United States
Environmental Protection
kl m m Agency
Febraary 2012
EP A/600/R-10/075B
Integrated Science Assessment
for Lead
National Center for Environmental Assessment-RTP Division
Office of Research and Development
U.S. Environmental Protection Agency
Research Triangle Park, NC

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DISCLAIMER
This document is the second external review draft for review purposes only and does not constitute
U.S. Environmental Protection Agency policy. Mention of trade names or commercial products does
not constitute endorsement or recommendation for use.
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CONTENTS
LEAD PROJECT TEAM	XXII
CLEAN AIR SCIENTIFIC ADVISORY COMMITTEE LEAD NAAQS REVIEW PANEL	XXIX
ACRONYMS AND ABBREVIATIONS	XXXI
Table I Aspects to aid in judging causality	Iv
Table II Weight of evidence for causal determination	Ivii
CHAPTER 1 EXECUTIVE SUMMARY	1-1
1.1	Introduction	 1-1
1.2	Lead Sources, Fate and Transport in the Environment, Human Exposure and Dose	 1-2
Figure 1-1 Conceptual model of multimedia Pb exposure.	 1-4
1.3	Integrative Overview of Health and Ecological Effects	 1-5
Figure 1-2 Schematic representation of the relationships between the various
MOAs by which Pb exerts its effects.	 1-6
1.3.1	Health Effects of Lead	 1-6
Table 1-1 Summary of causal determinations between exposure to Pb and
health outcomes	 1-7
1.3.2	Ecological Effects of Lead	 1-11
Table 1-2 Summary of Pb causal determinations for plants, invertebrates and
vertebrates	 1-12
1.4	Policy Relevant Considerations	 1-15
1.5	Summary	 1-18
CHAPTER 2 INTEGRATIVE SUMMARY	2-1
2.1	ISA Development and Scope	 2-1
2.2	Ambient Lead: Source to Concentration	 2-5
2.2.1	Sources, Fate and Transport of Ambient Lead	2-5
2.2.2	Monitoring and Concentrations of Ambient Air Lead	2-6
2.2.3	Ambient Lead Concentrations in Non-Air Media and Biota	2-9
Table 2-1 Ambient Pb Concentrations in Non-Air Media and Biota Considered
for Ecological Assessment	 2-10
2.3	Exposure to Ambient Lead	 2-10
2.4	Toxicokinetics	 2-11
2.5	Lead Biomarkers	 2-13
2.6	Health Effects	 2-14
Table 2-2 Summary of causal determinationsa between exposure to Pb and
health outcomes	 2-15
2.6.1	Nervous System Effects	2-15
2.6.2	Cardiovascular Effects	2-18
2.6.3	Renal Effects	2-19
2.6.4	Immune System Effects	2-21
2.6.5	Heme Synthesis and Red Blood Cell Function	2-23
2.6.6	Reproductive and Developmental Effects	2-24
2.6.7	Cancer	2-26
2.7	Ecological Effects of Lead	 2-27
2.7.1	Summary of Effects on Terrestrial Ecosystems	2-27
2.7.2	Summary of Effects on Aquatic Ecosystems	2-31
2.7.3	Determinations of Causality for Effects on Ecosystems 	2-36
Table 2-3 Summary of Pb causal determinations for plants, invertebrates and
vertebrates	 2-36
2.7.3.1	Effects on Physiological Stress	2-36
2.7.3.2	Hematological Effects	2-37
2.7.3.3	Neurobehavioral Effects	2-38
2.7.3.4	Effects on Development and Reproduction	2-39
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2.7.3.5	Effects on Growth	2-42
2.7.3.6	Effects on Survival 	2-43
2.7.3.7	Community and Ecosystem Effects 	2-45
2.8	Integration of Health and Ecological Effects	 2-47
Table 2-4 Summary of Causal Determinationsa for Health and Ecological
Effects	 2-48
2.8.1 Modes of Action Relevant to Downstream Health and Ecological Effects	2-49
Table 2-5 MOAs, their related health effects, and information on concentrations
eliciting the MOAs	 2-50
2.9	Policy Relevant Considerations	 2-54
2.9.1	Public Health Significance	2-54
Figure 2-1 The effect of a small shift in population mean on the proportion of
individuals in the population diagnosed with clinical disease (i.e., the
proportion to the right of the "Critical Line.")	 2-55
Table 2-6 Illustrative examples contrasting the effect sizes observed in
epidemiologic studies to effect sizes considered significant in a
clinical settinga	 2-58
2.9.2	Air Lead-to-Blood Lead Relationships	2-59
Table 2-7 Summary of Estimated Slopes for Blood Pb to Air Pb Relationships
in Humans	 2-61
2.9.3	Ecological Effects and Corresponding Lead Concentrations	2-62
2.9.4	Concentration-Response Functions for Health Effects	2-63
Figure 2-2 Comparison of associations between blood Pb and cognitive
function among various blood Pb strata.	 2-64
2.9.5	Pb Exposure and Nervous System Effects in Children	2-69
2.9.6	Populations Potentially At-Risk for Health Effects	2-71
Table 2-8 Factors evaluated that may determine populations potentially at
increased risk from lead	 2-72
2.9.6.1	Age and/or Lifestage	2-73
2.9.6.2	Sex 2-75
2.9.6.3	Genes	2-75
2.9.6.4	Pre-existing Conditions	2-75
2.9.6.5	Race and Ethnicity	2-76
2.9.6.6	Socioeconomic Status	2-76
2.9.6.7	Proximity to Pb Sources and Residential Factors	2-77
2.9.6.8	Lifestyle Factors	2-77
2.9.6.9	Nutritional Factors	2-77
2.9.6.10	Stress and Cognitive Reserve	2-78
2.9.6.11	Co-exposure of Lead with Metals and Chemicals	2-78
2.9.6.12	Summary of At-Risk Populations 	2-79
Table 2-9 Summary of evidence for factors that potentially increase the risk of
lead-related health effects	 2-79
2.10	Summary	 2-80
Table 2-10 Summary of evidence from epidemiologic, animal toxicological and
ecological studies on the effects associated with exposure to Pb	 2-80
CHAPTER 3 AMBIENT LEAD: SOURCE TO CONCENTRATION	3-1
3.1	Introduction	 3-1
3.2	Sources of Atmospheric Lead	 3-1
3.2.1	National Emissions Inventory	 3-2
Figure 3-1 Trends in Pb emissions (thousand tons) from stationary and mobile
sources in the U.S., 1970-2008.	 3-3
Figure 3-2 Trends in Pb emissions (thousand tons) from stationary and mobile
sources in the U.S., 1990-2008.	 3-4
Figure 3-3 Nationwide stationary and mobile source Pb emissions (tons) in the
U.S. by source sector in 2008. 	 3-5
Figure 3-4 County-level Pb emissions (tons) in the U. S. in 2008.	 3-6
3.2.2	Anthropogenic Sources	 3-6
Table 3-1 Pb compounds observed in the environment	 3-7
3.2.2.1 Lead Emissions from Piston-engine Aircraft Operating on Leaded Aviation Gasoline and
Other Non-Road Sources	 3-7
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3.2.2.2	Emissions from Metals Processing and Mining	3-8
3.2.2.3	Fossil Fuel Combustion	3-10
3.2.2.4	Waste Incineration 	3-11
3.2.2.5	Wood Burning	3-12
3.2.2.6	Roadway-Related Sources	3-13
3.2.2.7	Deposited Lead	3-15
Figure 3-5 Total U. S. Pb additives in on-road gasoline used in on-road vehicles,
1927-1995. 	 3-18
Figure 3-6 Estimated Pb aerosol inputs from on-road gasoline into 90 U. S.
urbanized areas (UAs), from 1950 through 1982.	 3-19
3.3	Fate and Transport of Lead	 3-19
Figure 3-7 Fate of atmospheric lead.	 3-20
3.3.1	Air 3-20
3.3.1.1	Transport	3-21
3.3.1.2	Deposition	 3-22
3.3.1.3	Resuspension of Lead from Surface Soil to Air after Deposition	3-24
Figure 3-8 Scales of turbulence within an urban environment.	 3-26
3.3.2	Water	 3-29
3.3.2.1	Lead Transport in Water and Sediment	3-30
3.3.2.2	Deposition of Lead within Bodies of Water and in Sediment	3-31
Table 3-2 Surface sediment Pb concentrations for various continental shelves	 3-33
3.3.2.3	Flux of Lead from Sediments	3-33
3.3.2.4	Lead in Runoff	3-35
3.3.3	Soil 3-43
3.3.3.1	Deposition of Lead onto Soil from Air	3-44
3.3.3.2	Sequestration of Lead from Water to Soil	3-45
3.3.3.3	Movement of Lead within the Soil Column	3-48
Figure 3-9 Schematic model summarizing the estimated flux of Pb within a
typical podzol profile from northern Sweden using data from
Klaminder et al. (2006a).	 3-50
Figure 3-10 Eh-pH diagram for Pb in shooting range soils, Jefferson National
Forest, VA.	 3-54
3.4	Monitoring of Ambient Lead	 3-55
3.4.1	Ambient Measurement Techniques	3-55
3.4.1.1	Sample Collection	3-55
Figure 3-11 Comparison of particle collection efficiency among different TSP
sampler types (Modified Andersen Sampler, Hi-volume Sampler (for
different incident wind direction (45°, 0 ), Prototype Dichotomous
Sampler, and Original Andersen Sampler).	 3-56
Table 3-3 Airborne Pb sampling methods	 3-58
3.4.1.2	Sample Analysis: Federal Reference and Federal Equivalence Methods	3-59
3.4.1.3	Other Analysis Methods for Total Lead	3-61
3.4.1.4	Sequential Extraction 	3-63
3.4.1.5	Speciation Techniques	3-64
3.4.1.6	Continuous Lead Monitoring	3-67
3.4.2	Ambient Network Design	3-68
3.4.2.1	NAAQS Monitoring Network	3-68
Figure 3-12 Map of Monitoring Sites in Current Pb NAAQS Monitoring Network.	 3-71
Figure 3-13 Fifteen U.S. locations where a study is currently being performed on
airport Pb emissions.	 3-72
Table 3-4 List of 15 airports included in the airport study	 3-73
3.4.2.2	Other Lead Monitoring Networks 	3-73
Figure 3-14 Pb-PM2.5 monitoring sites for CSN and IMPROVE networks.	 3-75
Figure 3-15 Pb-PM10 monitoring sites for NATTS network.	 3-75
3.5	Ambient Air Lead Concentrations	 3-76
3.5.1 Spatial Distribution of Air Lead	3-77
3.5.1.1 Variability across the U.S.	3-77
Table 3-5 Summary data for source-oriented Pb monitors across the U. S.,
2008-2010	 3-79
Table 3-6 Summary data for sites at which source-oriented statistics for one-
month and three-month annual site max are in the upper 90th
percentile, 2008-2010	 3-79
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Figure 3-16 Highest county-level source-oriented Pb-TSP concentrations
(/jg/m3), maximum 3-month average, 2008-2010.	 3-80
Table 3-7 Summary data for non-source-oriented Pb monitors across the U.S.,
2008-2010	 3-82
Table 3-8 Summary data for sites at which non-source-oriented statistics for
one-month and three-month annual site max are in the upper 90th
percentile, 2008-2010	 3-82
Figure 3-17 Highest county-level non-source-oriented Pb-TSP concentrations
(jjg/m3), maximum 3-month average, 2008-2010.		 3-83
Figure 3-18 Highest county-level Pb-PMio concentrations (fjg/mJ), maximum
3-month average, 2007-2009.		 3-84
Figure 3-19 Highest county-level Pb-PM2.5 concentrations (fjg/mJ), maximum
3-month average, 2007-2009.	 3-85
3.5.1.2 Intra-urban Variability	3-85
Table 3-9 Sample of U.S. near road Pb TSP monitors	 3-87
Figure 3-20 Time series of monthly average Pb-TSP concentration at five near-
road monitors.	 3-88
3.5.2	Temporal Variability	3-90
3.5.2.1	Multi-year Trends		3-90
Figure 3-21 National trends in Pb concentration (fjg/mJ), source-oriented FRM
monitors, 1990-2010.		 3-91
Figure 3-22 National trends in Pb concentration (fjg/m3), non-source-oriented
FRM monitors, 1990-2010.	 3-92
3.5.2.2	Seasonal Variations 		3-93
Figure 3-23 Monthly source-oriented Pb-TSP average (jjg/m3) over 12 months of
the year, 2008-2010.		 3-94
Figure 3-24 Monthly non-source-oriented lead-TSP average (fjg/mJ) over 12
months of the year, 2008-2010. 	 3-95
Figure 3-25 Monthly lead-PM-io average (jjg/m3) over 12 months of the year,
2007-2009. 		 3-95
Figure 3-26 Monthly iead-PM2 5 average (fjg/mJ) over 12 months of the year,
2007-2009. 	 3-96
3.5.3	Size Distribution of Lead-Bearing PM 	3-96
3.5.3.1	Airborne Pb Near Metals Industries	3-97
3.5.3.2	Airborne Pb Near Roadways	3-99
Figure 3-27 Comparison of urban background and near-road size fractions of
lead-bearing PM.	 3-101
3.5.3.3	Airborne Pb at Other Urban and Rural Sites	3-102
3.5.4	Lead Concentrations in a Multipollutant Context	3-103
Figure 3-28 Correlations of monitored Pb-TSP concentration with copollutant
concentrations, 2007-2008.	 3-104
Figure 3-29 Correlations of monitored Pb-TSP concentration with co-pollutant
concentrations, 2009. 	 3-104
Table 3-10 Correlations between Pb and copollutants, measured in TSP, PM10,
and PM2.5	 3-107
Figure 3-30 Correlations of monitored lead-PM2 5 concentration with copollutant
concentrations, 2007-2009.	 3-108
3.5.5	Background Lead Concentrations	3-108
3.6 Ambient Lead Concentrations in Non-Air Media and Biota	 3-111
3.6.1	Soils 3-112
Figure 3-31 Map of median Pb content in soil in New Orleans.	 3-115
3.6.2	Sediments 	3-118
Figure 3-32 WACAP data for Pb concentration in sediment at eight National
Parks and/or Preserves.	 3-120
Table 3-11 Sediment concentrations in various cities, prior to 2005	 3-122
Figure 3-33 Sediment core data (1992-1994) for the lakes and reservoirs along
the Apalachicola, Chattahoochee, and Flint River Basin (ACF),
which feeds from north of the Atlanta, GA metropolitan area into the
Gulf of Mexico at Apalachicola Bay in the Florida panhandle.	 3-122
Figure 3-34 Sediment core data (1975-1995) for the lakes and reservoirs along
the Apalachicola, Chattahoochee, and Flint River Basin (ACF),
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3.6.3	Rain 3-124
Figure 3-35
3.6.4	Snowpack	
which feeds from north of the Atlanta, GA metropolitan area into the
Gulf of Mexico at Apalachicola Bay in the Florida panhandle.	
Trends in Pb concentration in precipitation from various sites in
Norway over the period 1980-2005.	
Figure 3-36
3.6.5 Natural Waters
Box plots illustrating Pb concentration in snow melt at nine National
Parks and Preserves.
Figure 3-37 Boxplots of Pb concentration in surface waters measured at five
National Parks and Preserves.	
Table 3-12
Pb concentrations from stream food webs in mining-disturbed areas
of Missouri and the western United States
3.6.6 Vegetation
3.6.7
3.6.8
Figure 3-38
Figure 3-39
Aquatic Bivalves_
Boxplots of Pb concentration in lichen measured at seven National
Parks and Preserves.
Figure 3-41 Boxplots of Pb concentration in moose meat and liver measured at
Denali National Park and Preserve. 	
3.7 Summary and Conculusions	
3.7.1 Sources of Atmospheric Lead	
3.7.2
3.7.3
3.7.4
3.7.5
Fate and Transport of Lead
Ambient Lead Monitoring	
Ambient Air Lead Concentrations
Ambient Lead Concentrations in Non-Air Media and Biota
3.8 Chapter 3 Appendix (Supplemental Material)	
3.8.1 Variability across the U.S. 	
Table
3-13
Table
3-14
Table
3-15
Table
3-16
Table
3-17
Table
3-18
Table
3-19
Table
3-20
Table
3-21
Table
3-22
Table
3-23
Distribution of 1-month average Pb-TSP concentrations (fjg/m )
nationwide, non-source-oriented monitors, 2008-2010	
Distribution of 3-month moving average Pb-TSP concentrations
(fjg/m3) nationwide, source-oriented monitors, 2008-2010	
Distribution of 3-month moving average Pb-TSP concentrations
(jjg/m3) nationwide, non-source-oriented monitors, 2008-2010 _
Distribution of annual 1-month site maxima TSP Pb concentrations
(jjg/m3) nationwide, source-oriented monitors, 2008-2010	
Distribution of annual 1-month site maxima TSP Pb concentrations
(jjg/m3) nationwide, non-source-oriented monitors, 2008-2010	
Distribution of annual 3-month site maxima Pb-TSP concentrations
(fjg/m3) nationwide, source-oriented monitors, 2008-2010	
Distribution of annual 3-month site maxima Pb-TSP concentrations
(jjg/m3) nationwide, non-source-oriented monitors, 2008-2010	
One-month average Pb-TSP for individual county concentrations
nationwide (fjg/m ), source-oriented monitors, 2008-2010	
One-month average Pb-TSP for individual county concentrations
nationwide (fjg/m ), non-source-oriented monitors, 2008-2010	
Three-month moving average Pb-TSP for individual county
concentrations (fjg/m3) nationwide, source-oriented monitors, 2008-
2010	
Table 3-24 Three-month moving average Pb-TSP for individual county
concentrations (jjg/m3) nationwide, non-source-oriented monitors,
2008-2010	
3.8.2 Intra-urban Variability	
Figure 3-42 Pb TSP monitor and source locations within Los Angeles County,
CA (06-037), 2007-2009.	
3-123
3-125
3-126
3-128
3-129
3-130
3-132
3-132
3-133
Trends in regional pollution near a copper smelter in Canada and Pb
concentrations at the boundary of heartwood trees within roughly
75 km of the smelter.
Vertebrate Populations	
Figure 3-40 Boxplots of Pb concentration in fish fillet and liver measured at eight
National Parks and Preserves.
Distribution of 1-month average Pb-TSP concentrations (fjg/m )
nationwide, source-oriented monitors, 2008-2010	
3-135
3-135
3-136
3-137
3-138
3-138
3-138
3-139
3-140
3-141
3-142
3-143
3-143
3-143
3-145
3-147
3-149
3-150
3-151
3-152
3-152
3-154
3-155
3-156
3-158
3-159
3-162
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Figure 3-43
Figure 3-44
Table 3-25
Wind roses for Los Angeles County, CA, from meteorological data at
the Los Angeles International Airport, 1961-1990.	i^__
Box plots of annual and seasonal Pb TSP concentrations (fjg/mJ)
from source-oriented and non-source-oriented monitors within Los
Angeles County, CA (06-037), 2007-2009.	
Comparisons between Pb TSP concentrations from source-oriented
and non-source-oriented monitors within Los Angeles County, CA
(06-037), 2007-2009	
Figure 3-45 Pb TSP monitor locations within Hillsborough and Pinellas Counties,
FL (12-057 and 12-103), 2007-2009. 	
Figure 3-46 Wind roses for Hillsborough/Pinellas Counties, FL, obtained from
meteorological data at Tampa International Airport, 1961-1990.
Figure 3-47 Box plots of annual and seasonal Pb TSP concentrations (fjg/mT)
from source-oriented and non-source-oriented monitors within
Hillsborough and Pinellas Counties, FL (12-057 and 12-103),
2007-2009.
Table 3-26
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 3-48 Pb TSP Monitor locations within Cook County, IL (17-031),
2007-2009. 	
Figure 3-49
Wind roses for Cook County, IL, obtained from meteorological data
at O'Hare International Airport, 1961-1990. 	
Figure 3-50 Box plots of annual and seasonal Pb TSP concentrations (fjg/m )
from source-oriented and non-source-oriented monitors within Cook
County, IL (17-031), 2007-2009.	
Table 3-27
Figure 3-51
Figure 3-52
Figure 3-53
Table 3-28
Figure 3-54
Figure 3-55
Figure 3-56
Table 3-29
Correlations between Pb TSP concentrations from source-oriented
and non-source-oriented monitors within Cook County, IL (17-031),
2007-2009	
Pb TSP Monitor locations within Jefferson County, MO (29-099),
2007-2009. 	
Wind roses for Jefferson County, MO, obtained from meteorological
data at St. Louis/Lambert International Airport, 1961-1990.
Box plots of annual and seasonal Pb TSP concentrations (fjg/m )
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	
Pb TSP Monitor locations within Cuyahoga County, OH (39-035),
2007-2009. 	
Wind roses for Cuyahoga County, OH, obtained from meteorological
data at Cleveland/Hopkins International Airport, 1961-90.	
Box plots of annual and seasonal Pb TSP concentrations (fjg/mJ)
from source-oriented and non-source-oriented monitors within
Cuyahoga County, OH (39-035), 2007-2009.	
Correlations between Pb TSP concentrations from source-oriented
and non-source-oriented monitors within Cuyahoga County, OH
(39-035), 2007-2009	
Figure 3-57 Pb TSP Monitor locations within Sullivan County, TN (47-163),
2008-20102007-2009.
3.8.3
Figure 3-58 Wind roses for Sullivan County, TN, obtained from meteorological
data at Bristoi/Tri City Airport, 1961-90. 	
Figure 3-59 Box plots of annual and seasonal Pb TSP concentrations (jjg/m3)
from source-oriented monitors within Sullivan County, TN (47-163),
2007-2009. 	
Table 3-30 Correlations between Pb TSP concentrations from source-oriented
monitors within Sullivan County, TN (47-163), 2007-2009	
Lead Concentration in a Multipollutant Context	
Figure 3-60 Seasonal correlations of monitored Pb-TSP concentration with
copollutant concentrations, 2007-2008. 	
3-163
3-164
3-165
3-167
3-168
3-169
3-170
3-173
3-174
3-175
3-176
3-178
3-179
3-180
3-181
3-184
3-185
3-186
3-187
3-189
3-190
3-191
3-192
3-193
3-193
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Figure 3-61 Seasonal correlations of monitored Pb-TSP concentration with
copollutant concentrations, 2007-2008. 	 3-194
Figure 3-62 Seasonal correlations of monitored Pb-TSP concentration with
copollutant concentrations, 2009.	 3-195
Figure 3-63 Seasonal correlations of monitored Pb-TSP concentration with
copollutant concentrations, 2009.	 3-196
Figure 3-64 Seasonal correlations of monitored Pb-PM2.5 concentration with
copollutant concentrations, 2007-2009. 	 3-197
Figure 3-65 Seasonal correlations of monitored Pb-PM2.5 concentration with
copollutant concentrations, 2007-2009. 	 3-198
Table 3-31 Copollutant exposures for various trace metal studies	 3-199
CHAPTER 4 EXPOSURE, TOXICOKINETICS, AND BIOMARKERS	4-1
4.1	Exposure Assessment	 4-1
4.1.1	Pathways for Lead Exposure	4-1
Figure 4-1 Conceptual model of multimedia Pb exposure.	 4-3
Table 4-1 Estimates of Pb measurements for EPA Region 5 from the NHEXAS
study	 4-5
4.1.2	Environmental Exposure Assessment Methodologies	4-7
4.1.3	Exposure Studies	4-9
4.1.3.1	Airborne Lead Exposure	4-9
Table 4-2 Estimates of fixed effects multivariate modeling of Pb levels
measured during the NHEXAS-MD study	 4-10
Table 4-3 Comparison of personal, indoor, and outdoor Pb-PM measurements
from several studies	 4-13
4.1.3.2	Exposure to Lead in Soil and Dust	4-13
Table 4-4 Measurements of indoor dust Pb concentration from 2006-2011
studies	 4-14
4.1.3.3	Dietary Lead Exposure	4-18
Figure 4-2 Market basket survey results for Pb concentration in foods.	 4-19
Table 4-5 Pb bioaccumulation data for various plants. Bioaccumulation is
expressed as percent of Pb concentration in the plant to the Pb
concentration in the soil	 4-25
4.1.3.4	Occupational	4-27
4.1.3.5	Exposure to Lead from Consumer Products	4-28
Table 4-6 Pb content in various consumer products.	 4-28
4.2	Kinetics	 4-30
4.2.1	Absorption 	4-30
4.2.1.1	Inhalation	4-31
4.2.1.2	Ingestion	4-33
Figure 4-3 Estimated relative bioavailability (RBA, compared to Pb-acetate) of
ingested Pb in mineral groups.	 4-37
4.2.2	Distribution	4-39
4.2.2.1	Blood	 4-39
Figure 4-4 Plot of blood and plasma Pb concentrations measured in adults and
children.	 4-41
Figure 4-5 Relationship between Pb intake and blood Pb concentration in
infants (n = 105, age 13 weeks, formula-fed).	 4-42
Figure 4-6 Simulation of quasi-steady state blood and plasma Pb
concentrations in a child (age 4 years) associated with varying Pb
ingestion rates. 	 4-43
4.2.2.2	Bone 4-44
4.2.2.3	Soft Tissues	4-45
4.2.2.4	Fetus	4-46
4.2.2.5	Organic Lead	4-46
4.2.3	Elimination	4-47
4.3	Lead Biomarkers	 4-48
4.3.1	Bone Lead Measurements	4-51
4.3.2	Blood Lead Measurements	4-53
Figure 4-7 Simulation of temporal relationships between Pb exposure and blood
Pb concentration in children.	 4-56
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4.3.3	Urine Lead Measurements	4-57
Figure 4-8 Simulation of relationship between urinary Pb excretion and body
burden in adults.	 4-58
4.3.4	Lead in Other Biomarkers 	4-59
4.3.4.1	Teeth	 4-59
4.3.4.2	Hair 4-60
4.3.4.3	Saliva	 4-60
4.3.4.4	Serum 8-ALA and ALAD	4-61
4.3.5	Relationship between Lead in Blood and Lead in Bone 	4-61
4.3.5.1	Children	4-63
Figure 4-9 Simulation of relationship between blood Pb concentration and body
burden in children, with a constant Pb intake from age 2 to 5.	 4-65
Figure 4-10 Simulation of relationship between time-integrated blood Pb
concentration and cumulative Pb absorption in children.	 4-66
4.3.5.2	Adults 	 4-66
Figure 4-11 Simulation of relationship between blood Pb concentration, bone Pb
and body burden in adults.	 4-68
4.3.6	Relationship between Lead in Blood and Lead in Soft Tissues	4-72
Figure 4-12 Simulation of blood and soft tissue (including brain) Pb in children
and adults who experience a period of increased Pb intake.	 4-73
Figure 4-13 Simulation of blood and brain Pb in children and adults who
experience a period of increased Pb intake.	 4-75
Figure 4-14 Relationship between Pb in urine and Pb in blood.	 4-77
4.4	Observational Studies of Lead Biomarker Levels	 4-78
4.4.1	Lead in Blood	4-78
Figure 4-15 Temporal trend in blood Pb concentration.	 4-79
Table 4-7 Blood Pb concentrations in the U.S. population.	 4-80
Figure 4-16 Box plots of blood Pb levels among U.S. children (1-5 years old)
from the NHANES survey, 1988-2008.	 4-82
Figure 4-17 Blood Pb cohort means versus year of exam.	 4-82
Figure 4-18 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).	 4-83
Figure 4-19 Trends in 206Pb/04Pb isotope ratio in blood Pb (a) and trends in
blood Pb levels (b) among Australian study populations during the
period 1990-2000.	 4-87
4.4.2	Lead in Bone	4-87
Table 4-8 Epidemiologic studies that provide bone Pb measurements for non-
occupationally exposed populations	 4-88
Table 4-9 Epidemiologic studies that provide bone Pb measurements for
occupationally exposed populations	 4-94
4.4.3	Lead in Urine	4-96
Table 4-10 Urine Pb concentrations in the U.S. population	 4-97
4.4.4	Lead in Teeth	4-98
Figure 4-20 Comparison of relative temporal changes in tooth enamel Pb
concentration.	 4-99
4.5	Empirical Models of Lead Exposure-Blood Lead Relationships	 4-100
4.5.1 Air Lead-Blood Lead Relationships	4-100
Table 4-11 Summary of estimated slopes for blood Pb to air Pb relationships
in humans	 4-102
Figure 4-21 Predicted relationship between air Pb and blood Pb based on a meta
analysis of 18 studies.		 4-103
Figure 4-22 Blood Pb - air Pb slope (/jg/dL per fjg/mJ) predicted from various
epidemiologic studies (links available in Table 4-11).	 4-104
4.5.1.1	Children	4-104
Table 4-12 Environmental Pb levels and blood Pb levels in children in Trail,
British Columbia	 4-105
Table 4-13 U.S. gasoline Pb consumption and air Pb levels	 4-108
Table 4-14 Air Pb levels and blood Pb levels in children in Mumbai, India	 4-109
Figure 4-23 Predicted relationship between air Pb and blood Pb based on data
from Chicago, ILin children age 0-5 y (1974-1988).	 4-110
4.5.1.2	Adults 	4-110
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Table 4-15 Significant predictors of blood Pb concentration in bridge painters	 4-111
4.5.2 Environmental Lead-Blood Lead Relationships	4-112
Table 4-16 Linear model relating environmental Pb exposure and blood Pb
concentration in children	 4-114
Figure 4-24 Predicted relationship between soil Pb concentration and blood Pb
concentration in children based on data collected in the New	 4-115
Table 4-17 General linear model relating blood Pb concentration in children and
environmental Pb levels—Bunker Hill Superfund Site	 4-117
4.6	Biokinetic Models of Lead Exposure-Blood Lead Relationships	 4-117
4.7	Summary and Conclusions	 4-119
4.7.1	Exposure	4-119
4.7.2	Toxicokinetics	4-120
4.7.3	Lead Biomarkers	4-121
4.7.4	Air Lead-Blood Lead Relationships	4-123
CHAPTER 5 INTEGRATED HEALTH EFFECTS OF LEAD EXPOSURE	5-1
5.1	Introduction	 5-7
5.2	Modes of Action	 5-2
5.2.1	Introduction	 5-2
Figure 5-1 Schematic representation of the relationships between the various
MOAs by which Pb exerts its health effects.	 5-3
5.2.2	Altered Ion Status 	 5-3
5.2.2.1	Disruption of Caz+ Homeostasis	5-3
5.2.2.2	Disruption of Ion Transport Mechanisms	5-5
5.2.2.3	Displacement of Metal Ions and Perturbed Protein Function	5-9
Table 5-1 Enzymes and proteins potentially affected by exposure to Pb and the
metal cation cofactors necessary for their proper physiological
activity	 5-14
5.2.2.4	Mitochondrial Abnormality	5-15
5.2.3	Protein Binding	5-18
5.2.3.1	Intranuclear and Cytoplasmic Inclusion Bodies	5-18
5.2.3.2	Cytosolic Lead Binding Proteins	5-19
5.2.3.3	Erythrocytic Lead Binding Proteins	5-20
5.2.3.4	Metallothionein	5-21
5.2.4	Oxidative Stress	5-23
5.2.4.1	5-ALA Oxidation	5-23
5.2.4.2	Membrane and Lipid Peroxidation 	5-24
5.2.4.3	NAD(P)H Oxidase Activation 	5-25
5.2.4.4	Antioxidant Enzyme Disruption	5-26
5.2.4.5	Nitric Oxide Signaling	5-28
5.2.5	Inflammation	5-29
5.2.5.1 Cytokine Production	5-30
5.2.6	Endocrine Disruption	5-33
5.2.6.1	Hypothalamic-Pituitary-Gonadal Axis	5-33
5.2.6.2	Hypothalamic-Pituitary-Thyroid Axis	5-35
5.2.7	Cell Death and Genotoxicity	5-35
5.2.7.1	DNA Damage	5-36
5.2.7.2	Mutagenicity	5-39
5.2.7.3	Clastogenicity	5-40
5.2.7.4	Epigenetic Effects	5-44
5.2.7.5	Gene Expression	5-46
5.2.7.6	Apoptosis	 5-47
5.2.8	Summary	 5-48
Table 5-2 MOAs, their related health effects, and information on concentrations
eliciting the MOAs	 5-49
5.3	Nervous System Effects	 5-52
5.3.1	Introduction	 5-52
5.3.2	Cognitive Function and Learning	5-55
5.3.2.1 Epidemiologic Studies of Cognitive Function in Children	5-55
Figure 5-2 Associations of blood Pb levels with full-scale IQ (FSIQ) among
children.	 5-57
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Table 5-3
Figure 5-3
Table 5-4
Figure 5-4
Figure 5-5
Table 5-5
Figure 5-6
Additional characteristics and quantitative results for studies
represented in Figure 5-2	
Effect modification of the association between concurrent blood Pb
level and FSIQ by blood Mn level.	
Associations of blood Pb level with Bayley MDI in children ages
6 months to 3 yearsa	
Regression of fitted MDI score at 36 months on log-transformed cord
blood Pb level by sex.	
Standardized regression coefficients describing the associations of
blood Pb levels with specific indices of cognitive function in children. _
Additional characteristics and quantitative results for studies
represented in Figure 5-5	
Associations between childhood blood Pb levels and fourth grade
End-of-Grade (EOG) math scores.	
5.3.2.2
Figure 5-7 Greater reduction in EOG achievement test scores with increasing
blood Pb level in lower percentiles of the test score distribution.	
Toxicological Studies of Cognition, Memory and Learning	
Figure 5-8 Nervous system summary array of toxicological outcomes after Pb
exposure.	
Summary of findings from neurotoxicological exposure-response
array presented in Figure 5-8	
Table 5-6
Table 5-7
Figure 5-9
Figure 5-10
Summary of effects of maternal and lifetime Pb exposure on
Fl performance	
Mean basal corticosterone levels of female and male offspring
exposed to lifetime Pb.	
Changes in Fl performance (Fl overall performance, run rate, PRP)
in female offspring with maternal Pb exposure plus various stressors
(restraint, cold, novelty) in adulthood.	
5.3.2.3
5.3.2.4
5.3.2.5
Figure 5-11 Changes in Fl performance (Fl overall performance, run rate, PRP)
in male offspring with maternal Pb exposure plus various stressors
(restraint, cold, novelty) in adulthood.	
Toxicological Studies on the Effects of Chelation	
Integrated Summary of Cognitive Function in Children
Epidemiologic Studies of Cognitive Function in Adults
Table 5-8
Associations of blood and bone Pb levels with cognitive function in
adultsa
Table 5-9
Table 5-10
hyperactivity, and impulsivity.	
Additional characteristics and quantitative results for studies
presented in Figure 5-14	
Associations between blood Pb level and ADHD diagnosis or
diagnostic indices in children	
Figure 5-15 Adjusted odds ratios for Attention Deficit Hyperactivity Disorder
(ADHD) among U. S. children.	
5.3.3.2
5.3.3.3
5.3.3.4
5.3.3.5
Epidemiologic Studies of Mood and Psychiatric Effects in Adults 	
Toxicological Studies of Mood, Emotional, and Psychotic Changes 	
Figure 5-16 Animal toxicology evidence of possible Pb-induced contributors to
the development of mood and psychotic disorders.	
Figure 5-17 Schematic representation of the contribution of Pb exposure to the
development of a phenotype consistent with schizophrenia.	
5-58
5-63
5-67
5-69
5-74
5-75
5-82
5-83
5-84
5-85
5-86
5-91
5-94
5-97
_ 5-98
_ 5-99
5-100
5-103
5-106
Figure 5-12 Nonlinear association between patella Pb level and the relative
change over time in response latency on the pattern comparison
test.	
Figure 5-13 Exploration of nonlinear association of tibia Pb level with annual rate
of cognitive decline, by hemochromatosis (HFE) gene variant.	
5.3.3 Behavioral Effects 	
5.3.3.1 Epidemiologic Studies of Behavioral Effects in Children 	
Figure 5-14 Associations of blood Pb levels with behavioral indices of inattention,
Table 5-11 Associations between blood Pb level and misconduct and delinquent
behavior in children and young adultsa	
Toxicological Studies of Behavior	
Epidemiologic Studies of Mood in Children	
5-111
5-112
5-118
5-118
5-124
5-125
5-128
5-130
5-133
5-137
5-141
5-143
5-145
5-146
5-148
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Figure 5-18 Neurogenesis (production of new cells) in the rat dentate gyrus after
postnatal Pb exposure.	 5-149
5.3.3.6 Integrated Summary of Behavior and Mood	5-149
5.3.4	Sensory Organ Function 	5-151
5.3.4.1	Epidemiologic Studies of Sensory Organ Function in Children 	5-151
5.3.4.2	Epidemiologic Studies of Sensory Organ Function in Adults	5-152
5.3.4.3	Toxicological Studies of Sensory Organ Function 	5-154
Table 5-12 Summary of Pb-reiated retinal ERG studies	 5-157
Figure 5-19 Retinal a-wave and b-wave ERG amplitude in adult rodents after
prenatal and early postnatal Pb exposure.	 5-158
Figure 5-20 Retinal dopamine metabolism in adult control and gestationally lead-
exposed (GLE) rats. 	 5-160
5.3.5	Motor Function	5-161
5.3.6	Seizures in Animals	5-162
5.3.7	Neurodegenerative Diseases	5-163
5.3.7.1	Epidemiologic Studies of Neurodegenerative Diseases in Adults 	5-163
5.3.7.2	Toxicological Studies of Neurodegenerative Disease	5-167
5.3.8	Modes of Action for Lead Nervous System Effects	5-172
5.3.8.1	Effects on Brain Physiology and Activity	5-172
5.3.8.2	Oxidative Stress	5-174
5.3.8.3	Nitrosative Signaling and Nitrosative Stress	5-177
5.3.8.4	Synaptic Changes	5-177
5.3.8.5	Blood Brain Barrier	5-179
5.3.8.6	Cell Adhesion Molecules	5-181
5.3.8.7	Glial Effects	5-181
5.3.8.8	Neurotransmitters	5-182
5.3.8.9	Neurite Outgrowth	5-187
5.3.8.10	Epigenetics	5-188
5.3.8.11	Cholesterol and Lipid Homeostasis	5-189
5.3.9	Lifestage of Lead Exposure and Neurodevelopmental Deficits	5-189
Table 5-13 Associations of cognitive function and behavioral outcomes with
blood Pb levels measured at different lifestagesa	 5-192
Figure 5-21 Associations of cognitive function in children with different degrees
of changes in blood Pb levels over time.	 5-197
Table 5-14 Additional characteristics and quantitative results for studies
presented in Figure 5-21	 5-198
Figure 5-22 Estimated IQ in combined Cincinnati and Rochester cohorts, for
three patterns of blood Pb level levels from 1 through 6 years of age:
peak at 2 years (blue diamonds), peak at 5 years (black triangles),
and constant blood Pb level (white squares).	 5-199
5.3.10	Examination of the Lead Concentration-Response Relationship	5-200
Figure 5-23 Comparison of associations between blood Pb level and cognitive
function among various blood Pb strata.	 5-202
Table 5-15 Additional characteristics and quantitative results for studies
presented in Figure 5-23	 5-203
5.3.11	Confounding in Epidemiologic Studies of Nervous System Effects	5-209
5.3.12	Public Health Significance of Associations between Lead Biomarkers and Neurodevelopmental
Effects 	 5-213
Figure 5-24 Effect of blood Pb level on the proportion of the population with IQ
levels <70 and <80 points.	 5-216
5.3.13	Summary and Causal Determination	5-217
5.3.13.1	Cognitive Function in Children 	5-217
5.3.13.2	Behavior in Children	5-218
5.3.13.3	Other Nervous System Effects in Children	5-219
5.3.13.4	Factors that Modify Risk in Children	5-220
5.3.13.5	Nervous System Effects in Adults	5-221
5.3.13.6	Neurophysiological and Neurochemical Changes 	5-222
5.3.13.7	Lifestages and Duration of Lead Exposure	5-222
5.3.13.8	The Lead Concentration-Response Relationship	5-223
5.3.13.9	Evidence that Forms the Basis of the Causal Determination	5-224
5.4 Cardiovascular Effects	 5-225
5.4.1 Introduction	5-225
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5.4.2	Blood Pressure and Hypertension	5-227
5.4.2.1	Epidemiology	5-227
Figure 5-25 Concentration-response relationships (95% CI).	 5-229
Table 5-16 Additional characteristics and quantitative data for associations of
blood and bone Pb with BP measures for studies presented in
Figure 5-25	 5-230
Figure 5-26 Odds ratios (95% CI) for associations of blood and bone Pb with
hypertension prevalence and incidence.	 5-233
Table 5-17 Additional characteristics and quantitative data for associations of
blood and bone Pb with hypertension measures for results
presented in Figure 5-26	 5-234
Figure 5-27 The relationship between tibia Pb and estimated systolic BP (SBP)
for those with high self-reported stress versus those with low self-
reported stress.	 5-239
5.4.2.2	Toxicology	 5-241
Figure 5-28 Changes in BP after Pb exposure (represented as blood Pb level) in
unanesthetized adult rats across studies.	 5-242
Table 5-18 Characteristics of studies of blood Pb with BP measures in animals
presented in Figure 5-28	 5-243
5.4.2.3	Hypertension Modes of Action 	5-244
5.4.2.4	Summary of Blood Pressure and Hypertension	5-252
Figure 5-29 Change in systolic BP (SBP), in mmHg with 95% CI, associated with
a doubling in the blood Pb concentration.	 5-254
Figure 5-30 Prospective and cross-sectional increase in systolic BP (SBP) and
diastolic BP (DBP) and relative risk of hypertension per 10jjg/g
increase in bone Pb levels.	 5-255
5.4.3	Vascular Effects and Cardiotoxicity	5-256
5.4.3.1	Effects on Vascular Cell Types	5-256
5.4.3.2	Cholesterol	5-258
5.4.3.3	Heart Rate Variability	5-259
5.4.3.4	Peripheral Artery Disease	5-261
5.4.3.5	Ischemic Heart Disease	5-262
5.4.3.6	Atherosclerosis	5-263
Table 5-19 Characteristics and quantitative data for associations of blood and
bone Pb with other CVD measures in epidemiologic studies ordered
as they appear in the text	 5-264
5.4.3.7	Summary of Vascular Effects and Cardiotoxicity	5-265
5.4.4	Cardiovascular Function and Blood Pressure in Children	5-266
5.4.4.1	Introduction	5-266
Table 5-20 Studies of Children Cardiovascular Endpoints and Pb Biomarkers
Ordered as They Appear in the Text	 5-268
5.4.4.2	Cardiovascular Functioning in Children 		5-269
Figure 5-31 Children's adjusted total peripheral resistance (dyn-s/cm°) responses
to acute stress tasks, as a function of childhood Pb levels.	 5-270
5.4.4.3	Blood Pressure in Children	5-271
5.4.4.4	Summary of Child Cardiovascular Studies	5-272
5.4.5	Mortality	5-274
Figure 5-32 Multivariate adjusted relative hazards of all-cause and
cardiovascular mortality.	 5-276
Figure 5-33 Multivariate-adjusted relative hazard (left axis) of mortality
associated with blood Pb level between 1 yg/dL and 10jjg/dL.	 5-277
Figure 5-34 Relative risk of all cause mortality for different blood Pb levels
compared with referent level of 1.5 yg/dL (12.5th percentile).	 5-278
Figure 5-35 Associations between patella bone Pb level and the log of HR
(logHR) for all-cause, cardiovascular, and ischemic heart disease.	 5-280
Figure 5-36 Multivariate adjusted relative hazard (left axis) of mortality as a
function of blood Pb levels between 1 yg/dL and 15/jg/dL.	 5-281
5.4.5.1 Summary of Mortality	5-281
Figure 5-37 Hazard ratios for associations of blood Pb (closed markers) and
bone Pb (open markers) with all-cause mortality (black diamonds)
and cardiovascular mortality (blue circles).	 5-283
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Table 5-21 Additional characteristics and quantitative data for associations of
blood and bone Pb with CVD mortality for studies presented in
Figure 5-37	 5-284
5.4.6	Air Lead-Particulate Matter Studies	5-286
5.4.6.1	Cardiovascular Morbidity	5-286
5.4.6.2	Mortality	 5-287
5.4.7	Summary and Causal Determination	5-287
5.5	Renal Effects	 5-292
5.5.1	Introduction	5-292
5.5.1.1 Kidney Outcome Measures	5-294
5.5.2	Nephrotoxicity and Renal Pathology	5-295
5.5.2.1	Toxicology	 5-295
Figure 5-38 Concentration-response representation of the effect of Pb on renal
outcomes in animal toxicology studies.	 5-296
Table 5-22 Additional characteristics for results of toxicological studies
presented in Figure 5-38	 5-297
Table 5-23 Indicators of renal damage in male rats exposed to 50 ppm Pb for 40
and 65 days, starting at parturition	 5-299
Table 5-24 Effects of Pb on the kidney/renal system related to exposure
duration- evidence from animal toxicology studies	 5-306
5.5.2.2	Epidemiology in Adults	5-306
Figure 5-39 Kidney metric slopes for blood Pb or bone Pb.	 5-310
Figure 5-40 Percent change for kidney outcomes associated with blood Pb.	 5-312
Table 5-25 Additional characteristics and quantitative data for associations of
blood and bone Pb with kidney outcomes for results presented in
Figure 5-39 and Figure 5-40	 5-313
Table 5-26 Patient population studies: kidney function decline	 5-317
Table 5-27 Clinical randomized chelation trials in chronic kidney disease
patients	 5-321
Figure 5-41 Added variable plot of association between serum creatinine and
blood Pb in 267 Korean Pb workers in the oldest age fertile.	 5-324
5.5.2.3	Epidemiology in Children 	5-325
5.5.2.4	Associations between Lead Dose and New Kidney Outcome Measures	5-327
5.5.2.5	Reverse Causality	5-327
5.5.3	Modes of Action for Lead-Induced Nephrotoxicity	5-328
5.5.3.1	Altered Uric Acid	5-328
5.5.3.2	Oxidative Damage	5-329
5.5.3.3	Renal Gangliosides	5-333
5.5.3.4	Role of Metallothionein	5-333
5.5.4	Effects of Exposure to Lead Mixtures 	5-334
5.5.4.1	Lead and Cadmium	5-335
5.5.4.2	Lead, Cadmium, and Arsenic	5-337
5.5.4.3	Lead and Zinc	5-337
5.5.4.4	Lead and Mercury	5-338
5.5.5	Impact of Treatment with Antioxidants on Renal Lead Accumulation and Pathology	5-339
5.5.5.1	Treatment with Antioxidants	5-339
5.5.5.2	Treatment with Antioxidants plus Chelators	5-342
5.5.6	Summary and Causal Determination	5-343
5.6	Immune System Effects	 5-346
5.6.1	Introduction	5-346
Figure 5-42 Immunological pathways by which Pb exposure may increase risk of
immune-related diseases.	 5-348
5.6.2	Cell-Mediated Immunity	5-350
5.6.2.1	T Cells	 5-350
5.6.2.2	Lymphocyte Activation 	5-353
5.6.2.3	Delayed-type Hypersensitivity	5-354
5.6.2.4	Macrophages and Monocytes	5-355
5.6.2.5	Neutrophils	 5-357
5.6.2.6	Dendritic Cells	5-359
5.6.2.7	Natural Killer Cells	5-360
5.6.3	Humoral Immunity	5-361
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Table 5-28 Comparison of serum immunoglobulin levels and B cell abundance
among various blood Pb groups	 5-364
5.6.4	Immune-based Diseases	5-368
5.6.4.1	Host Resistance	5-368
5.6.4.2	Asthma and Allergy	5-369
Figure 5-43 Associations of blood Pb levels with asthma- and allergy-related
conditions in children.	 5-370
Table 5-29 Additional characteristics and quantitative results for studies
presented in Figure 5-43	 5-371
5.6.4.3	Other Respiratory Effects	5-374
5.6.4.4	Autoimmunity	5-376
5.6.4.5	Specialized Cells in Other Tissues	5-376
Figure 5-44 Specialized macrophages in nonlymphoid tissue may serve as a link
between Pb exposure and disease in multiple organ systems. 	 5-377
5.6.4.6	Tumors	 5-378
5.6.5	Modes of Action for Lead Immune Effects	5-378
5.6.5.1	Inflammation	5-378
5.6.5.2	Increased Prostaglandin E2 and Decreased Nitric Oxide	5-380
5.6.5.3	Cellular Death (Apoptosis, Necrosis)	5-382
5.6.5.4	Cytokine Production	5-383
5.6.6	Air-Lead Studies	5-386
5.6.7	Immune Effects of Lead within Mixtures 	5-388
5.6.8	Summary and Causal Determination	5-389
5.7	Effects on Heme Synthesis and Red Blood Cell Function	 5-394
5.7.1	Summary of Findings from 2006 Pb AQCD	5-394
5.7.2	Red Blood Cell Functions	5-396
5.7.2.1	Pb Uptake, Binding, and Transport into Red Blood Cells	5-396
5.7.2.2	Red Blood Cell Survival, Mobility, and Membrane Integrity	5-397
5.7.2.3	Red Blood Cell Hematopoiesis	5-404
5.7.2.4	Membrane Proteins	5-405
5.7.3	Red Blood Cell Heme Metabolism	5-406
Figure 5-45 Schematic representation of the enzymatic steps involved in the
heme synthetic pathway.	 5-406
5.7.3.1 Red Blood Cell 5-Aminolevulinic Acid Dehydratase	5-407
5.7.4	Other Heme Metabolism Enzymes 	5-408
5.7.5	Other Hematological Parameters	5-409
5.7.5.1	Energy Metabolism	5-409
5.7.5.2	Other Enzymes	5-410
5.7.6	Red Blood Cell Oxidative Stress	5-410
5.7.6.1	Oxidative Stress, Lipid Peroxidation, and Antioxidant Enzymes	5-411
5.7.6.2	Antioxidant Defense	5-413
5.7.7	Summary and Causal Determination	5-414
5.8	Reproductive and Developmental Effects	 5-417
5.8.1	Effects on Female Reproductive Function	5-418
5.8.1.1	Effects on Female Sex Endocrine System and Estrus Cycle	5-418
Table 5-30 Summary of recent epidemiologic studies of associations between
Pb levels and hormones for females	 5-419
5.8.1.2	Effects on Fertility	5-423
Table 5-31 Summary of recent epidemiologic studies of associations between
Pb levels and fertility for females	 5-424
5.8.1.3	Effects on Puberty	5-427
Table 5-32 Summary of recent epidemiologic studies of associations between
Pb levels and puberty for females	 5-428
Figure 5-46 Toxicological exposure-response array for reproductive effects
ofPb.	 5-432
Table 5-33 Toxicological concentration-response array summary for
reproductive effects of Pb presented in Figure 5-46	 5-433
5.8.1.4	Effects on Lactation	5-435
5.8.1.5	Summary of Effects on Female Reproductive Function	5-436
5.8.2	Effects on Male Reproductive Function	5-436
5.8.2.1 Effects on Sperm/Semen Production, Quality, and Function	5-437
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Table 5-34 Summary of recent epidemiologic studies of associations between
Pb levels and effects on sperm and semen	 5-437
5.8.2.2	Effects on Hormone Levels	5-444
Table 5-35 Summary of recent epidemiologic studies of associations between
Pb levels and hormones for males	 5-445
5.8.2.3	Fertility	5-450
5.8.2.4	Puberty	5-451
Table 5-36 Summary of recent epidemiologic studies of associations between
Pb levels and puberty for males	 5-452
5.8.2.5	Effects on Morphology and Histology of Male Sex Organs	5-455
5.8.2.6	Summary of Effects on Male Reproductive Function	5-455
5.8.3	Effects on Ovaries, Embryo Development, Placental function, and Spontaneous Abortions5-456
Table 5-37 Summary of recent epidemiologic studies of associations between
Pb levels and spontaneous abortions	 5-457
5.8.4	Infant Mortality and Embryogenesis	5-460
5.8.5	Birth Defects	5-460
Table 5-38 Summary of recent epidemiologic studies of associations between
Pb levels and neural tube defects	 5-461
5.8.6	Preterm Birth	5-462
Table 5-39 Summary of recent epidemiologic studies of associations between
Pb levels and preterm birth	 5-463
5.8.7	Low Birth Weight/Fetal Growth 	5-466
Table 5-40 Summary of recent epidemiologic studies of associations between
Pb levels and low birth weight and fetal growth	 5-467
5.8.8	Effects on Postnatal Stature and Body Weight	5-475
Table 5-41 Summary of recent epidemiologic studies of associations between
Pb levels and postnatal growth	 5-476
5.8.9	Toxicological Studies of Developmental Effects	5-483
5.8.9.1	Developmental Effects on Blood and Liver	5-483
5.8.9.2	Developmental Effects on Skin	5-484
5.8.9.3	Developmental Effects on the Retina	5-484
5.8.9.4	Developmental Effects on Teeth	5-485
5.8.10	Summary and Causal Determination	5-485
5.9	Effects on Other Organ Systems	 5-488
5.9.1	Effects on the Hepatic System	5-488
5.9.1.1	Summary of Key Findings of the Effects on the Hepatic System from the 2006 Lead
AQCD	 5-488
5.9.1.2	New Epidemiologic Studies	5-489
5.9.1.3	New Toxicological Studies	5-490
5.9.2	Effects on the Gastrointestinal System	5-496
5.9.2.1	Summary of Key Findings on the Effects on the Gastrointestinal System from the 2006
Lead AQCD	 5-496
5.9.2.2	New Epidemiologic Studies	5-497
5.9.2.3	New Toxicological Studies	5-497
5.9.3	Effects on the Endocrine System	5-498
5.9.3.1	Summary of Key Findings of the Effects on the Endocrine System from the 2006 Lead
AQCD	 5-498
5.9.3.2	New Epidemiologic Studies	5-498
5.9.3.3	New Toxicological Studies	5-501
5.9.4	Effects on Bone and Teeth	5-501
5.9.4.1	Summary of Key Findings of the Effects on Bone and Teeth from the 2006 Lead AQCD5-502
5.9.4.2	New Toxicological and Epidemiologic Studies	5-503
5.9.5	Effects on Ocular Health 	5-507
5.9.5.1	Summary of Key Findings of the Effects on Ocular Health from the 2006 Lead AQCD5-507
5.9.5.2	New Toxicological and Epidemiologic Studies	5-507
5.9.6	Effects on the Respiratory System	5-508
5.9.7	Summary	5-509
5.10	Cancer	 5-511
5.10.1 Cancer Incidence and Mortality	5-512
Table 5-42 Summary of recent epidemiologic studiesa of cancer incidence and
mortality	 5-513
5.10.1.1 Overall Cancer Mortality	5-517
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5.10.1.2	Lung Cancer	5-518
5.10.1.3	Brain Cancer	5-519
5.10.1.4	Breast Cancer	5-520
5.10.1.5	Other Cancers	5-521
5.10.1.6	Toxicological Models of Carcinogenicity	5-523
5.10.2	Cancer Biomarkers	5-524
5.10.3	DNA and Cellular Damage	5-525
5.10.3.1	Epidemiologic Evidence for DNA and Cellular Damage	5-525
5.10.3.2	Toxicological Evidence for DNA and Cellular Damage	5-526
5.10.4	Effects of Lead within Mixtures 	5-533
5.10.5	Modes of Action	5-534
5.10.5.1 Epigenetics	5-535
5.10.6	Summary and Causal Determination	5-536
CHAPTER 6 POTENTIALLY AT-RISK POPULATIONS	6-1
6.1	Physiological Factors that Influence the Internal Distribution of Lead	 6-2
6.2	Population Characteristics Potentially Related to Differential Lead Exposure	 6-4
6.2.1	Age 6-4
6.2.1.1	Early Childhood	6-4
Table 6-1 Blood Pb levels by age and sex, 2007-2008 NHANES	 6-5
Table 6-2 Percentage of children within six categories/brackets of blood Pb
levels, 1999-2004 NHANES	 6-6
6.2.1.2	Adulthood	6-7
6.2.2	Sex 6-7
6.2.3	Race and Ethnicity	6-8
Figure 6-1 Percent distribution of blood Pb levels by race/ethnicity among U. S.
children (1-5 years).	 6-9
Figure 6-2 Soil Pb concentration exposure among the population of three
parishes within greater metropolitan New Orleans.	 6-10
6.2.4	Socioeconomic Status (SES)	6-10
6.2.5	Proximity to Lead Sources	6-12
6.2.6	Residential Factors	6-13
Table 6-3 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	 6-14
6.3	Factors Potentially Related to Increased Risk of Lead Induced Health Effects	 6-15
Table 6-4 Summary of evidence for factors that potentially increase the risk of
lead-related health effects	 6-16
6.3.1	Age 6-16
6.3.1.1	Children	6-16
6.3.1.2	Older Adults	6-18
6.3.2	Sex 6-19
6.3.3	Genetics	6-22
6.3.3.1	Aminolevulinate Dehydratase	6-23
6.3.3.2	Vitamin D Receptor	6-24
6.3.3.3	Methylenetetrahydrofolate reductase	6-24
6.3.3.4	Apolipoprotein E	6-25
6.3.3.5	Hemochromatosis	6-25
6.3.3.6	Other Genetic Polymorphisms	6-25
6.3.4	Pre-existing Diseases/Conditions 	6-26
6.3.4.1	Autism	 6-27
6.3.4.2	Atopy	 6-27
6.3.4.3	Diabetes	6-27
6.3.4.4	Hypertension	6-28
6.3.5	Smoking	6-29
6.3.6	Race/Ethnicity	6-30
6.3.7	Socioeconomic Status	6-31
6.3.8	Body Mass Index	6-31
6.3.9	Alcohol Consumption	6-32
6.3.10	Nutritional Factors	6-32
6.3.10.1 Calcium	6-32
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6.3.10.2	Iron 6-32
6.3.10.3	Folate 	 6-33
6.3.10.4	Protein	6-33
6.3.11	Stress	6-33
6.3.12	Cognitive Reserve	6-34
6.3.13	Other Metal Exposure	6-34
6.3.13.1	Cadmium	6-35
6.3.13.2	Arsenic	6-35
6.3.13.3	Manganese	6-35
6.4 Summary	 6-36
CHAPTER 7 ECOLOGICAL EFFECTS OF LEAD 	7-1
7.1	Introduction to Ecological Concepts	 7-1
7.1.1	Ecosystem Scale, Function, and Structure	7-2
7.1.2	Ecosystem Services	 7-4
7.1.3	Critical Loads as an Organizing Principle for Ecological Effects of Atmospheric Deposition 7-5
7.1.4	Ecosystem Exposure, Lag Time and Re-entrainment of Historically Deposited Lead	7-6
Table 7-1 Comparison among several metals: Time to achieve 95% of steady
state metal concentration in soil; example in a temperate system	 7-8
7.2	Terrestrial Ecosystem Effects	 7-9
7.2.1	Introduction to Terrestrial Ecosystem Effects	7-9
7.2.2	Soil Biogeochemistry and Chemical Effects	7-10
7.2.2.1	pH, CEC and Salinity 	7-10
7.2.2.2	Organic Matter	7-11
7.2.2.3	Aging	 7-13
7.2.3	Bioavailability in Terrestrial Systems	7-14
Figure 7-1 Conceptual diagram for evaluating bioavailability processes and
bioaccessibility for metals in soil, sediment, or aquatic systems.	 7-16
Figure 7-2 Schematic diagram of the biotic ligand model.	 7-7 7
7.2.3.1	Terrestrial Plants	7-17
7.2.3.2	Terrestrial Invertebrates	7-26
7.2.3.3	Terrestrial Vertebrates 	7-30
7.2.3.4	Food Web	 7-32
7.2.4	Biological Effects	7-35
7.2.4.1	Terrestrial Plants and Lichen 	7-35
7.2.4.2	Terrestrial Invertebrates	7-40
7.2.4.3	Terrestrial Vertebrates 	7-44
7.2.5	Exposure and Response of Terrestrial Species	7-49
7.2.6	Community and Ecosystem Effects	7-52
7.2.7	Critical Loads in Terrestrial Systems	7-57
7.2.8	Soil Screening Levels	7-59
7.2.9	Characterization of Sensitivity and Vulnerability	7-60
7.2.9.1	Species Sensitivity	7-60
7.2.9.2	Nutritional Factors	7-61
7.2.9.3	Soil Aging and Site-Specific Bioavailability	7-61
7.2.9.4	Ecosystem Vulnerability	7-62
7.2.10	Ecosystem Services	7-63
7.2.11	Summary of Effects in Terrestrial Systems	7-64
7.2.11.1	Biogeochemistry and Chemical Effects	7-64
7.2.11.2	Bioavailability and Uptake	7-65
7.2.11.3	Biological Effects	7-66
7.2.11.4	Exposure Response	7-67
7.2.11.5	Community and Ecosystem Effects 	7-68
7.2.11.6	Critical Loads, Sensitivity and Vulnerability	7-68
7.3	Aquatic Ecosystem Effects	 7-69
7.3.1	Introduction to Aquatic Ecosystem Effects	7-69
7.3.2	Biogeochemistry and Chemical Effects of Pb in Freshwater and Saltwater Systems	7-69
7.3.2.1	Other Metals	7-72
7.3.2.2	Biofilm	 7-72
7.3.2.3	Carbonate	7-73
7.3.2.4	Dissolved Organic Matter (DOM)	7-74
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7.3.2.5 Sulfides	7-76
7.3.3	Introduction to Bioavailability and Biological Effects of Pb in Freshwater Ecosystems	7-76
7.3.4	Bioavailability in Freshwater Systems	7-78
7.3.4.1	Freshwater Plants and Algae	7-82
7.3.4.2	Freshwater Invertebrates	7-87
7.3.4.3	Freshwater Vertebrates	7-90
7.3.4.4	Food Web	 7-97
7.3.5	Biological Effects of Pb in Freshwater Systems	7-99
7.3.5.1	Freshwater Plants and Algae	7-100
7.3.5.2	Freshwater Invertebrates	7-103
7.3.5.3	Freshwater Vertebrates	7-107
7.3.6	Exposure and Response of Freshwater Species	7-118
7.3.7	Freshwater Community and Ecosystem Effects	7-122
7.3.8	Critical Loads in Freshwater Aquatic Systems	7-126
7.3.9	Characterization of Sensitivity and Vulnerability in Freshwater Systems 	7-127
7.3.9.1	Seasonally-Affected Physiological Changes	7-128
7.3.9.2	Increased Nutrient Uptake 	7-128
7.3.9.3	Temperature and pH	7-128
7.3.9.4	Lifestage	7-129
7.3.9.5	Species Sensitivity	7-130
7.3.9.6	Ecosystem Vulnerability	7-131
7.3.10	Introduction to Bioavailability and Biological Effects of Pb in Saltwater Ecosystems	7-132
7.3.11	Bioavailability of Pb in Saltwater Systems	7-133
7.3.11.1	Saltwater Plants and Algae	7-134
7.3.11.2	Saltwater Invertebrates	7-135
7.3.11.3	Saltwater Vertebrates	7-138
7.3.11.4	Marine Food Web	7-139
7.3.12	Biological Effects of Pb in Saltwater Systems	7-140
7.3.12.1	Saltwater Algae and Plants	7-140
7.3.12.2	Saltwater Invertebrates	7-141
7.3.12.3	Saltwater Vertebrates	7-143
7.3.13	Exposure and Response of Saltwater Species	7-145
7.3.14	Community and Ecosystem Effects in Saltwater Systems	7-146
7.3.15	Characterization of Sensitivity and Vulnerability in Saltwater Species 	7-147
7.3.15.1	Seasonally Affected Physiological Changes	7-148
7.3.15.2	Lifestage	7-149
7.3.15.3	Species Sensitivity	7-149
7.3.16	Ecosystem Services Associated with Freshwater and Marine Systems	7-150
7.3.17	Summary of Aquatic Effects	7-151
7.3.17.1	Biogeochemistry and Chemical Effects	7-152
7.3.17.2	Bioavailability	7-152
7.3.17.3	Biological Effects	7-154
7.3.17.4	Exposure and Response	7-155
7.3.17.5	Community and Ecosystem Effects 	7-156
7.3.17.6	Critical Loads, Sensitivity and Vulnerability	7-157
7.4 Causal Determinations for Ecological Effects of Lead	 7-157
Table 7-2 Summary of Pb causal determinations for plants, invertebrates and
vertebrates	 7-159
7.4.1	Causal Determinations for Lead in Terrestrial Systems	7-160
7.4.1.1	Physiological Stress-Terrestrial Biota	7-160
7.4.1.2	Hematological Effects-Terrestrial Biota	7-162
7.4.1.3	Neurobehavioral Effects-Terrestrial Biota	7-162
7.4.1.4	Developmental and Reproductive Effects-Terrestrial Biota	7-164
7.4.1.5	Growth Effects-Terrestrial Biota	7-166
7.4.1.6	Survival-Terrestrial Biota	7-167
7.4.1.7	Community and Ecosystem Level Effects-Terrestrial Biota	7-168
7.4.2	Causal Determinations for Lead in Aquatic Systems	7-171
7.4.2.1	Physiological Stress-Aquatic Biota	7-171
7.4.2.2	Hematological Effects-Aquatic Biota	7-172
7.4.2.3	Neurobehavioral Effects-Aquatic Biota	7-173
7.4.2.4	Developmental and Reproductive Effects-Aquatic Biota 	7-175
7.4.2.5	Growth Effects-Aquatic Biota	7-178
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7.4.2.6	Survival-Aquatic Biota	7-180
7.4.2.7	Community and Ecosystem Level Effects-Aquatic Biota	7-182
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Lead Project Team
Executive Direction
Dr. John Vandenberg (Director)—National Center for Environmental Assessment-RTP
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-
RTP 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. 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 (Pb Team Leader)—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
Dr. Meredith Lassiter—National Center for Environmental Assessment, Office of
Research and Development, U.S. Environmental Protection Agency, Research
Triangle Park, NC
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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
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
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
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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
Authors, Contributors, and Reviewers
Authors
Dr. Ellen Kirrane (Pb Team Leader)—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
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
Dr. Meredith Lassiter—National Center for Environmental Assessment, Office of
Research and Development, U.S. Environmental Protection Agency, Research
Triangle Park, NC
February 2012
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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
Dr. Marc Weisskopf—Department of Environmental Health and Department of
Epidemiology, Harvard School of Public Health, Harvard University, Boston,
Massachusetts
February 2012
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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
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 Tuttle—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
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. Christal Bowman—National Center for Environmental Assessment, Office of
Research and Development, U.S. Environmental Protection Agency, Research
Triangle Park, NC
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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, Darmouth 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. Douglas Johns—National Center for Environmental Assessment, Office of Research
and Development, U.S. Environmental Protection Agency, Research Triangle Park,
NC
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
February 2012
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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
Ms. Debra Walsh—National Center for Environmental Assessment-RTP 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
February 2012
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Clean Air Scientific Advisory Committee
Lead NAAQS Review Panel
Chair of the Environmental Protection Agency's Clean Air Scientific Advisory Committee
Dr. Jonathan M. Samet*, Department of Preventive Medicine at the Keck School of
Medicine, and Director of the Institute for Global Health at the University of Southern
California, Los Angeles, CA
Chair of the Lead Review Panel
Dr. Christopher H. Frey*, North Carolina State University, Raleigh, NC
Lead Review Panel Members
Dr. George A. Allen*, Northeast States for Coordinated Air Use Management
(NESCAUM), Boston, MA
Dr. Herbert Allen, University of Delaware, Newark, DE
Dr. Richard Canfield, Cornel University, Ithaca, NY
Dr. Deborah Cory-Slechta, University of Rochester, Rochester, NY
Dr. Cliff Davidson, Syracuse University, Syracuse, NY
Dr. Philip E. Goodrum, Environmental Resources Management (ERM), Dewitt, NY
Dr. Sean Hays, Summit Toxicology, Allenspark, CO
Dr. Philip Hopke, Clarkson University, Potsdam, NY
Dr. Susan Korrick, Harvard Medical School, Boston, MA
Dr. Michael Kosnett, University of Colorado Health Sciences Center, Denver, CO
Dr. Roman Lanno, Ohio State University, Columbus, OH
Dr. Richard L. Poirot, Vermont Agency of Natural Resources, Waterbury, VT
Dr. Joel Pounds, Battelle-Pacific Northwest National Laboratory, Richland, WA
Dr. Michael Rabinowitz, Harvard University, 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
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Science Advisory Board Staff
Mr. Aaron Yeow, Designated Federal Officer, Office of Administration, Science
Advisory Board Staff Office, U.S. Environmental Protection Agency, Washington,
DC 20004. Phone: 202-564-2050, Email: yeow.aaron@epa.gov
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Acronyms and Abbreviations
a	alpha
aT the extent of DNA denaturation per cell
A	Angstrom (10-10 meter)
AA African American; arachidonic acid, atomic
absorption
AALM All Ages Lead Model
AAS atomic absorption (spectrophotometry,
spectrometry, spectroscopy)
Ab amyloid-beta peptide
ABL atmospheric boundary layer
ACE angiotensin converting enzyme
ACh acetylcholine
ACP acid phosphatase
ACR acute to chronic ratio
Acyl-Co acyl-coenzyme A
AD axial diffusivity
ADHD attention deficit hyperactivity disorder
ADP adenosine diphosphate
AE anion exchanger
AF absorbed fraction; absorption fraction
AJG albumin/globulin
Ag silver
A-horizon: Topsoil horizon (surface soil)
AKI acute kidney injury
A1 aluminum
ALA aminolevulinic acid
ALAD aminolevulinic acid dehydratase;
ALAD 1-1: aminolevulinate delta-dehydratase 1-1
ALAD-2 aminolevulinate delta-dehydratase-2
ALD
aldehyde dehydrogenase
ALM
Adult Lead Methodology
ALP
alkaline phosphatase
ALT
alanine aminotransferase
AM
Alveolar macrophages
AMF
arbuscular mycorrhizal fungi
AMP
adenosine monophosphate
ANC acid neutralizing capacity; absolute neutrophil
counts
ANF atrial natriuretic factor
Angll renal angiotensin II
ANOVA analysis of variance
ANPR advance notice of proposed rulemaking
AP-1 activator protein-1
Apal polymorphism of the VDR in humans
APC antigen-presenting cell
APOE Apolipoprotein E
APRT adenine phosphoribosyltransferase
AQCD Air Quality Criteria Document
AQS (U.S. EPA) Air Quality System (database)
As arsenic
AST aspartate aminotransferase
ASV anode stripping voltammetry
ATLD ataxia-telangiectasia-like disorder
ATOFMSaerosol time-of-flight mass spectrometry
ATP adenosine-triphosphate
ATPase adenosine triphosphatase; adenosine
triphosphate synthase
ATSDR Agency for Toxic Substances and Disease
Research
Au gold
avg average
AVS acid-volitile sulfides
a-wave initial negative deflection in the
electroretinogram
AWQC Ambient Water Quality Criteria
P	Beta; Beta coefficient; regression coefficient;
standardized coefficient
3P-HSD 3-beta-hydroxysteroid dehydrogenase
17P-HSD: 17-beta-hydroxysteroid dehydrogenase
Ba barium
BAF bioaccumulation factors
BAL 2,3-dimercaptopropanol
BASC Behavior Assessment System for Children
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BASC-PRS: Behavior Assessment System for Children-
Parent Ratings Scale
BASC-TRS: Behavior Assessment System for Children-
Teacher Rating Scale
BCB blood cerebrospinal fluid barrier
B-cell Bone marrow-derived lymphocytes, B
lymphocyte
BCF bioconcentration factors
Bcl-x member of the B-cell lymphoma-2 protein
family
Bcl-xl B-cell lymphoma-extra large
B-horizon: subsoil horizon
bio biological
Bi2S3 bismuth (III) sulfide
BK biokinetics
BLM biotic ligand model
BMD benchmark dose; bone mineral density
BMDL benchmark dose limit
BMI body mass index
BMP bone morphogenetic protein
BMS Baltimore Memory Study
BMW battery manufacturing workers
BP blood pressure
BR bronchial responsiveness
BrdU bromo-2'-deoxyuridine
8-Br-GMPc: 8-bromo-cyclic guanosine monophosphate
Bs-horizon; subsoil horizon with accumulation of
sesquioxides
BSI Brief Symptom Inventory
BSID-II Bay ley Scale for Infant Development-II
BsmI polymorphism of the VDR in humans
Bt20 birth to 20 cohort
BUN blood urea nitrogen
bw body weight
b-wave initial positive deflection in the
electroretinogram
C	carbon; Celsius; soil or dry sediment Pb
concentration; Caucasian; Cysteine
Ca	calcium
Ca2+	calcium ion
CAA	Clean Air Act
CaBP	calcium binding protein
CaCl2 calcium chloride
CaC03 calcium carbonate; calcite
CaEDTA calcium ethylenediaminetetraacetic acid
CaMKII calmodulin-dependent protein kinase II
cAMP cyclic adenosine monophosphate
CASAC Clean Air Scientific Advisory Committee
CASM Comprehensive Aquatic Systems Model
CaS04 calcium sulfate
CaS04.2H20 : gypsum
CAT catalase
CBLI cumulative blood lead index
CBSA core based statistical area
CD cluster of differentiation
Cd cadmium
Cd(II) cadmium (II)
Cd2+ cadmium ion
CD3+ T lymphocyte
CD4+ T helper cell
CDC Centers for Disease Control
CEA carcinoembryonic antigen
CEC cation exchange capacity
cent central
cert. certiorari
cf	correction factor; latin abbreviation for
conferre (used as "compared with)
CFL	contanst flux layer
CFR	Code of Federal Regulations
cGMP	cyclic guanosine monophosphate
ChAT	chlorine acetyltranferase
CHD	coronary heart disease
CHL	Chinese hamster lung
CHO	Chinese hamster ovary cell line
C-horizon: Soil horizon underneath A- and B-horizons,
may contain lumps or shelves of rock and
parent material
CHV79	Chinese hamster lung cell line
CI	confidence interval
Cir.	circuit
CKD	chronic kidney disease
CKD-EPI: Chronic Kidney Disease Epidemiology
Collaboration
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CL confidence limit
CI chlorine
CI" chlorine ion
Cl2 molecular chlorine
CLACE 5: Fifth Cloud and Aerosol Characterization
Experiment in the Free Troposphere campaign
CLS Cincinnati Lead Study
CO carbon monoxide
C02 carbon dioxide
C032" carbonate ion
Co cobalt
CoA coenzyme A
COD coefficient of difference
Coeff coefficient
COMP aT: The percentage of sperm with increased
sensitivity to DNA denaturation
Con	control
Cone.	concentration
Cong.	congress
Corr	correlation
COX	cyclooxygenase; cytochrome oxidase subunits
COX-2	cyclooxygenase-2
cPLA2	cytosolic phopholipidase A2
CPRS-R	Conners' Parent Rating Scale-Revised
Cr	chromium; creatine
Cr III	chromium III
CRAC	Ca2+ release activated calcium
CRACI	calcium release activated calcium influx
CREB cyclic adenosinemonophosphate (cAMP)
response element-binding
CRP C-reactive protein
CSF colony-stimulating factor
CSN Chemical Speciation Network
CT zinc-adequate control
Cu copper
Cu(II) copper (II)
CV coefficient of variation
CVD cardiovasicular disease
CYP cytochrome
CYP 1A1, CyplAl: cytochrome P450 family
1 member A1
CYP 1A2, Cypl A2: cytochrome P450 family 1 member
A2
CYPP450: cytochrome P450
A	delta, difference, change
A5-3P-HSD : delta-5-3-beta-hydroxysteroid
dehydrogenase
5-ALA	5-aminolevulinic acid; delta-aminolevulinic
acid
8-ALAD : delta-aminolevulinic acid dehydratase
D2, D3	dopamine receptors
D50	size at 50% efficiency
d	day(s); depth
db, dB	decibel
DbH	dopamine beta-hyrdoxylase
DBP	diastolic blood pressure
dep	dependent
dev.	deviation
DEX	exogenous dexamethasone
DG	degenerate gyrus
2-dG	2-deoxyguanosine
DHAA	dehydroascorbate
diff	differentiation
DIT	developmental immunotoxicity
DMPS	2,3-dimercaptopropane-l-sulfonic acid
DMSA	dimercaptosuccinic acid
DMSO	dimethyl sulfoxide
DNA	deoxyribonucleic acid
DoAD	developmental origins of adult disease
DOC	dissolved organic carbon
DOM	dissolved organic matter
DP-109	metal chelator
DP-460	metal chelator
DR	diet-restricted
DRD4	dopamine 4 receptor
DRD4.7	dopamine 4 receptor repeat alleles
DRUM	Davis Rotating Unit for Monitoring
D-serine	neuronal signal
DSM-IV Diagnostic Statistical Manual-IV
DTH	delayed-type hypersensitivity
DTPA diethylene triamine pentaacetic acid;
techetium-diethylenetriamine-pentaacetic acid
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E
east; expression for exposure
E2
estradiol
e
exponential function
EC
endothelial cell
EC10
effect concentration for 10% of test population
ec20
effect concentration for 20% of test population
ec50
effect concentration for 50% of test population
ECG
electrocardiography; electrocardiogram
ECOD
7-ethoxycoumarin-o-deethy lase
Eco-SSLs: ecological soil screening levels
EDjo
effect dose for 10%) of population
EDTA
ethylenediaminetetraacetic acid
EFS
electrical field stimulus
EGF
epidermal growth factor
EGFR
epidermal growth factor receptor
eGFR
estimated glomerular filtration rate
Eh
electrochemical potential
E-horizon: soil horizon with eluviated or leached of

