g., the fraction of Pb that is easily and freely leachable from the sediment). Authors
noted that seasonal variations can alter the amount of Pb present in the exchangeable
fraction, and that Pb was more likely than Cd or Cu to remain tightly bound to sediments,
and therefore the relationship between total sediment Pb and Pb in aquatic plant tissues
was weaker (Ebrahimpour and Mushrifah. 2009).
Lemna sp., a free floating macrophyte, incubated in a water extract of waste ash
containing 19 (ig Pb/L accumulated 3.5 mg Pb/kg dry weight over 7 days of exposure.
Slight toxic effects, including suppression of growth, were observed over this exposure
period, but this may have been a result of exposures to multiple metals in the water
extract, including Cr, Mn, Cu, and Zn (Horvat et al., 2007). Lemna sp. was also
demonstrated to be effective in the biosorption of Pb from solution, even in the presence
of sediments (1 g per 700 mL water). Over 7 days of exposure to 3,600 and
7,000 (ig Pb/L, plant biomass was found to contain an average of 2,900 and 6,600 mg/kg
(wet weight) Pb, respectively, versus 200 and 300 mg/kg (dry weight) in sediment (Kurd
and Sternberg. 2008).
Young Typha latifolia, another rooted macrophyte, were grown in analytically verified
concentrations of 5,000 and 7,500 (ig/L Pb-spiked sediment for 10 days to determine
their value as metal accumulators. Within the exposure period, plants exposed to the
lower concentration were able to remove 89% of Pb, while 84% of the Pb present in the
higher treatment was taken up by T. latifolia. Pb concentrations measured in root and leaf
tissue ranged from 1,365 to 4,867 mg Pb/kg and 272 to 927 mg Pb/kg, respectively, and
were higher at the greater Pb exposure (Alonso-Castro et al.. 2009).
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Uptake studies available for aquatic macrophytes since the 2006 Pb AQCD include some
studies where Pb was measured in field collected plants growing in metal-contaminated
areas. Common reeds (Phragmites australis) grown in metal-impacted aquatic
environments in Sicily, Italy, preferentially accumulated Pb in root and rhizome tissues
(Bonanno and Lo Giudice. 2010). Pb concentrations in water and sediment averaged
0.4 (ig Pb/L and 2.7 mg Pb/kg. These levels yielded root and rhizome concentrations of
17 and 15 mg Pb/kg, respectively, whereas stem and leaf Pb concentrations were lower
(9.9 and 13 mg Pb/kg). These tissue concentrations were significantly correlated to both
water and sediment concentrations (Bonanno and Lo Giudice. 2010). Conversely, the
semi-aquatic plant Ammania baccifera, grown in mine tailings containing 35 to 78 mg
Pb/kg, did not accumulate analytically detectable levels of Pb in either root or shoot
tissues, despite the fact that other metals (Cu, Ni, Zn) were bioaccumulated (Das and
Maiti. 2007). This would indicate that at low/moderate environmental Pb concentrations,
some plant species may not bioaccumulate significant (or measurable) levels of Pb.
The average concentration of Pb in the tissues of rooted aquatic macrophytes (Callitriche
verna, P. natans, C. demersum, Polygonum amphibium, Veronica beccabungd) collected
from two metals-polluted streams in Poland (average sediment concentration 38 to 58 mg
Pb/kg) was less than 30 mg Pb/kg. Pb bioaccumulation in plants was significantly
correlated with sediment Pb concentrations (Samecka-Cymerman and Kempers. 2007). A
similar significant correlation was established between reed sweet grass root Pb
concentration and sediment Pb concentrations (Skorbiowicz. 2006).
Pb tissue concentrations of aquatic plants P. australis and Ludwigia prostrata collected
from wetlands containing an average of 52 mg Pb/kg in surficial sediments were
predominantly in root tissues, indicating poor translocation of Pb from roots. In the
former, Pb decreased from an average of 37 mg Pb/kg in roots to 17, 14, and
12 mg Pb/kg in rhizome, stem and leaf tissues, respectively, while L. prostrata Pb tissue
concentrations decreased from 77 mg Pb/kg in fibrous root to 7 and 43 mg Pb/kg in stem
and leaf tissues (Yang et al., 2008a). The authors proposed that this diminished transfer
ability explained the relatively low BCFs for Pb uptake in these two species, when
compared with those of other metals.
Despite no significant seasonal effect on surface water Pb concentrations, shining
pondweed (Potamogeton lucens), a rooted aquatic macrophyte grown in an urbanized
metal-contaminated lake in Turkey, exhibited seasonal alterations in Pb tissue
concentrations. Average water Pb concentrations were 28 (ig Pb/L in spring, 27 (ig Pb/L
in summer, and 30 (ig Pb/L in autumn. Over this same time period, root tissue Pb
concentrations significantly increased from 6 mg Pb/kg dry weight in spring, to 9 mg
Pb/kg dry weight in summer, and to 10 mg Pb/kg dry weight in autumn (Duman et al..
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2006). No differences were detected in stem Pb concentrations between spring and
summer (approximately 4 mg Pb/kg dry weight), but stem Pb concentrations were found
to be significantly higher in autumn (6 mg Pb/kg dry weight). In the same system,
P. australis plants accumulated the most Pb during winter: 103, 23, and 21 mg Pb/kg dry
weight in root, rhizome, and shoot tissue, respectively, in sediments containing 13 mg
Pb/kg dry weight. By contrast, Schoenoplectus lacustris accumulated maximum rhizome
and stem Pb concentrations of 5.1 and 7.3 mg Pb/kg dry weight in winter, but sequestered
the greatest amount of Pb in root tissues during the spring (30 mg Pb/kg dry weight) at a
comparable sediment concentration, 18 mg Pb/kg dry weight (Duman et al., 2007). The
authors suggest that this indicated that metal uptake was regulated differently between
species.
Tree species that inhabit semi-aquatic environments have also been shown to absorb Pb
from Pb-contaminated sediments. Bald-cypress trees (Taxodium distichum) growing in
sediments of a refinery-impacted bayou in Louisiana accumulated significantly greater
amounts of Pb than did trees of the same species growing in bankside soil, despite the
lower Pb concentrations of sediments. Bankside soils contained greater than 2,700 mg
Pb/kg versus concentrations of 10 to 424 mg Pb/kg in sediments, yet Pb concentrations in
trees averaged 4.5 and 7.8 mg Pb/kg tissue, respectively (Devall et al.. 2006). The authors
theorized that Pb was more readily released from sediments and that soil dispersion to the
swamp sediments provides additional, if periodic, loads of Pb into the system. Willow
seedlings planted in Pb-contaminated sediment were more effective at removing Pb from
the media than a diffusive gradient in thin film technique predicted (Jakl et al.. 2009).
The authors proposed that the plant's active mobilization of nutrients from soil during
growth also resulted in increased Pb uptake and sequestration.
Given that sediments are a significant sink for Pb entering aquatic systems, it is not
surprising that rooted macrophytes bioaccumulate significant quantities of the metal.
Although there are some similarities to Pb accumulation observed in terrestrial plants
(e.g., preferential sequestration of the metal in root tissue), Pb appears to be more
bioavailable in sediment than it is in soil. This may be a result of differences in plant
physiology between aquatic and terrestrial plants (e.g., more rapid growth or more
efficient assimilation of nutrients and ions from a water-saturated medium). While rooted
macrophytes are likely to be chronic accumulators of Pb sequestered in sediments, aerial
deposition of Pb into aquatic systems may result in pulsed inputs of labile Pb that would
be available for uptake by floating macrophytes and algae.
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6.4.4.2 Freshwater Invertebrates
Uptake and subsequent bioaccumulation of Pb in freshwater invertebrates varies greatly
between species and across taxa as previously characterized in the 2006 Pb AQCD. This
section expands on the findings from the 1986 Pb AQCD and 2006 Pb AQCD on
bioaccumulation and sequestration of Pb in aquatic invertebrates. In the case of
invertebrates, Pb can be bioaccumulated from multiple sources, including the water
column, sediment, and dietary exposures, and factors such as proportion of bioavailable
Pb, lifestage, age, and metabolism can alter the accumulation rate. In this section, new
information on Pb uptake from freshwater and sediments by invertebrates will be
considered, followed by a discussion on dietary and water routes of exposure and factors
that influence species-specific Pb tissue concentrations such as invertebrate habitat and
functional feeding group.
In a recent uptake study in freshwater mussels available since the 2006 Pb AQCD, the
Eastern elliptic mussel (Elliptic complanatd) was shown to accumulate Pb rapidly from
water and then reach an equilibrium with exposure level and tissue concentration by two
weeks following average daily exposures of 1, 4, 14, 57 or 245 (ig Pb/L as Pb nitrate
(Mosher et al. 2012). Tissue concentrations of Pb increased at an exposure-dependent
rate for the first 14 days and then did not change significantly for the remainder of the 28-
day exposure although mussels continued to accumulate Pb. At the end of the exposure
period, average Pb in tissue ranged from 0.33 to 898 mg Pb/kg. The authors concluded
that the mussels were likely eliminating Pb via pseudo feces and through storage of Pb in
shell.
The 2006 Pb AQCD (U.S. EPA. 2006b) summarized studies of uptake of Pb from
sediment by aquatic invertebrates and noted that sediment pore water, rather than bulk
sediment, is the primary route of exposure. However, a recent study suggests that in the
midge, Chironomus riparius, total metal concentrations in bulk sediment are better
predictors of metal accumulation than dissolved metal concentrations in sediment pore
water based on bioaccumulation studies using contaminated sediments from six different
sites (Roulier et al.. 2008a). Vink (2009) studied six river systems and found that, for a
range of metals, uptake by benthic organisms (the oligochaete, Limnodrilus (Family
Tubificidae) and the midge, C. riparius) from the sediment pore water (as compared with
surface water) was observed only occasionally, and solely for Pb. The physiological
mechanisms of Pb uptake are still unclear but it is suggested that uptake and elimination
of Pb obey different mechanisms than for other heavy metals.
The 2006 Pb AQCD recognized the potential importance of the dietary uptake pathway
as a source of Pb exposure for invertebrates. Specifically, in a study with the freshwater
amphipod Hyalella azteca, dietary exposure was found to contribute to the chronic
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toxicity of Pb, while acute toxicity was unaffected (Besser et al., 2004). Since the
2006 Pb AQCD, additional studies have considered the relative importance of water and
dietary uptake of Pb in aquatic invertebrates. A stable isotope technique was used to
simultaneously measure uptake of environmentally relevant concentrations of Pb
(10.4 (ig Pb/L) in the water column by the freshwater cladoceran D. magna directly from
water and through food, the green algae Pseudokirchneriella subcapitata. (Komjarova
and Blust 2009a). D. magna accumulated the metal from both sources, but the relative
proportion of uptake from each source changed over the exposure period. After the first
day of exposure, 12% of accumulated Pb was determined to have been absorbed from
dietary (algal) sources, but this percentage decreased by day four of exposure to 4%. Pb
absorbed from water exposure only resulted in Daphnia body burdens of approximately
62.2 mg Pb/kg dry weight (300 (imol Pb/kg dry weight), and was similar to the amount
absorbed by algae (Komjarova and Blust. 2009a). In a comparison of dietary and
waterborne exposure as sources of Pb to aquatic invertebrates, no correlation between Pb
uptake and dietary exposure was observed in the amphipod H. azteca fBorgmann et al..
2007).
Stable isotope analysis was to used measure uptake and elimination simultaneously in
net-spinning caddisfly larvae (Hydropsyche sp.) exposed to aqueous Pb concentrations of
0.2 (control) or 0.6 (ig Pb/L for 18 days (Evans et al., 2006). The measured uptake
constant for Pb in this study was 7.8 g/dry weight per day, and the elimination rate
constant of 0.15/day for Pb-exposed larvae was similar in both presence and absence of
the metal in the water. Tissue concentrations ranged from approximately 15 to 35 mg
Pb/kg. Hydropsychid Pb BCFs ranged from 41 to 65, and averaged 54, indicating a
relatively high accumulation when compared to other metals tested (average BCF of 17
for Cd, 7.7 for Cu, and 6.3 for Zn) (Evans et al.. 2006).
Recent reports on the tissue distribution of Pb in freshwater organisms generally support
the findings of the 2006 Pb AQCD that Pb is primarily sequestered in the gills,
hepatopancreas, and muscle. Uptake of Pb by the crayfish Cherax destructor exposed to
nominal concentration of 5,000 (ig Pb/L as Pb nitrate for 21 days resulted in
accumulation at the highest concentration in gill, followed by exoskeleton >mid-gut
gland >muscle >hemolymph (Morris et al.. 2005). Body burden analysis following 96
hour nominal exposure to 50, 100 and 500 (ig Pb/L as Pb nitrate in the freshwater snail
Biomphalaria glabrata indicated that bioaccumulation increased with increasing
concentrations of Pb and the highest levels were detected in the digestive gland (Ansaldo
et al.. 2006).
When the relationship between invertebrate habitat (epibenthic and benthic) and
environmental Pb bioaccumulation was investigated, De Jonge et al. (2010) determined
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that different environmental fractions of Pb were responsible for invertebrate uptake and
exposure. Pb uptake by benthic invertebrate taxa was not significantly correlated to AVS
Pb levels, but rather to total sediment concentrations (De Jonge et al., 2009). Conversely,
epibenthic invertebrate Pb body burdens were better correlated to AVS concentrations,
rather than total Pb sediment concentrations (De Jonge et al., 2010). For instance, the
biologically available Pb (e.g., bound to metal-rich granules or metallothioneins)
accumulated by the oligochaete Tubifex tubifex was determined to correlate with
sediment SEM-AVS Pb concentrations (De Jonge et al.. 2011). Similarly, Desrosiers et
al. (2008) reported that Pb accumulation by chironomid larvae from St. Lawrence river
sediments was significantly correlated to both total Pb and reactive Pb sediment
concentrations.
Both inter- and intra-specific difference in Pb uptake and bioaccumulation may occur in
macroinvertebrates of the same functional-feeding group. Cid et al. (2010) reported
significant differences in Pb bioaccumulation between field collected Ephoron virgo
mayflies and Hydro psyche sp, caddisflies, with only the mayfly exhibiting increased Pb
tissue concentrations when collected from Pb-contaminated sites; the caddisfly Pb tissue
concentrations were similar between reference and Pb-contaminated areas. The authors
also examined the lifestage specific accumulation of Pb for E. virgo mayflies, and
although there was no statistical difference in Pb tissue concentrations between different
lifestages, Pb bioaccumulation did change as mayflies aged (Cidet al.. 2010).
Reported BAF values for Pb in aquatic invertebrates from the 2006 Pb AQCD ranged
from 499 to 3,670 [Table AX7-2.3.2 (U.S. EPA. 2006c)1. Since the 2006 Pb AQCD,
additional BAF values have been established for invertebrates in field studies which tend
to be higher than BCF values calculated in laboratory exposures (Casas et al., 2008;
Gagnon and Fisher. 1997). A complicating factor in establishing BAF values is that
laboratory studies usually assess uptake in water-only or sediment only exposures while
field studies take into account dietary sources of Pb as well as waterborne Pb resulting in
BAF values that are frequently 100-1,000 times larger than BCF values for the same
metal and species (DeForest et al.. 2007). The EPA Framework for Metals Risk
Assessment states that the latest scientific data on bioaccumulation do not currently
support the use of BCFs and BAFs when applied as generic threshold criteria for the
hazard potential of metals (U.S. EPA. 2007c). See Section 6.3.3 for further discussion.
As reviewed by Wang and Rainbow (2008) and supported by additional studies reviewed
in the present document, there are considerable differences between species in the
amount of Pb taken up from the environment and in the levels of Pb retained in the
organism. The bioaccumulation and subsequent toxicity of Pb to aquatic organisms
(Section 6.4.5) are closely linked to the environmental fate of the metal under variable
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environmental conditions (Sections 2.3 and 6.2) as they are highly dependent upon the
proportion of free metal ions in the water column.
6.4.4.3 Freshwater Vertebrates
Uptake of Pb by vertebrates considered here includes data from fish species as well as a
limited amount of new information on amphibians and aquatic mammals. The
bioaccessibility and bioavailability of Pb is affected by abiotic and biotic modifying
factors considered in Sections 6.4.2 and 6.4.4. In fish, Pb is taken up from water via the
gills and from food via ingestion. Amphibians and aquatic mammals are exposed to
waterborne Pb primarily through dietary sources. In the 2006 Pb AQCD, dietary Pb was
recognized as a potentially significant source of exposure to all vertebrates since Pb
adsorbed to food, particulate matter and sediment can be taken up by aquatic organisms.
Since the 2006 Pb AQCD, tissue accumulation of Pb via gill and dietary uptake has been
further characterized in freshwater fish and new techniques such as the use of stable
isotopes have been applied to further elucidate bioaccumulation of Pb. For example,
patterns of uptake and subsequent excretion of Pb in fish as measured by isotopic ratios
of Pb in each tissue can determine whether exposure was due to relatively long term
sources (which favor accumulation in bone) or short term sources (which favors
accumulation in liver) (Miller et al.. 2005). Recent information since the 2006 Pb AQCD,
on uptake of Pb by fish from freshwater is reviewed below, followed by studies on
dietary uptake as a route of Pb exposure. Next, tissue accumulation patterns in fish
species are reported with special consideration of the anterior intestine as a newly
identified target of Pb from dietary exposures. Finally, studies that report Pb tissue
concentrations in amphibians, reptiles and freshwater mammals are considered.
Freshwater Fish
Pb uptake in freshwater fish is accomplished largely via direct uptake of dissolved Pb
from the water column through gill surfaces and by ingestion of Pb-contaminated diets.
According to the data presented in the 2006 Pb AQCD (U.S. EPA. 2006b). accumulation
rates of Pb are influenced by both environmental factors, such as water pH, DOC, and
Ca2+ concentrations, and by species-dependent factors, such as metabolism, sequestration,
and elimination capacities. The effects of these variables on Pb bioaccumulation in fish
are largely identical to the effects observed for invertebrates (discussed above).
Pb in fish is primarily found in bone, gill, blood, kidney and scales (Spry and Wiener.
1991). Since the 2006 Pb AQCD, multiple studies on uptake of Pb from water by fathead
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minnow and subsequent tissue distribution have been conducted. Spokas et al. (2006)
showed that Pb accumulates to the highest concentration in gill when compared to other
tissues over a 24-day exposure. This pattern was also observed in larval fathead minnows
exposed to 26 (ig Pb/L for 10-30 days, where gill exhibited the highest Pb concentration
compared to carcass, intestine, muscle and liver (Grosell et al.. 2006a). In the larval
minnows, Pb concentration in the intestine exhibited the highest initial accumulation of
all tissues on day 3 but then decreased for the remainder of the experiment while
concentrations in the other organs continued to increase. By day 30, gill tissue exhibited
the highest Pb concentration (approximately 120 mg Pb/kg), followed by whole fish and
carcass (whole fish minus gill, liver, muscle and intestine) Pb concentrations
(approximately 70 to 80 mg Pb/kg). However, in considering overall internal Pb body
burden, nearly 80% was largely concentrated in the bone tissue, while gill contributed
<5%.
In another study with fathead minnow, chronic (300 day) exposure to 120 (ig Pb/L
resulted in accumulation of approximately 41 mg Pb/kg tissue, although this number was
decreased from initial body burdens of greater than 104 mg Pb/kg at test initiation (Mager
et al.. 2010). Tissue distribution at 300 days was consistent with Grosell et al. (2006a)
with highest concentration in gill, followed by kidney, anterior intestine, and carcass.
Addition of humic acid and carbonate both independently reduced uptake of Pb in these
fish over the exposure time period. Interestingly, fathead minnow eggs collected daily
during 21 day breeding assays that followed the chronic exposure described above
accumulated similar levels of Pb from the test solutions regardless of Pb concentration or
water chemistry (e.g., addition of humic acid and carbonate) (Mager et al.. 2010). Direct
acute exposure from water rather than parental transfer accounted for the majority of the
Pb accumulation in eggs. Similarly, exposure offish to 32.5 (ig Pb/L in base water for
150 days resulted in fathead minnow whole body concentrations of approximately 31 mg
Pb/kg, with the most rapid accumulation rate occurring within the first 10 days of
exposure, followed by an extended period of equilibrium (Mager etal.. 2008). In this
same study, fish were tested in two additional treatments: 36.7 (ig Pb/L in hard water
(Ca2+ 500 (iM) or 38.7 (ig Pb/L in humic acid supplemented water (4 mg/L). While the
addition of humic acid significantly reduced Pb bioaccumulation in minnows (to
approximately 10.4 mg Pb/kg on a whole body basis), Ca2+ sulfate did not alter uptake.
Despite the fact that Ca2+-mediated Pb toxicity occurred in larval fathead minnow, there
was no concurrent effect on whole body Pb accumulation.
Uptake studies in other freshwater teleosts have generally followed the pattern of Pb
uptake described above for fathead minnow. In the cichlid, Nile tilapia (Oreochromis
niloticus), Pb accumulated significantly in gill (45.9 ±34.4 (ig/g dry weight at
2,070 (ig Pb/L), 57.4 ±26.1 (ig/g dry weight at 4,100 (ig Pb/L) and liver (14.3 (ig/g dry
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weight at 2,070 (ig Pb/L) and 10.2 (ig/g dry weight at 4,100 jig Pb/L) during a 14-day
nominal exposure (as Pb nitrate) (Atli and Canli. 2008). In rainbow trout exposed to
100 (ig Pb/L (as Pb acetate) for 72 hours, the accumulation in tissues was gill >kidney
>liver and this same pattern was observed in all concentrations tested
(100-10,000 ng Pb/L) (Suicmez et al.. 2006). In contrast to uptake in teleosts, in
Pb-uptake studies with the Chondrostei fish Chinese Sturgeon (Acipenser sinensis),
muscle tissue accumulated higher levels of Pb than gills (Hou et al., 2011).
Sloman et al. (2005) investigated the uptake of Pb in dominant-subordinate pairings of
rainbow trout exposed to 46 (ig/L or 325 (ig Pb/L (as Pb nitrate) for 48 hours. Significant
Pb accumulation in gill, liver, and kidney was only observed in the highest concentration.
Pb accumulated preferentially in liver of subordinate trout when compared to dominant
trout. Brown trout (Salmo truttd) exposed to aqueous Pb concentrations ranging from 15
to 46 (ig Pb/L for 24 days accumulated 6 mg Pb/kg dry weight in gill tissue and Pb
concentrations in liver tissue reached 14 mg Pb/kg dry weight. Interestingly, Pb in gill
tissue peaked on day 11 and decreased thereafter, while liver Pb concentrations increased
steadily over the exposure period, which may indicate translocation of Pb in brown trout
from gill to liver (Heier et al., 2009).
Zebrafish (Danio rerio) Pb uptake rates in gills, body (fish without gills) and whole body
were determined in fish exposed to water containing 5.2 ug 204Pb/L and sampled at 17,
22, 40 and 48h (Komjarova and Blust 2009c). This study demonstrated that zebrafish gill
tissue is the main uptake site for the metal, as Pb concentrations in these tissues were up
to eight times as high as that in other tissues. For example, the rate of Pb uptake in gill
tissue was significantly increased from 10 L/kg per hour to 35 L/kg per hour by
increasing pH from 6 to 7, and from 20 L/kg per hour to 35 L/kg per hour by increasing
Ca2+concentration from 0.1 to 0.5 mM.
The Eurasian silver crucian carp (Carassius auratus) collected from a pond containing an
average of 1,600 mg Pb/kg in the sediments exhibited increased average Pb whole body
burden of 36.5 mg Pb/kg dry weight (range 12 to 68 mg Pb/kg dry weight) (Khozhina
and Sherriff, 2008). Pb was primarily sequestered in skin, gill, and bone tissues, but was
also detected at elevated levels in muscle and liver tissues, as well as in eggs. Two fish
species (Labeo rohita and Ctenopharyngodon idella) collected from the Upper Lake of
Bhopal, India with average Pb concentration 30 (ig Pb/L in the water column contained
elevated Pb tissue concentrations (Malik et al., 2010). However, while liver and kidney
Pb concentrations were similar between the two species (1.5 and 1.1 mg Pb/kg tissue and
1.3 and 1.0 mg Pb/kg tissue for C. idella andZ. rohita, respectively), they accumulated
significantly different amounts of Pb in gill and muscle tissues. C. idella accumulated
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more than twice the Pb in these tissues (1.6 and 1.3 mg Pb/kg) than did L. rohita (0.5 and
0.4 mg Pb/kg).
The studies reviewed above generally support the conclusions of the 2006 Pb AQCD
(U.S. EPA. 2006b) that the gill is a major site of Pb uptake in fish and that there are
species-dependent differences in the rate and pattern of Pb accumulation. As indicated in
the 2006 Pb AQCD, exposure duration can be a factor in Pb uptake from water. In a
30-day exposure study, Nile tilapia fmgerlings had a three-fold increase in Pb uptake at
the gill on day 30 compared to Pb concentration in gill at day 10 and 20 (Kamaruzzaman
et al.. 2010). In addition to uptake at the gill, a time-dependent uptake of Pb into kidney
in rainbow trout exposed to 570 (ig Pb/L for 96 hours (Patel et al.. 2006) was observed.
Pb was accumulated preferentially in the posterior kidney compared to the anterior
kidney. A similar pattern was observed by Alves and Wood (2006) in a dietary exposure.
In catla (Catla catla) fmgerlings, the accumulation pattern of Pb was kidney >liver >gill
>brain >muscle in both 14 day and 60 day Pb exposures (Palaniappan et al.. 2009). In
multiple studies with fathead minnow at different exposure durations, tissue uptake
patterns were similar at 30 days (Grosell et al.. 2006a) and 300 days (Mager et al.. 2010).
In the larval minnows, Pb concentration in the intestine exhibited the highest initial
accumulation of all tissues on day 3 but then decreased for the remainder of the
experiment while concentrations in the other organs continued to increase (Grosell et al..
2006a). By day 30, gill tissue exhibited the highest Pb concentration followed by whole
fish and carcass (whole fish minus gill, liver, muscle and intestine). The most rapid rate
of Pb accumulation in this species occurs within the first 10 days of exposure (Mager et
al.. 2008). African catfish (Clarias gariepinus) exposed to nominal Pb concentrations of
50 to 1,000 (ig Pb/L (as Pb nitrate) for 4 weeks accumulated significant amounts of Pb in
heart (520-600 mg Pb/kg), liver (150-242 mg Pb/kg), and brain (120-230 mg Pb/kg)
tissues (Kudirat. 2008). Doubling the exposure time to 8 weeks increased sequestration of
Pb in these tissues as well as in skin (125-137.5 mg Pb/kg) and ovaries (30-60 mg Pb/kg).
Since the 2006 Pb AQCD, several studies have focused on dietary uptake of Pb in
teleosts. Metals have been shown to assimilate differently in tissues depending on the
exposure route (Rozon-Ramilo et al.. 2011; Meyer etal.. 2005). Alves et al. (2006)
administered a diet of three concentrations of Pb (7, 77 and 520 mg Pb/kg dry weight) to
rainbow trout for 21 days. Doses were calculated to be 0.02 (ig Pb/day (control),
3.7 (ig Pb/day (low concentration), 39.6 (ig Pb/day (intermediate concentration),and
221.5 (ig Pb/day (high concentration). Concentrations in the study were selected to
represent environmentally relevant concentrations in prey. After 21 days exposure to the
highest concentration, Pb accumulation was greatest in the intestine, followed by carcass,
kidney and liver leading the authors to hypothesize that the intestine is the primary site of
exposure in dietary uptake of Pb. All tissues, (gill, liver, kidney, intestine, carcass)
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sequestered Pb in a dose-dependent manner. The gills had the greatest concentration of
Pb on day 7 (8.0 mg Pb/kg tissue wet weight) and this accumulation decreased to
2.2 mg Pb/kg tissue wet weight by the end of the experiment suggesting that the Pb was
excreted or redistributed (Alves et al.. 2006). Furthermore, with increasing dietary
concentrations, the percentage of Pb retained in the fish decreased. Additionally, in this
study red blood cells were identified as a reservoir for dietary Pb. Plasma did not
accumulate significant Pb (0.012 mg Pb kg wet weight in the high dose), however, Pb
was elevated in blood cells (1.5 mg Pb kg wet weight in the high dose) (Alves et al..
2006).
Additional studies have supported the anterior intestine as a target for Pb in fish. Nile
tilapia exposed to dietary Pb for 60 days (105, 418, and 803 mg Pb/kg dry weight)
accumulated the greatest concentration of Pb in the intestine, followed by the stomach
and then the liver (Dai et al.. 2009a). The amount of Pb in tissue increased with
increasing dietary Pb concentration. In a 42 day chronic study of dietary uptake in
rainbow trout, fish fed 45 or 480 mg Pb/kg, accumulated Pb preferentially in anterior
intestine (Alves and Wood. 2006). Pb accumulation in the gut was followed by bone,
kidney, liver, spleen, gill, carcass, brain and white muscle (Alves and Wood. 2006). Ojo
and Wood (2007) investigated the bioavailability of ingested Pb within different
compartments of the rainbow trout gut using an in vitro gut sac technique. Although a
significant increase in Pb uptake was observed in the mid-intestines, this was determined
to be much lower than Pb uptake rates via gill surfaces. However, given that intestinal
uptake rate for Pb did not significantly differ from those derived for essential metals
(e.g., Cu, Zn, and Ni), this uptake route is likely to be significant when aqueous Pb
concentrations are low and absorption via gill surfaces is negligible (Ojo and Wood.
2007).
Following a chronic 63-day dietary exposure to Pb, male zebrafish had significantly
increased Pb body burdens, but did not exhibit any significant impairment when
compared with controls. Fish were fed diets consisting of field-collected Nereis
diversicolor oligochaetes that contained 1.7 or 33 mg Pb/kg dry weight. This resulted in a
daily Pb dose of either 0.1 or 0.4 mg Pb/kg (Boyle etal. 2010). At the end of the
exposure period, tissue from male fish reared on the high-Pb diet contained
approximately 0.6 mg Pb/kg wet weight, as compared with approximately 0.48 mg Pb/kg
wet weight in the low-Pb dietary exposure group. Pb level was elevated in female fish fed
the high-Pb diet, but not significantly so.
Ciardullo et al. (2008) examined bioaccumulation of Pb in rainbow trout tissues
following a 3-year chronic dietary exposure to the metal. Diet was determined to contain
0.19 mg Pb/kg wet weight. Fish skin accumulated the greatest Pb concentrations (0.02 to
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0.05 mg Pb/kg wet weight), followed by kidney, gills, liver, and muscle. Pb accumulation
in muscles (.005 mg Pb/kg) remained constant over all sampled growth stages (Ciardullo
et al., 2008). The authors concluded that dietary Pb was poorly absorbed by rainbow
trout. Comparison of dietary and water-borne exposures suggest that although
accumulation of Pb can occur from dietary sources, toxicity does not correlate with
dietary exposure, but does correlate with gill accumulation from waterborne exposure
(Alves et al., 2006). Comparison of uptake rates across the gut and gill have shown that
transporter pathways in the gill have a much higher affinity for Pb than do similar
pathways in the gut (Ojo and Wood. 2007).
Since the 2006 Pb AQCD, several field studies have considered Pb uptake and
bioaccumulation in fish as a tool for environmental assessment. Pb tissue concentrations
were elevated in several species offish exposed in the field to Pb from historical mining
waste, and blood Pb concentrations were highly correlated with elevated tissue
concentrations, suggesting that blood sampling may be a useful and potentially non-lethal
monitoring technique (Brumbaugh et al., 2005).
This review of the recent literature indicates that the primary and most efficient mode of
Pb absorption for freshwater fish is assimilation of labile Pb via gill surfaces; recent
research indicates that chronic dietary Pb exposure may result in some Pb
bioaccumulation although it is not the predominant route of exposure. Nevertheless, if
benthic invertebrates comprise a large portion of fish diets in chronically contaminated
systems, assimilated Pb loads may be significant. This was demonstrated by Boyle et al.
(2010). who showed that laboratory diets consisting of less than one third field-collected
Pb-contaminated invertebrates were sufficient to raise fish tissue Pb levels. However,
data from field sites suggest that fish accumulation of Pb from dietary sources is highly
variable and may be strongly dependent on the physiology of individual species and
absorption capacities.
Amphibians
Since the 2006 Pb AQCD, there are a few recent field measurements and laboratory-
based studies that consider uptake of Pb in amphibians. Whole body Pb measured in three
species of field-collected tadpoles in the Mobile-Tensaw River Delta in Alabama
averaged 1.19 mg Pb/kg dry weight in Rana clamitans, 0.65 mg Pb/kg dry weight in
Rana catesbeiana and 1.32 mg Pb/kg dry weight inHyla cinerea fAlbrecht et al.. 2007,).
Blood-Pb levels in Ozark hellbender salamanders (Cryptobranchus alleganiensis
bishopf), a candidate species for the Endangered Species Act, ranged from 0.044 to
0.055 mg/kg dry whole blood weight, in three rivers in Missouri (Huang et al.. 2010). In
the same study, Pb-blood levels were measured from Eastern hellbenders
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(Cryptobranchus alleganiensis alleganiensis), a species of concern, collected from four
rivers and ranged from 0.075 to 0.088 mg Pb/kg dry whole blood weight.
In a chronic laboratory-based study with tadpoles of the Northern Leopard frog (Rana
pipiens), Pb tissue concentrations were evaluated following exposures to 3, 10, and
100 (ig Pb/L from embryo to metamorphosis. The tadpole tissue concentrations ranged
from 0.1 to 224.5 mg Pb/kg dry mass and were positively correlated to Pb concentrations
in the water (Chen et al.. 2006b). Dose-dependent bioaccumulation of Pb was observed in
the livers of tadpoles of the African clawed frog (Xenopus laevis) exposed to nominal
concentrations ranging from 1.0 to 30,000 (ig Pb/L (3 to 115 mg Pb/kg wet weight) for
12 days (Mouchet et al.. 2007). Pb concentrations were measured in livers, bodies
without liver and whole bodies in Southern leopard frog (Rana sphenocephala) tadpoles
exposed to Pb in sediment (45 to 7,580 mg Pb/kg dry weight) with corresponding pore
water concentrations of 123 to 24,427 (ig Pb/L from embryonic stage to metamorphosis
(Sparling et al.. 2006). There was 100% mortality at 3,940 mg Pb/kg and higher. In all
body residues analyzed there was a significant positive correlation between Pb in
sediment and Pb in sediment pore water. Concentrations of Pb in liver were similar to
results with whole body and bodies without liver indicating that Pb is not preferentially
sequestered in liver.
Reptiles
Recent field surveys of Pb in water snakes since the 2006 Pb AQCD indicate that Pb is
bioaccumulated in several species. Water snakes spend time in terrestrial and aquatic
habitats and could potentially be exposed to atmospherically deposited-Pb in both
environments. Average Pb levels in whole body samples of Eastern Ribbon Snakes
(Thamnophis sauritus) collected from the Mobile-Tensaw River, a large watershed that
drains more than 75% of Alabama were 0.35 ± 0.12 mg Pb/kg dry weight) (Albrecht et
al.. 2007). Burger et al. (2007) measured Pb levels in blood, kidney, liver, muscle and
skin from water snakes, (Nerodia sepedon) collected from an urban/suburban canal in
New Jersey. Pb was highest in skin (0.467 mg Pb/kg wet weight) followed by kidney
(0.343 mg Pb/kg wet weight) blood (0.108 mg Pb/kg wet weight), muscle (0.103 mg
Pb/kg wet weight) and liver (0.063 mg Pb/kg wet weight). No interspecies differences
were observed in blood Pb (range 0.04 to 0.1 mg Pb/kg) from field-collected banded
water snakes (Nerodia fasciatd), brown water snakes (N. taxispilotd) and cottonmouth
(Agkistrodon piscivorus) from a reference area and an area contaminated by chemical and
radiation releases from the 1950's to the 1980's at the Department of Energy's Savannah
River site in South Carolina (Burger et al.. 2006). Cottonmouth and brown water snake
from the exposed site had significantly higher levels of Pb in tail muscle when compared
to the reference creek.
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Mammals
Pb bone levels in Eurasian otters (Lutra lutrd) measured in dead individuals collected in
southwest England fell by 73% between 1992 and 2004 (Chadwicket al.. 2011). Annual
mean bone Pb levels were 446 (ig Pb/kg in 1992 and 65 (ig Pb/kg in 2004. The 73%
decline of Pb in otter bones from 1992 to 2004 was found to coincide with legislative
controls on Pb emissions implemented in the U.K. starting in 1986. A positive correlation
with stream sediment Pb and bone Pb was also observed in this study. The strength of
this correlation decreased with increasing Ca2+ in streams.
6.4.4.4 Food Web
In the 2006 Pb AQCD, trophic transfer of Pb through aquatic food chains was considered
to be negligible (U.S. EPA. 2006c). Concentrations of Pb in the tissues of aquatic
organisms were found to be generally higher in algae and benthic organisms and lower in
higher trophic-level consumers indicating that Pb was bioaccumulated but not
biomagnified (U.S. EPA, 2006c; Eisler. 2000). Recent literature since the
2006 Pb AQCD, provides evidence of the potential for Pb to be transferred in aquatic
food webs. Other studies indicate Pb is decreased with increasing trophic level. This
section incorporates recent literature on transfer of Pb through freshwater aquatic food
chains including the application of stable isotope techniques to trace the accumulation
and dilution of metals through producers and consumers.
Pb was transferred through at least one trophic level in El Niagara reservoir,
Aguascalientes, Mexico, a freshwater ecosystem that lacks fishes (Rubio-Franchini et al..
2008). Pb was quantified in sediment (0.55 mg Pb/kg to 21 mg Pb/kg), water (5.8 to
39 (ig Pb/L), and zooplankton samples of this freshwater system. BAFs were calculated
for predatory and grazing zooplanktonic species. The BAF of the rotifer A. brightwellii
(BAF 49,300) was up to four times higher than the grazing cladocerans D. similis (BAF
9,022) andM micrura (BAF 8,046). According to the authors, since M. micrura are prey
for A. brightwellii this may explain the biomagnifications of Pb observed in the predatory
rotifer and provides evidence that Pb biomagnifies at intermediate trophic levels.
The relative contribution of water and food as source of trace metals including Pb was
investigated in the larvae of the alderfly Sialis velata fCroisetiere et al. 2006). Its prey,
the midge (C. riparius) was reared in the laboratory and then exposed to trace elements in
a metal-contaminated lake for one week prior to being fed to S. velata. During the one-
week exposure period of C. riparius to the contaminated water, five of six trace elements,
including Pb, reached steady state within C. riparius. Alderfly larvae were held in the lab
in uncontaminated lake water and feed one of the treated C. riparius per day for up to six
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days to measure Pb uptake via prey. A separate group of alderfly larvae were exposed
directly to the contaminated lake water for six days and fed uncontaminated C. riparius
while a third group was exposed to Pb via prey and water. Trace metal concentrations in
S. velata that consumed contaminated C. riparius increased significantly compared to
S. velata in water-only exposures. Food was concluded to be the primary source of Pb
(94%) to these organisms, not Pb in the water.
The trophic transfer of Pb from the sediment dwelling polychaete worm N. diversicolor
to the invertebrate polychaete predator Nereis virens provides additional evidence for
assimilation of Pb by a predator and the potential for further transport up the food chain
(Rainbow et al.. 2006). N. virens significantly accumulated Pb from a diet of
N. diversicolor and there was a significant inverse linear relationship between the trophic
transfer coefficient and prey Pb concentration. In the same study, another predator, the
decapod Palaemonetes varians, did not significantly accumulate Pb from N. diversicolor
indicating that trophic transfer is dependent on species-specific differences in metal
assimilation efficiencies and accumulation patterns.
In a recent dietary metal study, field-collected invertebrates representing ecologically
relevant sources of Pb were fed to zebrafish, to assess bioavailability of this metal via
food. The polychaete worm N. diversicolor was collected from two sites; an estuary
contaminated with Pb and a reference site with low metal concentrations (Boyle et al..
2010). Male zebrafish fed Pb-enriched N. diversicolor had significant increases in whole-
body Pb burden when compared to zebrafish fed prey from the reference site, brine
shrimp or flake food diets. There was a trend toward increased Pb levels in females under
the same dietary regimen. In this study, deposit feeding invertebrates were shown to
mobilize sediment-bound metals in the food chain since zebrafish were exposed only to
biologically incorporated metal.
The concentration of Pb in the tissues of various aquatic organisms was measured during
the biomonitoring of mining-impacted stream systems in Missouri. Generally, Pb
concentrations decreased with increasing trophic level: detritus contained 20 to 60 mg
Pb/kg dry weight, while periphyton and algae contained 1 to 30 mg Pb/kg dry weight;
invertebrates and fish collected from the same areas exhibited Pb tissue concentrations of
0.1 to 8 mg Pb/kg dry weight (Besser et al.. 2007). In addition, Pb concentrations in
invertebrates (snails, crayfish, and other benthos) were negatively correlated with Pb
concentrations in detritus, periphyton, and algae. Fish tissue concentrations, however,
were consistently correlated only with detritus Pb concentrations (Besser et al.. 2007).
Other studies have traced Pb in freshwater aquatic food webs and have found no evidence
of biomagnification of Pb with increasing trophic level. Watanabe et al. (2008) observed
decreasing Pb concentrations through a stream macroinvertebrate food web in Japan from
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producers to primary and secondary consumers. In a Brazilian freshwater coastal lagoon
food chain, Pb was significantly higher in invertebrates than in fishes (Pereira et al..
2010).
Introduction of exotic species into an aquatic food web may alter Pb concentrations at
higher tropic levels. In Lake Erie, the invasive round goby (Neogobius melanostomus)
and the introduced zebra mussel (Dreissena polymorphd) have created a new benthic
pathway for transfer of Pb and other metals (Southward Hogan et al.. 2007). The goby is
a predator of the benthic zebra mussel, while the endemic smallmouth bass (Micropterus
dolomieui) feed on goby. Since the introduction of goby into the lake, total Pb
concentrations have decreased in bass. The authors attribute this decrease of Pb in bass to
changes in food web structure, changes in prey contaminant burden or declines in
sediment Pb concentrations.
6.4.5 Biological Effects of Pb in Freshwater Systems
This section focuses on the studies of biological effects of Pb on freshwater algae, plants,
invertebrates, fish and other biota with an aquatic lifestage (e.g., amphibians) published
since the 2006 Pb AQCD. Key studies from the 1977 Pb AQCD, the 1986 Pb AQCD and
the 2006 Pb AQCD on biological effects of Pb are summarized where appropriate.
Waterborne Pb is highly toxic to aquatic organisms with bioavailability and subsequent
toxicity varying depending upon the species and lifestage tested, duration of exposure,
the form of Pb tested, and water quality characteristics (e.g., pH, alkalinity, DOC)
(Sections 6.4.2 and 6.4.3).
The 2006 Pb AQCD (U.S. EPA. 2006c) noted that the physiological effects of Pb in
aquatic organisms can occur at the biochemical, cellular, and tissue levels of biological
organization and include inhibition of heme formation, alterations of blood chemistry,
and decreases in enzyme levels. A review of the more recent literature corroborated these
findings, and added information about induction of oxidative stress by Pb, alterations in
chlorophyll, and changes in production and storage of carbohydrates and proteins. Recent
studies available since the 2006 Pb AQCD further consider effects of Pb on reproduction
and development, growth and survival of aquatic organisms. Alterations to these
endpoints can lead to changes at the community and ecosystem levels of biological
organization such as decreased abundance, reduced taxa richness, and shifts in species
composition (Section 6.1). Effects on reproduction, growth, and survival are reported in
additional species with some effects occurring in sensitive freshwater organisms at or
near ambient levels of Pb (Table 6-2). Because this review is focused on effects of Pb,
studies reviewed for this section include only those for which Pb was the only, or
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primary, metal to which the organism was exposed. Areas of research not addressed here
include literature related to exposure to Pb from ingestion of shot or pellets. Biological
effects of Pb on freshwater algae and plant species are considered below, followed by
information on effects on freshwater invertebrates and vertebrates. All reported values are
from exposures in which concentrations of Pb were analytically verified unless nominal
concentrations are stated.
6.4.5.1 Freshwater Plants and Algae
The toxicity of Pb to algae and plants has been recognized in earlier agency reviews of
this metal. In the 1977 Pb AQCD, differences in sensitivity to Pb among different species
of algae were observed and concentrations of Pb within the algae varied among genera
and within a genus (U.S. EPA. 1977). The 1986 Pb AQCD (U.S. EPA. 1986b) reported
that some algal species (e.g., Scenedesmus sp.) were found to exhibit physiological
changes when exposed to high Pb concentrations in situ. The observed changes included
increased numbers of vacuoles, deformations in cell organelles, and increased autolytic
activity. Effects of Pb on algae reported in the 2006 Pb AQCD included decreased
growth, deformation, and disintegration of algae cells, and blocking of the pathways that
lead to pigment synthesis, thus affecting photosynthesis. Observations in additional algal
species since the 2006 Pb AQCD, support these findings and indicate that Pb exposure is
associated with oxidative stress. All of these effects were observed at concentrations of
Pb that exceed those found currently in most surface waters (Table 6-2).
Recent studies available since the 2006 Pb AQCD, report additional mechanistic
information on Pb toxicity to freshwater macrophytes as well as further evidence for
effects on oxidative stress and growth endpoints. However, many of these studies were
conducted at nominal concentrations of Pb, complicating the comparisons to Pb
quantified in surface waters. Furthermore, their relevance to conditions encountered in
natural environments is difficult to establish since modifying factors of bioavailability,
such as DOC, are often absent from controlled exposures.
The effect of Pb exposure on the structure and function of plant photosystem II was
studied in giant duckweed, S. polyrrhiza fLing and Hong. 2009J. The Pb concentration of
extracted photosystem II particles was found to increase with increasing Pb
concentration, and increased Pb concentration was shown to decrease emission peak
intensity at 340 nm, amino acid excitation peaks at 230 nm, tyrosine residues, and
absorption intensities. This results in decreased efficiency of visible light absorption by
affected plants. The authors theorized that Pb2+ may replace either Mg2+ or Ca2+ in
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chlorophyll or the oxygen-evolving center, inhibiting photosystem II function through an
alteration of chlorophyll structure.
Pb exposure in microalgae species has been linked to several effects, including disruption
of thylakoid structure and inhibition of growth in both Scenedesmus quadricauda and
Anabaena flos-aquae fArunakumara and Zhang. 2008J. Arunakumara et al. (2008)
determined the effect of aqueous Pb on the algal species S. platensis using solutions of
Pb nitrate. Exposures at 3,440 (ig Pb/L stimulated 10-day algal growth; growth was
inhibited at higher concentrations of 6,830, 21,800, 32,800 and 44,500 (ig Pb/L by 5, 40,
49, and 78%, respectively. In addition to growth inhibition, algal chlorophyll a and b
content were significantly diminished at the three highest Pb exposures (Arunakumara et
al., 2008). Although no specific morphological abnormalities were linked to Pb exposure,
filament breakage was observed in S. platensis at Pb concentrations >50,000 (ig Pb/L.
Since the 2006 Pb AQCD, the production of reactive oxygen species following Pb
exposure has been measured directly in cells of the freshwater algae Chlamydomonas
reinhardtii at nominal concentrations of Pb as Pb nitrate (0.02 to 52 (ig Pb/L) with the
greatest response at 3.15 times more stained cells compared to the control sample
following an exposure of 2.5 hours (Szivak et al., 2009). Although this study provides
direct evidence for a mechanism of Pb-toxicity at the sub-organism level of biological
organization, the relevance of the exposure method to conditions encountered in natural
environments is unknown. The concentration data are not reliable in this case since Pb
concentrations were not quantified and the lowest reported values are below the
analytical detection limit for Pb.
At the time of the 1977 Pb AQCD, there was limited information available on Pb effects
on aquatic macrophytes. For plants in general, Pb was recognized to affect
photosynthesis, mitosis, and growth, however, the majority of studies reporting Pb
toxicity were not conducted with plants grown under field conditions (U.S. EPA. 1977).
The mechanism for Pb inhibition of photosynthesis was further elucidated in the 1986 Pb
AQCD. Additional evidence of Pb effects on plant growth was also observed, however,
the available studies were conducted under laboratory conditions at concentrations that
exceeded Pb levels in the environment except near smelters or roadsides (U.S. EPA.
1986b). In the 1986 Pb AQCD, EC50 values for plant growth were available for several
aquatic plants with the lowest EC50 of 1,100 (ig Pb/L in Azolla pinnata exposed to
Pb nitrate for 4 days. Effects of Pb on metabolic processes in aquatic plants reviewed in
the 2006 Pb AQCD (U.S. EPA. 2006b) included nitrate uptake, nitrogen fixation,
ammonium uptake and carbon fixation at concentrations of 20,000 (ig Pb/L and higher.
New information is available on Pb effects on oxidative stress endpoints such as changes
in antioxidant enzymes, lipid peroxidation, and reduced glutathione in aquatic plant,
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algae, and moss species exposed to Pb; however most evidence is from studies with high
concentrations where Pb was not quantified in the exposure media. An aquatic moss,
F. antipyretica, exhibited increased SOD and ascorbate levels following a 2-day exposure
to nominal Pb chloride solutions of concentrations of 20, 200, 2,070, 20,700 and
207,200 (ig Pb/L. When exposure duration was increased to 7 days, only SOD activity
remained significantly increased by Pb exposure (Dazy et al.. 2009). Bell-shaped
concentration-response curves were commonly observed for the induction of antioxidant
enzymes in F. antipyretica. The chlorophyll, carotenoid, and protein contents of the
aquatic macrophyte Elodea canadensis were significantly reduced following Pb
accumulation at nominal exposures of 1,000 10,000 and 100,000 (ig Pb/L (Dogan et al..
2009). This, along with the induction of some antioxidant systems and the reduction of
growth at the highest two exposures, indicated that exposure to the metal caused
significant stress, and that toxicity increased with exposure. In addition, native
Myriophyllum quitense exhibited elevated antioxidant enzyme activity (glutathione-S-
transferase, glutathione reductase, and peroxidase) following transplantation in
anthropogenically polluted areas containing elevated Pb concentrations. These were
correlated with sediment Pb concentrations in the range of 5 to 23 mg Pb/g dry weight
(Nimptsch et al., 2005).
Since the 2006 Pb AQCD, toxicity and oxidative stress were also observed in coontail
(C. demersum) rooted aquatic macrophytes following 7-day nominal exposures to
aqueous Pb 200 to 20,700 (ig Pb/L, with increasing effects observed with greater
exposure concentrations and times. Chlorosis and leaf fragmentation were evident
following a 7-day exposure to the highest concentration, while induction of antioxidant
enzymes (glutathione, superoxide dismutase, peroxidases, and catalase) was observed at
lower exposure concentrations and times. However, as the duration and concentration of
Pb exposure was increased, activities of these antioxidant enzymes decreased (Mishra et
al.. 2006b).
Sobrino et al. (2010) observed reductions in soluble starch stores and proteins with
subsequent increases in free sugars and amino acids in Lemna gibba plants exposed
nominally to Pb (50,000 to 300,000 (ig Pb/L); total phenols also increased with
increasing Pb exposure. Authors noted that this species exhibited similar responses under
extreme temperatures, drought, and disease. According to Odjegba and Fasidi (2006).
nominal exposure to 18,600 (ig Pb/L as Pb nitrate for 21 days was sufficient to induce a
gradual reduction of both chlorophyll and protein content in the macrophyte Eichhornia
crassipes.
Following 72-hour aqueous exposure to 8,495 (ig Pb/L as Pb nitrate, phytochelatin and
glutathione concentrations in the freshwater algae Scenedesmus vacuolatus were
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significantly increased over that of non-exposed algal cultures (Le Faucheur et al.. 2006).
The 72-hour Pb exposure also significantly reduced S. vacuolatus growth, and of all the
metals tested (Cu, Zn, Ni, Pb, Ag, As, and Sb), Pb was determined to be the most toxic to
the algae species. In the algae Chlamydomonas reinhardtii, phytochelatin concentrations
were lower than intracellular Pb and not sufficient to bind to accumulated metal
following 72-hour exposure (Scheidegger et al.. 2011).
In addition to oxidative stress responses, there is new information since the
2006 Pb AQCD on growth effects observed at high concentrations of Pb summarized in
Table 6-5. Growth effects at the species level can lead to effects at the population-level of
biological organization and higher (Section 6.1.1). Root elongation was significantly
reduced in a number of wetland plant species (Beckmannia syzigachne, Juncus effusus,
Oenanthe javanica, Cyperusflabelliformis, Cyperus malaccensis, and Neyraudia
reynaudiana) following nominal Pb exposures of 20,000 (ig Pb/L as Pb nitrate for 21
days (Deng et al.. 2009). Lemna sp. aquatic plants were determined to effectively
sequester aqueous Pb at nominal exposures of 5,000 and 10,000 (ig Pb/L in a 7-day
experiment, however, 15,000 (ig Pb/L resulted in plant mortality (Hurd and Sternberg.
2008). In another study with duckweed, Paczkowska et al. (2007) observed that nominal
Pb exposures of 2,070 to 20,700 (ig Pb/L for 9 days stimulated the growth of Lemna
minor cultures, although there was concurrent evidence of chlorosis and induction of
antioxidant enzymes. Additionally, Cd was found to be more toxic than Pb, although the
authors determined that this resulted from poor uptake of Pb by L. minor (Paczkowska et
al.. 2007). Pb exposure (as Pb nitrate) caused oxidative damage, growth inhibition, and
decreased biochemical parameters, including photosynthetic pigments, proteins, and
monosaccharides, in Wolffia arrhiza plants. Fresh weight of plants was reduced following
both 7- and 14-day exposures to Pb concentrations greater than 2,120 (ig Pb/L while
chlorophyll a content was decreased at 210 (ig Pb/L and higher (Piotrowska et al.. 2010).
Effects of Pb on algae reported in the 2006 Pb AQCD (U.S. EPA. 2006b) included
decreased growth, deformation and disintegration of algae cells, and blocking of the
pathways that lead to pigment synthesis, thus affecting photosynthesis. Observations in
additional algal species since the 2006 Pb AQCD support these findings. Effects on
plants supported by additional evidence in this review and evidence from previous
reviews include oxidative damage, decreased photosynthesis and reduced growth.
Elevated levels of antioxidant enzymes are commonly observed in aquatic plant, algae,
and moss species exposed to Pb. All of the observed effects on aquatic macrophytes and
algae occur at concentrations not typically encountered in surface waters of the U.S.
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6.4.5.2 Freshwater Invertebrates
Few studies on biological effects of Pb in freshwater invertebrates had been conducted at
the time of the 1977 Pb AQCD. One study reported an effect on reproduction in Daphnia
magna at 30 ng Pb/L (U.S. EPA. 1977). In the 1986 Pb AQCD (U.S. EPA. 1986b).
increased mortality was observed in the freshwater snail Lymnaea palustris as low as
19 (ig Pb/L and reproductive impairment was reported as low as 27 (ig Pb/L for
Daphnia sp. Population-level endpoints of Pb reviewed in the 2006 Pb AQCD included
reproduction, growth, and survival. Pb was recognized to be more toxic in longer-term
exposures than shorter-term exposures with chronic toxicity thresholds for reproduction
in water fleas (D. magna) ranging as low as 30 (ig Pb/L. In aquatic invertebrates, Pb has
also been shown to affect stress responses and osmoregulation (U.S. EPA. 2006c). Recent
evidence that supports previous findings of Pb effects on reproduction and growth in
invertebrates is reviewed here as well as limited studies on behavioral effects associated
with Pb exposure. Some of these effects are observed in the range of Pb values found in
surveys of U.S. surface waters (median 0.50 (ig Pb/L, range 0.04 to 30 (ig Pb/L), in the
U.S. based on a synthesis of NAWQA data reported in the previous 2006 Pb AQCD
(U.S. EPA. 2006c) (Table 6-2). The studies are generally presented in this section from
responses at the sub-organismal level of biological organization to consideration of
endpoints re levant to ecological risk assessment (growth, reproduction, survival).
Recent literature strengthens the evidence indicating that Pb affects enzymes and
antioxidant activity in aquatic invertebrates. These alterations at the sub-organismal level
may serve as biomarkers for effects at the organism level and higher. In invertebrate
species that have hemoglobin, ALAD activity can be measured as a biomarker for Pb
exposure. In the freshwater gastropod B. glabrata and the freshwater oligochaete
Lumbriculus variegatus a significant negative correlation between whole body tissue
ALAD enzyme activity and increasing Pb was observed following 48-hour exposure to
varying nominal concentrations of the metal (Aisemberg et al., 2005). The concentration
at which 50% of enzyme inhibition was measured was much lower in B. glabrata (23 to
29 (ig Pb/L) than in L. variegatus (703 (ig Pb/L). A significant negative correlation was
also observed between ALAD activity and metal accumulation by the organisms. Sodium
and potassium ATPase (Na+/K+ATPase) activity in gills of Eastern elliptic mussels was
significantly reduced following a 28-day exposure to 57 (ig Pb/L and 245 (ig Pb/L
(Mosher et al. 2012). A significant reduction in Na+ and significant increase in Ca2+ in
hemolymph was only observed at the highest concentration.
Studies of stress responses to Pb in invertebrates, conducted since the 2006 Pb AQCD,
include induction of heat shock proteins and depletion of glycogen reserves. Although
these stress responses are correlated with Pb exposure, they are non-specific and may be
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altered with exposure to any number of environmental stressors. Induction of heat shock
proteins in zebra mussel exposed to an average concentration of 574 (ig Pb/L for 10
weeks exhibited a 12-fold higher induction rate as compared to control groups (Singer et
al.. 2005). Energetic reserves in the freshwater snail B. glabrata in the form of glycogen
levels were significantly decreased by 20%, 57% and 78% in gonads compared to control
animals following 96-hour exposures to nominal concentrations of 50, 100 and
500 (ig Pb/L, respectively (Ansaldo et al.. 2006). Decreases in glycogen levels were also
observed in the pulmonary and digestive gland region at 50 and 100 (ig Pb/L treatment
levels. Pb did not exacerbate the effects of sustained hypoxia in the crayfish (C.
destructor) exposed to 5,000 (ig Pb/L for 14 days while being subjected to decreasing
oxygen levels in water (Morris et al.. 2005). The crayfish appeared to cope with Pb by
lowering metabolic rates in the presence of the metal.
The effect of Pb on osmoregulatory response has been studied since the 2006 Pb AQCD.
The combined effect of Pb and hyperosmotic stress on cell volume regulation was
analyzed in vivo and in vitro in the freshwater red crab, Dilocarcinus pagei (Amado et
al.. 2006). Crabs held in either freshwater or brackish water lost 10% of their body weight
after one day when exposed to 2,700 (ig Pb2+/L as Pb nitrate. This weight loss was
transient and was not observed during days 2-10 of the exposure. In vitro, muscle from
red crabs exposed to hyperosmotic saline solution had increased ninhydrin-positive
substances and muscle weight decreased in isosmotic conditions upon exposure to Pb
indicating that this metal affects tissue volume regulation in crabs although the exact
mechanism is unknown.
Behavioral responses of aquatic invertebrates to Pb reviewed in the 2006 Pb AQCD (U.S.
EPA. 2006b) included avoidance. A limited number of recent studies have considered
additional behavioral endpoints. Feeding rate of the blackworm L. variegatus was
significantly suppressed by day 6 of a 10 day sublethal test in Pb-spiked sediments
(Penttinen et al.. 2008) as compared to feeding rates at the start of the experiment.
However, this decrease of approximately 50% of the initial feeding rate was also
observed in the controls; therefore it is likely caused by some other factor other than Pb
exposure. Aqueous soil leachates containing multiple metals, including Pb, had no effect
on D. magna mobility. Authors noted that although some concentrations (13 to
686 (ig Pb/L) exceeded Canadian Environmental Quality Guidelines, no significant
correlation could be established between Pb exposure and D. magna mobility; in fact, the
cladocerans were more sensitive to Fe and Al in the leachate than to Pb (Chapman et al..
2010).
Alterations in reproductive and developmental endpoints at the species level can lead to
effects at the population-level of biological organization and higher (Section 6.1.1). For
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example, reduced fecundity may lead to a decreased population size and developmental
defects can compromise the ability of an organism to escape predation. Recent evidence
of reproductive and developmental effects of Pb on freshwater invertebrates available
since the 2006 Pb AQCD, include data from previously untested species as well as
further characterization of reproductive effects in commonly tested organisms such as
Daphnia sp (Table 6-5). However, many of these studies are conducted at nominal Pb
concentration complicating direct comparison to Pb quantified in freshwater
environments. Sublethal concentrations of Pb negatively affected the total number of
eggs, hatching success and embryonic survival of the freshwater snail B. glabrata
exposed to nominal concentrations of 50, 100, or 500 (ig Pb/L as Pb nitrate (Ansaldo et
al.. 2009). Following exposure of adult snails for 96 hours, adults were removed and the
eggs were left in the Pb solutions. The total number of eggs was significantly reduced at
the highest concentration tested (500 (ig Pb/L). Time to hatching was doubled and
embryonic survival was significantly decreased at 50 and 100 (ig Pb/L, while no embryos
survived in the highest concentration. Theegala et al. (2007) observed that the rate of
reproduction was significantly impaired in Daphnia pulex at >500 (ig Pb/L in 21-day
exposures at nominal concentrations of Pb. In a 21-day reproductive test in D. magna the
number of neonates born per female was significantly reduced at nominal concentrations
of 25, 250, and 2,500 (ig Pb/L (Ha and Choi. 2009). C. dubia reproduction was also
impacted by a seven-day exposure to 50 to 500 (ig Pb/L. Both DOC, and, to a lesser
degree, alkalinity were observed to ameliorate the effects of Pb on C. dubia reproduction.
As DOC increased from 100 (imol C/L to 400 and 600 (imol C/L, the calculated mean
EC50 values for C. dubia reproduction increased from approximately 25 (ig Pb/L to
200 (ig Pb/L and greater than 500 (ig Pb/L, respectively (Mageretal.. 2011 a).
Reproductive variables including average lifespan, rate of reproduction, generation time
and rate of population increase were adversely affected in the rotifer Brachionus patulus
under conditions of increasing turbidity and Pb concentration (Garcia-Garcia et al..
2007).
In larvae of the mosquito, Culex quinquefasciatus, exposed to 50 (ig Pb/L, 100 (ig Pb/L
or 200 (ig Pb/L (as Pb nitrate), exposure was found to significantly reduce hatching rate
and egg-production at all concentrations and larval emergence rate at 200 (ig Pb/L
(Kitvatanachai et al.. 2005). Larval emergence rates of 78% (FO), 86% (Fl) and 86% (F2)
were observed in the control group while emergence rates decreased in each generation
46% (FO), 26% (Fl) and 58% (F2) in mosquitoes reared in a concentration of
200 (ig Pb/L. The time to first emergence also increased slightly to 10 days in the
Pb-exposed group as compared to the control group where emergence was first observed
on day 9. In the F2 generation of parents exposed to 200 (ig Pb/L, the ratio of female to
male offspring was 3.6:1.0. No effects were observed on oviposition preference of adult
females, larval weight, or larval deformation.
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Impacts to growth can lead to effects at the population-level of biological organization
and higher (Section 6.1.1). As noted in the 2006 Pb AQCD (U.S. EPA. 2006b). Pb
exposure negatively affects the growth of aquatic invertebrates. Some studies reviewed in
the previous Pb AQCD suggested that juveniles do not discriminate between the uptake
of essential and non-essential metals (Arai et al., 2002). In recent literature (summarized
in Table 6-5). the freshwater pulmonate snail Lymnaea stagnalis has been identified as a
species that is extremely sensitive to Pb exposure. Growth of juveniles was inhibited at
EC20 <4 (ig Pb/L. (Grosell and Brix. 2009; Grosell et al.. 2006b). In L. stagnalis exposed
to 18.9 (ig Pb/L for 21 days, Ca2+ influx was significantly inhibited and model estimates
indicated 83% reduction in growth of newly hatched snails after 30 days at this exposure
concentration (Grosell and Brix. 2009). The authors speculate that the high Ca2+demand
of juvenile L. stagnalis for shell formation and interference of the Ca2+uptake pathway by
Pb result in the sensitivity of this species.
In a study of the combined effects of temperature (22 °C or 32 °C), nominal Pb
concentration (50, 100 and 200 (ig Pb/L as Pb chloride) and presence of a competitor, the
population growth rate of two freshwater rotifer species, Brachionus havanaensis and
B. mbens, as measured by quantifying the number of live rotifers for 15 days, responded
to presence of stressors (Montufar-Melendez et al.. 2007). At the lowest temperature,
B. rubens suppressed population growth of B. havanaensis at 50 (ig Pb/L and higher and
B. rubens population growth did not increase at any Pb concentration at 32 °C, a
temperature more suited for B. havanaensis. In situ toxicity testing with the woodland
crayfish (Orconectes hylas) indicated that crayfish survival and biomass were
significantly lower in streams impacted by Pb mining and that concentrations of Pb and
other metals in water, detritus, macroinvertebrates, fish and crayfish were significantly
higher at mining sites (Allert et al., 2009a).
Although Pb is known to cause mortality when invertebrates are exposed at sufficiently
high concentrations, species that are tolerant of Pb may not exhibit significant mortality
even at high concentrations of Pb. Odonates are highly tolerant of Pb with no significant
differences in survival of dragonfly larvae Pachydiplax longipennis and Erythemis
simplicicollis exposed for 7 days to nominal concentrations of Pb as high as
185,000 (ig Pb/L (Tollett et al.. 2009). This apparent tolerance to Pb may be even more
pronounced in natural environments where the presence of multiple modifying factors
(e.g., pH, alkalinity, hardness, DOC) influences Pb bioavailability. Other species are
more sensitive to Pb in the environment and these responses are reviewed in
Section 6.4.6.
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6.4.5.3 Freshwater Vertebrates
The 1977 Pb AQCD reported on Pb effects to domestic animals, wildlife and aquatic
vertebrates. The available Pb studies were from exposure to Pb via accidental poisoning
or ingestion of Pb shot (U.S. EPA. 1977). Studies on aquatic vertebrates reviewed in the
1986 Pb AQCD were limited to hematological, neurological and developmental
responses in fish (U.S. EPA. 1986b). In the 2006 Pb AQCD, effects on freshwater
vertebrates included recent data for fish specifically considering the effects of water
quality parameters on toxicity, as well as limited information on sensitivity of turtles and
aquatic stages of frogs to Pb (U.S. EPA. 2006c). Biological effects of Pb on freshwater
fish that have been studied since the 2006 Pb AQCD are reviewed here, and limited
recent evidence of Pb effects on amphibians are considered. This section presents recent
information available on the mechanism of Pb as a neurotoxicant in fish and effects of
this metal on blood chemistry. Additional mechanisms of Pb toxicity have been
elucidated in the gill and the renal system offish since the 2006 Pb AQCD. Further
supporting evidence of reproductive effects of Pb on fish is discussed along with limited
new information on behavioral effects of Pb.
Freshwater Fish
Evidence of toxicity of Pb and other metals to freshwater fish goes back to early
observations whereby contamination of natural areas by Pb mining lead to extirpation of
fish from streams (U.S. EPA. 1977). At the time of the 1977 Pb AQCD, documented
effects of Pb on fish included anemia, mucous secretion, functional damage to inner
organs, physical deformities and growth inhibition. Additionally, the role of temperature,
pH, hardness and other water quality parameters on Pb toxicity was discussed in the 1977
Pb AQCD. The 1986 Pb AQCD Q986b) reported that hematological and neurological
responses were the most commonly observed effects in fish and the lowest exposure
concentration causing either hematological or neurological effects was 8 (ig Pb/L. These
findings were additionally supported in the 2006 Pb AQCD, where observed effects of Pb
on fish included inhibition of heme formation, alterations in brain receptors, effects on
blood chemistry, and decreases in some enzyme activities (U.S. EPA. 2006c). Functional
responses resulting from Pb exposure included increased production of mucus, changes in
growth patterns, and gill binding affinities. According to Eisler (2000) and reviewed in
the 2006 Pb AQCD, the general symptoms of Pb toxicity in fish include production of
excess mucus, lordosis, anemia, darkening of the dorsal tail region, degeneration of the
caudal fin, destruction of spinal neurons, ALAD inhibition, growth inhibition, renal
pathology, reproductive effects, growth inhibition and mortality.
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Evidence of Pb effects on fish available since the 2006 Pb AQCD generally supports the
findings in previous Pb reviews and further elucidates the mechanisms of Pb-associated
toxicity on some physiological responses. At the sub-organism level, new information on
Pb effects on DNA, specific enzymes, ionoregulation and other biochemical responses is
presented followed by a discussion of new information on population-level endpoints
(i.e., growth reproduction summarized in Table 6-5).
Since the 2006 Pb AQCD evidence of direct interaction of Pb with fish DNA has become
available as well as additional studies on the genotoxic effects of Pb exposure to fish.
Hong et al. (2007a) observed covalent binding of Pb with kidney DNA from silver
crucian carp (Carassius auratus gibelio) though extended X-ray absorption fine structure
spectroscopy. This study suggests that exposure to Pb results in effects to DNA but the
exposure method (in vitro) makes it difficult to estimate the natural environmental
conditions that would be equivalent to the experimental one. In the freshwater fish
Prochilodus lineatus, blood, liver, and gill cells were sampled from fish treated with
nominal concentration of 5,000 (ig Pb/L as Pb nitrate for 6, 24 and 96-hours and then
DNA damage was assessed by comet assay (Monteiro et al.. 2011). DNA breaks were
observed in all cell types after 96-hour exposure. The concentrations used in this study
were high compared to Pb concentrations currently encountered in freshwater (Table
6-2), however, it presents supporting evidence for a possible mechanism of Pb toxicity to
fish.
Upregulation of antioxidant enzymes in fish is a well-recognized response to Pb
exposure. Since the last review, additional studies demonstrating antioxidant activity as
well as evidence for production of reactive oxygen species following Pb exposure are
available. Silver crucian carp injected with nominal concentration of 10, 20 or 30 mg
Pb/kg wet weight Pb chloride showed a significant increase in the rate of production of
superoxide ion and hydrogen peroxide in liver (Ling and Hong. 2010). In the same fish,
activities of liver SOD, catalase, ascorbate peroxidase, and glutathione peroxidase were
significantly inhibited. Both glutathione and ascorbic acid levels decreased and
malondialdehyde content increased with increasing Pb dosage, suggesting that lipid
peroxidation was occurring and the liver was depleting antioxidants. Although this
exposure pathway is unlikely to be relevant for air related deposition of Pb, it provides
evidence for the mechanism of toxicity (production of reactive oxygen species) and the
responses of antioxidant enzymes observed in this study are supported by findings in
studies from fish from nominal water-only exposures. For example, in the freshwater fish
Nile tilapia, liver catalase, liver alkaline phosphatase, Na+/K+ATPase, and muscle
Ca2+ATPase activities were quantified in various tissues following a 14-day exposure to
nominal concentrations (1,000, 2,000 and 4,000 (ig Pb/L) of Pb nitrate (Atli and Canli.
2007). Liver catalase activity significantly increased in the 1,000 and 4,000 (ig Pb/L
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concentrations while liver alkaline phosphatase activity was significantly increased only
at the 4,000 (ig Pb/L concentration. No significant change in alkaline phosphatase
activity was observed in intestine or serum. Ca2+ATPase activity was significantly
decreased in muscle. Na+/K+ATPase was elevated in gill in the highest concentration of
Pb while all concentrations resulted in significant decreases of this enzyme in intestine.
Serum alanine aminotransferase and aspartate aminotransferase activities were elevated
in Nile tilapia exposed to 50 (ig Pb/L in 4 and 21 day aqueous exposures while elevations
in alkaline phosphatase and lactate dehydrogenase were only observed at 21 days (Frratet
al.. 2011). In another study with Nile tilapia, Pb had no effect on glutathione measured in
liver, gill, intestine, muscle and blood and liver metallothionein levels following a 14-day
exposure to 1,000, 2,000 and 4,000 (ig Pb/L concentrations of Pb as Pb nitrate (Atli and
Canli. 2008).
Metabolic enzyme activity in teleosts has also been measured following dietary
exposures. Alves and Wood (2006) in a 42-day chronic dietary Pb study with 45 and
480 mg Pb/kg found that gill Na+/K+ATPase activity was not affected in rainbow trout
while increased Na+/K+ATPase was observed in the anterior intestine. Metabolic
activities measured in liver and kidney of Nile tilapia following 60-day dietary
administration of 100, 400, and 800 mg Pb/kg indicated that alanine transaminase,
aspartate transaminase, and lactate dehydrogenase activities significantly decreased in
kidney in a concentration-dependent manner (Dai et al.. 2009b) and increased in liver
with increasing concentration of dietary Pb. In a subsequent study using the same
exposure paradigm, the digestive enzymes amylase, trypsin and lipase in tilapia were
inhibited by dietary Pb in a concentration-dependent manner (Dai et al.. 2009a). Lesions
were also evident in histological sections from livers of Pb-exposed fish from this study
and included irregular hepatocytes, cell hypertrophy, and vacuolation although no
quantification of lesions by dose-group was presented.
There is also evidence for Pb exposure leading to changes in hepatic CYP450 content
although relevance of these in vitro and injection studies to air related exposures to Pb is
unknown. Pb was shown to inhibit hepatic cytochrome P450 in vitro in carp (C. carpio),
silver carp (Hypothalmichtys molitrix) and wels catfish (Silurus glanis) in a
concentration-dependent manner from 0 to 4 (ig/mL (Pb2+) (Henczova et al.. 2008). The
concentrations of Pb that resulted in 50% inhibition of EROD and 7-ethoxycoumarin-o-
deethylase (ECOD) isoenzymes varied with the fish species. Silver carp was the least
sensitive to the inhibitory effects of Pb (EROD 1.21, ECOD 1.52 jig Pb/mL) while carp
EROD activity was inhibited at 0.76 (ig Pb/mL. Interaction of Pb with cytochrome P450
was verified by spectral changes using Fourier Transform Infrared (FTIR) spectroscopy.
In the same study, CYP450 content was elevated and EROD isoenzyme activities were
decreased in vivo in silver carp for two days following an injection of 2 mg Pb/kg as
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Pb acetate and returned to control values by 6 days. Liver damage to African catfish
exposed to nominal concentrations of Pb (50-1,000 (ig Pb/L) for 4 or 8 weeks included
hepatic vacuolar degeneration followed by necrosis of hepatocytes (Adeyemo. 2008b).
The severity of observed histopathological effects in the liver was proportional to the
duration of exposure and concentration of Pb.
In environmental assessments of metal-impacted habitats, ALAD is a recognized
biomarker of Pb exposure (U.S. EPA. 2006c). For example, lower ALAD activity has
been significantly correlated with elevated blood Pb concentrations in wild caught fish
from Pb-Zn mining areas although there are differences in species sensitivity (Schmitt et
al.. 2007b: Schmitt et al.. 2005). Suppression of ALAD activity in brown trout
transplanted to a metal contaminated stream was linked to Pb accumulation on gills and
in liver in a 23-day exposure (Heier et al.. 2009). Alves Costa et al. (2007) observed
inhibition of ALAD in hepatocytes of the neotropical traira (Hoplias malabaricus)
following dietary dosing of 21 mg Pb/kg every 5 days for 70 days. Cytoskeletal and
cytoplasmic disorganization were observed in histopathological examination of affected
hepatocytes. In fathead minnow exposed to Pb in either control water (33 (ig Pb/L),
CaSO4 (37 (ig Pb/L) or (39 (ig Pb/L) humic acid-supplemented water for 30 days and
subsequently analyzed by quantitative PCR analysis there were no significant changes in
ALAD mRNA gene response leading the authors to speculate that water chemistry alone
does not influence this gene response (Mageretal.. 2008). In the same study, glucose-6-
phosphate dehydrogenase, glutathione-S-transferase and ferritin were upregulated, in
microarray analysis; however, no changes in whole body ion concentrations were
observed (Mager et al.. 2008).
In fish, changes in blood chemistry associated with Pb exposure were noted in the
2006 Pb AQCD (U.S. EPA. 2006b). however, only limited recent studies consider effects
on blood parameters. In a 70-day feeding study with traira exposed to dietary doses
(21 mg Pb/kg as Pb nitrate via prey [Astyanax sp.]) each five days (corresponding to
daily nominal doses of approximately 4 mg Pb/kg), there were no significant changes to
leukocytes or hemoglobin concentration and volume (Oliveira Ribeiro et al.. 2006).
Significant differences in area, elongation, and roundness of erythrocytes were observed
in the Pb-exposed individuals using light microscopy image analysis. Other studies
available since the 2006 Pb AQCD have only shown effects on blood chemistry at high
aqueous concentrations of Pb that are not representative of Pb concentrations in U.S.
surface waters. For example, in the African catfish packed cell volume decreased with
increasing nominal concentration of Pb (25,000 to 200,000 (ig Pb/L as Pb nitrate) and
platelet counts increased in a 96-hour exposure (Adevemo. 2007). Red blood cell counts
also decreased in some of the treatments when compared to controls, although the
response was not dose-dependent and so may not have been caused by Pb exposure.
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Disruption of ionoregulation is one of the major modes of action of Pb toxicity. The gill
has long been recognized as a target of Pb in teleosts. Acute Pb toxicity at the fish gill
primarily involves disruption of Ca2+ homeostasis as previously characterized in the
2006 Pb AQCD (Rogers and Wood. 2004: Rogers and Wood. 2003). In addition to this
mechanism, Pb was found to induce ionoregulatory toxicity at the gill of rainbow trout
through a binding of Pb with Na+/K+ATPase and rapid inhibition of carbonic anhydrase
activity thus enabling noncompetitive inhibition of Na+ and Cl" influx (Rogers et al..
2005). Alves et al. (2006) administered a diet of three concentrations of Pb (7, 77 and
520 mg Pb/kg dry weight) to rainbow trout for 21 days, and measured physiological
parameters including Na+ and Ca2+ influx rate from water. Dietary Pb had no effect on
brachial Na+ and Ca2+ rates except on day 8 where Na+ influx rates were significantly
elevated. These studies suggest that Pb is intermediate between purely Ca2+ antagonists
such as Zn2+ and Cd2+ and disrupters of Na+ and Cl" balance such as Ag+ and Cu2+. This
finding has implications for BLM modeling since it suggests that both Ca2+ and Na+ need
to be considered as protective cations for Pb toxicity. Indeed, protection from Pb toxicity
by both Na+ and Ca2+ has been documented in freshwater fish (Komjarova and Blust
2009b).
Additional experiments conducted since the 2006 Pb AQCD provide supporting evidence
for underlying mechanisms of Pb toxicity. It was previously established that long-term
exposures of Pb can impact gill structure and function. Histopathological observations of
gill tissue in the catfish (C. gariepinus) following an 8-week aqueous exposure to
nominal concentrations of Pb nitrate revealed focal areas of epithelial hyperplasia and
necrosis at the lower exposure concentrations (50 (ig Pb/L and 100 (ig Pb/L) (Adeyemo.
2008a). Hyperplasia of mucous cells and epithelial cells were apparent in the tissue from
fish exposed the highest concentrations of Pb in the study (500 (ig Pb/L and
1,000 (ig Pb/L). In vitro incubation of gill tissue from fathead minnow with Pb
concentrations of 2,500, 12,500 and 25,000 (ig Pb/L for 60 minutes decreased the ratio of
reduced glutathione to oxidized glutathione, indicating that lipid peroxidation at the gill
likely contributes to Pb toxicity at low water hardness (Spokas et al., 2006). It is difficult
to extrapolate these observations to natural environments due to the methods used for
exposure and the use of nominal exposure concentrations.
In addition to recent evidence of Pb interruption of Na+ and Cl" at the gill (Rogers et al.,
2005). Pb can interfere with the ionoregulation of Na+ and Cl" and tubular reabsorption of
Ca2+, Mg2+, glucose, and water in the teleost kidney (Patel et al.. 2006). Renal parameters
including urine flow rate, glomerular filtration rate, urine pH, and ammonia excretion
were monitored in a 96-hour exposure of rainbow trout to analytically verified
concentration of 1,200 (ig Pb/L as Pb nitrate. Rates of Na+ and Cl" excretion decreased by
30% by 48 hours while Mg excretion increased two-to-three fold by 96 hours. Urine flow
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rate was not altered by Pb exposure, although urinary Pb excretion rate was significantly
increased. After 24 hours of Pb exposure, the urine excretion rate of Ca2+ increased
significantly by approximately 43% and remained elevated above the excretion rate in the
control group for the duration of the exposure. Glomerular filtration rate significantly
decreased only during the last 12 hours of the exposure. Ammonia excretion rate
increased significantly at 48 hours as urine pH correspondingly decreased. At the end of
the experiment glucose excretion was significantly greater in Pb-exposed fish. Although
the exposures in this study approached the 96-hour LC50, nephrotoxic effects of Pb
indicate the need to consider additional binding sites for this metal in the development of
biotic ligand modeling (Patel et al., 2006). Additional evidence for Pb effects on ion
levels were observed in serum of Nile tilapia; Na+ and Cl" were decreased and K+ levels
were elevated following a 21 day nominal exposure to 50 (ig Pb/L as Pb nitrate (Frratet
Neurological responses offish to Pb exposure were reported in the 1986 Pb AQCD (U.S.
EPA, 1986b). Additional evidence of the neurotoxic effects of Pb on teleosts has become
available since the 2006 Pb AQCD. The mitogen-activated protein kinases (MAPK),
extracellular signal -regulated kinase (ERK)l/2 and p38MAPK were identified for the first
time as possible molecular targets for Pb neurotoxicity in ateleost (Leal et al.. 2006). The
phosphorylation of ERK1/2 and p38MAPK by Pb was determined in vitro and in vivo in the
catfish (Rhamdia queleri). R. quelen exposed to a nominal concentration of 1,000 (ig Pb/L
(as Pb acetate) for two days showed a significant increase in phosphorylation of ERK1/2
and p38MAPK in the nervous system. Incubation of cerebellar slices for 3 hours in 1,035
and 2,070 (ig Pb/L as Pb acetate also showed significant phosphorylation of MAPKs. The
observed effects of Pb on the MAPK family of signaling proteins have implications for
control of brain development, apoptosis and stress response. In the neotropical fish traira,
muscle cholinesterase was significantly inhibited after 14 dietary doses of 21 mg Pb/kg
wet weight (Rabitto et al., 2005). Histopathological observations of brains of African
catfish exposed to nominal concentrations of 500 (ig Pb/L or 1,000 (ig Pb/L Pb as
Pb nitrate for 4 weeks included perivascular edema, focal areas of malacia, and diffuse
areas of neuronal degeneration (Adevemo. 2008b). As in the observed effects of Pb on
gill function and ionoregulation, it is difficult to assess the significance of these findings
to fish in natural environments due to the methods used for exposure.
Evidence from the 2006 Pb AQCD (U.S. EPA. 2006b) and earlier Pb reviews indicate
that Pb can impair both cognitive and motor function in fish. Reduced locomotion and
foraging ability were observed in Chinese sturgeon juveniles exhibiting abnormal body
curvature following nominal exposure to either 800 or 1,600 (ig Pb/L for 1 12 days (Hou
et al.. 2011). Actual concentrations of Pb were quantified at the end of the 1 12-day
exposure period (30 to 50% of test media was renewed daily): 129 (ig Pb/L (200 (ig Pb/L,
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nominal), 458 (ig Pb/L (800 (ig Pb/L nominal), and 1,276 ng Pb/L (1,600 ng Pb/L,
nominal). These chondrostean fish gradually recovered from deformities during a
depuration period and were able to swim and forage effectively 6 weeks after transfer
into clean water
Since the 2006 Pb AQCD, several studies integrating behavioral and physiological
measures of Pb toxicity have been conducted on fish. Some of these observations are
reported to occur at concentrations of Pb reported in freshwater. Zebrafish embryos
exposed nominally to low concentrations of Pb as Pb chloride (2.0 and 6.0 (ig Pb/L
prepared from serial dilutions of a stock solution) until 24 hours post-fertilization and
then subsequently tested as larvae or adult fish exhibited behavioral disruptions in
response to mechanosensory and visual stimuli (Rice et al., 2011). Although Pb was not
measured in the water, Pb uptake in the embryos was quantified during the first 24 hours
post-fertilization (approximately 0.08 nM/100 embryos at 2.0 (ig Pb/L and 0.32 nM
Pb/100 embryos at 6.0 (ig Pb/L). Startle response time in larvae measured as maximum
head turn velocity and escape time decreased in a concentration-dependent pattern
following a directional, mechanical stimulus (tapping). The pattern of escape swimming
was altered in larvae of Pb-exposed embryos compared to controls. In the adult fish
hatched from Pb-exposed embryos (6.0 (ig Pb/L), visual response to a rotating black bar
against a white background (ability to detect contrast) was significantly degraded. These
findings provide evidence for behavioral effects of Pb at concentrations lower than
previously reported in fish (U.S. EPA. 2006c). however, aqueous exposure
concentrations were not analytically verified.
Sloman et al. (2005) investigated the effect of Pb on hierarchical social interactions and
the corresponding monoaminergic profiles in rainbow trout. Trout were allowed to
establish dominant-subordinate relationships for 24 hours, and then were exposed to
46 (ig Pb/L or 325 (ig Pb/L (Pb nitrate) for 48 hours to assess effects on behavior and
brain monoamines. In non-exposed fish, subordinate individuals had higher
concentrations of circulating plasma cortisol and telencephalic 5-hydroxyindoleacetic
acid/5-hydroxytryptamine (serotonin) (5-HIAA/5-HT) ratios. In the high concentration of
Pb, there was significant uptake of Pb into gill, kidney and liver when compared with the
control group and dominant fish appeared to have elevated hypothalamic 5-HIAA/5HT
ratios. Uptake of Pb into the liver was higher in subordinate fish when compared to the
dominant fish. No significant differences were observed in cortisol levels or behavior
after metal exposure.
Mager et al. (2010) conducted prey capture assays with 10 day old fathead minnow
larvae born from adult fish exposed to 120 (ig Pb/L for 300 days, then subsequently
tested in a breeding assay for 21 days. The time interval between 1st and 5th ingestion of
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10 prey items (Artemia nauplif) was used as a measure of behavior and motor function of
offspring of Pb-exposed fish. Larvae were offered 10 Artemia and the number ingested
within 5 minutes was scored. The number of larvae ingesting 5 Artemia decreased within
the time period in offspring of Pb-exposed fish as compared to the control group, leading
the authors to suggest this behavior is indicative of motor/behavioral impairment. In
another study with fathead minnows, swimming performance measured as critical aerobic
swim speed was significantly impaired in minnows in 24-hour acute (139 (ig Pb/L) and
chronic 33 to 57 day (143 (ig Pb/L) exposures, however, no significant difference in
swim speed was observed in chronic exposures to 33 (ig Pb/L (Mager and Grosell. 2011).
Alterations in reproductive and developmental endpoints at the species level can lead to
effects at the population-level of biological organization and higher (Section 6.1.1). For
example, reduced fecundity may lead to a decreased population size and developmental
effects may decrease the ability of a fish to escape predators or reduce spawning
mobility. Reproductive and developmental effects of Pb in fish have been reported for
several decades. In the 1977 Pb AQCD, second generation brook trout (Salvelinus
fontinalis) exposed to 235 or 474 (ig Pb/L were shown to develop severe spinal
deformities (lordoscoliosis) (U.S. EPA. 1977). Pb concentration of 120 (ig Pb/L produced
spinal curvature in rainbow trout (Oncorhynchus mykiss) and spinal curvatures were
observed in developing eggs of killifish as reviewed in the 1986 Pb AQCD (U.S. EPA.
1986b). Recent studies on reproductive effects of Pb in fish from oocyte formation to
spawning are summarized in Table 6-5.
Reproductive performance of zebrafish as measured by incidence of spawning, numbers
of eggs per breeding pair or hatch rate of embryos was unaffected following a 63 day diet
of field-collected Pb-contaminated polychaetes that were representative of a daily dose of
0.3-0.48 mg Pb/kg per day (dry weight diet/wet weight fish) through food (Boyle et al.
2010). Mager et al. (2010) conducted 21-day breeding exposures at the end of chronic
300 day toxicity testing with fathead minnow. Non-exposed breeders were switched to
water containing Pb and Pb-exposed breeders were moved to control tanks and effects on
egg hatchability and embryo Pb accumulation were assessed. Fish in the high Pb
concentration with HCO3" (113 (ig Pb/L) and DOC (112 (ig Pb/L) and the low Pb
concentration with HCO3" (31 (ig Pb/L) reduced total reproductive output, while a
significant increase in average egg mass was observed in the high Pb HCO3" and DOC
treatments as compared to egg mass size in controls and in low HCO3" and DOC
treatments with Pb. No significant differences were present between treatments in egg
hatchability.
The effects of metals on embryonic stage offish development in C. carpio and other
species were reviewed in Jezierska et al. (2009) and included developmental
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abnormalities during organogenesis as well as embryonic and larval malformations. The
authors concluded that the initial period of embryonic development, just after
fertilization, and the period of hatching are the times at which developing embryos are
most sensitive to metals. Additional nominal exposure studies provide supporting
evidence for embryo malformations associated with Pb-exposure. A significant
concentration-dependent increase in morphological malformations was observed in
African catfish embryos exposed to nominal concentrations of 100 (ig Pb/L, 300 (ig Pb/L
or 500 (ig Pb/L Pb nitrate from 6 hours post-fertilization to 168 hours post-fertilization
(Osman et al., 2007b). Hatching was delayed with increasing Pb concentration and hatch
success of the embryos decreased from 75% in the controls to 40% in the group exposed
to 500 (ig Pb/L. Chinese sturgeon exposed to nominal concentrations of 200 (ig Pb/L,
800 (ig Pb/L or 1,600 (ig Pb/L for 112 days (96 hour post-fertilized eggs through juvenile
stages) exhibited body curvatures in the two highest concentrations (Hou et al.. 2011).
During a 42 day depuration period in clean water following exposure, the degree of
curvature in affected individuals decreased with decreasing tissue concentrations of Pb.
Reproductive and endocrine effects of Pb have also been reported at the cellular level in
fish, including alterations in gonadal tissue and hormone secretions that are associated
with Pb-exposure, however, recent studies that report these effects are limited to
experiments where only nominal concentrations of Pb were tested. Histopathological
observations of ovarian tissue in the African catfish following an 8-week aqueous
exposure to Pb nitrate indicated necrosis of ovarian follicles at the lowest concentration
tested (50 (ig Pb/L) (Adeyemo. 2008a). Severe degeneration of ovarian follicles was
observed in the highest concentrations of 500 (ig Pb/L and 1,000 (ig Pb/L. Chaube et al.
(2010) considered the effects of Pb on steroid levels through 12 and 24 hour in vitro
exposures of post-vitellogenic ovaries from the catfish (Heteropneustes fossilis) to
nominal concentrations of Pb as Pb nitrate (0, 10, 100, 1,000, 3,000 and 10,000 (ig Pb/L).
Progesterone, 17-hydroxyprogesterone, 17, 20 beta-dihydroxyprogesterone,
corticosterone, 21-deoxycortisol and deoxycorticosterone were inhibited in a dose-
dependent manner. Pb was stimulatory on the steroids estradiol-17-(3, testosterone and
cortisol at low concentrations, and inhibitory at higher concentrations. The authors
propose that the disruption of steroid production and altered hormone secretion patterns
observed at the lower concentrations of Pb in this study are suggestive of the potential for
impacts to fish reproduction (Chaube et al.. 2010).
There is also evidence for alterations in steroid levels associated with Pb exposure in
other species offish although these studies were all conducted with nominal
concentrations of Pb and the actual exposure concentrations were not verified. Carp
(Cyprinus carpio) exposed for 35 days to nominal concentration of 410 (ig Pb/L
experienced altered plasma cortisol and prolactin levels. Plasma cortisol levels
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significantly increased throughout the study period while plasma prolactin increased up
to day 14 and then declined and was not significantly different from controls by the end
of the experiment (Ramesh et al., 2009). Cortisol levels were significantly decreased in
Nile tilapia exposed to 50 (ig Pb/L (nominal) for 4 days but were followed by a return to
control levels at 21 days of exposure (Firatetal. 2011). In a comparative study between
in vitro and in vivo estrogenic activity of Pb, vitellogenin was reported to be significantly
induced in juvenile goldfish (Carassius auratus) following 96-hour exposure to nominal
concentration of 0.2 and 0.02 (ig Pb/L when compared to control fish (Isidori et al..
2010). In the same study, estrogenicity of Pb was detected in vitro using a proliferation
assay with estrogen receptor-positive human MCF-7 cells. The estrogenic effects of Pb
reported by the authors were observed at concentrations at or below that of Pb typically
encountered in freshwaters, however, actual concentrations of Pb were not measured and
the reported concentrations were at or below analytical detection limits for Pb. The
observations of effects of Pb on vitellogenin are interesting; however, additional studies
are warranted considering the difficulty in maintaining these low concentrations of Pb.
The relevance of the observed in vitro activity to air related exposure to Pb in natural
environments is unknown.
Reduction of growth in fish was noted as an effect of Pb exposure in the 2006 Pb AQCD.
Recent studies available since the 2006 Pb AQCD do not present consistent evidence of
growth reduction in fish associated with Pb (Table 6-5). In a series of exposures in which
Ca2+, DOC and pH were varied to assess effects on Pb toxicity to fathead minnows,
Grosell et al. (2006a) observed a significant increase in growth in some groups exposed
to higher concentrations, however, the increase in body mass was noted to have occurred
in tanks with high mortality earlier in the exposure (Grosell et al.. 2006a). Fathead
minnows exposed to 33 (ig Pb/L to test swimming performance had significantly greater
body length and body mass compared to control fish following a mean Pb exposure
duration of 41 days (range 33 to 57 days) (Mager and Grosell. 2011). In 30 day chronic
tests in which a range of pH values (6.4, 7.5 and 8.3) were tested with low
(25-32 ng Pb/L), intermediate (82-156 jig Pb/L) and high (297-453 ng Pb/L)
concentrations of Pb, Mager et al. (20 lib) did not observe growth impairment in fathead
minnows at environmentally relevant concentrations of Pb. However, two 60-day early
lifestage tests with rainbow trout showed differences in LOEC for reduced growth
(Mebane et al., 2008). In the first test, a 69-day exposure, the LOECs for mortality and
reduced growth were the same (54 (ig Pb/L). In the second test, a 62-day exposure of Pb
to rainbow trout, the LOEC for fish length was 18 (ig Pb/L with an EC2o of >87 (ig Pb/L.
No effects on growth were observed in recently conducted feeding studies with fish.
Growth and survival were not significantly affected in juvenile rainbow trout, fathead
minnow and channel catfish (Ictalurus punctatus) fed a live diet of L. variegatus
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contaminated with Pb (846-1,000 (ig Pb/L per g dry mass for 30 days). (Erickson et al.,
2010). No effects on growth rates were observed in rainbow trout administered a diet
containing three concentrations of Pb (7, 77 and 520 mg Pb/kg dry weight) for 21 days
(Alves et al.. 2006) or in Nile tilapia fed diets with nominal concentration of 100, 400, or
800 mg/kg Pb dry weight for 60 days (Dai et al.. 2009b).