mineral and/or organic content
EI-MS
electron impact ionization mass spectrometry
eNOS
endothelial nitric oxide synthase
EOG
end-of-grade
EPA
U.S. Environmental Protection Agency
EPT
ephemeroptera, plecoptera, trichoptera
ER
endoplasmic reticulum
Erg-1
ether-a-go-go related gene
ERG
electroretinogram
ERK
extracellular signal regulated kinase
ERK1/2
extracellular signal-regulated kinases 1 and 2
EROD
7-ethoxyresorufin-o-deethylase
ESCA
electron spectroscopy for chemical analysis
ESI-MS
electrospray ionization mass spectrometry
ESRD
end stage renal disease
ET
endothelin
ET-1
vasoconstrictor endothelin-1
ETA-type receptors : endothelin type A receptors
EU European Union
EURO European emission standard
eV electronvolts
EXAFS X-ray absorption fine structure spectroscopy
F0	filial 0 generation
F i	first offspring generation
F2	second offspring generation
FAA	Federal Aviation Agency
FAI	free androgen index
FAS	apoptosis stimulating fragment
Fas-L	apoptosis stimulating fragment ligand
Fe	iron
Fe(III)	iron III
FEM	Federal equivalence method
FEV1	forced expiratory volume in 1 second
FI	fixed interval
FI-Ext	fixed interval with extinction
Fl	fluoride
Fokl	polymorphism of the VDR in humans
FR	Federal Register (Notice)
FrA	fractional anisotropy
FR-FI	fixed ratio-fixed interval
FRM	Federal reference method
FSH	follicle-stimulating hormone
FSIQ	full scale intelligence quotient (IQ)
FT3	free triodothyronine
FT4	free thyroxine
G	pregnancy; guanine
G2	gap 2 Phase
g, mg, kg, (j.g, ng, pg: Gram(s), milligram(s),
microgram(s), kilogram(s), nanogram(s),
picogram(s)
G93A	mouse model
GABA	y-aminobutyric acid; gamma aminobutyric acid
GABAergic: gamma aminobutyric acid-ergic
GAD	generalized anxiety disorder
GC	gas chromatography
G-CSF	granulocyte colony-stimulating factor
GD	gestational day
GEE	generalized estimating equations
GFAAS	graphite furnace atomic absorption
spectrometry
GFAP	glial fibrillary acidic protein
GFR	glomerular filtration rate
GGT	gamma-glutamyl transpeptidase
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GH	growth hormone
GI	gastrointestinal
GIS	Geographic Information System
G+L	pregnancy plus lactation
GLE	gestationally-lead exposed
GM	geometric mean
GMR	geometric mean blood lead ratio
GnRH	gonadotropin-releasing hormone
G6PD	glucose-6-phosphate dehydrogenase
GPEI	glutathione transferase P (GST-P) enhancer I
GPT	glutamate pyruvate transaminase
GPx	glutathione peroxidase
GPX1	gene encoding for glutathione peroxidase 1
GR	glutathione reductase
GRP78	glucose-regulated protein 78
GRP94	glucose-regulated protein 94
Grp	glucose-regulated protein
GSD	geometric standard deviation
GSH	glutathione
GSSG	glutathione disulfide
GST	glutathione S-transferase
GSTM1	glutathione S-transferase Mu 1
GST-P	glutathione transferase P
GTP	guanosine-5'-triphosphate; guanine
triphosphate
H	hydrogen
H+	hydrogen ion
h	hour(s)
ha	hectare
HAD	hydroxyalkenals
Hb	hemoglobin
HC5	acute toxicity hazardous concentration for 5%
of species
HC10	acute toxicity hazardous concentration for 10%
of species
HC1	hydrochloric acid
HC03"	bicarbonate; hydrogen carbonate
Hct	hematocrit
HDL	high-density lipoprotein
HF	hydrogen fluoride
HFE	hemochromatosis gene
HFE C282Y: hemochromatosis gene with C282Y
mutation
HFE H63D : hemochromatosis gene with H63D
mutation
Hg mercury
HgCl2 mercury(II) chloride
5-HIAA 5-hydroxyindoleacetic acid
HIV human immunodeficiency virus
HLA-DRB: human leukocyte antigen genes
HMEC human dermal microvascular endothelial cells
HMGR 3-hydroxy-3-methylglutaryl-CoA reductase
HMOX-1: heme oxygenase-1
HN03 nitric acid
HO-1 heme oxygenase; heme oxidase-1
H20 water
H202 hydrogen peroxide
HOME Home Observation for Measurement of the
Environment
HPA hypothalamic-pituitary-adrenal
HPb, h-Pb: high lead
HPG hypothalamic-pituitary-gonadal
HPLC high-performance liquid chromatography
HPRT hypoxanthine-guanine
phosphoribosyltransferase
HPT	hyperparathyroidism; hypothalamic-pituitary-
thyroid
HR	heart rate; hazard ratio
HRV	heart rate variability
hsp	heat shock proteins
5HT	serotonin
5-HT	5-hydroxy try ptamine
5-HT2B	5-hydroxy try ptamine receptor 2B
hTERT	telomerase reverse transcriptase
HVA	homovanillic acid
I	interstate
IARC	International Agency for Research on Cancer
IC50	half maximal inhibitory concentration
ICAP	inductively coupled argon plasma
ICP-AESinductively coupled plasma atomic emission
spectroscopy
ICPMS, ICP-MS: Inductively coupled plasma mass
spectrometry
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ICR	imprinting control region
ICRP	International Commission on Radiological
Protection
ID	identification
IDA	iron-deficiency anemia
IDE	insulin-degrading enzyme
IEPA	Illinois Environmental Protection Agency
IEUBK Integrated Exposure Uptake Biokinetic
IFN-y	interferon-gamma
Ig	immunoglobulin
IgA	immunoglobulin A
IgE	immunoglobulin E
IGF-1	insulin-like growth factor 1
IgG	immunoglobulin G
IgM	immunoglobulin M
IHD	ischemic heart disease
IL	interleukin
IL-ip	interleukin-1 Beta
IL-2	interleukin-2
IL-4	interleukin-4
IL-5	interleukin-5
IL-6	interleukin-6
IL-8	interleukin-8
IL-10	interleukin-10
IL-12	interleukin-12
IMPROVE: Interagency Monitoring of Protected Visual

Environment
IMT
intimal medial thickening
INL
inner neuroblastic layers of the retina
iNOS
inducible nitric oxide synthase
i.p.
intraperitoneal (route)
IQ
intelligence quotient
IQR
interquartile range
IRE1
inositol-requiring enzyme-1
ISA
Integrated Science Assessment
ISF
intake slope factor
ISL
inertial sublayer
ISO
International Standards Organization
i.v.
intravenous
IVBA
in vitro bioaccessibility
IVF in vitro fertilization
JNK jun N-terminal kinase
K	Kelvin; potassium; resupsension factor
K+ potassium ion
K0 5 concentration of free metal giving half
maximal metal-dependent release
KART Karters of American Racing Triad
Kd dissociation constant
Kd partition coefficient; ratio of the metal
concentration in soil to that in soil solution
kDa, kD kiloDalton
KEDI-WISC: Korean Educational Development
Institute-Wechsler Intelligence Scale for
Children
6-keto-PGFla: 6-keto-prostaglandin Fla (vasodilatory
prostaglandin)
keV kiloelectron volt
Ki-67 antigen, cell cycle and tumor growth marker
Kim-1 kidney injury molecule-1
Kinder-KITAP : Kinder-Testbatterie zur
Aufmerksamkeitspriifung fur Kinder
K-ras specific protooncogene
A	lambda; resuspension rate
L	length
L, mL, dL: liter(s), milliliter(s), deciliter(s)
LA-ICP-MS: laser ablation inductively coupled plasma
mass spectrometry
LC50 lethal concentration (at which 50% of exposed
organisms die)
LD50 lethal dose (at which 50% of exposed
organisms die)
LDH lactate dehydrogenase
LDL low-density lipoproteins
LFH-horizons: organic soil horizons located above well-

drained surface soil
LF/HF
low frequency to high frequency ratio
LH
luteinizing hormone
LHRH
luteinizing hormone releasing hormone
LINE
long interspersed nuclear element
LINE-1
long interspersed nucleotide elements-1
LLNA
local lymph node assay
In
natural logarithm
L-NAME: L-NG-nitroarginine methyl ester
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L-NOARG: L-nitroarginine
LOEC lowest-observed-effect concentration
log logarithm
LPb low lead
LPS lipopolysaccharide
LSO lateral superior olive
M metal
M, mM, |xM, nM: Molar, milliMolar, microMolar,
nanoMolar
m, cm, mm, |im, nm, km: meter(s), centimeter(s),
millimeter(s), micrometer(s), nanometer(s),
kilometer(s)
MAP mean arterial pressure
MAPK mitogen-activated protein kinase(s), MAP
kinase
MATC maximum acceptable toxicant concentration
max maximum, maxima
MBP myelin basic protein
MCH mean corpuscular hemoglobin
MCHC mean corpuscular hemoglobin concentration
MchDMSA: mono-cyclohexyl dimercaptosuccinic acid
MCL maximum containment level
MCP-1 monocyte chemotactic protein-1
MCV mean corpuscular volume
MD mean diffusivity
MDA malondialdehyde
MDD major depressive disorder
MDI Mental Development Index
MDL method detection limit
MDRD Modification of Diet in Kidney Disease
Med, med: median
MEK1 dual specificity mitogen-activated protein
kinase 1
MEK2 dual specificity mitogen-activated protein
kinase 2
Mg magnesium
Mg2+ magnesium ion
MHC major histocompatibility complex
MI myocardial infarction, "heart attack";
myocardial ischemia
ml myoinositol
min minimum; minima; minute(s)
MKK1/2 MAPK kinase 1 and 2
ML mixed layer
MMAD mass median aerodynamic diameter
MMF mycophenolate mofetil
mmHg millimeters of mercury
mmol, |imol, nmol: millimole(s), micromole(s),
nanomole(s)
MN	micronuclei formation; mononuclear
Mn	manganese
MNE	micronucleated erythrocytes per thousand
Mn02	manganese dioxide
Mo	molybdenum
mo	month(s)
MOUDI	multi-orifice uniform deposit impactor
MPb, m-Pb: moderate lead
MPO	myeloperoxidase
MRI	magnetic resonance imaging
mRNA	messenger ribonucleic acid
MRS	magnetic resonance spectroscopy
MS	maternal stress
MSC	mesenchymal cell
MSWI	municipal solid waste incineration
Mt	metallothionein
MTHFR	methylenetetrahydrofolate reductase
MTP	mitochondrial transmembrane pore
MW	molecular weight
MZ	marginal zinc
N	nitrogen; normal; north; number; population
n	number of observations
Na	sodium
Na+	sodium ion
NAAQS	National Ambient Air Quality Standards
NAC	N-acetyl cysteine; nucleus accumbens
Na2CaEDTA: calcium disodium
ethylenediaminetetraacetic acid
NaCl sodium chloride
NAD nicotinamide adenine dinucleotide
NADH nicotinamide adenine dinucleotide
dehydrogenase
NADP nicotinamide adenine dinucleotide phosphate
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NADPH, NAD(P)H: reduced nicotinamide adenine
dinucleotide phosphate
NAEC no-adverse-effect concentration
NAG N-acetyl-P-D-glucosaminidase; N-
acetylglucosamine
NaHC03 sodium bicarbonate; sodium hydrogen
carbonate
NANC non-adrenergic non-cholinergic
NAS Normative Aging Study
NASCAR: National Association for Stock Car
Automobile Racing
NATTS National Air Toxics Trends Station
NAWQANational Water Quality Assessment
NCAM neural cell adhesion molecule
NCEA National Center for Environmental Assessment
NCore National Core multi-pollutant monitoring
network
N.D. not detected
NDMARN-nitrosodimethylamine receptor
NE norepinephrine
NEC AT New England Children's Amalgam Trial
NEI National Emissions Inventory
NFI non-fixed interval
NF-kB nuclear factor kappa B
NGAL neutrophil gelatinase-associated lipocalin
NGF nerve growth factor
NH non-hispanic
NHANES: National Health and Nutrition Examination
Survey
NH4CI ammonium chloride
NHEJ non-homologous end joining
NHEXAS: National Human Exposure Assessment
Survey
NH4OAc ammonium acetate
7-NI 7-nitroinidazole
Ni nickel
NICA non-ideal competitive absorption
NIOSH National Institute for Occupational Safety and
Health
NIST National Institute of Standards and Technology
NK natural killer
NKF-K/DOQI: National Kidney Foundation - Kidney
Disease Outcomes Quality Initiative
NMDA	N-methyl-D-aspartate
NMR	nuclear magnetic resonance
nNOS	neuronal nitric oxide synthase (NOS)
NO	nitric oxide; nitrogen monoxide
N02	nitrogen dioxide
No.	number
NOAA	National Oceanic and Atmospheric
Administration
NOAEL no observed adverse effect level
NOEC no-observed-effect concentration
NOEL no-observed-effect level
NOS nitric oxide synthase; nitric oxide systems
NOx nitrogen oxides, oxides of nitrogen
(NO + NOz)
NP nanoparticle
NPSH nonprotein sulfhydryl
NQOl NAD(P)H-quinone oxidoreductase (genotype)
NRC National Research Council
NRCS Natural Resouces Conservation Service
Nrf2 nuclear factor erythroid 2-related factor 2
NS not specified
NTPDase: nucleoside triphosphate diphosphohydrolase
NW northwest
NYC New York City
NZ New Zealand
02	molecular oxygen
02~ superoxide
03	ozone
9-0-Ac-GD3: 9-0-acetylated-GD3
OAQPS U.S. EPA Office of Air Quality Planning and
Standards, in OAR
OAR U. S. EPA Office of Air and Radiation
OB S ob servations
OC organic carbon
OEPA Ohio Environmental Protection Agency
OH" hydroxide ion
1,25-(OH)2D3: 1,25-dihydroxy vitamin D
O-horizon: horizon forest floor, organic soil horizon
(above surface soil)
OLC osteoblast-like cells
OM organic matter
ONL outer neuroblastic layers of the retina
ONOO" peroxynitrate ion
OR odds ratio
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ORD U.S. EPA Office of Research and Development
OS offspring stress
OSHA Occupational Safety and Health Administration
OVA ovalbumin
8-oxo-dG: 8-hydroxy-2' -deoxyguanosine
P	percentile; phosphorus
P0	parent generation
P450 cytochrome P450
p	probability value; number of paried hourly
observations; statistical significance
PAD peripheral arterial disease
PAH(s) poly cyclic aromatic hydrocarbon(s)
Pb lead
203Pb lead-203 radionuclide
204Pb stable isotope of lead-204
206Pb stable isotope of lead-206
207Pb stable isotope of lead-207
208Pb stable isotope of lead-208
210Pb stable isotope of lead-210
Pb++ divalent Pb ion
Pb° elemental lead
Pb(II) lead (II)
Pb2+ lead ion
Pb(Ac)2 lead acetate
PbB blood lead concentration
PbBrCl lead bromochloride
Pb(C2H302)2: lead (II) acetate
PbCl+ lead chloride
PbCl2 lead chloride
PbCl3 lead (III) chloride; lead trichloride
PbCl4 lead (IV) chloride; lead tetrachloride
PbC03 cerrusite; lead carbonate
Pb(C03)2lead (IV) carbonate
Pb(C03)2(0H)2: hydrocerussite
PbCr04 lead (II) chromate
PbD floor dust lead
PhFe6(S04)4(0H)12: plumbjarosite
PBG porphobilinogen
Pb(N03)2: lead(II) nitrate
Pb-NS lead-no stress
PbO lead oxide; litharge; massicot
Pb02 lead dioxide
Pb(IV)02: lead dioxide
Pb304 minimum or "red Pb"
Pb(OH)2 lead hydroxide
Pb5(P04)3Cl: pyromorphite
Pb5(P04)30H: hydroxypyromorphite
PbS galena; lead sulfide; soil lead concentration
PbSe lead selenide
PbS04 anglesite; lead sulfate
Pb4S04(C03)2(0H)3: macphersonite
PbxS lead by stress
Pb5(V04)3Cl: vanadinite
PC 12 pheochromocytoma 12 (adrenal / neuronal cell
line)
PCA principal component analysis
PCE polychromatic erythrocyte
PCR polymerase chain reaction
Pet percent
PCV packed cell volume
PD Parkinson's Disease
PDI Psychomotor Development Index
PEC probable effect concentration
PEL permissible exposure limit
PER partial exfiltration reactor
PG prostaglandin
PGE2, PGE2: prostaglandin E2
PGF2 prostaglandin F2
pH relative acidity; Log of the reciprocal of the
hydrogen ion concentration
PHA polyhydroxyalkanoates
PHE phenylalanine
PIH pregnancy induced hypertension
PIQ performance intelligence quotient (IQ)
PIR poverty-income ratio
PIXE particle induced X-Ray emission; proton-
induced x-ray emission
PKC protein kinase C
PLP proteolipid protein
PM particulate matter
PMX 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.
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PM10 In general terms, particulate matter with a
nominal aerodynamic diameter less than or
equal to 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.
PM2 5 In general terms, particulate matter with a
nominal aerodynamic diameter less than or
equal to 2.5 (im; a measurement of fine
particles in regulatory terms, particles with an
upper 50%o 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.
PM10-2.5 In general terms, particulate matter with a
nominal aerodynamic diameter less than or
equal to 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%o 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%o 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.
PM10c 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, collocated low-volume (16.7 Lpm)
PM10 and PM2.5 reference method samplers.
p38MAPK: p38 mitogen-activated protein kinase(s)
PMN polymorphonuclear leukocyte
P5N pyrimidine 5'-nucleotidase
PND post natal day
POC particulate organic carbon
PP polypropylene; pulse pressure
ppb	parts per billion
ppm	parts per million
PRP	post-reinforcement pause
PS	dam stress; prenatal stress; phosphatidylserine
PSA	prostate specific antigen
PSA-NCAM: polysialylated-neural cell adhesion
molecule
PT	proximal tubule
PTFE	polytetrafluoroethylene
PTHrP	parathyroid hormone-related protein
PUFA	polyunsaturated fatty acid
PVC	polyvinyl chloride
PVD	peripheral vascular disease
Q	quantile; quartile; quintile
QRS	QRS complex in ECG
QT	QT interval in ECG
QTc	corrected QT Interval
p	rho; bulk density; correlation
pS	Pearson's r correlation coefficient
R	net drainage loss out of soil depth of concern;
Spearman correlation coefficient; upward
resuspension flux; correlation
r	Pearson correlation coefficient
R2	multiple regression correlation coefficient
r2	correlation coefficient
RAAS	renin-angiotensin-aldosterone system
RAC2	gene encoding for Rac2
RBA	relative bioavailability
RBC	red blood cell
RBP	retinol binding protein
RD	radial diffusivity
Ref	reference (group)
RI-RI	concurrent random interval
RL	repeated learning
220Rn	radon isotope
222Rn	stable isotope of radon-222
RNA	ribonucleic acid
ROI	reactive oxygen intermediate/superoxide anion;
regions of interest
ROS	reactive oxygen species
RR	relative risk; risk ratio
RSL	roughness sublayer (transition layer, wake
layer, interfacial layer)
rtPCR	reverse transcription polymerase chain reaction
a	sigma, standard deviation
S	south; sulfur; synthesis phase
SAB	U. S. EPA Science Advisory Board
SATs	Standard Assessment Tests
SBP	systolic blood pressure
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SCE	sister chromatid exchange
Sena	a-synuclein
SD	standard deviation
SDN	sexually dimorphic nucleus
SE	standard error
Se	selenium
sec	second(s)
SEM	scanning electron microscopy; simultaneously
extracted metal; standard error of the mean
SES	socioeconomic status
Sess.	session
SGA	small for gestational age
sGC	soluble guanylate cyclase
sGC-pi	soluble guanylate cyclase-beta 1
SGOT	serum glutamic oxaloacetic transaminase
SGPT	serum glutamic pyruvic transaminase
SHBG	sex hormone binding globulin
SHM	Stockholm humic model
siRNA	small interfering RNA
SJW	silver jewelry workers
SLAMS	State and Local Air Monitoring Stations
SMC	smooth muscle cells
SNAP-25: synaptosomal-associated protein 25
SNARE	soluble NSF attachment receptor
SNP	single-nucleotide polymorphism; sodium
nitroprusside
SNS	sympathetic nervous system
SO	stratum oriens
S02	sulfur dioxide
So	south
SOC	superior olivary complex
SOD	superoxide dismutase
SOD1	superoxide dismutase-1
SOF	study of osteoporotic fractures
SOM	self-organizing map
SP	spray painters
SP1, Spl	specificity protein 1
SPM	suspended particulate matter
SPT	skin prick test
SREBP-2: sterol regulatory element binding protein-2
S. Rep.	Senate Report
SRIXE	synchrotron radation induced X-ray emission
StAR	steroidogenic acute regulatory protein
STAT	signal transducer and activator of transcription
STAT3 signal transducer and activator of
transcription 3
STAT5 signal transducer and activator of
transcription 5
STD. Standard
ST Interval: measured from the J point to the end of the
T wave in an ECG
Syb synaptobrevin
Syn synaptophysin
Syt synaptotagmin
SZn supplemental zinc
T, t time
T3, T3 triiodothyronine
T4, T4 thyroxine
11/2 half-life (-lives); time required to reduce the
initial concentration by 50%
TBARS thioBarbituric acid reactive substances;
thiobarbituric acid-reactive species
T cell, T-cell: T lymphocyte
TE trace elements
TEC threshold effect concentrations
TF ratio of the metal concentration in plant to that
in soil; transferrin
TFIIIA transcription factor IIIA
Tg transgenic
TGF transforming growth factor
TGF-P p transforming growth factor
TGFpi, TGF-pi: pi transforming growth factor
TH tyrosine hydroxylase
TH1, Thl: T-derived lymphocyte helper 1
TH2, Th2: T-derived lymphocyte helper 2
Th T-helper lymphocyte
TIMP-1 tissue inhibitor of metalloproteinases-1
TIMS thermal ionization mass spectrometry
TLC Treatment of Lead-exposed Children (study)
T/LH testosterone/luteinizing hormone - measure of
Ley dig cell function
TNF tumor necrosis factor (e.g., TNF-a)
TNP-Ficoll: trinitrophenyl-ficoll
TNP-OVA: trinitrophenyl-ovalabumin
TPR total peripheral vascular resistance
TS	transferrin saturation
TSH thyroid stimulating hormone; total sulfhydryl
TSP	total suspended particles
TSS	total suspended solids
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TXB2 thromboxane
UA	urbanized area
UBL	urban boundary layer
UCL	urban canopy layer
UDPGT uridine diphosphate (UDP)-
glucuronosyltransferase(s)
U.K.	United Kingdom
U.S.	United States of America
USC	U.S. Code
U.S. EPA: U.S. Environmental Protection Agency
USF	uptake slope factor
USGS	U.S. Geological Survey
USL	urban surface layer
UUDS	urban dynamic driving schedule
UV	ultraviolet radiation
V	vanadium
V79	Chinese hamster lung cell line
VA	Veterans Administration
VAChAT vesicular acetylcholine transporter
VAMP-2 vesicle-associated membrane protein-2
VA-NAS Veterans Administration Normative Aging
Study
VDAC voltage-dependent anion channel
VDR vitamin D receptor
VGAT vesicular gamma aminobutyric acid (GABA)
transporter
VGLUT1: vesicular glutamate transporter 1
VIQ verbal intelligence quotient (IQ)
VLPb very low lead
VMAT2 vesicular monoamine transporter-2
V043" vanadate ion
VOC(s) volatile organic compound(s)
vs., v. versus
VSMC vascular smooth muscle cells
WACAP Western Airborne Contaminants Assessment
Project
WBC	white blood cell
WCST	Wisconsin Card Sorting Test
WHAM	Windermere humic aqueous model
WHO	World Health Organization
WIAT	Wechsler Individual Achievement Test
WISC	Weschler Intelligence Scale for Children
WISC-R Weschler Intelligence Scale for Children-
Revised
wk week(s)
WML white matter lesions
WPPSI-III: Wechsler Preschool and Primary Scales of
Intelligence-Ill
WPPSI-R: Weschler Preschool and Primary Scale of
Intelligence-Revised
WRAT	Wide Range Achievement Test
W/S	winter/summer
WT	wild type
wt.	weight
XAFS	X-ray absorption fine structure
XANES X-ray absorption near edge structure
XDH	xanthine dehydrogenase
Xy observed hourly concentrations for time period
i at site j
Xik observed hourly concentrations for time period
i at site k
XPS X-ray photoelectron spectroscopy
XRF X-ray fluorescence
yr	year(s)
Zn zinc
Zn2+ zinc ion
ZPP zirconium-potassium perchlorate; zinc
protoporphyrin
Z-score standard score
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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; information for individual NAAQS reviews is
available at www. epa.gov/ttn/naaq s. 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 publications on health or welfare1 effects of the pollutant, or from atmospheric and
exposure sciences fields. 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 CASAC during the external
review process. References considered for inclusion in the ISA can be found at
http ://hero. epa.gov/: the references cited in the ISA include a hyperlink to each reference.
1 Welfare effects as defined in Clean Air Act (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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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.
Each ISA builds upon the conclusions of previous assessments for the criteria 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 older
studies may be discussed in detail to reinforce key concepts and conclusions or for
reinterpretation in light of newer data. Older studies also are the primary focus in some
areas of the document where research efforts have subsided, or if these older 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
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?
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¦	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?
Considerations specific to particular disciplines include the following. 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 includes 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-
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
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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 our understanding of
interspecies differences and potential differences between healthy and susceptible 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.
In developing an ISA, EPA reviews and summarizes the evidence from: studies of
atmospheric sciences and exposure; the health effects evidence from toxicological,
controlled human exposure and epidemiologic studies; and ecological and welfare effects
evidence. In the process of developing the first draft ISA, EPA may convene a public
workshop in which EPA and non-EPA experts review the scientific content of
preliminary draft materials to ensure that the ISA is up to date and 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 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. Using the causal
framework described in the following section, EPA scientists consider aspects such as
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strength, consistency, coherence, and biological plausibility of the evidence, and develop
draft 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 CASAC 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.
EPA Framework for Causal Determination
EPA has developed a consistent and transparent basis to evaluate 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 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 (2006b). U.S. EPA
(2005c). and Centers for Disease Control and Prevention (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.
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Evaluating Evidence for Inferring Causation
The 1964 Surgeon General's report (
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be published than reports of null findings, and publication bias can also result in
overestimation of effect estimate sizes floannidis. 2008). For example, effect estimates
from single-city epidemiologic studies have been found to be generally larger than those
from multicity studies (Bell et al.. 2005).
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 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, field) and numerous disciplines
(e.g., community ecology, biogeochemistry and paleological/historical reconstructions).
Direct evidence of a relationship between pollutant exposures and human health effects
comes from controlled human exposure studies. Controlled human exposure 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. 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 controls, effectively controlling for many potential
confounders. However, controlled human exposure studies are limited by a number of
factors, including small sample size and short exposure time. For example, exposure
patterns relevant to understanding real-world exposures, especially long-term exposures,
are generally not practical to replicate in a laboratory setting. In addition, although
subjects do serve as their own controls, personal exposure to pollutants in the hours and
days preceding the controlled exposures may vary significantly between and within
individuals. Finally, controlled human exposure studies require investigators to adhere to
stringent health criteria for subjects included in the study, 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
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cardiovascular disease, these individuals must also be relatively healthy and may not
represent the most sensitive individuals in the population. In addition, the study design is
limited to exposures and endpoints that are not expected to result in severe health
outcomes. Thus, not observing an effect in controlled human exposure studies does not
necessarily mean that a causal relationship does not exist. While controlled human
exposure studies provide important information on the biological plausibility of
associations observed in epidemiologic studies, observed effects in these studies may
underestimate the response in certain populations.
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. "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: independent effects on health outcomes;
effects of the pollutant acting as an indicator of a copollutant or a complex ambient air
pollution mixture; 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 this
potential threat to inferential validity. A confounder is associated with both the exposure
and the effect; for example, confounding can occur between correlated pollutants that are
associated with the same effect.
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Several statistical methods are available to detect and control for potential confounders,
with none of them being 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, more
than one pollutant may exert similar health effects, resulting in independently observed
associations for multiple pollutants. 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 which use a consistent method to analyze
data from across locations with different levels of covariates can provide insight on
potential confounding by copollutants. 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. "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" (Rothman and Greenland. 1998). 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 stratified analyses.
For example, the presence of a pre-existing disease or indicator of low socioeconomic
status may be an effect modifier in causing 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
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and different from the overall estimate, indicating a different exposure-response
relationship may exist in populations represented by these variables.
Another key consideration for ambient air pollution epidemiologic studies is exposure
measurement error. There are several components that contribute to exposure
measurement error in these 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.
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 or susceptible 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.
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 easier to detect. However, the control conditions
may limit the range of responses (e.g., animals may not be able to seek alternative food
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sources), 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.
Intermediate between laboratory and field are studies that use environmental media
collected from the field to examine response in the laboratory, and experiments that are
performed in the natural environment while controlling for some 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
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 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, exposure doses or
concentrations, 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 (1965) articulated aspects of causality in epidemiology and public health
that have been widely used (Samet and Bodurow. 2008; I ARC. 2006b; U.S. EPA. 2005c;
CDC. 2004). These aspects (Hill. 1965) have been modified (Table I) for use in causal
1 The "aspects" described by 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.
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determinations specific to health and welfare effects for pollutant exposures (U.S. EPA.
2009V1 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 are taken into
account with the goal of producing an objective appraisal of the 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.
1 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
Consistency of the	An inference of causality is strengthened when a pattern of elevated risks is observed
observed association 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.
Coherence	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, 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.
Biological plausibility. 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 gradient
(exposure-response
relationship)
Strength of the observed
association
Experimental evidence
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).
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.
Temporal relationship of Evidence of a temporal sequence between the introduction of an agent, and appearance
the observed association of the effect, constitutes another argument in favor of causality.
Specificity of the	Evidence linking a specific outcome to an exposure can provide a strong argument for
observed association 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.
Analogy	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.
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. IS As use a
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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
IOM's Improving the Presumptive Disability Decision-Making Process for Veterans
(Samet and Bodurow. 2008). EPA's Guidelines for Carcinogen Risk Assessment (U.S.
EPA. 2005c'). and the U.S. Surgeon General's smoking report (CDC. 2004). The Health
Consequences of Smoking: A Report of the Surgeion General. This weight of evidence
evaluation is based on 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 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 and then conclusions are drawn based upon the integration of evidence from
across disciplines and also across the spectrum of related endpoints. In making causal
judgments, the ISA focuses on major outcome categories (e.g., respiratory effects,
vegetation growth), by evaluating the coherence of evidence across a spectrum of related
endpoints (e.g., health effects ranging from inflammatory effects to respiratory mortality)
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 major outcome category.
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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Table II
Weight of evidence for 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 be ruled out with
Causal	reasonable confidence. For example: a) controlled
relationship 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 replicated and
consistent high-quality studies by multiple
investigators.
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.
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)
Likely to be a observational studies show an association, but
causal	copollutant exposures are difficult to address
relationship 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 replicated and high-quality
studies by multiple investigators.
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
Suggestive of epidemiologic study shows an association with a
a causal given health outcome but the results of other
relationship studies are 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.
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Health Effects
Ecological and Welfare Effects
Evidence is suggestive of no causal relationship
with relevant pollutant exposures. Several
Not likely to adequate studies, covering the full range of levels Several adequate studies, examining relationships
be a causal of exposure that human beings are known to with relevant exposures, are consistent in failing to
relationship encounter and considering at-risk populations, are show an effect at any level of exposure,
mutually consistent in not showing an effect at any
level of exposure.
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 lead 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 our 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
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.
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?
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¦	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)?
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 across epidemiologic studies that use various methods to
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 susceptible 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.
Of particular interest is the shape of the concentration-response curve at and below the
level of the current standards. 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
(e.g., particulate matter [PM], ozone [03], 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. These
attributes of human population dose-response relationships have been extensively
discussed in the broader epidemiologic literature (Rothman and Greenland. 1998).
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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. These factors
may be intrinsic, such as genetic or developmental factors, race, gender, lifestage, or the
presence of pre-existing 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,
rather than at the national or even regional scale. Quantitative relationships therefore are
available site by site and may differ greatly between ecosystems.
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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 become adverse; for example, what percentage
change in a lung function measure represents an adverse effect. What 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
susceptible individuals.
For example, 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? (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 and 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
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
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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 scale 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 (IPC'C. 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 (IPC'C. 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. 2002b). 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 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. 2002b). Additionally, the
National Acid Precipitation Assessment Program (NAPAP) 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."
On a broader scale, ecosystem services may provide indicators for ecological impacts.
Ecosystem services are the benefits that people obtain from ecosystems (UNEP. 2003).
According to the Millennium Ecosystem Assessment, ecosystem services include:
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1	"provisioning services such as food and water; regulating services such as regulation of
2	floods, drought, land degradation, and disease; supporting services such as soil formation
3	and nutrient cycling; and cultural services such as recreational, spiritual, religious and
4	other nonmaterial benefits." For example, a more subtle ecological effect of pollution
5	exposure may result in a clearly adverse impact on ecosystem services if it results in a
6	population decline in a species that is recreationally or culturally important.
7
8
References
9	ATS (American Thoracic Society). (2000). What constitutes an adverse health effect of air pollution? This official
10	statement of the American Thoracic Society was adopted by the ATS Board of Directors, July 1999. Am J
11	Respir Crit Care Med 161: 665-673.
12	Bell. ML: Dominici. F: Samet. JM. (2005). A meta-analysis of time-series studies of ozone and mortality with
13	comparison to the national morbidity, mortality, and air pollution study. Epidemiology 16: 436-445.
14	CDC (Centers for Disease Control and Prevention). (2004). The health consequences of smoking: A report of the
15	Surgeon General. Washington, DC. http://www.surgeongeneral.gov/librarv/smokingconseauences/.
16	Fox. GA. (1991). Practical causal inference for ecoepidemiologists. J Toxicol Environ Health A 33: 359-373.
17	http://dx.doi.org/10.1080/15287399109531535.
18	Gee. GC: Pavne-Sturges. DC. (2004). Environmental health disparities: A framework integrating psychosocial and
19	environmental concepts. Environ Health Perspect 112: 1645-1653. http://dx.doi.org/10.1289/ehp.7074.
20	HEW (U.S. Department of Health, Education and Welfare). (1964). Smoking and health: Report of the advisory
21	committee to the surgeon general of the public health service. Washington, DC.
22	Hill. AB. (1965). The environment and disease: Association or causation? Proc R Soc Med 58: 295-300.
23	IARC (International Agency for Research on Cancer). (2006b). Preamble to the IARC monographs. Lyon, France.
24	http://monographs.iarc.fr/ENG/Preamble/.
25	loannidis. JPA. (2008). Why most discovered true associations are inflated. Epidemiology 19: 640-648.
26	http://dx.doi.org/10.1097/EDE.0b013e31818131e7.
27	IPCC (Intergovernmental Panel on Climate Change). (2007). Summary for policymakers. In: Impacts, Adaptation
28	and Vulnerability. Contribution of Working Group II to the Fourth Assessment Report of the
29	Intergovernmental Panel on Climate Change. Cambridge, UK: IPCC.
30	Rothman. KJ: Greenland. S. (1998). Modern epidemiology (2nd ed.). Philadelphia, PA: Lippincott, Williams, &
31	Wilkins.
32	Samet. JM: Bodurow. CC (Eds.). (2008). Improving the presumptive disability decision-making process for
33	veterans. Washington, DC: National Academies Press.
34	http://www.nap.edu/openbook.php7record id=l 1908.
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1	U.S. EPA (U.S. Environmental Protection Agency). (1998). Guidelines for ecological risk assessment. (EPA/630/R-
2	95/002F). Washington, DC. http://www.epa.gov/raf/publications/guidelines-ecological-risk-
3	assessmciil.htm.
4	U.S. EPA (U.S. Environmental Protection Agency). (2002b). A framework for assessing and reporting on ecological
5	condition: An SAB report. Washington, DC: US EPA.
6	U.S. EPA (U.S. Environmental Protection Agency). (2005c). Guidelines for carcinogen risk assessment.
7	(EPA/630/P-03/00IF). Washington, DC. http://www.epa.gov/cancerguidelines/.
8	U.S. EPA (U.S. Environmental Protection Agency). (2009). Integrated science assessment for particulate matter.
9	(EPA/600/R-08/139F). Research Triangle Park, NC.
10	http ://cfpub. epa. gov/ncea/cfm/recordisplav. cfm?deid=216546.
11	UNEP (United Nations Environment Programme). (2003). Millennium Ecosystem Assessment: Ecosystems and
12	human well-being: A framework for assessment. Washington, DC.
13
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PREFACE
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)(1) 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 (Lead Industries 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, "... [a]ttainability
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]).
Section 109(d)(1) 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
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... ." 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
On October 5, 1978, EPA initially 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 (|ig/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
Air Quality Criteria for Lead (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 Air Quality Criteria for Lead (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 (U.S. EPA. 1990b). 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. 1990b). 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 and conduct of the NAAQS review were contained in two
documents: Project Work Plan for Revised Air Quality Criteria for Lead (U.S. EPA.
20056); and Plan for Review of the National Ambient Air Quality Standards for Lead
(U.S. EPA. 2006e). 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.
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The scientific assessment for the review is described in the 2006 Air Quality Criteria for
Lead [2006 Pb AQCD; (U.S. EPA. 2006b VI. 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. 2007s). 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. 2006f). a draft of which also
received CASAC and public review. The final Staff Paper presented OAQPS staff s
evaluation of the public health and welfare policy implications of the key studies and
scientific information contained in the 2006 Criteria Document 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.
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).1 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 |ig/m3 was
established to protect against air Pb-related health effects, including IQ loss, 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
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).
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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.1
Revisions to the NAAQS were accompanied by revisions to the data handling
procedures, the treatment of exceptional events, and the ambient air monitoring and
reporting requirements; as well as emissions inventory reporting requirements.2 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 3.4.
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, 2011d)l. 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. 2011c) 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
1	The 2008 NAAQS forPb are specified at 40 CFR 50.16.
2	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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1	draft materials prepared during the ongoing development of the Pb ISA. The first external
2	review draft ISA for Lead was released on May 6, 2011 (U.S. EPA 2011). The CASAC
3	Pb Review Panel met at a public meeting on July 20, 2011 to review the draft ISA
4	(76 FR 36120). Subsequently, on December 9, 2011, the CASAC panel provided a
5	consensus letter for their review to the Administrator of the EPA (U.S. EPA 2011). The
6	current document, the second external review draft ISA for Lead, will be discussed at a
7	public meeting of the CASAC Pb Review Panel, and timely public comments received
8	will be provided to the CASAC panel. A Federal Register notice will inform the public of
9	the exact date and time of that CASAC meeting.
References
10	ICF (ICF International). (2006). Lead human exposure and health risk assessments and ecological risk assessment
11	for selected areas: Pilot phase: External review draft technical report. Research Triangle Park, NC: ICF
12	International.
13	U.S. EPA (U.S. Environmental Protection Agency). (1977). Air quality criteria for lead [EPA Report]. (EPA-600/8-
14	77-017). Washington, DC. http://nepis.epa.gov/Exe/ZvPURL,cgi?Dockev=20013GWR.txt.
15	U.S. EPA (U.S. Environmental Protection Agency). (1986a). Air quality criteria for lead [EPA Report]. (EPA/600/8-
16	83/028aF-dF). Research Triangle Park, NC. http://cfpub.epa. gov/ncea/cfm/recordisplav.cfm?deid=32647.
17	U.S. EPA (U.S. Environmental Protection Agency). (1986c). Lead effects on cardiovascular function, early
18	development, and stature: An addendum to U.S. EPA Air Quality Criteria for Lead (1986). (EPA-600/8-
19	83/028aF). Washington, DC: US EPA.
20	U.S. EPA (U.S. Environmental Protection Agency). (1989). Review of the national ambient air quality standards for
21	lead: Exposure analysis methodology and validation: OAQPS staff report. (EPA-450/2-89-011). Research
22	Triangle Park, NC: US EPA.
23	U.S. EPA (U.S. Environmental Protection Agency). (1990a). Air quality criteria for lead: Supplement to the 1986
24	addendum [EPA Report]. (EPA/600/8-89/049F). Washington, DC.
25	U.S. EPA (U.S. Environmental Protection Agency). (1990b). Review of the national ambient air quality standards
26	for lead: Assessment of scientific and technical information: OAQPS staff paper. (EPA-450/2-89-022).
27	Research Triangle Park, NC: US EPA.
28	U.S. EPA (U.S. Environmental Protection Agency). (1991). Strategy for reducing lead exposures. Washington, DC:
29	US EPA. http://www.epa.gov/ttn/naaas/standards/pb/data/leadstrategyl991.pdf.
30	U.S. EPA (U.S. Environmental Protection Agency). (2005e). Project work plan for revised air quality criteria for
31	lead. (NCEA-R-1465). Research Triangle Park, NC.
32	http://cfpub.epa.gov/ncea/cfm/recordisplav. cfm?deid=l 13963.
33	U.S. EPA (U.S. Environmental Protection Agency). (2006b). Air quality criteria for lead: Volume I of II.
34	(EPA/600/R-05/144aF). Research Triangle Park, NC: US EPA.
35	http://cfpub.epa.gov/ncea/CFM/recordisplav.cfm?deid= 158823.
36	U.S. EPA (U.S. Environmental Protection Agency). (2006d). Analysis plan for human health and ecological risk
37	assessment for the review of the lead national ambient air quality standards (draft). Research Triangle Park,
38	NC. http://www.epa.gOv/ttn/naaas/standards/pb/s pb cr txl.html.
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U.S. EPA (U.S. Environmental Protection Agency). (2006e). Plan for review of the national ambient air quality
standards for lead. Research Triangle Park, NC.
http://www.epa.gOv/ttn/naaas/standards/pb/s pb cr pd.html.
U.S. EPA (U.S. Environmental Protection Agency). (2006f). Review of the national ambient air quality standards
for lead: Policy assessment of scientific and technical information: OAQPS staff paper - first draft. (EPA-
452/P-06-002). Research Triangle Park, NC: US EPA.
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-452/R-07-
014a). Research Triangle Park, NC: US EPA.
http://www.ntis. gov/search/product.aspx?ABBR=PB2008102573.
U.S. EPA (U.S. Environmental Protection Agency). (201 lc). Integrated review plan for the national ambient air
quality standards for lead. (EPA-452/R-11-008). Research Triangle Park, NC.
U.S. EPA (U.S. Environmental Protection Agency). (201 Id). Integrated review plan for the national ambient air
quality standards for lead: External review draft. (EPA-452/D-11-001). Research Triangle Park, NC.
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CHAPTER 1
EXECUTIVE SUMMARY
1.1 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) national ambient air quality standard
(NAAQS) for Lead (Pb). In 2008, the levels of the the primary and secondary NAAQS
for Pb were lowered ten-fold, from the 1978 level of 1.5 (ig/m3, to a level of 0.15 (ig/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 primary standard was based on the much-expanded health effects evidence on
neurocognitive effects of Pb in children. The revised standard was established to protect
against air Pb-related health effects, including IQ loss, in the most highly exposed
children.
EPA has developed a process for evaluating the scientific evidence and drawing
conclusions and causal judgments regarding air pollution-related health and
environmental effects. The ISA development 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. 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
The process and causality framework are described in more detail in the Preamble to the
ISA. Considerations that are specific to the causal determinations drawn for the health
and ecological effects of Pb are described in Section 2.1 of the document.
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1.2 Lead Sources, Fate and Transport in the Environment, Human
Exposure and Dose
Emissions of Pb to ambient air have 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 standards. Emissions were estimated to be 970 tons in 2008, with
more than half emitted from piston-engine aircraft. Other sources of ambient air Pb
(beginning with the largest) include metals processing, fossil fuel combustion, other
industrial sources, roadway related sources, and historic Pb. During the same period that
saw this dramatic decrease in Pb emissions, ambient air Pb concentrations have also
declined. The median annual concentrations in 2010 (0.03 |ig/m3) were approximately
thirty times lower than they were in 1980. The sharpest drop in Pb concentration occurred
during 1980-1990 and concentrations continued to decline through 2010. Concentrations
near Pb sources as well as concentrations away from Pb sources have shown a sharp
decline (Section 2.2.2).
The indicator for the Pb NAAQS is Pb in total suspended particles (Pb-TSP). The Pb-
TSP indicator was retained in 2008 in recognition of the role of all PM sizes in ambient
air Pb exposures. 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. The size distribution of
Pb-bearing particulate matter (PM) varies substantially depending on the nature of Pb
sources and proximity of the monitors to the Pb sources. Coarse Pb-bearing PM deposits
to a great extent near its source, contributing to local soil Pb concentrations, while fine
Pb-bearing PM (e.g. aircraft emissions) can be transported long distances, potentially
contributing Pb to remote areas. Depending on local conditions, deposited particles may
be resuspended and redeposited multiple times before further transport becomes unlikely.
In regulatory Pb monitoring, ambient Pb sampled using the FRM for Pb-PMi0 is allowed
in certain instances where the expected Pb concentration does not approach the NAAQS
and no sources of ultracoarse Pb are nearby.
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 the highest concentrations near Pb sources, such as metal smelters.
After the phase-out of Pb from on-road gasoline, Pb concentrations have decreased
considerably in rain, snowpack, and surface waters. In contrast, Pb is retained in soils and
sediments, where it provides 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 and
indicate peak concentrations between 1960 and 1980 (when leaded on-road gasoline was
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at peak use). 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.
Human exposure to Pb involves multiple pathways (Figure 1-1) and is difficult to assess
because Pb has multiple sources in the environment. 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. Non-air-related
exposures to humans 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. 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 shoe tracking and hand-to-
mouth contact, which is the main pathway of childhood exposure to Pb. Infiltration of Pb
dust into indoor environments has been suggested, and Pb dust has been shown to persist
in indoor environments even after repeated cleanings. Measurements of particle-bound
Pb exposures reported in this assessment have shown that personal exposure
measurements of Pb concentration are typically higher than are indoor or outdoor
ambient Pb concentrations.
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AIR
' Newly Emitted Pb
Historically Emitted Pb
DOOR SOIL
\AND DUST
NATURAL WATERS
AND SEDIMENTS
Non-air Pb
^Releases,
O. Paint
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epidemiologic studies are Pb in bone (a metric of cumulative exposure) and Pb cord
blood (a metric of prenatal exposure).
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 in an individual 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 whose current
exposure to Pb has effectively ceased or greatly decreased, a slow decline in blood Pb
concentrations mainly reflects the gradual release of Pb from bone. Generally, bone Pb is
an index of cumulative exposure and body burden. Pb is sequestered in two bone
compartments; with Pb in trabecular bone exchanging more rapidly with the blood, than
Pb in cortical bone. Therefore, Pb in cortical bone a better marker of cumulative exposure
and Pb in trabecular bone more likely to be correlated with blood Pb, in adults and
children. There is evidence for maternal-to-fetal transfer of Pb in humans. 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.
1.3 Integrative Overview of Health and Ecological Effects
There is substantial overlap between the ecological and health endpoints related to Pb
exposure in that they can be mediated through multiple, interconnected modes of action
(Table 2-5). The principal cellular/subcellular effects contributing to modes of action for
human health and ecological endpoints are altered ion status, protein binding, oxidative
stress, inflammation, endocrine disruption and cell death/genotoxicity (Figure 1-2). The
commonalities with regard to the endpoints and modes of action allowed the integration
of the evidence across disciplines (Table 2-4). The mechanisms of Pb toxicity are likely
conserved from invertebrates to vertebrates, including humans in some organ systems.
See Section 2.8.1 for additional details.
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Altered Ion
Status
(5.2,2)
Exposure
Protein
Binding
(5,2,3)
Cell Death
Genotoxicity
(5.2,?)
Endocrine
Disruption
(5.2,6)
Inflammation
{5.2.5)
Oxidative
Stress
(5.2,4)
Note: The subsections where these MOAs are discussed are indicated in parentheses.
Figure 1-2 Schematic representation of the relationships between the
various MOAs by which Pb exerts its effects.
1.3.1 Health Effects of Lead
1	Evidence from epidemiologic and toxicological studies was considered in combination
2	with the evidence from other disciplines such as exposure sciences and toxicokinetics in