In one recent field study, faster growth rates were associated with lower whole-body trace
element concentrations in salmon (Salmo salar) across several streams in New
Hampshire and Massachusetts, U.S., regardless of whether accumulation was from prey
items or from water (Ward etal. 2010). In sites where conditions in the streams were
conducive to rapid salmon growth, Pb concentrations were 86% lower than in streams
where salmon were smaller.
Amphibians
Amphibians move between terrestrial and aquatic habitats and can therefore be exposed
to Pb both on land and in water. The studies reviewed here are all aquatic or sediment
exposures. Biological effects of Pb on amphibians in terrestrial exposure scenarios are
reviewed in Sections 6.3.3.3 and 6.3.4.3. Amphibians lay their eggs in or around water
making them susceptible to water-borne Pb during swimming, breeding and
development. In the 2006 Pb AQCD amphibians were considered to be relatively tolerant
to Pb. Observed responses to Pb exposure included decreased enzyme activity
(e.g., ALAD reduction) and changes in behavior summarized in Table AX7-2.4.3 (U.S.
EPA. 2006c). Since the 2006 Pb AQCD, studies conducted within two orders of
magnitude of the range of published Pb concentrations for surface waters and sediments
of the U.S. (Section 6.2.3) have indicated sublethal effects on tadpole endpoints including
growth, deformity, and swimming ability. Genotoxic and enzymatic effects of Pb
following chronic exposures have been assessed in laboratory bioassays; however, these
studies were limited to nominal exposures.
The genotoxic potential of Pb to larvae of the frog (X. laevis) was assessed by
determining the number of micronucleated erythrocytes per thousand (MNE) following a
12-day exposure to nominal concentrations of Pb as Pb nitrate (Mouchet et al.. 2007).
The lowest Pb concentrations withX laevis (10 and 100 (ig Pb/L) did not exhibit
genotoxic effects while both 1,000 and 10,000 (ig Pb/L significantly increased MNE to
14 and 202, respectively compared to the control (6 MNE). In another chronic genotoxic
study, erythrocytic micronuclei and erythrocytic nuclear abnormalities were significantly
increased with increasing Pb concentrations (700 (ig Pb/L, 1,400 (ig Pb/L,
14,000 (ig Pb/L, 70,000 (ig Pb/L) during 45, 60, and 75-day exposures of tadpoles Bufo
raddei (Zhang et al., 2007^). The authors noted that the erythrocytic micronuclei and
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erythrocytic nuclear abnormalities frequencies generally decreased with increasing
exposure time and that this may be indicative of regulation of genotoxic factors by
tadpoles.
Endpoints of oxidative damage were measured in testes of the black-spotted frog (Rana
nigromaculatd) treated with nominal concentrations of 100 (ig Pb/L, 200 (ig Pb/L,
400 (ig Pb/L, 800 (ig Pb/L or 1,600 ng Pb/L Pb nitrate by epidermal absorption for 30
days (Wang and Jia. 2009). All doses significantly increased MDA, a product of
oxidative stress, and glutathione levels were elevated in all but the lowest treatment
group. In the same study, damage to DNA assessed by DNA tail length showed effects at
>200 (ig Pb/L and DNA tail movement showed effects at >400 jig Pb/L. The authors
concluded that the effects on endpoints of oxidative stress and DNA damage detected in
testes indicated a possible reproductive effect of Pb to black-spotted frogs. The exposure
method and use of nominal concentration in this study make it difficult to determine the
relevance of this study to exposure scenarios under natural environmental conditions.
Various sublethal endpoints (growth, deformity, swimming ability, metamorphosis) were
evaluated in northern leopard frog (R. pipiens) tadpoles exposed to nominal
concentrations of 3, 10, and 100 (ig Pb/L as Pb nitrate from embryonic stage to
metamorphosis (Chen et al., 2006b). In this chronic study, the concentrations represent
the range of Pb found in surface freshwaters across the U.S. The lowest concentration of
3 (ig Pb/L approaches the EPA chronic criterion for Pb of 2.5 (ig Pb/L at a hardness of
100 mg/L or 4.5 (ig Pb/L at a hardness of 170 mg/L (U.S. EPA. 2002b). No effects were
observed in the lowest concentration. In the 100 (ig Pb/L treatment, tadpole growth rate
was slower (Gosner stages 25-30), 92% of tadpoles had lateral spinal curvature
(compared with 6% in the control) and maximum swimming speed was significantly
slower than the other treatment groups. In this study, Pb concentrations in the tissues of
tadpoles were quantified and the authors reported that they were within the range of
reported tissue concentrations from wild-caught populations.
The effects of Pb-contaminated sediment on early growth and development were assessed
in the southern leopard frog (Sparling et al., 2006). Tadpoles exposed to Pb in sediment
(45, 75, 180, 540, 2,360, 3,940, 5,520, and 7,580 mg Pb/kg dry weight) with
corresponding sediment pore water concentrations of 123, 227, 589, 1,833, 8,121, 13,579,
19,038 and 24,427 (ig Pb/L from embryonic stage to metamorphosis exhibited sublethal
responses to Pb in sediment at levels below 3,940 mg Pb/kg. There was 100% mortality
in the 3,940, 5,520 and 7,580 mg Pb/kg exposures by day 5. The authors noted that the
most profound effects of Pb on the tadpoles were on skeletal development. At 75 mg
Pb/kg, subtle effects on skeletal formation such as clinomely and brachydactyly were
observed. Skeletal malformations increased in severity at 540 mg Pb/kg and included
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clinodactyly, brachymely and spinal curvature and these effects persisted after
metamorphosis. At the highest concentration with surviving tadpoles (2,360 mg Pb/kg)
all individuals displayed severe skeletal malformations that impacted mobility. Other
sublethal effects of Pb observed in this study were reduced rates of early growth of
tadpoles at concentrations < 540 mg Pb/kg and increased time to metamorphosis in the
2,360mg Pb/kg (8,121(ig Pb/L sediment pore water) treatment.
Birds
As reviewed in Koivula and Eeva (2010) measurement of enzymes associated with
oxidative stress in birds is a well-established biomarker of exposure to metals, however,
little is known about the effects of this stress response in wild populations or at higher
levels of ecological organization. Changes in ALAD activity and other oxidative stress
biomarkers at low levels of Pb exposure were recently documented in mallards and coots
(Fulica atrd) from a lagoon in Spain impacted by Pb shot (Martinez-Haro et al., 2011).
ALAD ratio in mallards decreased linearly with blood Pb levels between 6 and
40 (ig Pb/dL, and at Pb levels of <20 (ig Pb/dL effects on several antioxidant enzymes
were observed in coots. Although the primary route of exposure to the birds was via
ingestion of Pb shot, effects were observed lower than 20 (ig Pb/dL, the background level
frequently applied to Pb exposures in birds (Martinez-Haro et al.. 2011; Brown et al..
2006).
Consideration of toxicity of Pb to vertebrate embryos that develop surrounded by a
protective egg shell has been expanded since the 2006 Pb AQCD. Pb treatment of
mallard duck (Anas platyrhynchos), eggs by immersion in an analytically verified
concentration of 100 (ig Pb/L for 30 minutes on day 0 of development did not increase
malformations or mortality of embryos (Kertesz and Fancsi. 2003). However, immersion
of eggs in 2,900 (ig Pb/L under the same experimental conditions resulted in increased
rate of mortality and significant malformations including hemorrhages of the body,
stunted growth, and absence of yolk sac circulatory system (Kertesz et al., 2006). The
second study was conducted to emulate environmental levels of Pb following a dam
failure in Hungary.
6.4.6 Exposure and Response of Freshwater Species
Evidence regarding exposure-response relationships and potential thresholds for Pb
effects on aquatic populations can inform determination of standard levels that are
protective of aquatic ecosystems. The Annex of the 2006 Pb AQCD (U.S. EPA. 2006c)
summarized data on exposure-response functions for invertebrates (Table AX7-2.4.1) and
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fish (Table AX7-2.4.2). The recent exposure-response studies in this section expand on
the findings from the 2006 Pb AQCD with information on newly-tested organisms
(including microalgae, invertebrate, amphibian and fish species) (Table 6-5 and Figure
6-5). Overall, new data for freshwater invertebrates generally support the previous
finding of sensitivity of juvenile lifestages and indicates some effects of Pb observed in
some species at concentrations of Pb reported in freshwater environments. (Table 6-2).
All reported values are from exposures in which concentrations of Pb were analytically
verified unless nominal concentrations are stated.
The aquatic macrophyte Lemna minor (duckweed) exhibited a EC50 for growth inhibition
of 6,800 (ig Pb/L in a 4-day exposure and 5,500 (ig Pb/L for a 7-day exposure to a range
of Pb concentrations from 100 to 9,970 (ig Pb/L (Dirilgen, 2011). Growth (measured as
biomass) was slightly increased at 100 and 200 (ig Pb/L and then decreased in subsequent
concentrations. In an assay using nominal concentrations of Pb the aquatic freshwater
microalgae Scenedesmus obliquus was significantly more sensitive to Pb exposure than
Chlorella vulgaris algae, although these authors stated that both appeared to be very
tolerant of the heavy metal. Laboratory 48-hour standard toxicity tests were performed
with both of these species and respective EC50 values of 4,040 and 24,500 (ig Pb/L for
growth as measured by cell division rate were derived (Atici et al.. 2008).
Exposure-response data for freshwater invertebrates provide evidence for effects of Pb at
concentrations of Pb encountered in U.S. surface waters. In the 2006 Pb AQCD, effects
of Pb-exposure in amphipods (H. aztecd) and water fleas (D. magnd) were reported at
concentrations as low as 0.45 (ig Pb/L. Effective concentrations for aquatic invertebrates
were found to range from 0.45 to 8,000 (ig Pb/L. Since the 2006 Pb AQCD, recent
studies have identified the freshwater snail L. stagnalis as a species that is extremely
sensitive to Pb exposure (Grosell and Brix. 2009; Grosell et al.. 2006b). Growth of
juvenile L. stagnalis was inhibited below the lowest concentration tested resulting in an
EC20 of <4 (ig Pb/L. In the same study, the NOEC was 12 jig Pb/L and the LOEC was
16 (ig Pb/L. In contrast, freshwater juvenile ramshorn snails M cornuarietis were less
sensitive to Pb with the same LOEC for hatching rate and LC50, calculated to be about
10,000 (ig Pb/L based on nominal exposure data (Sawasdee and Kohler. 2010).
Additional studies on Pb effects in aquatic invertebrates published since the
2006 Pb AQCD provide further evidence for differences in sensitivity of different
lifestages of aquatic organisms to Pb. In the freshwater mussel, L. siliquoidea (fatmucket)
a Pb concentration response was observed in which newly transformed (5-day-old)
juveniles were the most sensitive lifestage in a 96-hour toxicity test when compared to
acute and chronic results with other lifestages (Wang et al.. 2010f). The 96-hour EC50
values for the 5-day-old L. siliquoidea in two separate toxicity tests were 142 and
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298 (ig Pb/L (mean EC50 220 (ig Pb/L) in contrast to older juveniles (2 months old) with
an EC50 >426 (ig/L. The 24-hour median effect concentration for glochidia (larvae) of
L. siliquoidea in 48-hour acute toxicity tests was >299 (ig/L. A 28-day exposure chronic
value of 10 (ig Pb/L was obtained from 2-month-old L. siliquoidea juveniles, and was the
lowest genus mean chronic value ever reported for Pb (Wang etal.. 2010f). A 96-hour
test on newly transformed juveniles was also conducted on Lampsilis rafinesqueana
(Neosho mucket), a mussel that is a candidate for the endangered species list. The EC50
for this species was 188 (ig Pb/L.
Different lifestages of chironomids have been shown to have varying sensitivity to Pb
exposure in several studies available since the 2006 Pb AQCD. The acute toxicity of Pb
to first-instar C. riparius larvae was tested in soft water, with hardness of 8 mg/L as
CaCO3 (Bechard et al.. 2008). The 24-hour LC50 of 610 ng Pb/L for first instar
C. riparius larvae was much lower than previous values reported for later instars in
harder water. In a chronic test with Chironomus tentans, (8 day-old larvae exposed to Pb
until emergence [approximately 27 days]), the NOEC was 109, and the LOEC was
497 (ig Pb/L (Grosell et al.. 2006b). The EC2o for reduced growth and emergence of the
midge Chironomus dilutus was 28 (ig Pb/L, observed in a 55-day exposure, while the
same species had a 96-hour LC50 of 3,323 (ig Pb/L (Mebane et al.. 2008). In fourth
instars of the freshwater midge larvae Chironomus javanus the 24, 48, 72 and 96 hour
LC50 values were 20,490, 6,530, 1,690 and 720 (ig Pb/L, respectively (Shuhaimi-Othman
et al.. 20lie). This was comparable to the 96-hour LC50 (400 (ig Pb/L) in the midge
larvae Culicoides furens (Vedamanikam and Shazilli. 2008a). In the same study, the 96-
hour LC50 for Chironomusplumosus ranged from 8,300 (ig Pb/L to 16,210 (ig Pb/L
under different temperatures indicating the role of environmental factors in modulation of
toxicity and differences in sensitivity to Pb even among related species.
Cladocerans are commonly tested aquatic organisms, with data from three species:
D. magna, D. pulex and Ceriodaphnia dubia, representing approximately 70% of
available metal toxicological literature on this group (Wong et al., 2009). Recent studies
have been conducted with C. dubia and acute toxicity values for other cladocerans as
well as sublethal endpoints for D. magna are available. In a series of 48 hour acute
toxicity tests with C. dubia conducted in a variety of natural waters across North
America, LC50 values ranged from 29 to 1,180 (ig Pb/L and were correlated with DOC
(Esbaugh et al.. 2011). Median lethal concentrations forMoina micrura (LC50
690 (ig Pb/L), Diaphanosoma birgei (LC50 3,160 (ig Pb/L), andAlona rectangular (LC50
7,060 (ig Pb/L) indicate differences in sensitivity to Pb in these freshwater cladocerans
from Mexico (Garcia-Garcia et al., 2006). Several additional studies available since the
2006 Pb AQCD report exposure response values for Daphnia that are based an nominal
data: an acute study of Pb with D. pulex identified a 48-hour LC50 of 4,000 (ig/L for this
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species (Theegala et al., 2007) and the EC50 for swimming inhibition in neonate
D. magna exposed to Pb nitrate for 24 hours was 18,153 (ig Pb/L (Ha and Choi. 2009).
Rotifers are among the most sensitive aquatic genera to Pb with wide variation in LC50
values reported between species (Perez-Legaspi and Rico-Martinez. 2001). For example,
in the rotifer genus Lecane, a 22-fold difference in LC50 values was observed in 48-hour
exposure to Pb between L. hamata, L. luna and L. quadridentata. (Perez-Legaspi and
Rico-Martinez. 2001). L. luna was most sensitive to Pb toxicity with a 48-hour LC50 of
140 (ig Pb/L. In a 48-hour toxicity test with the rotifer Brachionus calyciflorus, an NOEC
(194 ng Pb/L), an LOEC (284 jig Pb/L), and an EC20 of 125 ng Pb/L was established for
this species (Grosell et al.. 2006b). The freshwater rotifer Euchlanis dilatata 48 hour
LC50 was 35 (ig Pb/L using neonates hatched from asexual eggs (Arias-Almeida and
Rico-Martinez. 2011). In this study the authors estimated the NOEC to be 0.1 (ig Pb/L
and the LOEC to be 0.5 (ig Pb/L. In contrast, for rotifer Brachionus patulus neonates, the
24-hour LC50 was 6,150 (ig Pb/L, however, this value was based on nominal exposures
(Garcia-Garcia et al., 2007).
Exposure-response assays on other freshwater species have been conducted since the
2006 Pb AQCD. The 24-hour LC50 for larvae of C. quinquefasciatus mosquitoes was
180 ng Pb/L (Kitvatanachai et al.. 2005). A 48-hour LC50 of 5,200 ng Pb/L was observed
in water-only exposures of the blackworm Lumbriculus variegatus fPenttinen et al..
2008,). In the mayfly Baetis tricaudatus, the 96-hour LC50 was 664 (ig Pb/L (Mebane et
al.. 2008). An EC2o value of 66 (ig Pb/L was derived for B. tricaudatus by quantifying
the reduction in the number of molts over a 10-day exposure to Pb (Mebane et al., 2008).
The number of molts was significantly less than the control (average of 14 molts over 10
days) at concentrations of 160 (ig Pb/L and higher with the lowest number of molts
(average of 5.3 molts over 10 days) observed in the highest concentration (546 (ig Pb/L).
In the freshwater ostracod Stenocypris major, the 96-hour LC50 was 526 (ig Pb/L
(Shuhaimi-Othman et al.. 201 Ib). In another freshwater crustacean, the prawn
Macrobrachium lancesteri, the 96-hour LC50 was 35 (ig Pb/L in soft water (<75 mg/L as
CaCO3) (Shuhaimi-Othman et al.. 201 la).
In the studies reviewed for the 2006 Pb AQCD, freshwater fish demonstrated adverse
effects at concentrations ranging from 10 to >5,400 (ig Pb/L, generally depending on
water quality parameters (e.g., pH, hardness, salinity) (U.S. EPA. 2006c). Pb tended to be
more toxic in longer-term exposures and correlated to Pb-uptake in tissues. Table
AX7-2.4.2 of the 2006 Pb AQCD summarizes effects of Pb to fish. A series of studies
published since the 2006 Pb AQCD have been conducted and have further elucidated the
influence of water chemistry parameters on Pb uptake and toxicity in fathead minnow
resulting in additional dose-response data for this species. Grosell et al. (2006b)
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conducted a series of 30-day exposures with larval fathead minnow in which varying
concentrations of Ca2+ (as CaSO4) and DOC were tested. The effects of reduced pH (6.7)
and increased pH (8.1) compared to a control pH of 7.4 on Pb toxicity were also assessed
in this study. DOC, CaSO4 and pH influenced Pb toxicity considerably over the range of
water parameters tested. For example, The 30-day LC50 for low hardness (19 mg
CaCO3/L) in basic test water (base water) was 39 (ig dissolved Pb/L and the highest LC50
value (obtained from the protection from increased concentrations of DOC and CaSO4)
was 1,903 (ig dissolved Pb/L (Grosell et al.. 2006a). This range in LC50 values for larval
fathead minnows for differing concentrations of DOC and hardness clearly demonstrates
the importance of the chemistry of the exposure medium to Pb toxicity.
Mager et al. (2010) conducted 300-day chronic toxicity tests at 35 and 120 (ig Pb/L with
fathead minnow under conditions of varied DOC and alkalinity to assess the effects of
these water quality parameters on fish growth and Pb-uptake. In additional tests with
fathead minnow, Mager et al. (20lib) conducted both 96-hour acute and 30-day chronic
tests to further characterize Ca2+, DOC, pH, and alkalinity values on Pb toxicity.
Increased Ca2+, DOC and NaHCO3 concentration afforded protection to minnows in
acute studies. The role of pH in Pb toxicity is complex and likely involves Pb speciation
and competitive interaction of FT with Pb2+ (Mager et al.. 201 Ib). In a series of 96-hour
acute toxicity tests with fathead minnow conducted in a variety of natural waters across
North America, LC50 values ranged from 41 to 3,598 (ig Pb/L and no Pb toxicity
occurred in three highly alkaline waters (Esbaugh et al.. 2011).
In the 2006 Pb AQCD, fish lifestage was recognized as an important variable in
determining the sensitivity of these organisms to Pb. Recent data available since the
2006 Pb AQCD (U.S. EPA. 2006c) support the findings of increased sensitivity of
juvenile fish to Pb when compared to adults. Acute (96-hour) and chronic (60-day) early-
lifestage test exposures were conducted with rainbow trout to develop acute-chronic
ratios (ACR's) for this species (Mebane et al.. 2008). Two early-lifestage chronic tests
were conducted, the first with an exposure range of 12-384 (ig Pb/L (69 days) at 20 mg
CaCO3/L water hardness and the second with an exposure range of 8 to 124 (ig Pb/L
(62 days) and a water hardness of 29 mg CaCO3/L. In the 69-day test, the following
chronic values were observed for survival: NOEC=24 (ig Pb/L, maximum acceptable
toxicant concentration (MATC)=36 (ig Pb/L, EC10=26 (ig Pb/L, EC20=34 (ig Pb/L, and
LC50=55 (ig Pb/L. Results from the 62-day test, with fish length as the endpoint, were
NOEC=8 (ig Pb/L, MATC=12 (ig Pb/L, EC10=7(ig Pb/L, EC20=102 (ig Pb/L and
LC50=120 (ig Pb/L. In acute tests with rainbow trout run concurrently with the chronic
tests, 96-hour LC50 values were 120 and 150 (ig Pb/L, respectively. Data from this study
resulted in ACR's for trout lower than previously reported. The low ACR values were
due to the acute tests which produced LC50 values that were 10 to 25 times lower than
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earlier studies with trout (Mebane et al., 2008). The authors speculated that the lower
LC50 values were due to the age of the fish used in the study (two to four week old fry)
and that testing with larger and older fish may not be protective of more sensitive
lifestages.
There have been only a few recent exposure-response studies in amphibians since the
2006 Pb AQCD. Southern leopard frog tadpoles exposed to Pb in sediment (45 to
7,580 mg Pb/kg dry weight) with corresponding sediment pore water concentrations from
123 to 24,427 (ig Pb/L from embryonic stage to metamorphosis exhibited concentration-
dependent effects on survival (Sparling et al.. 2006). The LC50 value for Pb in sediment
was 3,738 mg Pb/kg, which corresponds to 12,539 (ig Pb/L in sediment pore water. In the
same study, concentration-dependent effects on skeletal development were observed. The
40 day-EC50 for deformed spinal columns in the tadpoles was 1,958 mg Pb/kg
(corresponding to 6,734 (ig Pb/L sediment pore water) and the 60 day-EC50 was 579 mg
Pb/kg (corresponding to 1,968 (ig Pb/L sediment pore water) (Sparling et al.. 2006).
6.4.7 Freshwater Community and Ecosystem Effects
As discussed in the 1986 Pb AQCD (U.S. EPA. 1986b) and the 2006 Pb AQCD (U.S.
EPA, 2006b). exposure to Pb is likely to have impacts in aquatic environments via effects
at several levels of ecological organization (organisms, populations, communities, or
ecosystems). These effects resulting from toxicity of Pb would be evidenced by changes
in species composition and richness, in ecosystem function, and in energy flow. The
2006 Pb AQCD concluded that, in general, there was insufficient information available
for single materials in controlled studies to permit evaluation of specific impacts on
higher levels of organization (beyond the organism). Furthermore, Pb rarely occurs as a
sole contaminant in natural systems making the effects of Pb difficult to ascertain. New
information on effects of Pb at the population, community, and ecosystem level is
reviewed below.
In laboratory studies reviewed in the 2006 Pb AQCD and in more recent studies, Pb
exposure has been demonstrated to alter predator-prey interactions, as well as feeding and
avoidance behaviors. In aquatic ecosystems there are field studies reviewed in the 1977
Pb AQCD (U.S. EPA. 1977). the 1986 Pb AQCD (U.S. EPA. 1986b). the
2006 Pb AQCD (U.S. EPA. 2006b) and more recent studies that report reductions of
species abundance, richness or diversity particularly in benthic macroinvertebrate
communities coexisting with other metals where the sources of Pb were from mining or
urban effluents. Additionally, field studies have linked Pb contamination to reduced
primary productivity and respiration, and to altered energy flow and nutrient cycling.
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However, because of the complexity inherent in defining such effects, there are relatively
few available population, community, or ecosystem level studies that conclusively relate
Pb exposure to aquatic ecosystem effects. In addition, most of the available work is
related to point-source Pb contamination, with very few studies considering the effects of
diffuse Pb pollution. Both plant species and habitat type were determined to be factors
affecting the rate of Pb accumulation from contaminated sediments. While the rooted
aquatic plant E. canadensis was observed to accumulate the highest concentrations of Pb,
the authors concluded that submerged macrophytes (versus emergent plants) as a group
were the most likely to accumulate Pb and other heavy metals (Kurilenko and
Osmolovskava. 2006). This would suggest that certain types of aquatic plants, such as
rooted and submerged species, may be more susceptible to aerially-deposited Pb
contamination, resulting in shifts in plant community composition as a result of Pb
pollution.
Alteration of macrophyte community composition was demonstrated in the presence of
elevated surface water Pb concentrations at three lake sites impacted by mining effluents
(Mishra et al.. 2008). A total of 11 species of macrophytes were collected. Two sites
located 500 meters and 1,500 meters downstream from the mine discharge point (study
sites 2 and 3) exhibited similar dissolved Pb concentrations (78 to 92 (ig Pb/L, depending
on season) and contained six and eight unique macrophyte species, respectively. The site
nearest the discharge point of the mine effluent (study site 1) had the highest Pb
concentrations (103 to 118 (ig Pb/L) and the lowest number of resident macrophyte
species; these included E. crassipes, L. minor, Azolla pinnata and S. polyrrhiza. Based on
analysis of plant tissue Pb concentrations, the authors theorized that certain species may
be more able to develop Pb tolerant eco-types that can survive at higher Pb
concentrations (Mishra etal.. 2008). In field studies available for certain freshwater
habitats, exposure to Pb has been shown to result in significant alterations of invertebrate
communities. Macroinvertebrate community structure in mine-influenced streams was
determined to be significantly correlated to Pb sediment pore water concentrations.
Multiple invertebrate community indices, including Ephemeroptera, Plecoptera,
Trichoptera (EPT) taxa richness, Missouri biotic index, and Shannon-Wiener diversity
index, were integrated into a macroinvertebrate biotic condition score (Poulton et al..
2010). These scores were determined to be significantly lower at sample sites
downstream from mining sites where Pb pore water and bulk sediment concentrations
were elevated. Sediment Pb, Cd, and Zn levels were inversely correlated to mussel taxa
richness in the Spring River basin encompassing sites in Kansas, Missouri and Oklahoma
overlapping a former Pb and Zn mining and processing area (Angelo et al.. 2007). In sites
upstream of the mining area, 21 to 25 species of mussels were present whereas in sites
downstream, only 6 to 8 species were observed.
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Rhea et al. (2006) examined the effects of multiple heavy metals in the Boulder River,
Montana, watershed biofilm on resident macroinvertebrate assemblages and community
structure, and determined that, among all the metals, biofilm Pb concentrations exerted
the greatest influence on the macroinvertebrate community indices. Pb biofilm
concentrations were significantly correlated with reduced EPT taxa richness, reduced
EPT abundance, and an increase in Diptera species abundance. Interestingly, Pb
concentrations in invertebrate tissues were correlated to an increase in Hydropsychidae
caddisfly abundance, but this may have resulted from the intrinsically high variability in
tissue Pb concentrations. The authors concluded that Pb-containing biofilm represented a
significant dietary exposure for impacted macroinvertebrate species, thus altering
invertebrate community metrics (Rhea et al.. 2006).
Kominkova and Nabelkova (2005) examined ecological risks associated with metal
contamination (including Pb) in small urban streams. Although surface water Pb
concentrations in monitored streams were determined to be very low, concentrations of
the metal in sediment were high enough to pose a risk to the benthic community (e.g., 34
to 101 mg Pb/kg). These risks were observed to be linked to benthic invertebrate
functional feeding group, with collector-gatherer species exhibiting larger body burdens
of heavy metals than other groups (Kominkova and Nabelkova. 2005). In contrast,
benthic predators and collector-filterers accumulated significantly lower metals
concentrations. Consequently, it is likely that sediment-bound Pb contamination would
differentially affect members of the benthic invertebrate community, potentially altering
ecosystem dynamics.
Invertebrate functional feeding group may also affect invertebrate Pb body burdens in
those systems where Pb bioconcentration occurs. The predaceous zooplanktonic rotifer,
A. brightwellii collected from a Pb-impacted freshwater reservoir in Mexico, contained
384 ng Pb/mg and exhibited a water-to-tissue BCF of 49,344. The authors theorized that
Pb biomagnification may have been observed in this case because the cladoceran
M. micrura is both a known Pb accumulator and a favorite prey item of the rotifer
(Rubio-Franchini et al.. 2008). They showed thatM micrura had twice the Pb body
burden of D. similis, another grazing cladoceran species present in the reservoir. These
two species exhibited average Pb tissue concentrations of 57 and 98 ng Pb/mg,
respectively, with respective water column BCFs of 9,022 and 8,046. Conversely, an
examination of the simultaneous uptake of dissolved Pb by the algae P. subcapitata and
the cladoceran D. magna suggests that the dietary exposure route for the water column
filter-feeder is minor. Although Pb accumulated in the algal food source, uptake directly
from the water column was determined to be the primary route of exposure for D. magna
fKomjarova and Blust. 2009cj.
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For many invertebrate species, sediment Pb concentrations may be the most important
driver in determining Pb uptake. For instance, while Hg and Cd body burdens in lentic
invertebrates were affected by lake ecological processes (e.g., eutrophication), a similar
effect was not observed for Pb concentrations in crayfish tissue, despite a high variability
between sites (Larsson et al., 2007). Although this may be a result of differing
bioaccumulation tendencies, the authors suggested that other factors, including the
potential for sediment exposures, may be responsible for Pb uptake in lentic
invertebrates.
A field survey of fishes in the Viburnum Trend Pb-Zn mining district in southeast
Missouri available since the 2006 Pb AQCD, found that species richness and species
density of riffle-swelling benthic fishes were negatively correlated with metal
concentrations in pore water and in fish in mining impacted streams (Allert et al.. 2009b).
Density of Ozark sculpin (Coitus hypselurus) and banded sculpin (Coitus carolinae) were
positively correlated with distance from mining sources.
In addition to the ecological effects discussed above, there is additional evidence that Pb
exposure could alter bacterial infection (and potentially disease transmission) in certain
fish species. Following 96-hour exposures to 4,000 (ig Pb/L, bacterial density in Channa
punctatus fish was observed to be significantly altered when compared to non-exposed
fish. Bacteria population densities in fish spleen, gills, liver, kidneys and muscle tissues
were higher following Pb exposure, with bacterial abundance in the gills too numerous to
quantify (Pathak and Gopal. 2009). In addition, bacteria inhabiting Pb-exposed fish were
more likely to exhibit antibacterial resistance than colonies isolated from non-exposed
fish. Although the mechanism remains unknown, this study suggests that Pb exposure
may increase the likelihood of infection in fish, potentially affecting fish abundance and
recruitment.
In summary, despite the fact that alterations of macrophyte communities may be highly
visible effects of increased sediment Pb concentrations, several recently published papers
propose that ecological impacts on invertebrate communities are also significant, and can
occur at environmental Pb concentrations lower than those required to impact plant
communities. High sediment Pb concentrations were linked to shifts in amphipod
communities inhabiting plant structures, and potentially to alterations in ecosystem
nutrient processing through selective pressures on certain invertebrate functional feeding
groups (e.g., greater bioaccumulation and toxic effects in collector-gatherers versus
predators or filter-feeders). Increased sediment pore water Pb concentrations were
demonstrated to likely be of greater importance to invertebrate communities, as well.
Interestingly, recent research also suggests that Pb exposure can alter bacterial
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infestations in fish, increasing both microbial density and resilience, and potentially
increasing the likelihood of serious disease outbreak.
6.4.8 Critical Loads in Freshwater Systems
The general concept and definition of critical loads is introduced in Section 6.1.3 of this
chapter [also see Section 7.3 of the 2006 Pb AQCD (U.S. EPA. 2006c)1. Critical load
values are linked to critical limits of Pb for endpoints/receptors of interest in the
ecosystems, such as blood Pb. Some important critical limits for Pb in aquatic ecosystems
are discussed in this section along with information on aquatic critical loads for Pb.
Unit World Models (UWM) have been used to calculate critical loads for metals in
aquatic ecosystems. These models couple an ecotoxicity model, the BLM, to a
speciation/complexation model, the Windermere Humic Adsorption Model (WHAM),
then to the multi-species fate model, TRANsport-SPECiation (TRANSPEC). Gandhi et
al. (2011) applied a UWM approach to estimate speciation/complexation, fate and critical
loads using lakes of three different trophic status. A high percentage of colloidal-bound
Pb was found in the eutrophic and mesotrophic lakes (75-80%) versus the oligotrophic
lakes (2%), owing to the high affinity of Pb to DOM. Pb concentrations were lowest for
mesotrophic and highest for oligotrophic systems. Critical loads were not calculated for
Pb; however, for the other metals tested the critical load was lowest in the oligotrophic
and highest in the eutrophic systems.
A critical load of 39.0 g Pb/m2 per year (0.19 mol Pb/ m2 per year) was calculated for a
generalized lake in the Sudbury area of the Canadian Shield using TICKET-UWM based
on acute toxicity data for D. magna. (Farley et al., 2011). The model was set up to
calculate critical loads of metals by specifying free metal ion activity or the critical biotic
ligand concentration. This critical load for Pb was much higher than for Cu, Ni and Zn
and the authors attribute this difference to the strong binding of Pb to particulate organic
matter and the sequestration of PbCO3 in sediment.
Given the heterogeneity of ecosystems exposed to Pb, and the differences in expectations
for ecosystem services attached to different land uses, it is expected that there will be a
range of critical load values for Pb for soils and waters within the U.S. Refer to
Section 7.3.6 of the 2006 Pb AQCD for additional discussion of critical loads of Pb in
aquatic systems.
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6.4.9 Characterization of Sensitivity and Vulnerability in Freshwater
Systems
Data from the literature indicate that exposure to Pb may affect survival, reproduction,
growth, metabolism, and development in a wide range of freshwater aquatic species.
Often, species differences in metabolism, sequestration, and elimination rates control
relative sensitivity and vulnerability of exposed organisms. Diet and lifestage at the time
of exposure also contribute significantly to the determination of sensitive and vulnerable
populations and communities. Further, environmental conditions in addition to those
discussed as affecting bioavailability (Sections 6.4.3 and 6.4.4) may also alter Pb toxicity.
The 2006 Pb AQCD (U.S. EPA. 2006b) reviewed the effects of genetics, age, and body
size on Pb toxicity. While genetics appears to be a significant determinant of Pb
sensitivity, effects of age and body size are complicated by environmental factors that
alter metabolic rates of aquatic organisms. A review of the more recent literature
corroborated these findings, and identified seasonally-affected physiological changes and
lifestage as other important determinants of differential sensitivity to Pb.
6.4.9.1 Seasonally-Affected Physiological Changes
A study by Duman et al. (2006) identified species and seasonal effects of Pb uptake in
aquatic plants. P. australis accumulated higher root Pb concentrations than S. lacustris.
Additionally, the P. australis Pb accumulation factor was significantly higher during the
winter versus other seasons, while the Pb accumulation factor for S. lacustris was greatest
in spring and autumn. The Pb accumulation factor for a third species, P. lucens, was
greatest in autumn (Duman et al.. 2006). Most significantly, these changes in
bioaccumulation were not linked with biomass increases, indicating that species-
dependent seasonal physiological changes may control Pb uptake in aquatic macrophytes
(Duman et al.. 2007). Significant interspecies differences in Pb uptake were observed for
plants representing the same genus (Sargassum), indicating that uptake of Pb by aquatic
plants also may be governed by highly species-dependent factors (Jothinayagi and
Anbazhagan. 2009).
Heier et al. (2009) established the speciation of Pb in water draining from a shooting
range in Norway and looked at the time-dependent accumulation in brown trout. They
found that elevated concentrations of Pb, particularly high molecular weight (>10 kDa)
positively charged Pb species associated with organic matter, were correlated with high
flow episodes and accumulation of Pb on gills and in the liver. Thus, high flow episodes
can remobilize metals from a catchment and induce stress to aquatic organisms.
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6.4.9.2 Increased Nutrient Uptake
Singh et al. (2010) proposed that metal-resistant plants have the capacity to not only up-
regulate antioxidant synthesis, but also have the ability to increase nutrient consumption
and uptake to support metal sequestration and detoxification via production of
antioxidants (Singh etal.. 2010). Therefore, it is likely that such plant species would be
significantly less susceptible to Pb exposure than those species without those abilities.
6.4.9.3 Temperature and pH
Water temperature also appears to affect the toxicity of Pb to aquatic organisms, with
higher temperatures leading to greater responses. Pb toxicity to crayfish increased 7 to
10% when the water temperature was increased by 4 °C, and by 14% when the
temperature increased by 7 °C. The authors determined that the increased toxicity was a
result of the negative impact of Pb on crayfish respiration, which was exacerbated by the
lower dissolved oxygen concentrations at higher water temperatures (Khan et al.. 2006).
In a study of the combined effects of temperature and Pb concentration on two freshwater
rotifer species, Brachionus havanaensis and B. rubens, population growth was measured
in three nominal concentrations of Pb as Pb chloride (50, 100 and 200 (ig Pb/L) for 15
days at either 22 °C or 32 °C (Montufar-Melendez et al.. 2007). At 22 °C, population
growth of B. havanaensis was suppressed by B. rubens regardless of Pb treatment. At the
higher temperature, there was no population increase of B. rubens at any Pb
concentration. In the controls, population growth rates of B. havanaensis, but not
B. rubens, increased with an increase in temperature. These studies highlight the role of
temperature in Pb toxicity in organisms adapted to low temperatures.
The sequestration ability of L. minor macrophytes was similarly impacted by increased
surface water temperature; plants absorbed a maximum Pb concentration of 8.6 mg /g at
30 °C, while uptake at 15 °C was only 0.3 mg/g (Uysal and Taner. 2009). Decreased pH
was also demonstrated to increase the uptake of environmental Pb in aquatic plants
(Wang etal.. 2010b; Uysal and Taner. 2009). Lower pH was shown to result in increased
sensitivity to Pb in juvenile fathead minnows in 30-day exposure to Pb of varying
concentrations (Grosell et al., 2006a). Additionally, Birceanu et al. (2008) determined
that fish (specifically rainbow trout) were more susceptible to Pb toxicity in acidic, soft
waters characteristic of sensitive regions in Canada and Scandinavia. Hence, fish species
endemic to such systems may be more at risk from Pb contamination than fish species in
other habitats.
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6.4.9.4 Lifestage
It is clear that certain stages of a life cycle are more vulnerable to Pb. A comparison of
C. riparius Pb LC50 values derived from toxicity tests with different instars indicates a
significant effect of lifestage on Pb sensitivity for aquatic invertebrates. Bechard et al.
(2008) calculated a first instar C. riparius 24-hour LC50 value of 613 (ig Pb/L, and
contrasted this value with the 24-hour and 48-hour LC50 values derived using later instar
larvae—350,000 and 200,000 (ig Pb/L, respectively. This disparity would suggest that
seasonal co-occurrence of aquatic Pb contamination and sensitive early instars could have
significant population-level impacts (Bechard et al.. 2008). Similarly, Wang et al. (201 Of)
demonstrated that the newly transformed juvenile mussels, L. siliquoidea and
L. rafinesqueana, at 5 days old were more sensitive to Pb exposure than were glochidia
or two to six month- old juveniles, suggesting that Pb exposure at particularly sensitive
lifestages could have a significant influence on population viability (Wang etal. 2010f).
Evidence for differences in susceptibility to Pb at distinct lifestages is also available for
freshwater fish. In chronic (60-day) early-lifestage test exposures conducted with
rainbow trout to develop ACR's for this species the study resulted in ACR's for rainbow
trout lower than previously reported due to the acute tests which produced LC50 values
that were 10 to 25 times lower than earlier studies with trout (Mebane et al.. 2008). The
authors speculated that the lower LC50 values were due to the age of the fish used in the
study (two to four week old fry) and that testing with larger and older fish may not be
protective of more sensitive lifestages. Post-hatching stages of the African catfish were
more sensitive than the embryonic stage to Pb-exposure and the authors attributed this
apparent protective effect to the presence of a hardened chorion in embryos (Osman et
al.. 2007a).
6.4.9.5 Species Sensitivity
Species-specific Ca2+ requirements have been shown to affect the vulnerability of aquatic
organisms to Pb. The snail, L. stagnalis, exhibits an unusually high Ca2+ demand due to
CaCO3 formation required for shell production and growth, and exposure to Pb prevents
the uptake of needed Ca2+, leading to toxicity. Consequently, aquatic species that require
high assimilation rates of environmental Ca2+ for homeostasis are likely to be more
sensitive to Pb contamination (Grosell and Brix. 2009). Grosell and colleagues also noted
that reduced snail growth following chronic Pb exposure was likely a result of reduced
Ca2+ uptake (Grosell et al.. 2006b).
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There is some indication that molting may comprise an additional sequestration and
excretion pathway for aquatic animals exposed to Pb (Soto-Jimenez et al.. 201 la:
Mohapatra et al.. 2009; Tollett et al.. 2009; Bergevand Weis. 2007). Libellulidae
dragonfly nymphs (Tollett et al.. 2009) have been shown to preferentially sequester Pb in
exoskeleton tissue. Consequently, aquatic arthropod species and those species that shed
their exoskeleton more frequently may be able to tolerate higher environmental Pb
concentrations than non-arthropods or slow-growing molting species, as this pathway
allows them to effectively lower Pb body burdens.
In contrast, the effect of Pb exposure on fish bacterial loads demonstrated by Pathak and
Gopal (2009) suggest that infected fish populations may be more at risk to the toxic
effects of Pb than healthier species. Aqueous Pb was demonstrated to both increase
bacteria density in several fish organs and to improve the likelihood of antibacterial
resistance (Pathak and Gopal 2009).
Tolerance to prolonged Pb exposure may develop in aquatic invertebrates and fish. Multi-
generational exposure Pb appears to confer some degree of metal tolerance in
invertebrates such as C. plumosus larvae; consequently, previous population Pb
exposures may decrease species' susceptibility to Pb contamination (Vedamanikam and
Shazilli. 2008b). However, the authors noted that metal tolerant larvae were significantly
smaller than larvae reared under clean conditions, and that transference of Pb-tolerant
C. plumosus larvae to clean systems resulted in a subsequent loss of tolerance. Evidence
of acclimation to elevated Pb in fathead minnow was suggested in the variations in
ionoregulatory parameters that were measured on day 10 and 30 in fish exposed to
115 (ig Pb/L for 30 days. At the end of the experiment, whole body Ca2+ was elevated
while Na+ and K+ recovered from elevated levels at 30 days (Grosell et al., 2006a).
A series of species sensitivity distributions constructed by Brix et al. (2005) in freshwater
systems indicated that sensitivity to Pb was greatest in crustacean species, followed by
coldwater fish, and warmwater fish and aquatic insects, which exhibited a similar
sensitivity. Further, analysis of both acute and chronic mesocosm data sets indicated that
Pb-contaminated systems exhibited diminished species diversity and taxa richness
following both types of exposure (Brix et al.. 2005). Wong et al. (2009) constructed Pb
species sensitivity distributions for both cladoceran and copepod freshwater species. A
comparison of the two curves indicated that cladoceran species, as a group, were more
sensitive to the toxic effects of Pb than were copepods, with respective hazardous
concentration values for 5% of the species (HC5) values of 35 and 77 (ig Pb/L. This
difference in sensitivities would indicate that cladoceran species are more likely to be
impacted at lower environmental Pb concentrations than copepods, potentially altering
community structures or ecosystem functions (Wong et al., 2009).
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6.4.9.6 Ecosystem Vulnerability
Relative vulnerability of different freshwater ecosystems to effects of Pb can be inferred
from the information discussed above on species sensitivity and the influence of water
quality variables on the bioavailability and toxicity of Pb. It is, however, difficult to
categorically state that certain freshwater plant, invertebrate or vertebrate communities
are more vulnerable to Pb than others, since toxicity is dependent on many variables and
data from field studies are complicated by co-occurrence of other metals and alterations
of pH, such as in mining areas. Aquatic ecosystems with low pH and low DOM are likely
to be the most sensitive to the effects of atmospherically-deposited Pb. Examples of such
systems are acidic, soft waters such as sensitive regions in Canada and Scandinavia
(Birceanu et al., 2008). In the U.S., aquatic systems that may be more sensitive to effects
of Pb include habitats that are acidified due to atmospheric deposition of pollutants,
runoff from mining activities or lakes and streams with naturally occurring organic acids.
Hence, fish and invertebrate species endemic to such systems may be more at risk from
Pb contamination than corresponding species in other habitats.
6.4.10 Ecosystem Services Associated with Freshwater Systems
Pb deposited on the surface of, or taken up by organisms has the potential to alter the
services provided by freshwater biota to humans although the directionality of impacts is
not always clear. For example, aquatic macrophytes provide a service by sequestering Pb.
At the same time, the uptake of Pb by plants may result in toxicological effects associated
with Pb exposure and decreased capacity of wetland species to remove contaminants. At
this time, few publications address Pb impacts on ecosystem services associated with
freshwater systems and most studies focus on wetlands rather than lakes and streams. Pb
can affect the ecological effects in each of the four main categories of ecosystem services
(Section 6.1.2) as defined by Hassan et al. (2005). These effects are sorted into ecosystem
services categories and summarized here:
• Supporting: food for higher trophic levels, biodiversity
• Provisioning: clean drinking water, contamination of food by heavy metals,
decline in health offish and other aquatic species
• Regulating: water quality
• Cultural: ecosystem and cultural heritage values related to ecosystem integrity
and biodiversity, wildlife and bird watching, fishing
Freshwater wetlands are sinks for atmospheric Pb as well as Pb from terrestrial runoff
(Landre et al., 2010; Watmough and Dillon, 2007). Several studies have addressed the
response of natural wetlands to Pb (Odum. 2000; Gambrell. 1994). Recent reviews of
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pollution control (Mander and Mitsch. 2009) or removal of metals (Marchand et al.,
2010) by constructed wetlands and phytoremediation of metals by wetland plants (Rai.
2008) indicate that these systems can remove Pb from the aquatic environment and are
important for water quality, sediment stabilization, nutrient cycling and shelter for
aquatic biota. The use of plants as a tool for immobilization of Pb and other metals from
the environment is not limited to wetland species. Recent advances in the
phytoremediation of metals are reviewed in Dickinson et al. (2009). The impact of Pb on
ecological services provided by specific components of aquatic systems has been
considered in a limited number of studies. For example, Theegala et al. (2007) discuss the
high uptake rate of Pb by D. pulex as the basis for a possible Daphnia-based remediation
for aquatic systems.
6.4.11 Synthesis of New Evidence for Pb Effects in Freshwater Ecosystems
This synthesis of the effects of Pb on freshwater ecosystems covers information from the
publication of the 2006 Pb AQCD to present. It is followed in Section 6.4.12 by
determinations of causality that take into account evidence dating back to the 1977 Pb
AQCD.
Evidence assessed in the present document supports the findings of the previous Pb
AQCDs that waterborne Pb is highly toxic to freshwater organisms, with toxicity varying
with species and lifestage, duration of exposure, form of Pb, and water quality
characteristics. The studies that are available for freshwater plants, invertebrates and
vertebrates include studies where Pb concentration was analytically verified and those
that reported nominal concentrations (Table 6-5). Many of the studies that report nominal
concentrations in media are uptake studies that subsequently quantify Pb in tissues,
however, measurement of Pb in water or sediment at the beginning of an exposure is
desirable when comparing laboratory studies to concentrations of Pb in freshwater
systems. As reported in Section 6.2.3 and Table 6-2. the median and range of Pb
concentrations in surface waters (median 0.50 (ig Pb/L, range 0.04 to 30 (ig Pb/L) and
sediments (median 28 mg Pb/kg dry weight, range 0.5 to 12,000 mg Pb/kg dry weight) in
the U.S. based on a synthesis of NAWQA data was reported in the previous
2006 Pb AQCD (U.S. EPA. 2006c).
Recent studies provide further evidence for the role of modifying factors such as pH,
DOC and hardness on the effects of Pb on plants, invertebrates and vertebrates. The same
Pb concentration added to water or sediment produces far greater effects under some
conditions, than others. Many studies reviewed in the ISA included concentrations that
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were higher than Pb found near contaminated areas. However, when multiple
concentrations were used, effects gradually increased with increasing Pb exposure.