3	determining the causal relationships for the health outcomes discussed in this assessment.
4	The evidence related to the specific endpoints comprising the weight of the evidence for
5	the organ system-specific causal determination are summarized in Table 1-1 below. Table
6	1-1 also specifies the sections sections in Chapter 2 and Chapter 5 where detailed
7	information can be found.
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Table 1-1 Summary of causal determinations between exposure to Pb and
health outcomes
Outcome3
Causality Determination
Nervous System Effects (Section 2.6.1)
Causal Relationship
Cardiovascular Effects (Section 2.6.2)
Causal Relationship
Renal Effects (Section 2.6.3)
Causal Relationship
Immune System Effects (Section 2.6.4)
Causal Relationship
Effects on Heme Synthesis and Red Blood Cell Function (Section 2.6.5)
Causal Relationship
Reproductive and Developmental Effects (Section 2.6.6)
Causal Relationship
Cancer (Section 2.6.7)
Likely Causal Relationship
aBased upon the framework described in the Preamble, a determination of causality was made for a broad outcome category (e.g., nervous system effects)
by evaluating the coherence of evidence across disciplines and across a spectrum of related endpoints. However, the evidence on which the causal
judgment is based, including the strength of evidence for the individual endpoints within the major outcome category, is characterized within the discussion.
Causal determinations were made within approximately 1-2 orders of magnitude of current levels.
Nervous System Effects
Recent epidemiologic and toxicological studies substantiated the strong body of evidence
presented in the 2006 Pb AQCD that Pb exposure is associated with nervous system
effects. The weight of evidence clearly supports associations of increases in blood Pb
levels with decrements in cognitive function in children, i.e., full-scale IQ and various
measures of learning and memory. Several studies find that there is a larger incremental
effect of Pb on cognition at lower blood Pb levels. Studies have not provided evidence of
a threshold for Pb-related impaired cognition. Epidemiologic and toxicological evidence
also clearly demonstrates Pb-associated increases in behavioral problems, in particular,
inattention and impulsivity in children. In epidemiologic studies, associations with
cognitive function and behavior were observed after adjustment for a range of potential
confounding variables, but most commonly, parental IQ, parental education, and other
SES-related variables. In children, the weight of evidence supports cognitive function
decrements and behavioral problems in association with concurrent blood Pb levels.
Associations also are observed with prenatal, early childhood, and childhood average
blood Pb levels, thus uncertainty remains regarding the lifestage of exposure within
childhood that is associated with the greatest risk. The weight of toxicological evidence
demonstrates neurodevelopmental effects with prenatal and early postnatal Pb exposures
that can have effects persisting to adulthood. The biological plausibility for
epidemiologic and toxicological findings for effects on cognitive function and behavior is
provided by evidence characterizing underlying mechanisms, including Pb-induced
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changes in neurogenesis, synaptogenesis and synaptic pruning, long term potentiation,
and neurotransmitter function. Based most heavily on cognitive function decrements and
inattention in children, the collective body of evidence integrated across epidemiologic
and toxicological studies is sufficient to conclude that there is a causal relationship
between Pb exposures and nervous system effects (Section 2.6.1).
Cardiovascular Effects
The 2006 Pb AQCD concluded that there was a relationship between blood Pb and bone
Pb and cardiovascular effects in adults, in particular increased blood pressure (BP) and
increased incidence of hypertension. Building on this strong body of evidence, recent
epidemiologic and toxicological studies substantiated the evidence that long-term Pb
exposure is associated with cardiovascular effects in adults with the largest body of
evidence demonstrating associations of Pb with increased BP and hypertension. The
epidemiologic evidence is strengthened by several recent prospective studies that find
associations between biomarkers of Pb and BP and hypertension and by effect estimates
that are observed after adjustment for multiple potential confounding factors. However,
there is uncertainty regarding the level, timing, frequency, and duration of Pb exposure
contributing to the observed associations since the populations studied were likely to
have higher past than recent Pb exposure. Animal toxicological studies provide
mechanistic evidence to support the biological plausibility of Pb-induced hypertension,
including Pb-induced oxidative stress, activation of renin-angiotensin aldosterone system
(RAAS), altered sympathetic activity, and vasomodulator imbalance. Collectively, the
evidence integrated across epidemiologic and toxicological studies as well as across the
spectrum of other cardiovascular endpoints examined is sufficient to conclude that there
is a causal relationship between Pb exposures and cardiovascular health effects
(Section 2.6.2).
Renal Effects
New epidemiologic and toxicological studies evaluated in the current review support or
expand upon the strong body of older evidence on the effect of Pb on the renal system
presented in the 2006 Pb AQCD. The weight of epidemiologic evidence consistently
demonstrates a relationship between higher blood Pb level and kidney dysfunction
(e.g., lower creatinine clearance, higher serum creatinine, and lower glomerular filtration
rate [GFR]) in nonoccupationally-exposed adults. Associations between Pb biomarker
levels and renal function are observed after adjustment for multiple potential confounding
factors such as age, sex, comorbid cardiovascular conditions, body mass index, smoking,
and alcohol use; however, there is uncertainty regarding the level, timing, frequency, and
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duration of Pb exposure contributing to the observed associations, since the populations
studied were likely to have higher past exposure than recent exposure. Coherence for
epidemiologic findings is provided by observations from animal models that Pb exposure
for longer than 6 months decreases GFR and increases serum creatinine. The weight of
evidence from animal studies indicates that Pb induces histopathological changes,
including tubular atrophy and sclerosis. By demonstrating Pb-induced renal oxidative
stress, inflammation, mitochondrial dysfunction, apoptosis, and glomerular hypertrophy,
toxicological studies provide biological plausibility for the associations observed in
epidemiologic studies between blood Pb levels and kidney dysfunction. Collectively, the
evidence integrated across epidemiologic and toxicological studies as well as across the
spectrum of renal outcomes is sufficient to conclude that there is a causal relationship
between Pb exposures and renal health effects (Section 2.6.3).
Immune System Effects
As described in the 2006 Pb AQCD, rather than producing overt cytotoxicity or
pathology, Pb exposure was found to be associated with alterations in several subclinical
parameters related to cellular and humoral immunity (Figure 5-42). The principal
findings are Pb-induced increased production of Th2 cytokines, suppressed production of
Thl cytokines, increased inflammation, and elevated IgE, with the weight of evidence
provided by toxicological studies. Collectively, these findings are coherent with the
observed effects of Pb exposure on decreasing responses to antigens (e.g., DTH, bacterial
resistance) in animals. Both toxicological studies in animals and epidemiologic studies in
children provide evidence for Pb-associated increases in IgE. The toxicological and
epidemiologic findings for the Pb-induced effects on Th2 cytokines, IgE, and
inflammation provide biological plausibility for the associations observed for blood Pb
levels with asthma and allergic conditions in children. Associations with asthma and
allergy were observed after considering potential confounding by several factors,
including, SES and allergen exposure. The blood Pb levels and Pb exposure lifestage,
magnitude, frequency, and duration associated with immune effects are not well
characterized in children or adults. Epidemiologic studies of children and adults primarily
examined concurrent blood Pb levels. The consistency and coherence of findings across
the continuum of related immune parameters that demonstrate a stimulation of Th2
responses in toxicological studies combined with the supporting epidemiologic evidence
in children are sufficient to conclude that there is a causal relationship between Pb
exposures and immune system effects (Section 2.6.4).
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Heme Synthesis and Red Blood Cell Function
Recent epidemiologic and toxicological studies support findings from the 2006 Pb
AQCD which concluded that Pb exposure was associated with effects on developing red
blood cells (RBCs) as well as heme synthesis in children and occupationally exposed
adults. The principal findings for RBC survival and function from recent studies are the
consistent Pb-induced alterations in several inter-connected hematological parameters
such as hemoglobin (Hb), hematocrit (Hct), and mean corpuscular volume (MCV)
observed across multiple studies. The effects on heme synthesis occurred through the
inhibition of multiple key enzymes, most notably aminolevulinic acid dehydratase
(ALAD), the enzyme that catalyzes the second, rate-limiting step in heme biosynthesis.
For heme synthesis, the largest body of evidence indicates decreased ALAD activity in
association with Pb exposure or blood Pb levels in occupationally-exposed adults or in
children (3-12 years old), respectively. Evidence for oxidative stress, an important mode
of action by which Pb may exert effects, primarily comes from studies of occupationally-
exposed adults and of children, the level, frequency, timing and duration of exposure
associated with altered RBC survival and function, heme synthesis, or the state of
oxidative stress in RBCs is uncertain in both adults and in children. The consistency of
findings in epidemiologic studies investigating effects in occupationally-exposed adults
and children, and the coherence of findings in the toxicological literature and coherence
across the disciplines is sufficient to conclude that there is a causal relationship exists
between Pb exposures and effects on heme synthesis and RBC function (Section 2.6.5).
Reproductive and Developmental Effects
Supporting conclusions from the 2006 Pb AQCD, recent toxicological and epidemiologic
literature provides strong evidence that Pb exposure is associated with effects on
reproduction and development, and expands evidence for additional endpoints. The
weight of the evidence supports the association of Pb exposure with delayed onset of
puberty in both males and females; and detrimental effects on sperm and semen quality in
occupationally-exposed males and in laboratory animals. In cross-sectional
epidemiologic studies of girls (ages 6-18 years) consistent associations with delayed
pubertal development (measured by age at menarche, pubic hair development, and breast
development) were observed. In boys (ages 8-15 years), fewer studies were conducted
but associations with delayed puberty were observed in most. There is uncertainty with
regard to the exposure frequency, timing, duration and level that contributed to these
observed association in studies of adolescents. The collective body of evidence integrated
across epidemiologic and toxicological studies (which examined Pb-induced detrimental
effects on sperm and on delayed onset of puberty), is sufficient to conclude that there is a
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causal relationship between Pb exposures and reproductive and developmental effects
(Section 2.6.6).
Cancer
The previous epidemiologic studies included in the 2006 Pb AQCD provided very limited
evidence suggestive of Pb exposure associations with cancers of the lung and stomach
and genotoxic effects in humans. The conclusions of the 2006 Pb AQCD were consistent
with those of International Agency for Research on Cancer (IARC) and the National
Toxicology Program. The animal toxicological literature continues to provide the
strongest evidence for an association between cancer and Pb exposure, with some
supporting evidence provided by the epidemiologic literature. Evidence from
toxicological studies demonstrates a relationship between Pb and cancer,
genotoxicity/clastogenicity or epigenetic modification. Carcinogenicity in previous
animal toxicology studies with 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.
Recent occupational studies of Pb exposure and lung cancer reported no associations. The
majority of epidemiologic studies of brain cancer had null results overall, but positive
associations between Pb exposure and brain cancer were observed among individuals
with certain genotypes (e.g. ALAD2). The collective body of evidence integrated across
toxicological and epidemiologic studies is sufficient to conclude that there is a likely
causal relationship between Pb exposure and cancer (Section 2.6.7).
1.3.2 Ecological Effects of Lead
The 2006 Pb AQCD and previous assessments found that the most commonly observed
effects in terrestrial organisms included decreased survival, reproduction and growth as
well as development and behavior. In aquatic invertebrates, effects of Pb exposures
included inhibition of ALAD, reduced reproduction, growth and survival. Evidence of
toxicity to fish (e.g. heme formation, alterations in brain receptors, effects on blood
chemistry and hormonal systems, and decreases in some enzyme activities) was also
observed. The effects of Pb can be observed in endpoints common to both terrestrial and
aquatic biota. These effects can be observed at multiple levels of biological organization,
starting at the cellular and subcellular level, then at the whole organism level, and finally
at the community and ecosystem level. Effects of Pb on physiological stress, blood, and
neurobehavior may increase susceptibility to other stressors and impact fitness of
individual organisms. Effects on development, reproduction, growth and survival may
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lead to effects on communities and ecosystems. The relationship of exposure and
responses is difficult to quantitatively characterize because of the influence of multiple
environmental variables on both Pb bioavailability and toxicity, and the substantial
species and lifestage differences in Pb sensitivity.
A brief discussion of the conclusions regarding Pb effects on physiological stress,
hematology, neurobehavior, development, reproduction, growth, survival and community
and ecosystem level effects is provided below and summarized in Table 1-2. Causal
determinations were based on consideration and integration of information on
biogeochemistry, bioavailability, biological effects, and exposure-response relationships
of Pb in terrestrial and aquatic environments. A detailed discussion for all relevant
welfare effects is provided in Chapters 2 and 7.
Table 1-2 Summary of Pb causal determinations for plants, invertebrates and
vertebrates
Effect3
Terrestrial
Aquaticb
Physiological Stress-All organisms (Section 2.7.3.1)
Causal
Causal
Hematological Effects-Invertebrates (Section 2.7.3.2)
Inadequate
Causal
Hematological Effects-Vertebrates (Section 2.7.3.2)
Causal
Causal
Neurobehavioral Effects-Invertebrates and Vertebrates (Section 2.7.3.3)
Likely Causal
Likely Causal
Developmental and Reproductive Effects-Plants (Section 2.7.3.4)
Inadequate
Inadequate
Developmental and Reproductive Effects-Invertebrates and Vertebrates (Section 2.7.3.4)
Causal
Causal
Growth-Plants (Section 2.7.3.5)
Causal
Causal
Growth-Invertebrates (Section 2.7.3.5)
Inadequate
Causal
Growth-Vertebrates (Section 2.7.3.5)
Inadequate
Inadequate
Survival-Plants (Section 2.7.3.6)
Inadequate
Inadequate
Survival- Invertebrates and Vertebrates (Section 2.7.3.6)
Causal
Causal
Community and Ecosystem Level Effects (Section 2.7.3.7)
Likely Causal
Likely Causal
aEffects observed at or near ambient levels of Pb were emphasized, and studies generally within 1 -2 orders of magnitude above current conditions
were considered in the body of evidence for terrestrial and aquatic ecosystems.
bCausal determinations for aquatic biota are based primarily on evidence from freshwater organisms.
Effects on physiological stress
Upregulation of antioxidant enzymes and increased lipid peroxidation 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, and of reduction in individual fitness. Markers of
oxidative damage and antioxidant activity have been observed in a wide range of species
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in terrestrial and aquatic environments when Pb is present, and following laboratory
exposures. Evidence is sufficient to conclude that there is a causal relationship between
Pb exposures and physiological stress in terrestrial and aquatic plants, invertebrates and
vertebrates (Section 2.7.3.1).
Hematological Effects
Changes in hematological variables including ALAD activity, blood cell counts and
serum profiles are associated with Pb exposure in both aquatic and terrestrial animal taxa.
Evidence from both field and laboratory studies suggest that ALAD is an indicator for Pb
exposure across a wide range of biota, and it is commonly recognized as such. Studies
conducted across the last two decades have indicated that hematological responses are
associated with Pb in the environment. Evidence is sufficient to conclude that there is a
causal relationship between Pb exposures and hematological effects in aquatic
invertebrates and vertebrates. Evidence is inadequate to establish a causal relationship
between Pb exposures and hematological effects in terrestrial invertebrates
(Section 2.7.3.2).
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
thermoregulation, and prey capture. Alterations to these behaviors may decrease the
overall fitness of the organism. Evidence from laboratory studies has shown effects of Pb
on nervous system endpoints in both aquatic and terrestrial animal taxa. Overall, the
evidence from terrestrial and aquatic systems is sufficient to conclude that there is a
likely causal relationship between Pb exposures and neurobehavioral effects in terrestrial
and aquatic invertebrates and vertebrates (Section 2.7.3.3).
Effects on Development and Reproduction
Various endpoints measured in multiple taxa of terrestrial and aquatic organisms have
documented effects of Pb on development, fecundity and hormone homeostasis. In plants,
few studies have addressed reproductive effects of Pb exposure. Decreased reproduction
at the organismal level can result in declining abundance and/or extirpation of
populations, decreased taxa richness, and decreased relative or absolute abundance at the
community level. Among the animal species tested, aquatic invertebrates were the most
sensitive to Pb with respect to reproduction. Effects of Pb on reproduction in
invertebrates and vertebrates indicate that Pb is affecting fecundity of Pb-exposed
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organisms in both aquatic and terrestrial habitats, and the evidence is sufficient to
conclude that there is a causal relationship between Pb exposures and reproductive effects
in terrestrial and aquatic invertebrates and vertebrates. The evidence is inadequate to
conclude that there is a causal relationship between Pb exposures and reproductive effects
in plants (Section 2.7.3.4).
Effects on Growth
Effects on growth observed at the species level may 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 plants, although they are typically observed in laboratory studies with high exposure
concentrations or in field studies near point sources. Evidence for Pb effects on growth in
invertebrates has been observed most extensively in freshwater aquatic taxa, with
inhibition in sensitive species occurring near the current range of Pb in surface waters. In
general, juvenile organisms are more sensitive than are adults. There are limited data on
growth effects in vertebrates. Evidence is sufficient to conclude that there is a causal
relationship between Pb exposures and growth effects in aquatic and terrestrial plants and
aquatic invertebrates. Evidence is inadequate to to conclude that there is a causal
relationship between Pb exposures and growth effects in terrestrial invertebrates and in
terrestrial and aquatic vertebrates (Section 2.7.3.5).
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 point 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 taxa, with survival
adversely impacted in a few species at concentrations occurring near point sources, or at
concentrations near common ambient levels. Terrestrial invertebrates typically tolerate
higher concentrations of Pb. Limited studies with vertebrates indicated adverse effects of
Pb on survival at concentrations typically higher than ambient Pb in the environment,
although juvenile organisms appear to be more sensitive than adults. The evidence is
inadequate to conclude that there is a causal relationship between Pb exposure and
survival in terrestrial and aquatic plants. Evidence is sufficient to conclude that there is a
causal relationship between Pb exposures and survival in terrestrial and aquatic
invertebrates and vertebrates (Section 2.7.3.6).
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Community and Ecosystem Effects
At higher levels of ecological organization, exposure to Pb may alter community and
ecosystem structure and function. Decreases in abundance, reduced species diversity and
shifts in community composition have been observed following Pb exposure in
laboratory and field experiments, but considerable uncertainties exist in generalizing to
natural ecosystem level effects. Assessment of Pb-specific exposure-response
relationships is difficult at the whole ecosystem level because potential confounders such
as other metals, physico-chemical variables, or various stressors cannot be controlled.
These factors make it difficult to quantify relationships between ambient concentrations
of Pb and ecosystem responses. The cumulative evidence that has been reported for
effects of Pb at the highest levels of ecological organization, and for species-level
endpoints with direct relevance to population and ecosystem level effects
(i.e., development and reproduction, growth, survival) is sufficient to conclude that there
is a likely causal relationship between Pb exposures and the alteration of species richness,
species composition and biodiversity in terrestrial and aquatic ecosystems
(Section 2.7.3.7).
1.4 Policy Relevant Considerations
Public Health Significance
The concept of population risk is relevant to the interpretation of findings for the
continuously-distributed subclinical health endpoints frequently studied in association
with Pb biomarkers in the assessment of their public health significance. A seemingly
small increase in the mean of a continuously distributed health index may push the most
susceptible group in the population above a critical cut point on the continuum of disease
development, such that their condition meets the clinical definition of a disease.
Moreover, small changes at the population level could translate into large numbers of
clinical events if a large population is affected. Pb-associated changes in subclinical
indices of disease may also increase an individual's risk for health effects that are of
greater clinical consequence, thus of greater public health concern. For example, a
downward shift in the mean IQ value can also reduce the proportion of the population
achieving very high IQ scores. Additional small increases in blood pressure or decreases
in renal function that are associated with Pb biomarkers, may shift the population mean
resulting in a larger proportion of the population that is diagnosed with hypertension or
chronic kidney disease, respectively.
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Air Lead-to-Blood Lead Relationships
A limited number of epidemiological studies evaluated relationships between air Pb and
blood Pb (Section 2.9.2). Much of the pertinent earlier literature described in the 1986 Pb
AQCD and more recent studies provide data from which estimates of the blood Pb-air Pb
slope can be derived for children. The range of estimates from these studies is
2 to 9 (ig/dL per (ig/m3. The differences in the estimates across studies may reflect model
selection by the investigators (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) as well as inclusion parameters
(e.g. soil Pb) that 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).
Ecological Effects and Corresponding Lead Concentrations
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 2.9.3). 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. In
addition, 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.
Concentration-Response Relationships for Health Effects
With each successive assessment to-date, the epidemiologic and toxicological study
findings show that progressively lower blood Pb levels or Pb exposures are associated
with cognitive deficits and behavioral impairments (Section 2.9.4). Compelling evidence
for a steeper slope for the relationship between blood Pb level and children's IQ at lower
blood Pb levels was presented in the 2006 Pb AQCD based on the international pooled
analysis of seven prospective cohort studies. A subsequent reanalysis of these data
focusing on the shape of the concentration-response function and several recent studies
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support these findings. The majority of the epidemiologic evidence from stratified
analyses comparing the lower and the higher ends of the blood Pb distributions indicates
larger negative slopes at lower blood Pb levels (Figure 2-2). Several lines of toxicological
evidence support the possibility of lower and higher Pb exposures acting through
differential activation of mechanisms underlying cognition. The shapes of concentration-
response relationships were examined in a more limited number epidemiologic studies of
other effects and findings were mixed (Section 2.9.4).
Pb Exposure and Nervous System Effects in Children
Young children, do not have lengthy exposure histories and consequently the
interpretation of associations with blood Pb levels for this age group may be less
complicated compared to older age groups. Several lines of evidence inform the
interpretation of study findings as they relate to aspects of exposure that can be attributed
to the cognitive and behavioral effects of Pb observed in young children (Section 2.9.5).
Epidemiologic studies find associations of cognitive function and behavior with prenatal,
early-childhood, lifetime average, and concurrent blood Pb levels as well as with
childhood tooth Pb levels. However, the weight of epidemiologic evidence supports
associations of concurrent blood Pb level. Studies of children up to three years of age that
found associations with concurrent blood Pb levels also tended to find associations with
prenatal cord or maternal blood Pb levels. Thus, both postnatal child and maternal Pb
exposures may contribute to lower cognitive function in young children. Exposures that
are reflected by concurrent blood Pb measured when children are older, by cumulative
blood Pb levels or by tooth Pb levels have also been demonstrated to be associated with
neurodevelopmental deficits throughout school-age and into adolescence. These findings
are consistent with the understanding that the nervous system continues to develop
throughout childhood.
Potentially At-Risk Populations
The NAAQS are intended to provide an adequate margin of safety for both the population
as a whole and those groups with unique factors that make them potentially at increased
risk for health effects in response to ambient air pollutant exposure. The most well-
substantiated at-risk population for the effects of Pb exposure is children. 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. Recent epidemiologic
studies of infants/children detected increased risk of Pb-related health effects, and this
was supported by toxicological studies. Synaptic pruning, which is active throughout
early childhood (ages 1-4 years), may underlie the elevated risk of neurodevelopmental
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effects in young children. Evidence is less consistent for other factors that may increase
Pb-related risk including sex, genetics, pre-existing disease (e.g., hypertension),
race/ethnicity, SES, nutrition, stress and co-exposure to other metals (Table 2-9).
1.5 Summary
Overall, the evidence evaluated for the current review expands upon findings of the 2006
Pb AQCD, which concluded that there was a strong body of evidence substantiating the
health effects of Pb and strong evidence of effects on some ecological endpoints. Table
2-10 of Chapter 2 summarizes the main conclusions from the 2006 Pb AQCD and
compares them to the findings from this assessment, including causality determinations,
regarding the health and ecological effects of Pb.
Nervous system effects, specifically cognition and behavior, in children are the effects
that are best substantiated as occurring at the lowest blood lead concentrations. Some new
epidemiologic and toxicological findings indicate that some effects (e.g., behavioral
impairments) are observed in populations of children with lower mean blood lead levels
than in previous assessments (e.g. <2 |_ig/dL). Causal relationships were also determined
between Pb exposure and several adult outcomes including renal and cardiovascular
effects, for which the evidence strongly suggests that cumulative exposure plays a role.
As lead exposures were generally higher in the past than they are today, uncertainties
exist, regarding the relative importance of recent versus past exposure in the development
of the lead-related health effects in the adult populations studied.
With regard to the ecological effects of Pb at or near concentrations currently present in
the environment, there is new evidence to support previous findings of effects on growth
and on reproduction and development in aquatic invertebrates as well as effects on
survival in both invertebrate and vertebrate aquatic biota. These effects are observed at or
near ambient levels of Pb in only a few species with greater toxicity generally associated
with early lifestages. Hematological and stress related responses were also associated
with elevated Pb levels in polluted areas in some terrestrial and aquatic species. In both
aquatic and terrestrial systems, uptake of Pb into flora and fauna 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 of biological organization.
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CHAPTER 2
INTEGRATIVE SUMMARY
2.1 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 NAAQS. The Integrated Review Plan (IRP) for the National
Ambient Air Quality Standards for Lead (U.S. EPA. 201 lc) identifies a series of
policy-relevant questions (in Chapter 3) that provide the framework for this assessment,
and which frame the entire review of the NAAQS for Pb, and thus are informed by both
science and policy considerations. The plans and underlying questions for the ISA are
included in the IRP. The ISA organizes, presents, and integrates the scientific evidence,
which is considered along with findings from any 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, 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 ecological effects of Pb in aquatic and
terrestrial ecosystems, discussion of endpoints common to plants, invertebrates
and vertebrates and consideration of uncertainties in relating atmospheric Pb
concentrations to ecological 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 developed a process for evaluating the scientific evidence; and drawing
conclusions and causal judgments regarding air pollution-related health and
environmental effects. The ISA development 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. The process and
causality 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.
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EPA initiated the current review of the NAAQS in April 2010 with a call for information
from the public ("75 FR 20843). Literature searches were conducted routinely to identify
studies published since the last review, focusing on studies published from 2006 (close of
previous scientific assessment) through September 2011. References that were considered
for inclusion or cited in this ISA can be found at http://hero.epa.gov/lead.
This ISA evaluates relevant epidemiologic, animal toxicological, and ecological 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. With regard to
the causal determinations drawn for human health effects of Pb, population-based
epidemiology studies were emphasized over occupational studies. Recent occupational
studies were considered insofar as they added to the evidence base for an outcome for
which sufficient numbers of population-based studies were unavailable or, to the extent
that 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).
Evidence from toxicological studies of effects observed in experimental animals at doses
that were relevant to, or somewhat above, those currently experienced by the U.S. general
population were emphasized. The extent to which studies of higher concentrations were
considered varied by major outcome category, but generally studies with blood Pb levels
within one order of magnitude above the upper end of the distribution of U.S. blood Pb
levels were considered (i.e., toxicological studies reporting blood Pb levels less than
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about 100 (.ig/dL).1 The majority of these studies reported blood Pb levels below
approximately 30-40 (ig/dL, however. Studies with higher blood Pb levels were
considered to the extent that they provided useful information to inform modes of action
or mechanisms or kinetics. For toxicological studies where blood Pb levels were not
measured, judgments regarding how to 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.
Relevant concentrations for drawing causality judgments for ecological 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 biota. Effects
observed at or near ambient levels of Pb were emphasized and studies generally within
one to two orders of magnitude above current conditions were considered in the body of
evidence for terrestrial and aquatic 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. For aquatic biota, generally, the number of studies available
on freshwater organisms is greater than on saltwater organisms, covering more taxa as
well as more endpoints. The causal determinations for aquatic biota are, therefore, based
primarily on evidence from freshwater species. When available, data from saltwater
ecosystems are discussed separately under the appropriate causal endpoint. For most of
the endpoints under consideration, evidence is inadequate to establish causality in
saltwater species. When sufficient evidence is available for marine organisms, data on
concentrations at which effects are observed are presented in Section 7.4.2.
The causal statements for terrestrial and aquatic effects are arranged according to
ecologically meaningful levels of biological organization (organism, population,
community, ecosystem). As recognized in EPA's Framework for Ecological Risk
Assessment (U.S. EPA. 1992). and in the adverse outcome pathway (AOP) framework
(Anklev et al.. 2010) endpoints that are measured at one level of ecological 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 survival and to developmental and
reproductive endpoints that can be used in ecological risk assessment, i.e., at the
population level and higher. Fecundity, growth, and survival are organism-level attributes
that lead to population-level (e.g., abundance, production, extirpation), community-level
(taxa richness, relative abundance) and ecosystem-level responses (Anklev et al. 2010;
1 Studies considered were generally within an order of magnitude above 10 |ig/dL (i.e., 100 |ig/dL). which was used
to indicate the upper end of the blood Pb distribution in the U.S. population. Currently approximately 1.4% of U.S.
children overall have blood Pb levels that exceed 10 ng/dL (Jones et al.. 2009a'): however, the proportion of
individuals with blood Pb levels that exceed this concentration varies depending on factors including age and/or
residence near a source of Pb exposure (Section 4.1).
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Suteret al.. 2005). In the case of Pb, physiological stress, hematological effects and
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. Biomarkers are typically used in epidemiologic studies of the health effects
associated with Pb as an index of exposure or dose. Consequently, the timing, frequency,
level, and duration of Pb exposure(s) associated with the observed health effects are
uncertain. Evidence from animal toxicological studies is also drawn upon to understand
and interpret Pb exposures in epidemiologic studies. Animal toxicological studies that are
relied upon to inform our understanding of the exposures needed to induce health effects
in humans include chronic exposure studies (i.e., over 10% of the lifespan of the animal),
long-term exposure studies (e.g., greater than 4 weeks in rodents) as well as acute or
short-term exposure studies (e.g., less than 4 weeks in rodents). In this ISA, short-term
human exposures are generally defined to include exposures of months (e.g., < one year)
while long-term human exposures include those greater than one year in duration. In
addition, information including the age of the population studied, study period and study
location can be used to aid in the interpretation of findings from epidemiologic studies
because Pb exposures have declined over time and exposures vary depending on
proximity to Pb sources.
This 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
Beyond judgments regarding causality are questions relevant to quantifying health or
environmental risks based on the understanding of the quantitative relationships between
pollutant exposures and health or ecological effects. Once a determination is made
regarding the causal relationship between the pollutant and outcome category, important
questions regarding quantitative relationships include:
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¦	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 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
whole, with several additional sections including: (Section 2.1) a discussion of the
assessment development and scope; (Section 2.8) an integration of the evidence across
the disciplines of health and ecology; (Section 2.9) a discussion of policy-relevant
considerations; and, (Section 2.10) an overall summary. This ISA itself is composed of
six chapters including this integrative summary. Chapter 3 highlights key concepts or
issues relevant to understanding the sources, ambient concentrations, and fate and
transport of Pb in the environment. 0 summarizes key concepts and recent findings on Pb
exposures, toxicokinetics, and biomarkers reflecting Pb exposure and body burden.
Chapter 5 presents a discussion of the MOA of Pb and evaluates and integrates
epidemiologic and animal toxicological information on health effects related to Pb
exposures, including nervous system, cardiovascular, renal, immunological, reproductive
and developmental, and cancer outcomes. Chapter 6 summarizes the evidence on
potentially susceptible populations. Chapter 7 evaluates ecological effects evidence that
is relevant to the review of the secondary NAAQS for Pb.
2.2 Ambient Lead: Source to Concentration
2.2.1 Sources, Fate and Transport of Ambient Lead
The findings of this review with respect to sources of atmospheric Pb build upon those
from the 2006 Pb AQCD (U.S. EPA. 2006b). which 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 76% from 1995
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to 2008, at which time emissions were estimated at 970 tons per year. As was the case at
the time of the last Pb NAAQS review, data from the 2008 National Emissions Inventory
(NEI) (U.S. EPA. 201 la) illustrate that piston-engine aircraft emissions comprise the
largest share (56%) of total atmospheric Pb emissions; the 2008 NEI estimated that 550
tons of Pb were emitted from this source. Other sources of ambient air Pb, beginning with
the largest, include metals processing, fossil fuel combustion, other industrial sources,
roadway related sources, and historic Pb. Global atmospheric Pb deposition peaked in the
1970s, followed by a decline. See Section 3.2 for additional information.
The atmosphere is the main environmental transport pathway for Pb. On a local scale
atmospheric Pb is primarily associated with coarse particulate matter (PM), and on a
global scale atmospheric Pb is primarily associated with fine PM. Both wet and dry
deposition are important removal mechanisms for atmospheric Pb. Because Pb in fine
particles is typically fairly soluble, wet deposition is more important for fine Pb. In
contrast, Pb associated with coarse particles is usually insoluble and 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 to remote areas. Depending on
local conditions, once deposited particles may be resuspended and redeposited before
reaching a site where further transport is unlikely, especially for dry deposition. See
Section 3.3 for additional information.
Environmental distribution of Pb occurs mainly through the atmosphere, from which it is
deposited into surface waters and soil. Surface waters act as an important reservoir, with
Pb lifetimes in the water column largely controlled by deposition and resuspension of Pb
in sediments. Substantial amounts of Pb from vehicle wear and building materials can
also be transported by runoff waters to surface waters and sediments without becoming
airborne. Pb containing sediment particles can be remobilized into the water column, and
sediment concentrations tend to follow those in overlying waters. See Section 3.3 for
additional information.
2.2.2 Monitoring and Concentrations of Ambient Air Lead
The indicator for the Pb NAAQS is Pb in total suspended particles (Pb-TSP). The
Pb-TSP indicator was retained in 2008 in recognition of the role of all PM sizes in
ambient air Pb exposures. 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 particle
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size with an uncertain upper size limit (Wedding et al. 1977). In recognition of the steep
spatial gradients associated with sources of ultra-coarse 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 ultra-coarse Pb are
nearby. A new FRM for Pb-PM10 has been implemented in which ambient air is drawn
through an inertial particle size separator for collection on a polytetrafluoroethylene
(PTFE) filter. The Pb-PM10 FRM is free of wind direction bias. Several FEMs have also
been approved. The Pb-TSP FRM is based on flame atomic absorption spectroscopy
(AAS), while the Pb-PM10 FRM is based on x-ray fluorescense spectroscopy (XRF).
Inductively-coupled plasma mass spectrometry (ICPMS) is under consideration as a new
FRM for Pb-TSP. See Section 3.4.1 for additional information.
Pb monitoring requirements have experienced several changes since publication of the
2006 Pb AQCD (U.S. EPA. 2006b). The current Pb monitoring network design
requirements include two types of FRM monitoring sites: source-oriented and
non-source-oriented. 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. At a minimum, monitoring is
required near sources that emit 0.50 tons/year or more of Pb unless it can be determined
that the source will not contribute to an ambient concentration exceeding 50% of the Pb
NAAQS. With the December 2010 completion of action on regulatory requirement of Pb
monitoring, one-year of Pb-TSP FRM monitoring is also required near 15 specific
airports to gather additional information on the likelihood of NAAQS exceedances near
airports due to combustion of leaded aviation gasoline. Non-source-oriented 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, including the different particle size ranges sampled in the different
networks. Depending on monitoring network, Pb is monitored in TSP, PMi0, or PM2 5.
For the purpose of analyzing data for the ISA, monitors reporting to the AQS were
considered to be source-oriented if they were designated in AQS as source-oriented, or
they were located within 1 mile of a 0.5 ton per year or greater source, as noted in the
2005 NEI (U.S. EPA. 2008a) or the 2008 NEI (U.S. EPA. 2011a). See Section 3.4 for
additional information.
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Ambient air Pb concentrations have declined drastically over the period 1980-2010. The
median annual concentrations have dropped by 97% from 0.87 |ig/m3 in 1980 to 0.03
(.ig/ni1 in 2010. While the sharpest drop in Pb concentration occurred during 1980-1990, a
declining trend was observed between 1990 and 2010. There was an 84% reduction in the
median annual source-oriented Pb concentration and a 85% reduction in the median
annual non-source-oriented Pb concentration for 1990-2010. Recent estimates for the
natural contribution to background Pb are -0.3 to 1 ng/m3. These estimates exceed
estimates of natural background presented in the 1986 Pb AQCD by a factor of 2 to 20.
The more recent estimate still indicates that background airborne Pb concentrations are
well below current ambient concentrations. See Section 3.5 for additional information.
AQS data for source-oriented and non-source-oriented FRM monitoring were analyzed
for 2008-2010. For source-oriented monitors, the three-month rolling average was above
the level of the NAAQS in twenty counties across the U.S. The three-month rolling
average was never above the level of the NAAQS for any of the non-source-oriented
FRM monitors. Pb concentrations, seasonal variations, inter-monitor correlations, and
wind data were analyzed for six counties: Los Angeles County, CA;
Hillsborough/Pinellas Counties, FL; Cook County, IL; Jefferson County, MO; Cuyahoga
County, OH; and Sullivan County, TN. Spatial and temporal variability of Pb
concentrations in each county were commonly high. Meteorology, 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. See Section 3.5 for additional
information.
Size distribution of Pb-bearing PM varied substantially depending on the nature of Pb
sources and proximity of the monitors to the Pb sources. Variation in the correlation of
size fractionated Pb samples among different land use types may be related to differences
in sources across land use types; for example, ultra-coarse Pb-PM has been observed near
industrial sites. Additionally, Pb concentrations exhibited varying degrees of association
with other criteria pollutant concentrations. Overall, Pb was moderately associated with
PM2 5, PMio and N02. Pb was moderately associated with CO in fall and winter only. The
poorest associations were observed between Pb and 03. Among trace metals, the
strongest association was with Zn. Moderate associations with Pb concentrations were
observed for Br, Cu, and K. Such correlations may suggest some common sources
affecting the concentrations of various pollutants. See Section 3.5 for additional
information.
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2.2.3
Ambient Lead 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.
After the 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. 2006^. In contrast, Pb is retained in soils
and sediments, where it provides 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 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. Although 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 emissions reductions from other sources.
Information on ambient Pb concentrations in non-air media and biota is reported in
Section 3.6, and concentrations considered in the interpretation of the ecological evidence
are tabulated in Table 2-1. 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 2-1 Ambient Pb Concentrations in Non-Air Media and Biota Considered
for Ecological Assessment
Media
Pb Concentration
Years Data Obtained
References
Soil
National Average: 18.9 mg Pb/kg
Range of state averages: 5-38.6 mg Pb/Kg
1961-1976
U.S. EPA (2007d. 2006b. 2003b)
Sediment1
Median: 28 mg Pb/kg (dry weight)
1996-2001
Mahler et al. 12006). U.S. EPA (2006b)
Fresh Surface V\Mer
Median: 0.5 pg Pb/L; Max: 30 pg Pb/L
2001-2003
U.S. EPA ('2006b')
(Dissolved Pb)a
Range: 0.0003-0.075 pg Pb/L
2002-2007
Field and Sherrell (2003). Landers et al. (2008)
Saltwaterb
Range: 0.01 - 27 pg Pb/L
Range: 0.005-0.04 pg Pb/L
Dates not available
Sadiq (1992)
Leland and Kuwabara (1985)
Vegetation
Lichens: 0.3-5 mg Pb/kg
2002-2007
Landers etal. (2008)

Grasses: 31% (percent of soil Pb in grass)
1980s-2000s
Vandenhove et al. (2009)
Vertebrate
Fish: Geometric Mean:
0.54 mg Pb/kg (dry weight)
Maximum:
23 mg Pb/kg (dry weight)
2001-2003
U.S. EPA (2006b)

Fish: 0.0033-0.97 mg Pb/kg
2002-2007
Landers et al. (2008)

Moose: 0.021-0.23 mg Pb/kg


"Based on synthesis of NAWQA data reported in 2006 Pb AQCD (U.S. EPA. 2006b)
bData 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.
2.3 Exposure to Ambient Lead
Exposure data considered in this assessment build upon the conclusions of the 2006 Pb
AQCD (2006b). which found air Pb concentrations in the U.S. and associated biomarkers
of exposure to have decreased substantially following the ban on Pb in on-road gasoline
as well as an earlier ban on Pb in house-hold paints. 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).
Non-air-related exposures include occupational exposures, hand-to-mouth contact with t
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 exposure to Pb-containing
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paint and home age (often a surrogate for the presence of Pb paint) are important
residential factors that increase risk of elevated blood Pb (Sections 2.9.6.7 and 6.2.6).
Most Pb biomarker studies do not indicate species or isotopic signature, and so non-air
exposures are reviewed in this section because they can also contribute to Pb body
burden. See Section 4.1 for additional information on exposure to ambient Pb.
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 3.4. Additionally, high spatial variability of the Pb concentrations in
various media also can contribute to exposure error, as described in the 2009 PM ISA
(U.S. EPA. 2009). 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; error in measurements and
assumptions thus have the potential to propagate through the exposure models.
Section 4.1 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 shoe tracking and hand-to-mouth contact,
which is the main pathway of childhood exposure to Pb. Recent data by Yamamoto et al.
(2006) have shown that the size distribution of particles collected on children's hands
have a mode around 40 |_un with the upper tail of the distribution extending to
200-300 (.un. Infiltration of Pb dust into indoor environments has been suggested, and Pb
dust has been shown to persist in indoor environments even after repeated cleanings.
Measurements of particle-bound Pb exposures reported in this assessment have shown
that personal exposure measurements of Pb concentration are typically higher than indoor
or outdoor ambient Pb concentrations. These findings regarding personal exposure to Pb
dust may be related to the personal cloud effect.
2.4 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%)
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may be the more biologically labile and toxicologically 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 (ig/day-kg) 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 4.2 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 and elimination in soft tissues is much faster than in bone. Pb
accumulates in bone regions undergoing the most active calcification at the time of
exposure. 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 andWittmers. 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'Flahcrtv.
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 deeper bone regions where it is relatively
inert, particularly in adults. These bone compartments are much more labile in infants
and children than in adults as reflected by half-times for movement of Pb from bone into
the plasma (e.g., cortical half-time = 0.23 years at birth, 3.7 years at 15 years of age, and
23 years in adults; trabecular half-time = 0.23 years at birth, 2.0 years at 15 years of age,
and 3.8 years in adults) (Leggett. 1993). See Section 4.2 for additional details.
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. See Section 4.2 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
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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 4.3 for additional details.
2.5 Lead Biomarkers
Overall, trends in blood Pb levels have been decreasing among U.S. children and adults
over the past 20 years (Section 4.4). The median blood Pb level for the entire U.S.
population is 1.2 (ig/dL and the 95th percentile blood Pb level was 3.7 (ig/dL, based on
the 2007-2008 NHANES data (NCHS. 2010). Among children aged 1-5 years, the
median and 95th percentiles were slightly higher at 1.4 (ig/dL and 4.1 (ig/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 making 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. See Section 4.3 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
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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
that is often not feasible in epidemiologic studies. Other biomarkers have been utilized to
a lesser extent (e.g., Pb in teeth). See Section 4.3.
Z6 Health Effects
This section evaluates the evidence from toxicological and epidemiologic studies that
examined the health effects associated with exposure to Pb and integrates that evidence
across these disciplines. The results from the health studies are also considered in
combination with the evidence from other disciplines (e.g., toxicokinetics) for the causal
determinations (Section 2.1) made for the health outcomes discussed in this assessment.
In the following sections a discussion of the causal determinations will be presented for
the health effects for which sufficient evidence was available to conclude a causal or
likely to be causal relationship (Table 2-2). A more detailed discussion of the underlying
evidence used to formulate each causal determination can be found in Chapter 5 of this
document.
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Table 2-2 Summary of causal determinations9 between exposure to Pb and
health outcomes
Outcome
Causality Determination3
Nervous System Effects (Section 5.3.13)
Causal Relationship
Cardiovascular Effects (Section 5.4.7)
Causal Relationship
Renal Effects (Section 5.5.6)
Causal Relationship
Immune System Effects (Section 5.6.8)
Causal Relationship
Effects on Heme Synthesis and Red Blood Cell Function (Section 5.7.7)
Causal Relationship
Reproductive and Developmental Effects (Section 5.8.10)
Causal Relationship
Cancer (Section 5.10.6)
Likely Causal Relationship
aBased upon the framework described in the Preamble, a determination of causality was made for a broad outcome category (i.e., nervous
system effects) by evaluating the coherence of evidence across disciplines and across a spectrum of related endpoints. However, the
evidence on which the causal judgment is based, including the strength of evidence for the individual endpoints within the major outcome
category, is characterized within the discussion. Causal determinations were made within approximately 1-2 orders of magnitude of
current levels.
Recent epidemiologic and toxicological studies substantiated the strong body of evidence
presented in the 2006 Pb AQCD that Pb exposure is associated with nervous system
effects. The weight of epidemiologic and toxicological evidence clearly supports
associations of higher blood Pb levels with decrements in cognitive function in children,
i.e., full-scale IQ and various measures of learning and memory. In epidemiologic
studies, these associations were substantiated in children ages 1 to 11 years and in
populations with mean blood Pb levels between 2 and 7 (ig/dL. Observation of a
supralinear concentration-response relationship and associations with mean (or quantile)
blood Pb levels < 2 (ig/dL do not provide evidence for a threshold for effects of Pb
exposure on the developing nervous systems of children. Epidemiologic and
toxicological evidence clearly demonstrates Pb-associated increases in behavioral
problems, in particular, inattention and impulsivity (prior symptoms of ADHD).
Associations are substantiated in children ages 1 to 12 years with mean concurrent blood
Pb levels of 2 to 5 (ig/dL. In epidemiologic studies, associations with cognitive function
and behavior were observed after adjustment for a range of potential confounding
variables, but most commonly, parental IQ, parental education, and other SES-related
variables.
2.6.1 Nervous System Effects
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In animals, the weight of evidence demonstrates effects on cognition and behavior with
prenatal and early postnatal Pb exposures that resulted in blood Pb levels of 10 to
40 (ig/dL. In children, the weight of evidence supports cognitive function decrements and
behavioral problems in association with concurrent blood Pb levels. Associations also are
observed with prenatal, early childhood, and childhood average blood Pb levels, thus
uncertainty remains regarding the lifestage of exposure within childhood that is
associated with the greatest risk. The uncertainty regarding the frequency, timing
duration and level of exposure contributing to the observed associations is greater in
studies of older children compared to younger children. The weight of toxicological
evidence demonstrates neurodevelopmental effects on cognition and behavior with
prenatal and early postnatal Pb exposures that can have effects persisting to adulthood.
The biological plausibility for epidemiologic and toxicological findings for effects on
cognitive function and behavior is provided by evidence characterizing underlying
mechanisms, including Pb-induced changes in neurogenesis, synaptogenesis and synaptic
pruning, long term potentiation, and neurotransmitter function. Based most heavily on
cognitive function decrements and inattention in children, the collective body of evidence
integrated across epidemiologic and toxicological studies is sufficient to conclude that
there is a causal relationship between Pb exposures and nervous system effects
(Section 5.3.13).
New evidence demonstrates associations of blood Pb level with Attention Deficit
Hyperactivity Disorder (ADHD) in children (8-17 years old) and this evidence is
consistent with findings demonstrating the effect of Pb on inattention and impulsivity in
children. Several new epidemiologic studies indicate associations between higher
concurrent blood Pb level and higher prevalence or incidence of ADHD diagnosis and its
contributing diagnostic indices, whereas previous evidence was inconsistent. The
biological plausibility for associations with ADHD is strongly supported by the large
epidemiologic and toxicological evidence base demonstrating Pb-associated increases in
inattention and impulsivity, both of which are primary symptoms of ADHD.
The evidence for Pb-associated effects on other nervous system endpoints was also
evaluated. Compared to the body of evidence on behavior and inattention, a smaller but
equally consistent evidence base indicated associations of concurrent and early childhood
blood Pb levels with social misconduct in children and delinquent behaviors in
adolescents and young adults (Table 5-11). Associations of blood Pb levels with ADHD,
misconduct, and delinquency were observed in populations of children with a wide range
of blood Pb levels, 1 to 11 (ig/dL, all similar in the strength of evidence. While the
different behavioral indices are examined, Pb exposure also was found to affect behavior
(decreased ability to escape predators or capture prey) in aquatic and terrestrial species
(Section 7.2 and Section 7.3).
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A few new toxicological studies augmented the evidence for Pb-related effects to the
visual system by demonstrating retinal changes in male rodent offspring in association
with lower blood Pb levels (<15 (ig/dL) than previously examined (20 to > 100 j^ig/dL)
(Section 5.3.4.3). A small body of epidemiologic evidence together with a large historical
base of toxicological evidence indicated associations of Pb biomarkers or exposure with
impaired auditory function. Associations were found in children ranging from 4 to
19 years in age and with mean concurrent blood Pb levels of 7-12 (ig/dL
(Section 5.3.4.1). While mood and emotional state have been examined less frequently
compared with inattention and misconduct, several studies found associations of
biomarkers of cumulative Pb exposure (i.e., tooth or childhood average blood Pb) and
concurrent blood Pb levels with parental or teacher reports of withdrawn behavior or
depression in children with mean blood Pb levels 8-28 (ig/dL (Section 5.3.3.3). These
findings in children are supported by a small body of toxicological studies in which
prenatal plus lactational Pb exposure resulted in depression-like behavior in rodent
models. Studies also reported associations of early childhood average and concurrent
blood Pb levels with lower fine and gross motor function in children ages 3 to 17 years
(Section 5.3.5). A common observation across studies was finding that biomarkers of Pb
exposure were associated with decrements in multiple neurodevelopmental outcomes,
including cognitive function, externalizing behaviors, internalizing behaviors, and motor
function, within the same population. These findings in combination with previously
discussed cognitive and behavioral effects indicate that Pb exposure affects a broad
spectrum of neurodevelopmental effects in children.
In adults, the frequency, duration, timing and level of Pb exposure implicated in nervous
system effects remain uncertain. Among occupationally-exposed adults, a spectrum of
nervous system effects is associated with concurrent blood Pb level (> 14 (ig/dL), which
reflects both current and cumulative exposure. However, in adults without occupational
exposure, cognitive performance is more strongly associated with tibia Pb levels than
blood Pb levels, which indicates an effect of long-term, cumulative Pb exposures. Based
on a smaller body of epidemiologic studies, blood and bone Pb levels were associated
with essential tremor and PD, respectively, in adults (Section 5.3.7.1). However, in these
case-control studies, it is difficult to establish temporality between Pb exposure and
disease. Support for epidemiologic findings for PD is provided by toxicological evidence
for Pb-induced decreased dopaminergic cell activity in the substantia nigra, which
contributes to the primary symptoms of Parkinson's disease. Whereas evidence for
association with Alzheimer's disease in adults is weak, developmental Pb exposures of
monkeys (early postnatal, PND 1-400) and rats (lactational) has been shown to induce
formation of amyloid plaques, pathology that underlies Alzheimer's disease
(Section 5.3.7.2). A small body of studies of behavior in adults examined and found
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associations of blood and tibia Pb levels with depression and anxiety symptoms
(Section 5.3.3.6).
2.6.2 Cardiovascular Effects
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 blood pressure (BP)
and increased incidence of hypertension (U.S. EPA. 2006^. Building on this strong body
of evidence, recent epidemiologic and toxicological studies substantiated the evidence
that long-term Pb exposure is associated with cardiovascular effects in adults with the
largest body of evidence demonstrating associations of Pb with increased BP and
hypertension. The epidemiologic evidence is strengthened by several recent prospective
studies that find associations between biomarkers of Pb and BP and hypertension and by
effect estimates that are observed after adjustment for multiple potential confounding
factors. The weight of epidemiologic evidence supported associations in adults with mean
concurrent blood Pb levels less than 5 (ig/dL. As these outcomes in epidemiologic studies
were most often observed in adults with likely higher past than current Pb exposures,
uncertainty exists as to the Pb exposure level, timing, frequency, and duration
contributing to the observed associations. Recent epidemiologic studies found that bone
Pb level, a metric of cumulative exposure, is strongly related to hypertension risk in
adults with mean bone Pb levels greater than 20 jj.g/g. However, uncertainties also exist
as to the specific Pb exposure conditions that contributed to the associations. The weight
of animal evidence also demonstrates an increase in BP after long-term (i.e., greater than
4 weeks in rodents) exposure to Pb. Whereas the majority of studies examined and found
increases in BP in animals with mean blood Pb levels greater than 10 (ig/dL (Table 5-18),
a recent study found elevated BP in animals with a mean blood Pb level of 2 (ig/dL (Tsao
et al. 2000). Also, animal toxicological studies provided mechanistic evidence to support
the biological plausibility of Pb-induced hypertension, including Pb-induced oxidative
stress, activation of RAAS (renin-angiotensin-aldosterone system), altered sympathetic
activity, and vasomodulator imbalance. Collectively, the evidence integrated across
epidemiologic and toxicological studies as well as across the spectrum of other
cardiovascular endpoints examined is sufficient to conclude that there is a causal
relationship between Pb exposures and cardiovascular health effects (Section 5.4.7).
Studies in the 2006 Pb AQCD (U.S. EPA. 2006^ also found associations between Pb
biomarkers of exposure and other cardiovascular diseases such as IHD, cerebrovascular
disease, peripheral vascular disease, and cardiovascular disease related mortality;
however, the available evidence was limited. Recent epidemiologic and toxicological
studies continue to provide evidence in adults for blood Pb-associated increased
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atherosclerosis, thrombosis, IHD, PAD, arrhythmia, and cardiac contractility in
populations with mean blood Pb levels >2.5 (ig/dL (Table 5-19). Further, animal
toxicological evidence continued to build on the evidence supporting the mechanisms
leading to these cardiovascular system responses, as well as Pb-induced changes in BP
and hypertension. Enhanced understanding of Pb-induced oxidative stress including NO
inactivation, endothelial dysfunction leading to altered vascular reactivity, activation of
the RAAS, and vasomodulator imbalance provides biological plausibility for the
consistent associations observed between higher blood and bone Pb levels and greater
cardiovascular effects.
Recent epidemiologic studies of adults also investigated the interaction of Pb biomarkers
with genetic variants in associations with cardiovascular effects. 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 (Figure 5-26).
Additionally, bone Pb concentration was associated with larger increases in PP, which is
a good predictor of cardiovascular morbidity and mortality and an indicator of arterial
stiffness, among adults with the HFE H63D and/or C282Y variant and there was
evidence of HFE and transferrin gene variants, related to iron metabolism, impacting the
associations of bone Pb levels with cardiovascular effects, evaluated by QT interval
changes (Figure 5-26). The evidence for genetic factors to potentially increase the risk of
Pb related health effects was limited (Table 6-4). New evidence extended the potential
continuum of Pb-related cardiovascular effects by demonstrating associations of blood Pb
and bone Pb with both cardiovascular and all-cause mortality with follow-up periods
ranging between 8 and 12 years. The associations of Pb with cardiovascular morbidity
observed in both epidemiologic and toxicological studies support recent epidemiologic
findings of increased Pb-associated cardiovascular mortality. Since the populations
enrolled in the epidemiologic studies were largely composed of adults with likely higher
past than current Pb exposures, uncertainty exists as to the Pb exposure level, timing,
frequency, and duration contributing to the observed associations.
2.6.3 Renal Effects
The 2006 Pb AQCD 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 follow-up of 4 to 15 years in association with higher baseline blood and bone Pb
levels (U.S. EPA. 2006^. Uncertainty remained on the contribution of past Pb exposures
to associations observed in adults, the renal effects of Pb in children, and the implication
of hyperfiltration. Because blood Pb level in nonoccupationally-exposed adults reflects
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both recent and past Pb exposures, the magnitude, timing, frequency, and duration of Pb
exposure contributing to the observed associations remains uncertain.
New epidemiologic and toxicological studies evaluated in the current review support or
expand upon the strong body of evidence presented in the 2006 Pb AQCD (U.S. EPA.
2006b). The weight of epidemiologic evidence demonstrates consistently a relationship
between higher blood Pb level and kidney dysfunction (e.g., lower creatinine clearance,
higher serum creatinine, and lower GFR [glomerular filtration rate]) in
nonoccupationally-exposed adults with mean concurrent or baseline blood Pb levels of
2-10 (ig/dL. A few analyses find higher blood Pb levels to be associated with a greater
longitudinal decrease in kidney function over time (4-15 years), suggesting that past Pb
exposures may contribute to ongoing renal effects. The epidemiologic evidence is
strengthened by associations between Pb biomarker levels and renal function that are
observed after adjustment for multiple potential confounding factors such as age, sex,
comorbid cardiovascular conditions, BMI (body mass index), smoking, and alcohol use.
Coherence for epidemiologic findings is provided by observations in animal models that
Pb exposure for greater than 6 months decreases GFR and increases serum creatinine.
The weight of evidence in animal studies indicates that Pb induces histopathological
changes, including tubular atrophy and sclerosis. Overall, reduced renal function and
increased kidney damage in animals are observed with chronic Pb (> 4 weeks) exposure
that results in blood Pb levels > 20 (ig/dL. By demonstrating Pb-induced renal oxidative
stress, inflammation, mitochondrial dysfunction, apoptosis, and glomerular hypertrophy,
toxicological studies provide biological plausibility for the associations observed in
epidemiologic studies between blood Pb levels and kidney dysfunction. Collectively, the
evidence integrated across epidemiologic and toxicological studies as well as across the
spectrum of renal outcomes is sufficient to conclude that there is a causal relationship
between Pb exposures and renal health effects (Section 5.5.6).
In addition, data on the effects of Pb on the kidney in children were reported in a recent
NHANES analysis of adolescents, ages 12-20 years, which observed an association
between higher concurrent blood Pb (mean: 1.5 (ig/dL) and lower cystatin C-based eGFR
(Section 5.5.6). These findings are consistent with results from a rodent model study in
which a low dose of Pb (50 ppm) administered from birth resulted in renal impairment
(elevated serum creatinine as compared to control rats), but these observations require
confirmation by measurement of GFR and renal pathology. This recent epidemiologic
study along with several previous studies that included children with higher Pb exposures
(due to residence near sources, Pb poisoning, or parental occupational exposure) provide
evidence that renal function in children may be affected by Pb exposure. The NHANES
adolescents, however, likely had higher Pb exposures earlier in childhood, thus, the
magnitude, timing, frequency, and duration of Pb exposure contributing to the observed
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association is uncertain. Research on the association of Pb with kidney function in the
occupational setting is less consistent than that in environmentally exposed populations
(Section 5.5.2.25.5.2.2). The observation 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 reflects limitations of the study design.
2.6.4 Immune System Effects
The 2006 Pb AQCD presented consistent evidence for immune system effects associated
with Pb exposure (U.S. EPA. 2006b). Rather than producing overt cytotoxicity or
pathology, Pb exposure was found to be associated with alterations in several subclinical
parameters related to cellular and humoral immunity (Figure 5-42). Recent toxicological
and epidemiologic studies support the strong body of evidence presented in the 2006 Pb
AQCD that Pb exposure may be associated with a broad spectrum of changes in both
cell-mediated and humoral immunity that cumulatively promote a Th2 phenotype and
hyperinflammatory state (U.S. EPA. 2006b). The principal findings are Pb-induced
increased production of Th2 cytokines, suppressed production of Thl cytokines,
increased inflammation, and elevated IgE, with the weight of evidence provided by
toxicological studies. Collectively, these findings are coherent with the observed effects
of Pb exposure on decreasing responses to antigens (e.g., DTH, bacterial resistance) in
animals. Both toxicological studies in animals and epidemiologic studies in children
provide evidence for Pb-associated increases in IgE. The toxicological and epidemiologic
findings for Th2 cytokines, IgE, and inflammation provide biological plausibility for the
associations observed for blood Pb levels with asthma and allergic conditions in children.
Associations with asthma and allergy were observed after considering potential
confounding by several factors, including, SES and allergen exposure. Animal studies
found a range of immune effects with prenatal exposure in juvenile animals and
long-term postnatal (> 4 weeks) Pb exposures in adult animals. The blood Pb levels and
Pb exposure lifestage, magnitude, frequency, and duration associated with immune
effects are not well characterized in children or adults. Epidemiologic studies of children
and adults primarily examined concurrent blood Pb levels. Little information was
provided on concentration-response functions. In epidemiologic studies, higher IgE and
higher asthma prevalence were examined and found in children with blood Pb levels >
10 (ig/dL. The consistency and coherence of findings across the continuum of related
immune parameters that demonstrate a stimulation of Th2 responses in toxicological
studies combined with the supporting epidemiologic evidence in children are sufficient to
conclude that there is a causal relationship between Pb exposures and immune system
effects (Section 5.6.8).
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The toxicological evidence for Pb-induced production of Th2 cytokines provides
biological plausibility for the evidence linking Pb exposure with elevated IgE levels
(Section 5.6.8). An increase in IL-4 from activated Th2 cells induces differentiation of B
cells into antibody-producing cells, thereby amplifying B cell expansion to secrete IgE,
IgA, and IgG. However, epidemiologic studies did not find blood Pb level to be
associated consistently with B cell abundance. In addition to T cell responses, a
prominent effect of Pb exposure, as demonstrated in an extensive historical toxicological
evidence base, was the induction of macrophages into a hyperinflammatory state as
characterized by enhanced production of ROS, suppressed production of NO, enhanced
production of TNF-a, and excessive metabolism of arachidonic acid into
immunosuppressive metabolites (e.g., PGE2). Consistent with these observations, a
previous epidemiologic study examined and found greater release of ROS and lower
release of NO from macrophages, primarily in children with concurrent blood Pb levels
10.31-47.49 (ig/dL (Pineda-Zavaleta et al. 2004). Misregulated inflammation represents
one of the major modes of action for Pb-induced immune effects. Toxicological studies
provide evidence for the modulation of inflammatory cell function, production of
pro-inflammatory cytokines and metabolites, enhanced inflammatory chemical
messengers, and pro-inflammatory signaling cascades. In addition to the associations
reported with IL-4, epidemiologic evidence for Pb effects on inflammation is limited to a
few recent studies in nonoccupationally-exposed adults in which concurrent blood Pb
level was associated with other indicators of inflammation such as CRP and IL-6
(Section 5.6.5.1). The studies commonly adjusted for potential confounding by age, sex,
BMI, and smoking status. However, because only concurrent blood Pb levels in adults
were examined, there is uncertainty regarding the magnitude, timing, frequency, and
duration of Pb exposures that contributed to the observed associations.
The effects of Pb exposure on macrophages also suggest a role for the immune system in
mediating Pb-associated effects in multiple other physiological systems. A small body of
new toxicological studies indicated Pb-induced changes in specialized macrophages in
nonlymphoid tissue such as alveolar macrophages, testicular macrophages, and brain
microglia (Section 5.6.4.5); however, these studies primarily used the i.p. route to
administer Pb. Thus, the relevancy of observations to those expected from typical routes
of human exposure is not clear.
In the large body of studies of adults (mostly males) with occupational Pb exposures
(Section 5.6.2.5), the most consistent findings were decreased neutrophil functionality in
workers with mean blood Pb levels 21-71 (ig/dL. Recent epidemiologic studies provided
new evidence in adults without occupational Pb exposures; however, each examined a
different immune endpoint, for example, IgE, eNO, IL-6. These endpoints were
associated with concurrent blood Pb levels in populations of adults with mean blood Pb
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levels of 1.9 (ig/dL to 7 (ig/dL; however, there is uncertainty regarding the contributions
of current Pb exposures and cumulative Pb stores in bone. The small body of
epidemiologic studies of nonoccupationally-exposed adults examined different endpoints
and found associations with concurrent blood Pb levels, which are influenced by current
Pb exposures as well as cumulative Pb stores in bone (Section 5.6).
2.6.5 Heme Synthesis and Red Blood Cell Function
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 blood Pb
>	40 (ig/dL. Pb-induced anemia is thought to occur due to decreased RBC life span and
effects on hemoglobin (Hb) synthesis. The exact mechanism for these effects was not
known, although Pb-induced changes on iron uptake or inhibition of enzymes in the heme
synthetic pathway may be responsible. Pb was also observed to exert effects on heme
synthesis through the inhibition of multiple key enzymes, most notably ALAD
(aminolevulinic acid dehydratase), the enzyme that catalyzes the second, rate-limiting
step in heme biosynthesis (Figure 5-45 presents a schematic representation of the heme
biosynthetic pathway). Decreased RBC ALAD activity was concluded to be the most
sensitive measure of human Pb exposure, in that measurement of ALAD activity is
correlated with blood Pb levels. Oxidative stress was identified as an important potential
mechanism of action by which Pb exposure induced effects on RBCs (U.S. EPA. 2006b').
Recent epidemiologic and toxicological studies support findings from the 2006 Pb
AQCD. The principal finding regarding RBC survival and function are consistent
Pb-induced alterations in several inter-connected hematological parameters such as Hb
(hemoglobin), Hct (hematocrit), and MCV (mean corpuscular volume) across multiple
studies, with the weight of evidence provided by epidemiological studies in
occupationally-exposed adult populations and children. In occupationally-exposed adults,
these findings are most substantiated in populations with current blood Pb levels
>	20 (ig/dL, although effects on hematological parameters were observed in some
occupationally-exposed populations at concurrent blood Pb levels in the range of
5-7 (ig/dL. The weight of evidence in adult rodents exposed long-term to Pb
(i.e., > 4 weeks) is coherent with epidemiologic studies regarding decrements in
hematological parameters at blood Pb levels as low as 6.6-7.1 |_ig/dL in rats and mice.
Regarding alterations in heme synthesis, the largest body of evidence again is provided
by decreased ALAD activity in association with Pb exposure or blood Pb levels in
occupationally-exposed adults or in children (3-12 years old), respectively. In the
occupationally-exposed adult populations, the observation of decreased ALAD activity
was most often observed in populations with concurrent blood Pb levels >15 (ig/dL.
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Animal toxicological studies also provide to the weight of evidence regarding altered
ALAD activity, with effects seen in adult animals exposed for 3-4 weeks with blood Pb
levels as low as 6.5 (ig/dL. The weight of evidence for oxidative stress (i.e., increased
lipid peroxidation or alterations in antioxidant enzyme levels) primarily comes from
epidemiological studies in occupationally-exposed adults and children. The majority of
evidence for increased oxidative stress in Pb-exposed adults comes from occupational
cohorts with concurrent blood Pb levels >15 (ig/dL. In children, concurrent blood Pb
levels of 7-22 (ig/dL were associated with measures of oxidative stress. The frequency,
timing and duration of exposure necessary to alter RBC survival and function, heme
synthesis, or the state of oxidative stress in RBCs is uncertain in both adults and children.
The consistency of findings in epidemiologic studies investigating effects in
occupationally-exposed adults and in children, and the coherence of findings in the
toxicological literature and coherence across the disciplines, is sufficient to conclude that
there is a causal relationship between Pb exposures and heme synthesis and RBC
function (Section 5.7.7).
2.6.6 Reproductive and Developmental Effects
Recent toxicological and epidemiologic literature provides strong evidence that Pb
exposure is associated with effects on reproduction and development supporting
conclusions of the 2006 Pb AQCD and expanding evidence for additional endpoints. The
weight of the evidence supports the association of Pb exposure with delayed onset of
puberty in both males and females and detrimental effects on sperm and semen quality in
occupationally-exposed males and in laboratory animals. In cross-sectional
epidemiologic studies of girls (ages 6-18 years) with mean and/or median concurrent
blood Pb levels less than 5 (ig/dL consistent associations with delayed pubertal
development (measured by age at menarche, pubic hair development, and breast
development) were observed. Toxicological studies of rodents indicate that prenatal and
lactational exposures to Pb can cause a delay in the onset of female puberty at blood Pb
levels as low as 8 (ig/dL. Recent studies show that pubertal onset is one of the more
sensitive markers of Pb exposure with effects observed after maternal exposures leading
to blood Pb levels in the pup of 3.5-13 (ig/dL. In boys (ages 8-15 years), fewer studies
were conducted but associations were observed in most. Male animal toxicology studies
have reported delayed sexual maturity as measured with prostate weight, among other
outcomes, seeing significant decrements at blood Pb levels of 34 (ig/dL. There is
uncertainty with regard to the exposure frequency, timing, duration and level that
contributed to these observed association in these studies of adolescents. Multiple studies
were performed in areas, which were contaminated with other chemicals as well as Pb.
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Additionally, Pb exposure has been shown to have detrimental effects on sperm. These
were observed in epidemiologic studies at population mean blood Pb levels of 30 j^ig/dL
and greater among men occupationally exposed (mean blood Pb levels in study controls
around 10 (ig/dL) and in animal toxicological studies with rabbits exposed to
subcutaneous Pb 3 times per week for 15 weeks with blood Pb levels of 20 (ig/dL. The
collective body of evidence integrated across epidemiologic and toxicological studies
with a focus on the strong relationship observed with detrimental effects on sperm and
delayed pubertal onset is sufficient to conclude that there is a causal relationship between
Pb exposures and reproductive and developmental effects (Section 5.8.10).
The evidence on hormonal influences and fecundity are less consistent (Section 5.8.1,
Section 5.8.2). Pb-mediated changes in levels or function of reproductive and growth
hormones have been demonstrated in past and more recent toxicological studies; however
the findings are inconsistent. Recent toxicological studies suggest that oxidative stress is
a major contributor to the toxic effects of Pb on male and female reproductive systems.
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 an association between increased generation of ROS and germ cell
injury as evidenced by destruction of 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.
Consistent conclusions in the 2006 Pb ISA recent studies of pregnancy outcomes were
reported mixed results. Inconsistent evidence of a relationship with Pb was available for
preterm birth and little evidence was available to study the associations with spontaneous
abortions (Section 5.8.6, Section 5.8.3). The 2006 Pb AQCD included a few studies that
reported potential associations between Pb and neural tube defects, but the recent
epidemiologic studies found no association. Some associations were observed between
Pb and low birth weight (Section 5.8.7) when epidemiologic studies used measures of
maternal bone Pb or air exposures, but the associations were less consistent when using
maternal blood Pb 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). Effects of Pb exposure during early development on toxicological studies
included reduction in litter size, implantation, birth weight and postnatal growth
(Section 5.8.4). Findings from epidemiologic studies of postnatal growth are inconsistent
(Section 5.8.8). Toxicological studies demonstrated that the effects ofPb exposure during
early development include impairment of retinal development and alterations in the
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developing hematopoietic and hepatic systems (Section 5.8.9). Negative developmental
outcomes were also noted including effects on the eyes and teeth (Section 5.8.9).
2.6.7 Cancer
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 classifications) based on
stronger evidence in animal studies than human studies, and organic Pb compounds as
not classifiable (Group 3 of IARC classifications) (IARC. 2006a; Rousseau et al. 2005).
Additionally, the National Toxicology Program has listed Pb and Pb compounds as
"reasonably anticipated to be human carcinogens" (NTP. 2004). The typical cancer
bioassays used 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.
The animal toxicological literature continues to provide the strongest evidence for Pb
exposure and cancer with some supporting evidence provided by the epidemiologic
literature. Evidence from toxicological studies demonstrates an association between Pb
and cancer, genotoxicity/clastogenicity or epigenetic modification. Carcinogenicity in
historical animal toxicology studies with 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
large 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, Pb exposure was found to be associated with stomach cancer, but there was only
one recent study on stomach cancer and Pb exposure, which reported mixed findings.
Similarly, some studies in the 2006 Pb AQCD reported associations between Pb exposure
and lung cancer (U.S. EPA. 2006b). More recent occupational studies of Pb exposure and
lung cancer reported no associations. The majority of epidemiologic studies of brain
cancer had null results overall, but positive associations between Pb exposure and brain
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cancer were observed among individuals with certain genotypes (e.g., ALAD2). In
toxicological studies, chromosomal aberrations after Pb exposure are most often reported
with Pb chromate exposure, which is likely due to toxicity of the chromate moiety.
Mechanistic understanding of Pb and its effect on cancer and genotoxicity is expanding
through toxicological work focusing on antioxidants and other proteins that sequester Pb
or reduce its bioavailability. The collective body of evidence integrated across
toxicological and epidemiologic studies is sufficient to conclude that there is a likely
causal relationship between Pb exposure and cancer (Section 5.10.6).
2.7 Ecological Effects of Lead
This section evaluates the evidence from studies of ecological effects associated with
exposure to Pb. Causal determinations are developed for the ecological outcomes
discussed in this assessment, in combination with evidence from other disciplines
(e.g., fate and transport) where relevant. Pb effects on terrestrial and aquatic systems
(Section 2.7.1 and Section 2.7.2) are summarized from Chapter 7 in which the effects on
terrestrial and aquatic ecosystems are presented separately. These sections are followed
by a summary of the evidence for the causal determinations (Section 2.7.3) and
consideration of atmospheric deposition of Pb as related to ecological effects
(Section 2.9.3). The ecological causal determinations are integrated across endpoints
(physiological stress, hematological effects, neurobehavioral effects, development and
reproduction, growth, survival) common to both terrestrial and aquatic biota (Table 2-4).
Where the causal determination varies substantially between types of organisms
(typically between plants and other organisms or between aquatic and terrestrial biota),
the divergence is noted.
2.7.1 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 7.4.2.3). 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). ALAD was recognized as a sensitive indicator of Pb
exposure in rats and waterfowl in the 1977 Pb AQCD and now is regarded as a biomarker
of exposure in many terrestrial organisms. According to the 2006 Pb AQCD and
supported by evidence from previous Pb AQCDs, commonly observed effects of Pb on
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terrestrial organisms include decreased survival, reproduction, and growth, as well as
effects on development and behavior. Furthermore, the toxicity of Pb to terrestrial
vertebrates and invertebrates varies with bioavailability of the metal, species and
lifestage, duration of exposure, form of Pb, and soil characteristics.
In plants, Pb effects have also been studied for several decades. 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). Furthermore, most of the Pb taken
up by plants from soil, remains in the roots and that distribution to other portions of the
plant is variable among species (U.S. EPA. 1977V 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). A study reviewed in the 1977 Pb AQCD provided evidence for Pb effects on
forest-nutrient cycling and shifts in arthropod community composition in the vicinity of a
smelting complex. A number of ecosystem-level effects, including decreased species
diversity, changes in floral and faunal community composition, and decreasing vigor of
terrestrial vegetation have subsequently been reported near Pb-point sources (U.S. EPA.
2006b. 1986b. 1977; Watson et al.. 1976V
Pb in terrestrial ecosystems is either deposited directly onto plant surfaces, or
incorporated into 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
variables: (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 details have been 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 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.
There is evidence over several decades of research previously reviewed in Pb AQCDs
and in recent studies reviewed in this ISA that Pb accumulates in terrestrial plants,
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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 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. Multiple new studies showed
that in trees, the latter is likely to be very substantial. One study attempted to quantify the
proportion of atmospheric Pb up directly from the atmosphere, and suggested it amounts
to 50% of the Pb contained in Scots Pine in Sweden (klaminder et al.. 2005). 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. Studies with trees found that soil Pb generally
is translocated to other parts.
Since the 2006 Pb AQCD, various species of terrestrial snails have been found to
accumulate Pb from both diet and soil (U.S. EPA. 2006b). New 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 between
species of earthworms that burrow in different soil layers. The rate of accumulation in
each of these species may result from differences in interacting factors such as pH and
CEC between layers. 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 point
sources.
There are a few new studies of Pb bioavailability and uptake in birds since the 2006 Pb
AQCD (U.S. EPA. 2006b). Several found tissue levels in birds that indicated exposure to
Pb, but none of the locations for these studies was in proximity to point 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 showed that the form of
Pb determined accumulation. New 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.
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
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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 low concentrations of Pb (2.5 (jM), 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.
Effects on amphibian 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. 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.
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New evidence reviewed in Section 7.2.3 and Section 7.2.4 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, a few studies of
exposure-response have been conducted with earthworms, and results have been
inconsistent (U.S. EPA. 2006b).
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 in the presence of Pb, attributed to decreased uptake of Pb by plants,
although both mycorrhizal fungus and plant were negatively affected. 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 recent studies have been conducted using mixtures of metals, but have
tried 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 a factor of 8, while avian
blood NOECs varied by a factor of 50 (Buekers et al. 2009). Protective effects of dietary
Ca have been found in plants, birds, and invertebrates.
2.7.2 Summary of Effects on Aquatic Ecosystems
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. 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 of fish from streams
(U.S. EPA. 1977). Observed responses of fish to Pb reported in the 1986 Pb AQCD and
the 2006 Pb AQCD include 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). In the 1986 Pb AQCD and 2006 Pb AQCD,
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additional evidence for Pb toxicity was available for aquatic invertebrates. Inhibition of
ALAD, and reduced reproduction, growth, and survival were reported. Few studies of Pb
toxicity to saltwater organisms are reported in previous Pb AQCDs. The toxicity of Pb to
saltwater and freshwater organisms varies with bioavailability of the metal, species and
lifestage, duration of exposure, form of Pb, and water quality parameters (U.S. EPA.
2006b. 1986a). As reviewed in the 2006 Pb AQCD, the biotic ligand model (BLM)
constitutes an attractive means of quantifying factors that affect bioavailability, such as
pH, dissolved organic carbon (DOC), and alkalinity (U.S. EPA. 2006b').
The toxicity of Pb to aquatic 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 reported and concentrations of Pb varied among genera
and within a genus, an observation subsequently found across aquatic taxa (U.S. EPA.
1977). At the time of the 1977 Pb AQCD, the information available on effects of Pb on
aquatic plants was limited. For plants in general, Pb was recognized to affect
photosynthesis, mitosis, and growth, but at concentrations higher than typically found in
the environment (U.S. EPA. 1977). Findings from field studies of aquatic communities in
the vicinity of Pb point sources 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.
Effects of Pb observed in aquatic organisms are tied to terrestrial systems via watershed
processes (Section 3.3). Atmospherically-derived Pb can enter aquatic systems through
runoff from terrestrial systems or via direct deposition over a water surface. Once Pb
enters surface waters, its fate and bioavailability are influenced by Ca2+ concentration,
pH, alkalinity, total suspended solids, and dissolved organic carbon (DOC, including
humic acids). In sediments, Pb bioavailability may be influenced by the presence of other
metals, sulfides, Fe and Mn oxides and physical disturbance. In many, but not all aquatic
organisms, Pb dissolved in the water can be the primary exposure route to gills or other
biotic ligands. As recognized in the 2006 Pb AQCD and further supported in this review,
chronic exposures to Pb may also include dietary uptake, and there is an increasing body
of evidence showing that differences in uptake and elimination of Pb vary with species.
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 conclusion that processes such as Pb
adsorption, complexation, and chelation alter bioavailability to aquatic biota. Given the
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low solubility of Pb in water, bioaccumulation by aquatic organisms may preferentially
occur via exposure routes other than direct absorption from the water column; which
include ingestion of contaminated food and water, uptake from sediment pore waters, or
incidental ingestion of sediment (U.S. EPA. 2006^.
There are considerable differences between species in both the amount of Pb taken up
from the environment and in the amount retained in the organism. Closely related species
can vary greatly in bioaccumulation of Pb and other non-essential metals. Recent studies
on uptake of Pb by aquatic plants and algae support the findings of previous Pb AQCDs
that all 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 in responses to Pb in water and sediments. In invertebrates, Pb can be
accumulated from multiple sources, including the water column, sediment, and dietary
exposure. Since the last review, new 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. 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 several species.
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 are supported in this review. The anterior intestine
has been newly identified as a site of uptake of Pb through dietary exposure studies.
There are few new studies on Pb uptake by amphibians and mammals. 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 (U.S. EPA. 2006b).
Evidence in this ISA further supports the findings of the previous Pb AQCDs that
waterborne Pb is highly toxic to aquatic organisms, with toxicity varying with species
and lifestage, duration of exposure, form of Pb, and water quality characteristics. Effects
of Pb on algae reported in the 2006 Pb AQCD are further supported by evidence from
additional species. They include decreased growth, deformation and disintegration of
cells, and blocking of the pathways that lead to pigment synthesis, thus affecting
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photosynthesis. Effects on plants supported by additional evidence in this review 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.
Since the 2006 Pb AQCD, there is additional evidence for Pb effects on antioxidant
enzymes, lipid peroxidation, stress response and osmoregulation in aquatic invertebrates.
Studies of reproductive and developmental effects of Pb in this review provide further
support for findings in the 2006 Pb AQCD. These new studies include reproductive
endpoints for rotifers and freshwater snails as well as multigenerational effects of Pb in
mosquito larvae and marine amphipods. As reviewed below, growth effects are observed
at lower concentrations in some aquatic invertebrates since the 2006 Pb AQCD,
especially in juvenile organisms. Behavioral effects of Pb in aquatic invertebrates
reviewed in this ISA include decreased valve closing speed in scallops and slower
feeding rate in blackworms.
Evidence in this ISA supports the findings of reproductive, behavioral, and growth effects
in previous Pb AQCDs, as well as effects on blood parameters in aquatic vertebrates.
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 tadpole endpoints including growth, deformity, and
swimming ability. 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.
Concentration-response data from plants, invertebrates and vertebrates are consistent with
findings in previous AQCDs of species differences in sensitivity to Pb in aquatic systems
(Section 7.3.17.4). In this ISA, as in previous AQCDs, aquatic plant growth was shown to
be adversely affected by Pb exposure. The lowest EC50 for growth observed in marine
microalgae and freshwater microalgae was in the range of 100 (.ig Pb/L. In the 2006 Pb
AQCD, concentrations at which effects were observed in aquatic invertebrates ranged
from 5 to 8,000 (.ig Pb/L. Several studies in this review have provided evidence of effects
at lower concentrations. Among the most sensitive species, growth of juvenile freshwater
snails (/.. stagnalis) was inhibited at an EC2o of <4 (.ig Pb/L (Grosell and Brix. 2009;
Grosell et al.. 2006a'). A chronic value of 10 (.ig Pb/L obtained in 28-day exposures of
2-month-old Lampsilis siliquoidea juveniles was the lowest genus mean chronic value
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ever reported for Pb (Wang et al.. 20 lOe). 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 (Esbaimh et al.. 2011).
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. However, adverse
effects in fish observed in recent studies fall within the range of concentrations observed
in the previous Pb AQCD.
Since the 2006 Pb AQCD, additional evidence for community and ecosystem level
effects of Pb have been observed primarily in microcosm studies or field studies near
point sources (mining, effluent) with other metals present. Ecological effects associated
with Pb, reported in previous Pb AQCDs, include alteration of predator-prey dynamics,
species richness, species composition, and biodiversity. New studies in this ISA 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 the species
sensitivity, lifestage, or seasonally-affected physiological state. Aquatic ecosystems with
low pH and low DOM are likely to be the most sensitive to the effects of
atmospherically-deposited Pb.
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2.7.3
Determinations of Causality for Effects on Ecosystems
Table 2-3 Summary of Pb causal determinations for plants, invertebrates and
vertebrates
Effect
Terrestrial
Aquatic9
Physiological Stress-All organisms
Causal
Causal
Hematological Effects-Invertebrates
Inadequate
Causal
Hematological Effects-Vertebrates
Causal
Causal
Neurobehavioral Effects-Invertebrates and Vertebrates
Likely Causal
Likely Causal
Developmental and Reproductive Effects-Plants
Inadequate
Inadequate
Developmental and Reproductive Effects-Invertebrates and Vertebrates
Causal
Causal
Growth-Plants
Causal
Causal
Growth-Invertebrates
Inadequate
Causal
Growth-Vertebrates
Inadequate
Inadequate
Survival-Plants
Inadequate
Inadequate
Survival- Invertebrates and Vertebrates
Causal
Causal
Community and Ecosystem Level Effects
Likely Causal
Likely Causal
"Causal determinations for aquatic biota are based primarily on evidence from freshwater organisms.
2.7.3.1 Effects on Physiological Stress
Evidence is sufficient to conclude that there is a causal relationship between Pb
exposures and physiological stress in terrestrial and aquatic plants, invertebrates and
vertebrates (Sections 7.4.1.1 and 7.4.2.1).
Endpoints associated with physiological stress received no consideration prior to the
2006 Pb AQCD and this ISA. Studies published since the 2006 Pb AQCD 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 and aquatic plants, invertebrates and vertebrates that support, and expand upon
findings in the previous Pb AQCD (U.S. EPA. 2006^.
In this ISA and the 2006 Pb AQCD, 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
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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 and limited transplantation
studies with aquatic plants indicate elevated antioxidant enzyme activity associated with
Pb levels measured in sediments at polluted sites. There is considerable evidence for
antioxidant activity in invertebrates, including gastropods, mussels, and crustaceans, in
response to Pb exposure. Some recent evidence for invertebrate antioxidant responses in
aquatic 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 aquatic 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. Glycogen levels in the
freshwater snail Biomphalaria glabrata were significantly decreased following 96-hour
exposures at near environmentally-relevant concentrations (50 (.ig Pb/L and higher)
(Ansaldo et al.. 2006).
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 aquatic organisms. Furthermore, these responses are
also observed in experimental animal studies, and in humans.
2.7.3.2 Hematological Effects
Based on observations in both terrestrial and aquatic 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 aquatic invertebrates and terrestrial and aquatic
vertebrates (Section 7.4.1.2, Section 7.4.2.2). The evidence is inadequate to conclude that
there is a causal relationship between Pb exposures and hematological effects in
terrestrial invertebrates.
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Recent studies add support to the strong body of evidence presented in previous Pb
AQCDs that Pb exposure is associated with hematological responses in aquatic
invertebrates and terrestrial and aquatic vertebrates. In environmental assessments of
metal-impacted habitats, ALAD is a recognized biomarker of Pb exposure (U.S. EPA.
2006b). ALAD activity is negatively correlated with total Pb concentration in bivalves.
Lower ALAD activity has been significantly correlated with elevated blood Pb levels in
fish and mammals as well. In the 1986 Pb AQCD, decreases in RBC ALAD activity
following Pb exposure were well documented in birds and mammals (U.S. EPA. 1986a).
Further evidence from the 2006 Pb AQCD and this review suggests 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
amphibians and fish, and has been identified in bacteria. New field studies of ALAD
activity include observations in songbirds and owls near historical mining areas. In
addition to consideration of ALAD activity, there is new 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 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 exposure to Pb and hematological effects in humans and laboratory animals
(Section 2.6.5 and Section 5.7).
2.7.3.3 Neurobehavioral Effects
Overall, the evidence from terrestrial and aquatic systems is sufficient to conclude that
there is a likely causal relationship between Pb exposures and neurobehavioral effects in
terrestrial and aquatic invertebrates and vertebrates (Section 7.4.1.3, Section 7.4.2.3).
Evidence from laboratory studies reviewed in Chapter 7 and previous Pb AQCDs have
shown adverse effects of Pb on neurological endpoints in both aquatic and terrestrial
animal taxa. Studies that consider mode-of-action and molecular targets of Pb toxicity in
biota are now available for a few species. New 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. New evidence in this ISA includes reports of behavioral
responses across a larger variety of organisms including reptiles and fish larvae born
from Pb-exposed adults, while some impairments in feeding and escaping behaviors were
reported for the first time. More evidence has become available for marine organisms.
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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 p3SVIAI>K as possible molecular targets for Pb neurotoxicity in
catfish (Leal et al. 2006V Additionally, there is new evidence for neurotoxic action of Pb
in invertebrates with exposure to Pb observed to cause changes in the morphology of
GABA motor neurons in nematodes (Caenorhabditis elegans) (Du and Wang. 2009V
Decreased food consumption of Pb-contaminated diet has been demonstrated in some
invertebrates (snails) and vertebrates (lizards, pigs, fish). Behavioral effects in grunt fish
Haemulon 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..
2011a). These fish were fed white shrimp exposed to Pb via brine shrimp that were
initially fed microalgae cultured 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). 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. Other behavioral effects affected
by Pb exposure include increased hyperactivity in fish and hypoxia-like behavior in
frogs.
These findings show strong coherence with findings from studies in laboratory animals
described in Section 2.6.1 and Section 5.3 of the ISA that show that Pb induces changes
in attention, increased response rates and motor function. The evidence presented in those
sections is sufficient to conclude that there is a causal relationship between Pb exposure
and neurobehavioral effects (Section 5.3).
2.7.3.4 Effects on Development and Reproduction
Evidence in this review and the previous Pb AQCDs from invertebrate and vertebrate
studies indicate that Pb is affecting reproductive performance in multiple species. Various
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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 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 (Suteret
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 and aquatic invertebrates and vertebrates (Section 7.4.1.4, Section 7.4.2.4). The
evidence is inadequate to conclude that there is a causal relationship between Pb
exposures and developmental and reproductive effects in plants (Section 7.4.2.4),
In terrestrial invertebrates, new developmental and reproductive endpoints 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; however, these
studies have been conducted at Pb concentrations that exceed Pb soil concentrations
found in most U.S. locations. In terrestrial vertebrates, new 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. Recent studies in freshwater rotifers and cladocerans provide
additional evidence for reproductive and developmental effects of Pb at concentrations at
or near ambient Pb levels in some aquatic species. New evidence in frogs and fish
continue to support developmental and reproductive effects of Pb in aquatic vertebrates
reported in earlier Pb AQCDs. Few studies are available that specifically address
reproductive effects of Pb exposure in 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
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 and at concentrations that typically exceed Pb levels
in U.S. soils.
For aquatic invertebrates, reproductive effects were reported to begin at 19 (ig Pb/L for
the freshwater snail Lymnea palustris and 27 (.ig Pb/L for Daphnia sp. as reported in the
1986 Pb AQCD (U.S. EP A. 1986b). Several new studies of snails, clams, and rotifers
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support previous findings of adverse impacts to embryonic development. 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 aquatic vertebrates there is evidence for reproductive and developmental effects of Pb.
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 fish, there is new evidence in this
ISA of Pb effects on steroid profiles. 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.
Additional observed impacts of Pb on reproductive endpoints in fish include decreased
oocyte diameter and density in toadfish, associated with elevated Pb levels in gonad.
In terrestrial vertebrates, evidence from Chapter 7 and in previous Pb AQCDs indicates
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 5), which is coherent with the partial evidence from mammals in the
wild. 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. New toxicological data
support trans-generational effects, a finding that is also an area of emerging interest in
ecology. The evidence presented in Section 5.8 is sufficient to conclude that there is a
causal relationship between Pb exposure and reproductive effects in humans and
laboratory animals.
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. Studies in this ISA include exposure-response
experiments showing exposure-dependent responses with exposure increasing from
background level to levels greater than near point sources. For some aquatic species,
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recent evidence supports previous findings of reproductive and developmental effects of
Pb and differential lifestage response at near ambient concentrations of Pb.
2.7.3.5 Effects on Growth
Evidence is sufficient to conclude that there is a causal relationship between Pb
exposures and effects on growth in aquatic and terrestrial plants and aquatic invertebrates
(Section 7.4.1.5, Section 7.4.2.5). Evidence is inadequate to establish a causal
relationship between Pb exposures and effects on growth in terrestrial invertebrates and
in terrestrial and aquatic vertebrates (Section 7.4.2.5).
Alterations in the growth of an organism can impact population, community and
ecosystem level variables. Evidence for effects of Pb on growth is strongest in aquatic
and terrestrial plants. In invertebrates, evidence for effects of Pb on growth has been
observed most extensively in aquatic taxa, with inhibition in sensitive species occurring
near the current range of Pb concentrations in surface waters. 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 new studies reporting effects at < 10 (.ig Pb/L, specifically in snails and
mussels. Also, growth effects in frogs are reported at lower concentrations in this ISA
than in earlier reviews.
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
laboratory and greenhouse toxicity studies have reported effects on plants. These effects
are typically observed in laboratory studies with high exposure concentrations or in field
studies near point sources. Pb has been shown to affect photosystem II with the
hypothesized mechanism being that Pb may replace either Mg or Ca in chlorophyll,
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 (.ig 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 Pb (U.S. EPA. 2008b) and the 2006 Pb AQCD. In the
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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
provide additional evidence for effects on growth of aquatic invertebrates at < 10 (ig
Pb/L. 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; Grose 11 et
al.. 2006a'). 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 (/.. siliquoidea) juveniles
were the most sensitive Lifestage (Wang et al.. 20106). 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 previous 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 new study reviewed in this ISA 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 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.
2.7.3.6 Effects on Survival
The evidence is sufficient to conclude that there is a causal relationship between Pb
exposures and survival in terrestrial and aquatic invertebrates and vertebrates
(Section 7.4.1.6, Section 7.4.2.6). The evidence is inadequate to conclude that there is a
causal relationship between Pb exposure and survival in terrestrial and aquatic plants
(Section 7.4.1.6, Section 7.4.2.6). In terrestrial vertebrates and invertebrates, evidence for
Pb effects on survival is primarily supported by previous Pb AQCDs with no new 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 measured in U.S. surface waters in cladocerans,
amphipods and rotifers. In aquatic vertebrates, there is new evidence for effects in fish at
<100 (ig Pb/L. Pb is generally not phytotoxic to aquatic or terrestrial plants at
concentrations found in the environment away from point sources, probably due to the
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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 Section 7.2.5,
Section 7.3.6, and Section 7.3.17.4 of this ISA (and in the 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 point-sources, 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 at concentrations occurring near
point-sources, or at concentrations near ambient levels. These impacted species may
include endangered species or candidates for the endangered species list, such as the
freshwater mussel Lampsilis rafmesqueana (Neosho mucket), The EC50 for foot
movement (a measure of viability) for newly transformed juveniles of this species was
188 (.ig Pb/L. Freshwater biota that exhibit sensitivity to Pb in the 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 VI. include some species of gastropods, amphipods, cladocerans, and
rotifers although the toxicity of Pb is highly dependent upon water quality variables such
as DOC, hardness, and pH. Other aquatic invertebrates such as odonates may be tolerant
of Pb concentrations that greatly exceed environmental levels.
Terrestrial invertebrates typically tolerate higher concentrations of Pb. In the 1986 Pb
AQCD it was reported that Pb at environmental concentrations occasionally found near
roadsides and smelters (10,000 to 40,000 |ag 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 OM 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 though a
simulated marine food chain in which the primary producer, the microalgae T. suecica,
was exposed to 20 |ag Pb/L and subsequently fed to brine shrimp Artemia franciscana,
(mean Pb content 12 to 15 |_ig Pb/g) which were consumed by white-leg shrimp
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Litopenaeus vannamei, itself consumed by grunt fish H. scudderi representing the top of
the marine food chain (Soto-Jimenez et al.. 201 la). 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 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-day) as reviewed for the
development of Eco-SSL's (U.S. EPA. 2005b'). The NOAELs for survival ranged from
3.5 to 3,200 mg Pb/kg • day.
In aquatic vertebrates there is considerable historic information on Pb toxicity to
freshwater fish. New studies support findings in previous AQCDs of Pb effects on fish
survival and indicate effects at lower concentrations in 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 (Esbaimh et al.. 2011). Thirty day
LC50 values for larval fathead minnows ranged from 39 to 1,903 |_ig Pb/L in varying
concentrations of DOC, CaS04 and pH (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 CaC03, a value much lower than in testing with older fish
(Mebane et al.. 2008).
2.7.3.7 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 ecological 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 there is a likely causal relationship
between Pb exposures and the alteration of species richness, species composition and
biodiversity in terrestrial and aquatic ecosystems (Section 7.4.1.7, Section 7.4.2.7).
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Ecosystem-level studies are complicated by the confounding of Pb exposure with other
factors such as the presence of other trace metals and acidic deposition. 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. At higher
levels of ecological organization, these effects may change species abundance and
community structure. Uptake of Pb into aquatic and terrestrial organisms and its effects
on survival, growth, physiological stress, blood, neurobehavior, and developmental 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 point
sources 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 aquatic systems, the literature focuses on evaluating ecological stress
from Pb originating from urban and mining effluents rather than atmospheric deposition.
In 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. For example, frogs and toads lack avoidance response while
snails and fish avoid higher concentrations of Pb. New 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 ecological 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. New 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.
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In aquatic ecosystems, Pb effects reviewed in the 2006 Pb AQCD 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. Exposure to three levels of sediment Pb
contamination in a microcosm experiment significantly reduced nematode diversity and
resulted in profound restructuring of the community structure (Mahmoudi et al. 2007).
Sediment-bound Pb contamination appears to differentially affect members of the benthic
invertebrate community, potentially altering ecosystems 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.
2.8 Integration of Health and Ecological Effects
The health and ecological effects considered for causal determination are summarized in
the Table 2-4. The health outcomes were nervous system, cardiovascular, renal, immune,
effects on heme synthesis and RBC function, reproductive effects, and cancer. The
ecological endpoints considered for causal determination were: physiological stress,
hematological effects, neurobehavioral effects, developmental and reproductive effects,
growth, survival, and community and ecosystem level effects. 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.
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Table 2-4 Summary of Causal Determinations9 for Health and Ecological
Effects
Outcome/Effect
Human Health
Causal Determination
Ecological Receptors
Causal Determination
Neurological Effects'5
Causal Relationship
Likely Causal Relationship: Invertebrates and Vertebrates
Cardiovascular Effects
Causal Relationship
N/Ae
Renal Effects
Causal Relationship
N/Ae
Immune System Effects
Causal Relationship
N/Ae
Heme Synthesis and
RBC Function0
Causal Relationship
Causal Relationship: Terrestrial Vertebrates,
Aquatic Invertebrates and Vertebrates
Reproductive Effects and
Birth Outcomed
Causal Relationship
Causal Relationship: Invertebrates and Vertebrates
Inadequate to Infer Causal Relationship: Plants
Cancer
Likely to be a causal relationship
N/Ae
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
cardiovascular effects.)
Causal Relationship: Invertebrates and Vertebrates
Inadequate to Infer Causal Relationship: Plants
Growth
N/Ae (There is evidence from toxicological and
epidemiologic studies of Pb effects on fetal and
postnatal growth, which was considered in
determining the causal association between Pb
exposure and reproductive and developmental
effects.)
Causal Relationship: Plants and Aquatic Invertebrates
Inadequate to Infer Causal Relationship: Vertebrates and
Terrestrial Invertebrates
Physiological Stress
N/Ae (In Human Health, oxidative stress was
considered as a 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
Community and
Ecosystem Level Effects
N/Ae
Likely Causal Relationship
aBased upon the framework described in the Preamble, a determination of causality for health effects was made for a broad outcome category (i.e., nervous
system effects) by evaluating the coherence of evidence across disciplines and across a spectrum of related endpoints. However, the evidence on which
the causal judgment is based, including the strength of evidence for the individual endpoints within the major outcome category, is characterized within the
discussion. Causal determinations were made within approximately 1-2 orders of magnitude of current levels.
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 human epidemiologic evidence evaluated included a wider range of
health endpoints such as cognition.
cThe health hematological effects considered in the determination of causality were primarily heme synthesis and RBC function. The ecological evidence
considered for the causal determination included heme synthesis, blood cell count, and altered serum profiles.
d Reproductive health effects, including effects on sperm, as well as birth outcomes such as spontaneous abortion, were considered in the causal
determination. 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., for health effects no causal determination for this specific endpoint, considered within the causal determination for the larger
outcome category; for ecological effects there was no comparable endpoint.
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2.8.1 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 2.7) 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 5.2 has the potential to contribute to the
development of a number of Pb-induced health effects (Table 2-5). 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
2-5) 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.
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Table 2-5 MOAs, their related health effects, and information on concentrations
eliciting the MOAs
Mode of Action
[Related Health Effects
(ISA Section)]
Concentrations or doses (Conditions)"
Blood Pb
Dose
Altered Ion Status
[All Health Effects of Pb]
3.5 pg/dL
(mean in cord blood; association with cord blood Ca2t-
ATPase pump activity)
Huel et al. (2008)
0.00005 pM free Pb2t
(In vitro; 30 minutes; calmodulin activation assay)
Kern et al. (2000)
Protein Binding
[Renal (5.5), Heme Synthesis and RBC
Function (5.7)]
17.0 pg/dL
(concurrent mean in adult workers with wildtype
metallothionein expression; increased BP susceptibility)
Chen et al. (2010a)
50 pM Pb glutamate
(In vitro; 24 hours; increased nuclear protein in
neurological cell)
Klann and Shelton (1989)
Oxidative Stress
[All Heath Effects of Pb]
5.4 pg/dL
(concurrent mean in adult male workers; decreased CAT
activity in blood)
Conterato et al. (In Press)
0.1 pM Pb-acetate
(In vitro; 48 hours; decreased cellular GSH in
neuroblastoma cells)
Chetty et al. (2005)
Inflammation
[Nervous System (5.3), Cardiovascular
(5.4), Renal (5.5), Immune (5.6),
Respiratory (5.6.4), Hepatic (5.9.1)]
2.5 pg/dL
(concurrent minimum in adult males; increased serum
TNF-a and blood WBC count)
Kim et al. (2007)
0.01 pM Pb-acetate
(In vitro; 48 hours; increased cellular PGE2 in
neuroblastoma cells)
Chetty et al. (2005)
Endocrine Disruption
[Reproductive and Developmental Effects
(5.8), Endocrine System (5.9.3), Bone
and Teeth (5.9.4)]
1.7 pg/dL
(concurrent minimum in women with both ovaries
removed; increased serum FSH)
Krieg (2007)
10 pMPb nitrate
(In vitro; 30 minutes; displaced GHRH binding to rat
pituitary receptors)
Lau et al. (1991)
Cell Death/Genotoxicity
[Cancer (5.10), Reproductive and
Developmental Effects (5.8), Bone and
Teeth (5.9.4)]
3.3 pg/dL
(concurrent median in adult women; increased rate of
HPRT mutation frequency)
Van et al. (2004)
0.03 pM Pb-acetate
(In vitro; 18 hours; increased formation of micronuclei)
Bonacker et al. (2005)
a 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.
1	Ecosystem studies have presented evidence for the occurrence of many of these modes of
2	action in animals, and to some degree in plants, however the connection to ecological
3	outcomes must usually be inferred because ecological studies are typically not designed
4	to address mode of action directly. The level at which Pb elicits a specific effect is more
5	difficult to establish in terrestrial and aquatic systems due to the influence of
6	environmental variables on Pb bioavailability and toxicity and substantial species
7	differences in Pb susceptibility.
8	The alteration of cellular ion status (including disruption of Ca2+ homeostasis, altered ion
9	transport mechanisms, and perturbed protein function through displacement of metal
10	cofactors) appears to be the major unifying mode of action underlying all subsequent
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modes of action in plants, animals, and humans (Figure 5-1). 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 calcium-dependent cell signaling,
such as 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 of fish disrupts the Na+ and CI" 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 affected by Pb are highly conserved across species accounting for the common
response seen in human health and ecological studies. This evolutionarily conserved
response to Pb is likely the result of the 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.
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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 '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 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
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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 will 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 animals, including cattle; however the results
of these studies are mixed and require further investigation.
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 caused 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 (>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 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 following Pb exposure.
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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.
2.9 Policy Relevant Considerations
2.9.1 Public Health Significance
The weight of the epidemiologic and toxicological evidence clearly supports the causal
relationship of Pb with decrements in cognitive function and behavior in young children
(Section 5.3.2 and Section 5.3.13). Associations of blood Pb level with higher IgE,
asthma and allergy (Section 5.6.8) and delays in the pubertal development in children
(Section 5.8.10), as well as associations of blood and bone Pb with increased blood
pressure, hypertension (Section 5.4.7), and kidney function (Section 5.5.6) in adults are
also well-substantiated in the evidence base. These endpoints were among those
comprising the weight of the evidence for determining causal relationships for the organ
system effects.
The rationale for establishing the public health significance of the various health
endpoints associated with Pb exposure is multifaceted. The 2006 Pb AQCD concluded
that neurological effects in children and cardiovascular effects in adults were among the
effects best substantiated as occurring at blood Pb levels as low as 5-10|ig/dL (or possibly
lower), and that these categories of effects were clearly of the greatest public health
concern. Evidence of newly demonstrated immune and renal system effects in the general
population was also noted to be of potential public health concern during the previous
review (U.S. EPA. 2006b). No safe level of Pb exposure has been identified in current or
previous assessments and any threshold for Pb neurotoxicity would have to exist at
distinctly lower levels than those associated with the lowest blood Pb concentrations
examined in the epidemiologic studies included in this assessment. Recent epidemiologic
studies of children continue to find associations with several health endpoints in
populations with lower mean blood Pb levels than previously reported (Chapter 5, Table
2-9).
The concept of population risk is relevant to the interpretation of findings for the
continuously-distributed subclinical health endpoints frequently studied in association
with Pb biomarkers in the assessment of their public health significance. The World
Health Organization 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
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this definition, even decrements in health status that are not severe enough to result in the
assignment of a clinical diagnosis might be undesirable if they 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. The American Thoracic Society (ATS) discusses concepts
related to understanding the shift in a population distribution of a health endpoint (ATS.
2000). As shown in Figure 2-1, a seemingly small increase in the mean of a continuously
distributed health index may push the most susceptible group in the population above a
critical cut point on the continuum of disease development, such that their condition
meets the clinical definition of a disease. Moreover, small changes at the population level
could translate into large numbers of clinical events if a large population is affected.
Shift in Population Mean
C
O
Critical
Line
ro
3
Q.
O
Q_
- B
a)
Q_
Health Outcome
Figure 2-1 The effect of a small shift in population mean on the proportion of
individuals in the population diagnosed with clinical disease
(i.e., the proportion to the right of the "Critical Line.")
As shown in Table 2-6, the size of the Pb-associated effects observed in epidemiologic
studies may be small relative to the effect sizes that are considered clinically relevant.
Still, these small shifts in population means are often significant from a public health
perspective. For example, small changes in IQ or inattention are often significant at the
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population level, especially in the two tails of the outcome distribution. Further, a small
shift in the population mean may result in a substantial increase in the number of
individuals functioning in the low range of the IQ distribution who may fall into the IQ
range associated with increase risk of educational, vocational, and social failure after
experiencing a Pb-associated decrement in IQ. In cohorts of children where Pb-associated
decrements in cognition and attention were observed, lower academic performance,
antisocial behavior, or delinquent behavior assessed in adolescence or in early adulthood
were also observed in association with blood Pb levels. Studies also find that higher
blood Pb level is associated with measures of inattention and hyperactivity as well as
with ADHD or other diagnostic indices used in a clinical setting (Section 5.3.3.1). A
downward shift in the mean IQ value can also reduce the proportion of the population
achieving very high IQ scores. It is shown that interventions that shift the population
mean, in a beneficial direction, by an amount that is without clinical consequence for an
individual produce substantial decreases in the proportion of individuals with clinically
significant disorders. This evidence, which informs the public health significance of the
evidence on the nervous system effects of Pb, is summarized in detail in Section 5.3.12.
It is also important to note that the change in a population mean observed in an
epidemiologic study may be small compared to its standard error of measurement for the
test. As noted in the 2006 Pb AQCD, statistics that pertain to individual-level data should
not be used to draw inferences about group-level data. Measurement error affects the
likelihood of detecting an association and is not relevant to the size of the association that
is detected. If a study is large enough it will have adequate statistical power to detect
small changes. Bias may be introduced if the measurement error of the outcome is highly
correlated with the exposure and there is no evidence indicating that individuals with
higher blood Pb levels test systematically lower than their true IQ.
Pb-associated changes in other subclinical indices also increase an individual's risk for
health effects that are of greater clinical consequence, thus of greater public health
concern. Small increases in blood pressure or decreases in GFRthat are associated with
Pb biomarkers, may shift the population mean resulting in a larger proportion of the
population that is diagnosed with hypertension or CKD, respectively. Results from the
Framingham Heart Study show that higher levels of blood pressure, even within the
nonhypertensive range, impose increased rates of cardiovascular disease (Kannel. 2000a.
b). A continuous graded increase in cardiovascular risk is observed as blood pressure
increases, with no evidence of a threshold value. Most events arise not in the most severe
cases, but mainly in those with high normal blood pressure (i.e., mild hypertension).
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. In addition to the small increases in blood pressure associated
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with Pb, Pb-associated effects on cardiovascular morbidity outcomes such as ischemic
heart disease (Section 5.4.3.5) and mortality (Section 5.4.5) have been observed. The
high correlation between blood pressure and clinical cardiovascular outcomes combined
with the high prevalence of cardiovascular disease in the U.S. adult population translate
into a large increase in the prevalence of conditions in the population. In addition to the
small changes in markers of kidney function, which are potentially significant from a
public health perspective, Navas-Acien (2009) and Munter et al. (2003) both report
increased risk of clinically significant CKD and albuminuria in association with Pb
among NHANES III participants. CKD results in substantial morbidity and mortality,
and, even at earlier stages than those requiring kidney dialysis or transplantation, is an
important risk factor for heart disease. 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.
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Table 2-6 Illustrative examples contrasting the effect sizes observed in
epidemiologic studies to effect sizes considered significant in a
clinical setting.9
Organ System
Endpoint and Findings from Epidemiologic
Studies
Clinical Significance Context
Nervous
IQ: 1 |jg/dL increase in blood Pb is typically associated
with small decrement in IQ (e.g., < 3 point decrement
in FSIQ score)
IQ below 70 is considered a disability by the Social Security Disability
Insurance Administrationb
Cardiovascular
Systolic Blood Pressure (SBP): Doubling of blood Pb
is associated with an approximate 1 mmHg increase in
blood pressure (Nawrot et al.. 20021
Hypertension is defined as systolic over diastolic blood pressure that
is greater than 140/90. Pre-hypertension is blood pressure that is 120
to 139/80 to 89. Hypertension is a risk factor for other cardiovascular
diseases and mortality.