Effects at lower concentrations can be implied from many studies since an exposure-
response relationship to Pb was observed, although uncertainty remains in relating these
findings to reported concentrations of Pb in freshwater. Many studies only report an LC50
value when an LCio would be more relevant for consideration of effects on organisms
since an effect occurring at the LC50 would most likely not maintain a stable population.
Most available studies only report acute toxicity and are conducted at higher
concentrations of Pb than found in sampling from U.S. surface waters (Table 6-2).
however, exposure to Pb in freshwater systems is most likely characterized as a chronic
low dose exposure.
Plants
Most recent studies on effects of Pb in freshwater algal species reviewed in
Section 6.4.5.1 were conducted with nominal media exposures and effect concentrations
greatly exceeded Pb reported in surface water. In studies where Pb was quantified, effect
concentrations for growth (EC50) for aquatic macrophytes were much higher than
currently reported ambient Pb, however, some sublethal endpoints such as effects on
chlorophyll were observed at lower concentrations. For example, chlorophyll a content
was significantly inhibited at 210 (ig Pb/L and higher in W. arrhiza (Piotrowska et al.,
2010). An increase in biomass was reported in L. minor exposed to 100 or 200 (ig Pb/L
with inhibition observed at higher concentrations (Dirilgen. 2011). There were also
numerous studies conducted at nominal Pb concentration that report effects on enzyme
activities and protein content in freshwater aquatic plant species. Exposure-response
relationships in which increasing concentrations of Pb lead to increasing effects were
consistently observed for freshwater aquatic plants.
Recent studies on bioavailability of Pb in aquatic plants and algae support the findings of
previous Pb AQCDs that plants tend to sequester larger amounts of Pb in their roots than
in their shoots and provide additional evidence for species differences in
compartmentalization of sequestered Pb and responses to Pb in water and sediments.
Given that atmospherically-derived Pb is likely to become sequestered in sediments,
uptake by aquatic plants is a significant route of Pb removal from sediments, and a
potential route for Pb mobilization into the aquatic food web. Although there are some
similarities to Pb accumulation observed in terrestrial plants (e.g., preferential
sequestration of the metal in root tissue), Pb appears to be more bioavailable in sediment
than it is in soil.
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Invertebrates
The largest body of evidence for effects of Pb at or near concentrations of this metal
found in surveys of surface waters of the U.S. is for invertebrates and recent studies
reviewed in Sections 6.4.5.2 and 6.4.6 further support this observation. Exposure-
response relationships in which increasing concentrations of Pb lead to increasing effects
were consistently observed for freshwater invertebrates. Among the most sensitive
species, growth of juvenile freshwater snails L. stagnalis was inhibited at an EC2o of
<4 (ig Pb/L. (Grosell and Brix. 2009; Grosell et al. 2006b). A chronic value of
10 (ig Pb/L obtained in 28-day exposures of 2-month-old fatmucket mussel,
L. siliquoidea juveniles was the lowest genus mean chronic value ever reported for Pb
(Wang et al.. 2010f). The 96-hour EC50 values for 5-day-old juveniles in two separate
toxicity tests with this species were 142 and 298 (ig Pb/L (mean EC50 220 (ig Pb/L).
Recent studies (Sections 6.4.5.2 and 6.4.6) have further elucidated the role of water
quality on Pb toxicity. In freshwater invertebrates some effects were observed at
concentrations occasionally encountered in U.S. surface waters (Table 6-2). In a 7-day
exposure of the cladoceran C. dubia to 50 to 500 (ig Pb/L, increased DOC leads to an
increase in mean EC50 for reproduction ranging from approximately 25 (ig Pb/L to
>500 (ig Pb/L (Mager et al.. 201 la). The 48-hour LC50 values for the cladoceran C. dubia
tested in eight natural waters across the U.S. varied from 29 to 1,180 (ig Pb/L and were
correlated with DOC (Esbaughetal.. 2011).
Additional new evidence reviewed in Sections 6.4.5.2 and 6.4.6 for effects near the upper
range of concentrations of Pb available from surveys of U.S. surface waters include
studies with rotifer, midge and mayfly species. The freshwater rotifer E. dilatata 48 hour
LC50 was 35 (ig Pb/L using neonates hatched from asexual eggs (Arias-Almeida and
Rico-Martinez. 2011). An EC2o for reduced growth and emergence of the midge
C. dilutus was reported to be 28 (ig Pb/L, observed in a 55-day exposure, while the same
species had a 96-hour LC50 of 3,323 (ig Pb/L (Mebane et al.. 2008) The ECio for molting
in the mayfly B. tricaudatus was 37 (ig Pb/L (Mebane et al., 2008). All of these effect
concentrations provide additional evidence for Pb effects on freshwater invertebrates.
Vertebrates
For freshwater fish (Sections 6.4.5.3 and 6.4.6). most recent studies available since the
2006 Pb AQCD, were conducted with fathead minnow, P. promelas, or rainbow trout,
O. mykiss. In a series of 96-hour acute toxicity tests with fathead minnow conducted in a
variety of natural waters across North America, LC50 values ranged from 41 to
3,598 (ig Pb/L (Esbaugh et al.. 2011). Reproductive effects associated with water quality
parameters were also noted with this species (Mager et al.. 2010). In trout, no effects of
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Pb were observed in dietary studies. In chronic aqueous exposures with trout the
following endpoints were reported: NOEC=24 (ig Pb/L, ECi0=26 (ig Pb/L,
EC2o=34 (ig Pb/L, and LC50=55 (ig Pb/L. In a separate test with the same species an
NOEC=8 jig Pb/L, EC10=7(ig Pb/L, EC20=102 ng Pb/L and LC50=120 jig Pb/L were
reported. In acute tests with rainbow trout run concurrently with the chronic tests,
96-hour LC50 values were 120 and 150 (ig Pb/L, respectively (Mebane et al.. 2008).
These reported effects provide additional evidence for toxicity of Pb to fish and chronic
NOEC and ECio values reported for trout, a sensitive species, are within the upper range
of Pb currently reported in U.S. surface waters (Table 6-2).
In Section 6.4.5.3. a study with the frog R. pipiens exposed nominally to Pb, tissue
concentrations were quantified at the end of the study and found to be in the range of Pb
tissue concentrations in wild-caught tadpoles. Growth rate was significantly slower in the
100 (ig Pb/L nominal concentration and more than 90% of tadpoles developed lateral
spinal curvature. Time to metamorphosis was also delayed at this treatment level.
Food Web
In the 2006 Pb AQCD, trophic transfer of Pb through aquatic food chains was considered
to be negligible. Concentrations of Pb in the tissues of aquatic organisms were generally
higher in algae and benthic organisms than in higher trophic-level consumers indicating
that Pb was bioaccumulated but not biomagnified (U.S. EPA. 2006c: Eisler. 2000). Some
studies published since the 2006 Pb AQCD, (see Section 6.4.4.4) support the potential for
Pb to be transferred in aquatic food webs, while other studies indicate that Pb
concentration decreases with increasing trophic level (biodilution).
Community and Ecosystem Effects
New evidence of effects of Pb at the community and ecosystem levels of biological
organization reviewed in Section 6.4.7 include shift in community composition in
macrophytes. Effects on reproduction, growth or survival (summarized in Table 6-5) may
lead to effects at the population-level of biological organization and higher. Additional
evidence for community and ecosystem level effects of Pb have been observed primarily
in microcosm studies or field studies with other metals present.
6.4.12 Causal Determinations for Pb in Freshwater Systems
In the following sections, organism-level effects on reproduction and development,
growth and survival are considered first since these endpoints can lead to effects at the
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population level or above and are important in ecological risk assessment.
Neurobehavioral effects are considered next followed by sub-organismal responses
(hematological effects, physiological stress) for which Pb has been shown to have an
impact in multiple species and across taxa, including humans. Causal determinations for
terrestrial, freshwater and saltwater ecological effects are summarized in Table 6-3.
6.4.12.1 Reproductive and Developmental Effects-Freshwater
Biota
Evaluation of the findings in previous Pb AQCDs and recent literature on Pb effects in
aquatic fauna indicates that exposure to Pb is associated with reproductive effects at or
near ambient concentrations of this metal (Table 6-2) in some freshwater species.
Impaired fecundity at the organismal level can result in a decline in abundance and/or
extirpation of populations, decreased taxa richness, and decreased relative or absolute
abundance at the community level (Suter et al. 2005; U.S. EPA. 2003a). Various
endpoints have been measured in freshwater organisms to assess the effect of Pb on
fecundity, development and hormone homeostasis. However, there are typically only
limited studies available from different taxa. Recent evidence available since the
2006 Pb AQCD for effects of Pb on reproductive endpoints in freshwater invertebrates
and vertebrates is summarized in Table 6-5.
There are no studies reviewed in the ISA or previous Pb AQCDs on development and
reproductive effects of Pb in freshwater aquatic algae or macrophytes.
Experimental data from freshwater invertebrates provide evidence for increasing
reproductive effects associated with increasing exposure to Pb. The exposure-response
relationship is used in judging causality (Table I of the Preamble). Reproductive effects
of Pb in freshwater aquatic invertebrates are well-characterized in previous Pb AQCDs,
the draft Ambient Aquatic Life Water Quality Criteria for Lead report (U.S. EPA. 2008b).
and in the current ISA and have been observed at or near current ambient concentrations
(median 0.5 jig Pb/L, range 0.04 to 30 jig Pb/L) (U.S. EPA. 2006c) in some species in
laboratory exposures. In the 1986 Pb AQCD reproductive effects were reported to begin
at 19 (ig Pb/L for the freshwater snail Lymnaea palustris and 27 (ig Pb/L for Daphnia sp.
(U.S. EPA. 1986b). In the 2006 Pb AQCD (U.S. EPA. 2006b) the number of neonates per
surviving adult was significantly decreased in the amphipod H. azteca during chronic
42-day exposures to Pb (Besser et al., 2005). In the group exposed to Pb in water-only
exposures, the LOEC for reproductive effects was 16 (ig Pb/L while in amphipods
receiving both water-borne and dietary Pb the LOEC for reproduction was 3.5 (ig Pb/L.
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New evidence in freshwater invertebrates (Table 6-5 and Section 6.4.5.2) show
consistency of the observed association between reproductive endpoints and Pb exposure.
In the freshwater rotifer B. calyciflorus, reproductive output was measured as total
number of individuals and intrinsic growth rate. The EC2o for number of rotifers was
125 (ig Pb/L and the 48 hour EC2o for intrinsic rate of population increase was
307 (ig Pb/L with an NOEC of 194 (ig Pb/L (Grosell et al.. 2006b). In a 7-day exposure
of the cladoceran C. dubia to 50 to 500 (ig Pb/L, increased DOC leads to an increase in
mean EC50 for reproduction ranging from approximately 25 (ig Pb/L to >500 (ig Pb/L
(Mageretal.. 201 la). Additional reproductive impairment endpoints for freshwater
cladocerans are reported in Table 6 of the draft Ambient Aquatic Life Water Quality
Criteria for Lead report (U.S. EPA. 2008b). It is not clear how these laboratory-derived
values for freshwater invertebrates compare to Pb exposures in natural systems due to the
role of modifying factors (i.e., pH, hardness, and DOC) which affect Pb speciation and
bioavailability, however, results under controlled conditions have consistently shown
reproductive effects of Pb in sensitive taxa (amphipods, cladocerans) at concentrations at
or near Pb quantified in freshwater environments.
In freshwater aquatic vertebrates there is evidence for reproductive and developmental
effects of Pb. Pb exposure in frogs has been demonstrated to delay metamorphosis,
decrease larval size and produce skeletal malformations. For example, in northern
leopard frog R. pipiens exposed to nominal concentrations of 100 (ig Pb/L from
embryonic stage to metamorphosis, maximum swimming speed was significantly slower
than other treatment groups and 92% of tadpoles exposed to 100 (ig Pb/L had lateral
spinal curvature (compared with 6% in the control) (Chen et al., 2006b). Pb tissue
concentrations were quantified in frogs following exposure and fell within the range of
tissue concentrations in wild-caught tadpoles.
The weight of evidence for reproductive and developmental effects in freshwater
vertebrates is from studies with fish. Pb AQCDs have reported developmental effects in
fish, specifically spinal deformities in brook trout (Salvelinus fontinalis) exposed to
119 ng Pb/L for three generations (U.S. EPA. 1977). and in rainbow trout as low as
120 ng Pb/L (U.S. EPA. 1986b). Reproductive behaviors of fathead minnows including
reduced time spent in nest preparation by males, increased interspawn periods and
reduced oviposition by females was observed following a 4-week exposure to
500 ng Pb/L (Weber. 1993). In the 2006 Pb AQCD (U.S. EPA. 2006b). decreased
spermatocyte development in rainbow trout was reported at 10 (ig Pb/L and, in fathead
minnow testicular damage occurred at 500 (ig Pb/L. In a recent study, reproductive
effects in fathead minnows were influenced by water chemistry parameters (alkalinity
and DOC) in breeding exposures following 300 day chronic toxicity testing with Pb
(Mageretal.. 2010). Specifically, in fish treated in both 35 and 120 (ig Pb/L with HCO3"
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and with 120 (ig Pb/L with DOC, total reproductive output was decreased and average
egg mass production increased as compared to egg mass size in controls and in low
HCO3" and DOC treatments with Pb. No significant differences were present between
treatments in egg hatchability. In a feeding study, Reproductive performance was
unaffected in zebrafish exposed to Pb-via consumption of contaminated prey (Boyle et
al.. 2010). In fish, there is evidence for alteration of steroid profiles and additional
reproductive parameters although most of the available studies were conducted using
nominal concentrations of Pb.
Observations of Pb toxicity to reproductive and developmental endpoints in freshwater
fauna are further supported by evidence in terrestrial invertebrates and vertebrates
(Section 6.3.12.1), marine invertebrates (Section 6.4.21.1) and from laboratory animals
(Section 4.8). Pb appears to affect multiple endpoints associated with reproduction and
development in aquatic invertebrates and vertebrates. A few sensitive invertebrate taxa
(amphipods, cladocerans) have been identified where effects are observed in laboratory
studies at concentrations of Pb that occur in the environment. Overall, there is a dearth of
information on reproductive effects of Pb in natural environments; however, the weight
of evidence is sufficient to conclude that there is a causal relationship between Pb
exposures and developmental and reproductive effects in freshwater invertebrates and
vertebrates. In freshwater plants, the evidence is inadequate to conclude that there is a
causal relationship between Pb exposures and plant developmental and reproductive
effects.
6.4.12.2 Growth Effects-Freshwater Biota
Alterations in the growth of an organism can impact population, community and
ecosystem level variables. Growth is a commonly measured endpoint in aquatic plants,
however, reported effects typically occur at concentrations that exceed Pb quantified in
freshwater habitats. Growth effects of Pb on plants include visible growth responses and
reduction of photosynthetic rate, inhibition of respiration, cell elongation, root
development or premature senescence (U.S. EPA. 1986b). In the 2006 Pb AQCD (U.S.
EPA. 2006b). both freshwater algae and plants had EC50 values for growth in the range of
1,000 to > 100,000 (ig Pb/L with minimal inhibition of growth observed as low as
100 (ig Pb/L (U.S. EPA. 2006c). The most sensitive aquatic macrophyte reported in the
2006 Pb AQCD was A. pinnata with an EC50 for relative growth rate of 1,100 (ig Pb/L
following a 4-day exposure to Pb (Gauretal. 1994). An LOEC of 25 (ig Pb/L for
reduced chlorophyll in Coontail (Ceratophyllum demersum), and 50 (ig Pb/L in Cattail
(T. latifolid) following 12-day exposure to Pb (as Pb acetate) were the lowest reported
concentrations of growth-related effects in freshwater plants in the draft Ambient Aquatic
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Life Water Quality Criteria for Lead report (U.S. EPA. 2008b). and were near the upper
range of Pb values reported from sampling of U.S. surface waters (Table 6-2). Additional
growth studies in freshwater algae and plants summarized in Table 6 of the draft Ambient
Aquatic Life Water Quality Criteria for Lead report and Table 6-5 of the present
document report growth effects in laboratory studies at concentrations that exceed
measured levels of Pb in the aquatic environment (U.S. EPA. 2008b).
Most of the evidence for growth effects of Pb in freshwater biota is for invertebrates.
Some of these studies report inhibition of growth in sensitive species occurring at or near
the current upper range of Pb in surface waters (median 0.50 (ig Pb/L, range 0.04 to
30 (ig Pb/L) (U.S. EPA. 2006c). Growth effects of Pb on aquatic invertebrates were
reviewed in the draft Ambient Aquatic Life Water Quality Criteria for Lead report (U.S.
EPA. 2008b) and in the 2006 Pb AQCD. The lowest reported LOEC for growth in the
2006 Pb AQCD (16 (ig Pb/L) was in amphipods (H. aztecd) in a 42-day chronic exposure
(Besser et al. 2005).
Recent studies provide additional evidence for effects on growth of freshwater aquatic
invertebrates at < 10 (ig Pb/L. Growth effects observed in invertebrates underscores the
importance of lifestage to overall Pb sensitivity. In general, juvenile organisms are more
sensitive than adults. Growth of juvenile freshwater snails L. stagnalis was inhibited
below the lowest concentration tested resulting in an EC2o <4 (ig Pb/L (Grosell and Brix.
2009; Grosell et al.. 2006b). In the same study, the NOEC was 12 jig Pb/L and the LOEC
was 16 (ig Pb/L. The authors indicated that the observed effect level for Pb was very
close to the current U.S. EPA water quality criteria for Pb (3.3 (ig Pb/L normalized to test
water hardness) (Grosell and Brix. 2009). In the freshwater mussel, fatmucket
(L. siliquoidea) juveniles were the most sensitive lifestage (Wang et al., 2010f). In this
study, growth of juvenile mussels at the end of a 28-day exposure in 17 (ig Pb/L was
significantly reduced from growth in the controls. A chronic value of 10 (ig Pb/L in 2-
month-old fatmucket juveniles was the lowest genus mean-chronic value ever reported
for Pb. The ECi0 and EC2o for reduced growth and emergence of the midge C. dilutus in
a 55-day exposure were 28 (ig Pb/L and 55 (ig Pb/L, respectively, while the same species
had a 96-hour LC50 of 3,323 ng Pb/L (Mebane et al.. 2008) The EC10 and EC20 for
molting in the mayfly B. tricaudatus were 37 (ig Pb/L and 66 (ig Pb/L, respectively
(Mebane et al., 2008). In natural freshwater systems the effects of Pb are influenced by
additional factors (i.e., pH, hardness, and DOC) which may modulate the toxicity of Pb
observed under laboratory conditions.
Evidence for growth effects of Pb in freshwater aquatic vertebrates is limited to a few
studies in amphibians and fish. In the 2006 Pb AQCD growth effects of Pb were reported
in frogs at concentrations typically higher than currently found in the environment. A
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recent study supports findings of growth effects in frogs and suggests that these effects
may be occurring at lower concentrations: the growth rate of tadpoles of the northern
leopard frog exposed nominally to 100 (ig Pb/L from embryo to metamorphosis was
slower than the growth rate of the controls (Chen et al.. 2006b). In this study, Pb
concentrations in the tissues of tadpoles were quantified and the authors reported that
they were within the range of reported tissue concentrations reported in wild-caught
populations.
Reports of Pb-associated growth effects in freshwater fish are inconsistent (Mager. 2012).
In a review cited in the 2006 Pb AQCD, general symptoms of Pb toxicity in fish included
growth inhibition (Eisler. 2000) however, other studies with Pb have shown no effects on
growth (Mager. 2012). In the studies reviewed for the current ISA no growth effects were
observed in fish exposed to Pb via dietary intake. Recent aqueous exposure studies with
fathead minnows showed significant increases in body length and body mass following
chronic low Pb exposure, however, the authors noted that some effects were observed in
tanks with high mortality early in the exposure (Mager and Grosell. 2011; Grosell et al.,
2006a). Other studies with fathead minnows have shown growth reductions with Pb
exposure, however, concentrations of observed effects typically exceeded the 96-hour
LC50 value (Mager. 2012: Mager etal.. 2010: Grosell et al.. 2006a). Two 60-day early
lifestage tests with rainbow trout showed differences in LOEC for reduced growth
(TVIebane et al.. 2008). In the first test, a 69-day exposure, the LOECs for mortality and
reduced growth were the same (54 (ig Pb/L). In the second test, a 62-day exposure of Pb
to rainbow trout, the LOEC for fish length was 18 (ig Pb/L with an EC2o of >87 (ig Pb/L.
Evidence of effects of Pb exposure on growth in terrestrial plants (Section 6.3.12.2) is
highly coherent with evidence from freshwater plants. Although there is a lack of
evidence in freshwater plants for growth effects at concentrations of Pb typically
encountered in U.S. surface waters, several studies suggest that minimal growth
inhibition can occur within one to two orders of magnitude of the reported range for
freshwater. Due to the concentration-response relationship observed between Pb exposure
and freshwater plants, growth is likely impacted at lower, more ecologically relevant
ECio or LOEC values, than the typically reported EC50 values which may not be suitable
for a maintaining a sustainable population.
There is a large body of evidence to support growth effects of Pb on aquatic plants at
concentrations that greatly exceed those typically found in U.S. surface waters. Less
evidence is available at current concentrations of Pb measured in U.S. surface waters and
within one to two orders of magnitude above the range of these measured values. The
available evidence is, however, sufficient to conclude that a causal relationship is likely
to exist between Pb exposures and growth effects in freshwater plants. The evidence is
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sufficient to conclude that there is a causal relationship between Pb exposures and growth
effects in aquatic invertebrates. Available evidence is inadequate to conclude that there is
a causal relationship between Pb exposures and growth effects in aquatic vertebrates.
6.4.12.3 Survival-Freshwater Biota
The relationship between Pb exposure and survival has been well demonstrated in
freshwater species as presented in Section 6.4.5 and Table 6-5 of the present document
and in the previous Pb AQCDs. Pb exposure can either result in direct lethality or
produce sublethal effects that diminish survival probabilities. Survival is a biologically
important response that can have a direct impact on population size. The concentration at
which there is 50% mortality of test organisms (LC50) is one of the most commonly
reported measures of toxicity. LC50 is more sensitive to changes in exposure than
estimates of lethal concentration for other percentages of the population such as LCio (the
concentration at which there is 10% mortality), and can thus be estimated with greater
precision. However, LC50 alone may be less informative for consideration of effects at
ecologically-relevant concentrations. It is most commonly used to describe acute toxicity,
whereas Pb effects on ecosystem receptors may also be characterized as a result of
chronic and cumulative exposures. Furthermore, a scenario in which 50% of a population
does not survive is likely not a sustainable population. From the LC50 data on Pb in this
review and previous Pb AQCDs, a wide range of sensitivity to Pb is evident across taxa
and within genera. However, the LC50 is usually much higher than current environmental
levels of Pb in the U.S, even though physiological dysfunction that adversely impacts the
fitness of an organism often occurs at concentrations well below lethal ones. When
available, LCio, LC2o, NOEC, or LOEC are therefore reported.
There are no studies reported in the previous Pb AQCDs or the current ISA for aquatic
plants that indicate phytotoxicity at current concentrations of Pb in freshwater
environments.
There are considerable data available on toxicity of Pb to aquatic invertebrates as
reviewed in the previous Pb AQCDs and the draft Ambient Aquatic Life Water Quality
Criteria for Lead report (U.S. EPA. 2008b. 1985V Table AX7-2.4.1 from the
2006 Pb AQCD summarizes LC50 data and other endpoints for freshwater and marine
invertebrates (U.S. EPA. 2006c). Recent studies available since the 2006 Pb AQCD and
the draft Ambient Aquatic Life Water Quality Criteria for Lead document, that report
mortality data are summarized in Table 6-5. Freshwater invertebrates are generally more
sensitive to Pb exposure than other taxa, with survival impacted in a few species at or
near concentrations that are encountered in aquatic environments (Table 6-2). These
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impacted species may include candidate or endangered species. For example, the
freshwater mussel L. rafinesqueana (Neosho mucket), is a candidate for the endangered
species list. The EC50 for foot movement (a measure of viability) for newly transformed
juveniles of this species was 188 (ig Pb/L. (Wang et al.. 2010f).
Most of the evidence for Pb effects on survival in freshwater invertebrates is from
sensitive species of gastropods, amphipods, cladocerans and rotifers (Sections 6.4.5.2 and
6.4.6). In some of these organisms, increased mortality is observed in the upper range of
Pb concentration values found in surveys of U.S. surface waters (median 0.50 (ig Pb/L,
range 0.04 to 30 (ig Pb/L) (U.S. EPA. 2006c). although the toxicity of Pb is highly
dependent upon water quality variables such as DOC, hardness and pH. In the 1986 Pb
AQCD, increased mortality was reported in the freshwater gastropod Lymnaeapalutris at
Pb concentration as low as 19 (ig Pb/L effectively reducing total biomass production
(Borgmann et al.. 1978). Toxicity testing with amphipods under various water conditions
indicate these organisms are sensitive to Pb at <10 (ig Pb/L (U.S. EPA. 2006c) and the
present document). A 7 day LC50 of 1 (ig Pb/L was observed in soft water with the
amphipod H. azteca fBorgmann et al.. 2005j. In this same species, the 96-hour LC50 for
Pb at pH of 5 was 10 (ig Pb/L (Mackie. 1989). In 42-day chronic exposures ofH. azteca
exposed to Pb via water and diet, the LC50 was 16 (ig Pb/L (Besser et al.. 2005). At
higher pH and water hardness, amphipods are less sensitive to Pb (U.S. EPA, 2006c). In a
series of 48 hour acute toxicity tests with the cladoceran C. dubia conducted in a variety
of natural waters across North America, LC50 values ranged from 29 to 1,180 (ig Pb/L
(NOEC range 18 to <985 (ig Pb/L) and were most significantly influenced by DOC and
water ionic strength (Esbaugh et al., 2011). In the 2006 Pb AQCD the range of 48 hour
LC50 values for C. dubia were reported from 280 to >2,700 (ig Pb/L when tested at
varying pH levels (U.S. EPA. 2006c). In the rotifer genus Lecane, a 22-fold difference in
LC50 values was observed in 48-hour exposure to Pb between L. hamata, L. luna and
L. quadridentata. (Perez-Legaspi and Rico-Martinez. 2001). L. luna was most sensitive to
Pb toxicity with a 48-hour LC50 of 140 (ig Pb/L. In neonate rotifers, E. dilatata the 48-
hour LC50 was 35 (ig Pb/L (Arias-Almeida and Rico-Martinez. 2011). A wide range of
LC50 values were reported for chironomid species (Table 6-5). however, the available
evidence suggests these freshwater invertebrates are less sensitive to Pb than amphipods,
cladocerans and rotifers. Other freshwater invertebrates such as odonates may be tolerant
of Pb concentrations that greatly exceed concentrations of Pb reported in environmental
media. Some invertebrates are able to detoxify Pb such as through sequestration of Pb in
the exoskeleton which is subsequently molted.
There is considerable historic information on Pb toxicity to freshwater fish. Early
observations from mining areas where Pb and other metals were present indicated local
extinction offish from streams (U.S. EPA. 1977). The lowest LC50 for fish reported in
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the 1977 Pb AQCD was 1,000 (ig Pb/L in soft water for rainbow trout O. mykiss
(reclassified from Salmo gairdneri) following 96-hour exposure to Pb (U.S. EPA. 1977).
Additional LC50 values for freshwater fish are summarized in the 1985 Ambient Water
Criteria for Lead (U.S. EPA. 1985) and the draft Ambient Aquatic Life Water Quality
Criteria for Lead report (U.S. EPA. 2008b). An LC50 of 236 jig Pb/L adjusted to a total
hardness of 50 mg/L CaCO3 was reported for O. mykiss in the draft Ambient Aquatic Life
Water Quality Criteria for Lead report.
More recently reviewed studies using fish have considered the role of water quality
variables and bioavailability on Pb toxicity (Table 6-5). Higher toxicity tends to occur in
acidic waters where more free-Pb ion is available for uptake. The interactive effects of Pb
concentration and water quality variables on toxicity may result in equivalent toxicity for
a broad range of Pb concentrations. In a series of 96-hour acute toxicity tests with
juvenile fathead minnow conducted in a variety of natural waters across North America,
LC50 values ranged from 41 to 3,598 (ig Pb/L and no Pb toxicity occurred in three highly
alkaline waters (Esbaughetal.. 2011). In the 2006 Pb AQCD, the 96-hr LC50 values in
fathead minnow ranged from 810->5,400 (ig Pb/L in varying pH and hardness (U.S.
EPA. 2006c).
Decreased survival is also a function of age of the fish. Thirty day LC50 values for larval
fathead minnows ranged from 4.5 to 1,903 (ig Pb/L in varying concentrations of DOC,
CaSO4 and pH (Grosell et al., 2006a). In a recent study of rainbow trout fry at 2 to 4
weeks post swim-up, the 96-hour LC50 was 120 (ig Pb/L at a hardness of 29 mg/L as
CaCO3, a value much lower than in testing with older fish (Mebane et al. 2008). In the
same study, two chronic (>60 day) tests were conducted with rainbow trout and the
NOECs for survival were 24 and 87 (ig Pb/L and the LOECs were 54 and 125 (ig Pb/L,
respectively. In contrast to aqueous exposures, 30 day dietary studies with rainbow trout
fathead minnow, and channel catfish fed a live diet of L. variegatus contaminated with Pb
showed no statistically significant effects on survival (Erickson et al.. 2010).
Freshwater fish are less sensitive to Pb than freshwater invertebrates, however, recent
studies have highlighted the importance of considering pH, hardness, and additional
modifying factors in assessing toxicity since effects can vary over several orders of
magnitude. Fish mortalities occur above the concentrations of Pb encountered in U.S.
surface waters although, in some cases, the observed effects may be just above the upper
measured range of Pb in some aquatic environments (Table 6-2). Furthermore, although
LC50 values are the most commonly reported, effects are occurring at lower
concentrations. A more relevant indication of exposure impacts would be an LCi0 or
LOEC, however, these values are not always provided. The evidence is sufficient to
conclude that there is a causal relationship between Pb exposures and survival in
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freshwater invertebrates and vertebrates. The evidence is inadequate to conclude that
there is a causal relationship between Pb exposures and survival in freshwater plants.
6.4.12.4 Neurobehavioral Effects-Freshwater Biota
Evidence from laboratory studies and limited data from field studies reviewed in this
chapter, in the draft Ambient Aquatic Life Water Quality Criteria for Lead report, which
updates the 1985 Ambient Water Quality Criteria for Lead (U.S. EPA. 1985). and in
previous Pb AQCDs have shown effects of Pb on neurological endpoints in aquatic
animal taxa. These include changes in behaviors that may decrease the overall fitness of
the organism such as avoidance responses, decreased ability of an organism to capture
prey or escape predators, and alterations in feeding behaviors. Evidence of alteration in
behaviors at the level of the organism is a potential endpoint for effects at population or
community levels of biological organization (U.S. EPA. 2003a).
In the 1977 Pb AQCD behavioral impairment of a conditioned response (avoidance of a
mild electric shock) in goldfish was observed as low as 70 (ig Pb/L (Weir and Hine.
1970). In the 2006 Pb AQCD several studies were reviewed in which Pb was shown to
affect predator-prey interactions, including alteration in prey size choice and delayed prey
selection in juvenile fathead minnows following 2-week pre-exposure to 500 (ig Pb/L
(Weber. 1996). In limited studies available on worms, snails, tadpoles, hatchling turtles
and fish there is evidence that Pb may affect the ability to escape or avoid predation. For
example, in the tubificid worm T. tubifex the 96 hour EC50 for immobilization was
42 (ig Pb/L (Khangarot 1991). Some organisms exhibit behavioral avoidance while
others do not seem to detect the presence of Pb (U.S. EPA. 2006c). Additional behavioral
endpoints reported in the draft Ambient Aquatic Life Water Quality Criteria for Lead
document include an EC50 of 140 (ig Pb/L for feeding inhibition in the freshwater
cladoceran C. dubia and deceased learning acquisition in bullfrogs at 500 (ig Pb/L
(31.51 (ig Pb/L adjusted to a total hardness of 50 mg/L CaCO3). All of these effects occur
at concentrations that exceed Pb concentration values found in surveys of U.S. surface
waters although within the range of Pb detected near some mining-disturbed areas (Table
2-11).
Recent information since the 2006 Pb AQCD provides evidence for Pb impacts on
behaviors that may affect feeding and predator avoidance in freshwater environments at
concentrations near the range of Pb detected in U.S. surface waters (Table 6-2 and
Section 6.4.5.3). Prey capture ability was decreased in 10 day old fathead minnow larvae
born from adult fish exposed to 120 (ig Pb/L for 300 days, then subsequently tested in a
21-day breeding assay (Mager et al. 2010). Another study in fish reported effects at
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low (ig Pb/L concentration, however, the findings are not considered as strong evidence
for causality since exposure concentrations in water were not analytically verified.
Specifically, zebrafish embryos exposed nominally to concentrations of Pb (2.0 to
6.0 (ig Pb/L) until 24 hours post-fertilization and then subsequently tested as larvae
exhibited decreased startle response time and altered pattern of escape swimming (Rice et
al.. 2011). In adult fish raised from the exposed embryos (6.0 (ig Pb/L), the ability to
detect visual contrast was degraded. Although this study was conducted with nominal
concentration of Pb in media, uptake of Pb by embryos was quantified and more Pb was
measured in tissues of embryos exposed to the higher concentration of Pb when
compared to the lower exposure concentration. Additional studies are needed in fish to
support these initial findings of effects on ecologically relevant behavioral impairments.
Findings in laboratory animals support the limited evidence for neurobehavioral effects
of Pb in freshwater invertebrates and vertebrates. In animal toxicological studies Pb
induced changes in learning and memory (Section 4.3.2.3). New behaviors induced by
exposure to Pb reviewed in Chapter 4 that are relevant to effects of Pb observed in
freshwater systems include effects on visual and auditory sensory systems and changes in
structure and function of neurons and supporting cells in the brain. Mechanisms that
include the displacement of physiological cations, oxidative stress and changes in
neurotransmitters and receptors are also reviewed. Central nervous system effects in fish
recognized in previous Pb AQCDs include effects on spinal neurons and brain receptors.
New evidence from this review identifies the MAPKs ERK1/2 and p38MAPK as possible
molecular targets for Pb neurotoxicity in catfish (Leal et al.. 2006). Evidence in terrestrial
ecosystems (Section 6.3.12.4) is not as extensive, but it is highly coherent with findings
in aquatic ecosystems. Overall, the evidence from available studies on neurobehavioral
effects of Pb in aquatic systems is limited, but sufficient to conclude that a causal
relationship is likely to exist between Pb exposures and neurobehavioral effects in aquatic
invertebrates and vertebrates.
6.4.12.5 Hematological Effects-Freshwater Biota
Hematological responses are commonly reported effects of Pb exposure in aquatic
invertebrates and vertebrates. Anemia was recognized as a symptom of chronic Pb
poisoning in fish in the 1977 Pb AQCD and has been subsequently reported in various
fish species using common hematological endpoints (e.g., red blood cell counts,
hematocrit, Hb concentrations) (Mager. 2012). In the 1986 Pb AQCD, hematological
effects of Pb exposure on fish included decrease in red blood cells and inhibition of
ALAD (U.S. EPA. 1986b). Inhibition of ALAD activity under various test conditions is
reported in Table 6 of the the draft Ambient Aquatic Life Water Quality Criteria for Lead
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report, for freshwater fish species (O. mykiss (Rainbow Trout), S. fontinalis (Brook
Trout), C. auratus (Goldfish) and Lepomis gibbosus (Pumpkinseed)) (U.S. EPA. 2008b).
In these studies, Rainbow Trout was the most sensitive with inhibition of ALAD activity
reported in multiple studies within the upper range of Pb in surface waters of the U.S.
(median 0.5 ng Pb/L, range 0.04 to 30 jig Pb/L) (U.S. EPA. 2006c).
Laboratory studies with freshwater invertebrates have also indicated considerable species
differences in ALAD activity in response to Pb. For example, the concentration at which
50% ALAD inhibition was measured in the freshwater gastropod B. glabrata (23 to
29 (ig Pb/L) was much lower than in the freshwater oligochaete L. variegatus
(703 (ig Pb/L) based on nominal exposure data (Aisemberg et al.. 2005).
Findings in laboratory studies are additionally supported by evidence from field-collected
organisms providing coherence to the observations of Pb effects on ALAD activity. In
environmental assessments of metal-impacted habitats, ALAD is a recognized biomarker
of Pb exposure (U.S. EPA. 2006c). ALAD activity is negatively correlated with total Pb
concentration in freshwater bivalves, and lower ALAD activity has been correlated with
elevated blood Pb levels in field-collected fish as well. Further evidence from the
2006 Pb AQCD and this review of Pb effects on ALAD enzymatic activity, including
effects in bacteria, amphibians and additional field and laboratory studies on freshwater
fish, confirms that the decreased activity in this enzyme is an indicator for Pb exposure
across a wide range of taxa and that a common mode of action is likely for invertebrates
and vertebrates. The finding that the hematological system is a target for Pb in natural
systems is also supported by some evidence of Pb-induced alterations of serum profiles
and changes in white blood cell counts in fish (U.S. EPA. 2006c) and amphibians. This
evidence is strongly coherent with evidence from terrestrial vertebrates
(Section 6.3.12.5). It is also coherent with observations from human epidemiologic and
animal toxicology studies (Section 4.7) where there is consistent evidence that exposure
to Pb induces adverse effects on hematological endpoints, including altered heme
synthesis mediated through decreased ALAD and ferrochelatase activities, decreased red
blood cell survival and function, and increased red blood cell oxidative stress. The overall
weight of epidemiologic and toxicological evidence for humans was sufficient to
conclude that a causal relationship exists between exposure to Pb and hematological
effects (Section 4.7).
Based on observations in freshwater organisms and additionally supported by findings in
terrestrial systems, saltwater invertebrates (Section 6.4.21.5). and by toxicological and
epidemiologic evidence on human health effects, evidence is sufficient to conclude that
there is a causal relationship between Pb exposures and hematological effects in
freshwater vertebrates. Evidence is sufficient to conclude that a causal relationship is
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likely to exist between Pb exposures and hematological effects in freshwater
invertebrates.
6.4.12.6 Physiological Stress-Freshwater Biota
Building on the body of evidence presented in the 2006 Pb AQCD (U.S. EPA. 2006c)
recent studies provide consistent and coherent evidence of upregulation of antioxidant
enzymes and increased lipid peroxidation associated with Pb exposure within one or two
orders of magnitude above current or ambient conditions in many species of freshwater
plants, invertebrates and vertebrates. A few studies provide evidence of effects at
concentrations of Pb encountered in some sediments of the U.S. (Table 6-2). In aquatic
plants, increases of antioxidant enzymes with Pb exposure occur in algae, mosses, and
floating and rooted aquatic macrophytes. Most available evidence for antioxidant
responses in aquatic plants are from laboratory studies lasting from 2 to 7 days and at
concentrations higher than typically found in the environment. However, data from
transplantation experiments from non-polluted to polluted sites indicate that elevated
enzyme activities are associated with Pb levels measured in sediments. For example, the
freshwater macrophyte Myriophyllum quitense exhibited elevated antioxidant enzyme
activity (glutathione-S-transferase, glutathione reductase, peroxidase) following
transplantation in anthropogenically polluted areas containing elevated Pb concentrations.
These were correlated with sediment Pb concentrations in the range of 5 to 23 mg Pb/g
dry weight (Nimptsch et al., 2005).There is evidence for antioxidant activity in response
to Pb exposure in freshwater invertebrates (i.e., bivalves). Markers of oxidative damage
are also observed in fish, amphibians and mammals in laboratory studies. Across all
organisms, there are differences in the induction of antioxidant enzymes that appear to be
species-dependent.
Additional stress responses to Pb in a few aquatic invertebrates have been reported since
the 2006 Pb AQCD, and included elevated heat shock proteins, osmotic stress, lowered
metabolism and decreased glycogen levels associated with Pb exposure. Although these
stress responses are correlated with Pb exposure, they are non-specific and may be altered
with exposure to any number of environmental stressors. Heat shock protein induction
has been observed in zebra mussels exposed to 500 (ig Pb/L for 10 weeks (Singer et al..
2005). Crayfish exposed for 14 days to 500 (ig Pb/L exhibited a Pb-induced
hypometabolism under conditions of environmental hypoxia in the presence of the metal
(Morris et al., 2005). Glycogen levels in the freshwater snail B. glabrata were
significantly decreased following 96-hour exposures at 50 (ig/L and higher (Ansaldo et
al.. 2006).
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Evidence for stress responses observed in freshwater plants, invertebrates and vertebrates
is coherent with findings in terrestrial species (Section 6.3.12.6) and saltwater
invertebrates (Section 6.4.21.6). It is also coherent with evidence from human and
experimental animal studies of oxidative stress following impairment of normal metal ion
functions (Section 4.2.4). Upregulation of antioxidant enzymes and increased lipid
peroxidation are considered to be reliable biomarkers of stress, and provide evidence that
Pb exposure induces a stress response in those organisms which may increase
susceptibility to other stressors and reduce individual fitness. Evidence is sufficient to
conclude that a causal relationship is likely to exist between Pb exposures and
physiological stress in freshwater aquatic plants, invertebrates and vertebrates.
6.4.12.7 Community and Ecosystem Level Effects-Freshwater
Biota
Most direct evidence of community- and ecosystem-level effects in freshwater systems is
from heavily contaminated sites where Pb concentrations are higher than typically
observed environmental concentrations for this metal. Impacts of Pb on aquatic habitats
that receive runoff from contaminated areas have been studied for several decades. For
aquatic systems, the literature focuses on evaluating ecological stress from Pb originating
from urban and mining effluents rather than atmospheric deposition. Ecosystem-level
field studies are complicated by the confounding of Pb exposure with other factors such
as the presence of trace metals and acidic deposition and by the variability inherent in
natural systems. In natural systems, Pb is often found co-existing with other stressors, and
observed effects may be due to cumulative toxicity.
In laboratory studies and simulated ecosystems, where it is possible to isolate the effect
of Pb, this metal has been shown to alter competitive behavior of species, predator-prey
interactions and contaminant avoidance. These dynamics may change species abundance
and community structure at higher levels of ecological organization. Uptake of Pb into
aquatic and terrestrial organisms and subsequent effects on mortality, growth,
developmental and reproductive endpoints at the organism level are expected to have
ecosystem-level consequences, and thus provide consistency and plausibility for causality
in ecosystem-level effects.
In aquatic ecosystems, field studies reviewed in the 2006 Pb AQCD (summarized in
Table AX7-2.5.2) and more recent studies report reductions of species abundance,
richness or diversity. This is particularly the case for benthic macroinvertebrate
communities where sources of Pb were mining or urban effluents, and Pb coexisted with
other metals. The results often indicate a correlation between the presence of one or more
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metals and the negative effects observed. For example, in the 2006 Pb AQCD, the Coeur
d'Alene River watershed in Idaho, U.S. was used as a case study for Pb effects at the
population and community level. Significant negative correlations were observed
between Pb in water column and total taxa richness and EPT taxa richness in the river. In
a simulated aquatic microcosm a reduction in abundance and richness of protozoan
species was observed with increasing Pb concentration from 50 to 1,000 (ig Pb/L
(Fernandez-Leborans and Antonio-Garcia. 1988).
Since the 2006 Pb AQCD, there is further evidence for effects of Pb in sediment-
associated communities. Sediment-bound Pb contamination appears to differentially
affect members of the benthic invertebrate community, potentially altering ecosystem
dynamics in small urban streams (Kominkova and Nabelkova, 2005). Although surface
water Pb concentrations in monitored streams were determined to be very low,
concentrations of the metal in sediment were high enough to pose a risk to the benthic
community (e.g., 34 to 101 mg Pb/kg). These risks were observed to vary with benthic
invertebrate functional feeding group, with collector-gatherer species exhibiting larger
body burdens of heavy metals than benthic predators and collector-filterers.
In a recent study conducted since the 2006 Pb AQCD, changes to aquatic plant
community composition have been observed in the presence of elevated surface water Pb
concentrations at three lake sites impacted by mining effluents. The site with highest Pb
concentration (103-118 (ig Pb/L) had lowest number of aquatic plant species when
compared to sites with lower Pb concentrations (78-92 (ig Pb/L) (Mishraet al.. 2008).
Certain types of plants such as rooted and submerged aquatic plants may be more
susceptible to aerially deposited Pb resulting in shifts in Pb community composition.
High Pb sediment concentrations are linked to shifts in amphipod communities inhabiting
plant structures.
Avoidance response to Pb exposure varies widely in different species and this could
affect community composition and structure and species abundance. For example, frogs
and toads lack avoidance response while snails and fish avoid higher concentrations of Pb
(U.S. EPA. 2006c).
In the Annex to the 2006 Pb AQCD, the Coeur d'Alene River basin in Idaho was
presented as a case study for a watershed impacted by Pb and other metals. A significant
negative correlation was observed between Pb in water column (0.5 to 30 (ig Pb/L) and
total taxa richness, EPT taxa richness, and the number of metal-sensitive mayfly species
(Maret et al., 2003). Additional lines of evidence including mine density, metal
concentrations, and bioaccumulation in caddisfly tissue were included. Since the
2006 Pb AQCD, additional research at this site and model development has resulted in
further characterization of the effects of Pb on waterfowl and other aquatic organisms in
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this heavily contaminated ecosystem. Mean Pb concentrations in Coeur d'Alene sediment
range from 14 to 5,009 mg Pb/kg dry weight (Spears et al.. 2007). Modeling of sediment
and Pb levels in waterfowl predict a sediment Pb effects range of 147-944 mg Pb/kg dry
weight and a mortality effects level of 1,652 mg/kg dry weight (Spears et al.. 2007). In a
6-week feeding study with mallard (Anas platyrhynchos) ducklings, ingestion of
Pb-contaminated sediments from the Coeur d' Alene basin was shown to result in
decreased brain growth and altered brain chemistry (Douglas-Stroebel et al., 2004). These
findings support previous observations of altered behavior and hematological,
hepatotoxic, and histopathological endpoints in waterfowl from Lake Coeur d'Alene that
ingest Pb contaminated sediments and vegetation during feeding.
In addition to the evidence from microcosm and field studies presented above, effects on
reproduction (Section 6.4.12.1). growth (Section 6.4.12.2) and survival (Section 6.4.12.3)
have been clearly demonstrated in freshwater species. These endpoints can have effects at
the population-level and community-level of biological organization which may lead to
ecosystem-level impacts. Although the evidence is strong for effects of Pb on growth,
reproduction and survival in certain species in experimental settings at or near the range
of Pb concentrations reported in surveys of U.S. freshwater systems, considerable
uncertainties exist in generalizing effects observed under small-scale, particular
conditions up to predicted effects at the ecosystem level of biological organization. In
many cases it is difficult to characterize the nature and magnitude of effects and to
quantify relationships between ambient concentrations of Pb and ecosystem response due
to presence of multiple stressors, variability in field conditions and to differences in Pb
bioavailability at that level of organization. Bioavailability of Pb is influenced by pH,
alkalinity, total suspended solids, and DOC among other factors and can vary greatly in
natural environments. Nevertheless, evidence of ecosystem effects in aquatic systems is
coherent with similar evidence in terrestrial systems, and based on the cumulative
evidence from laboratory studies and field observations, a causal relationship is likely to
exist between Pb exposures and the alteration of species richness, species composition
and biodiversity in freshwater aquatic ecosystems.