Glomerular Filtration Rate (GFR): Increased blood Pb
level is associated with reductions in GFR
(e.a.. Akesson et al. (20051 reports a decrease of-2.0
mL/min (95% CI: -3.2, -0.9) per 1 |jg/dL increase in
blood Pb)
GFR <60mL/min/1,73m2 for > 3 months indicates Chronic Kidney
Disease (CKD). Estimated GFR (eGFR) can be calculated using
several different equations. An equation using serum cystatin C rather
than serum creatinine may be a more sensitive marker of early kidney
damage.
Renal8
Serum Creatinine: Increased blood Pb is associated
with increased serum creatinine Ce.a.. (Kim et al..
19961 observed an increase 0.030 mg/g (95% CI:
0.011, 0.049) among those with peak blood Pb levels
< 10 |jg/dL).
Creatinine > 30 mg/g creatinine indicates albuminuria.

Creatinine Clearance: [Increased blood Pb is
associated with decreased creatinine clearance
Ce.a.. Akesson et al. (20051 reports a -1.8 mL/min
(95% CI: -3.0, 0.7) per 1 |jg/dL increase in blood Pb)
Normal values range from 97 to 137 mL/min (males) and 88 to 128
mL/min (females). Abnormal results indicate poor kidney function.
aEffect sizes based on general population studies
information on Adult Mental Disorders can be found at: http://www.socialsecuritv.gov/disabilitv/professionals/bluebook/12.00-MentalDisorders-
Adult.htm
Rather than producing overt cytotoxicity or pathology, blood Pb level is associated with
alterations in several subclinical parameters related to cellular and humoral immunity
(Figure 5-42). Increases in IgE may contribute to more consequential health conditions
including allergy and asthma and the small changes observed in epidemiologic studies
may result in a population shift and increased prevalence of these conditions. Several
studies found associations between blood Pb levels and asthma and allergy in children
(Section 5.6.4.2). Immune changes may also be associated with autoimmune diseases
later in life as well as reduced capacity to combat certain viral infections and cancer.
Although the public health significance of a delay in onset of puberty is less clear, there
is some limited evidence that delay in puberty is associated with other health outcomes
(Gilsanz etal.. 2011; Naves et al.. 2005).
There is a new body of literature suggesting there may be relationship between maternal
bone Pb, cord Pb and/or concurrent blood Pb and cardiovascular effects in children
including increased blood pressure, Total Peripheral Resistance (TPR) responses to acute
stress tasks, and acute stress-induced autonomic and cardiovascular dysregulation
(Section 5.4.4). This evidence, however, is difficult to interpret with respect to the risk
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for health effects because cardiovascular and renal effects are associated with long-term
exposures to Pb and there is uncertainty regarding the frequency, timing, duration and
level of exposure associated with the effects. There is some evidence that the
pathogenesis of CVD begins in childhood (Kaoukii et al.. 2006). Although compensatory
mechanisms may be more active in children compared to adults, these early
cardiovascular effects may persist and worsen if Pb exposure persists over the long-term.
A recent study that strengthens the evidence of an association between blood Pb and
altered renal function in children is included in this review. This NHANES analysis in
adolescents found that association between higher blood Pb and lower cystatin C-based
eGFR (Section 5.5.6). This association is reported in a population with mean blood Pb
levels less than 5 j^ig/dL: however, those enrolled in the study were likely to have higher
past than recent Pb exposure and the pattern and level of their exposure was not assessed
in the study.
In addition to the long-term effects of Pb, early life exposure to Pb (postnatally) may
permanently alter normal development of the CNS to increase risk of pathology in adults
(Section 5.3.2.2). For example connections between developmental exposure to Pb with
early life programming and the resulting inflammation-associated DNA damage with
neurodegeneration have been demonstrated in rats. Additionally, there is animal
toxicological evidence for a Pb exposure contribution to Alzheimer's disease pathology
through the generation of neuronal plaques is in early life (Section 5.3.7.2). Early life Pb
exposure is also associated with impaired auditory function (Section 5.3.4.3).
2.9.2 Air Lead-to-Blood Lead Relationships
The 1986 Pb AQCD described epidemiological studies of relationships between air Pb
and blood Pb. Much of the pertinent earlier literature described in the 1986 Pb AQCD
was drawn from a meta-analysis by Brunekreef (1984). Based on the studies available at
that time that considered multiple air-related Pb exposure pathways in the aggregate, 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 (ig/dL) and high air Pb levels
(much greater than 10 |ig/m3) prevailed relative to lower environmental exposures which
showed lower blood Pb and air Pb concentrations (<30 (ig/dL and <3 (.ig/nr1). 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 (ig/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 2-7, Table
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4-11). The range of estimates from these studies is 2-9 (ig/dL per (.ig/nr', which
encompasses the estimate from the Brunekreef (1984) meta-analysis of (3-6 (ig/dL per
(ig/m3). 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 across all 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. See Section 4.5.1 for a detailed discussion of studies that
inform air Pb-to blood-Pb relationships.
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Table 2-7 Summary of Estimated Slopes for Blood Pb to Air Pb Relationships
in Humans
Reference
Study Methods
Model Description
Blood Pb-Air Pb
Slope3
Children Populations
Brunekreef et al. (1984)
Location: Various countries
Model: Log-Log
All children:

Years: 1974-1983
Blood Pb: 5-41 fjg/dL
4.6 (1.5)c

Subjects: Children (varying age ranges, n>190,000)
Analysis: Meta analysis of 18 studies
(mean range for studies)
AirPb: 0.1-24 fjg/m3
(mean range for studies)
Children <20
|jg/dL:
4.8 (0.54)d
Hayes et al. (1994)
Location: Chicago, IL
Years: 1974-1988
Model: Log-Log
Blood Pb: 12-30 fjg/dL


Subjects: 0.5-6 yr (n = 9,604)
Analysis: Regression of quarterly median blood Pb and
quarterly mean air Pb
(annual median range)
AirPb: 0.05-1.2 fjg/m
(annual mean range)
8.2 (0.62)8
Hilts et al. (2003)
Location: Trail, BC
Years: 1989-2001
Model: Linear
Blood Pb: 4.7-11.5 fjg/dL


Subjects: 0.5-6 yr (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
(annual geometric mean range)
AirPb: 0.03-1.1 fjg/m3
(annual geometric mean range)
6.5 (0.48)f
Ranft et al. (2008)
Location: Germany
Years: 1983-2000
Model: Multivariate Log-Linear
Blood Pb: 2.2-13.6 fjg/dL


Subjects: 6-11 yr (n = 843)
Analysis: Pooled regression 5 cross-sectional studies
(5th-95th percentile)
Air Pb: 0.03-0.47 fjg/m3
(5th-95th percentile)
3.2 (0.1 )g
Schnaas et al. (2004)
Location: Mexico City
Years: 1987-2002
Model: Log-Log
Blood Pb: 5-12 fjg/dL


Subjects: 0.5-1 Oyr (n = 321)
Analysis: Regression of lifetime blood Pb from longitudinal
blood Pb measurements and annual average air Pb data
(annual GM range)
Air Pb: 0.07-2.8 fjg/m3
(annual mean range in yr of
birth)
2.2 (0.4)h
Schwartz and Pitcher
Location: Chicago, IL
Model: Linear

(1989). U.S. EPA (1986a)
Years: 1976-1980
Subjects: Black children, 0-5 yr (n = 5,476)
Analysis: Chicago blood Pb screening, gasoline consumption
data, and Pb concentrations in gasoline
Blood Pb: 18-27 fjg/dL
(mean range)
AirPb: 0.36-1.22 fjg/m3
(annual maximum quarterly
mean)b
8.6 (0.75)'
Tripathi et al. (2001)
Location: Mumbai, India
Years: 1984-1996
Model: Linear
Blood Pb: 8.6-14.4 fjg/dL


Subjects: 6-10 yr (n = 544)
Analysis: Regression of blood Pb and air Pb data
(regional GM range)
AirPb: 0.11-1.18 fjg/m3
(regional GM range)
3.6 (0.45)J
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Reference
Study Methods
Model Description
Blood Pb-Air Pb
Slope3
Adult Populations
Rodrigues et al. (2010)
Location: New England, U.S.
Years: 1994-1995
Subjects: Adult bridge painters (n=84,1 female)
Analysis: Regression analysis of blood Pb and air Pb data
(personal monitors) collected during work performing various
job-related tasks
Model: Log-log
Blood Pb: 16.1 pg/dL
(GM, 1,7 GSD)
Air Pb: 58 pg/m3
(GM, 2.8 GSD)
0.01 (58)k
Mixed Child-Adult Populations
Schwartz and Pitcher
(1989). U.S. EPA (1986a)
Location: U.S.
Years: 1976-1980
Subjects: 0.5-74 yr (n = 9,987)
Analysis: NHANES blood Pb, gasoline consumption data and
Pb concentrations in gasoline
Model: Linear
Blood Pb: 11-18 pg/dL
(mean range)
AirPb: 0.36-1.22 pg/m3
(annual maximum quarterly
mean)
9.3 (0.75)'
aSlope is predicted change in blood Pb (|jg/dL per pg/m3) evaluated at ± 0.01 pg/m3 from central estimate of air Pb for the study
(shown in parentheses)
"Based on data for U.S. (1986 Pb AQCD).
cln(PbB) = In(PbA) * 0.3485 + 2.853
dln(PbB) = In(PbA) x 0.2159 + 2.620
eln(PbB) = In(PbA) x 0.24 +3.17
fPbB = PbA x 6.5
gPbB =1.5 x EXP(0.9361 x (PbA-0.1)/0.44), where 1.5 |jg/dL is the background PbB, and 0.1 pg/m3 is the median PbA for the study; model also
adjusted for soil Pb concentration, which may reduce estimated slope
hln(PbB) = Ln(PbA) x 0.213 +1.615 for the 1987 cohort, see text in Section 4.5.1 for more study details.
'PbB = PbA x 8.6
'PbB =Pb A x 3.6
kln(PbB) = In(PbA) x 0.05 +2.12
'PbB = PbA x 9.63
GM, geometric mean; GSD, geometric standard deviation; PbB, blood Pb concentration (|jg/dL); PbA, air-Pb concentration (|jg/m3)
2.9.3 Ecological Effects and Corresponding Lead Concentrations
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, biota). The proportion of observed effects of Pb attributable to Pb from
atmospheric sources is difficult to assess due to a lack of information not only on
bioavailability, as affected by the specific characteristics of the receiving ecosystem, but
also on deposition, and on kinetics of Pb distribution in ecosystems 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,
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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.
2.9.4 Concentration-Response Functions for Health Effects
With each successive assessment to-date, the epidemiologic and toxicological study
findings show that progressively lower blood Pb levels or Pb exposures are associated
with cognitive deficits and behavioral impairments (Section 5.3.9). C-R functions have
been examined across different populations and the interpretation of these functions
depends on several factors including lifestage and Pb biomarker used for the analysis.
Compelling evidence for a steeper slope for the relationship between blood Pb level and
children's IQ at lower blood Pb levels was presented in the 2006 Pb AQCD based on 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). and several individual studies
noted in Figure 2-2. The majority of the epidemiologic evidence from stratified analyses
comparing the lower and the higher ends of the blood Pb distributions indicates larger
negative slopes at lower blood Pb levels.
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Study	Blood Pb Timing Outcome Blood Pb	Sample size
stratum (|jg/dL)
Lanphearetal. (2000) Concurrent	Readingscore All	4853
<10	4681
<7.5	4526
<5	4043
<2.5	2467
Tellez-Rojoetal. (2006) Concurrent
Bellingerand Needleman
Peak
(2003)
Canfield etal. (2003) Lifetime avg
Lanphearetal. (2005) Concurrent
BayleyMDI
FSIQ
FSIQ
FSIQ
Jusko (2011)	Peak	FSIQ
Kordasetal. (2006)	Concurrent	Maths'
Schwartz (1994)	Early childhood FSIQ
>10
<10
5-10
<5
>10
<10
All
<10s
>10a
<10a
>7.5a
<7.5s
All
< 10
all
<10
>15
<15
90
294
101
193
NR
NR
172
101
1089
244
1230
103
174
96
532
293
NR
NR
Change in Cognitive Function Score perl |jg/dL
increase in blood Pb level (95% CI)
Note: Studies are presented in order of increasing mean blood Pb level.
aStrata refer to peak blood Pb level measured in child at any point during follow up.
FSIQ = full-scale IQ, MDI = mental development index. Effect estimates are standardized to a 1 |jg/dL increase in blood Pb level.
Black symbols represent effect estimates among all subjects or in highest blood Pb stratum. Blue symbols represent effect estimates in lower
blood Pb strata.
Figure 2-2 Comparison of associations between blood Pb and cognitive
function among various blood Pb strata.
Relatively few studies examined the shape of the concentration-response relationship
between Pb in blood or bone and effects in adults. There is uncertainty regarding the
frequency, duration, timing and level of exposure contributing to the blood Pb or bone Pb
levels in the adult population studied. Some of the populations examined (e.g., NHANES,
NAS) are likely to have higher past than recent Pb exposure. Other populations
(e.g., worker populations) studied have ongoing exposure to Pb. As described elsewhere
in the document (Section 4.3, Section 5.3, Section 5.4, Section 5.5), 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.
In NHANES analyses describing the C-R function for blood Pb level and neurocognitive
effects in adults, only log-linear models were used to fit the data (Krieg et al.. 2010;
Krieg and Butler. 2009; Krieg et al. 2009). With regard to C-R relationships for
cumulative exposure metrics, nonlinearity in the BMS and NAS cohorts of
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community-dwelling older adults was examined with the use of quadratic terms,
penalized splines, or visual inspection of bivariate plots (Bandeen-Roche et al.. 2009;
Weisskopf et al.. 2007a; Shih et al.. 2006). There was some evidence for nonlinearity for
some but not all cognitive tests or subjects in the NAS cohort (Section 5.3.10). In
analyses of the BMS cohort, observations of a statistically nonsignificant quadratic term
(Shih et al.. 2006) or spline (Bandeen-Roche et al.. 2009) for tibia Pb indicated that a
linear model adequately fit the relationship between tibia Pb level and various tests of
cognitive performance.
Concentration-response relationships were examined in several epidemiologic studies of
blood pressure and mortality (Section 5.4.5). With regard to the concentration-response
relationship, meta-analysis of human studies found that each doubling of blood Pb level
(between 1 and >40 (ig/dL 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). However, there is uncertainty in the concentration-response relationship to
cardiovascular endpoints at lower blood Pb levels since most studies model a linear
relationship. 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. 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 superior fit to data in non-nested models. Davidson and MacKinnon
(1981) returned a p-value of 0.013 in favor of the natural log blood Pb level, over the
linear blood Pb level specification. This analysis indicates that systolic BP increase in this
cohort is better described as a logarithmic function of blood Pb level within the blood Pb
level range of the study than by a linear function. Animal toxicological studies provide
support for this concentration response relationship. Few studies that focused on
Pb-induced hypertension in experimental animals have included more than two exposure
concentrations; however these few studies appear to have a supralinear (concave
downward) dose response.
Studies investigating both all-cause and cardiovascular mortality report both linear and
non-linear relationships (Section 5.4.5). Findings from NHANES analyses were mixed
with Schober et al. (2006). reporting a linear association between the relative hazard for
all cause mortality with blood Pb, and Menke et al. (2006). reporting a non-linear
relationships between blood Pb level and the hazard ratio for all cause, MI, stroke, and
cancer mortality. There is uncertainty regarding the frequency, duration, timing and
magnitude of exposure contributing to the blood Pb levels among the NHANES
population studied; individuals in this population are likely to have higher past than
recent exposures. Weisskopf et al. (2009) examined the concentration-response
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relationships of patella and tibia Pb with mortality. Concentration-response relationships
were approximately linear for patella Pb on the log HR scale for all CVD, but appear
nonlinear for IHD (p <0.10). The peak HR is shown around 60 jj.g/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. 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.
Concentration response information was provided in a small number of studies of
Pb-related nephrotoxicity in the occupational setting (Weaver et al.. 2003b; Ehrlich et al..
1998V Data in 267 Korean Pb workers in the oldest age tertile (mean age = 52 years)
revealed no threshold for a Pb effect (beta = 0.0011, p = <0.05; regression and lowess
lines shown), however the mean blood Pb level in this population was 32 (ig/dL (Weaver
et al.. 2003b).
Non-linear concentration/exposure response relationships or attenuation of these
relationships at higher exposure levels is reported in the occupational literature for a
variety of effects. Explanations for this phenomenon include greater exposure
measurement error, competing risks, and saturation of biological mechanisms at higher
exposure levels, and exposure-dependent variation in other risk factors (Stavner et al..
2003). With respect to Pb exposure, different biological mechanisms may operate at
different exposure levels and/or there may be a lower incremental effect of Pb due to
covarying risk factors such as low SES, poorer caregiving environment, and other higher
environmental exposures. The 2006 Pb AQCD also considered the explanation for the
supralinear concentration response function postulated by Bowers and Beck (2006). who
stated that "a supralinear slope is a required outcome of correlations between a data
distribution where one is lognormally distributed and the other is normally distributed."
The 2006 Pb AQCD determined that, while the conclusions drawn by Bowers and Beck
may be true under certain conditions, their assumptions (e.g., that IQ are scores forced
into a normal distribution) were not generally the case in the epidemiologic analyses
showing a supralinear concentration response function. To support this conclusion, the
2006 Pb AQCD cited Hornung et al. (2006). which provided evidence that the IQ data
used in the pooled analysis of seven studies by Lanphear et al. (2005) were not
normalized and a log-linear model (a linear relationship between IQ and the log of blood
Pb) provided the best fit.
The factors contributing to the supralinear relationship between blood Pb levels and
neurocognitive function in children has not been examined widely in epidemiologic
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studies to date. However, in several populations, higher blood Pb levels have been
measured in susceptible groups such as those with higher poverty, greater exposure to
tobacco smoke, lower parental education, and lower birth weight (Lanphcar et al.. 2005;
Lanphear et al.. 2000). It has been suggested that in populations of low SES, poorer
caregiving environment, and greater social stress, the incremental effect of Pb exposure
may be attenuated due to the overwhelmingly larger effects of these other risk factors
(Schwartz. 1994). Several studies found significant associations of these
sociodemographic risk factors with neurocognitive deficits, and Miranda et al. (2009)
found that indicators of SES (i.e., parental education and enrollment in a free/reduced fee
lunch program) accounted for larger decrements in EOG scores than did blood Pb level
(Figure 5-7). Few studies have compared Pb effect estimates among groups in different
sociodemographic strata, and the limited data are mixed. Greater Pb-associated
neurocognitive deficits were reported in low-SES groups by Bellinger et al. (1990). In a
meta-analysis of eight studies, Schwartz (1994) found a smaller decrement in IQ per
1 (ig/dL increase in blood Pb level for studies in disadvantaged populations (-2.7 points
[95% CI: -5.3, -0.07]) than for studies in advantaged populations (-4.5 points [95% CI:
-5.6, -2.8]). It is important to note that blood Pb level is associated with deficits in
neurocognitive function in both higher and lower SES groups; however, it is unclear what
differences there are between groups in the decrement per unit increase in blood Pb and
whether these differences can explain the nonlinear concentration-response relationship.
In support of epidemiologic findings, toxicological studies provided some evidence that
compared with control or higher Pb exposures (e.g., 150 ppm), lower Pb exposures
(e.g., 50 ppm) during gestation and lactation (possibly with offspring stress co-exposure)
impaired learning and memory as indicated by increased FI responses on schedule
controlled behavior tests (Rossi-George et al. 2011; Corv-Slechta. 1994). The 2006 Pb
AQCD did not identify a biological mechanism for a steeper slope at lower than at higher
blood Pb levels but such a mechanism was not ruled out. In fact, several lines of
toxicological evidence support the possibility of lower and higher Pb exposures acting
through differentially activating mechanisms underlying cognition.
As detailed in Section 5.3.10, the shape of the C-R relationship for dopamine varied
across studies. However, relative to higher Pb exposures, lower Pb exposures were found
to reduce LTP (Gilbert et al.. 1999) and glutamate release in the hippocampus (Laslev
and Gilbert. 2002). LTP is one indication of synaptic plasticity (Section 5.3.8.4) and is
considered to mediate learning and memory. Glutaminergic neurotransmission via its
NMDA receptor has been implicated in learning and memory (Section 5.3.8.8). Thus,
these differential responses to lower versus higher Pb exposures may provide mechanistic
understanding and additional biological plausibility for the nonlinear associations of Pb
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exposure or blood Pb levels with neurodevelopmental outcomes observed in animals and
children.
While epidemiologic studies have not examined widely the shape of the
concentration-response relationship for other nervous system effects, as detailed in
Section 5.3.10, toxicological studies have found nonlinear C-R relationships for diverse
outcomes such as rod and retinal effects in the visual system (Giddabasappa et al.. 2011;
Fox et al.. 2010). motor function (Lcasure et al.. 2008). and hippocampal neurogenesis
(Fox et al.. 2008; Gilbert et al.. 2005).
The supralinear concentration-response relationship widely documented for Pb is
consistent with the lack of a threshold for Pb-associated nervous system effects in
children because a smaller effect estimate would be expected at lower blood Pb levels if a
threshold existed. However, an important limitation of previous studies in terms of
identifying whether a threshold exists, is the limited examination of effects in populations
or blood Pb strata with blood Pb levels more comparable to the current U.S. population
mean. While Schwartz (1994) did not find evidence for a threshold in the Boston study
data, the mean blood Pb in that population was 6.5 (ig/dL, and 56% of subjects had a
blood Pb level >5 j^ig/dL. Recent studies indicate a downward shift in the distribution of
blood Pb levels (i.e., 50% of subjects in the 2001-2004 NHANES population had a blood
Pb <1 (ig/dL (Braun et al.. 2008). Additionally, more sensitive quantification methods
have improved the detection limits, for example, from 0.6 (ig/dL in 1999-2002 NHANES
to 0.025 (ig/dL in 2003-2004 NHANES allowing categorization of children in multiple
blood Pb quantiles below 1 (ig/dL (Braun et al.. 2008). Consequently, the examination of
populations with large proportions of subjects at very low blood Pb levels has improved
the ability to discern a threshold for Pb-associated nervous system effects. In the
2001-2004 NHANES population, Braun et al. (2008) found higher odds ratios for
conduct disorder and ADHD among children with blood Pb levels 0.8-1.0 (ig/dL (2nd
quartile) compared with children with blood Pb levels 0.2-0.7 (ig/dL (1st quartile).
However, the large proportions of adolescents in NHANES analyses who were born the
1970s may have had higher past Pb exposures that contributed to associations observed
with concurrent blood Pb levels. Nonetheless, several recent studies reported associations
between blood Pb levels and deficits in cognitive and behavioral endpoints in children
ages 8-11 years with mean or quantile blood Pb levels <2 (ig/dL (C'ho et al.. 2010; Kim et
al.. 2009b; Miranda et al.. 2009). In comparisons of various quantiles of blood Pb,
Miranda et al. (2009) reported lower EOG scores in children in North Carolina with
blood Pb levels of 2 j^ig/dL compared with children with blood Pb levels of 1 (ig/dL.
Collectively, these new findings in children, as summarized in this document, do not
provide evidence for a threshold for the nervous system effects of Pb in the ranges of
blood Pb levels examined to date.
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2.9.5
Pb Exposure and Nervous System Effects in Children
As discussed in various sections across this document, there are uncertainties related to
the frequency, timing, duration and level of Pb exposure contributing to the health effects
observed in epidemiologic studies of adults and older children who are likely to have
higher past than recent Pb exposure. Both blood Pb and bone Pb, which are the most
common metrics of exposure/dose used in epidemiologic studies, are integrative
measures and do not allow these aspects of exposure to be distinguished. Bone Pb is a
measure of cumulative exposure in adults and is not typically used in studies of children.
As discussed in Section 4.3.5, blood Pb may reflect both recent exposures and Pb
released from the bone; the relative proportion of blood Pb from recent versus past
exposure is uncertain in the absence of information on exposure timing and duration.
Young children, however, do not have lengthy exposure histories and consequently the
interpretation of associations with blood Pb levels for this age group may be less
complicated compared to older age groups. Several lines of evidence, which are
discussed below, inform the interpretation of study findings as they relate to aspects of
exposure that can be attributed to the neurocognitive and behavioral effects of Pb
observed in young children.
A common limitation of epidemiologic studies is the potentially high correlation between
Pb exposure metrics at different ages in childhood and with maternal Pb exposure
metrics. In longitudinal studies of the effect of Pb in children, blood Pb levels remained
relatively stable over time. Thus, it is difficult to distinguish effects of past and
concurrent exposures. Further, concurrent blood Pb levels in children, although highly
affected by recent exposure, are also influenced by their past exposure (including prenatal
exposure) due to the rapid growth-related bone turnover in children. Thus, concurrent
blood Pb level in children also may reflect cumulative dose (Section 4.3.5.1). There is
some evidence that the influence of maternal Pb levels on postnatal blood Pb level is
substantially reduced soon after birth. Simon et al. (2007) followed a cohort of 13
children living near an Australian smelter from birth through 36 months. In general,
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. There was a good correlation between child blood Pb level
and child hand Pb loading (R2 = 0.70) in this study, indicating the influence of concurrent
Pb exposures on blood Pb during the early childhood years.
In epidemiologic studies, associations of cognitive function and behavior have been
observed with prenatal, early-childhood, lifetime average, and concurrent blood Pb levels
as well as with childhood tooth Pb levels and it is difficult to ascertain which lifestage
within childhood is associated with the greatest risk of Pb-associated effects on cognition
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(Section 5.3.9, Table 5-13, Table 5-14). Epidemiologic studies of Bayley MDI following
children beginning in utero and through three years of age indicate that short-duration
exposures in utero or during early childhood are associated with cognitive decrements in
young children. Among studies with prenatal and concurrent blood Pb measurements,
some found stronger associations for prenatal blood Pb levels (Hu et al. 2006; Bellinger
etal.. 1984). and other found stronger associations for concurrent blood Pb levels
("Wasserman et al.. 1998; Wasserman et al.. 1992). Studies that found associations with
concurrent blood Pb levels also tended to find associations with prenatal cord or maternal
blood Pb levels. Thus, both postnatal child and maternal Pb exposures may contribute to
lower cognitive function in young children. In addition, exposures that are reflected by
concurrent blood Pb measured when children are older, by cumulative blood Pb levels or
by tooth Pb levels have also been demonstrated to be associated with neurodevelopmental
deficits throughout school-age and into adolescence. These findings are consistent with
the understanding that the nervous system continues to develop throughout childhood.
However, the weight of epidemiologic evidence supports associations of concurrent
blood Pb level with neurodevelopmental effects (i.e., cognition and behavior) in children
(Sections 5.3.2.1 and 5.3.3.1).
The persistence of neurodevelopmental effects from low-level Pb exposure was also
considered in the 2006 Pb AQCD, 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 range of effects with prenatal and early
postnatal Pb exposures which persist to adulthood (Sections 5.3.2.1 and 5.3.2.4). 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 short-term Pb exposures during prenatal
and early childhood periods. 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.
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). Recent toxicological studies continue to point to an early life window
in which Pb exposure can contribute to pathological brain changes consistent with
Alzheimer's disease. The same Pb exposure (i.e., dose and duration equivalent to the
early life exposure, to adult rodents) did not induce neurofibrillary tangles in the aged
animals. However, blood Pb is not generally associated with Alzheimer's disease in
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epidemiologic studies of adults. In other studies, recent evidence indicates associations
between early life ALAD activity, a biomarker of Pb exposure, and schizophrenia later in
adulthood when schizophrenia typically manifests. Consistent with these findings,
toxicological studies have shown NMDA pathway disruption with Pb exposure;
manipulation of this pathway is known to be associated with schizophrenia. There is also
toxicological evidence for Pb-induced emotional changes in males and the occurrence of
depression changes in females. Sensitive windows of early life Pb exposure have been
associated with persistent changes in adulthood as demonstrated with animal models of
neurodegeneration, i.e., neurofibrillary tangle formation, and with epidemiologic findings
for the psychotic disorder schizophrenia. These effects are 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.
2.9.6 Populations Potentially At-Risk for Health Effects
The NAAQS are intended to provide an adequate margin of safety for both the population
as a whole and those groups with unique factors that make them potentially at increased
risk for health effects in response to ambient air pollutants. Interindividual variation in
human responses to air pollution exposure suggests that some populations are at
increased risk for detrimental effects of ambient exposure to an air pollutant.
Epidemiologic studies reporting results from stratified analyses designed to identify
effect measure modifiers comprise the overall weight of evidence for conclusions related
to the increased risk of specific populations to the Pb-related health effects presented in
Section 6.3. Supporting evidence from toxicological studies provide biological
plausibility for conclusions and summarizes factors that potentially influence Pb levels
within the body (Section 6.1) and factors related to differential Pb exposures
(Section 6.2). The factors that were evaluated are listed in Table 2-8.
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Table 2-8 Factors evaluated that may determine populations potentially at
increased risk from lead

Physiological Factors that Influence the Population Characteristics Potentially
Factors Potentially Related to Increased
Internal Distribution of Lead Related to Differential Lead Exposure
Risk of Lead Induced Health Effects
Age (Section 6.1 and Section 4.4) Age (Section 6.2.1)
Age (Section 6.3.1, and Section 5.3.8)
Sex (Section 6.2.2)
Sex (Section 6.3.2)
Genetics (Section 6.1, Section 4.1, Section 4.2)
Genetics (Section 6.3.3)
Pre-existing Disease (Section 6.3.4)
Race and Ethnicity (Section 6.2.3)
Race and Ethnicity (Section 6.3.6)
Socioeconomic Status (Section 6.2.4)
Socioeconomic Status (Section 6.3.7)
Proximity to Pb Sources (Section 6.2.5,

Section 3.2, Section 3.3, Section 3.5, and

Section 4.1)

Residential Factors (Section 6.2.6)
Lifestyle Factors
Smoking (Section 6.3.5)
Body Mass Index (Section 6.3.8)
Alcohol Consumption (Section 6.3.9)
Nutrition (Section 6.1 and Section 4.2)
Nutrition (Section 6.3.10)
Stress (Section 6.3.11)
Cognitive Reserve (Section 6.3.12)
Co-exposure (Section 6.1 and Section 4.2)
Other metals (Section 6.3.13)
1	Studies are included in Section 6.3 only if analyses were conducted to identify the
2	presence or absence of effect measure modification. By virtue of their design some cohort
3	studies, including cohort studies of pregnant women or other populations or lifestages
4	with no comparison group, are discussed in the endpoint-specific sections rather than in
5	Chapter 6. This integrative summary, however, draws on evidence relating to potentially
6	at-risk populations and lifestages appearing throughout this document. Also,
7	physiological factors that influence the internal distribution of Pb, population
8	characteristics potentially related to differential Pb exposure and factors potentially
9	related to increased risk of Pb induced health effects are considered together in this
10	integrative summary.
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2.9.6.1
Age and/or Lifestage
Infancy and Early Childhood
It is recognized that Pb can cross the placenta to affect the developing nervous system of
the fetus (Section 4.2.2.4). Further, elevated blood Pb levels among mothers present a
potential exposure to their children through breast milk. This transfer of Pb from mother
to fetus is observed to occur partly from the remobilization of the mother's bone stores
(Section 4.2.2.4). Other biokinetic factors that vary by age, including bone turnover and
absorption, also affect blood Pb levels. Typically, children have increased exposure to Pb
compared with adults because children's behaviors and activities include increased hand-
to-mouth contact, crawling, outdoor play, and poor hand-washing that typically result in
increased ingestion compared with adults. Blood Pb among different age groups are
shown in Table 6-1. Blood Pb levels are highest among the young children and decrease
with increasing age of the child.
There is evidence of increased susceptibility to the neurocognitive effects of Pb exposure
during several lifestages throughout childhood (Section 5.3.2.1 to Section 5.3.8, and
Section 6.3.1.1). Exposure to environmental toxicants during prenatal and/or early
postnatal development may alter the normal course of morphogenesis and maturation that
occurs in utero and early in life, resulting in changes that affect structure or function of
the central nervous system via altered neuronal growth and/or synaptogenesis/pruning
structure (Section 5.3.9). Synaptic pruning, which is active throughout early childhood
(ages 1-4 years), may underlie the elevated risk of young children to environmental
exposures. Overall evidence indicates early childhood as a lifestage of increased risk for
Pb-related health effects with epidemiologic studies report associations with cognition
(Section 5.3.9) among the youngest age groups (6 months to 3 years). Toxicological
studies provide support for the younger age groups being especially sensitive to Pb
exposure. 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 (Fullmer et
al.. 1985). Further, blood Pb level is associated with effects to the immune system
(Section 5.6) of children. Findings of a recent study 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 (Section 5.6.2.1 and Section 5.6.3).
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Adolescents
The 2007-2008 NHANES data suggest that sex-based differences in blood Pb levels are
not substantial until adolescence at which time blood Pb levels of boys are higher than
girls. Anatomical understanding of brain development in the typically or normally
developing child also helps to inform findings with associated pathologies by providing
grounds for comparison. Findings from MRI studies have been used to detail anatomical
changes within separate brain segments in the normally developing child and adolescent,
from age 3-30 (Section 5.3.9). Throughout adolescence, normal brain development
remains dynamic. Volumes of specific regions vary largely by sex and age, but also
include inter-individual variation. Studies have linked concurrent blood Pb level as well
as other blood Pb metrics) in adolescents to decrements in cognitive function, delinquent
or criminal behavior (Section 5.3.3.1). Delays in puberty onset (Section 5.8.1.3) and renal
effects (Section 5.5.1.1) are also observed in association with concurrent blood Pb level
in cross-sectional studies of adolescents. However, the populations of older children
studied generally had higher past exposures, which may have influenced the findings of
these studies.
Adulthood
There is uncertainty regarding the frequency, duration, timing and magnitude of exposure
contributing to the blood Pb levels among the adult population studied. Associations of
both blood Pb and/or bone Pb with blood pressure, hypertension and mortality, renal,
immune, hematological, and reproductive effects. Blood Pb and bone Pb levels tend to be
higher in older adults (>65 years) compared with the general population. Higher average
and median bone and blood Pb 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 4.2). In recent studies, age was
specifically examined as an effect modifier of the association of Pb with mortality
(Section 5.4.5), cognition (Section 5.3.2.5) and blood pressure (Section 5.4.2) in adults.
Results for age-related modification of the association between Pb and mortality were
mixed and no difference by age was observed for the associations between Pb and
cognitive function or blood pressure. Toxicological studies have shown increases in
Pb-related health effects by age that may be relevant to nervous system effects in humans
(Section 5.3.1).
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2.9.6.2
Sex
Multiple associations between Pb and various health endpoints have been examined for
effect measure modification by sex (Section 6.3.2). Some studies reported differences
between the associations for males and females; overall, the findings were mixed. For
example, 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. Toxicological studies continue to demonstrate
increased susceptibility of males to Pb for specific endpoints such as sensory function,
balance, liver hyperplasia, obesity, memory and gross motor skills. Males and females
show differential susceptibility in other toxicological studies across a wide array of
endpoints including of behavior, stress hormone homeostasis, and depression with some
endpoints having greater effects in one sex versus the other based on the exposure
paradigm.
2.9.6.3 Genes
The 2006 Pb AQCD observed that genetic polymorphisms s have been implicated as
influencing the absorption, retention, and toxicokinetics of Pb in humans, with the
majority of discussion focused on the aminolevulinate dehydratase (ALAD) and vitamin
D receptor (VDR) polymorphisms (U.S. EPA. 2006b). Studies published since then
continue to indicate that variants of the VDR and the ALAD polymorphism have been
associated with varied internal Pb concentrations. Overall, studies of ALAD observed
increased Pb-related health effects associated with certain gene variants. Other genes,
such as VDR, MTHFR, APOE, 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 (Section 6.3.3).
2.9.6.4 Pre-existing 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, atopy, diabetes, and hypertension; however, there were generally
few studies available that evaluated conditions such as autism and atopy. The 2006 Pb
AQCD concluded that diabetes and hypertension may result in increased risk of
Pb-associated declines in renal function (U.S. EPA. 2006b). However, a recent study
indicates that diabetes was not related to increased risk of Pb-related cardiovascular
outcomes, and there were mixed results regarding the potential for hypertension to
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increase the risk of Pb-related effects on renal function and cardiovascular effects
(Section 6.3.4).
2.9.6.5 Race and Ethnicity
Higher blood Pb and bone Pb levels among African Americans has been well
documented (U.S. EPA. 2006b). Recent studies are consistent with those previous
findings (Section 6.2.3). Greater tibia Pb, but not patella Pb or blood Pb levels, was
reported in the Baltimore population studied by Theppeang et al. (2008b). This finding
may indicate greater historical exposure among African Americans compared to
Caucasians. In addition, a higher proportion of blacks compared with other race-ethnicity
groups has been reported in Census blocks where soil Pb concentration is higher, whereas
the proportion of whites was higher in low soil Pb Census blocks for a study of New
Orleans soil exposure (Campanella and Mielke. 2008). The results of recent
epidemiologic studies suggest that there may be race/ethnicity-related increased risk for
cardiovascular effects and delayed puberty, 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 cofounded by other factors, such as
socioeconomic status (Section 6.3.6).
2.9.6.6 Socioeconomic Status
Socioeconomic factors have sometimes been associated with Pb biomarkers, although
these relationships have not always been consistent (U.S. EPA. 2006b). On a national
level, the gap between income levels with respect to blood Pb has been decreasing.
However, blood Pb level of children remains higher depending on income, enrollment in
Medicaid and poverty income ratio (PIR), which is the ratio of family income to the
poverty threshold appropriate for a given family size (Section 6.2.4). Although there are a
limited number of studies, there is some evidence that Pb-associated neurocognitive
effects may be larger in magnitude among lower SES populations (Section 6.3.7). There
is also 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). In a meta-analysis, Schwartz (1994) found that in studies in
higher SES populations, blood Pb was associated with a greater decrement in IQ than in
low SES populations
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2.9.6.7
Proximity to Pb Sources and Residential Factors
Proximity to an industrial Pb source or Superfund site likely contributes to higher Pb
exposures (U.S. EPA. 2006b). Recent studies show that soil Pb concentrations may be
higher in some urban areas as a result of contemporaneous and/or historical Pb sources
(Section 6.2.5). Consistent with findings of the 2006 Pb AQCD (U.S. EPA. 2006b). child
blood Pb level was significantly associated with renovation activities in older homes as
well as housing factors (e.g., year of construction, floor and windowsill condition). It has
also been observed that windowsill dust Pb was significantly associated with the presence
of deteriorated indoor paint. Other studies of renovation activities involving removal of
Pb paint have included measurements of highly elevated blood Pb among children and
adult workers. Living in a home built prior to 1950, where old home age is a surrogate for
presence of Pb paint, has also been shown to be a significant predictor of blood Pb.
(Sections 4.1.3.2 and 6.2.6).
2.9.6.8 Lifestyle Factors
Body mass index (BMI), obesity and alcohol consumption have been examined in
epidemiologic and toxicological studies. Modification of associations between Pb with
mortality and HRV was not observed by BMI or obesity (Section 6.3.8). There is limited
evidence on the potential for alcohol consumption as a modifier of Pb-related effects
(Section 6.3.9) and mixed findings on whether smoking modifies the relationship
between Pb and health effects (Section 6.3.5).
2.9.6.9 Nutritional Factors
It is well established that diets sufficient in minerals such as calcium, iron, and zinc offer
some protection from Pb exposure by preventing or competing with Pb for absorption in
the GI tract (Section 6.1). The 2006 Pb AQCD included studies that indicated individuals
with iron-deficiency and malnourishment had greater inverse associations between Pb
and cognition/intellect (U.S. EP A. 2006b). and recent epidemiologic and toxicological
studies continue to observe greater Pb-related health effects among individuals with low
levels of iron (Section 6.3.10). Calcium intake from diet and vitamin D supplement can
modify blood Pb levels in women of various ages. 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. There is
little recent evidence available on the potential for zinc or other as modifiers of Pb-related
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health effects associations. A recent study in China reported that children who regularly
consumed breakfast had lower blood Pb levels than those children that did not eat
breakfast (Liu et al.. 201 lb). Recent toxicological studies r indicate that diets designed to
limit or reduce caloric intake and induce weight loss have been associated with increased
blood Pb levels and an epidemiologic analysis reported that regular consumption of
breakfast reduced blood Pb levels in children, compared with children who did not eat
breakfast (Section 6.1).
2.9.6.10 Stress and Cognitive Reserve
Animal toxicology findings described in the 2006 Pb AQCD demonstrated interactions
between Pb exposure and stress. Pb-exposed animals reared in cages with enriched
environments (toys) perform better in the Morris water maze than their Pb-exposed
littermates who were reared in isolation. New findings indicate a potentiating effect of
stress on behavior and memory at low-dose Pb exposures. Although examined in a
limited number of studies, recent epidemiologic studies observed modification of the
association between Pb and various nervous system and cardiovascular effects by
stress-level and also indicated that maternal self-esteem may attenuate the negative
effects observed of Pb on MDI and PDI scores. Toxicological studies have demonstrated
that early life exposure to Pb and maternal stress can result in altered responses in
multiple systems, including altered corticosterone and neurotransmitter levels.
Additionally, toxicological studies have demonstrated that immune stress also affects
associations with Pb (Section 6.3.11).
2.9.6.11 Co-exposure of Lead with Metals and Chemicals
Other exposures have also been studied to assess how they affected the uptake and
absorption of Pb. Recent toxicological studies that have examined the addition of arsenic
(As) to Pb and Cd mixtures report increases in bioavailability of Pb (Section 6.1). The
2006 Pb AQCD reported that the majority of studies examined other metals as potential
confounders rather than effect measure modifiers (U.S. EPA. 2006^. Recent
epidemiologic studies have, however, begun to explore the possible interaction between
Pb and other metals. These studies report some stronger associations between Pb and
various health endpoints with co-exposure to Cd, As and Mn (Section 6.3.13); some
studies indicate that coexposures may increase bioavailability of Pb (Section 6.1). Since
Pb is acid soluble, fluoridation may increase Pb concentration in water through leaching
from pipes and Pb solder (Section 4.1.3.3)
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Table 2-9 Summary of evidence for factors that potentially increase the risk of
lead-related health effects

Factor Evaluated
Potentially Related to Increased Risk
Age (Section 6.3.1)
Children
Sex (Section 6.3.2)
Males,® Females®
Genetics (Section 6.3.3)
ALADa, VDRa* DRD4a* GSTMf, TNF-aa, eNOSa, APOEa, HFEa
Pre-existing Disease (Section 6.3.4)
Autism3, Atopyab, Hypertensionb
Smoking (Section 6.3.5)
Smokers'
Race/Ethnicity (Section 6.3.6)
Non-Hispanic Blacks', Hispanics3
Socioeconomic Status (SES) (Section 6.3.7)
Low SESa
Nutrition (Section 6.3.10)
Iron deficiency8
Stress (Section 6.3.11)
High stress®
Cognitive Reserve (Section 6.3.12)
Low cognitive reserveab
Other Metals (Section 6.3.13)
High/co-exposure to Cda, Asa, Mna
a Evidence for this factor was limited.
'Possible mediator
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2.9.6.12 Summary of At-Risk Populations
As shown in Table 2-9 (Table 6-4) the most well-substantiated at risk population for the
effects of Pb exposure is children. Evidence for all other factors was limited. 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. Recent epidemiologic
studies of infants/children detected increased risk of Pb-related health effects, and this
was supported by toxicological studies. Synaptic pruning, which is active throughout
early childhood (ages 1-4 years), may underlie the elevated risk of neurodevelopmental
effects in young children to environmental exposures. Other factors that have been
studied with regard to the influence on Pb-related risk are listed in Table 2-9 below.

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2.10 Summary
Table 2-10 summarizes the main conclusions from the 2006 Pb AQCD and from this
assessment, including causality determinations, regarding the health and ecological
effects of Pb.
Table 2-10
Summary of evidence from epidemiologic, animal toxicological and
ecological studies on the effects associated with exposure to Pb
Outcome/Effect
Conclusions from the 2006 Pb AQCD
Conclusions from the (2012-2nd Draft) Pb ISA
Health Outcomes:
Nervous Svstem
Effects
The 2006 Pb AQCD did not integrate the evidence
across specific endpoints within the nervous system to
make conclusions regarding role of Pb in causing effects
for the organ system as a whole.
Based most heavily on cognitive function decrements and
behavioral problems in children, the collective body of evidence is
sufficient to conclude that there is a causal relationship between
Pb exposures and nervous system effects.
Neurocognitive
Function
and Learning
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 |jg/dL, and possibly
lower. Prenatal, early childhood, lifetime average, and
concurrent blood Pb levels were associated with
decrements in IQ and specific indices of learning and
memory; however, 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 new evidence emphasizes
associations of concurrent blood Pb levels with specific
indices of neurocognitive function (e.g., verbal skills,
memory, learning visuospatial processing). The weight of
evidence supports associations in populations with mean
blood Pb levels 2-7 |jg/dL.
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.
Consistent with previous studies, the weight of evidence in
new studies indicates associations of cumulative Pb
exposure metrics with cognitive performance in
environmentally-exposed adults. However, as these
outcomes are observed in adults with likely higher past Pb
exposures, uncertainty exists as to the Pb exposure level,
frequency, duration, and timing contributing to the observed
associations
Neurobehavioral
Effects
Children:
Several epidemiologic studies reported associations
between Pb exposure and effects that ranged from
inattention and aggression in children to delinquent
behaviors and criminal activities in adolescents and
young adults. Most studies examined blood Pb levels
measured earlier in childhood (means ~ 10 |jg/dL), tooth
Pb, or bone Pb. There was little examination of
concurrent blood Pb levels. Uncertainty remained
regarding whether Pb exposure was an independent
predictor of neurobehavioral effects. Results from studies
of ADHD were inconclusive. Suggestive relationship for
both blood and bone Pb with withdrawn behavior and
anxiety symptoms in children.
Recent studies in children continue to support associations
of Pb exposure with a range of effects, largely inattention
and hyperactivity, and also misconduct delinquent behavior.
In new studies, the weight of evidence supports
associations with concurrent blood Pb in populations with
lower mean blood Pb levels (2-5 |jg/dL) than those in
previous studies. New evidence indicates associations of
concurrent blood Pb levels with ADHD diagnosis and
contributing diagnostic indices in populations with mean
blood Pb levels 2-4 |jg/dL.
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Outcome/Effect
Conclusions from the 2006 Pb AQCD
Conclusions from the (2012-2nd Draft) Pb ISA
Neurobehavioral,
Mood and
Psychiatric Effects
Adults:
Associations of Pb exposure with behavioral outcomes
were not widely examined in environmentally-exposed
adults.
In the few available new studies of environmentally-
exposed adults, blood and bone Pb levels are associated
with symptoms of depression and anxiety.
Sensory Organ
Function
The selective action of Pb on retinal rod cells and bipolar
cells is well documented in earlier AQCDs. There was
coherence between the extensive animal and the limited
available human literature reporting associations
between concurrent blood Pb levels (population means
7-12 |jg/dL) and decreased auditory function in children.
The limited available new epidemiologic studies on sensory
organ function in children examined children with high
blood Pb levels ((means > 30 |jg/dL. A recent toxicological
study found retinal effects in male rodents who had been
exposed at gestation through 10 days of age at exposures
resulting in lower blood Pb levels (~12 |jg/dL) than those at
which effects had been previously reported.
Neurodegenerative
Diseases
In the limited body of epidemiologic studies, blood and
bone Pb levels were not associated with Alzheimer's
disease or Amyotrophic Lateral Sclerosis among
environmentally-exposed adults. Each study had
sufficient limitations.
A few case-control studies report associations between
bone Pb levels and PD in environmentally-exposed adults.
Recent toxicological evidence suggests that early life Pb
exposure may be associated with neurodegeneration in
adult animals.
The 2006 Pb AQCD did not integrate the evidence
across specific endpoints within the cardiovascular
system to make conclusions regarding role of Pb in
causing effects for the organ system as a whole.
Cardiovascular Effects
The weight of the evidence continues to support and
expand upon the strong body of evidence that Pb exposure
is associated with increased blood pressure and
hypertension in adults. The evidence is sufficient to
conclude that there is a causal relationship between Pb
exposures and cardiovascular effects; however, as these
outcomes are observed in adults with likely higher past Pb
exposures, uncertainty exists as to the Pb exposure level,
frequency, duration, and timing contributing to the observed
associations.
A meta-analysis of numerous epidemiologic studies
estimated that a doubling of blood Pb level (e.g., from 5
Blood Pressure to 10 |jg/dL) was associated with a 1 mmHg increase in
systolic BP and a 0.6 mmHg increase in diastolic BP."
Recent epidemiologic and toxicological studies continue to
support associations between long-term Pb exposure and
increased BP. Associations of increased BP with blood and
bone Pb concentrations are observed in populations with
lower mean blood Pb levels.
Hypertension
Cardiovascular
Mortality
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.
The evidence for an association of Pb with
cardiovascular mortality is limited but supportive.
Recent studies, including those using bone Pb as a metric
of cumulative exposure, continue to demonstrate
associations of hypertension with Pb levels in adults at
lower population Pb concentrations. Recent studies have
emphasized the interaction of cumulative exposure to Pb
with other factors including stress.
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.
Renal Effects
The 2006 Pb AQCD did not integrate the evidence
across specific endpoints within the renal system to
make conclusions regarding role of Pb in causing effects
for the organ system as a whole. 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 evaluated
in the current review support or expand upon the strong
body of evidence indicating that Pb exposure is associated
with kidney dysfunction (e.g., lower creatinine clearance,
higher serum creatinine, and lower GFR) in
nonoccupationally-exposed adults. The evidence is
sufficient to conclude that there is a causal relationship
between Pb exposures and renal health effects; however,
as these outcomes are most often observed in adults with
likely higher past Pb exposures, uncertainty exists as to the
Pb exposure level, frequency, duration, and timing
contributing to the observed associations.
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Outcome/Effect Conclusions from the 2006 Pb AQCD	Conclusions from the (2012-2nd Draft) Pb ISA
The 2006 Pb AQCD did not integrate the evidence
across specific endpoints within the immune system to
make conclusions regarding role of Pb in causing effects
for the organ system as a whole.
The consistency and coherence of findings across the
continuum of related immune parameters that demonstrate
a stimulation of Th2 responses in toxicological studies
combined with the supporting epidemiologic evidence in
children are sufficient to conclude that there is a causal
relationship between Pb exposures and immune system
effects.
Children:
Immune System
Effects
Several epidemiologic studies suggested that Pb
exposure may be associated with effects on cellular and
humoral immunity in children. The principal effects
demonstrated were decreases in T cell abundance and
increases in serum immunoglobulin E levels with
concurrent blood Pb levels > 10 |jg/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 inflammation and decreased
responses to antigens (e.g., delayed type
hypersensitivity, bacterial resistance).
A small body of new studies supports the previous
evidence that Pb exposure is associated with immune
effects in children. New 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 establishes the biological
plausibility for epidemiologic observations of associations
with asthma, allergy and inflammation-related effects in
other organ systems.
Adults:
In the large body of studies in occupationally-exposed
adults, the most consistent findings were reduced
neutrophil functionality in workers with blood Pb levels >
30 |jg/dL. Pb exposure-associated immune effects were
not widely examined in environmentally-exposed adults.
Several toxicological studies found a Pb-induced shift to
Th2 cytokine production and a hyperinflammatory
phenotype of macrophages in adult animals with long-
term (> weeks) exposure.
A small body of available studies provides new evidence
for immune effects in environmentally-exposed adults.
Specific endpoints varied among studies but were
consistent with increased inflammation. Associations were
observed in populations with a wide range of mean
concurrent blood Pb levels (3-22 ^jg/dL)._However, as these
outcomes are observed in adults with likely higher past Pb
exposures, uncertainty exists as to the Pb exposure level,
frequency, duration, and timing contributing to the observed
associations. A few available toxicological studies find Pb-
associated immune effects in adult mice.
Children:
Heme Synthesis and
RBC Function
Pb exposure was associated with disruption in heme
synthesis with increases in blood Pb levels of
approximately 20 |jg/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 |jg/dL Pb at 1 year of age, 50 |jg/dL at 3 years
of age, and 75 |jg/dL at 5 years of age.
Recent epidemiologic studies provide strong 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. Associations were observed in populations
with mean (or median) blood Pb concentrations as low as
approximately 5 |jg/dL.
Adults:
Pb exposure was associated with disruption in heme
synthesis with increases in blood Pb levels of
approximately 20 |jg/dL sufficient to halve ALAD activity
and inhibit ferrochelatase. Exposures to Pb resulting in
blood concentrations < 40 |jg/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 strong evidence
exposure to Pb is associated with numerous deleterious
effects on the hematological system, including altered
hematological parameters (Hb, MCV, MCH, RBC count),
perturbed heme synthesis mediated through decreased
ALAD and ferrochelatase activities, decreased
erythropoiesis, and oxidative stress.
The 2006 Pb AQCD did not integrate the evidence	The collective body of evidence integrated across
across specific endpoints within the reproductive system epidemiologic and toxicological studies with a focus on the
to make conclusions regarding role of Pb in causing strong relationship observed with negative effects on sperm
Reproductive Effects effects for the organ system as a whole.	and delayed pubertal onset is sufficient to conclude that
and Birth Outcomes	there is a causal relationship between Pb exposures and
reproductive and developmental effects.
Children:
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Outcome/Effect
Conclusions from the 2006 Pb AQCD
Conclusions from the (2012-2nd Draft) Pb ISA
No epidemiologic studies of delayed onset of puberty
were reviewed.
Recent toxicological and epidemiologic studies provide
strong evidence for delayed onset of puberty in males and
females. The weight of evidence for delayed onset of
puberty was among children ages 6-18 years with
mean/median blood Pb levels less than 5 |jg/dL.
Adults:
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 and epidemiologic studies provide
strong evidence for effects on sperm. Evidence on
pregnancy outcomes was inconsistent and less coherent
across disciplines for preterm birth, spontaneous abortion,
low birth weight, birth defects, hormonal influence and
fecundity.
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 coexposures (e.g., to
arsenic, cadmium), smoking, and dietary habits. The
2003 NTP and 2004 IARC reviews concluded that Pb
and Pb compounds were probable carcinogens, based
on limited evidence in humans and sufficient evidence in
animals. 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.The evidence i is
sufficient to conclude that there is a likely causal
relationship between Pb exposure and cancer.
Ecological/Welfare Effects:
Terrestrial Organisms:
Pb exposure may cause lipid peroxidation and changes
in glutathione concentrations. There are species
differences in resistance to oxidative stress.
Physiological 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.
Experimental exposures increasing from background
concentrations to concentrations higher than near point
sources result in increasing effects. The evidence is
sufficient to conclude a causal relationship between Pb
exposure and physiological stress in terrestrial plants,
invertebrates and vertebrates.
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
includes upregulation of antioxidant enzymes, production of
reactive oxygen species and increased lipid peroxidation
associated with Pb exposure in additional species of
aquatic plants, invertebrates and vertebrates and
decreased glycogen levels in freshwater snails. 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. The evidence is
sufficient to conclude a causal relationship between Pb
exposure and physiological stress in aquatic plants.
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Outcome/Effect
Conclusions from the 2006 Pb AQCD
Conclusions from the (2012-2nd Draft) Pb ISA
Terrestrial Organismps:
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.
Hematological Effects
Consistent with previous studies, the weight of the
evidence in new studies continues to support findings of Pb
effects on heme synthesis and ALAD enzyme activity. New
studies in birds near historical mining areas and altered
serum profiles and blood cell counts in vertebrates provide
evidence for additional species and hematological
endpoints potentially effected by Pb. The evidence is
sufficient to conclude a causal relationship between Pb
exposure and hematological effects in invertebrates and
vertebrates in terrestrial ecosystems.
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 |jg Pb/L.
Consistent with previous studies, the weight of the
evidence in new 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 new studies on altered blood cell counts in vertebrates.
The evidence is sufficient to conclude a causal relationship
between Pb exposure and hematological effects in
invertebrates and vertebrates in aquatic ecosystems.
Terrestrial Organisms:
Neurobehavioral
Effects
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. New
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). The evidence is sufficient to conclude
that there is a likely causal relationship between Pb
exposure and neurobehavioral effects in terrestrial
invertebrates and vertebrates.
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. New
studies identify possible molecular targets for Pb
neurotoxicity in fish and provide additional evidence for Pb
effects on behaviors in aquatic organisms that may impact
predator avoidance (swimming). The evidence is sufficient
to conclude that there is a likely causal relationship
between neurobehavioral effects in aquatic invertebrates
and vertebrates.
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Outcome/Effect Conclusions from the 2006 Pb AQCD	Conclusions from the (2012-2nd Draft) Pb ISA
Terrestrial Organisms:
No information on reproduction in plants.	There are an insufficient number of studies that consider
Pb effects on plant reproduction. Thus the evidence is
inadequate to infer a causal relationship for terrestrial
plants
Recent studies in a few taxa 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 level to levels
greater than near point sources. Data on terrestrial species
is supported by toxicological data from mammals in the
context of human health research. The evidence is
sufficient to conclude a causal relationship in terrestrial
invertebrates and vertebrates.
Limited evidence in invertebrates and vertebrates.
Developmental and
Reproductive Effects
Aquatic Organisms:
No reviewed studies on reproductive effects in aquatic
plants.
Reproductive and developmental effects reported in a
few species of invertebrates at <50 |jg Pb/L and in fish at
<150 |jg Pb/L
Recent evidence supports previous findings of reproductive
and developmental effects of Pb in aquatic biota and
differential lifestage response at near ambient
concentrations of Pb in some organisms. The evidence is
inadequate to conclude a causal relationship for plants,
and sufficient to conclude a causal relationship in
invertebrates and vertebrates.
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 some plant species' biomass in soil amended
with 30 mg Pb/kg soil. The evidence is adequate to
conclude a causal relationship in plants.
Limited studies considered effects on growth in
invertebrates and vertebrates. The evidence is inadequate
to conclude a causal relationship for terrestrial
invertebrates and vertebrates.
Aquatic Organisms:
Evidence for growth effects in algae, aquatic plants and
aquatic invertebrates
Most primary producers experience EC50s for growth in
the range of 1,000 to 100,000 |jg Pb/L
The weight of the evidence continues to support growth
effects of Pb in aquatic plants and invertebrates.Recent
studies on growth in invertebrates find effects of Pb at
lower concentrations than previously reported. The
evidence is sufficient to conclude a causal relationship
between Pb exposure and growth in aquatic plants and
invertebrates. Growth inhibition in one species of
freshwater snail was observed at <4 |jg Pb/L in juveniles.
Lowest genus mean chronic value for Pb reported at 10 |jg
Pb/L in a freshwater mussel.
The evidence is inadequate to conclude a causal
relationship between Pb exposure growth effects in aquatic
vertebrates.
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Outcome/Effect
Conclusions from the 2006 Pb AQCD
Conclusions from the (2012-2nd Draft) Pb ISA
Terrestrial Organisms:
Survival
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 weighUday) (U.S. EPA.
2005b).
Recent studies in invertebrates and vertebrates support
previous associations between Pb exposure and mortality.
The evidence is inadequate to conclude a causal
relationship between Pb exposure and survival for
terrestrial plants. The evidence is sufficient to conclude a
causal relationship in terrestrial vertebrates and
invertebrates.
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 |jg Pb/L dependent upon water quality variables
(i.e., DOC, hardness, pH).
Range of 96-hour LC50 values in fathead minnow: 810-
>5,400 |jg Pb/L
The weight of evidence continues to support Pb effects on
survival of aquatic invertebrates and vertebrates and
indicates that there are effects in a few species at lower
concentrations than previously reported.
New evidence for effects in a few invertebrates: at
<20 |jg Pb/L
New evidence in fish for impacts to survival at <100 |jg
Pb/L dependent upon water quality parameters and
lifestage
96- hour LC50 values as low as 41 |jg Pb/L in fathead
minnows tested in natural waters from across the U.S.
The evidence is inadequate to conclude a causal
relationship between Pb exposure and survival for aquatic
plants. The evidence is sufficient to conclude a causal
relationship between Pb exposure and survival for aquatic
invertebrates and vertebrates.
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.
New evidence for effects of Pb in soil microbial
communities add to the body of evidence for effects at
higher levels of biological organization. However, most
evidence for Pb toxicity to terrestrial 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
terrestrial organisms and subsequent effects on
reproduction, growth and survival at the species level is
likely to lead to effects at the population, community and
ecosystem level. The evidence is sufficient to conclude that
there is a likely causal relationship between Pb exposure
and the alteration of species richness, species composition
and biodiversity in terrestrial ecosystems.
Aquatic Ecosystems:
Most evidence of community and ecosystem level effects
is from near point 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).
New evidence for Pb effects on sediment-associated and
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 is likely to lead to
effects at the population, community and ecosystem level.
The evidence is sufficient to conclude that there is likely to
be a causal relationship between Pb exposure and the
alteration of species richness, species composition and
biodiversity in aquatic ecosystems.
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CHAPTER 3 AMBIENT LEAD: SOURCE TO
CONCENTRATION
3.1 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 new 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 3.2 provides an overview of the sources of ambient air Pb. Section 3.3 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 3.4. 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 3.5.
Concentrations of Pb in non-air media and biota are described in Section 3.6.
3.2 Sources of Atmospheric Lead
The following section reviews updated National Emissions Inventory (NEI) data from
2008 (U.S. EPA. 201 la), which is the most recently available quality-assured Pb
emissions data and compares these emissions data with those from previous years. 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 ambient
airborne Pb readily condenses to PM. The mobile source category included combustion
products from organic Pb antiknock additives used in piston-engine aircraft (hereafter
referred to piston-engine aircraft emissions).
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3.2.1
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 the use of 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. 200710. 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
3-1 and Figure 3-2. 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 76%,
from 4,100 tons in 1995 to 970 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 18% (170 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 56% 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 10% of emissions (100 tons) in 2008 (U.S. EPA.
201 la. 2008a).
1 This is the most recent version of the 2008 National Emissions Inventory and is posted separately from version 1
and version 1.5. The piston-engine aircraft emissions inventory can be obtained from the following site:
http://www.epa.gov/ttnchiel/net/2008inventorv.html under the link for Aircraft, Locomotive, and Commercial
marine Vessel sources.
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250
200
o
a
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£ 100
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a.
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I Highway Vehicles
I Metals Processing
I Fuel Combustion
I Piston Engine Aircraft
[ Miscellaneous
I
1970
1975 1980 1985
1990
1995
1999
2002
2005
2008
Source: U. S. EPA f2011a. 2008a)
Figure 3-1 Trends in Pb emissions (thousand tons) from stationary and
mobile sources in the U.S., 1970-2008.
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g	¦ Highway Vehicles
¦	Metals Processing
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w	¦ Miscellaneous
llhi.
1990	1995	1999	2002	2005	2008
Source: U. S. EPA (2011a. 2008a)
Figure 3-2 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 3-3. Piston-engine aircraft produced more than half of all emissions
(550 tons). Metal working and mining contributed 170 tons (18%) of Pb emissions in
2008, followed by industrial fuel combustion (15%), other industry (8%), and
miscellaneous contributions from agriculture, solvent utilization, and operation of
commercial marine vessels and locomotives (3%) (U.S. EPA. 2011a). 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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0	100	200	300	400	BOO	600
Pb Emissions (Tons)
Source: U. S. EPA (2011a)
Figure 3-3 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
3-4 for the continental U.S. The emissions levels, shown in units of tons, vary over
several orders of magnitude. Ninety-four percent of U.S. counties had 2008 emissions
below 1 ton, and 50% of counties had 2008 emissions below 0.044 tons. This category
included all counties emitting more than 20 tons of Pb in 2008. Jefferson County, MO
was the highest emitting county, with over 34 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 occur at approximately 20,000 airports across the
U.S. Airports tend to be more numerous around 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-engine aircraft are
estimated to emit cumulatively greater than one ton of Pb in 2008 U.S. EPA (2011a).
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2008 NEI Pb Emissions by County
Tons Pb Emitted Annually
0 00 - °02
| 0.03 - 0.04
| | 0.05 - 0.08
| | 0.09 -1.60	' <3
Source: U.S. EPA (2011a)
Figure 3-4 County-level Pb emissions (tons) in the U.S. in 2008.
3.2.2 Anthropogenic Sources
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Anthropogenic Pb source categones are organized below in order of magnitude with
regard to the sum of emissions nationally reported on the 2008 NEI (U.S. EPA. 2011a).
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 3-1 to serve as a reference for the
categories of Pb sources described in Sections 3.2.1 and 3.2.2.

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Table 3-1 Pb compounds observed in the environment
Emission Source
Observed Pb Compounds

PbS (Galena)
Minerals
PbO (Litharge, Massicot)

Pb304 ("Red Pb")

PbS04 (Anglesite)

Pb°, PbS

PbS04, PbO
Smelting aerosols
PbCOj

Pb silicates

PbS
Coal combustion aerosols


PbSe

Pb°, PbO, Pb02 (Above 1,150 K)
Coal combustion flue gases
PbCI2 (Low rank coals, above 1,150 K)

PbS04 (Below 1,150 K)
\No od combustion
o
o
_Q
Q_
\Naste incineration aerosols
PbCI2, PbO

PbCOj

PbS04

[PbFe6(S04)4(0H)12]
Soils near mining operations
[Pb5(P04),CI]

[Pb4S04(C0,)2(0H)3]

PbS-Bi2S3

Pb oxides, silicates

PbBr2
Piston-engine aircraft emissions, racing vehicle exhaust (combustion of leaded fuel)
Alkyl Pb

PbBrCI-NH4CI, PbBrCI-2NH4CI
Roadside dust
PbS04, Pb°, PbS04(NH4)S04, Pb304, PbO-PbS04 and 2PbCOrPb(OH)2
Brake wear, wheel weights
Pb°
Aircraft engine wear
Pb°
Source: Biggins and Harrison (1980, 19791: U.S. EPA (2006b).
3.2.2.1 Lead Emissions from Piston-engine Aircraft Operating on
Leaded Aviation Gasoline and Other Non-Road Sources
1	The largest source of Pb in the NEI, in terms of total emissions nationally, is emissions
2	from piston-engine aircraft operating on leaded aviation gasoline (U.S. EPA. 2011a).
3	Levin et al. (2008) point out that emissions from piston-engine aircraft are exempt from
4	reporting to the EPA Toxic Release Inventory. As outlined in Table 3-1, there are several
5	forms of Pb emitted from engines operating on leaded fuel. Dynamometer testing has
6	indicated that Pb emissions from piston engines operating on leaded fuel can occur in the
7	particulate and gaseous forms. For example, Gidney et al. (2010) performed
8	dynamometer testing on automobiles operating on standard gasoline and on gasoline with
9	low levels of organometallic additives. Tetraethyl Pb was included since it is still used in
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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 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 (Shotvk 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. were estimated to range from
1.6 g/sec of fuel during taxi-out to 15.3 g/sec of fuel during run-up preflight check for
single-engine aircraft and 5.1 g/sec during taxi and 50 g/sec during preflight run-up check
for twin-engine aircraft (C'arr 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 3.2.1. 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).
3.2.2.2 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 (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
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abundant species, with important contributions from Pb oxides. Pb concentrations
downwind of the smelter were roughly thirty-five times higher than upwind in this study
(0.625-0.880 (ig/m3).
Fugitive emissions from secondary Pb processing can be substantial over the course of a
year, but they are difficult to estimate. Thurston et al. (2011) performed source
apportionment of PM2 5 found that Pb-PM2 5 concentrations from the Chemical Speciation
Network (CSN) were associated with the metals industry along with Zn-PM2 5. Goyal et
al. ("2005) estimated fugitive emissions using concentration data obtained from samplers
sited in close vicinity of secondary Pb recovery 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 facilities in
Florida, Texas, and New York. Depending on the model used, median fugitive emissions
were estimated to be 1.0 x 10~6to4.4 * 10"5 g Pb/m2-sec at the Florida site, 9.4 * 10"7to
2.0 x 10"6 g/m2-sec for the Texas site, and 8.8 x 10"7to 1.1 x 10"6 g/m2-sec 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"6to 1.4 x 10"1 g Pb/sec, the Texas site median
ranging from 8.4 x 10"2to 8.6 x 10"2 g/sec, and the New York site ranging from 8.4 x 10"3
to 1.0 x 10"2 g/sec. Additionally, the Bayesian hierarchical model was used to estimate
fugitive Pb emissions nationwide using concentration data as prior information.
Nationwide median fugitive emissions were estimated to be 9.4 x 10"7 to
3.3 x 10"6 g/m2-sec. Recently, speciation of emissions from a battery recycling facility
indicated that PbS was most abundant, followed by Pb sulfates (PbS04 and PbS04-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
major component of the steel making process in southern France (Sammut et al.. 2010).
Cerussite, a Pb carbonate (PbC03-2H2o), 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 03, 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
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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, Zheng et al. (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 PM10-2.5 and PM2 5. They estimated that mine waste
was responsible for 88% of Pb in PM10_2 5 samples and 40% of Pb in PM2 5 samples.
3.2.2.3 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. PbS04, also derived from coal combustion, has low water solubility
(Barrett et al. 2010). 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 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 tons Pb/yr to the atmosphere. At the time of the Murphy et al. (2007)
study, there were no emission factors for crude oil or residual oil but these represent
potentially large sources (up to 100-500 tons/year and up to 25-700 tons/year,
respectively). As part of recent rulemaking, EPA has developed a draft Pb emission
factor of 1.3x10"5 lb/MMBtu for boilers larger than 25 MW that use #2 or #6 fuel oil
(U.S. EPA. 201 lb).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. 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; Pacvna and Pacvna. 2001). The Pb isotope
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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.
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. described in the 2006 Pb AQCD (U.S. EPA. 2006b) which indicated that for
West Virginia, higher emissions from power stations occurred in summer months. The
increased energy use in summer periods in the U.S. may be attributable to increased
requirements for air-conditioning.
3.2.2.4 Waste Incineration
Waste incineration studies suggest that the Pb content vary by industrial or municipal
waste stream. For example, Ogulei et al. (2006) performed positive matrix factorization
of PM2 5 and gaseous copollutants for Baltimore, MD and observed that 63% of Pb in
PM2 5 was attributed to waste incineration. Other prevalent compounds associated with
incineration included N03", EC, Cd, Cu, Fe, Mn, Se, Zn, 03, and N02 (note that CI was
not observed in this study). A study by Moffet et al. (2008a) 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 CI 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 3.4.1.5. 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
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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 -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/207Pb ratios for waste incineration Pb emitted in
European countries were 1.14-1.16 and 2.43 respectively (de la Cruz et al.. 2009).
3.2.2.5 Wood Burning
Another potentially uncontrollable source is Pb deposited historically in forests and
remobilized during forest fires. 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 S04 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. Other components strongly associated with the Quebec wildfires
included N03", OC, EC, Cd, Mn, Zn, 03, 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-PMio 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. 2011a).
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Several studies have explored the chemical properties of biomass emissions. Obernberger
et al. (2006) simualated 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 (Zuna et al.. 2011).
3.2.2.6 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. On pavement they may be ground into dust by
the pounding forces of traffic (Root. 2000). For example, Schauer et al. (2006) measured
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Pb emissions in two traffic tunnels and found that the Pb-PM2 5 concentration did not
exceed 17% of the Pb-PMi0 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 PM2.5 composition from brake
dust and found concentrations that were low but statistically significantly greater than
zero for Pb in PMi0 (0.02 ± 0.01 mg/g) and Pb in PM2 5 (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 et al. (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 (.un.
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 Pb were
emitted from brake wear each year, compared with 2.6 kg of Pb from tire tread wear; an
estimated 549 kg 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 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.
Unleaded Fuel
Unleaded fuel contains Pb as an impurity within crude oil (Pacvna et al.. 2007). Schauer
et al. (2006) measured Pb in PM25 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 <1-3 mg/kg Pb in samples of lubricating oil.
Hu et al. (2009a) also measured the size distribution of Pb emissions during
dynamometer testing of heavy duty diesel vehicles with different driving patterns and
control technologies. An urban dynamic driving schedule (UDDS) designed to mimic
urban stop-go driving conditions, was simulated in two cases to produce 80 and 241 ng
Pb/km driven, depending on the control technology used. Respectively, 54% and 33% of
those emissions were smaller than 0.25 |im in MMAD.
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3.2.2.7
Deposited Lead
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 notable source of airborne Pb. The concentration of
historically deposited Pb in soil is expected to be the greatest in soil next to roads.
Section 3.5.1.2 includes recent information on ambient air concentrations of Pb-TSP,
sampled at 6 m AGL at a distance of 500 m from the heavily trafficked 1-405 and 10 m
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 (Lankev 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 PM2 5 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 PM2 5 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 Pb concentrations. Such
mechanisms are described in more detail in Section 3.3. As described in Sections 3.2.2.6
and 3.6.1, there are many contemporary 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.
Lead 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 Herwiinen et al.. 2007; Deng and
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Jennings. 2006). Pb in industrial soils is described in Section 3.6.1. 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 consistent with Pb-humate as the
major form of Pb in airborne PM, suggestive of soil resuspension since the local near-
surface soils had high humic content (Pingitore et al.. 2009).
Lead 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 older findings.
Mielke and Gonzales (2008) sampled exterior paint chips from 25 homes in New
Orleans, LA, and they found elevated Pb levels in 24 of the 25 tested exterior paints
(median: 36,000 mg/kg). 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 comparison 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 also
affect the availability of Pb in paints. Edwards et al. (2009) performed experiments to
simulate one week of exposure of Pb-based paints to highly elevated levels of 03
(11.3 ± 0.8 mg/kg or 150 times the level of the 8-hour NAAQS) and N02
(11.6 ± 0.9 mg/kg, or 220 times the level of the annual NAAQS). Following N02
exposure, the Pb availability in wipe samples increased by a median of 260% (p < 0.001),
and following 03 exposure, the Pb availability 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. One month later, Pb dust loadings were still elevated in alleys (18%) and
sidewalks (18%), although they had decreased in streets by 29%. 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
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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.
Lead 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. (2010a'). These
estimates are presented in Figure 3-5. 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
c 150,000
I 100,000
50,000
^ ^ ^	J> -A* -A* ^	J>
Year
Source: Reprinted with permission of Pergamon Press, Mielke et al. (2010a).
Note: Estimates were derived from the proceedings of the U.S. Senate hearings on the Airborne Pb Reduction Act of 1984, S. 2609
(19841 and the U.S. Geological Survey Pb end use statistics CUSGS. 20051.
Figure 3-5 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 (.un mass median
aerodynamic diameter (MMAD) Pb-PM probably settled locally. EPA (1986a) indicated
that 35% of the Pb-PM at that time were < 0.25 |im 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.. 2010b). Figure 3-6 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. The
implication of this figure is that the soil Pb concentration in these areas will be
proportional to the magnitude of historic on-road emissions in each city. It is recognized
that the amount of soil turnover since 1982 may have varied substantially among the
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cities illustrated in Figure 3-6, depending on the amount of highway construction in those
cities. As noted in Section 3.2.2.6 , 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 gg/m3 (Section 3.5.3.2).
9*
'	/l	_ £1
• |
" Jjf
•	* * m /
6 A32	49#	40 64 W19 24 ~
V s—( »I* *
f
*"	•» » ^ 31
>	•«	*97
84	*54
•61
•2'
*85	•"	037
#18	^
I 58
56
U S Urbanized Areas	•» 0	Q	^
Pb 1950 -1982 (metric tons)	#28	l
•	608	vv	98V
•	652 - 4,570	^	79#
•	4.662 - 8,108	•	87« ^
0 8.403 - 16.614	W10
% 16,623 - 91,876	^
# 149,938
Source: Reprinted with permission of Pergamon Press, Mielke et a!. (2010a)
Note; The numbers on the map are rankings of each UA. The size of each dot refers to the magnitude of motor vehicle gasoiine-
related emissions for each group of LJAs.
Figure 3-6 Estimated Pb aerosol inputs from on-road gasoline into 90 U.S.
urbanized areas (UAs), from 1950 through 1982.
%
An
|42 W
039
3.3 Fate and Transport of Lead
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
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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. Figure 3-7 describes
relevant Pb transport pathways through environmental media discussed in this chapter
and their relationship to key environmental exposure pathways for which some or all of
the Pb is processed through the atmosphere. This discussion includes new research on
atmospheric transport of Pb, atmospheric deposition and resuspension of Pb, Pb transport
in surface waters and sediments, and Pb transport in soil.
Biota
Air
Human
Exposure
Source
Soil
Runoff
Agriculture +
Livestock
Surface Water
Groundwater
Sediment
Drinking Water
Transport
to Sea
Note: Media through which Pb is transported and deposited are shown in bold.
Figure 3-7 Fate of atmospheric lead.
3^3/1 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
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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 has not only led to lower atmospheric concentrations
in areas impacted by vehicles (Section 3.5), but a pervasive pattern of decreasing
atmospheric Pb deposition and decreasing concentrations in other environmental media
even at great distances from sources.
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 wet deposition as 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/sec and
dry deposition fluxes ranging from 0.04 to 4 mg/m2-yr. Precipitation concentrations
ranged mostly from 0.5 to 60 |ig/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-yr. 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-yr 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. 2006^ concluded that resuspension by wind and traffic
contribute to airborne Pb near sources.
3.3.1.1 Transport
New 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
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of coarse geogenic mineral particles and concluded that toxic metals were often
associated with coarse PM. 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 PM 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. PbS04 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, CaC03, a PM mineral
component, to form Pb3(C03)2(0H)2, Pb(C03) and Ca(S04)2-H20 on the surface of the
calcite (Falgavrac et al.. 2006). In laboratory experiments, (Ishizaka et al.. 2009) also
showed that PbS04 could be converted to PbC03 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 chemical
characteristics of Pb can be altered by this process due to accompanying transformations.
3.3.1.2 Deposition
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. As a
consequence, it is possible that efforts to reduce acidity 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
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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. This could also explain
the greater importance of dry deposition near sources because coarse mode PM makes a
greater contribution to PM mass. Although recent observations are consistent with these
trends 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 (.un (-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 PMi0 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 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. 2006^. but in industrialized urban areas, they are considerably
greater. The following studies presented measurements of dry deposition flux obtained by
capturing deposited particles onto a sampling substrate. Hence, these measurements did
not provide information on net deposition following resuspension of deposited material.
Resuspension processes and measurements thereof are described in Section 3.3.1.3. For
example, 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/ni2 d) than for suburban New Brunswick
(mean: 8 |_ig/m2d). Sabin and Schiff (2008) (2008) measured dry Pb deposition flux
along a transect from Santa Barbara to San Diego, CA in 2006 and observed and
observed a range of 0.52-14 (.ig/nrd 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/nrd). The
second highest median Pb flux was observed at San Diego Bay, a military port
(3.3 |_ig/m2d). 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.
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Elements from anthropogenic sources, including Pb, were generally associated with fine
PM. In a study of Tokyo Bay (Sakata and Asakura. 2008). reported an average dry
deposition velocity of 1.06 cm/sec, which is at 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-yr, more than 10 times the upper bound of the range reported in the 2006 Pb
AQCD (U.S. EPA. 2006b).
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-yr; this was lower
than the value of 1.30-1.90 mg/m2-yr 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 took place before the start of the Watmough and
Dillon (2007) study.
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, new 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.
3.3.1.3 Resuspension of Lead 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 3.2.1,
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
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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 PM10.
Results from several studies have suggested a minor contribution from resuspension to
airborne Pb concentration is indicated by a smoothed soil Pb concentration profile that
decreases with distance 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). The smoothed profile is suggested to be consistent with
continual Pb resuspension and deposition due to atmospheric turbulence. 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
Section 3.5.3, 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 3-8, 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 3-8), 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 m/s) traffic on a test road and observed scales
of 0.6-2.7 m when winds were perpendicular to the test road and scales of 1.8-2.7 m
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, interfacial layer)
UCl - Urban canopy layer
USL - Urban surface layer
ML - Mixed layer
CFL - Constant flux layer
(i.e., inertial sublayer - ISL)
Outer layer




CFL/ISL
53 ^ ^ ^
_mJUl f5TRflj5l IU1: =
ill
jhua


h,


uU) = uhe*"H-"
I *uci
Street
canyon
Urban can
length L
Street
Canyon
Recirculation
Source: Reprinted with permission of Annual Reviews, Fernando (2010)
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.
Figure 3-8
Scales of turbulence within an urban environment.
Recent research oil 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.. 201 1); concentrations are typically
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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 relationship 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 over time 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. In their model simulations showed that 2.5 (.un and
10 |_im particles reached a maximum height of 0.04-0.06 m above ground level (AGL),
while 50 |_im particles reached a maximum of 0.2 m AGL and 75 |_im particles reached at
least 0.4 m AGL. 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 displacement than smaller particles. At the same time, lateral travel distance
following resuspension tended to decrease linearly with increasing particle size (0.5 |_im
particle: median lateral distance estimated at 3.75 m; 100 |_im particle: median lateral
distance estimated at 0.5 m), reflecting the counteracting force of gravity. Harris and
Davidson (2008) estimated that the combination of forces caused a maximum height of
0.05 m to occur for particles with diameters around 75-100 (im, depending on turbulent
wind conditions. For all cases simulated, the resuspension and deposition were estimated
to occur overtime 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
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.
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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 et al. (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 (C'ho et al.. 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 m 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 (.un with a mass median diameter of
greater than 60 |_im (Yang et al.. 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.
New 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 etal. 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 (Lara-C'azenave et al.. 1994). 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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3.3.2 Water
As described in the 2006 Pb AQCD (U.S. EPA. 2006b). atmospheric deposition has been
identified as the largest source of Pb in surface waters, but urban runoff and industrial
discharge are also important. 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. 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 and
manganese 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 PbC03 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 environmental removal by sulfides is particularly important.
Although atmospheric deposition was identified as the largest source of Pb in surface
waters in the 2006 Pb AQCD (U.S. EPA. 2006b). runoff from storms was also identified
as an important source. A substantial portion of Pb susceptible to runoff originates from
atmospheric deposition. Runoff from buildings due to paint, gutters, roofing materials
and other housing materials were also 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. In several studies Pb in runoff was
consistently mostly PM, with a relatively small dissolved fraction. Runoff release was
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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.
3.3.2.1 Lead Transport in Water and Sediment
Recent publications provide additional detail regarding Pb adsorption on iron 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 (Pokrovskv 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 (Stolpc 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 (.un pore-size filter, and that these metals had a higher affinity for organo-
mineral Fe-Al colloids (Pokrovskv 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 iron 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 material that passes through a 0.45 |_im pore-size filter (Pokrovskv et al..
2010). Pb adsorbed on colloidal surfaces rather than incorporated into particle structure is
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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 iron 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.
3.3.2.2 Deposition of Lead within Bodies of Water and in
Sediment
As described in the 2006 Pb AQCD (U.S. EPA. 2006^. 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). Urbanized areas tend to have greater aquatic Pb loads, 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
complexation. In peatland areas, such as those in upland areas of the U.K., organic acids
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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)1. 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 m/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 3-2).
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Table 3-2 Surface sediment Pb concentrations for various continental shelves
Location
Digestion solution
Pba (mg/kg)
East China Sea
HCI/HNOj/HF
10-49 (27)a
Mediterranean, Israel coast
HNOj
9.9-20
Aegean Sea
HCI/HNOj/HF
21-44 (34)
Banc d'Arguin, Mauritania
HCI/HNOj/HF
2.8-8.9
Campeche shelf, Gulf of Mexico
HCI/HNOj
0.22-20 (4.3)
Laptev Sea, Siberia
HCI/HNOj/HF
12-22
Pechora Sea, Russia
Not reported
9.0-22 (14)
'Values in parentheses are the average, where calculable
Source: Data from Fang et al. (2009) and references therein.
3.3.2.3 Flux of Lead 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). The effect of
pH change on Pb release from lake sediments has also been examined, revealing that
1.8 protons (H ) 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 and results indicated that release of Pb from suboxic and anoxic zones of the
sediment act as a Pb source to the overlying water of the lake (Sengor et al.. 2007).
Bacardit et al. (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,
while sediment flux accounted for 98-99% of Pb outputs. In this paper, sediment was
only modeled as an output.
Sediment resuspension from marine environments is similarly important, with
disturbance of bed sediments by tidal action in estuarine areas resulting in a general
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greater capacity for re-suspension of PM. 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.
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 stormwater 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 stormwater 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.
2006b). and 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
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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.
3.3.2.4 Lead 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
runoff 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
sewage systems (urban runoff and domestic wastewater) 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
PM phases as well as bedload, material moved by rolling, sliding, and saltating along the
bottom of a stream (Birch and McC'readv. 2009).
Several new studies reported that the size distribution of PM transported in runoff is
relatively uniform. Characterization of the roadside dust in Australia showed that Pb in
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PM was approximately uniformly distributed among PM size fractions of up to 250 (.un.
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
PbCr04 comprised 8-34% of total Pb with the highest concentrations being found in the
largest and smallest size fractions. Pb(C03)2 and Pb(OH)2 were measured in the two
middle size fractions, while PbO and PbS04 were present in the largest and smallest size
fractions (Barrett et al. 2010).
Murakami et al. (2007) also emphasized the importance of PbCr04 as an important
species of Pb 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 Cr and 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 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 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.
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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). The study suggested that interaction of
the rainfall-runoff with the grassy area may have resulted in removal of PM-bound Pb
and hence in the association of Pb with DOM. 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 fraction 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, Pb was strongly correlated with TSP, no
relationship between total suspended solids and rainfall, rain intensity, antecedent
dry days or runoff event duration were observed, and traffic volume also did not appear
to have an effect They concluded that runoff composition from site to site could be highly
variable. Most other studies, however, 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 PM 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 (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 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 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 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 (U.S. EPA. 2006b). The relocation of snow means that the area
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receiving the snowmelt is not necessarily the same area that which 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 j^im )/l iter 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 coarser PM (179-542 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 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 et al.. 2010. 2009; Watmough and Dillon. 2007) and
nearby Kawagama Lake, Canada (Shotvk and krachler. 2010). Results indicated that
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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 it
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 rapid
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 that had been transported from throughout the
soil profile and 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, particularly since higher concentrations of previously deposited anthropogenic Pb
are usually found in the near-surface sections of upland U.K. soils [e.g., (Farmer et al..
2005)1.
In another study Rothwell et al. (2007a) observed stormflow Pb concentrations 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
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DOC concentrations (Rothwell et al.. 2007a). In a separate study, Rothwell et al. (2007b)
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 Catt (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 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% Pb was derived from past on-road gasoline
emissions and 33-43% was from hydraulic Au mining sediment. The flow into the
Sacramento River from these sources is an ongoing process. Periods of high surface
runoff, however, mobilize additional fluxes of Pb from these two sources and carry them
into the river. These pulses of Pb, driven by rainfall events, suggest a direct link between
local climate change and transport of toxic metals in surface waters (Dunlap et al.. 2008).
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Rothwell et al. (2007a) 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. (2007a) 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
increase as a result of climate change and/or other anthropogenic impacts (flood
management, reservoir removal).
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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 PM size and composition, indicating
that size distributions for Pb-containing PM in runoff water varies 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 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
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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 OM. Runoff also follows a
two-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, and the understanding of Pb runoff has
become more complete since publication of the 2006 Pb AQCD (U.S. EPA. 2006b).
3.3.3 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 CI
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
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
fraction of Pb was associated with Mn and Fe oxides or carbonate.
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3.3.3.1
Deposition of Lead 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,
and soil Pb concentrations are typically on the order of 100 mg/kg (Mielke et al.. 2010a').
Removal and translocation of Pb in soil is an ongoing process.
High Pb soil concentrations 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 urban soils due to ongoing emissions from nearby
mining and smelting activities are continuing to impact on the urban environment (Tavlor
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-yr to about 3 mg/m2-yr with fair agreement
between locations in Canada, Scandinavia, and Scotland and showed a substantial
decrease compared to when leaded on-road gasoline was in widespread use (Shotbolt et
al.. 2008; Watmough and Dillon. 2007; Fowler et al.. 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. 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 3.3.3.2. High Pb
concentrations at greater altitudes were also found in Japan, especially above 600 m
(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
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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 was 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 indicated little change in soil Pb concentrations
since the phase-out of leaded on-road gasoline, consistent with the high retention reported
for Pb in soils.
3.3.3.2 Sequestration of Lead 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
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 et al.. 2010).
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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 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 high quality
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 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).
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 Mn02 occurred even when Mn02 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.
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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). PbS04 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 roadsides 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(C03)2(0H)2, indicating that sequestration of Pb via
formation of oxycarbonate minerals is only effective at near-neutral to alkaline pH values
(Figure 3-9 in Section 3.3.3.3).
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 and concluded that Pb adsorption to the bacterial cell walls may be important
with respect to Pb transport in soils. This new area of research provides important
evidence 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
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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 (Xic 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.
3.3.3.3 Movement of Lead 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
(Shotvk 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.
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
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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 S or 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).
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 3-9). The estimated annual losses via
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percolating soil water were -2 mg/m2-yr (Klaminder et al.. 2008a) and so the annual loss,
assumed to be from the mor layer, was greater than the atmospheric input of
~0.5 mg/nr-yr. 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. (Watmough 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.
Bs
BJC