6.4.13 Introduction to Bioavailability and Biological Effects of Pb in Saltwater
Ecosystems
Saltwater ecosystems include salt marsh, estuaries, embayments, beaches, and other
coastal areas; and encompass a range of salinities from just above that of freshwater to
that of seawater. These ecosystems may receive Pb contributions from direct atmospheric
deposition and/or via runoff from terrestrial systems. A range of 0.01 to 27 (ig Pb/L
including coastal areas, estuaries and open ocean was reported by Sadiq (1992) with the
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higher values associated with sites involving human activity, however, these values are
not specific to the U.S. (Table 6-2). In an earlier publication, levels of Pb in the North
Atlantic and North Pacific surface waters ranged from 0.005 to 0.4 (ig Pb/L but the range
of values in coastal waters and estuaries were approximately equal to the range of Pb in
freshwater (Leland and Kuwabara. 1985). Additional information on Pb levels in water is
available in Sections 6.2.3 and 2.6. The 2006 Pb AQCD provided an overview of
regulatory considerations for water and sediments in addition to consideration of
biological effects and major environmental factors that modify the response of marine
organisms to Pb exposure. Regulatory guidelines for Pb in saltwater have not changed
since the 2006 Pb AQCD and are summarized below. This section is followed by new
information on bioavailability and biological effects of Pb in saltwater since the
2006 Pb AQCD.
The most recent ambient water quality criteria for Pb in saltwater were released in 1985
(U.S. EPA. 1985) by the EPA Office of Water which employed empirical regressions
between observed toxicity and water hardness to develop hardness-dependent equations
for acute and chronic criteria. These criteria are published pursuant to Section 304(a) of
the Clean Water Act and provide guidance to states and tribes to use in adopting water
quality standards for the protection of aquatic life and human health in surface water. The
ambient water quality criteria for Pb are currently expressed as a criteria maximum
concentration (CMC) for acute toxicity and criterion continuous concentration (CCC) for
chronic toxicity (U.S. EPA. 2009b). In saltwater, the CMC is 210 (ig Pb/L and the CCC
is 8.1 (ig Pb/L. The 2006 Pb AQCD summarized two approaches for establishing
sediment criteria for Pb based on either bulk sediment or equilibrium partitioning as
reviewed in the present document in Section 6.4.2.
In the following sections, recent information available since the 2006 Pb AQCD on Pb in
marine and estuarine ecosystems will be presented. Throughout the sections, brief
summaries of conclusions from the 1977 Pb AQCD (U.S. EPA. 1977). the 1986 Pb
AQCD (U.S. EPA. 1986b) and the 2006 Pb AQCD (U.S. EPA. 2006b) are included
where appropriate. Recent research on the bioavailability and uptake of Pb into saltwater
organisms including plants, invertebrates and vertebrates is presented in Section 6.4.14.
Toxicity of Pb to marine flora and fauna including growth, reproductive and
developmental effects (Section 6.4.15) are followed with data on exposure and response
of saltwater organisms (Section 6.4.16). Responses at the community and ecosystem
levels of biological organization are reviewed in Section 6.4.17 followed by
characterization of sensitivity and vulnerability of saltwater ecosystem components
(Section 6.4.18) and a discussion of ecosystem services (Section 6.4.19). The saltwater
sections conclude with a synthesis of the new data for Pb effects on saltwater plants,
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invertebrates and vertebrates (Section 6.4.20) and causal determinations based on
evidence from previous Pb AQCDs and recent studies (Section 6.4.21).
6.4.14 Bioavailability of Pb in Saltwater Systems
Bioavailability was defined in the 2006 Pb AQCD as "the proportion of a toxin that
passes a physiological membrane (the plasma membrane in plants or the gut wall in
animals) and reaches a target receptor (cytosol or blood)". In 2007, EPA took cases of
bioactive adsorption into consideration and revised the definition of bioavailability as
"the extent to which bioaccessible metals absorb onto, or into, and across biological
membranes of organisms, expressed as a fraction of the total amount of metal the
organism is proximately exposed to (at the sorption surface) during a given time and
under defined conditions" (U.S. EPA. 2007c)
Factors affecting bioavailability of Pb to marine organisms are the same as those in
freshwater systems (Sections 6.4.2 and 6.4.4). However, although routes of exposure and
physiological mechanisms for storage and excretion influence uptake of metals by all
organisms, they may be different in marine organisms, particularly for ion transport
mechanisms (Niyogi and Wood. 2004). Marine environments are characterized by higher
levels of ions, such as Na+, Ca2+, and Mg2+, which compete for potential binding sites on
biotic ligands such as gills, thereby generally reducing the effective toxicity of metal ions
as compared to freshwater environments. However, because the concentrations of these
ions are relatively constant, bioavailability may be more predictable in marine systems
that are little influenced by freshwater than in freshwater systems, varying mostly with
amount and type of dissolved organic matter. In estuaries and embayments, changing
salinities and proximity to anthropogenic loading of pollutants add to the complexity of
predicting Pb speciation in these dynamic systems. BLMs (Figure 6-3) now being
developed for marine organisms are functionally similar to those applied to freshwater
organisms (Section 6.4.4).
Although in freshwater systems the presence of humic acid is considered to
reduce the bioavailable fraction of metals in freshwater, there is evidence that
DOC/DOM does not have the same effect on free Pb ion concentration in marine systems
(see Section 6.4.2.4 for detailed discussion). For the sea urchin P. lividus, the presence of
humic acid increased both the uptake and toxicity of Pb possibly by enhancing uptake of
Pb via membrane Ca2+ channels (Sanchez-Marin et al. 2010b). This also was observed in
the marine diatom Thalassiosira weissflogii, where humic acids absorbed to cell surfaces
increased metal uptake (Sanchez-Marin et al., 2010b). Formation of a ternary complex
that is better absorbed by biological membranes was another proposed mechanism that
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could describe the increased bioavailability to marine invertebrates of Pb bound to humic
acid (Sanchez-Marin et al.. 2007).
Sanchez-Marin et al. (2011) subsequently have shown that different components of DOM
have different effects on Pb bioavailability in marine systems. Their initial research using
Aldrich humic acid found that increasing humic acid concentrations increased Pb uptake
by mussel gills and increased toxicity to sea urchin larvae in marine environments
(Sanchez-Marin et al.. 2007). In contrast, a subsequent investigation found that fulvic
acid reduced Pb bioavailability in marine water (Sanchez-Marin et al.. 2011). The
contradictory effects of different components of DOM on marine bioavailability likely
reflect their distinct physico-chemical characteristics. More hydrophobic than fulvic acid,
humic acid may adsorb directly with cell membranes and enhance Pb uptake through
some (still unidentified) mechanism (Sanchez-Marin et al.. 2011). Pb AVS-measurements
were also determined to accurately predict uptake by mussels (Mytilus sp.) in the
presence of 2.5 to 20 mg/L fulvic acid (Sanchez-Marin et al.. 2011). However, the effects
of DOM on Pb bioavailability to mussels were underpredicted by AVS Pb concentration
measurements, potentially as a result of adsorption of DOM-Pb complexes.
Based on the above, BLMs (see Section 6.4.4 and Figure 6-3) used to predict
bioavailability of Pb to aquatic organisms (Pi Toro et al.. 2005). may require
modifications for application to marine organisms. Of particular importance is the finding
that in marine aquatic systems, surface water DOM was found to increase (rather than
decrease) uptake of Pb by fish gill structures, potentially through the alteration of
membrane Ca2+ channel permeability. Veltman et al. (2010) proposed integrating BLM
and bioaccumulation models in order to more accurately predict metal uptake by fish and
invertebrates, and calculated metal absorption efficiencies for marine fish species from
both types of models. They noted that affinity constants for Ca2+, Cd, Cu, Na, and Zn
were highly similar across different aquatic species, including fish and invertebrates
(Veltman et al.. 2010). These findings suggest that the BLM can be integrated with
bioaccumulation kinetics to account for both environmental chemical speciation and
biological and physiological factors in both marine and freshwater systems.
6.4.14.1 Saltwater Plants and Algae
In the 1977 Pb AQCD, the cordgrass Spartina alterniflora was found to reduce by a small
amount the quantity of Pb in sediments (U.S. EPA. 1977). Limited data on marine algal
species reviewed in the 1986 Pb AQCD and 2006 Pb AQCD provided additional
evidence for Pb uptake. Recent data available since the 2006 Pb AQCD includes Pb
bioaccumulation studies conducted with five species of marine algae, (Tetraselmis chuii,
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Rhodomonas satinet, Chaetoceros sp., Isochrysis galbana, and Nannochloropsis
gaditana). In this study it was demonstrated that bioaccumulation rates varied with
species following 72-hour exposure to Pb. /. galbana accumulated the lowest
concentrations of Pb (0.01 and 0.6 pg Pb/cell at water concentrations of 51 and
6,348 (ig Pb/L), while Chaetoceros sp. was observed to be the most efficient Pb
bioaccumulator, adsorbing 0.04 and 54 pg Pb/cell at 51 and 6,348 (ig Pb/L (Debelius et
al.. 2009).
Recent uptake studies of Pb in plants associated with marine environments are also
available. The roots of two salt marsh species, Sarcocornia fruticosa and Spartina
maritima, significantly accumulated Pb, to maximum concentrations of 2,870 mg Pb/kg
and 1,755 mg Pb/kg, respectively (Caetano et al., 2007). Roots had similar isotopic
signature to those of sediments in vegetated zones indicating that Pb uptake by plants
reflects the input in sediments. BCFs for Pb in root tissue from mangrove tree species
range between 0.09 and 2.9, depending on the species and the habitat, with an average
BCF of 0.84. The average BCF for mangrove species leaf tissue was considerably less
(0.11), as these species are poor translocators of Pb (MacFarlane et al.. 2007).
6.4.14.2 Saltwater Invertebrates
Uptake and subsequent bioaccumulation of Pb in marine invertebrates varies greatly
between species and across taxa as previously characterized in the 2006 Pb AQCD. This
section expands on the findings from the 2006 Pb AQCD on bioaccumulation and
sequestration of Pb in saltwater invertebrates. In the case of invertebrates, Pb can be
bioaccumulated from multiple sources, including the water column, sediment, and dietary
exposures, and factors such as proportion of bioavailable Pb, lifestage, age, and
metabolism can alter the accumulation rate. In this section, new information on Pb uptake
and subsequent tissue and subcellular distribution will be considered, followed by a
discussion on dietary and water routes of exposure and strategies for detoxification of Pb
in marine invertebrates.
In marine invertebrates, sites for Pb accumulation include the gill and digestive
gland/hepatopancreas. The gills were the main sites of Pb accumulation in pearl oyster,
Pinctada fucata followed by mantle, in 72-hour exposures to 103.5 (ig Pb/L (Jing et al..
2007). Following a 10-day exposure to 2,500 (ig Pb/L as Pb nitrate, accumulation of Pb
was higher in gill than digestive gland ofMytilus edulis: after a 10 day depuration, Pb
content was decreased in the gills and digestive gland of these mussels (Einsporn et al..
2009). In blue crabs, Callinectes sapidus, collected from a contaminated and a clean
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estuary in New Jersey, U.S., the hepatopancreas was found to be the primary organ for Pb
uptake (Reichmuth et al.. 2010).
There is more information now on the cellular and subcellular distribution of Pb in
invertebrates than there was at the time of writing the 2006 Pb AQCD. Specifically,
localization of Pb at the ultrastructural level has been assessed in the marine mussel
(M edulis) through an antibody-based detection method (Einsporn et al.. 2009; Einsporn
and Koehler. 2008). Dissolved Pb was detected mainly within specific lysosomal
structures in gill epithelial cells and digestive gland cells and was also localized in nuclei
and mitochondria. Transport of Pb is thought to be via lysosomal granules associated
with hemocytes (Einsporn et al.. 2009). In the digestive gland of the variegated scallop
(Chlamys varia), Pb was also mainly bound to organelles, (66% of the total metal burden)
(Bustamante and Miramand. 2005). In the digestive gland of the cephalopod Sephia
officinalis, (cuttlefish) most of the Pb was found in the organelles (62%) (Bustamante et
al.. 2006). In contrast, only 7% of Pb in the digestive gland of the octopus (Octopus
vulgaris) was associated with the fraction containing nuclei, mitochondria, lysosome and
microsomes: the majority of Pb in this species was found in cytosolic proteins (Raimundo
et al.. 2008).
Metian et al. (2009) investigated the uptake and bioaccumulation of 210Pb in variegated
scallop and king scallop to determine the major accumulation route (seawater or food)
and then assess subsequent tissue distribution. Dietary Pb from phytoplankton in the diet
was poorly assimilated (<20%) while more than 70% of Pb in seawater was retained in
the tissues. In seawater, 210Pb was accumulated more rapidly in variegated scallop than
king scallop and soft tissue distribution patterns differed between the species. Variegated
scallop accumulated Pb preferentially in the digestive gland (50%) while in king scallop,
Pb was equally distributed in the digestive gland, kidneys, gills, gonad, mantle, intestine,
and adductor muscle with each tissue representing 12-30% of 210Pb body load. An
additional test with Pb-spiked sediment in king scallop showed low bioaccumulation
efficiency of Pb from spiked sediment.
Recently, several studies have attempted to establish biodynamic exposure assessments
for various contaminants. In an in situ metal kinetics field study with the mussel
M. galloprovincialis, simultaneous measurements of metal concentrations in water and
suspended particles with mussel biometrics and physiological indices were conducted to
establish uptake and excretion rates in the natural environment (Casas et al., 2008). The
mean logarithmic ratio of metal concentration in mussels (ng/kg of wet-flesh weight) to
metal concentration in water (ng/L) was found to be 4.3 inM galloprovincialis, based on
the rate constants of uptake and efflux in a series of transplantation experiments between
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contaminated and clean environments. Equilibrium concentrations of Pb in mussels
leveled out at approximately 30 days with a concentration of 6.7 mg Pb/kg.
The protective barrier against Pb toxicity formed by the egg structure in some
invertebrates was recognized in the 2006 Pb AQCD. Consideration of toxicity of Pb to
embryos that develop surrounded by a protective egg shell has been expanded since the
2006 Pb AQCD. In a study with cuttlefish (S. officinalis) eggs, radioisotopes were used to
assess the permeability of the egg to Pb at low exposure concentrations (210Pb activity
concentration corresponding to 512 (ig/L Pb) (Lacoue-Labarthe et al., 2009). Retention
and diffusion properties of the cuttlefish egg change throughout the development of the
embryo and since the eggs are fixed on substrata in shallow coastal waters they may be
subject to both acute and chronic Pb exposures. In the radiotracer experiments, 210Pb was
never detected in the internal compartments of the egg during the embryonic
development stage, while concentrations in the eggshell increased throughout the 48-day
exposure. These results are consistent with a study of cuttlefish eggs collected from
natural environments in which Pb was only detected in the eggshell. These studies
indicate that the cuttlefish egg provides a protective barrier from Pb toxicity (Miramand
et al.. 2006).
Aquatic invertebrate strategies for detoxifying Pb were reviewed in the 2006 Pb AQCD
and include sequestration of Pb in lysosomal-vacuolar systems, excretion of Pb by some
organisms, and deposition of Pb to molted exoskeleton. Molting of the exoskeleton can
result in depuration of Pb from the body (see Knowlton et al. (1983) and Anderson et al.
(1997). as cited in the 2006 Pb AQCD). New research has provided further evidence of
depuration of Pb via molting in invertebrates. Mohapatra et al. (2009) observed that Pb
concentrations in body tissues were lower in the newly molted mud crabs (Scylla serrata)
than in the pre-molt, hard-shelled crabs. However, the carapace of hard shelled crabs had
lower concentrations of Pb than the exuvium of the soft shell crabs, leading the authors to
speculate that some of the metal might be partially excreted during the molting process,
rather than entirely through shedding of the previous exoskeleton. Bergey and Weis
(2007) showed that differences in the proportion of Pb stored in exoskeleton and soft
tissues changed during intermolt and immediate postmolt in two populations of fiddler
crabs (Ucapugnax) collected from New Jersey. One population from a relatively clean
estuary eliminated an average of 56% of Pb total body burden during molting while
individuals from a site contaminated by metals eliminated an average of 76% of total Pb
body burden via this route. Pb distribution within the body of crabs from the clean site
shifted from exoskeleton to soft tissues prior to molting. The authors observed the
opposite pattern of Pb distribution in fiddlers from the contaminated site where larger
amounts of Pb were depurated in the exoskeleton. The exact dynamics of Pb depuration
through molting in crabs are thus still not completely characterized.
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6.4.14.3 Saltwater Vertebrates
Saltwater Fish
In comparison to freshwater fish, fewer studies have been conducted on Pb uptake in
marine fish. Since marine fish drink seawater to maintain osmotic homeostasis, Pb can be
taken up via both gills and intestine (Wang and Rainbow. 2008). Pb was significantly
accumulated in gill, liver, plasma, kidney, rectal gland, intestine, skin, and muscle of the
elasmobranch spotted dogfish (Scyliorhinus canicula) exposed to 2,072 (ig Pb/L for one
week (De Boeck et al., 2010). In contrast to Pb distribution patterns in freshwater
teleosts, high Pb concentrations were present in this species in the skin and rectal gland.
Egg cases of the spotted dogfish exposed to 210Pb in seawater for 21 days, accumulated
radiolabeled Pb rapidly and the metal was subsequently detected in embryos indicating
the permeability of shark eggs to Pb in coastal environments (Jeffree et al., 2008). A
study of Pb bioaccumulation in five marine fish species (Chloroscombrus chrysurus,
Sardinella aurita, Ilisha africana, Galeoides decadactylus, Caranx latus) found that
C. chrysurus was an especially strong bioaccumulator, yielding Pb concentrations of 6 to
10 mg Pb/kg (Gnandi et al., 2006). However, C. chrysurus metal content was not
correlated to the Pb concentrations along the mine tailings gradient from which they were
collected (8.5 and 9.0 \ig Pb/L for minimum and maximum tissue concentrations,
respectively). This lack of correlation was also observed for fish species that were
considered to be weaker Pb bioaccumulators, indicating that unidentified sources of Pb
(e.g., in sediments or in dietary sources) may be contributing to Pb uptake by marine fish.
In grunt fish H. scudderi, exposed to Pb via dietary uptake through a simulated marine
food chain, mean total Pb body burden increased from 0.55 to 3.32 mg Pb/kg in a 42-day
feeding study (Soto-Jimenez et al.. 201 Ib). Pb was accumulated to the highest relative
concentration in liver with less than 3% of total Pb accumulated in gills. Most of the Pb
based on total body mass was accumulated in skeleton, skin, scales and muscle.
The 2006 Pb AQCD considered detoxification mechanisms in fish including mucus
production and Pb removal by shedding of scales in which Pb is chelated with keratin.
Since the 2006 Pb AQCD review, additional Pb detoxification mechanisms in marine fish
have been further elucidated. Mummichog (Fundulus heteroclitus) populations in metal-
polluted salt marshes in New York exhibited different patterns of intracellular
partitioning of Pb although body burden between sites was not significantly different
(Goto and Wallace. 2010). Mummichogs at more polluted sites stored a higher amount of
Pb in metal rich granules as compared to other detoxifying cellular components such as
heat-stable proteins, heat-denaturable proteins and organelles.
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Marine Mammals
Studies that consider uptake of Pb in aquatic mammals are limited. Kannan et al. (2006)
compared trace element concentrations in livers of free-ranging sea otters (Enhydra lutris
nereis) found dead along the California coast. They detected Pb in all individuals
sampled (N=80) in a range of 0.019 to 1.06 mg Pb/kg. The otters were classified by cause
of death (infectious causes, non-infectious causes, emaciated condition) and trace element
patterns of tissue distribution were compared. Livers from emaciated otters had
significantly elevated levels of Pb compared to non-diseased individuals.
6.4.14.4 Marine Food Web
As discussed in Section 6.4.4.4 trophic transfer of Pb through aquatic food chains was
considered to be negligible in the 2006 Pb AQCD (U.S. EPA. 2006c). Measured
concentrations of Pb in the tissues of aquatic organisms were found to be generally higher
in algae and benthic organisms and lower in higher trophic-level consumers, indicating
that Pb was bioaccumulated but not biomagnified (U.S. EPA. 2006c; Eisler. 2000).
Recent literature since the 2006 Pb AQCD, provides evidence of the potential for Pb to
be transferred in marine food webs while other studies indicate Pb is decreased with
increasing trophic level. This section incorporates recent literature on transfer of Pb
through marine food chains.
In a dietary study using environmentally realistic concentrations of Pb in prey through
four levels of a simplified marine food chain, biological responses including decreased
growth and survival and changes in behavior were observed at different trophic levels.
However, the concentration of Pb did not increase along the trophic gradient (Soto-
Jimenez et al.. 201 Ib; Soto-Jimenez et al., 201 la). The base of the simulated food chain
was the microalgae Tetraselmis suecica (phytoplankton) grown in 20 (ig Pb/L.
Pb-exposed cultures of T. suecica had significantly less cell divisions per day (growth),
biomass and total cell concentrations than control microalgae at 72 hours of exposure.
The microalgal cultures were then fed to Artemiafranciscana (crustacean, brine shrimp)
which were then fed to Litopenaeus vannamei (crustacean, whiteleg shrimp) and finally
to Haemulon scudderi (fish, grunt). Effects on behavior, growth and survival were
observed in shrimp and in grunt fish occupying the intermediate and top levels of the
simulated marine food chain. The authors speculate that the species used in the simulated
food chain were able to regulate and eliminate Pb (Soto-Jimenez et al.. 201 Ib).
Partial evidence for biomagnification was observed in a subtropical lagoon in Mexico
with increases of Pb concentration occurring in 14 of the 31 (45.2%) of trophic
interactions considered (Ruelas-Inzunza and Paez-Osuna. 2008). The highest rate of
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transference of Pb as measured in muscle tissue occurred between the prey species
whiteleg shrimp (Litopenaeus vannamei) and mullet (Mugil cephalus) to pelican
(Pelecanus occidentalis).
Other studies have traced Pb in marine food webs and have found no evidence of
biomagnification of Pb with increasing trophic level. In the southeastern Gulf of
California, Mexico, Pb was not positively transferred (biomagnification factor <1)
through primary producers (seston, detritus) and 14 consumer species in a lagoon food
web (Jara-Marini et al., 2009). In a planktonic food web in Bahia Blanca estuary,
Argentina, Pb levels in macrozooplankton and mesozooplankton exhibited temporal
fluctuations, however no biomagnification was observed between mesozooplankton and
macrozooplankton (Fernandez Severini et al., 2011). It is important to note, however, that
even in the absence of biomagnification, aquatic organisms can bioaccumulate relatively
large amounts of metals and become a significant source of dietary metal to their
predators (Fairbrother et al.. 2007: Reinfelder et al.. 1998V
6.4.15 Biological Effects of Pb in Saltwater Systems
This section focuses on the studies of biological effects of Pb on marine and estuarine
algae, plants, invertebrates, fish and mammals published since the 2006 Pb AQCD. Key
studies from the 1977 Pb AQCD, the 1986 Pb AQCD and the 2006 Pb AQCD on
biological effects of Pb are summarized where appropriate. Biological effects of Pb on
saltwater algae and plant species are considered below, followed by information on
effects on marine invertebrates and vertebrates. Alterations to reproduction, growth and
survival of saltwater organisms can lead to changes at the community and ecosystem
levels of biological organization such as decreased abundance, reduced taxa richness, and
shifts in species composition (Section 6.1). New evidence for Pb effects on reproduction,
growth and survival in saltwater plants, invertebrates and vertebrates is summarized in
Table 6-6. In general, Pb toxicity to saltwater organisms is less well characterized than
toxicity of Pb in freshwater ecosystems due to the fewer number of available studies on
marine species. Because this review is focused on effects of Pb, studies reviewed for this
section include only those for which Pb was the only, or primary, metal to which the
organism was exposed. All reported values are from exposures in which concentrations
of Pb were analytically verified unless nominal concentrations are stated.
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6.4.15.1 Saltwater Algae and Plants
New evidence on toxicity of Pb to marine algae indicates that species exhibit varying
sensitivities to Pb in saltwater. The lowest 72-hour EC50 for growth inhibition reported
for marine algae was 105 (ig Pb/L in Chaetoceros sp (Debelius et al.. 2009). The
microalgae T. suecica, had statistically significant decreased biomass yield, growth rate
and cell count following 72 hours nominal exposure to 20 (ig Pb/L (Soto-Jimenez et al..
201 Ib). Pb tested at nominal concentrations up to -2,000 (ig Pb/L over a 14-day period
did not affect photosynthetic activity in seven species of marine macroalgae
(Ascophyllum nodosum, Fucus vesiculosus, Ulva intestinalis, Cladophora rupestris,
Chondrus crispus, Palmaria palmate, Polysiphonia lanosd) as measured by pulse
amplitude modulation chlorophyll fluorescence yield although Pb was readily
accumulated by these species (Baumann et al.. 2009). In a recent review of the
production of phytochelatins and glutathione by marine phytoplankton in response to
metal stress, Kawakami et al. (2006) included several studies in which Pb exposure was
shown to induce glutathione and phytochelatin at high concentrations in a few species.
6.4.15.2 Saltwater Invertebrates
No studies with marine invertebrates were reviewed in the 1977 Pb AQCD or the 1986
Pb AQCD. Effects of Pb on marine invertebrates reported in the 2006 Pb AQCD included
impacts on embryo development in bivalves with an EC50 of 221 (ig Pb/L for
embryogenesis, gender differences in sensitivity to Pb in copepods and increasing
toxicity with decreasing salinity in mysids. Survival, growth, and reproduction are
affected by Pb in marine organisms. Pb has also been shown to affect stress responses,
antioxidant activity, and osmoregulation.
Recent literature strengthens the evidence indicating that Pb affects enzymes and
antioxidant activity in marine invertebrates. Most of these studies only report nominal
concentrations of Pb. Activity of enzymes associated with the immune defense system in
the mantle of pearl oyster were measured at 0, 24, 48 and 72 hour nominal exposure to
100 (ig Pb/L (Jing et al.. 2007). Activity of AcPase, a lysosomal marker enzyme, was
detected at 24 hours and subsequently decreased. Phenoloxidase activity was depressed
compared with controls and remained significantly lower than control after 72 hours of
exposure to Pb. Increased SOD activity was observed in the mantle but decreased with
time, although always remaining higher than in the control animals (Jing et al.. 2007).
Activity of Se-dependent glutathione peroxidase did not change with Pb exposure. SOD,
catalase, and glutathione peroxidase were significantly reduced at environmentally
relevant concentrations of Pb (2 (ig Pb/L as measured in Bohai Bay, China) in the
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digestive gland of the bivalve Chlamys farreri fZhang et al., 2010^). In contrast,
Einsporn et al. (2009) observed no change in catalase activity in the digestive gland and
gill of blue mussel M edulis following nominal exposure to 2,500 (ig Pb/L as Pb nitrate
for 10 days and again following a 10 day depuration period. However, in this same
species, glutathione-S-transferase activity was elevated in the gills after Pb exposure and
remained active during depuration while no changes to glutathione-S-transferase activity
were observed in the digestive gland. In black mussel (M galloprovincialis) exposed 10
days to sublethal nominal concentrations of Pb, fluctuations in SOD activity were
observed over the length of the exposure and MDA levels were increased in mantle and
gill (Vlahogianni and Valavanidis. 2007). Catalase activity was decreased in the mantle
of these mussels but fluctuated in their gills, as compared with the control group. In the
bivalve C. farreri exposed to Pb, there was induction of lipid peroxidation measured as
MDA of 24% and a 37% reduction in 7-ethoxyresorufm-o-deethylase (EROD) activity
when compared to controls (Zhang et al., 2010b). In red fingered marsh crab,
Parasesarma erythrodactyla, collected from sites along an estuarine lake in New South
Wales, Australia, elevated glutathione peroxidase activity was correlated with individuals
with higher metal body burdens (MacFarlane et al.. 2006).
ALAD is a recognized biomarker of exposure across a wide range of taxa including
bacteria (Korean et al., 2007). invertebrates and vertebrates. Since the 2006 Pb AQCD,
there are additional studies measuring changes in ALAD activity in field-collected
bivalves and crustaceans from saltwater habitats. In the bivalve Chamelea gallina
collected from the coast of Spain, ALAD inhibition was greater with higher
concentrations of Pb measured in whole tissue (Kalman et al., 2008). In another study
conducted in Spain, ALAD activity was negatively correlated with total Pb concentration
in seven marine bivalves (C. gallina, Mactra corallina, Donax trunculus, Cerastoderma
edule,M. galloprovincialis, Scrobicularia plana and Crassostrea angulata). However,
the authors of this study indicated the need to consider variability of responses between
species when using ALAD as a biomarker for Pb (Company et al.. 2011). Pb content
varied significantly among species and was related to habitat (sediment versus substrate)
and feeding behavior.
Behavioral responses of aquatic invertebrates to Pb reviewed in the 2006 Pb AQCD
included avoidance. A limited number of recent studies have considered additional
behavioral endpoints in marine organisms. Valve closing speed was used as a measure of
physiological alterations due to Pb exposure in the Catarina scallop (Sobrino-Figueroa
and Caceres-Martinez. 2009). The average valve closing time increased from under one
second in the control group to 3 to 12 seconds in juvenile scallops exposed to analytically
verified concentrations of Pb as Pb nitrate (40 (ig/L to 400 (ig/L) for 20 days. Damage to
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sensory cilia of the mantle was observed following microscopic examination of
Pb-exposed individuals.
Since the 2006 Pb AQCD, limited studies on marine invertebrates have indicated effects
of Pb on reproduction. In a long term (approximately 60 days) sediment
multigenerational bioassay with the estuarine-sediment dwelling amphipod Elasmopus
laevis, onset to reproduction was significantly delayed at 118 mg Pb/kg compared to
controls. In the higher concentrations, start of offspring production was delayed further;
4 days in 234 mg Pb/kg and 8 days in 424 mg Pb/kg (Ringenary et al., 2007). Fecundity
and time of first offspring production was also reduced with increasing Pb concentration
in sediment above 118 mg Pb/kg. The authors indicate that this concentration is below
the current marine sediment regulatory guideline for Pb (218 mg Pb/kg sediment)
(NOAA. 1999) and that reproductive endpoints are more sensitive than survival in this
species. Exposure of gametes to Pb prior to fertilization resulted in a decrease of the
fertilization rates of the marine polychaete Hydroides elegans (Gopalakrishnan et al..
2008). In sperm pretreated in 97 (ig Pb/L filtered seawater for 20 minutes, fertilization
rate decreased by approximately 70% compared to controls. In a separate experiment,
eggs were pretreated with Pb prior to addition of an untreated sperm suspension. The
fertilization rate of eggs pretreated in 48 (ig Pb/L filtered seawater decreased to 20% of
the control. In another test with H. elegans in which gametes were not pre-treated, but
instead added directly to varying concentrations of Pb for fertilization, there appears to be
a protective effect following fertilization due to the formation of the fertilization
membrane during the first cell division that may prevent Pb from entering the oocytes
(Gopalakrishnan et al., 2007).
As noted in the 2006 Pb AQCD and supported by recent studies, Pb exposure negatively
affects the growth of marine invertebrates. Wang et al., (2009d_) observed growth of
embryos of the Asian Clam (Meretrix meretrix) was significantly reduced by Pb with an
EC50 of 197 (ig/L. In juvenile Catarina scallop, Argopecten ventricosus, exposed to Pb
for 30 days, the EC50 for growth was 4,210 (ig Pb/L (Sobrino-Figueroa et al., 2007). Rate
of growth of the deposit feeding Capitella sp. polychaetes decreased significantly from
the controls in 3 and 6-day exposures, however, the observed changes did not exhibit a
clear dose response with increasing Pb concentration (Horng et al.. 2009).
Although Pb is known to cause mortality when invertebrates are exposed to sufficiently
high concentrations, some species may not exhibit significant mortality even at high
concentrations. In a 10-day Pb-spiked sediment exposure (1,000 mg Pb/kg and
15 (ig Pb/L dissolved Pb in pore water), 100% of individuals of the Australian estuarine
bivalve Tellina deltoidalis survived (King et al.. 2010). In the deposit feeding Capitella
sp., polychaetes, exposure to varying concentrations of Pb associated with spiked
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sediment up to 870 mg Pb/kg had no effect on survival (Horng et al., 2009). No
differences in adult survival were observed in 28 and 60 day sediment exposures to a
range of Pb concentrations from 58 mg Pb/kg to 424 mg Pb/kg in the amphipod E. laevis
(Ringenary et al. 2007). Other species are more sensitive to Pb and these responses are
reviewed in Section 6.4.16.
6.4.15.3 Saltwater Vertebrates
Saltwater Fish
There is a dearth of information in previous Pb AQCDs on Pb effects in saltwater fish.
Recent data available since the 2006 Pb AQCD include a study with a marine
elasmobranch. De Boeck et al. (2010) exposed the spotted dogfish (S. canicula) to
2,072 (ig Pb/L for one week and measured metallothionein induction in gill and liver
tissue, and the electrolytes Na, K, Ca2+ and Cl, in plasma. No effects were observed in
Pb-exposed fish for any of the physiological variables measured in this study, although
Pb was detected in all organs (De Boeck et al.. 2010).
Since the 2006 Pb AQCD, several studies integrating behavioral and physiological
measures of Pb toxicity have been conducted on marine fish. The ornate wrasse
(Thalassoma pavo) was exposed nominally to sublethal (400 (ig Pb/L) or a maximum
acceptable toxicant concentration (1,600 (ig Pb/L) dissolved in seawater for one week to
assess the effects of Pb on feeding and motor activities (Giusi et al.. 2008). In the
sublethal concentration group, hyperactivity was elevated 36% over controls. In the high
concentration, a 70% increase in hyperactivity was observed and hyperventilation
occurred in 56% of behavioral observations. Elevated expression of heat shock protein
70/90 orthologs was detected in the hypothalamus and mesencephalic areas of the brains
of Pb-treated fish and neuronal damage was observed in the posterior hypothalamic area
and optic tectum. No changes in feeding activity were noted between non-treated and
treated fish.
Additional behavioral studies in fish consider effects of dietary Pb. The grunt fish
H. scudderi, occupying the top level of a simulated marine food chain, exhibited lethargy
and decreased food intake during the last week of a 42-day feeding study (Soto-Jimenez
et al., 20 lib). The fish were fed white shrimp exposed to Pb via brine shrimp that were in
turn fed microalgae cultured at a nominal concentration of 20 (ig Pb/L. Pb was quantified
in shrimp and fish. The authors noted a few of the fish exposed to Pb via dietary transfer
through the food chain were observed surfacing and speculated that this behavior was air
breathing as a response to stress.
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Evidence for reproductive effects of Pb in saltwater fish is limited to a field study in
which decreased oocyte diameter and density in the toadfish (Tetractenos glaber) were
associated with elevated levels of Pb in the gonad offish collected from contaminated
estuaries in Sydney, Australia (Alquezar et al.. 2006). The authors state this is suggestive
of a reduction in egg size which ultimately may lead to a decline in female reproductive
output.
Mammals
Although Pb continues to be detected in tissues of marine mammals in U.S. coastal
waters (Bryan et al., 2007; Stavros et al., 2007; Kannan et al.. 2006) few studies exist that
consider biological effects associated with Pb exposure. Pb effects on immune variables,
including cell viability, apoptosis, lymphocyte proliferation, and phagocytosis were tested
in vitro on phagocytes and lymphocytes isolated from the peripheral blood of bottlenose
dolphin (Turslops truncates) (Camara Pellisso et al., 2008). No effects on viability of
immune cells, apoptosis, or phagocytosis were observed in 72-hour exposure to nominal
concentrations of 1,000, 10,000, 20,000 and 50,000 (ig Pb/L. Proliferative response of
bottlenose dolphin leukocytes was significantly reduced at 50,000 (ig Pb/L, albeit by only
10% in comparison to the control. This in vitro exposure with nominal concentrations of
Pb is likely not relevant for assessing effects of atmospherically-deposited Pb on marine
mammals, however, no additional studies were available for review on the effects on Pb
on these organisms.
6.4.16 Exposure and Response of Saltwater Species
Evidence regarding exposure-response relationships and potential thresholds for Pb
effects on saltwater populations can inform determination of standard levels that are
protective of marine ecosystems. The Annex of the 2006 Pb AQCD (U.S. EPA. 2006c)
summarized data on exposure-response functions for invertebrates (Table AX7-2.4.1)
(Table AX7-2.4.2). The recent exposure-response studies reviewed in this section expand
on earlier findings with information on microalgal and invertebrate species. Studies
specific to growth, reproduction and survival endpoints are summarized in Table 6-6. All
reported values are from exposures in which concentrations of Pb were analytically
verified unless nominal concentrations are stated.
A series of 72-hour Pb toxicity tests were conducted with five marine microalgae species
(T. chuii, R. salina, Chaetoceros sp., /. galbana and N. gaditand) to determine the relative
Pb sensitivities as measured by growth inhibition. The respective 72-hour EC50 values
derived were 2,640, 900, 105, 1,340, and 740 (ig Pb/L (Debelius et al.. 2009). The
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authors noted that species cellular size, sorption capacity, or taxonomy did not explain
differences in sensitivity to Pb, leaving the mechanism of response still open to question.
In the deposit feeding polychaete, Capitella sp. an LOAEL of 85 mg Pb/kg sediment was
established in 3 day and 6 day growth experiments (Horng et al., 2009). Other studies of
marine invertebrates published since the 2006 Pb AQCD (U.S. EPA. 2006c) have
indicated differences in sensitivity of different lifestages of aquatic organisms to Pb. In a
series of seawater and sediment exposures using adult and juvenile amphipods Melita
plumulosa, juveniles were more sensitive to Pb than adults (King et al., 2006). In the
seawater-only exposures, the 96-hour LC50 was 1,520 (ig Pb/L for juveniles and
3,000 (ig Pb/L for adults. In comparison,10 day juvenile sediment test results were LC50
1,980, NOEC 580 and LOEC 1,020 mg Pb/kg dry weight compared to the LC50, NOEC,
and LOEC value for the adults exposed in sediment (3,560 mg Pb/kg dry weight). A 24-
hour LC50 of 4,500 (ig Pb/L for adult black mussel (M galloprovincialis) suggests that,
in general, juvenile bivalves are more sensitive to Pb exposure than adults although this
value was based on nominal exposure data (Vlahogianni and Valavanidis. 2007).
Since the 2006 Pb AQCD, Pb toxicity to larval stages of marine species has been
assessed at sublethal and lethal concentrations. The effective concentrations at which Pb
resulted in 50% of abnormal embryogenesis of the Asian clam (M meretrix) was
297 (ig Pb/L. The 96-hour LC50 for larvae of the same species was 353 (ig Pb/L (Wang et
al., 2009d). In comparison, juvenile Catarina scallop (A. ventricosus) had a LC50 of
830 (ig Pb/L in a 96-hour exposure (Sobrino-Figueroa et al.. 2007). In the marine
polychaete H. elegans, EC50 values of gametes, embryos, larvae (blastulato trochophore
and larval settlement), and adults, exhibited dose-responses to Pb that reflected the
differential sensitivity of various lifestages of this organism (Gopalakrishnan et al..
2008). The EC50 values for sperm and egg toxicity were 380 and 690 (ig Pb/L
respectively. Larval settlement measured as the metal concentration causing 50%
reduction in attachment was most sensitive to Pb with an EC50 of 100 (ig Pb/L, while the
EC50 for abnormal development of embryos was 1,130 (ig Pb/L. The LC50 values for
adult worms in 24-hour and 96-hour tests were 25,017 and 946 (ig Pb/L, respectively.
Manzo et al. (2010) established a LOEC of 500 (ig Pb/L and a maximum effect at
3,000 (ig Pb/L in an embryotoxicity assay with sea urchin P. lividus exposed to nominal
concentrations of Pb. The EC50 for developmental defects in this species was
1,250 (ig Pb/L with a NOEL of 250 (ig Pb/L. In a study using nominal concentrations of
Pb, morphological deformities were observed in 50% of veliger larvae of blacklip
abalone (Haliotis rubra) at 4,100 (ig Pb/L following a 48-hour exposure, suggesting this
species is not as sensitive to Pb as other marine invertebrate larvae (Gorski and
Nugegoda. 2006).
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6.4.17 Community and Ecosystem Effects in Saltwater Systems
As discussed in the 1986 Pb AQCD and the 2006 Pb AQCD (U.S. EPA. 2006c).
exposure to Pb is likely to have impacts in aquatic environments via effects at several
levels of ecological organization (organisms, populations, communities, or ecosystems).
But fewer studies explicitly consider community and ecosystem-level effects in marine
and brackish waters than in freshwater. Reduced species abundance and biodiversity of
protozoan and meiofauna communities were observed in laboratory microcosm studies
with marine water and marine sediments reviewed in the 2006 Pb AQCD as summarized
in Table AX7-2.5.2 (U.S. EPA. 2006c). In a laboratory study with larval mummichogs
reviewed in the 2006 Pb AQCD, feeding and predator avoidance behaviors were altered
in this marine fish species following a 4-week exposure to Pb. Observations from field
studies reviewed in the 2006 Pb AQCD included findings of a negative correlation
between Pb and species richness and diversity indices of macroinvertebrates associated
with estuary sediments and changes in species distribution and abundance in fish,
crustaceans and macroinvertebrates correlated with Pb levels in marine sediments. The
2006 Pb AQCD concluded that, in general, information from controlled studies for single
pollutants was insufficient to permit evaluation of specific impacts on higher levels of
organization (beyond the organism). In studies from natural saltwater ecosystems, Pb
rarely occurs as a sole contaminant making its effects difficult to ascertain. New
information on effects of Pb at the population, community and ecosystem level in coastal
ecosystems is reviewed below.
The faunal composition of seagrass beds in a Spanish coastal saltwater lagoon was found
to be impacted by Pb in sediment, plants, and biofilm (Marin-Guirao et al., 2005).
Sediment Pb concentrations ranged from approximately 100 to 5,000 mg Pb/kg and
corresponding biofilm concentrations were 500 to 1,600 mg Pb/kg, with leaf
concentrations up to 300 mg Pb/kg. Although multiple community indices (abundance,
Shannon-Wiener diversity, Simpson dominance index) did not vary from site to site,
multivariate analysis and similarity analysis indicated significant differences in
macroinvertebrate communities between sites with different sediment, biofilm, and leaf
Pb concentrations. Differences were largely attributable to three amphipod species
(Microdeutopus sp., Siphonoecetes sabatieri, Gammarus sp.). This indicates that,
although seagrass abundance and biomass were unaffected by Pb exposure, organisms
inhabiting these plants still may be adversely impacted.
Caetano et al. (2007) investigated the mobility of Pb in salt marshes using total content
and stable isotope signature. They found that roots had similar isotopic signature to
sediments in vegetated zones indicating that Pb uptake by plants reflects the input in
sediments. At one site, there was a high anthropogenic Pb content while at the other
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natural mineralogical sources dominated. The roots ofS.fruticosa and S. maritima
significantly accumulated Pb, having maximum concentrations of 2,870 mg Pb/kg and
1,755 mg Pb/kg, respectively, indicating that below-ground biomass played an important
role in the biogeochemical cycling of Pb.
Exposure to three levels of sediment Pb contamination (322, 1,225, and 1,465 mg Pb/kg
dry weight) had variable effects on different species within a marine nematode
community (Mahmoudi et al.. 2007). Abundance, taxa richness, and species dominance
indices were altered at all Pb exposures when compared with unexposed communities.
Further, while the species Oncholaimellus mediterraneus dominated control communities
(14% of total abundance), communities exposed to low and medium Pb concentrations
were dominated by Oncholaimus campylocercoides (36%) andMarylynnia stekhoveni
(32%), and O. campylocercoides (42%) and Chromadorina metulata (14%), respectively.
Communities exposed to the highest Pb sediment concentrations were dominated by
Spirinia gerlachi (41%) and Hypodontolaimus colesi (29%). Given this, the authors
concluded that exposure to Pb significantly reduced nematode diversity and resulted in
profound restructuring of the community structure.
In another laboratory microcosm experiment with nematodes, nematode diversity and
community structure was altered with a mean number of 8 genera present in microcosms
contaminated with Pb compared to the control with 20 genera. The spiked sediments used
in the study were collected from the Swartkop River estuary, South Africa. Pb (3 to
6,710 mg Pb/kg sediment dry weight) was tested alone and in combination with Cu, Fe,
and Zn (Gyedu-Ababio and Baird. 2006). The synergistic effect of the four metals on
nematode community structure was greater than the individual metals and the effects of
Pb could not be distinguished from Cu, Fe and Zn.
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6.4.18 Characterization of Sensitivity and Vulnerability in Saltwater Species
Species differences in metabolism, sequestration, and elimination rates have been shown
to control relative sensitivity and vulnerability of exposed organisms and effects on
survival, reproduction, growth, metabolism, and development. Diet and lifestage at the
time of exposure also contribute significantly to the determination of sensitive and
vulnerable populations and communities. Further, environmental conditions in addition to
those discussed as affecting bioavailability may also alter Pb toxicity. The
2006 Pb AQCD (U.S. EPA. 2006c) reviewed the effects of genetics, age, and body size
on Pb toxicity. While genetics appears to be a significant determinant of Pb sensitivity,
effects of age and body size are complicated by environmental factors that alter metabolic
rates of saltwater organisms. A review of the more recent literature corroborated these
findings, and identified seasonal physiological changes and lifestage as other important
determinants of differential sensitivity to Pb.
6.4.18.1 Seasonally Affected Physiological Changes
Couture et al. (2010) investigated seasonal and decadal variations in Pb sources to
mussels (M. edulis) from the French Atlantic shoreline. Pb concentrations in the mussels
were 5-66 times higher than the natural background value for the north Atlantic. The
206Pb/207Pb signature indicated that the bioaccumulated Pb was anthropogenic in origin.
The signature was not, however, the same as that emitted in western Europe, as a result of
leaded gasoline combustion, although that was a major emission source to the atmosphere
during a large part of the study period (1985-2005). Instead, it was most similar to that of
Pb released into the environment from wastewater treatment plants, municipal waste
incinerators and industries such as metal refineries and smelters. Thus continental runoff
rather than atmospheric deposition was identified as the main source of Pb to the French
coastal area. The strong seasonal variations in 206Pb/208Pb were used to conclude that
resuspension of Pb triggered by high river runoff events was a key factor affecting
bioaccumulation of Pb inM edulis.
In another monitoring study, Pearce and Mann (2006) investigated variations in
concentrations of trace metals in the U.K. including Pb in the shells of pod razor shell
(Ensis siliqud). Pb concentration varied from 3.06-36.2 mg Pb/kg and showed a regional
relationship to known sources, e.g., former metal mining areas such as Cardigan Bay,
Anglesey, and industrial activity in Liverpool Bay. Seasonal variations were also found
for Pb in both Cardigan Bay and Liverpool Bay, relating to increased winter fluxes of Pb
(and other metals) into the marine environment. In contrast, levels of Pb and other metals
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were highest in summer and lowest in winter in oysters Crassostrea corteziensis collected
from Sonora, Mexico (Garcia-Rico et al.. 2010).
Carvalho et al. (2011) quantified 210Pb inM galloprovincialis sampled at coastal
locations in Portugal and noted that the apparent seasonal fluctuation in radionuclide
concentrations in mussel soft tissues was mostly attributable to changes in physiological
condition (i.e., fat content, gonadal development) and not to radionuclide body burden
fluctuation. The authors caution that since concentrations of contaminants are dependent
upon tissue composition, corrections for mussel physiological condition are need to
compare results from different seasons and different locations.
6.4.18.2 Lifestage
Lifestages of the marine polychaete H. elegans including embryogenesis, sexual
maturation, and offspring development were shown to be differentially affected by Pb
exposure. Pb water concentrations of 91 (ig Pb/L and greater significantly affected
fertilization and embryonic development, but the greatest effects were exhibited by 24-
hour-old larvae (Gopalakrishnan et al.. 2007). The authors suggested that timing of Pb
exposure may have different impacts on marine polychaete populations, if life cycles are
offset (Gopalakrishnan et al., 2007). Further, given that the adult lifestage is sedentary,
reduction of the mobile early lifestage as a result of Pb exposures may disproportionally
affect sessile polychaetes. For instance, larval settlement was significantly reduced at Pb
exposures of 48 (ig Pb/L and greater (Gopalakrishnan et al.. 2008).
6.4.18.3 Species Sensitivity
Both inter- and intra-specific differences in Pb uptake and bioaccumulation may occur in
macroinvertebrates of the same functional feeding group. Data from 20 years of
monitoring of contaminant levels in filter-feeding mussels of the Mytilus species and
Crassostrea virginica oysters in coastal areas of the U.S. through the National Oceanic
and Atmospheric Administration (NOAA) Mussel Watch program indicate that Pb is on
average three times higher in mussels than in oysters (Kimbrough et al., 2008). Limpet
(Patella sp.) from the Lebanese Coast had Pb BAF values ranging from 2,500 to 6,000
and in the same field study Pb BAF values for a mussel (Brachidontes variabilis) ranged
from 7,500-8,000 (Nakhle et al.. 2006).
There is some indication that molting may comprise an additional sequestration and
excretion pathway for aquatic animals exposed to Pb (Soto-Jimenez et al.. 201 la;
Mohapatra et al.. 2009: Tollett et al.. 2009: Bergevand Weis. 2007). Crab species
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U. pugnax (Bergey and Weis. 2007) and Scylla serrata fMohapatra et al., 2009A and
white shrimp L. vannamei (Soto-Jimenez et al.. 201 la) have been shown to sequester Pb
preferentially in exoskeleton tissue, where it is later shed along with other tissue.
Consequently, aquatic arthropod species and those species that shed their exoskeleton
more frequently may be able to tolerate higher environmental Pb concentrations than
non-arthropods or slow-growing molting species, as this pathway allows them to
effectively lower Pb body burdens.
Some tolerant species offish (e.g., mummichog) have the ability to sequester
accumulated Pb in metal-rich granules or heat-stable proteins (Goto and Wallace. 2010).
Fish with such abilities are more likely to thrive in Pb-contaminated environments than
other species.
6.4.19 Ecosystem Services Associated with Saltwater Systems
Pb deposited on the surface of (or taken up by) organisms has the potential to alter the
services provided by saltwater biota to humans although the directionality of impacts is
not always clear. For example, oysters and mussels provide a service by sequestering Pb.
At the same time, the uptake of Pb by these bivalves may result in toxicological effects
associated with Pb exposure and decreased value of shellfish as a commodity. At this
time, a few publications address Pb impacts on ecosystem services associated with
saltwater ecosystems. Pb can affect the ecological effects in each of the four main
categories of ecosystem services (Section 6.1.2) as defined by Hassan et al. (2005). These
effects are sorted into ecosystem services categories and summarized here:
• Supporting: food for higher trophic levels, biodiversity
• Provisioning: contamination of food by heavy metals, decline in health offish
and other aquatic species
• Regulating: water quality
• Cultural: ecosystem and cultural heritage values related to ecosystem integrity
and biodiversity, wildlife and bird watching, fishing
A few recent studies explicitly consider the impact of Pb and other heavy metals on
ecosystem services provided by salt marsh (Gedan et al.. 2009) and estuaries (Smith et
al.. 2009b). These systems are natural sinks for metals and other contaminants. Pb can be
toxic to salt marsh plant species and decaying plant detritus may result in resuspension of
Pb into the aquatic food chain (Gedan et al.. 2009). Salt marsh and estuaries provide
habitat and breeding areas for both terrestrial and marine wildlife and are locations for
bird watching. Using a modeling approach designed to assess the degree of risk of Pb and
Hg to wading birds in estuarine habitats in the U.K., the authors found a high probability
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that Pb poses an ecologically relevant risk to dunlin, Calidris alpina fSmith et al.. 20091x).
However, the authors noted that a major source of uncertainty in this study was the
NOAEL values for Pb.
The impact of Pb on ecological services provided by specific components of aquatic
systems has been considered in a limited number of studies. Recent research has
suggested that dietary Pb (i.e., Pb adsorbed to sediment, particulate matter, and food) may
contribute to exposure and toxicity in primary and secondary order consumers (including
humans). Aquatic fauna can take up and bioaccumulate metals. If the bioaccumulating
species is a food source, the uptake of metals may make it toxic or more dangerous for
people or other wildlife to consume. For example, oysters and mussels bioaccumulate Pb
from anthropogenic sources, including atmospheric deposition, and are a food source that
is widely consumed by humans and wildlife (Couture et al.. 2010). Their capacity to
bioaccumulate Pb makes them good bioindicators of environmental contamination and
they have been used as monitors of coastal pollutants by the NOAA Mussel Watch
program since 1986. Although bioaccumulation may render aquatic fauna toxic to
consumers, bioaccumulation is a way to sequester the metals and remove them from
waters and soils. Sequestration for this purpose is itself an ecosystem service and has
been quantified. For example, the total ecological services value of a constructed
intertidal oyster (Crassostrea sp.) reef in improving water quality and sequestering metals
including Pb was calculated in the Yangtze River estuary to be about $500,000 per year
(Quan et al.. 2009).
6.4.20 Synthesis of New Evidence for Pb Effects in Saltwater Systems
This synthesis of the effects of Pb on saltwater ecosystems covers information from the
publication of the 2006 Pb AQCD (U.S. EPA. 2006c) to present. It is followed in
Section 6.4.21 by determinations of causality that take into account evidence dating back
to the 1977 Pb AQCD. In general, evidence for toxicity to saltwater organisms is less
well characterized than toxicity of Pb in freshwater ecosystems due to the fewer number
of available studies on marine species. The studies that are available for marine plants,
invertebrates and vertebrates include studies where Pb concentration was analytically
verified and those that reported nominal concentrations (Table 6-6). Many of the studies
that report nominal concentrations in media are uptake studies that subsequently quantify
Pb in tissues; however, measurement of Pb in water or sediment at the beginning of an
exposure is desirable when comparing laboratory studies to concentrations of Pb in
marine systems. In Section 6.2.3 and Table 6-2. a range of 0.01 to 27 (ig Pb/L was
reported for saltwater, including estuaries and open ocean, with the higher values
associated with sites involving human activity (Sadiq. 1992).
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Most studies on marine organisms reviewed in the present document included
concentrations that were higher than Pb encountered in seawater. However, when
multiple concentrations were used, effects generally increased with increasing Pb
exposure. Effects at lower concentrations can be implied from many reported studies
since an exposure response relationship to Pb was observed. In marine and estuarine
systems, exposure to Pb from air is most likely characterized as a chronic low dose
exposure, however, most studies only report an acute LC50 value when an LCi0 would be
a more appropriate measurement for consideration of effects on organisms since an effect
occurring at the LC50 value would most likely not maintain a stable population.
Plants
Only a few studies were available since the 2006 Pb AQCD that consider effects of Pb on
marine algae (Section 6.4.15.1). A 72-hour EC50 for growth inhibition was reported in the
marine algae Chaetoceros sp. at 105 (ig Pb/L (Debelius et al.. 2009). A study with the
green alga T. suecica reports a statistically significant decease in growth rate, total dry
biomass and final cell concentration between control cultures and algae cultured in
20 (ig Pb/L (Soto-Jimenez et al.. 201 Ib). Both of these studies suggest growth effects at
or near the highest recorded values of Pb in seawater (27 (ig Pb/L), however, effects are
likely to occur at lower concentrations since only EC50 values are reported.
Invertebrates
In saltwater invertebrates (Section 6.4.15.2 and 6.4.16) there are studies that consider
Pb-effects on supporting endpoints (stress responses, hematological effects and
neurobehavior) as well as studies that assess Pb impacts to reproduction, growth, and
survival; endpoints that have the potential to alter population, community and
ecosystem—levels of biological organization. Many studies, especially those that
consider enzymatic responses to Pb exposure, were conducted with nominal Pb
concentrations. Two of these studies; Jing et al. (2007) and Zhang et al. (201 Ob).
considered Pb nominal exposures at 100 (ig Pb/L or lower and reported significant
decreases in antioxidant enzyme activity. The Zhang et al. (201 Ob) study observed effects
on enzymatic activity at a nominal exposure of 2 (ig Pb/L. Although these effects are near
reported Pb concentrations in seawater they were not analytically verified.
Other studies that report sub-organismal responses in saltwater organisms have quantified
Pb exposure. Field studies with bivalves collected off the coast of Spain correlated
ALAD activity with measured levels of Pb in tissue (Company et al.. 2011; Kalman et al..
2008). An increase in valve closing time with increasing Pb exposure in the range of 40
to 400 (ig Pb/L was observed in the scallop, A.ventricosus (Sobrino-Figueroa and
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Caceres-Martinez. 2009). Although the concentrations in this study exceed reported
levels of Pb in seawater, the lower range is near 27 (ig Pb/L reported by Sadiq (1992).
Evidence for effects on reproduction, growth, and survival in marine invertebrates (Table
6-6) are primarily from studies in which Pb in the exposure media was quantified. In the
amphipod, E. laevis, onset to reproduction was significantly delayed at 118 mg/Pb kg
sediment; a concentration that the authors indicate is below the current marine sediment
regulatory guideline for Pb (218 mg Pb/kg sediment) (Ringenary et al.. 2007; NOAA.
1999). In the same study, no effects of Pb on adult survival in 28 and 60 day sediment
exposures were observed. In another study with amphipods, juvenile M. plumosa were
more sensitive than adults in 10-day sediment exposures with an NOEC of 580 mg Pb/kg
dry weight compared to an NOEC of 3,560 mg Pb/kg dry weight for adults (King et al..
2006). Effects of Pb on gametes of the marine polycheate H. elegans were observed at
48 (ig Pb/L (Gopalakrishnan et al.. 2008). a concentration near the upper range of Pb in
seawater reported by Sadiq (1992). Specifically, fertilization rate of eggs pretreated with
48 (ig Pb/L decreased to 20% of control. Life stages ofH. elegans varied in their
sensitivity to Pb with the most sensitive period being larval settlement with an EC50 of
100 jig Pb/L.
There are only a few recent studies that considered effects of Pb on growth of marine
invertebrates (Sections 6.4.15.2 and 6.4.16). In the polychaete Capitella sp. growth was
decreased significantly from controls, however, there was not a clear dose-response
relationship between increasing Pb concentrations and observed effects (Horng et al..
2009). The authors reported a LOAEL of 85 mg Pb/kg in the sediment exposure. In the
Asian ClamM meretrix, an EC50 of 197 (ig Pb/L was reported for growth (Wang et al..
2009d). Other marine invertebrate growth effects were observed at much higher Pb
concentrations (Table 6-6).
Survival was a less sensitive endpoint in marine invertebrates than reproduction or
growth with no effects reported at concentrations typically observed in seawater (Table
6-6). In the amphipod M plumulosa an NOEC of 400 (ig Pb/L for juveniles and an
NOEC of 850 (ig Pb/L was reported for adults in 96-hour seawater only exposures (King
et al.. 2006). In 10 day sediment tests with the same species, juveniles were also more
sensitive than adults. Other concentrations at which survival effects were reported in
marine invertebrates also greatly exceeded concentrations of Pb typically found in
seawater.
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Vertebrates
There is not sufficient new evidence for saltwater vertebrates especially for reproductive,
growth and survival endpoints that may have relevance to the population-level of
biological organization and higher.
Food Web
Some studies published since the 2006 Pb AQCD (see Section 6.4.14.4) support the
potential for Pb to be transferred in saltwater food webs, while other studies have found
no evidence for biomagnification.
Ecosystem Level Effects
Evidence for effects at higher levels of biological organization in saltwater habitats is
primarily supported by observations in a small number of microcosm and field studies
where shifts in community structure are the most commonly observed effects of Pb
(Section 6.4.17). Effects on reproduction, growth or survival (summarized in Table 6-6)
may lead to effects at the population-level of biological organization and higher.
6.4.21 Causal Determinations for Pb in Saltwater Systems
In the following sections, organism-level effects on reproduction and development,
growth and survival are considered first since these endpoints can lead to effects at the
population level or above and are important in ecological risk assessment.
Neurobehavioral effects are considered next followed by sub-organismal responses
(hematological effects, physiological stress) for which Pb has been shown to have an
impact in multiple species and across taxa, including humans. Causal determinations for
terrestrial, freshwater and saltwater ecological effects are summarized in Table 6-3.
6.4.21.1 Reproductive and Developmental Effects-Saltwater Biota
Reproductive effects of Pb have been reported in a few marine organisms and the
majority of the available studies are with invertebrate species. In a study reviewed in the
2006 Pb AQCD (U.S. EPA. 2006c). embryo development in two commercial bivalves
Ruditapes decussatus and M. galloprovincialis was inhibited by Pb (Beiras and
Albentosa. 2003). In R. decussatus an EC50 range of 156 to 312 (ig Pb/L and LOEC of
156 (ig Pb/L were observed for inhibition of embryonic development while in
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M. galloprovincialis the EC50 was 221 (ig Pb/L and the LOEC was 50 (ig Pb/L. Larvae of
the mussel M edulis were sensitive to Pb exposure with an EC50 of 476 (ig Pb/L for
abnormal development of embryos following 48-hour exposure to Pb during
embryogenesis (Martin et al.. 1981). The LOEC for embryogenesis in the marine bivalve
M. galloprovincialis was 50 (ig Pb/L with an EC50 for embryogenesis of 221 (ig Pb/L
(Beiras and Albentosa. 2003).
Recent evidence for reproductive effects of Pb on marine invertebrates is summarized in
Table 6-6. In the marine polychaete H. elegans an EC50 of 261 (ig Pb/L was observed for
unhatched or abnormal larvae following 20 hour incubation with Pb (Gopalakrishnan et
al.. 2008). The EC50 for the metal concentration causing 5% reduction in larval
attachment was 100 (ig Pb/L. The EC50 values for sperm and egg toxicity were 380 and
692 (ig Pb/L, respectively. The EC50 for embryogenesis in the clamM meretrix was
297 (ig Pb/L (Wang et al., 2009d). In a multigenerational bioassay with the marine
amphipod E. laevis, statistically significant delays in onset of reproduction (4 to 8 days),
sexual maturation and first offspring were observed at concentrations of 188 mg Pb/kg
sediment and higher (Ringenary et al.. 2007). The authors indicate that this concentration
is below the current sediment regulatory guideline for Pb (218 mg Pb/kg sediment)
(TSfOAA. 1999) and that reproductive effects are a more sensitive endpoint than lethality.
Although LC50 values are typically reported for Pb effects on reproductive endpoints in
saltwater invertebrates, a concentration dependent relationship between reproductive
impairment and increasing concentration of Pb is reported in most studies. This exposure-
response relationship implies that effects on reproduction are occurring at concentrations
lower than the LC50 value.
Reproductive effects are only characterized in a few species and endpoints for marine
systems. The weight of the current evidence for reproductive effects is limited to
laboratory-based studies with saltwater invertebrates in which observed effects occur at
Pb concentrations that are higher than Pb concentrations encountered in the marine
environment. Evidence for reproductive effects of Pb on marine plant species is limited to
one study on the red alga (Champia parvuld) reviewed in the draft Ambient Aquatic Life
Water Quality Criteria for Lead report (U.S. EPA. 2008b). In one study from a saltwater
fish, field-collected smooth toadfish (T. glaber) from metal contaminated estuaries in
Sydney, Australia had elevated Pb levels in gonad and decreased oocyte diameter and
density. Evidence is, therefore, inadequate to conclude that there is a causal relationship
for reproductive effects in saltwater plants, and vertebrates. The available studies on
marine invertebrates are suggestive that there is a causal relationship between Pb
exposure and reproductive effects.
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6.4.21.2 Growth Effects-Saltwater Biota
There are few studies that measure growth effects of Pb on marine organisms; available
information is limited to marine flora and invertebrates. Growth studies in saltwater plant
species are summarized in Table 4 and Table 6 of the draft Ambient Aquatic Life Water
Quality Criteria for Lead report (U.S. EPA. 2008b) and Table 6-6 of the present
document.
Diatoms are among the most sensitive algae; however, growth effects are typically
observed at concentrations of Pb higher than the range of values available from saltwater
locations [0.01 to 27 (ig Pb/L, (Sadiq. 1992)1. In studies available since the draft Ambient
Aquatic Life Water Quality Criteria for Lead report, the lowest 72-hour EC50 for growth
inhibition reported in marine diatoms was 105 (ig Pb/L in Chaetoceros sp (Debelius et al..
2009) and the growth of the green alga T. suecica exposed nominally to 20 (ig Pb/L was
40% lower than control cultures (Soto-Jimenez et al.. 201 Ib). The microalgae was the
base of a simulated marine food chain including primary, secondary and tertiary level
consumers and effects on survival were observed at the higher trophic levels that
originated from Pb exposure via consumption of the primary producer. The majority of
growth effects reported in saltwater algae exceed concentrations of Pb in seawater by
several orders of magnitude. Effects of Pb on growth in two species of brown algae,
Fucus vesiculosus and Fucus serratus are summarized in Table 6 of the draft Ambient
Aquatic Life Water Quality Criteria for Lead report (U.S. EPA. 2008b). Concentrations
where growth impairment was observed in these species greatly exceed available values
for Pb measured in seawater.
In saltwater invertebrates, evidence for growth effects is limited to a few species at
concentrations that exceed Pb concentrations reported in seawater. Growth inhibition in
the bivalve Macoma balthica (EC 50 =45 3.4 (ig Pb/L) is reported in Table 6 of the draft
Ambient Aquatic Life Water Quality Criteria for Lead report (U.S. EPA. 2008b). Recent
studies include Wang et al., (2009d_) in which observed growth of embryos of the Asian
Clam (M meretrix) was significantly reduced by Pb with an EC50 of 197 (ig Pb/L. In
juvenile Catarina scallop, A. ventricosus, exposed to Pb for 30 days, the EC50 for growth
was 4,210 (ig Pb/L (Sobrino-Figueroa et al.. 2007). Rate of growth of the deposit feeding
polychaete Capitella sp. exposed to Pb-spiked sediments from polluted estuaries
decreased significantly from the control; however, changes were inconsistent with
increasing concentration of Pb (Horng et al.. 2009). Evidence is therefore inadequate to
conclude that there is a causal relationship between Pb exposure and growth effects in
saltwater plants, invertebrates and vertebrates.
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6.4.21.3 Survival-Saltwater Biota
There are no studies reported in the previous Pb AQCDs or the current ISA for aquatic
plants that indicate phytotoxicity at or near current concentrations of Pb in saltwater [0.01
to 27 ng Pb/L, (Sadiq. 1992)1.
Mortality data for saltwater invertebrate species are summarized in the draft Ambient
Aquatic Life Water Quality Criteria for Lead report (U.S. EPA. 2008b) and reported
LC50 values greatly exceed Pb concentrations encountered in seawater. Recent studies
available since the 2006 Pb AQCD, and the draft Ambient Aquatic Life Water Quality
Criteria for Lead document, that report mortality data are summarized in Table 6-6. In
general, marine fauna are less sensitive to this metal than freshwater fauna and the
highest toxicity is observed in juveniles. A 144-hour LC50 of 680 (ig Pb/L was reported
for juvenile scallop A. ventricosus (Sobrino-Figueroa et al.. 2007) and a 96-hour LC50 of
353 (ig Pb/L for embryos of the clamM meretrix (Wang et al.. 2009d). In the amphipod
M. plumulosa, juveniles were more sensitive to Pb than adults in 96 hour seawater-only
exposures and 10 day sediment exposures (King et al.. 2006). The 96-hour LC50 was
1,520 ng Pb/L and the NOEC was 400 (ig Pb/L for juveniles in comparison to adults (96-
hour LC50 =3,000 (ig Pb/L; NOEC= 1,680 ng Pb/L). In the 10-day sediment exposures,
the NOEC for juveniles was 580 mg Pb/kg dry weight compared to an adult NOEC of
3,560 mg Pb/kg dry weight. In 10-day exposures to Pb nitrate spiked sediment, all
individuals of the bivalve T. deltoidalis survived at 1,000 mg/Pb kg with 15 (ig Pb/L
dissolved in pore water (King etal.. 2010). No effects on survival were observed in either
the amphipod E. laevis exposed 60 days to Pb-spiked sediment up to 424 mg Pb/kg
(Ringenary et al., 2007), or in the polychaete Capitella sp. exposed to sediment for 3 or 6
days up to 871 mg Pb/kg (Horng et al.. 2009).
Effects of Pb on survival have been demonstrated though a simulated marine food chain
in which the primary producer, the microalgae T. suecica, was exposed nominally to
20 (ig Pb/L and subsequently fed to brine shrimp A. franciscana, (mean Pb content 12 to
15 mg Pb/kg) which were consumed by white-leg shrimp L. vannamei, itself consumed
by grunt fish H. scudderi representing the top of the marine food chain (Soto-Jimenez et
al.. 20 lib). Survival of brine shrimp was 25 to 35% lower than the control and both
white shrimp and grunt fish had significantly higher mortalities than controls.
Data on Pb toxicity to eight species of marine fishes are summarized in Table 1 of the
draft Ambient Aquatic Life Water Quality Criteria for Lead report (U.S. EPA. 2008b). All
of the LC50 values for these fish (range 1,500 to 315,000 \ig Pb/L) greatly exceed
concentrations of Pb reported in seawater. Additionally, in the 2006 Pb AQCD (U.S.
EPA. 2006c) the acute toxicity of Pb to plaice (Pleuronectes platessa) was reported to
range from 50 (ig Pb/L to 300,000 (ig Pb/L depending on the form of Pb (Eisler. 2000).
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The existing evidence on toxicity of Pb to marine vertebrates is limited to laboratory-
based studies conducted under different salinities and exposure conditions. Considerable
uncertainties exist in applying laboratory observations to actual conditions in the field
where other modulating factors can affect Pb bioavailability and toxicity.
Although evidence exists for increased mortality of marine fish at very high
concentrations of Pb, the focus of the causal determinations are on studies where effects
were observed within one to two orders of magnitude of Pb measured in the environment
(Table II of the Preamble). Evidence is therefore inadequate to conclude that there is a
causal relationship between Pb and survival in saltwater plants, invertebrates, and
vertebrates.
6.4.21.4 Neurobehavioral Effects-Saltwater Biota
In marine organisms evidence for neurobehavioral effects of Pb is limited to a few studies
on bivalves and fish. In a study reviewed in the 2006 Pb AQCD (U.S. EPA. 2006c). prey
capture rate and predator avoidance was affected in mummichogs starting at 300 (ig Pb/L
Weis and Weis. 1998). Recent studies support previous findings of decreased ability to
escape predation associated with Pb exposure. In juvenile Catarina scallops exposed to
Pb (40 (ig/L to 400 (ig/L) for 20 days, the average valve closing time increased from
under one second in the control group to 3 to 12 seconds in juvenile scallops A decrease
in valve closing speed in these bivalves may impact escape swimming behaviors
important for predator avoidance (Sobrino-Figueroa and Caceres-Martinez. 2009).
Behavioral effects in grunt fish H. scudderi, occupying the top level of a simulated
marine food chain included lethargy and decreased food intake in a 42-day feeding study
(Soto-Jimenez et al.. 201 Ib). These fish were fed white shrimp exposed to Pb via brine
shrimp that were initially fed microalgae cultured at a nominal concentration of
20 (ig Pb/L. In the same study, surfacing, reduction of motility, and erratic swimming
were observed in the white shrimp after 30 days of exposure to Pb via diet. The ornate
wrasse, T. pavo, was exposed nominally to sublethal (400 (ig Pb/L) or a maximum
acceptable toxicant concentration (1,600 (ig Pb/L) dissolved in seawater for one week to
assess the effects of Pb on feeding and motor activities (Giusi et al., 2008). In the
sublethal concentration group, hyperactivity was elevated 36% over controls. In the high
concentration, a 70% increase in hyperactivity was observed and hyperventilation
occurred in 56% of behavioral observations, however, no changes in feeding activity
were noted between non-treated and treated fish.
Most of the evidence for neurobehavioral changes in marine organisms is observed with
concentrations of Pb that exceed the range of Pb values available for saltwater of 0.01 to
6-259
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27 (ig Pb/L (Sadiq. 1992)1. with the exception of the food chain study discussed above in
which behavioral effects were observed in shrimp and their fish predators following
ingestion of microalgae cultured in nominal concentration of 20 (ig Pb/L and then
quantified in prey (Soto-Jimenez et al.. 201 Ib). Marine species are typically
underrepresented in toxicity testing of behavioral endpoints with metals. There are
considerable uncertainties in applying observations from laboratory-based studies to field
scenarios including the role of environmental factors such as salinity and DOM on Pb
bioavailability. Evidence is inadequate to conclude that there is a causal relationship
between Pb exposures and neurobehavioral endpoints in saltwater invertebrates and
vertebrates.
6.4.21.5 Hematological Effects-Saltwater Biota
Evidence for hematological effects of Pb on saltwater organisms is limited primarily to
field monitoring studies on bivalves. Several recent field studies using a multi-biomarker
approach to study the sources and impacts of Pb in marine environments have measured
ALAD activity in bivalve species and found positive correlations between increased
tissue Pb levels and ALAD inhibition (e.g., Company et al. (2011). Kalman et al (2008)).
Generally, these studies have noted that Pb content varies significantly among species
and is related to habitat and feeding behavior. There is precedent, especially in Europe,
for the inclusion of ALAD as a biomarker of exposure to Pb in marine invertebrates. The
mechanism of ALAD inhibition in response to Pb exposure is likely mediated through a
common pathway in both marine and freshwater invertebrates (Section 6.4.12.5) as well
as in terrestrial species (Section 6.3.12.5) and humans (Section 4.7). Evidence is
therefore, suggestive of a causal relationship between Pb exposure and hematological
effects in saltwater invertebrates. Evidence is inadequate to conclude that there is a causal
relationship between hematological effects and saltwater vertebrates.
6.4.21.6 Physiological Stress-Saltwater Biota
Most studies on physiological stress responses in marine invertebrates are laboratory-
based exposures where effects are observed at Pb concentrations that exceed those known
to occur in seawater [0.01 to 27 (ig Pb/L (Sadiq. 1992). Table 6-2]. However, some
recent evidence for invertebrate antioxidant response in bivalves and crustaceans
indicates effects may occur at Pb concentrations that are detected in the marine
environment. For example, SOD, catalase, and glutathione peroxidase activities were
significantly reduced in the digestive gland of the marine bivalve C. farreri at 2 (ig Pb/L
(as measured in Bohai Bay, China) (Zhang et al.. 2010b). In red fingered marsh crabs,
6-260
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P. erythrodactyla collected from an estuarine lake in Australia, elevated glutathione
peroxidase activity was correlated with individuals with higher metal body burdens
(MacFarlane et al.. 2006).
Additional evidence from environmental monitoring studies that compared biomarker
responses between reference and contaminated sites indicated a correlation between the
amount of Pb with changes in antioxidant enzyme activity [e.g., (Serafim etal. 2011;
Cravo et al.. 2009)1. Marine bivalves are the organisms typically sampled for these
biomonitoring studies since both metals and enzymatic activities can be readily measured
in these invertebrates. Although these studies show clear evidence of alterations in
antioxidant stress markers in response to marine pollution, these effects cannot be
attributed solely to Pb in the environment due to the presence of other metals and
contaminants. Evidence for stress responses in marine organisms is typically limited to
invertebrates, however, elevated expression of heat shock protein orthologs were reported
for the first time in the hypothalamic and mesencephalic brain regions of Pb-treated fish
(Giusi et al.. 2008).
Evidence for physiological stress responses in saltwater invertebrates are supported by
evidence in freshwater species (Section 6.4.12.6) and terrestrial species (Section 6.3.12.6)
as well as in humans and experimental animal studies of oxidative stress following
impairment of normal metal ion functions (Section 4.2.4). Stress responses may increase
susceptibility to other stressors and reduce individual fitness. Evidence is suggestive of a
causal relationship between Pb exposures and physiological stress in saltwater
invertebrates. The evidence is inadequate to conclude that there is a causal relationship
between Pb exposure and physiological stress in saltwater plants and vertebrates.
6.4.21.7 Community and Ecosystem Level Effects-Saltwater Biota
No studies on community and ecosystem level effects of Pb in marine systems were
reviewed in the 1977 Pb AQCD (U.S. EPA. 1977). or the 1986 Pb AQCD (U.S. EPA.
1986a). Observations from field studies reviewed in the 2006 Pb AQCD (U.S. EPA.
2006c) included findings of a negative correlation between Pb and species richness and
diversity indices of macroinvertebrates associated with estuary sediments (summarized in
Table AX7-2.5.2 of the 2006 Pb AQCD). Additional findings in marine environments
included changes in species distribution and abundance in fish, crustaceans and
macroinvertebrates correlated with Pb levels in marine sediments.
New evidence for community and ecosystem level effects of Pb in saltwater ecosystems
includes laboratory microcosm studies as well as observations from field-collected
sediments, biofilm and plants in which changes in community structure were observed. In
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a recent study, significant differences in macroinvertebrate communities associated with
seagrass beds were reported between sites with different sediment, biofilm, and leaf Pb
concentrations (Marin-Guirao et al.. 2005). Sediment Pb concentrations ranged from
approximately 100 to 5,000 mg Pb/kg and corresponding biofilm concentrations were
500 to 1,600 mg Pb/kg, with leaf concentrations up to 300 mg Pb/kg. In a laboratory
microcosm experiment conducted with estuarine sediments from South Africa, total
meiofauna density decreased (range 3 to 5 taxa) after 32 days in Pb-treated (1,886 to
6,710 (ig/Pb g sediment dry weight) sediments compared to 9 taxa in the control (3 (ig/Pb
g sediment dry weight) (Gyedu-Ababio and Baird. 2006). In a microcosm experiment,
exposure to three levels of sediment Pb contamination (322, 1,225, and 1,465 mg Pb/kg
dry weight) significantly reduced marine nematode diversity and resulted in profound
restructuring of the community structure (Mahmoudi et al.. 2007).
There is not sufficient information at this time to characterize and to quantify
relationships between ambient concentrations of Pb and response in saltwater
communities and ecosystems. Fewer studies are available for saltwater organisms when
compared to freshwater systems. There are likely differences in uptake and
bioaccumulation in marine species due to physiological characteristics for adaptation in
salt water. Additional uncertainties in evaluating the effects of Pb in marine environments
include the presence of multiple stressors, inherent natural variability, and differences in
Pb bioavailability across saltwater ecosystems. Evidence is inadequate to establish if
there is a causal relationship between Pb exposures and the alteration of species richness,
species composition and biodiversity in saltwater ecosystems.
6.5 Causal Determinations for Ecological Effects of Pb
This section summarizes the key conclusions regarding causality for welfare effects of
Pb. Causal determinations for reproductive, growth, survival, neurobehavioral,
hematological and physiological stress endpoints are presented separately for terrestrial,
freshwater, and saltwater organisms (Sections 6.3.12. 6.4.12. and 6.4.21). In Section
1.7.3. causal determinations for the same endpoints are further integrated across
terrestrial, freshwater, and saltwater taxa. Evidence considered in establishing causality
was drawn from findings presented in the 1977 (U.S. EPA. 1977). 1986 (U.S. EPA.
1986a) and 2006 Pb AQCDs (U.S. EPA. 2006c). integrated with an exhaustive review of
more recent evidence. The causal statements for terrestrial, freshwater and saltwater
effects are divided into two categories: (1) endpoints that are commonly used in
ecological risk assessment (reproduction, growth and survival) because they clearly can
lead to population-level (e.g., abundance, production, extirpation), community-level (taxa
richness, relative abundance) and ecosystem-level effects (Ankley etal.. 2010; Suter et
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al.. 2005). and (2) organism and sub-organism responses such as neurobehavioral effects,
hematological effects and physiological stress. There are many different effects at the
molecular and cellular levels, and chronic toxicity of Pb in ecosystems is thus likely
attained through multiple modes of action. Furthermore, the effects of Pb on ecosystems
necessarily begin with some initial effects at the molecular level of specific organisms
within the ecosystem (U.S. EPA. 1986b).
Experimental settings for studies used in making causal determinations for the ecological
effects of Pb include controlled exposures in the laboratory, microcosm experiments and
field observations. Controlled exposure studies in laboratory or small-to medium-scale
field settings provide the most direct evidence for causality, but their scope of inference
may be limited. In contrast, microcosms and field studies where exposure is not
controlled include potentially confounding factors (e.g., other metals) or factors known to
interact with exposure (e.g., pH), thus increasing the uncertainty in associating effects
with exposure to Pb specifically. A large majority of the available studies of Pb
exposures are laboratory toxicity tests on single species, in which an organism is exposed
to a known concentration of Pb and the effect on a specific endpoint is evaluated. These
studies provide evidence for a temporal sequence between Pb exposure and an effect, an
aspect important in judging causality (Table I of the Preamble). As detailed in the
Framework for Causality (see Preamble), coherence between different types of studies
also provides strong support to a determination of causality. Evidence from laboratory
studies conducted under controlled conditions provides the largest amount of information
used in the causal determinations summarized in Table 6-3. but their coherence with
microcosm and field-based studies plays an important role in those determinations.
Biological gradients (Table I of the Preamble) are often found in studies of the effects of
Pb, and add support to causality where present. For some ecological endpoints, support
for causal determinations is additionally supported by toxicological findings reviewed in
the chapters of the ISA that evaluate evidence for human health effects associated with
Pb exposure, particularly when a common mode of action is documented.
The amount of Pb in ecosystems is a result of a number of inputs and it is not currently
possible to determine the contribution of atmospherically-derived Pb from total Pb in
terrestrial, freshwater or saltwater systems. The causal determinations are, therefore, not
specific to Pb from atmospheric deposition since atmospherically-derived Pb may
ultimately be present in water, sediments, soils and biota (Section 6.2 and Figure 6-1).
The causal determinations encompass findings of studies at concentrations of Pb reported
from environmental media (Table 6-2). and up to one to two orders of magnitude above
the range of these values (Table II of the Preamble). Studies at the upper range of Pb
concentrations are generally conducted at and near heavily exposed sites such as mining
and metal industries-disturbed areas. Studies at those higher concentrations were used
6-263
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only when they were part of a range of concentrations that also included more typical
values, or when they informed understanding of modes of action and illustrated the wide
range of sensitivity to Pb across taxa.
The exposure values at which Pb elicits a specific effect in terrestrial and aquatic systems
are difficult to establish, due the influence of other environmental variables on Pb
bioavailability and toxicity and to substantial differences among biological species in
their sensitivity to Pb. In the 1977 Pb AQCD (U.S. EPA. 1977). no correlation could be
established between toxic effects in invertebrates, fish, birds or small mammals and
environmental concentrations of Pb. At the time of the 1986 Pb AQCD additional data
were available on toxicity but there was still little information on the exposure values that
can cause toxic effects in small mammals or birds (U.S. EPA, 1986b). In the
2006 Pb AQCD (U.S. EPA. 2006c) several studies on effects of Pb exposure on natural
ecosystem structure and function advanced the characterization of Pb levels in the
environment that occur near contaminated sites (i.e., smelters, mining, industry).
According to the 2006 Pb AQCD, natural terrestrial ecosystems near significant Pb
sources exhibited a number of ecosystem-level effects, including decreased species
diversity, changes in floral and faunal community composition, and decreasing vigor of
terrestrial vegetation. These findings were summarized in Table AX7-2.5.2 of the Annex
to the 2006 Pb AQCD (U.S. EPA. 2006c). The 2006 Pb AQCD concluded that, in
general, there was insufficient information available for single materials in controlled
studies to permit evaluation on higher levels of biological organization (beyond the
organism). Furthermore, Pb rarely occurs as a sole contaminant in natural systems
making the effects of Pb difficult to ascertain. Recent information available since the
2006 Pb AQCD, includes additional field studies in both terrestrial and aquatic
ecosystems, but the connection between air concentration and ecosystem exposure
continues to be poorly characterized for Pb and the contribution of atmospheric Pb to
specific sites is not clear.
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Table 6-3 Summary of Pb causal determinations for plants, invertebrates and
vertebrates.
Level
Effect
Terrestrial3 Freshwater3 Saltwater3
Community
and Ecosystem
1 1 Population-Level Endpoints
Sub-organismal Organism-Level Responses
Responses
Community and Ecosystem Effects
Reproductive and Developmental Effects-Plants
Reproductive and Developmental Effects-
Invertebrates
Reproductive and Developmental Effects-
Vertebrates
Growth-Plants
Growth-Invertebrates
Growth-Vertebrates
Survival-Plants
Survival- Invertebrates
Survival- Vertebrates
Neurobehavioral Effects-Invertebrates
Neurobehavioral Effects- Vertebrates
Hematological Effects-Invertebrates
Hematological Effects-Vertebrates
Physiological Stress-Plants
Physiological Stress-Invertebrates
Physiological Stress-Vertebrates
Likely Causal
Inadequate
Causal
Causal
Causal
Likely Causal
Inadequate
Inadequate
Causal
Likely Causal
Likely Causal
Likely Causal
Inadequate
Causal
Causal
Likely Causal
Likely Causal
Likely Causal
Inadequate
Causal
Causal
Likely Causal
Causal
Inadequate
Inadequate
Causal
Causal
Likely Causal
Likely Causal
Likely Causal
Causal
Likely Causal
Likely Causal
Likely Causal
Inadequate
Inadequate
Suggestive
Inadequate
Inadequate
Inadequate
Inadequate
Inadequate
Inadequate
Inadequate
Inadequate
Inadequate
Suggestive
Inadequate
Inadequate
Suggestive
Inadequate
Conclusions are based on the weight of evidence for causal determination in Table II of the ISA Preamble. Ecological effects
observed at or near ambient Pb concentrations measured in soil, sediment and water in the most recent available studies (Table
6-2). were emphasized and studies generally within one to two orders of magnitude above the reported range of these values were
considered in the body of evidence for terrestrial (Section 6.3.12). freshwater (Section 6.4.12) and saltwater (Section 6.4.21)
ecosystems.
6-265
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6.6 Supplemental Material
Table 6-4 Recent evidence for Pb effects on terrestrial plants, invertebrates and vertebrates; growth,
reproduction and survival.
Species
Exposure Exposure
Concentration Concentration
(Nominal) (Measured)
Exposure
Method
Modifying
factors Effects on Endpoint
Reference3
(Published since
Effect the 2006 Pb
Concentration AQCD)
Plants
Buckwheat
(Fagopyrum
esculentum)
Contaminated
soil:
HCI extractable:
6,643 mg Pb/kg
Acetate
extractable:
832 mg Pb/kg
Water leachate:
0.679 mg Pb/kg
Control soil:
HCI extractable:
5 mg Pb/kg
Acetate
extractable: ND
Water leachate:
ND
Plants were
grown for 8
weeks in
contaminated
soil collected
from a shooting
range, and
control soil.
Contaminated Growth:
soil No effect on growth
Sand: 62.3% Survival:
Silt: 36.7% No effect on survjval
Clay: 1.0%
pH: 6.0
CEC: 13.0
Control soil
Sand: 87.7%
Silt: 12.3%
Clay: ND
pH: 6.3
CEC: 7.6
Tamura et al.
(2005)
6-266
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Table 6-4 (Continued): Recent evidence for Pb effects on terrestrial plants, invertebrates and vertebrates; growth,
reproduction and survival.
Species
Canola
(Brassica
napus)
Chinese
cabbage
(Brassica
pekinensis)
Corn (Zea
mays)
Exposure Exposure
Concentration Concentration Exposure Modifying
(Nominal) (Measured) Method factors
0; 22; 45; and Plants of four
67 mg Pb/kg cultivars were
grown for 40
days in soil
amended with
Pb chloride.
46; 874; 1,703 Plants were
mg Pb/kg dry grown for 12
soil days in soil
amended with
Pb acetate.
0; 0.007; 0.7; 7 Seeds were
mg Pb/L germinated on
paper soaked
in P-sulfate.
Plants were
grown for 21
days in washed
sand with
Pb-sulfate
amended
nutrient
solution.
Effect
Effects on Endpoint Concentration
Growth:
Shoot and root dry weight
decreased with
increasing Pb
Zn, Cu, Fe, Mn content
decreased with
increasing Pb.
N, P, K, and Ca2+ content
decreased to a lesser
degree.
Growth:
Shoot biomass
decreased with
increasing Pb (91% and
84% of lowest exposure).
Reproduction:
Germination%,
germination index, plant
decreased with
increasing Pb.
Growth:
Shoot length, plant dry
weight, water use
efficiency decreased with
increasing Pb.
Reference3
(Published since
the 2006 Pb
AQCD)
Ashraf et al.
(2011)
Xiong et al.
(2006)
Ahmad et al.
(2011)
6-267
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Table 6-4 (Continued): Recent evidence for Pb effects on terrestrial plants, invertebrates and vertebrates; growth,
reproduction and survival.
Species
Grass pea
(Lathyrus
sativus)
Lettuce
(Lactuca
sativa)
Exposure Exposure
Concentration Concentration
(Nominal) (Measured)
16; 31; 63; 125;
188mgPb/L
2,000 mg
Pb/kg, but soil
was mixed with
50% V/V
vermiculite
following
amendment
with Pb nitrate.
Tissue Pb:
3.22-233
mg Pb/kg
Exposure
Method
Plants were
grown in soil
amended with
Pb nitrate.
Plants were
grown for 40
days in 21 soils
with varying
native CEC,
OC, pH, and
amorphous Fe
and Al oxides,
which were
then all
amended with
Pb nitrate and
mixed with
50% VAX
\J U /O V/V
vermiculite.
Modifying
factors
After amendment:
n|_|- Q Q 7 Q
[jrn. G.O- / . o
CEC:
3.01-32.04
cmolc/kg
OC:
5 - 30 g/kg
Fe/AI oxides:
0.009-0.195
mol/kg
Effect
Effects on Endpoint Concentration
Reproduction:
Germination decreased
with increasing Pb
(control 100%, highest
exposure 30%).
Chromosomal
abnormalities increased
with increasing Pb
(control 0%, highest
exposure 72%).
Growth:
Shoot length decreased
with increasing Pb
(highest exposure was
50% of control).
Reproduction:
r^&rminatinn ^0 QQQ/f.
OtM 1 1 III IClUUI 1 >JU CJZ. /O
Germination was greater
in amended soils.
Growth:
2.5 -88. 5% of control
In the presence of the
same amount of
Pb(NO3)2,OCwasthe
main determinant of
effects, although CEC
had a strong influence,
but mediated by its effect
on pH and Fe/AI oxides
Reference3
(Published since
the 2006 Pb
AQCD)
Kumar and
Tripathi (2008)
Dayton et al.
(2006)
6-268
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Table 6-4 (Continued): Recent evidence for Pb effects on terrestrial plants, invertebrates and vertebrates; growth,
reproduction and survival.
Species
Mustard
(Brassica
juncea)
Radish
(Raphanus
sativus)
Wheat
( Triticum
aestivum)
Lettuce
(Lactuca
sativa)
Exposure Exposure
Concentration Concentration Exposure Modifying
(Nominal) (Measured) Method factors
0;31; 62; 124; Plants were
186; 249; 311 grown for 60
mg Pb/L days in field
soil amended
with
Pb acetate.
0; 21; 105 Plants were
mg Pb/L grown for 35
days in sand
with a full
nutrient
solution
amended with
Pb nitrate.
69-9,714 Wheat plants pH: 4.25-7.26
mg Pb/kg were grown for OC: 6 2-47.6%
6 weeks in
undisturbed
core samples
from four
locations in
each two
Pb-contaminat
ed sites
Lettuce seeds
were
germinated in
the leachate.
Effect
Effects on Endpoint Concentration
Growth:
Root and shoot length
decreased with
increasing Pb, and the
decrease was greater
with time.
After 60 days, roots were
two times longer in
controls than in the
highest Pb exposure
shoot length was 75%
greater.
Growth:
Leaf area, root volume,
shoot and root dry weight
decreased with
increasing Pb.
(total dry weight at
21 mg Pb/L was 30%
smaller than control, 52%
smaller at 105 mg Pb/L).
Growth:
No effects were found on
germination or growth of
either species.
Reference3
(Published since
the 2006 Pb
AQCD)
John et al.
(2009)
Gopal and Rizvi
(2008)
Chapman et al.
(2010)
6-269
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Table 6-4 (Continued): Recent evidence for Pb effects on terrestrial plants, invertebrates and vertebrates; growth,
reproduction and survival.
Species
Exposure Exposure
Concentration Concentration Exposure Modifying
(Nominal) (Measured) Method factors
Effects on Endpoint
Effect
Concentration
Reference3
(Published since
the 2006 Pb
AQCD)
Invertebrates
Cabbage
aphid
(Brevicoryne
brassicae)
0.87 mg Pb/L in
watering
solution used
for plants
Aphids were
reared for
several
generations on
radish and
cabbage plants
grown in soil
amended with
Pb nitrate.
Reproduction:
In aphids fed
Pb-contaminated plants,
development time was
longer, and relative
fecundity and rate of
population increase were
lower than in control
aphids.
Survival:
Mortality was higher in
exposed aphids, both
adults and offspring.
Gorur (2007)
6-270
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Table 6-4 (Continued): Recent evidence for Pb effects on terrestrial plants, invertebrates and vertebrates; growth,
reproduction and survival.
Species
Collembolan
(Folsomia
Candida)
Collembolan
(Folsomia
Candida)
Exposure Exposure
Concentration Concentration
(Nominal) (Measured)
Approximate
range was 12
mg Pb/kg soil to
15,000
mg Pb/kg soil;
Pore water
approximately
0.002 Pb/L to
1,OOOmgPb/L
(Concentrations
were measured,
but not
reported).
0; 100; 200;
400; 800; 1,600;
3,200 mg Pb/kg
dry soil
Exposure
Method
Springtails
were reared for
28 days in soil
collected at
seven locations
along each of
three transects
with increasing
Pb concentratio
ns within each
transect
(21 locations).
Lowest
concentration
soils from each
of the three
transects were
then amended
with Pb nitrate
to match the
gradient, and
one set of the
amended
samples were
then leached,
for a total of
57 concentratio
ns of Pb.
Springtails
were reared for
10 days in field
soil amended
with
Pb chloride.
Modifying
factors
pH was constant
in transects, but
decreased with
increasing
addition of
Pb(NO3)2 in both
amended and
amended-and-
leached soils.
pH decreased by
3 units in the
highest addition,
regardless of
subsequent
leaching.
Effects on Endpoint
Reproduction:
Reproduction decreased
by up to 50% in transect
soils
Amended soils
Pb concentrations 2,207
mg Pb/kg or lower never
had a significant effect on
reproduction.
Reproduction:
Hatching success
decreased with
increasing Pb.
Effect
Concentration
Transect A 28 day
ECso in mg Pb/kg
dry weight:
native: >5,690
amended: 2,570
amended and
leached: 2,060
Transect B 28 day
ECso in mg Pb/kg
dry weight:
native: >14,400
amended: 3,210
amended and
leached: 2,580
Transect C 28 day
ECso in mg Pb/kg
dry weight:
native: >5,460
amended: 2,160
amended and
leached: 2,320
ECso (hatching):
2,361
mg Pb/kg dry soil
Reference3
(Published since
the 2006 Pb
AQCD)
Lock et al.
(2006)
Xu et al. (2009b)
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Table 6-4 (Continued): Recent evidence for Pb effects on terrestrial plants, invertebrates and vertebrates; growth,
reproduction and survival.
Species
Collembolan
(Sinella
curviseta)
Collembolan
(Paronychiur
us kimi)
Exposure Exposure
Concentration Concentration
(Nominal) (Measured)
0; 100; 200;
400; 800; 1,600;
3,200 mg Pb/kg
dry soil
Toxicity run:
100; 500; 1,000;
2,000 mg Pb/kg
Reproduction
run: 0; 250;
500; 1,000;
2,000; 3,000
mg/kg Pb
Exposure Modifying
Method factors
Springtails
were reared for
28 days in field
soil amended
with
Pb chloride.
Springtails
were reared for
28 days on
artificial soil
amended with
Pb chloride in
two separate
runs.
Effects on Endpoint
There was a small but not
significatnt effect of Pb on
growth, and no effect on
survival and reproduction.
Survival:
Survival decreased with
increasing Pb.
Reproduction:
Offspring production and
instantaneous rate of
increase values
decreased with
increasing Pb.
Effect
Concentration
Survival:
LCio/ECio:
1,838 mg Pb/kg
Reproduction:
LCio/ECio:
642 mg Pb/kg
EC50:
3, 21 2 mg/kg Pb
Body Size:
LCio/ECio:
4,094 mg Pb/kg
Survival LCso:
7 day:
1, 322 mg Pb/kg
28 day:
1, 299 mg Pb/kg
FP=n
CV>50
28 day: 428 mg
Pb/kg
NOEC
reproduction:
EC50 28 day: 428
mg Pb/kg
NOEC: 250 mg
Pb/kg
LOEC: 500 mg
Pb/kg
Reference3
(Published since
the 2006 Pb
AQCD)
Xu et al. (2009a)
Son et al. (2007)
6-272
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Table 6-4 (Continued): Recent evidence for Pb effects on terrestrial plants, invertebrates and vertebrates; growth,
reproduction and survival.
Species
Collembolan
s (Sinella
coeca,
Folsomia
Candida)
Exposure
Concentration
(Nominal)
10; 50; 100;
500; 1,000 mg
Pb/kg
Exposure
Concentration Exposure Modifying
(Measured) Method factors
Springtails
were reared for
42 or 45 days
in artificial soil
amended with
Pb nitrate.
Effects on Endpoint
S. coeca:
Survival:
Mortality significantly
increased with increasing
concentration in adult
population.
Reproduction:
Juvenile production not
significantly compromised
at 10-500 mg Pb/kg,
reduced at 1,000 mg
Pb/kg
Effect
Concentration
LC50:
S. coeca: Could
not be determined
F. Candida: Could
not be determined
EC 50
reproduction:
S. coeca:
490 mg Pb/kg Pb
on dry soil
Reference3
(Published since
the 2006 Pb
AQCD)
Menta et al.
(2006)
F. Candida:
Survival:
Increase in mortality with
increasing concentration
Reproduction:
Juvenile production not
significantly reduced
between 10-500 mg
Pb/kg, significant effect at
1,000 mg Pb/kg
F. Candida:
Could not be
calculated with
accuracy; Ranged
from 500-1,000
mg Pb/kg
6-273
-------
Table 6-4 (Continued): Recent evidence for Pb effects on terrestrial plants, invertebrates and vertebrates; growth,
reproduction and survival.
Species
Earthworm
(Eisenia
andrei)
Earthworm
(Eisenia
fetida)
Exposure Exposure
Concentration Concentration
(Nominal) (Measured)
2,000 mg Pb/kg
in soil
Internal
concentration of
Pbin
earthworms
varied among
the amended
soils, between
29 and 782 mg
Pb/kg dry
weight.
18-9,311
mg Pb/kg
Exposure
Method
Earthworms
were reared for
28 days in 21
soils with
varying native
CEC.OC, pH,
and amorphous
Fe and Al
oxides, which
were then all
amended with
Pb nitrate.
Earthworms
were reared for
14 days in
OECD-
standard
toxicity testing
soil with
7 concentration
s of Pb, either
one or 10
earthworms per
container.
Modifying
factors
After amendment
pH:
3.8-7.8
CEC:
3.01 -32.04
f^rnnl flcn
ui 1 luic/ r\y
/— \/-^ .
oc.
5 - 30 g/kg
Fe/AI oxides:
0.009-0.195
mol/kg
pH decreased
with increasing
Pb nitrate addition
NHs concentration
increased with
Pb concentration
and time
Reference3
(Published since
Effect the 2006 Pb
Effects on Endpoint Concentration AQCD)
Survival: Bradham et al.
Mortality ranged between (2006)
Oand 100%.
In the presence of
potentially lethal amounts
of Pb, the main
determinant of mortality
was pH, with little or no
effect from OC, CEC, or
Fe/AI oxides.
Reproduction:
Reproduction relative to
controls ranged between
Oand 167%.
Effects of Pb on
reproduction are
dependent principally on
Fe/AI oxides, with some
influence of CEC
Survival: LCso (multiple- Currie et al.
Mortality increased from 0 occupancy): (2005)
to 1 00% with increasing 2,662 mg Pb/kg at
Pb, with 100% reached at 7 ^ays anc|
4, 500 mg Pb/kg after 7 0 * _.„ .
days and 14 days. 2^589 mg Pb/kg at
Number of worms per 14 days or
container had no effect 2,827 mg Pb/kg at
on mortality. both 7 and 14
Growth: daVs
Worm weight decreased
with increasing Pb, and
faster in multiple-worm
containers.
6-274
-------
Table 6-4 (Continued): Recent evidence for Pb effects on terrestrial plants, invertebrates and vertebrates; growth,
reproduction and survival.
Species
Earthworm
(Eisenia
fetida)
Earthworm
(Eisenia
fetida)
Exposure Exposure
Concentration Concentration
(Nominal) (Measured)
0; 300; 711; Mean: 79% of
1,687; 2, 249 nominal
mg Pb/kg
Soil 1:
0; 355; 593;
989; 1,650 mg
Pb/kg
Soil 2:
59; 297; 593;
2,965 mg Pb/kg
Soil 3:
386; 771; 1,929;
3,857 mg Pb/kg
Exposure
Method
Earthworms
were reared for
14 days in soil
amended with
five levels of
Pb nitrate and
artificially aged.
Earthworms
were reared for
28 days in
three soils
amended with
five levels of
Pb nitrate
without aging,
after which
they were
removed from
the containers.
Containers
were then kept
in the same
conditions for
another 28
days, after
which cocoons
were extracted.
Modifying
factors
pH 6.72 prior to
amendment
OC 0.7%
CEC11 meq/100
g
PH
6.72; 5.48; 6.75
(prior to
amendment)
OC
0.7; 1.2; 5.2%
CEC
11; 8; 27
meq/100g
Reference3
(Published since
Effect the 2006 Pb
Effects on Endpoint Concentration AQCD)
Survival: Jones et al.
Mortality was only (2009b)
observed at the highest
exposure.
Reproduction: Jones et al.
Soil 1 ' (2009b)
Juvenile and cocoon
count decreased from 19
and 45, respectively, to
near 0 with increasing Pb.
Soil 2:
Cocoon count decreased
to 40% of control at
highest Pb.
Soil 3:
Cocoon count was 0 at all
concentrations.
6-275
-------
Table 6-4 (Continued): Recent evidence for Pb effects on terrestrial plants, invertebrates and vertebrates; growth,
reproduction and survival.
Species
Earthworm
(Pheretima
guillelmi)
Earthworms
(Eisenia
andrei,
Lumbricus
rubellus,
Aporrectode
a caliginosa)
Exposure Exposure
Concentration Concentration
(Nominal) (Measured)
Toxicity run:
0, 1,000; 1,400;
2,000; 2,800;
3,800; 5,400;
7,500 mg Pb/kg
dry weight
Sublethal
toxicity run:
1,000; 1,400;
1,800; 2,500
mg Pb/kg dry
weight
0; 1,000; 3,000;
4,000; 5,000;
7,500; 10,000
mg Pb/kg
Exposure
Method
Earthworms
were reared for
14 days in
OECD-
standard soil
amended with
Pb nitrate.
There were two
runs with
different
concentrations.
Earthworms
were reared for
28 days in
sterilized
Kettering Loam
watered with
Pb nitrate
solution.
Modifying
factors
Temperature:
20 °C (Ł. andrei);
15°C(L rubellus
and A. caliginosa)
pH
Day 7: 4.57-5.83,
Day 28: 4.71-5.83,
increasing with
decreasing Pb
Effects on Endpoint
Survival:
Mortality increased with
increasing Pb.(0% in
control, 100% at 7, 500
mg Pb/kg after 14 days).
Growth:
Weight decreased with
increasing Pb (2.5 g in
control, 1.4 g at 5.400 mg
Pb/kg after 14 days).
Reproduction:
Sperm abnormalities
increased with increasing
Pb (10% of control, 21%
at 2, 500 mg Pb/kg after
14 days).
Growth:
Weight decreased with
increasing concentration
and time, severity of
weight decrease varied
with species.
Survival:
Mortality increased with
increasing concentration
and time, and varied with
species
100% mortality for all
species at higher
concentrations after 28
days.
Reference3
(Published since
Effect the 2006 Pb
Concentration AQCD)
LCso : Zheng and Li
4,285 mg Pb/kg at (2QQE)
7 days and
3,207 mg Pb/kg at
*1 A r\^\ic~
14 days
LCso: Langdon et al.
Ł. andrei: (2005)
5,824 mg Pb/kg
L. rubellus:
2,867 mg Pb/kg
A. caliginosa:
2,747mg Pb/kg
C f* •
EC so-
Łafi/Vf*Of "
. CfffUf d.
2,841 mg Pb/kg
1,303 mg Pb/kg
A. caliginosa:
1, 208 mg Pb/kg
6-276
-------
Table 6-4 (Continued): Recent evidence for Pb effects on terrestrial plants, invertebrates and vertebrates; growth,
reproduction and survival.
Exposure
Concentration
Species (Nominal)
Nematode
(Caenorhabd
itis elegans)
Nematode 5; 10; 16; 21
(Caenorhabd mg Pb/L
itis elegans)
Nematode 0.5; 16; 41 mg
(Caenorhabd Pb/L
itis elegans)
Exposure
Concentration Exposure Modifying
(Measured) Method factors
0.5; 10; 21 Nematodes at
mg Pb/L various
developmental
stages were
exposed to
Pb(NO3)2 for
four hours in
water.
Late larval
nematodes
(L4) were
exposed for
one or three
days.
Nematodes
were placed for
48 hours in
growing
medium with
4 concentration
sofPb.
Nematodes
were placed for
three days in
growth medium
amended with
Pb nitrate.
Effect
Effects on Endpoint Concentration
Reproduction:
Brood size decreased
with increasing Pb, but
the decrease was smaller
with increasing
developmental age.
Generation time
increased with increasing
Pb, and the increase was
smaller with increasing
developmental age.
These effects were
greater in late larval
nematodes when
exposure duration
increased from four hours
to one and three days.
Survival:
No effect
Growth:
Life span, body size
decreased with
increasing Pb.
Reproduction:
Generation time and
brood size increased with
increasing Pb.
All effects were present
and of comparable
magnitude in progeny of
exposed nematodes.
Reference3
(Published since
the 2006 Pb
AQCD)
Guo et al. (2009)
Vigneshkumar et
al. (2013)
Wang and Yang
(2007)
6-277
-------
Table 6-4 (Continued): Recent evidence for Pb effects on terrestrial plants, invertebrates and vertebrates; growth,
reproduction and survival.
Exposure
Concentration
Species (Nominal)
Snail
(Achatina
achatina)
Snail
(Achatina
achatina)
Snail (Theba 0; 50; 100; 500;
pisana) 1,000; 5,000;
10,000; 15,000
mg Pb/kg
Exposure
Concentration
(Measured)
1.33;
70.98;
134.61;
339.40;
674. 86;
1,009.22;
1, 344.39 mg
Pb/kg
0.56;
20.37;
200.42;
1,200.30
mg Pb/kg
Exposure
Method
Snails were
reared for 12
weeks on a diet
amended with
Pb chloride.
Snails from
laboratory
source were
reared for
12 weeks in
bottomless
enclosures at
four locations
within the
grounds of an
abandoned
battery factory.
Snails were
reared for 5
weeks on
Pb-amended
diet.
Modifying
factors
pH: 4.42 -6.29,
decreasing with
increasing Pb
OC: 1.39-3.45%,
decreasing with
increasing Pb
CEC: 3.32-5.37
cmol/kg,
increasing with
increasing Pb.
Effect
Effects on Endpoint Concentration
Survival:
no effect
Growth:
Small decrease in feeding
at highest exposure,
small decrease in weight
gain with increasing Pb
(over 12 weeks, snails in
the highest exposure
gained 12% less weight
than in the lowest
exposure).
Growth:
Feeding, weight gain and
shell thickness all
decreased with
increasing Pb (13; 17;
and 19% lower in highest
exposure than in lowest).
Growth:
Feeding and weight gain
decreased with
increasing Pb and time
(snails in 0 added Pb
gained 45% more weight
than in highest Pb).
Survival:
No effect
Reference3
(Published since
the 2006 Pb
AQCD)
Ebenso and
Ologhobo
(2009a)
Ebenso and
Ologhobo
(2009b)
EI-Gendy et al.
(2011)
6-278
-------
Table 6-4 (Continued): Recent evidence for Pb effects on terrestrial plants, invertebrates and vertebrates; growth,
reproduction and survival.
Species
Exposure
Concentration
(Nominal)
Exposure
Concentration
(Measured)
Exposure
Method
Modifying
factors
Effects on Endpoint
Effect
Concentration
Reference3
(Published since
the 2006 Pb
AQCD)
Snails
(Cantareus
aspersus,
Helix
aspersa)
Total Soil Pb:
1740-2060 mg
Pb/kg
CaCI2
extractable;
4-80 mg Pb/kg
Dissolved
(estimated):
0.007-0.09 mg
Pb/L
Snails were
reared for 7 - 9
weeks in field
soil amended
with varying
amounts of
Pb-sulfate,
clay, peat, and
CaCO3.
Clay content
11 -16%
Organic matter
1.2-10%
PH
4.6-7.49
Growth:
No effect
Survival:
No effect
Pauget et al.
(2011)
Vertebrates
Japanese
quail
(Coturnix
coturnix
japonica)
Drinking water
Pb:
0; 5; 50 mg
Pb/L
Tissue Pb:
0;1.1; 10.7
mg/kg wet
weight
Quail were
given
Pb-amended
water for 7
weeks.
Growth: Nain and Smits
Feed intake and growth (2011)
rate not affected.
Survival:
Morbidity/mortality was
lower in highest exposure
than in control.
Incidence of pericarditis,
airsacculitis, perihepatitis,
and arthritis was lower in
highest exposure than in
control.
6-279
-------
Table 6-4 (Continued): Recent evidence for Pb effects on terrestrial plants, invertebrates and vertebrates; growth,
reproduction and survival.
Species
Pied
flycatchers
(Ficedula
hypoleuca)
Pig (Sus
domestica)
Exposure Exposure
Concentration Concentration
(Nominal) (Measured)
Blood Pb in
nestlings at
mining site
while active:
41 mgPb/100
kg wet weight;
after closing:
29 mg Pb/100
kg wet weight.
Blood Pb in
nestlings at
reference site
while active:
2mg Pb/100 kg
wet weight;
after closing:
0.4 mg Pb/100
kg wet weight.
Feed Pb
control:
unreported
exposed:
10 mg Pb/kg
Blood Pb
control:
1.44|jg/dL
exposed:
2.08 |jg/dL
Exposure Modifying
Method factors
Data were
collected in
wild flycatchers
near a Pb mine
and at a
reference site
for three years
while the mine
was active, and
for three years
five years after
mine closing.
Pigs were
reared for 120
days with
Pb-sulfate-
amended feed.
Effect
Effects on Endpoint Concentration
Reproduction:
Clutch size and breeding
success were lower at the
mine site, but did not
change after closure of
the mine (clutch size
5.6 reference,
4.9 mining site;
breeding success
80% reference site,
76% mining site).
Nestling mortality was
higher at the mine site,
and increased after
closure
(5% reference site,
11% mining site while
active; 11% reference
site, 26% mining site after
closure).
Growth:
Significant decrease in
body weight, average day
gain, average day feed
intake, and feed
efficiency.
Increase in feed
conversion ratio.
Reproduction:
No effect on ovary and
uterus weight
Reference3
(Published since
the 2006 Pb
AQCD)
Berglund et al.
(2010)
Yu et al. (2005)
"References included are those which were published since the 2006 Pb AQCD.
6-280
-------
Xuetal. (2009a)
Xuetal. (2009a)
Sonetal. (2007)
Currieetal. (2005)
Zhengand Li(2009)
Xuetal. (2009a)
Langdon et al. (2005)
Locketal. (2006)
Xuetal. (2009b)
Xuetal. (2009a)
Sonetal. (2007)
Mentaetal. (2006)
Sinella curviseta
Sinella curviseta
Paronychiurus kimi
Eisenia fetida
Pheretima guilelmi
Sinella curviseta
Eisenia andrei
Lumbricus rubellus
Apporectodea caliginosa
Folsomia Candida
(meansof 3 transects)
Folsomia Candida
Sinella curviseta
Paronychiurus kimi
Sinella coeca
/
FP
bL10
Reproducti
Survival
i r
LL50
Reproducti
Growth Inh
Reproducti
V B
on
on
ibiti
on
0
«
A
A
A
A
A
A
on
A
A
A
A
•
«
amended soil
amended soil
amended soil, 7 days
amended soil, 28 days
amended soil, 7 days
amended soil, 14 days
amended soil, 7 days
amended soil, 14 days
amended soil
amended soil
amended soil
amended soil
contaminated soil
amended soil
amended soil, leached
amended soil
amended soil
amended soil
amended soil
0.1
10
100
1000
10000
log (mgPb/kg soil)
Note: Detailed summaries of the individual studies referenced above, including experimental conditions, modifying factors and
responses at higher concentrations are summarized in Table 6-4. Data in Figure 6-4 are limited to EC and LC values published
since the 2006 Pb AQCD. Reference concentrations "A" and "B" are the National mean background soil Pb sampled between 1961
and 1997, as reported in U.S. EPA (2007d. 2006b. 2003b), and the 95th percentile of background soil Pb sampled between 1961
and 1976 in the contiguous U.S. (Shacklette and Boerngen. 1984) (see Table 6-2).
Figure 6-4 Subset of concentration-response data reported in Table 6-4 for
Pb effects on growth, reproduction, and survival in some
terrestrial invertebrates.
6-281
-------
Table 6-5 Recent evidence for Pb effects
reproduction, and survival.
Species
Concentration Exposure Method
on freshwater
Modifying
factors
plants, invertebrates and vertebrates; growth,
Effects on Endpoint Effect Concentration
Reference3
(Published
since the
2006 Pb
AQCD)
Algae/Plants
Blue-green algae
(Spirulina
(Arthrospira)
platensis)
Microalgae
(Scenedesmus
obliquus)
Microalgae
(Chlorella vulgaris)
5,000; 10,000; 10-day exposure to
30,000; 50,000; Pb nitrate in Zarrouk
and liquid medium.
1 00,000 ug Pb/
L (nominal
concentration at
day 0 and then
Pb in media
was measured
every two days
thereafter).
5, 000 to 48 and 96-hour
300,000 ug Pb/ acute toxicity test
L with Pb nitrate in
(nominal) BG11 medium.
Temperature:
25 ± 1 °C
pH:
7.0
Light dark cycle
of 14:10 hours.
Plants were
incubated at
normal room
temperature
(not provided)
Growth: LC5o:
10-day algal growth (measured 75,340 ug/L
turbidimetrically at 560 nm) was
stimulated by 3.7% in the lowest
concentration, growth was
inhibited at higher concentrations
of 30,000; 50,000; and
100,000 ug Pb/L by 40; 49; and
78%, respectively. Chlorophyll a
and b content were significantly
diminished at the three highest
exposures.
Growth: 48 hour EC 5o:
In growth studies (measured as 4,040 ug Pb/L
cell division rate) S. obliquus was S. obliquus
significantly more sensitive to Pb
exposure than C. vulgaris
K y 48hourEC50:
24,500 ug/L
C. vulgaris
Arunakumar
a et al.
(2008)
Atici et al.
(2008)
6-282
-------
Table 6-5 (Continued): Recent evidence for Pb effects on freshwater plants, invertebrates and vertebrates; growth,
reproduction and survival.
Species
Duckweed
(Lemna minor)
Duckweed
(Lemna minor)
Duckweed
(Wolffia arrhiza)
Waterweed
(Elodea
canadensis)
Concentration
100; 200; 490;
900; 2,000;
5,020; 7,990;
and
9,970 ug Pb/L
(measured)
2,070; 10,360;
20,700; and
1 03,600 ugPb/
L
(nominal)
210; 2,120;
20,720; and
207,200 ug Pb/
L (analytically
verified)
1,000; 10,000;
and
1 00,000 ug/L
(nominal)
Exposure Method
4 day or 7-day
exposures to
Pb chloride in static
test conditions with
Jacob culture
medium under
continuous
illumination.
9-day exposures to
Pb nitrate in a
growth chamber on
Knopp's medium
under a 14-hour
photoperiod.
14-day exposure to
Pb as Pb nitrate in
sterile 1/50 dilution
of Hutner's medium,
day to night cycle of
16:8 hours.
Plants were exposed
5 days to Pb as
Pb acetate in a 10%
nutrient solution and
then assayed for
pigment content,
total ascorbic acid,
and protein content.
Modifying
factors
Temperature:
25±2°C
pH
6.0
Temperature:
25±0.5°C
pH:
7.0
Temperature:
25-27°C
pH
6.5-6.7
10% Hoagland &
Arnon nutrient
solution
Effects on Endpoint Effect Concentration
Growth: 4dayEC50:
Growth (measured as biomass) 6,800 ug Pb/L
of the duckweed was promoted 7 ^ay EC50'
up to 103% at 100 ug Pb/L and 5 50o Ug/L '
200 ug Pb/L. However, growth
was inhibited monotonically at all
other test levels with increasing
concentrations. Overall, the
relative growth rate was reduced
to 37-38% at the highest
concentration.
Growth: At lower Pb doses,
growth was slightly stimulated.
Fresh weight was lower by 65%
at the highest dose. Pb-induced
chlorosis occurred and the
enzymes of the antioxidative
system were modified due to Pb
exposure in all concentrations.
Growth:
Biomass decreased
proportionally with increasing Pb
concentration; Chlorophyll a
content was significantly inhibited
at 210 ug Pb/L and greater;
carotenoid, monosaccharide and
protein content significantly
decreased in higher
concentrations.
Growth:
The chlorophyll, carotenoid, and
protein contents of Ł. canadensis
were significantly reduced
following Pb accumulation.
Reference3
(Published
since the
2006 Pb
AQCD)
Dirilgen
(2011)
Paczkowska
et al. (2007)
Piotrowska
et al. (2010)
Dogan et al.
(2009)
6-283
-------
Table 6-5 (Continued): Recent evidence for Pb effects on freshwater plants, invertebrates and vertebrates; growth,
reproduction and survival.
Species
Wetland plants
(Beckmannia
syzigachne,
Alternanthera
philoxeroides,
Juncus effusus,
Oenanthe
javanica,
Cyperus
flabelliformis,
Cyperus
malaccensis,
Polypogon
fugax,
Leersia
hexandra,
Panicum
paludosum,
Neyraudia
reynaudiana)
Modifying
Concentration Exposure Method factors
20,000 |jg/L Field-collected tillers
(nominal) or seedlings (from
Metals were various locations in
analyzed in China) for each
plant samples species were used
in 21-day
experiments to
determine Pb
tolerance as inferred
from measuring the
elongation of the
longest root in a
hydroponic system
in a Pb nitrate
solution.
Effects on Endpoint Effect Concentration
Growth:
Root elongation was significantly
reduced in a number of wetland
species (6. syzigachne,
J. effusus, O. javanica,
C. flabelliformis, C. malaccensis,
and N. reynaudiana). Metal
tolerance was related to root
anatomy and spatial pattern of
radical oxygen loss.
Reference3
(Published
since the
2006 Pb
AQCD)
Deng et al.
(2009)
Invertebrates
Rotifer
(Brachionus
calyciflorus)
67; 194; 284; Cysts of rotifers Temperature:
390; and were obtained from 25 ± 1 °C
700 ug Pb/L Florida Aqua farms p|_|.
(measured) in Dade City,
Florida, U.S. Tests S'19
with Pb nitrate were
performed in total
darkness for 48
hours.
Reproduction: EC 20 for number of
The total number of rotifers and rotifers: 125 ug Pb/L
the intrinsic rate of population 48-hour EC2o for
increase exhibited concentration- intrinsic rate of
dependent responses at the end population increase:
of the 48 hour incubation period. 307 ug Pb/L
NOEC: 194ugPb/L
LOEC: 284 ug Pb/L
Grosell et al.
(2006b)
6-284
-------
Table 6-5 (Continued): Recent evidence for Pb effects on freshwater plants, invertebrates and vertebrates; growth,
reproduction and survival.
Species
Rotifer
(Brachionus
patulus)
Rotifer
(Euchlanis
dilatata)
Concentration
1,250; 2,500;
4,000; 5,000;
and
8.000 |jg Pb/L
(nominal) for
acute toxicity
tests.
Chronic
exposures used
nominal
concentration of
60 and
600 ug Pb/L
with varying
turbidity levels.
0.1; 0.5; 50;
100; 250; 1,000;
2,500 ug Pb/L
(analytically
verified, actual
concentrations
not reported)
Exposure Method
24-hour exposures
to Pb chloride in the
presence and
absence of
sediments using
rotifers originally
isolated from the
Chimaliapan
wetland, Toluca,
Mexico.
Three week chronic
toxicity tests were
also conducted.
48-hour acute
toxicity tests with
rotifer neonates
exposed to
Pb nitrate in
synthetic moderately
hard water. Adult
rotifers were
collected in a
reservoir in
Aguascalientes,
Mexico.
Modifying
factors
Temperature:
20 °C
Temperature:
25±2°C
pH:
7.5
Hardness:
80-1 00 mg/L
PaPPU
WC1WW3
Effects on Endpoint
Reproduction:
In chronic tests, net reproductive
rate and rate of population
increase decreased under
conditions of increasing turbidity
and Pb concentration.
Survival:
24-hour LC5o reported for this
species. In chronic tests,
average life span and life
expectancy at birth decreased
under conditions of increasing
turbidity and Pb concentration.
Survival:
Based on 48-hour LCso
E. dilatata is among the most
sensitive rotifer species to Pb.
E. dilatata may be a more
suitable test organism for
ecotoxicology in Mexico, where
this study was conducted,
instead of D. magna, a species
that is not been found in Mexico
reservoirs.
Effect Concentration
24-hour LC5o:
6,150ug/L
48-hour
NOEC: 0.1 ug/L
LOEC: 0.5 ug/L
LC 50 : 35 ug/L
(estimated from
analytically verified
concentrations)
Reference3
(Published
since the
2006 Pb
AQCD)
Garcfa-Garcfa
et al. (2007)
Arias-Almeida
and Rico-
Martinez
(2011)
6-285
-------
Table 6-5 (Continued): Recent evidence for Pb effects on freshwater plants, invertebrates and vertebrates; growth,
reproduction and survival.
Species
Cladoceran
(Ceriodaphnia
dubia)
Cladoceran
(Ceriodaphnia
dubia)
Cladoceran
(Diaphanosoma
birgei)
Concentration
Measured but
not reported.
Predicted
concentration of
major Pb
chemical
species in the
natural water
bioassays is
provided in
Table 4 of
Esbaugh et al.
(2011).
A range of 5 to
6Pb
concentrations
(measured but
not reported)
were prepared
with varying pH,
hardness and
alkalinity.
2,000 to
5,500 ug Pb/L
(analytically
verified)
Exposure Method
Acute toxicity of Pb
to C. dubia was
assessed in 48-hour
exposures to two lab
generated reference
waters and eight
natural waters from
across North
America selected to
include a range of
water quality
parameters. Waters
were spiked with
varying
concentrations of Pb
as Pb nitrate.
Chronic 7-day static
renewal 3 times per
week in 2:1
dechlorinated,
aerated tap
waterdeionized
water to determine
the effects of
hardness (as CaSO4
and MgSO4),
alkalinity, pH, and
DOM on Pb toxicity.
24-hour exposure to
Pb chloride in
moderately hard
water.
Modifying
factors
Temperature:
26°C
Water chemistry
of the field-
collected waters
are reported in
Table 1 of
Esbaugh et al.
(2011) including
pH (range 5.5 to
8.5), Ca (range
24to1,934uM),
DOC (range 36
to 1,244uM)and
hardness (range
4 to 298 mg/L)
Temperature:
25 °C
pH:
6.4-8.2
1— 1 o i-rJ n & C C '
ndl Ul Icoo.
22-524 mg/L
Temperature:
23 °C
pH:
7.0-7.5
Effects on Endpoint
Survival:
LCso values ranged from 29 to
1,180 ug Pb/L. Sensitivity to Pb
varied greatly with water
chemistry. DOC was correlated
with protection from acute
toxicity.
Survival:
DOM and alkalinity have a
protective effect against chronic
toxicity of Pb. CaSO4 and
MgSO4 do not have a protective
influence of water hardness; Pb
toxicity increased at elevated
Ca2+ and Mg2+. Low pH
increases the toxicity of Pb.
Reproduction:
Increased DOC leads to an
increase in mean ECso for
reproduction ranging from
approximately 25 ug Pb/L to
>500 ug Pb/L.
Survival:
LCso reported
Reference3
(Published
since the
2006 Pb
Effect Concentration AQCD)
48-hour LCso range: Esbaugh et al.
29 to 1,1 80 ug Pb/L (2011)
NOEC range: 18 to
<985 ug Pb/L.
LOEC range: 52 to
1, 039 ug Pb/L
Control base water Mager et al.
EC20: (201 1a)
45 ug Pb/L
5.0 mM CaSC-4 EC2o:
22 ug Pb/L
32 mg/L DOM EC20:
COO I in Dk/l
OZO Ug ru/L
2.5 mM NaHCOs EC20:
73 ug Pb/L
Additional EC20 and all
ECso values were
reported in the study.
24 hour LC5o Garcfa-Garcfa
3,1 60 ug Pb/L et a I. (2006)
6-286
-------
Table 6-5 (Continued): Recent evidence for Pb effects on freshwater plants, invertebrates and vertebrates; growth,
reproduction and survival.
Species
Cladoceran
(Moina micrura)
Cladoceran
(Alona
rectangular)
Cladoceran
(Daphnia magna)
Cladoceran
(Daphnia pulex)
Concentration
1,000 to
8,000 |jg Pb/L
(analytically
verified)
3,000 to
1 0,000 |jg Pb/L
(analytically
verified)
Acute test:
Concentrations
not provided
Chronic test:
25 ug Pb/L
250 |jg Pb/L
2,500 |jg Pb/L
(nominal)
Acute test:
250; 500; 1,000;
2,000;
5,000 |jg Pb/L
(nominal)
Chronic test:
250; 500;
1, 000 |jg Pb/L
(nominal).
Exposure Method
24-hour exposure to
Pb chloride in
moderately hard
water
24-hour exposure to
Pb chloride in
moderately hard
water
24 hour acute
toxicity test and 21
day toxicity test with
Pb nitrate, static
renewal every two
days.
48-hour acute
toxicity test and two
21 -day exposures to
Pb nitrate in
dechlorinated tap
water.
Modifying
factors
Temperature:
23 °C
pH:
7.0-7.5
Temperature:
23 °C
pH:
7.0-7.5
Temperature:
20 ± 1 °C
Effects on Endpoint Effect Concentration
Survival: 24 hour LC so
M. micrura was more sensitive to 690 ug Pb/L
Pb than D. birgei and
A. rectangular.
Survival: 24 hour LC so
A. rectangular was more 7,000 ug Pb/L
resistant to Pb than the other
species tested
Reproduction: 24-hour EC50
Significant concentration- (immobility):
dependent decrease in number 18 153 ug Pb/L
of neonates per female;
Significant long-term effects on
reproduction. Negative
correlation between hemoglobin
gene expression and
reproduction outcomes.
Reproduction: 48-hour LC 50:
Reproduction rates (cumulative 4 goo ug Pb/L
neonates) significantly
decreased at 1,000 ug Pb/L in
the first chronic toxicity test and
at 500 ug Pb/L in a second test.
Survival:
LCso values reported
Reference3
(Published
since the
2006 Pb
AQCD)
Garcfa-Garcfa
et al. (2006)
Garcfa-Garcfa
et al. (2006)
Ha and Choi
(2009)
Theegala et
al. (2007)
6-287
-------
Table 6-5 (Continued): Recent evidence for Pb effects on freshwater plants, invertebrates and vertebrates; growth,
reproduction and survival.
Species
Ostracod
(Stenocypris
major)
Midge
(Chironomus
tentans)
Concentration
475; 1,160;
3,410; 4, 829;
8,972 ug Pb/L
(measured)
42; 109; 497;
724 ug Pb/L
(measured)
Exposure Method
96-hour static
renewal with
Pb nitrate in
dechlorinated tap
water. Ostracods
were collected from
a filter system of a
fish pond in Bangi,
Selangor, Malaysia.
27 day exposure
with 8-day old larvae
to Pb-nitrate in a
mixture of tap water
and deionized water
supplemented with
Ca2+ (as CaSO4)
Modifying
factors Effects on Endpoint
Temperature: Survival:
28-30 °C LC50 values reported
pH:
6.5± 0.01
Conductivity:
244.3 ±0.6
uS/cm
DO: 6.3 ±0.06
mg/L
Total hardness:
15.6 mg/L as
CaCO3
Light dark cycle
of 12:12 hours.
Temperature: Survival:
23 ± 1 °C Significantly reduced at the two
pl_|. highest concentrations
7 Q
/ .y
Reference3
(Published
since the
2006 Pb
Effect Concentration AQCD)
24-hour LCso: Shuhaimi-
6 583 uq Pb/L Othman et al.
4«h ir <2011b)
4o-nour LOso-
2,886 ug Pb/L
72-hour LCso:
1,491 ugPb/L
96-hour LC50:
526 ug Pb/L
NOEC: 109 ug Pb/L Grosell et al.
LOEC: 497 ug Pb/L (2Q06b)
6-288
-------
Table 6-5 (Continued): Recent evidence for Pb effects on freshwater plants, invertebrates and vertebrates; growth,
reproduction and survival.
Species
Midge
(Chironomus
dilutus, formerly
C. tentans)
Midge
(Chironomus
riparius)
Concentration Exposure Method
29; 57; 75; 1 1 5; 96-hour static
128; renewal test
152ugPb/L 20 day midge life
(measured) cyde test in
Pb chloride spiked
water, flow through,
and emergence at
55 days.
The C. dilutus
culture was initially
started with egg
cases from Aquatic
Biosystems, Fort
Collins, CO, USA.
Logarithmic 24-hour acute
range from 0 to toxicity tests with
25,000 ug Pb/L first-instar larvae
(nominal) exposed to
Pb nitrate in
synthetic soft water.
C. riparius culture
was from egg
masses from
Environment
Canada.
Modifying
factors
Average ± SD
(range):
Temperature:
22.2 ± 1.0 °C
(1 9.7-24.4 °C)
pH:
7.26 ±0.21
(6.9-7.7)
Hardness:
32 ± 3.2 mg/L
as CaCO3
Alkalinity:
31 ± 3.0 mg/L
as CaCO3
Conductivity:
76 ± 4.9 us
DO:
7.8 ± 0.8 mg/L
Temperature:
20 °C
Water hardness:
8 mg/L of
CaCO3,
Effects on Endpoint
Growth:
Growth and emergence
decreased as concentration
increased.
Reproduction:
No effect
Survival:
No effect
Survival:
Concentration-dependent
decrease in survival with
increasing Pb.
Of the five metals tested in the
study (Cd, Cu, Pb, Ni, Zn), Pb
was most toxic to first instar
C. riparius.
Reference3
(Published
since the
2006 Pb
Effect Concentration AQCD)
96-hour LC5o: Mebane et al.
3,323 ug Pb/L (2008)
Survival
NOEC: 152ugPb/L
LOEC:>152ug Pb/L
MATC: >1 52 ug Pb/L
Weight
NOEC: 57 ug Pb/L
LOEC: 75 ug Pb/L
MATC: 65 ug Pb/L
EC10: 15 ug Pb/L
EC20: 28 ug Pb/L
Fecundity
NOEC: 152ugPb/L
LOEC: >1 52 ug Pb/L
MATC: >1 52 ug Pb/L
ECi0: >152 ug Pb/L
EC20: >152 ug Pb/L
Emergence
NOEC: 115ug Pb/L
LOEC: 128ug Pb/L
MATC: 121 ug Pb/L
ECi0: 28 ug Pb/L
EC20: 55 ug Pb/L
24-hour LC 5o: Bechard et al.
613 ug Pb/L (2008)
6-289
-------
Table 6-5 (Continued): Recent evidence for Pb effects on freshwater plants, invertebrates and vertebrates; growth,
reproduction and survival.
Species Concentration
Midge 430; 580,;
(Chironomus 1,330; 2,460;
j'avanus) 7>670 U9 Pb/L
(measured)
Midge 100; 290; 510;
(Culicoides furens) 80°'<
2,800 ug Pb/L
(measured)
Exposure Method
4-day exposure with
fourth instar larvae
to Pb nitrate in
aerated, filtered,
dechlorinated tap
water with static
aerated renewal at
2 days. Larvae were
collected from a filter
system of a fish
pond in Bangi,
Selangor, Malaysia.
Series of 96-hour
exposures to
Pb chloride in
dechlorinated water
under several
temperature ranges.
Modifying
factors
Temperature:
28-30 °C
pH:
6.51 ±0.01
Conductivity:
244.3 ±0.6
uS/cm
DO:
6.25 ± 0.06 mg/L
Total hardness
(Mg2+ and Ca2+):
15.63 ±2.74
mg/L
as CaCO3
1st experiment:
Temperature:
25-28 °C
2nd experiment:
Temperature:
20-26 °C
3rd experiment:
Temperature:
10, 15, 20, 23,
25, 28, 30, 35,
40 ± 0.5 °C
Effects on Endpoint
Survival:
LCso values reported for this
species.
Survival:
Higher and lower temperatures
brought about increased
toxicities.
LCso values generally increased
in 10-25°C and decreased in
28-40°C.
40°C temperature produced
100% mortality.
Reference3
(Published
since the
2006 Pb
Effect Concentration AQCD)
24 hour LCso: Shuhaimi-
20,490 ug Pb/L Othman et al.
>18hourLC-o' (201 1c)
6,530 ug Pb/L
72-hour LC50:
1, 690 ug Pb/L
96 hour LC50:
790 ug Pb/L
96 hour LCso values: Vedamanikam
28-25 °C*: 400 ug Pb/L and Shazilli
26-20 °C*: 300 ug Pb/L
25 °C: 400 ug Pb/L
35 °C-25 °C*:
500 ug Pb/L
23 °C: 700 ug Pb/L
20 °C: 400 ug Pb/L
15°C:400|jgPb/L
10°C:357|jgPb/L
"temperature decreased
over the duration of the
experiment
6-290
-------
Table 6-5 (Continued): Recent evidence for Pb effects on freshwater plants, invertebrates and vertebrates; growth,
reproduction and survival.
Species
Midge
(Chironomus
plumosus)
Oligochaete worm
(Lumbriculus
variegatus)
Concentration
3,000; 5,400;
8,200; 30,000;
54,000 |jg Pb/L
(measured)
1,300; 3,200;
8,000; 20,000;
50,000 |jg Pb/L
(nominal)
Exposure Method
Series of 96-hour
exposures to
Pb chloride in
dechlorinated water
under several
temperature ranges.
24 and 48-hour
exposures to
Pb nitrate spiked
water from
Lake Vesijarvi,
Finland
Modifying
factors
1st experiment:
Temperature:
25-28°C
2nd experiment:
20-26°C
3rd experiment:
10, 15,20,23,
25, 28, 30, 35,
40±0.5°C
Temperature:
20 °C
Effects on Endpoint
Survival:
Higher and lower temperatures
brought about increased
toxicities
40°C temperature produced
100% mortality
LC5o values generally increased
in 10-25°C and decreased in
28-40°C
Survival:
48-hour LCso reported
Effect Concentration
96 hour LCso values:
28-25°C*:
1 6,200 ug Pb/L
26-20°C*:
8,300 ug Pb/L
25 °C: 9,500 ug Pb/L
35 °C: 700 ug Pb/L
30 °C: 700 ug Pb/L
28 °C: 900 ug Pb/L
25 °C: 900 ug Pb/L
23 °C: 700 ug Pb/L
20 °C: 600 ug Pb/L
15°C:600ug Pb/L
10°C:500ug Pb/L
"temperature decreased
over the duration of the
experiment
48-hour LC5o:
5,200 ug Pb/L
Reference3
(Published
since the
2006 Pb
AQCD)
Vedamanikam
and Shazilli
(2008a)
Penttinen et
al. (2008)
6-291
-------
Table 6-5 (Continued): Recent evidence for Pb effects on freshwater plants, invertebrates and vertebrates; growth,
reproduction and survival.
Species
Mayfly
(Baetis
tricaudatus)
Mosquito
(Culex
quinquefasciatus)
Concentration
69; 103; 160;
222; 350;
546 |jg Pb/L
(measured)
Acute test:
100; 150; 200;
250 ug Pb/L
(analytically
verified)
50; 150; and
200 ug Pb/L
for reproductive
studies
Exposure Method
96-hour static
renewal test and 10
day chronic study
with aerated
Pb chloride spiked
water, static renewal
every 48 hours.
Mayflies were
collected from the
South ForkCoeur
d'Alene River, Idaho.
24-hour acute
toxicity test and
several tests to
assess reproductive
endpoints. All tests
were conducted with
Pb nitrate in distilled
water
Modifying
factors
Mean ± SD
Temperature:
9.3 ± 0.67 °C
ni_i-
pi-i.
60/1 + r\ -to
.OH Z U. 1 O
Hardness:
20.7±0.58mg/L
oc PaPm
do \^a\^\~JO
Alkalinity:
19.8 ± 1.04mg/L
oc PaPm
do \^a\^\~JO
Conductivity:
47.7 ± 1.72 MS
DO:
10.1 ±0.45mg/L
Temperature:
25 ± 2 °C
pH:
7
Effects on Endpoint
Growth:
Consistent dose-dependent
reductions in mayfly growth (as
number of molts); growth
decreased with increased Pb
exposure.
Survival:
96-hour EC5o reported for this
species. Reduced molting
endpoint more sensitive than
mortality endpoint.
Reproduction:
Hatching rate significantly
decreased, lower emergence
rates, larval development from
L1 to adults took longer.
Survival:
24 hour LC5o reported
Effect Concentration
96-hour EC50
664 ug Pb/L
Survival:
NOEC: 222 ug Pb/L
LOEC: 350 ug Pb/L
MATC: 279 ug Pb/L
EC10: 169 ug Pb/L
EC20: 230 ug Pb/L
Molting :
NOEC: 103ugPb/L
LOEC: 160ug Pb/L
MATC: 130ugPb/L
ECi0: 37 ug Pb/L
EC20: 66 ug Pb/L
Survival:
24 hour LC5o:
180ugPb/L
Reference3
(Published
since the
2006 Pb
AQCD)
Mebane et al.
(2008)
Kitvatanachai
et al. (2005)
6-292
-------
Table 6-5 (Continued): Recent evidence for Pb effects on freshwater plants, invertebrates and vertebrates; growth,
reproduction and survival.
Species
Neosho mucket
(Lampsilis
rafinesqueana)
Concentration
Concentrations
were measured
and used to
calculate ECso
values, reported
in supplemental
data.
Exposure Method
24 and 48-hour
exposure with 5 day
old juveniles
obtained from adults
collected from
Spring River, KS,
U.S.
Modifying
factors
Temperature:
20 ± 1 °C
pH:
7.2-7.6
DOC:
>7.0 mg/L
Hardness:
40-48 mg/L
as CaCO3
Alkalinity:
30-35 mg/L
as CaCO3
Effects on Endpoint
Survival:
Neosho mucket is a candidate
species for U.S. federal
endangered and threatened
status. Toxicity testing with newly
transformed juveniles indicated
that this species is sensitive to
Pb exposure.
Effect Concentration
24 hour EC50:
5 day old juveniles
>507 ug Pb/L
48 hour EC50:
5 day old juveniles:
188ugPb/L
Reference3
(Published
since the
2006 Pb
AQCD)
Wang et al.
(201 Of)
6-293
-------
Table 6-5 (Continued): Recent evidence for Pb effects on freshwater plants, invertebrates and vertebrates; growth,
reproduction and survival.
Species Concentration
Fatmucket mussel For 28-day
(Lampsilis exposure:
siliquoidea) 0.04; 2.9; 6.1;
17; 36;
83 ug Pb/L
(measured)
Exposure Method
24 and 48-hour
exposure with
glochidia, 96-hour
exposure with 5 day
old, and 2 or 6
month old juveniles
and 28-day
exposure with 2 or 4
month old mussels
in reconstituted soft
water. Tests were
conducted with
glochidia and
juveniles obtained
from adults collected
from the Silver Fork
of Perche Creek,
MO, U.S.
Modifying
factors
Temperature:
20 ± 1 °C
pH:
7.2-7.6
DOC:
>7.0 mg/L
Hardness:
40-48 mg/L
as CaCO3
Alkalinity:
30-35 mg/L
as CaCO3
Effects on Endpoint
Growth:
Growth of juvenile mussels in the
17 ug Pb/L concentration was
statistically significantly reduced
compared to growth in the
controls at the end of 28 days.
Growth was not assessed in the
higher concentrations due to
mortality.
Survival:
The 24-hour EC5o values for
glochidia and 96-hour ECso
values for 2 and 6 month old
juveniles were much higher than
96 hour LC5o value for 5 day old
newly transformed juveniles.
Genus mean chronic value was
the lowest value ever reported
for Pb. Survival was based on
foot movement within a 5-minute
observation period.
Effect Concentration
24 and 48 hour EC50:
glochidia
>400ug Pb/L (test 1)
>299 ug Pb/L (test 2)
48 hour EC5o:
5 day old juveniles
465 ug Pb/L (test 1)
392 ug Pb/L (test 2)
96 hour EC50:
5 day old juveniles
1 42 ug Pb/L (test 1)
298 ug Pb/L (test 2)
24 and 48 hour ECso:
2 month old juveniles
>426 ug Pb/L
4 day EC5o:
>83 ug Pb/L
10 day EC50:
>83 ug Pb/L
21 day EC50:
29 ug Pb/L
28 day EC5o:
20 ug Pb/L
Reference3
(Published
since the
2006 Pb
AQCD)
Wang et al.
(201 Of)
28-day NOEC
Juvenile fatmucket:
6.1 ug/L
28-day LOEC
Juvenile fatmucket:
17 ug/L
Genus mean chronic
value
1 Dug Pb/L
6-294
-------
Table 6-5 (Continued): Recent evidence for Pb effects on freshwater plants, invertebrates and vertebrates; growth,
reproduction and survival.
Species
Snail
(Lymnaea
stagnalis)
Snail
(Lymnaea
stagnalis)
Snail
(Marisa
cornuarietis)
Concentration
4; 12; 16; 42;
113; and
245 |jg Pb/L
(measured)
1st
Experiment:
<0.5 (control),
2.7 and
1 8.9 |jg Pb/L
(measured)
2nd
Experiment:
1.3 and
7.5 ug Pb/L
(measured)
5,000;10,000;
and
1 5,000 ug Pb/L
(nominal)
Exposure Method
30-day exposure
with newly hatched
snails (<24-hour old)
in artificial fresh
water with
Pb nitrate.
Pb exposures were
performed with
juvenile snails (~ 1
g) for 21 days and
then 14 days in
dechlorinated tap
water under flow-
through conditions.
5-day, 6-day, and
10-day exposure to
Pb chloride in
deionized or double-
distilled water. Snail
strain used for egg
production was from
the Zoological
Institute in Frankfurt,
Germany.
Modifying
factors
Temperature:
23 ± 1 °C
Dechlorinated
City of Miami tap
water
([Na+]~1.1
mmol/L
[Ca2+] -0.31
mmol/L
[Of] -1.03
mmol/L
[HCO3~] -0.68
mmol/L,
[DOC]~200
umol/L
pH~7.7 at room
temperature
Temperature:
24 ± 1 °C
pH:
-7.5
Conductivity:
-800 uS/cm
Effects on Endpoint Effect Concentration
Growth: EC2o <4 ug Pb/L
Newly hatched snails exhibited NOEC' 12 ug Pb/L
greatly reduced growth in . ___ „„ „, „
response to Pb exposure LOEC: 16 M9 Pb/L
Survival:
No Pb-induced mortality was
observed.
Growth:
In juveniles exposed to 18.9 ug/L
Pb for 21 days, Ca2+ influx was
significantly inhibited and model
estimates indicated 83%
reduction in growth of newly
hatched snails after 30 days at
this exposure concentration
Survival:
No Pb-induced mortality was
observed
Reproduction: LOEC: 10,000 ug Pb/L
Significant delay in hatching at
1 0,000 ug Pb/L
Growth:
Significantly delayed
development (reduced visible
tentacles, eye formation) at
1 5,000 ug Pb/L.
No effect on fresh weight.
Survival:
Significantly increased mortality
at 15, 000 ug Pb/L
Reference3
(Published
since the
2006 Pb
AQCD)
Grosell et al.
(2006b)
Grosell and
Brix (2009)
Sawasdee
and Kohler
(2010)
6-295
-------
Table 6-5 (Continued): Recent evidence for Pb effects on freshwater plants, invertebrates and vertebrates; growth,
reproduction and survival.
Species
Snail
(Biomphalaria
glabrata)
Prawn
(Macrobrachium
lancesteri)
Concentration Exposure Method
50; 100; and 96-hour acute
500 ug Pb/L laboratory bioassays
(nominal)
22; 31; 48; 126; 4-day exposures in
170ugPb/L Pb chloride in
(measured) aerated, filtered,
dechlorinated tap
water, static aerated,
renewal at 2 days.
Prawns were
purchased from
aquarium shops in
Bani, Selangor,
Malaysia.
Modifying
factors
Dechlorinated
continuously
aerated tap
water:
Temperature:
22 °C
pH:
7.1 ±0.2
Total hardness:
OC _|_ O ,-v-i,-
DO ± o mg
CaCOs/L
Alkalinity:
29 ± 2 mg
CaCOs/L
Conductivity"
230 ± 17 uS
Temperature:
28-30 °C
pH:
6.51 ±0.01
Conductivity:
244.3 ± 0.6
uS/cm
DO:
6.25 ± 0.06 mg/L
Total hardness
(Mg2+ and Ca2+):
15.63 ±2.74
mg/L
as CaCO3
Effects on Endpoint Effect Concentration
Reproduction:
Significant decrease in number
of eggs laid at 500 ug Pb/L.
Survival:
Embryonic survival was 12% of
the number of eggs laid by the
control group at 100 ug Pb/L.
Time to hatching increased 3 fold
from the control. No embryos
survived the highest
concentration.
Survival: 24 hour LC so
LC5o increased with decrease in 85.9 ug Pb/L
mean exposure concentration 43 nour LCso'
58.5 ug Pb/L
72 hour LC5o:
45.5 ug Pb/L
96 hour LCso
35.0 ug Pb/L
Reference3
(Published
since the
2006 Pb
AQCD)
Ansaldo et al.
(2009)
Shuhaimi-
Othman et al.
(2011 a)
6-296
-------
Table 6-5 (Continued): Recent evidence for Pb effects on freshwater plants, invertebrates and vertebrates; growth,
reproduction and survival.
Species
Crayfish
(Orconectes hylas)
Concentration
Reference sites:
0.03 |jg Pb/L
Mining sites:
0.12 to
1.59|jgPb/L
Downstream
sites
0.03 to
0.04 |jg Pb/L
Exposure Method
in situ 28-day
exposure with
juvenile crayfish in
streams impacted by
Pb mining and
reference sites in
Missouri, USA.
Modifying
factors
Water quality
parameters were
measured at
each site
Temperature
23 to 26 °C
pH:
7.9 to 8.1
Conductivity:
282 to 858
DO:
6.3to8.4mg/L)
Effects on Endpoint Effect Concentration
Survival:
Crayfish survival and biomass
were significantly lower in
streams impacted by Pb mining
Metal concentrations were
negatively correlated with caged
crayfish survival.
Reference3
(Published
since the
2006 Pb
AQCD)
Allert et al.
(2009a)
Alkalinity 141 to
182 mg/Las
CaCO3
Turbidity: 0.4 to
0.6 NTU
Sulfate:
0.3to304mg/L
Other metals
were present
downstream of
mining sites
6-297
-------
Table 6-5 (Continued): Recent evidence for Pb effects on freshwater plants, invertebrates and vertebrates; growth,
reproduction and survival.
Species
Concentration
Exposure Method
Modifying
factors
Effects on Endpoint Effect Concentration
Reference3
(Published
since the
2006 Pb
AQCD)
Vertebrates
Fathead minnow
(Pimephales
promelas)
Measured:
Mean ± SEM
Tap low Pb:
28 ±
1.1 |jgPb/L
Tap high Pb:
m^ +
1 \J\J Z
4.8 |jg Pb/L
HCO3" low Pb:
31 ±
1.2 |jg Pb/L
HCO3" high Pb:
113±
4.6 |jg Pb/L
Humic low Pb:
•^n +
ou z
1 A i in Ph/l
I .H |Jlj r U/ 1_
Humic high Pb:
112±
4.5 |jg Pb/L
4-day, 10-day,
30-day, 150-day,
and 300-day
exposures in
Pb nitrate spiked
dechlorinated tap
water with static
renewal to study the
effects of DOC and
alkalinity on Pb
toxicity. Breeding
assays (21 days)
were also
performed.
Temperature:
22 ±1 °C
Tap H2O:
Hardness:
91 mg/L
pH:
8.1
+500 uM
NaHCO3:
DOC:
257 uM
Hardness:
93 mg/L
pH:
8.3
NaHCO3:
Hardness:
93 mg/L
pH:
8.3
+4 mg/L HA:
Hardness:
93 mg/L
pH:
8 .-i
.0
Humic:
Hardness:
93 mg/L
pH:
8.0
Growth:
No statistically significant growth
differences observed at any age
due to water chemistry alone;
DOC addition strongly protected
against Pb accumulation;
increased alkalinity reduced
whole body Pb burdens;. Growth
inhibited at 4 days, but recovered
by 30 days in high Pb
concentration.
Reproduction:
HCOs- reduced 21 day total
reproductive output (reduced
clutch size and number of
clutches produced); addition of
HCOs" alone actually increased
reproductive output; significantly
higher fecundity in HCOs"
treatment; Egg attachment low in
both tap water and HCO3"
treatments; HA promoted
attachment; No statistically
significant differences in egg
hatchability. HCO3"and humic
acid treatments increased
average egg mass; no effects on
hatchability in the HCO3" and
humic acid treatment;
Mager et al.
(2010)
6-298
-------
Table 6-5 (Continued): Recent evidence for Pb effects on freshwater plants, invertebrates and vertebrates; growth,
reproduction and survival.
Species
Concentration Exposure Method
Modifying
factors
Effects on Endpoint
Effect Concentration
Reference3
(Published
since the
2006 Pb
AQCD)
Fathead minnow
(Pimephales
promelas)
<0.1 -3,605
|jg Pb/L
(measured)
Pb was
quantified in fish
tissues in a
separate set of
experiments.
30 day flow through
exposure to
Pb nitrate to
determine to the
effects of Ca2+,
humic acid and pH
(6.3 and 8.3) on Pb
accumulation and
toxicity in juvenile
fathead minnows.
Exposure media
were made up
from a base-
water consisting
of 2:1 deionized
water:
dechlorinated tap
water.
Temperature: 23
°C and had
various levels of
Ca2+, humics,
and pH values:
0.5; 1;2mM
Cai+
2; 4; 8; 16 mg
humic
6.3; 8.3 pH
Growth:
No growth inhibition was
observed in any treatment. An
increase in growth was observed
in groups exposed to higher Pb
concentrations where there were
high initial mortalities.
Survival:
For most treatments, mortalities
occurred during the first 5 to 7
days of exposure. The lowest
tolerance was observed at low
pH (6.8). Addition of DOC or
CaSCU decreased Pb toxicity.
30 day LC50, EC2o,
and LOEC values*;(low
hardness basic test water
[19 mg CaCO3 /L], [Base
water])
LC 50 in ug Pb/L:
Base water: 39
0.5 mM Ca2+: 91
1.0mMCa2+: 104
2 mg humic: 255
4 mg humic: 443
8 mg humic: 832
16 mg humic: 1903
pH 6.3: 4.5
pH8.3: 13
EC 20 in ug Pb/L:
Base water: 22
0.5 mM Ca2+: 47
1.0mMCa2+51
2 mg humic: 189
4 mg humic: 319
8 mg humic: 736
16 mg humic: 1729
pH 6.3: 2.1
pH8.3: 8.7
LOEC in ug Pb/L:
Base water: 62
0.5 mM Ca2+: 40
1.0mMCa2+: 107
2 mg humic: 199
4 mg humic: 475
8 mg humic: 919
16 mg humic: 1751
pH 6.3: 6.2
pH8.3: 15
*4 day and 10 day LC50,
LC2o, and LOEC values also
reported in the paper.
Grosell et al.
(2006a)
6-299
-------
Table 6-5 (Continued): Recent evidence for Pb effects on freshwater plants, invertebrates and vertebrates; growth,
reproduction and survival.
Species
Fathead minnow
(Pimephales
promelas)
Fathead minnow
(Pimephales
promelas)
Concentration
Measured but
not reported.
Figure 1 of
Esbaugh et al.
(2011) plots the
relationship
between
dissolved and
nominal Pb
concentrations
in three waters
with low Pb
solubility.
Predicted
concentration of
major Pb
chemical
species in the
natural water
bioassays is
provided in
Table 4 of
Esbaugh et al.
(2011)
0.2 ±0.1 ugPb/
L
(control)
33 ± 4 ug/L
(chronic low)
143 ± 14ug/L
(chronic high)
(measured)
Exposure Method
Acute toxicity of Pb
to juvenile
P. promelas (<24
hours old) was
assessed in 96-hour
static renewal
exposures to two lab
generated reference
waters and seven
natural waters from
across North
America selected to
include a range of
water quality
parameters. Waters
were spiked with
varying
concentrations of Pb
as Pb nitrate.
33 to 57-day
exposures in
dechlorinated tap
water and deionized
water to Pb nitrate to
assess swimming
performance.
Modifying
factors
Temperature: 26
°C
Water chemistry
of the field-
collected waters
are reported in
Table 1 of
Esbaugh et al.
(2011) including
pH (range 5.5 to
8.5), Ca (range
24to1,934uM),
DOC (range 36
to 1,244uM)and
hardness (range
4 to 298 mg/L)
Temperature:
21 ± 1 °C
pH:
7.50 ±0.03
Total CO2:
543 ± 69 umol/L
DOC:
108± 4 umol
carbon/L
Hardness:
26 ± 3 mg/L
Effects on Endpoint Effect Concentration
Survival: 96-hour LCso range:
LC50 values ranged from 41 to 41 to 3>598 M9 Pb/L.
3,598 ug Pb/L. NOEC: range
DOC had the strongest 14 to 2>271 M9 Pb/L.
protective effect. LOEC: range
The lowest LC50 occurred in the 42 to 5>477 M9 Pb/L
pH 5.5 water.
No Pb toxicity was observed in
three alkaline natural waters.
Growth:
Fish from the chronic low
exposure were significantly
larger than control fish
(increased mass and body
length).
Reference3
(Published
since the
2006 Pb
AQCD)
Esbaugh et
al. (2011)
Mager and
Grosell
(2011)
6-300
-------
Table 6-5 (Continued): Recent evidence for Pb effects on freshwater plants, invertebrates and vertebrates; growth,
reproduction and survival.
Species
Fathead minnow
(Pimephales
promelas)
Zebrafish
(Danio rerio)
Concentration
Fish were feed
L variegatus
exposed via
water to
628 |jg Pb/L.
Mean daily
dietary dose of
0.417(0.3-0.48)
or 0.1
(0.07-0.14)
mg Pb/kg
per day
(measured)
Exposure Method
Juvenile fish fed a
live diet of the
oligochaete
L. variegatus for
30 days
contaminated with
Pb.
63 day dietary
exposure with
Pb-enriched
polychaete, Nereis
diversicolor.
Adult zebrafish were
fed a daily dose of
1% flake food (dry
wet diet/wet weight
fish), 1% brine
shrimp, and 1%
N. diversicolor
collected from either
Gannel Estuary,
Cornwall, U.K., an
estuary with legacy
Pb contamination, or
Blackwater Estuary,
Essex, U.K.,
(reference site)
Modifying
factors
Water from Lake
Superior that was
subsequently
filtered and UV-
treated.
Temperature:
25 °C
pH:
7.5-8.0,
Hardness:
45-50 mg
CaCO3,
Alkalinity :
40-45 mg
CaCOs/L
Temperature:
29 ± 1 °C
Effects on Endpoint Effect Concentration
Growth:
Not significantly affected
Survival:
Not significantly affected
Reproduction:
No impairment observed to
incidence of spawning, numbers
of eggs per breeding pair or
hatch rate of embryos compared
with pre-exposure levels. Metal
analyses revealed significant
increases in whole-body Pb
burdens of male fish fed
polychaetes from the
contaminated estuary.
Reference3
(Published
since the
2006 Pb
AQCD)
Erickson et
al. (2010)
Boyle et al.
(2010)
6-301
-------
Table 6-5 (Continued): Recent evidence for Pb effects on freshwater plants, invertebrates and vertebrates; growth,
reproduction and survival.
Species
Tilapia
(Oreochromis
niloticus)
Channel catfish
(Ictalurus
Punctatus)
Concentration
Mean
concentration of
Pb in food
pellets: 100;
400; and 800
mg Pb/kg dry
weight
(nominal)
Fish were fed
L variegatus
exposed via
water to
576 |jg Pb/L
Exposure Method
Fish obtained from
an unpolluted fish
farm in Hangzhou,
China were held in
tanks with
dechlorinated tap
water and were fed
diets with Pb nitrate
twice daily for 60
days.
Juvenile fish fed a
live diet of the
oligochaete
L. variegatus for
30 days
contaminated with
Pb.
Modifying
factors Effects on Endpoint Effect Concentration
Temperature: Growth:
25 ± 1 °C No effects on growth were found.
pH:
7.1-7.5 Survival:
DO: Exposure to Pb-contaminated
7 5-7 8 mq/L diets did not result in mortality.
Alkalinity:
109 mg CaCOs/L
Hardness:
118 mg CaCO3 /
L
Water from Lake Growth:
Superior that was Not significantly affected
subsequently Survival:
filtered and UV- Not significantly affected
treated.
Temperature: 25
°C
pH:
7.5-8.0,
Hardness:
45-50 mg
CaCOs/L
Alkalinity:
40-45 mg
CaCOs/L
Reference3
(Published
since the
2006 Pb
AQCD)
Dai et al.
(2009b)
Erickson et
al. (2010)
6-302
-------
Table 6-5 (Continued): Recent evidence for Pb effects on freshwater plants, invertebrates and vertebrates; growth,
reproduction and survival.
Species
African catfish
(Cl arias
gariepinus)
Rainbow trout
(Oncorhynchus
mykiss)
Concentration
100; 300; and
500 |jg Pb/L
(nominal)
Pb was
quantified in
tissues following
exposure.
In diet:
6.9 mg Pb per g
dry mass)
(L. variegatus
exposed via
sediments)
Exposure Method
24; 42; 90; 138; and
162 hour embryo
exposure to
Pb nitrate in
dechlorinated tap
water.
These intervals
corresponded to 30;
48,; 96; 144; and
168 hours post-
fertilization.
Juvenile fish fed a
live diet of the
oligochaete
L. variegatus for
30 days
contaminated with
Pb of varying
concentrations of
Pb nitrate
Modifying
factors
Temperature:
24 °C
pH:
8.0
Conductivity:
700 us/cm
Oxygen:
on QE;O/.
yu-yu /o
saturation
Water from Lake
Superior that was
subsequently
filtered and UV-
treated.
Temperature:
11 °C
pH:
7.5-8.0
Hardness:
45-50 mg
CaCOs/L
Alkalinity :
40-45 mg
CaCOs/L
Reference3
(Published
since the
2006 Pb
Effects on Endpoint Effect Concentration AQCD)
Growth: Osman et al.
Malformations observed in (2007b)
exposed embryos (malformed
embryos only survived shortly
after hatching), delay in
development.
Reproduction:
Concentration-dependent delay
in hatching, reduced percentage
of embryos completing egg stage
period from 75% in control to
40% in 500 ug Pb/L.
Growth: Erickson et
Not significantly affected al. (2010)
Survival:
Not significantly affected
6-303
-------
Table 6-5 (Continued): Recent evidence for Pb effects on freshwater plants, invertebrates and vertebrates; growth,
reproduction and survival.
Species
Rainbow trout
(Oncorhynchus
mykiss)
Concentration
Control Pb-free
diet of 0.06 mg
Pb/kg dry
weight, and
three different
diets of 7; 77;
and 520 mg Pb
dry weight (0.02
(control), 3.7;
39.6; and 221. 5
mg Pb/day dry
weight
calculated for
food
consumption)
Pb also
quantified in
tissues
Exposure Method
21 -day exposure to
juvenile rainbow
trout via diet
amended with
Pb nitrate. Fish were
held in aerated tanks
with dechlorinated
water
Modifying
factors
Temperature:
11-13 °C
pH
7.5-8.0
ndi ui icbo.
140 ppm
as CaCO3
Effects on Endpoint Effect Concentration
Growth:
No effects on growth rates were
observed in rainbow trout
administered a diet containing
three concentrations of Pb.
Dietary Pb was poorly absorbed.
Comparison of dietary and water-
borne exposures suggest that
toxicity does not correlate with
dietary exposure, but does
correlate with gill accumulation
from waterborne exposure.
Survival:
Not significantly affected by
dietary Pb
Reference3
(Published
since the
2006 Pb
AQCD)
Alves et al.
(2006)
6-304
-------
Table 6-5 (Continued): Recent evidence for Pb effects on freshwater plants, invertebrates and vertebrates; growth,
reproduction and survival.
Species Concentration Exposure Method
Rainbow trout ELS (early life 96-hour static
(Oncorhynchus stage) test 1: renewal acute
mykiss) 12; 24; 54; 143; toxicity test with
and swim-up stage fry (2
384 Mg Pb/L to 4 weeks post-
(measured) hatch) reared from
ELS test 2: e99s used in the
g. -|g. 37. QJ. chronic studies.
and Two 60+ day ELS
124 Mg Pb/L exposures were
(measured) conducted in a flow-
through system
using temperature
controlled water
from Little North
Fork of the South
ForkCoeur d'Alene
river in ELS 1 (69
days) and water
from the South Fork
in ELS 2 (62 days).
Modifying
factors
ELS1:
Temperature
9.8±0.6°C
pH:
6.75 ±0.4
Hardness:
19.7± 1.5 mg/L
as CaCOs
Conductivity:
45.8 ±2.2 MS
Alkalinity:
19.6 ±2.2 mg/L
as CaCOs
T~\/~\.
DO.
10.2 ± 0.7 mg/L
ELS 2:
Temperature
12.5 ± 0.9 °C
pH:
7.19 ± 0.3
Hardness:
29.4 ± 3.6 mg/L
as CaCOs
Conductivity:
69.1 ±7.4 MS
Alkalinity:
27 mg/L± 2.1 as
CaCO3
DO:
9.2 ± 0.9 mg/L
Effects on Endpoint
Growth:
In ELS 1, growth generally
decreased as concentration
increased, with fish in the highest
surviving treatment
(143 Mg Pb/L) exhibiting severely
stunted growth that was
statistically different from the
control.
In ELS 2, growth increased in the
highest treatment with a slight
reduction in length and wet
weight in intermediate
exposures.
Survival:
In ELS 1 Survival decreased as
concentration increased with
complete mortality before the
end of the test in the highest
treatment (384 Mg Pb/L).
In ELS 2, survival decreased
significantly in the highest
treatment with high survival in
intermediate exposure. In both
tests, the greatest number of
mortalities occurred around or
shortly after the swim-up stage.
Effect Concentration
96-hour LC50:
120 Mg Pb/L
ELS 1: Survival
NOEC: 24 Mg Pb/L
LOEC: 54 Mg Pb/L
MATC: 36 Mg Pb/L
ECi0: 26 Mg Pb/L
EC20 : 34 M9 Pb/L
ELS1:Weight
NOEC: 24 Mg Pb/L
I OFP- 54 un Ph/l
L\U/^W . \j^ My ' '-'' i—
MATC: 36 Mg Pb/L
ECi0: 39 Mg Pb/L
EC20: 55 Mg Pb/L
ELS1: Length
NOEC: 54 Mg Pb/L
LOEC: 143Mg Pb/L
MATC: 88 Mg Pb/L
ECi0: 64 Mg Pb/L
EC20: 98 Mg Pb/L
ELS 2: Survival
NOEC: 87 Mg Pb/L
LOEC: 125MgPb/L
MATC: 104MgPb/L
ECi0: 108Mg Pb/L
EC20: 113Mg Pb/L
ELS 2: Weight
NOEC: 37 Mg Pb/L
LOEC: 87 Mg Pb/L
MATC: 57 Mg Pb/L
ECi0: 7 Mg Pb/L
EC20: >87 Mg Pb/L
Reference3
(Published
since the
2006 Pb
AQCD)
Mebane et
al. (2008)
6-305
ELS 2: Length
NOEC: 8 Mg Pb/L
LOEC: 18MgPb/L
MATC: 12MgPb/L
ECi0: >87 Mg Pb/L
EC20: >87 ug Pb/L
-------
Table 6-5 (Continued): Recent evidence for Pb effects on freshwater plants, invertebrates and vertebrates; growth,
reproduction and survival.
Species
Southern leopard
frog
(Rana
sphenocephala)
Concentration
Sediment:
45; 75; 180;
540; 2,360;
3,940; 5,520;
7,580 mg Pb/kg
dry weight
Corresponding
sediment pore
water:
123; 227; 589;
1,833; 8,121;
13,579; 19,038;
24,427 |jg Pb/L
(measured)
Modifying
Exposure Method factors
20; 40; 61; and Mean (SD)
82-day exposures in Temperature:
Pb acetate spiked 21 6 °C
sediment collected
from wetland, static P ^
renewal twice per
week. DOC:
6.08 mg/L
Conductivity:
168 uS/L
Hardness:
7.30 mg Ca2+/L
Ammonia:
0.39 mg/L
Sediment:
Organic carbon:
8.25%
Sand: 22.4%
Silt: 38.4%
Clay: 39.1%
Effects on Endpoint
Growth:
Snout-vent length and body
mass increased through time in
all treatments; skeletal
deformations (spinal
deformations, digits truncated
and twisted, long bones curved
and truncated) increased with Pb
content and length of exposure
Survival:
Exposure to > 3,940 mg/kg
sediment Pb (13,579 ug/L) pore
water) killed all tadpoles within 5
days; tadpoles that reached
climax stage (Gosner 42) had no
difference in survival among
treatments through the
completion of metamorphosis.
Effect Concentration
40 day EC50
(for deformed spinal
columns in sediment):
1,958 mg/kg Pb
Corresponding EC5o
(for deformed spinal
columns in pore water):
6,734 ug/L
40 day EC50
(for deformed spinal
columns in sediment):
579 mg Pb/kg,
and
1,968 ug Pb/L (in pore
water).
Sediment: LC5o:
3,728 mg Pb/kg
(Corresponding to a
Pore water LCso of:
1 2,539 ug Pb/L).
Reference3
(Published
since the
2006 Pb
AQCD)
Sparling et
al. (2006)
6-306
-------
Table 6-5 (Continued): Recent evidence for Pb effects on freshwater plants, invertebrates and vertebrates; growth,
reproduction and survival.
Species
Northern leopard
frog
(Rana pipiens)
Concentration
3; 10; and
100|jgPb/L
(nominal)
(Pb was
measured in
tissues at the
end ofthe
study. Pb tissue
concentrations
ranged from 0.1
to 224.5 mg
Pb/kg dry mass
and fell within
the range of
tissue
concentrations
in wild-caught
Exposure Method
Northern leopard
frogs were exposed
to Pb as Pb nitrate in
dechlorinated water
from the embryonic
stage to
metamorphosis
(>66 days post-
hatching)
Modifying
factors
Temperature:
21 to 22 °C
PH
7.9
Hardness:
170mg/L
as CaCO3
Effects on Endpoint Effect Concentration
Growth:
Tadpole growth was significantly
slower in the early stages in
100 ug Pb/L treatment. More
than 90% of tadpoles in the
100 ug Pb/L treatment
developed lateral spinal
curvature, whereas almost all the
tadpoles in the other groups
were morphologically normal. No
significant effect of Pb exposure
was found on percentage
metamorphosis, snout-vent
length, mortality, or sex ratio.
Time to metamorphosis was
delayed in 100 ug Pb/L
treatment.
Reference3
(Published
since the
2006 Pb
AQCD)
Chen et al.
(2006b)
tadpoles).
References included are those which were published since the 2006 Pb AQCD.
6-307
-------
Mebaneetal. (2008)
Groselletal (2006b)
Mebaneetal. (2008)
Groselletal. (2006b)
Mageretal. (2011a)
Arias-Almedia and Rico-Martinez (2011)
Esbaughetal. (2011)
Vedamanikam and Shazilli (2008a)
Kitvatanachai etal. (2005)
Shuhaimi-Othman etal. (2011a)
Groselletal (2006a)
Esbaughetal. (2011)
Mebaneetal. (2008)
Mageretal. (2011a)
Wang etal. (2010f)
Baetis trica udatus (mayfly)
Oncorhynchus mykiss (Rainbow trout)
Baetis trica udatus (mayfly)
Baetis trica udatus (mayfly)
Oncorhynchus mykiss (Rainbow trout)
Lymnaeastagnalis (snail)3
Baetis trica udatus (mayfly)
Brachionus calyciflorus (rotifer)
Ceriodaphniadubia (cladoceran)
Euchlanis dilatata (rotifer)
Ceriodaphniadubia (cladoceran)
Culicoides furens (midge)
Culexquinquefasciatus (mosquito)
Macrobrachium lancesteri (prawn)
Pimephales promelas (fathead minnow)
Pimephales promelas (fathead minnow)
Oncorhynchus mykiss (Rainbow trout)
Ceriodaphniadubia (cladoceran)
Lampsilisrafinesqueana (mussel)
Lampsilissiliquoidea (mussel )ti
AJ B
• i .—
1 "-1-10 Survival
i A
* \LC
I l~'"10 Growth Inhibition
|
\ Survival
; r-f~
1 L("20 Growth Inhibition
| «
1 Reproduction
1 •
' 1 C
\ '-'-5° Survival
| i
I
1 A
|
j rc
\ 50 Reproduction
1 . , «
s Survival
I «
• EC ALC
A
A
«
A
A
A
0
9
0
A
^
A
A
A
A
A
A
A
A
A
•
0
1
10 day
early life stage test 1 (69 days)
early life stage test 2 (62 days)
10 day
10 day
early life stage test 1 (69 days)
early life stage test 2 (62 days)
30 day
10 day
number of rotifers (48 hr)
7 day in base water
7day5.0mM CaSO4
7 day 2.5 mM NaHCO3
48 hr
48hrc
A 96hrat26-20°Cb
24 hr
96 hr
30 day (low hardness water)
30 day 0.5 mM Ca2*
30 day 1.0 mM Ca2*
30 day 2 mg humic acid
30 day pH 6.3
30 day pH 8.3
96hrc
96 hr (swim upstage fry)
7dayc
48 hr (5 day old juveniles)
96 hr (5 day old juveniles)
> 96 hr (5 day old juveniles)
21 day (2 month old juveniles)
28 day (2 month old juveniles)
log ng Pb/L
"Reported as <4 |jg Pb/L
bTemperature was decreased over the course of the exposure
""Reported as a range, lowest value shown
dSurvival was assessed by foot movement.
Notes: Reference concentration "A" is the Median value of Pb reported from freshwater;
Reference concentration "B" is the Maximum value of Pb reported from freshwater.
Data in Figure 6-5 are limited to EC and LC values published since the 2006 Pb AQCD that are within one order of magnitude or
less of the maximum value of Pb (30 ug Pb/L) reported for fresh surface water (minimum 0.04 ug Pb/L, median 0.5 ug Pb/L,
95th percentile 1.1 ug Pb/L) based on a synthesis of National Ambient Water Quality Assessment data collected from 1991-2003
[(U.S. EPA. 2006b): see Table 6-21.
Detailed summaries of the individual studies referenced above, including experimental conditions, form of Pb, modifying factors and
responses at higher concentrations are summarized in Table 6-5.
Figure 6-5 Subset of concentration-response data reported in Table 6-5 for
Pb effects on growth, reproduction, and survival in some
freshwater invertebrates and fish.
6-308
-------
Table 6-6 Recent evidence for Pb effects on saltwater plants, invertebrates, and vertebrates: growth,
reproduction, and survival.
Species
Concentration
Exposure
Method
Modifying
factors
Effects on Endpoint
Effect Concentration
Reference3
(Published since
the 2006 Pb
AQCD)
Algae
Microalgae
(Tetraselmis chuii,
Rhodomonas
salina, Chaetoceros
sp., Isochrysis
gal ban a,
Nannochloropsis
gaditana)
Microalgae
(Tetraselmis
suecica)
50; 100; 250;
500; 800; 1,000;
1,600; 3,000; and
6,000 |jg Pb/L,
(nominal). Five
nominal
concentrations
were analytically
verified: 51; 225;
824; 1,704;
6,348 |jg Pb/L
(measured)
20 |jg Pb/L
(nominal)
T. suecica in this
study was then
fed to Artemia
franciscana
(mean Pb content
12 to 15 mg
Pb/kq)
i *ji ixy y.
Populations of
each microalgal
species were
exposed for 72
hours to ten
progressively
increasing
nominal
concentrations
of Pb in filtered
seawater
72-hour
exposure to
Pb nitrate in
filtered natural
seawater from
Mazatlan Bay,
Mexico. This
was the first
step in a four-
level food chain.
Temperature:
20 ± 1 °C
pH
8.0
Temperature:
28±2°C
Salinity:
34. 6 ± 1.2ppt
nl-l-
pn.
7 Q 8 9
/ . \y o .Ł.
DO saturation:
90-95%
(>7 mg/L)
Growth:
Growth inhibition (as measured
by flow cytometry) was reported
for each species. Species
cellular size, sorption capacity,
or taxonomy did not explain
differences in sensitivity to Pb.
Growth:
Mean final cell concentrations,
growth rate and total dry
biomass were significantly
reduced (40% lower than
control cultures). Effects on
primary, secondary and tertiary
consumers were observed
following Pb-exposure via
T. suecica at the base of a
simulated marine food chain.
EC50:
T. chuii:
2,640 ug Pb/L
D oa//nŁr
r\, oa////a.
900 ug Pb/L
Chaetoceros sp.:
105ugPb/L
/. galbana'.
1, 340 ug Pb/L
N. gaditana'.
740 ug Pb/L
Debelius et al.
(2009)
Soto-Jimenez
et al. (201 1b)
6-309
-------
Table 6-6 (Continued): Recent evidence for Pb effects on saltwater plants, invertebrates, and vertebrates: growth,
reproduction, and survival.
Species
Exposure
Concentration Method
Modifying
factors
Effects on Endpoint
Effect Concentration
Reference3
(Published since
the 2006 Pb
AQCD)
Invertebrates
Polychaete
(Hydroides elegans)
Polychaete
(Hydroides elegans)
91; 245; 451; 24-hour
4,443; 9,210; and exposure of
41,060 ug Pb/L fertilized eggs to
(measured) Pb chloride.
Assay was
stopped at
2 hours to
assess effects
on blastula.
48; 97; 201; 407; A series of
803; experiments
1 ,621 ug Pb/L were performed
(measured) from 20 minutes
to 4 days in
Pb chloride
using
polychaetes
collected from
seawater in
Chennai, India.
Temperature:
27 ± 1 °C
DO
(86.5%)
Salinity
(34 ± 1 ppt)
pH (8.1 ±0.1)
Carbonate
24.5 mg/L.
Temperature:
27 ± 1 °C
DO:
7-9 mg/L
Salinity
O A -L. *1 ««4-
34 ± 1 ppt
pH:
8.1 ±0.1
Carbonate:
22.5 mg/L
Reproduction:
Exposure to Pb caused a
significant decrease in the
number of embryos developing
normally to blastula after 2 to 3
hours of exposure to Pb.
Reproduction:
Fertilization rate decreased by
70% in sperm pretreated with
97 ug Pb/L for 20 minutes.
Fertilization rate of eggs
pretreated in 48 ug Pb/L
decreased to 20% of control.
Life stages of H. elegans varied
in their sensitivity to Pb.
Gametes, embryo and larvae
were more sensitive than adults
with the larval settlement period
being most sensitive to Pb
exposure.
Survival:
LCso reported for adults
EC50
Fertilization
membrane stage :
30,370 ug Pb/L
Blastula
1, 429 ug Pb/L
24 hourtrochophore
larva:
231 ug Pb/L.
EC50
Sperm toxicity:
380 ug Pb/L
Egg toxicity:
692 ug Pb/L
Embryo toxicity:
1,1 30 ug Pb/L
Blastula to trochophore:
261 ug Pb/L
Larval settlement:
100 ug Pb/L
Adult 96-hour LC5o :
946 ug/L
Gopalakrishnan
et al. (2007)
Gopalakrishnan
et al. (2008)
6-310
-------
Table 6-6 (Continued): Recent evidence for Pb effects on saltwater plants, invertebrates, and vertebrates: growth,
reproduction, and survival.
Species
Polychaete
(Ca pit ell a sp.)
Amphipod
(Elasmopus laevis)
Concentration
85; 137; 251;
392; 487; 738;
871 mg Pb/kg
(measured)
30-mg Pb/kg
(control whole-
sediment),
58 mg Pb/kg;
118mg Pb/kg;
234 mg Pb/kg;
424 mg Pb/kg
(measured)
Exposure
Method
3 and 6-day
exposure for
growth
experiments,
96-hour
exposure to
Pb chloride
spiked sediment
from
Chi-kou Estuary,
Taiwan
Multi-
generational
bioassay with
amphipods
collected in
Jamaica Bay,
New York
exposed
60+ days to
sediment spiked
with Pb acetate
in filtered
seawater. 10-
day and 28-day
bioassays were
also conducted.
Modifying
factors
Aerating
circulating
seawater
Temperature:
20±2°C
O^i II « !4-\ i-
Salinity.
•^no/.
ou /o
Temperature:
19-24°C
Salinity:
27-29 g/L
DO:
>6.57 mg/L
Effects on Endpoint
Growth:
Significant differences among
growth rates of Capitella sp. in
different levels of
Pb-contaminated sediments,
with the exception of 251 mg/kg
treatment in the 6-day
experiment. Growth rates
deceased significantly from the
control in the 3-day experiment
but changes were inconsistent
with increasing Pb
concentration.
Survival:
No effect
Reproduction:
Fecundity was reduced as
sediment Pb concentration
increased.
Onset of reproduction and
reproduction were delayed as
Pb concentration increased.
Survival:
No differences in adult survival
among the Pb concentrations
tested in 28-day and 60-day
exposures.
Reference3
(Published since
the 2006 Pb
Effect Concentration AQCD)
Growth LOAEL: Horng et al.,
85 mg Pb/kg (2009)
Fecundity and time of first Ringenary et al.
offspring was significantly (2007)
reduced with increasing
sediment concentration
above 118 mg Pb/kg.
Onset to reproduction
significantly delayed at 1 1 8
mg Pb/kg and delayed
further at higher tested
concentrations.
6-311
-------
Table 6-6 (Continued): Recent evidence for Pb effects on saltwater plants, invertebrates, and vertebrates: growth,
reproduction, and survival.
Species
Amphipod
(Melita plumulosa)
Brine shrimp
(Artemia
franciscana)
Concentration
Water-only tests
ranged from 0 to
4,000 |jg Pb/L
(analytically
verified)
Adult Sediment
Test:
500; 1,000;
2,000; 4,000 mg
Pb/kg dry weight
(analytically
verified)
Juvenile
Sediment Test:
500; 1,000; 2,000
mg Pb/kg dry
weight
(analytically
verified)
Mean Pb content
12 to 15 mg
Pb/kg from
dietary exposure
toPb.
Exposure
Method
Juveniles and
adults were
tested in 96-
hour seawater
only or 10 day
static-non-
renewal
exposure spiked
sediment
collected from
intertidal mud
flats, Woronora
River, New
South Wales,
Australia. Adults
were also tested
in 10-day
seawater only
exposures.
This was the
second step in a
four-level food
chain.
A. franciscana
feeding on
Tetraselmis
suecica cultured
in 20 ug/L Pb,
as Pb nitrate.
Modifying
factors
Temperature
21 ± 1 °C,
Salinity
30 ± 1%,
pH
r1 '
7.2-8.2,
Ammonia
(total)
<3 mg N/L
Temperature:
28 ± 2°C
Salinity:
34.6 ± 1.2ppt
nl_l •
pH.
7 Q « 9
/ . y-o.^i
DO saturation:
90-95%
(>7 mg/L)
Effects on Endpoint
Survival:
Juvenile amphipods were more
sensitive to Pb than adults in
seawater and sediment
exposures.
Growth:
A tendency toward lower
biomass yields was reported
(significant only on day of final
harvest).
Survival:
A tendency toward lower
survival was reported
(significant only on day of final
harvest).
Effect Concentration
96-hour seawater-only
Adults:
LC50 3,000 ug Pb/L
NOEC 850 ug Pb/L
LOEC 1, 680 ug Pb/L
Juveniles:
LC50 1,530ug Pb/L
NOEC 400 ug Pb/L
LOEC 600 ug Pb/L
Seawater-only 10 days:
Adults:
LC50 1,270ug Pb/L
NOEC 1 90 ug Pb/L
LOEC 390 ug Pb/L
10 days Sediment-only
Adults:
LC50 NOEC, LOEC
>3,560, mg Pb/kg
Juveniles:
LC5o 1,980 mg Pb/kg
NOEC 580 mg Pb/kg
LOEC 1, 020 mg Pb/kg
Dry biomass was 195 mg/L
in control cultures and 153
mg/L in cultures fed Pb
exposed T. suecica. Mean
cell count (individuals/L) on
days 19-23 (harvest) was
320 in control and 255 in
A. franciscana cultures fed
Pb-exposed T. suecica.
Reference3
(Published since
the 2006 Pb
AQCD)
King et al.
(2006)
Soto-Jimenez
et al. (201 1b)
6-312
-------
Table 6-6 (Continued): Recent evidence for Pb effects on saltwater plants, invertebrates, and vertebrates: growth,
reproduction, and survival.
Species
Shrimp
(Litopenaeus
vannamei)
Sea urchin
(Paracentrotus
lividus)
Mussel
(Mytilus
gailopro vin cialis)
Clam
(Meretrix meretrix)
Concentration
Pbin
exoskeleton,
hepatopancreas,
muscle, and
remaining tissues
was quantified on
days 0; 15; 28;
and 42 of the
dietary study
50 to
5,000 ug Pb/L
(nominal)
3,500; 4,500;
5,500;
6,000 ug Pb/L
(nominal)
2; 20; 197; 1,016;
7,1 58 ug Pb/L
Exposure
Method
This was the
third step in a
four-level food
chain.
L vannamei, fed
A. franciscana
(mean Pb
content 12 to 15
mg Pb/kg)
feeding on
T. suecica
cultured in
20 ug/L Pb as
Pb nitrate.
Gametes and
embryos
exposed 48 to
50 hours in
filtered seawater
to Pb nitrate
from adults
collected from
the Bay of
Naples, Italy.
24-hour static
aerated
exposure
in seawater with
Pb acetate with
mussels
collected from a
mussel farm in
Greece.
24-hour and 96-
hour toxicity test
with Pb nitrate in
Modifying
factors
Temperature:
28 ± 2°C
Salinity:
34. 6 ± 1.2ppt
pH:
7.9-8.2
DO saturation:
90-95%
(>7 mg/L)
Filtered sea
water
Temperature:
18± 1 °C
Salinity: 38%
pH: 8 ±0.2
Temperature:
25 ± 2°C
Salinity:
35%
DO:
7-8 mg/L
Temperature:
28 ± 1°C
Salinity:
Effects on Endpoint
Growth:
Tendency toward decreased
total length and weight was
reported (significant only on day
of final harvest)
Survival:
Tendency toward lower survival
was reported (significant only on
day of final harvest)
Growth:
Up to LOEC concentration,
defects observed in plutei were
mainly reduction in size (20%);
above LOEC concentration,
developmental defects were
mainly larvae affected in
skeletal or gut differentiation up
to 2,000 ug Pb/L, where arrest
of development started to
increase.
Survival:
Mortality at high Pb
concentrations
Reproduction:
Embryo development inhibited
by increasing Pb concentrations
Effect Concentration
Total mean length in the
shrimp fed the
experimental diet was 13
mm and wet weight was
7.1 g compared to mean
total length (14.8 mm) and
wet weight (8.5 g) at day
42.
Mean survival in the shrimp
fed the experimental diet
ranged from 67 to 78%
compared to control
survival (84 to 90%) at day
42.
EC 50
1, 250 ug Pb/L
NOEL
250 ug Pb/L
LOEC
500 ug Pb/L
24-hour LC5o
4,500 ug Pb/L
Embryogenesis ECso:
297 ug Pb/L
Reference3
(Published since
the 2006 Pb
AQCD)
Soto-Jimenez
et al. (201 1b)
Manzo et al.
(2010)
Vlahogianni
and
Valavanidis
(2QQ7)
Wang et al.
(2009d)
6-313
-------
Table 6-6 (Continued): Recent evidence for Pb effects on saltwater plants, invertebrates, and vertebrates: growth,
reproduction, and survival.
Species
Scallop
(Argopecten
ventricosus)
Concentration
(measured)
For growth:
10; 100; 1,000;
and 10,000 ug/L
(analytically
verified)
For survival:
280; 560; 1120;
2,250; and
5,000 ug/L
(analytically
verified)
Exposure
Method
filtered seawater
using gametes
from adults
collected from
Wenzhou, China
and held under
laboratory
conditions.
144 hour
(survival) or 30
day (growth)
exposure to
Pb nitrate (static
48 hours) with
juvenile
A. ventricosus
hatched from
laboratory
cultures held at
Universidad
Autonoma de
Baja California
Sur, Mexico.
Modifying
factors
20%
pH:
7.8
Temperature:
20°C
Salinity:
36 ± 1%
DO:
>4 mg/L
Effects on Endpoint
Growth:
Significant concentration-
dependent growth inhibition in
larvae. Larval metamorphosis
rate decreased, no adverse
effect on larval settlement at
20.4 ug Pb/L
Survival:
Significant concentration-
dependent survival inhibition of
embryos over time
Growth:
Juvenile growth rates and
weight were significantly
reduced at high concentrations
ofPb
Survival:
Juvenile mortality was
significantly different than
control at 96 hour LCso
Reference3
(Published since
the 2006 Pb
Effect Concentration AQCD)
Growth:
EC50:
197ugPb/L
Metamorphosis:
>7,160ugPb/L
48-hour LC5o:
>7, 160 ug Pb/L
96-hour LC5o:
353 ug Pb/L
ECso for growth: Sobrino-
4, 21 Dug Pb/L Figueroa et al.
(2007)
72 -hour LCs0:
4,690 ug Pb/L
96-hour LCso:
830 ug Pb/L
120-hour LC50:
680 ug Pb/L
144-hour LC50:
680 ug Pb/L
6-314
-------
Table 6-6 (Continued): Recent evidence for Pb effects on saltwater plants, invertebrates, and vertebrates: growth,
reproduction, and survival.
Species
Bivalve
(Tellina deltoid alls)
Exposure
Concentration Method
1,000 mg Pb/kg 10 day direct
(analytically exposure to
verified) Pb nitrate
spiked sediment
collected from
Woronora River,
New South
Wales,
Australia. Adults
used in the test
were collected
from Lane Cove
River, Sydney,
Australia and
held in filtered
seawater.
Modifying
factors
Temperature:
21 ± 1 °C
Salinity
30 ± 1%,
PH
7.2-8.2
Ammonia
(total)
<3 mg N/L
Reference3
(Published since
the 2006 Pb
Effects on Endpoint Effect Concentration AQCD)
Survival: 10 Day NOEC King et al.
All individuals survived. Porewater: 15 ug Pb/L (2010)
dissolved Pb
Sediment: 1,000 mg Pb/kg
Vertebrates
Toadfish
(Tetractenos glaber)
Measured but not Field-collected
reported fish-Pb sampled
from sediments
collected in two
reference and
two metal
contaminated
estuaries near
Sydney,
Australia
Temperature :
15to16°C
Salinity:
29 to 31%,
pH:
8/1 tn ft fi
.H LU O.U,
amongst
estuaries.
Sediment pH:
7.0-7.5
Organic
matter:
1.5-2.1%
Reproduction: N/A Alquezaret al.
Decreased oocyte diameter and (2006)
density in the toadfish were
associated with elevated levels
of Pb in the gonads of field
collected fish; authors state that
this is suggestive of a reduction
in egg size.
6-315
-------
Table 6-6 (Continued): Recent evidence for Pb effects on saltwater plants, invertebrates, and vertebrates: growth,
reproduction, and survival.
Species
Grunt fish
(Haemulon
scudderi)
Concentration
Mean total Pb
body burden
increased from
0.55 to 3.32 mg
Pb/kg during the
feeding
experiment.
Exposure
Method
42-day dietary
exposure from
simulated
marine food
chain-shrimp,
L vannamei, fed
A franciscana
(mean Pb
content
12-1 5 mg Pb/kg)
feeding on
T. suecica
cultured in
20 ug/L Pb as
Pb nitrate.
Modifying
factors
Temperature:
28±2°C
Salinity:
34.6 ± 1.2ppt
n|_|. 7 Q Q 0
[jn. / . y-o.^i
DO saturation:
90-95%
(>7 mg/L)
Effects on Endpoint
Survival:
No significant differences
observed in intermediate and
final length, mean wet weight, or
Fulton's condition factor: Final
survival significantly lower;
mean total Pb body burden
increased.
Effect Concentration
Mean survival in the fish
fed the experimental diet
ranged from 65 to 75%
compared to control diet
survival (88 to 91%).
Reference3
(Published since
the 2006 Pb
AQCD)
Soto-Jimenez
et al. (201 1b)
References included are those which were published since the 2006 Pb AQCD.
6-316
-------
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