Vt 1 BliUt-
Loss of lead
Buildup of
lead
i—> Present atmospheric deposition (mg rrr2yr1)
—* Plant uptake (mg rrr2 yr1)
* Soil water flux (mg rrr2 yr1)
Notes The atmospheric deposition rate is from (Klaminder et ai.. 2006a1. the plant uptake rates from (Klaminder et al.. 20051 and
estimated soil-water fluxes from (Klaminder et al.. 2006b1.
Figure 3-9 Schematic model summarizing the estimated flux of Pb within a
typical podzol profile from northern Sweden using data from
Klaminder et al. (2006a).
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Klaminder et al. (2008b) investigated in more detail the distribution and isotopic
signature of Pb persisted 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 concentrations changes within the
mor layer.
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.
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
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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/yr and 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 on non-forest soils and enhanced
mobility on Fe and OM colloids. Pb was strongly retained on acidic Mediterranean soil
columns, with association of Pb with the exchangeable, OM and crystalline Fe oxide
fractions appearing to favor mobility while 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 m/yr (Erel. 1998). The authors attributed Pb
movement within the soil column to advection and concluded that the soil 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
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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 PbS04, which may move more easily through soil and enter
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 were more Pb was mobile in
weathered topsoil than in the unweathered subsoil. 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 (Schuwirth et al.. 2007). 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 3-10, 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).
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1.0
Pb02
Plattnerite
,2+
0.5
Pbco3
>
Cerussite
0.0
Pb(OH)3"
-0.5
2
0
3
5
6
8
9 10 11 12 13 14
1
4
7
PH
Source: Reprinted with permission of Elsevier Publishing, Scheetz and Rimstidt (2009)
Figure 3-10 Eh-pH diagram for Pb in shooting range soils, Jefferson National
Forest, VA.
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
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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.
3.4 Monitoring of Ambient Lead
3.4.1 Ambient Measurement Techniques
3.4.1.1 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 3.4.2 below): (1) Reference Method forthe 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
PMio 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
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FRM for Pb-PM10 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, a directional difference of 45 degrees
can result in a nearly two-fold difference in 15 (.un 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/h
wind speed to 22 |_im at 24 km/h (Rodes and Evans. 1985). Figure 3-11 illustrates the
effect of sampler orientation on collection efficiency as a function of particle size.
Figure 3-11 Comparison of particle collection efficiency among different TSP
sampler types (Modified Andersen Sampler, Hi-volume Sampler
(for different incident wind direction (45°, 0°), Prototype
Dichotomous Sampler, and Original Andersen Sampler).
8 % turbulence
MOOIFIEO I CFM ANDEFSEN
}hI - VOLUME SAMPLER
PROTOTYPE DICHOTOMOUS SAMPLER
ORIGINAL 1 CFM ANDERSEN
PARTICLE DIAMETER , C jum )
Source: Reprinted with permission of the American Chemical Society; (Wedding et al.. 19771
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Some existing commercially available sampler inlets are designed to collect particles
larger than 10 (.un 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
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 PMi0 samplers).
The spatial scale for which ambient air Pb samples are representative varies depending on
particle sizes present, as discussed further in Section 3.5.3. 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,
PMio.2.5 particles travel much shorter distances before settling out than finer particles
(U.S. EPA. 2009). As a result, spatial and temporal heterogeneity is greater for PMi0.2.5
than for PM25, 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 previous AQCD (U.S. EPA. 2006b). It has also been acknowledged in
previous AQCDs, with a lengthy discussion appearing in the 1977 AQCD (U.S. EPA.
1986b. 1977).
The Pb-PMio FRM sample collection method specifies use of a high-volume PMi0
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 PM10 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.
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Other 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 3.4.2, and other sampling methods that have been used to measure airborne Pb
concentrations in research studies unrelated to network applications. Some of the more
commonly used methods are listed in Table 3-3. Most of these methods have been
described in considerable detail in the 2004 PM AQCD (U.S. EPA. 2004). Table 3-3 also
lists key conditions of capture for each method, including particle size, inlet type,
collection medium, and flow rate.
Table 3-3
Airborne Pb sampling methods



Sampler
Network
Sampler Type
Mass Median
Aerodynamic
Diameter
Inlet Type
Collection
Medium
Typical
Flow Rate
Reference
High Volume TSP
Pb-FRM
Single
Channel
TSP
None
Glass
1.13 m3/min
U.S. EPA
(2011e)
Low Volume PM,0
PM-FRM, NATTS
Single
Channel
< 10 nm
Louvered Inlet +
PM10
Impactor
Teflon
16.67 L/min
U.S. EPA
(2011e)
PM2.5
PM-FRM
Single Channel
< 2.5 |im
WINS Impactor
Teflon
16.67 L/min
U.S. EPA
(2011 el
Met One SASS
CSN
Multiple Channel
< 2.5 |im
Cyclone
Teflon
6.7 L/min
MetOne (2009)
IMPROVE
IMPROVE
Multiple Channel
< 2.5 |im
Cyclone
Teflon
22.8 L/min
IMPROVE
(2001)
MOUDI
None
Multistage
Impactor
8 stages
0.056-18 |im
Impactor
Teflon
30 L/min
(Maroleetal..
1991)
Noll Impactor
None
Multistage
Impactor
4 stages
< 108 |im
Impactor
Coated Mylar
Rotating arm
(Noll. 1970)
SEAS
None
Slurry
< 1.2 |im
Impactor
Slurry
90 L/min
(Pancrasetal..
2006)
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 micoro-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
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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.
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 PMi0 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.
3.4.1.2 Sample Analysis: Federal Reference and Federal
Equivalence Methods
As described in Section 3.4.1.1., measurements for determinations ofNAAQS
compliance must be made with FRMs or FEMs. As of October 12, 2011, 1 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 (HNO3),
facilitated by heat, or by a mixture of HN03 and hydrochloric acid (HC1) facilitated by
ultrasonication. The Pb content of the sample is analyzed by atomic absorption
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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.
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
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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
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.
3.4.1.3 Other Analysis Methods for Total Lead
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
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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 et al.. 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 (Finlavson-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
mostly in the form of Pb sulfate (Oi et al. 2006). XPS analysis is a surface technique that
is suitable only to a depth of 20-50A.
Other Total Lead Methods
Anodic stripping voltammetry, atomic emission spectroscopy, and colorimetry have also
been used for measurement of atmospheric Pb (Finlavson-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 amalgam on a mercury 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 atomic absorption spectroscopy 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
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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.
3.4.1.4 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 (Tessieret 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 and manganese oxides, bound to OM,
and residual. Extraction was carried out with successively stronger solutions, starting
with magnesium chloride 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
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. 2009). 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 (Granev et al. 2004; Kvotani and Iwatsuki. 2002; Wang et
al.. 2002b). Compared to other elements, a large fraction of total Pb is soluble (Granev 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; Povkio et al.. 2007; Sillanpaa et al.. 2005) and locations (Perrino et al.. 2010;
Dos Santos et al.. 2009; Cizmecioglu and Muezzinoglu. 2008; Dahl et al.. 2008; Sato et
al.. 2008; Annihnldi 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;
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Gutierrez-Castillo et al.. 2005; Heal et al.. 2005). leading to a better understanding of
mobility characteristics of Pb in airborne PM.
3.4.1.5 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
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 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 quantify Pb complexed with humic
substances from soil in road dust (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- and HPLC-ICPMS
Environmental analytical methods for organolead compounds prior to 2000 were
generally time consuming and costly, requiring extraction, derivatization, and detection
(Ouevauviller. 2000). These have been thoroughly reviewed (Pvrzviiska. 1996) and
method intercomparison studies have been conducted (Ouevauviller. 2000). More
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recently, speciation of organometallic compounds in environmental samples has usually
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 (Hirncr. 2006).
Chromatographic systems in common use are gas chromatography and high performance
liquid chromatography. Detection systems most commonly used are ICPMS, electron
impact ionization mass spectrometry (EI-MS), and electrospray ionization mass
spectrometry (ESI-MS) (Himer. 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 (Ouevauviller. 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
(Yogi andHeumann. 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 (Popcrcchna and Heumann. 2005; Jitaru et al.. 2004; Leal-
Granadillo et al.. 2000).
Lead Isotope 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. 2010V 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
Mo ran. 2010; Noble et al.. 2008; Hsu et al.. 2006; Widorv. 2006). Shotyk and Krachler
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(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 GIS mapping of Pb isotopic 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
2°8pb/2°6pb obtained from PM samples, 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
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the 206Pb/204Pb ratio decreased over time with increasing contributions of industrial, soil,
smoke, and sea spray sources. For the most part, these sources were not substantial
contributions to Pb-PM2 5 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.
3.4.1.6 Continuous Lead 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
(kidwe 11 and Ondov. 2004. 2001) uses direct steam injection to promote condensational
growth of sampled 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 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). Other high time resolution methods
suitable for Pb analysis in PM are under development, including near real-time XRF
analysis.
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 (.un (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 (Snvder et al.. 2009; Johnson et al. 2008; Bein et al.. 2007;
Reinard et al.. 2007; Peknev et al. 2006) or specifically targeted (Salcedo et al.. 2010;
Moffet et al. 2008b; Murphy et al. 2007) Pb measurements.
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3.4.2
Ambient Network Design
Four national monitoring networks collect data on Pb concentrations in ambient air and
report it to the Air Quality System (AQS). 1 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 3.4.2.2.
Measurements among these networks are not directly comparable in all cases because of
method differences, including the PM size range sampled (TSP, PM10, or PM2 5). Data
from these various networks are presented in Section 3.5.1 to provide information on
ambient Pb concentrations in different size ranges.
3.4.2.1 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).2 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
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/).
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).
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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. The Pb NAAQS measurements required at these sites may be as Pb-TSP
or Pb-PMio (75 FR81126).
Monitoring agencies are also required to conduct non-source-oriented Pb monitoring at
each National Core multipollutant monitoring network (NCore)2 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
neighborhood-scale lead 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 lead 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 m.
¦	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 m 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.
1	The 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	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,
DC, will have at least one urban station.
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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, are required to be operational by
December 27, 2011 (75 FR 81126). When the December 2010 Pb network requirements
are fully implemented, the Pb NAAQS monitoring network is expected to consist of
approximately 270 required monitors including approximately 210 source-oriented
monitors and 60 non-source-oriented monitors. Figure 3-12 shows the estimated
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 0.5 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
sites (Figure 3-13 and Table 3-14) are required to be operational no later than December
27, 2011.
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• Previously existing sites and any newly required 1 tpy source-oriented sites
¦ Potential new 0.50 tpy source-oriented sites1
~ NCore (non-source-oriented) sites
1Based on 2008 National Emission Inventory lead emission estimates
Figure 3-12 Map of Monitoring Sites in Current Pb NAAQS Monitoring
Network.
1 Estimates for source-oriented monitors are based on Pb emissions estimates in the 2008 National Emissions
Inventory.
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Quality assured results of this study were not available in time for this assessment. Note that the two Santa Clara Co., CA airports
are not distinguishable on the map.
Figure 3-13 Fifteen U.S. locations where a study is currently being performed
on airport Pb emissions.
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Table 3-4 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
3.4.2.2 Other Lead 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 PM2 5 is monitored
as part of the CSN and IMPROVE networks, and Pb in PMi0 as a part of the National Air
Toxics Trends (NATTS) networks (Figure 3-14 and Figure 3-15). These networks are
designed to meet different objectives than those of the Pb NAAQS monitoring network.
The purpose of the CSN is to monitor PM2 5 species to assist in understanding PM2 5
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 assists states in formulating
their emission control strategies, even if the network is not compliance-oriented. Pb is
one of 33 elements in PM2 5 collected on Teflon filters every third day and analyzed by
energy dispersive XRF spectrometry.
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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, lead in PM2 5 is monitored with is determined
by XRF, including Pb (University of California Davis. 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 10-3.5 (U.S. EPA. 1999V Pb is one of seven core inorganic HAPs
collected on Teflon filters and analyzed by ICPMS. As of December 2009, the network
consisted of 27 monitoring stations, including 20 urban and 7 rural stations operating on a
one in six day sampling frequency.
Pb monitoring is also conducted at NCore monitoring sites. Monitoring for Pb-PM2 5 is
currently being conducted at NCore sites as part of the larger CSN (described above). As
described in Section 3.4.2, monitoring for Pb-PMi0 is required to be operational at NCore
sites by December 27, 2011. Methods for Pb in PMi0.2 .5 are being developed as part of the
PM10_2 5 speciation pilot project and may be implemented at some NCore sites in the
future.
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Hawaii
CSN
IMPROVE
Puerto Rico &
Virgin Islands
Alaska
0 250 500	1,000 Miles 0 70 140 280 Miles 0	245	490
' c?
0 25 50 100 Miles
Figure 3-14 Pb-PM2.5 monitoring sites for CSN and IMPROVE networks.
NATTS Monitors
Counties
United States
Figure 3-15 Pb-PMi0 monitoring sites for NATTS network.
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3.5
Ambient Air Lead Concentrations
The following section summarizes data on ambient air Pb concentrations during the years
2008-2010. Data used in this chapter are from the period 2008-2010. Data from SLAMS
(i.e., source-oriented and non-source-oriented Pb-TSP) are presented. Data from other
networks, such as the CSN, IMPROVE, and the NATTS networks, are also presented for
scientific purposes to characterize ambient PM component concentrations in different
size fractions and for different settings. Measurements among these networks are not
directly comparable in all cases because there are method differences, including the PM
size range sampled.
The 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. 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 (U.S. EPA. 201 la). The remainder of FRM monitors reporting
to the AQS were classified as non-source-oriented. 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. For this analysis, 120 FRM monitors were considered source-oriented,
while 184 were considered to be non-source-oriented.1'2 However, the number of source-
oriented and non-source-oriented monitors differed for each analysis year because there
were changes in monitor siting.
The section begins with a description of concentrations observed in Pb-TSP
measurements obtained for NAAQS compliance at source-oriented and non-source-
oriented monitors across the U.S. Additionally, concentrations of Pb-PM10 and Pb-PM25
1	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).
2	The requirement for monitoring near sources emitting 0.5 or 1.0 tons/year 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)).
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are presented. Next, seasonal patterns and multi-year trends of Pb concentration are
presented for the U.S. It is notable that Pb concentrations have declined substantially over
the past 40 years; this is described further in Section 3.5.2. An examination of the peer-
reviewed literature is provided to evaluate the size distribution of Pb-bearing airborne PM
under varied ambient conditions. The relationship between Pb concentration and
concentrations of copollutants are presented. Finally, estimates of transcontinental
transport of Pb to the U.S. and current background Pb levels are provided. Summary
information is presented within this section, and detailed data are included in an
Appendix (Section 3.8) to this chapter.
3.5.1 Spatial Distribution of Air Lead
3.5.1.1 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 PM2 5
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. The Pb-PMi0
data are obtained from the NATTS network, and the Pb-PM2 5 data are from the CSN.
Although the Pb-PMi0 and Pb-PM2 5 data are not from compliance networks, data are
presented from these networks because the additional data presents a picture of 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 concentrations1 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 3-16 illustrates that the level of the NAAQS was exceeded in twenty
counties where source-oriented monitoring was performed. Summary statistics for the
monitor-specific one-month and three-month averages for these monitors are presented
below in Table 3-5, and detailed statistics for the one-month and three-month averages
are provided in Table 3-13, Table 3-15, Table 3-17, Table 3-19, Table 3-21, and Table
3-23 in the Appendix (Secton 3.8). The mean was skewed toward the 75th percentile of
1 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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the distribution for both the monthly and three-month data sets. The primary difference
between the one-month average and three-month rolling average data sets occurs at the
upper tails of the distribution. Data for sites at which one-month or three-month annual
site maximal were in the upper 90th percentile for 2008-2010 are presented in Table 3-6.
The highest monthly and three-month average concentrations occurred in Iron Co., MO,
Herculaneum, MO (Jefferson Co.), and Los Angeles, CA. The highest one-month annual
site max value occurred in Cook County, IL in 2008 followed by Iron County, MO in
2008. The highest three-month annual site max concentrations occurred in Herculaneum
in 2008, Los Angeles in 2008, and Iron County, MO in 2008. The majority of monitors
reported data that did not exceed the NAAQS during this three-year period, but the
generally higher values at a subset of the source-oriented monitoring locations tended to
skew the nationwide distribution of Pb concentrations upwards.
1 The one-month annual site max is defined as the highest one-month average for a monitoring site over a given
year, and the three-month annual site max is defined as the highest three-month average for a monitoring site over a
given year.
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Table 3-5 Summary data for source-oriented Pb monitors across the U.S.,
2008-2010

Mean, |jg/m3 Median, |jg/m3
95th%, (jg/m3 99th%, (jg/m3
Max, |jg/m3
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
Table 3-6 Summary data for sites at which source-oriented statistics for one-
month and three-month annual site max are in the upper 90th
percentile, 2008-2010
County
AQS
Highest 1-mo
Mean, |jg/m
Highest 3-mo
Mean3, [jg/m
Highest 1-mo
Annual Site Max,
[jg/m3 (Year)
Highest 3-mo
Annual Site Max,
[jg/m3 (Year)
Pike, AL
011090003
1.3
1.2

1.2 (2008)
Los Angeles, CA
060371405
2.9
2.5
2.9 (2008)
2.5 (2008)
Hillsborough, FL
120571066



1.8(2008)
Cook, IL
180350009


4.4 (2008)
2.2 (2008)
Iron, MO
290930016
4.2
2.5
4.2 (2008)
2.5 (2008),
2.1 (2009)
Iron, MO
290930021
2.6
1.9
2.6 (2008),
2.4 (2009)
1.9(2009)

290990004


2.4 (2008),

Jefferson, MO

2.4
2.0
1.6(2009),
1.6(2010)
2.0 (2008)
Jefferson, MO
290990011


1.5(2008)

Jefferson, MO
290990015
3.1
2.9
3.1 (2008)
2.9 (2008)
Jefferson, MO
290990020
2.2
0.99
2.2 (2008)

Jefferson, MO
290990021
1.6
1.1
1.6(2009),
1.6(2010)

Jefferson, MO
290990022
0.86



Jefferson, MO
290999001
1.6
1.2
1.6(2009)
1.2 (2009)
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.
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trm m!
rV i O-Q r>-W-rbA!>0-V-'cTs PvOA.
¦ft'V drJ^Sto^ST^i!^
ft #f /
;PdMS3r^^r:A
Concentration:
¦	>= 1.55 |jg/m3 (5 counties)
¦	0.76 - 1.54 |jg/m3 (2 counties)
0.16 - 0.75 |Jg/m3 (13 counties)
¦	0.06 - 0.15 |jg/m3 (13 counties)
<= .05 |jg/m3 (17 counties)
~ no data
Figure 3-16 Highest county-level source-oriented Pb-TSP concentrations
(jjg/m3), maximum 3-month average, 2008-2010.
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 3-17 illustrates that the level of the NAAQS was never
exceeded at non-source-oriented monitors. Summary statistics are presented below in
Table 3-7, and detailed statistics for the one-month and three-month average and maxima
non-source-oriented Pb-TSP concentrations are provided in Table 3-14, Table 3-16,
Table 3-18, Table 3-20, Table 3-22, and
Table 3-24 in the Appendix (Section 3.8). The mean was slightly higher than the median
for both the monthly and three-month data sets. The primary difference between the one-
month average and three-month rolling average data sets occurs at the maxima of the
distributions. Data for sites at which national maxima were reached for 2008-2010 are
presented in Table 3-8. This table shows that all non-source-oriented monitor results were
below the NAAQS. The highest monthly and three-month average concentrations
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occurred at the same site in Cambria County, PA. The highest annual site max 1-month
value occurred in Columbiana County, OH in 2010, followed by Cambria County, PA in
2009. The Columbiana and Cambria sites were also above the 90th percentile annual
3-month site max Pb concentrations.
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Table 3-7 Summary data for non-source-oriented Pb monitors across the U.S.,
2008-2010

Mean, |jg/m3
Median, |jg/m3
95th%, [jg/m3
99th%, (jg/m3
Max, |jg/m3
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
Table 3-8 Summary data for sites at which non-source-oriented statistics for
one-month and three-month annual site max are in the upper 90th
percentile, 2008-2010
County
AQS
Highest Monthly
Mean, |jg/m
Highest 3-mo
Mean3, [jg/m
Highest Monthly Annual
Site Max,
[jg/m3 (Year)
Highest 3-mo Annual
Site Max, pg/rn (Year)
Cook, IL
170310022
0.070
0.051
0.070 (2009), 0.062 (2010)
0.048 (2008), 0.047 (2009), 0.051
(2010)
Cook, IL
170310026
0.052
0.046

0.046 (2008)
Cook, IL
170316003
0.040
0.036


Madison, IL
171193007


0.066 (2008)

Saint Francois, MO
291870006
0.089
0.054
0.089 (2010)
0.054 (2010)
Saint Francois, MO
291870007
0.054
0.041


Saint Louis, MO
291892003

0.055
0.066 (2008)
0.055 (2008)
Columbiana, OH
390290019


0.14(2010)
0.057 (2010)
Columbiana, OH
390290022


0.065 (2010)
0.044 (2010)
Cambria, PA
420210808
0.13
0.073
0.058 (2008), 0.13(2009)
0.049 (2008), 0.073 (2009)
Delaware, PA
420450002
0.048
0.047

0.047 (2010)
Westmoreland, PA
421290007
0.053
0.048

0.048 (2008)
El Paso, TX
481410002
0.087

0.087 (2010)

El Paso, TX
481410033


0.057 (2009)

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.
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r \ ,
\ \ j
Concentration
¦ 0.06 - 0.15 |jg/m3 (4 counties)
<= .05 |Jg/m3 (43 counties)
~ no data
Figure 3-17 Highest county-level non-source-oriented Pb-TSP concentrations
(jjg/m3), maximum 3-month average, 2008-2010.
Concentrations of Pb Measured using PM10 Monitors (for HAP
Concentrations and Trends)
1	Figure 3-18 displays maximum 3-month averages for Pb-PMn, concentrations for 36
2	counties in which measurements were obtained. Among the 36 counties in which PMi0
3	monitoring was conducted, only one county, Gila County, AZ, reported concentrations
4	above 0.076 |ig/nr\ Three other counties reported concentrations greater than
5	0.016 (.ig/nr1: Wayne County, MI, Boyd County, KY, and the county of St. Louis City,
6	MO.
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2007-2009 Pb-PM10 County Maximum 3-Month Mean
Concentration:
¦	>= 0,076 ng/m3 (1 county)
¦	0.016 - 0.075 ng/m3 (3 counties)
¦	0.006 - 0.015 ng/m3 (17 counties)
<= .005 ng/m-" (15 counties)
~ no data
Figure 3-18 Highest county-level Pb-PMio concentrations (jjg/m3), maximum
3-month average, 2007-2009.
Concentrations of Pb Measured using PM2.s Monitors (for Speciation
Concentrations and Trends)
1	Figure 3-19 displays maximum 3-month average county-level data for Pb in PM2 s
2	concentrations for 323 counties in which PM2 5 measurements were obtained for
3	speciation in the CSN and IMPROVE networks. The data presented here are not
4	compared to the NAAQS because P\P > monitors are not deployed for the purpose of
5	evaluating compliance for the NAAQS. Among the 323 counties in which PM2s
6	monitoring was conducted, only eleven counties reported concentrations greater than
7	0.016 (.ig/m5: Jefferson, AL, San Bernardino, CA, Imperial, CA, Wayne, MI, Jefferson,
8	MO, Erie, NY, Lorain, OH, Allegheny, PA, Berks, PA, Davidson, TN, and El Paso, TX.
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15
Concentration:
2007-2009 Pb-PM2 5 County Maximum 3-Month Mean
-wivi fAg/ui (xi counties)
¦ 0.006 - 0.015 |jg/m3 (71 counties)
<= .005 Mfi/m11 (241 counties)
~ no data
Figure 3-19 Highest county-level Pb-PM2.5 concentrations (pg/m3), maximum
3-month average, 2007-2009.
3.5.1.2 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, Pb-PM tends to settle quickly over short distances after
becoming airborne because Pb has relatively high density; 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. Intra-
urban variability in Pb concentrations reported to AQS was described in detail in the
Appendix in Section 3.8.1 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
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36
Cleveland). All six counties contain source-oriented monitors. Maps and wind roses are
presented in the 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.
When collectively reviewing the data from the six counties, it became apparent that
spatial and temporal variability of Pb concentrations were commonly high. 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, the 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 3.3, 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 for the monitoring data analyzed in the Appendix.
Airborne Pb near Roads
Five monitors were selected from the TSP network to examine Pb concentrations in the
near road environment. The monitors were located in Los Angeles County, CA
(06-037-4002), Riverside County, CA (06-065-1003), Cook County, IL (17-031-0052,
17-031-6003), and Suffolk County, MA (25-025-0002). 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 are 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. The monitoring sites are
described in Table 3-9, and time series of Pb-TSP monthly concentration for all five
monitors are shown in Figure 3-20. The annual average over the two sites that were
reporting data in 1980, the first year presented in Figure 3-20, was 0.90 |_ig/nr\ This Pb-
TSP concentration from 1980 likely reflected the influence of Pb emissions from leaded
automobile gasoline (Figure 3-6 for annual national consumption of leaded motor vehicle
gasoline). By 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 m to Interstate-290 at a
monitor height of 2 AGL. The multi-site average was not substantially larger than the
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1	maximum three-month rolling average of 0.012 (ig/m3 for non-source-oriented monitors
2	for the period 2008-2010, and the Pb-TSP concentration varied little over the period
3	2001 -2010. Note that the monitor heights were 2-6 m AGL, which may be higher than
4	the vertical distance likely traveled by some particles (depending on particle size)
5	following initial re suspension (see Section 3.3.1.3).
Table 3-9 Sample of U.S. near road Pb TSP monitors
County,
State
Site ID
Latitude
Longitude
2008-2010
Design Value
(Hg/m?b
Monitor
Height (m
AGL)
Distance from Roads
Surrounding Area
Los
Angeles, CA
06-037-4002
33.82376
-118.18921
0.01
6
500 m to I-405 (San Diego
Freeway), 10 mto Long
Beach Blvd
High intensity residential,
urban
Riverside,
CA
06-065-1003
33.94603
-117.40063
0.01
4
Within 20 m of intersection
of Magnolia and Arlington
Ave.
High intensity residential,
mixed use urban
Cook, IL
17-031-0052
41.96548
-87.749928
0.02c
5
Near to intersection of I-90
and I-94, 80 m to I-90, 200
mto 1-94,70 mto railroad
Located at public utilities
water pumping station, high
density residential urban
Cook, IL
17-031-6003
41.872202
-87.826165
0.04c
2
Less than 10 mto I-290
(Dwight D. Eisenhower
Expressway)
Parking lot of Circuit Court
of Cook County, ZA
surrounded by Concordia
Cemetary
Suffolk, MA
25-025-0002
42.348873
-71.097163
0.02c
5
95 mto I-90, inside median
of Commonwealth Ave.
High intensity urban, mixed
use residential and
commercial
aThe level of the 2008 NAAQS for lead is 0.15 micrograms per cubic meter (|jg/m3) 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 or equivalent data reported by States,
Tribes, and location agencies to EPA's Air Quality System (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 lead-TSP average data are available at this site for this 3-year period; the value shown here is the highest valid 3-month
mean.
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17
18
19
1.6
1.4
	060374002
060651003
170310052
o 0.6
170316003
0.4
0.2
Note: Monitor IDs: Los Angeles, CA: 06-037-4002; Riverside, CA: 06-065-1003; Cook, IL: 17-031-0052, 17-031-6003; Suffolk, MA:
25-025-0002.
Figure 3-20 Time series of monthly average Pb-TSP concentration at five
near-road monitors.
Airborne Pb near Airports
Levin et al. (2008) summarized findings from environmental protection departments of
the State of Illinois, the U.S., and Canada regarding ambient Pb concentrations at and
near airports. Data presented in the Canadian report yielded median air Pb-PMn, level of
0.01 for ten sample days (with 37 24-hour and 3 11-hour samples) compared with an
average reported background level of 0.007 |ig/nr (Conor Pacific Environmental
Technologies Inc. 2000). Median and average values were derived from all samples
collected at the airport, both upwind and downwind from aircraft activity. The maximum
24-hour concentration measured in this 10-day study was 0.13 |ig/nr. The Illinois report
noted that air Pb concentrations were elevated downwind of O'Harc airport compared
with upwind levels (Illinois Environmental Protection Agency. 2002).
More recently, Carr et al. (2011) performed TSP monitoring and dispersion modeling of
Pb emissions at the Santa Monica Airport and surrounding neighborhood in Santa
Monica, CA. Ambient sampling was conducted in March and July, 2009. During winter,
measurements of 24-hour Pb-TSP concentrations upwind of the airport ranged from
0.003 to 0.011 (.ig/nr1 and at the two downwind locations, concentrations ranged from
0.025 to 0.083 (.ig/nr1 approximately 70m from the location where piston engine aircraft
emit the majority of Pb during ground-based operations and 0.028 to 0.050 (.ig/nr1 in an
adjacent neighborhood roughly 85m further downwind. During summer, 24-hour Pb-TSP
concentrations upwind from aircraft operations ranged from 0.001 to 0.006 |ag/nr\ Three
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locations downwind were monitored, the first of which was the same location as the
winter site 70m from the location where piston engine aircraft emit the majority of Pb
during ground-based operations. This location had concentrations ranging from 0.034 to
0.062 (ig/m3. At a site 100 m further downwind from this location, concentrations ranging
from 0.027 to 0.044 |_ig/nr' and at a distance of 175 m downwind, a 24-hour concentration
of 0.023 (ig/m3 was measured. Modeling results 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 and is
therefore not subject to Pb monitoring to evaluate compliance with the Pb NAAQS,
which is described further in Section 3.4.2.
Airborne Pb at Urban and Rural Sites
A number of studies have characterized how Pb-bearing PM is distributed over the
neighborhood scale in the air. Yu et al. (2011) measured Pb-PMi0 concentration at four
rooftop sites (10-13 m AGL) within Paterson, NJ: background, near-road, industrial, and
commercial. Interstate-80 and Route 19 were both within 0.8 km from the near-road site.
The industrial site was located 0.1-1 km from a metal recovery plant, a plating facility
that emits Ni and Zn compounds, and another facility emitting Cu compounds. The
commercial site was proximal to several restaurants and dry cleaners. Average
concentrations at each site were: 2.95 ng/m3 (background), 5.61 ng/m3 (near road), 6.48
ng/m3 (industrial), and 6.58 ng/m3 (commercial) to yield a coefficient of variation
coefficient of variation (CV, defined as the standard deviation of site measurements
divided by the average) of 31.3%. Weekday and weekend ambient Pb-PMi0
concentrations were not significantly different (p = 0.45). 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 and 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-PM2 5
concentration among the sites varied 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 an
upper cutpoint of 29 (jm 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. Sabin et al. (2006a') also reported deposition flux at the same sites, which
ranged from 8.3 to 29 |ig/m2-day at the rural sites, with a CV of 48%, and was 1.4 ng/m3
at the rural site. Li et al. (2009a') 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-
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PM2 5 concentration at three Baltimore sites. Average Pb-PM2 5 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. However, the authors do not specify which
industrial facilities were proximal to each monitor. 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.
3.5.2 Temporal Variability
The following section presents 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.
3.5.2.1 Multi-year Trends
Pb-TSP concentrations have declined substantially during the years 1980-2010. For
source and non-source monitors combined, the annual average across Pb-TSP monitors
reporting three-month average site max air Pb concentrations have dropped by 89% from
1.3 |ig/m3 in 1980 to 0.14 (ig/m3 in 2010. The median concentrations have declined by
97% from 0.87 (ig/m3 in 1980 to 0.03 (ig/m3 in 2010. While the sharpest drop in Pb
concentration occurred during 1980-1990 as a result of the phase-out of Pb antiknock
agents in on-road fuel, a declining trend can also be observed between 1990 and 2010
following reductions in industrial use and processing of Pb, as described in Section 3.2.1.
In 1990, the average Pb concentration was 0.84 (.ig/ni3 and the median Pb concentration
was 0.25 (.ig/ni3 to yield 84% and 71% reductions, respectively, from 1990 to 2010.
Average concentrations in these calculations are heavily influenced by the source-
oriented monitors in the network. New Pb concentration data from expansion of the
source-oriented portion of the network in 2010 will allow for greater assessment of
changes of Pb concentrations on nationwide statistics and trends. Figure 3-21 and Figure
3-22 show ambient Pb concentrations from 1990 to 2010 for source-oriented monitors
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and non-source-oriented monitors, respectively. In both cases concentration data are
consistent with a downward trend, and concentrations were considerably lower at the end
of the period than at the beginning of the period.
8.S
7.5
6.5
5 4.5
o 3.5
K 2.5
1978-2008 Level of the NAAQS
1.5
0.5 2008 Level of the NAAQS	
0 <9 Sj,	k>0	r>0	*?0
s> ¦%> ¦%> °o °j v °3 °s °s °> °
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1
2
3
4
5
6
7
8
9
10
11
12
13
14
0.35
0.3
--.0.25
cm
2008 Level of the NAAQS
5 0.15
Q.
Q.
0.05
\>	%>%, \> %o %,\-%W %
Note: Annual average of Pb-TSP monitors reporting three-month average site max. Concentration is shown by the solid black line,
median concentration is shown by the solid blue line, and the 10th and 90th percentiles are shown by the dashed lines.
Figure 3-22 National trends in Pb concentration (jjg/m3), non-source-oriented
FRM monitors, 1990-2010.
For source-oriented monitors, the annual average across Pb-TSP monitors reporting
three-month average site max decreased from 1.7 (.ig/rn1 to 0.27 (.ig/nr1 (84% decline) and
upper 90th percentile concentrations decreased from 4.9 |ag/irf to 0.73 |ag/irf (85%
decline) over the 20-year period. A portion of the decrease can be attributed to reductions
in emissions from the Herculaneum, MO smelter between 2001 and 2002 (U.S. EPA.
2010c). An abrupt decrease in average concentrations between these years is evident in
Figure 3-21.
For non-source-oriented monitors, the annual average across Pb-TSP monitors reporting
three-month average site max decreased from 0.12 |ag/irf to 0.018 |ag/irf (85% decline)
and upper 90th percentile concentrations decreased from 0.32 |ag/irf to approximately
0.04 (ig/m3 (88% decline) over the 20-year period. Average concentrations near
stationary sources were 5 to 8 times typical concentrations from non-source-oriented
monitoring locations between 1990 and 1999; during the subsequent decade, average
source-oriented Pb concentrations were 13 to 24 times higher than non-source
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19
concentrations (U.S. EPA. 2010c'). This differential likely reflects the absence of Pb
emissions from automobiles during 2000-2009.
3.5.2.2 Seasonal Variations
This section outlines seasonal variability among Pb monitors on a nationwide basis.
Seasonal variation may provide insight related to differential influences of sources and
climate throughout a year. Additionally, the magnitude of concentrations within the
monthly data distributions and of variations between months sheds light on the influence
of season as well as on differences between source-oriented, non-source-oriented, PMi0,
and PM2 5 data.
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 3-23), Pb-TSP
from non-source-oriented monitors (Figure 3-24), Pb-PM10 (Figure 3-25), and Pb-PM2 5
(Figure 3-26). For source-oriented Pb-TSP (Figure 3-23), monthly average concentrations
were determined from between 146 and 154 samples in each month. For non-source-
oriented TSP (Figure 3-24), 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 average. 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.
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0.9-
0.7-
0.6-
0.5-
0.4-
0.3-
0.2-
0.0-
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 3-23 Monthly source-oriented Pb-TSP average (jjg/m3) over 12 months
of the year, 2008-2010.
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0.05-
0.04-
E
§> 0.03 -
0.02 -
0.01 -
0.00-
Jan
—I—
Feb
Mar
Apr
May
Jun
—r~
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 3-24 Monthly non-source-oriented lead-TSP average (jjg/m3) over 12
months of the year, 2008-2010.
0.028
0.026 -
0.024 -
0.022
0.020 -
0.018-
0.016
0.014-
0.012
0.010-
0.008 -
0.006 -
0.004
0.002 -
0.000
Jun Jul
Month
Nov Dec
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 3-25 Monthly lead-PMio average (jjg/m ) over 12 months of the year,
2007-2009.
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0.010-
0.009 -
0.008 -
pj. 0.007-
E
1	0.006-
¦I 0.005 -
2
o 0.004 -
o
c
° 0.003 -
0.002 -
0.001 -
0.000 -
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 3-26 Monthly lead-PM2.5 average (jjg/m3) over 12 months of the year,
2007-2009.
For both Pb-PMn, (Figure 3-25) and Pb-PM2 5, (Figure 3-26) 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.
3.5.3 Size Distribution of Lead-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 3.4.1), 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,
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.
T
T
^ X
T
T
—i	1	
Nov Dec
Jan Feb Mar Apr
May
Jul
Aug Sep
Oct
Month
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Because atmospheric lifetime is dependent on particle size, as described in
Section 3.3.1.3 and in U.S. EPA (2009). 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. 2009; 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).
Size-selective monitoring data from the literature are examined in this section. Size
distribution data enhances understanding of the relationship between sources and
characteristics of airborne Pb-bearing PM and hence informs 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 subsections
below, 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.
3.5.3.1 Airborne Pb Near Metals Industries
Size distributions of Pb-bearing PM have been measured near several active and closed
industrial 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 Multi-Orifice Uniform Deposit Impactor
(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 (.un. In the industrial area, 27%
of Pb-PM were larger than PMi0 (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). Bein et al. (2006)
measured the size distribution of Pb in PM from the Pittsburgh Supersite using rapid
single particle mass spectrometry and a MOUDI. The Pittsburgh, PA, Supersite had
seventeen major PM sources within a 24-km radius; 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
(Peknev 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
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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. Concentrations of Pb-PM3 2 measured by the MOUDI were 5.6 ng/m3 in
October, 2001 and 5.0 ng/m3 in March, 2002; measurements by the hi-vol were 4.4 ng/m3
in October, 2001 and 6.5 ng/m3 in March, 2002. Singh et al. (2002) measured the mass
distribution of Pb-PM10 in the coarse and fine PM size ranges (cut points range:
0.10-10 |_im) for the Downey industrial site along the Alameda industrial corridor in Los
Angeles 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 with 34% of the particles in this size bin. 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 |im bin
and the 350 nm-1 |im bin, comprising 42% and 26% of the mass measured as PMi0,
respectively. Pb in the fine range only comprised 13% of the particles in the 100-350 nm
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.
Industrial operations associated with Pb emissions include metal works and incineration.
Dall'Osto et al. (2008) measured the size distribution of Pb emissions from a steel works
facility in a coastal town within the United Kingdom (U.K.). A MOUDI was employed to
measure Pb concentrations in the coarse to fine PM size range (cut points range:
0.196-18 The size distribution was multimodal with a primary mode around 1 |im at
a concentration of 40 ng/m3, a secondary mode around 300 nm at a concentration of 25
ng/m3, and a very small additional mode around 5 |im at a concentration of 7 ng/m3. This
multimodal distribution was thought to be associated with sintering and steel working
processes, from which Pb was emitted. Weitkamp et al. ("2005) measured Pb-bearing
PM2 5 concentrations across the river from a coke plant in the Pittsburgh, PA area. Pb was
measured to comprise 0.088% of the PM2 5 mass (avg. Pb-PM2 5: 53 ng/m3), and
background-corrected Pb mass concentration was reasonably correlated with background-
corrected PM2 5 mass concentration (R2 = 0.55). Pekey et al. (2010) measured PM2 5 and
PM10 concentrations in a heavily industrialized area of Kocaeli City, Turkey and obtained
an average PM2 5 concentration of 47 ng/m3 during summer and 72 ng/m3 during winter.
Average PM10 concentration was 78 ng/m3 during summer and 159 ng/m3 during winter,
to produce PM2 5/PMi0 ratios of 0.60 during summer and 0.45 during winter.
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3.5.3.2
Airborne Pb Near Roadways
Traffic-induced turbulence can be a cause of resuspension of Pb-bearing particles from
deposited contemporaneous wheel weights, industrial emissions, or historic sources. Pb
in resuspended road dust exhibits a bimodal size distribution, with fine mode likely from
vapor phase condensation from combustion engines and the coarse mode from vehicle
wear (U.S. EPA. 2006b). However, mass is predominantly associated with coarse PM.
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 m from the 1-405 Freeway in
the southern California air basin; data from Sabin et al. (2006b) are displayed in Figure
3-27; note that the Noll rotary impactor has been shown to sample with high efficiency
for particles up to 100 (.un in diameter (Noll. 1970). For both the urban background and
near-road sites, the largest fraction was from PM sampled below the 6 (j,m cut point, but
the near-road Pb-PM distribution appeared bimodal with a mode in the largest size
fraction. Overall size fractions, the near-road site had a Pb concentration of 17 ng/m3,
compared with an urban background concentration of 9.7 ng/m3. 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 et al. (2011) used an eight-stage
MOUDI (cut point range: 0.18-18 |_im) to measure roadside PM 5 m from the New Jersey
Turnpike in Carlstadt, NJ and speciated the samples. They observed a bimodal
distribution of the Pb concentration in summer and a trimodal distribution in winter. 85%
of the Pb-PM mass smaller than 2.5 |_im during the summer and 68% was within the
2.5 |_im fraction in the winter. Pb-PM mass measured in this study ranged from 1.2-2.8
ng/m3. Similarly, Zereini et al. (2005) observed that roughly 80% of particle-bound Pb
measured with a MOUDI was smaller than 5.8 |_im for an urban main street (avg. conc.:
33 ng/m3), and more than 90% were smaller than 5.8 |_im for a rural area included in that
study (avg. conc.: 12 ng/m3). However, in a study of automotive emissions in a traffic
tunnel, Lough et al. (2005) measured that 85% of Pb measured with a MOUDI was in the
PMio, with just 39% in the PM2 5 fraction and 20% in the PMi fraction. In a near-road
study conducted in Raleigh, NC, Hays et al. (2011) note that the concentration of Pb in
ultrafine, fine, and coarse size ranges was roughly constant at 50 mg/kg; similar to Lough
et al. (2005). mass concentrations were 0.4 ± 0.4 ng/m3, 1.4 ± 0.6 ng/m3, and 0.1 ± 0.02
ng/m3 for PM10_2.5, PM2 5.01, and PM0.1, 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 3.2.2.6). 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
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on-road gasoline emissions. Chen et al. (2010b) measured Pb in PM10.2 5, PM2 5.01, and
PM0.1 using a MOUDI at a roadside location and in a tunnel in Taipei, Taiwan in 2008.
While roadside and tunnel concentrations of PM10 and PM2 5 were roughly equivalent
around 20-30 ng/m3, Pb in PM0.1 was approximately 15 times higher in the tunnel (during
the hours 9:00 AM - 9:00 PM) than by the roadside (tunnel: 20 (ig/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 at various
traffic and background sites in Birmingham, U.K.. captured at the stage below a 0.5 |im
cutpoint and on the 1.5-3.0 (j,m 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-PM05 and 10 ng/m3 for Pb-PMi 5_3 0. Roadside concentrations were low.
During the day, Birmili et al. (2006) measured 0.4 ng/m3 for Pb-PM0 5 and 1.2 ng/m3 for
Pb-PMi 5.3 0. At night, roadside concentrations reduced to 0.17 ng/m3 for Pb-PM0 5 and
0.6 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-PM0 5 and 0.84 ng/m3 for Pb-PMi 5.
3 0. Remote background concentrations were on 0.16 ng/m3 for Pb-PM0 5 and 0.03 ng/m3
for Pb-PMi 5.3 0. Briiggemann et al. (2009) measured roadside distribution of Pb in PM in
Dresden, Germany 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 |im size
bin. When winds came from the east, the total concentration was approximately
22 ng/m3, compared with a concentration of approximately 13 ng/m3 when winds came
from the west. Total winter concentrations of Pb were approximately 26 ng/m3, while
summertime concentrations were roughly 11 ng/m3.
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I Urban Background
¦ Near-Road
Source: Adapted, with permission of Elsevier Publishing, Sabin et al. (2006b).
Figure 3-27 Comparison of urban background and near-road size fractions of
lead-bearing PM.
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
PMio at a roadside sampler in Birmingham, U.K.. The size distribution was unimodal
with approximately 2% of the Pb mass (0.5 ng/nr) above the 10 |im cut point, 53% of the
Pb mass (14 ng/m3) in the 0.2-1.0 (mi bin, and 24% (7 ng/m3) collected below the 0.2 (mi
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 (|3 = 0.067,
R2 = 0.38) as well as PM10 (|3 = 0.26, R2 = 0.35). Reactivity of NOx may contribute to the
somewhat low values for R2 in these models. Bruggemann et al. (2009) observed a
unimodal Pb size distribution with 51% of the mass in the 0.42-1.2 (.im size bin.
Observed Pb-PMn, concentration was 17 ng/nr\ During winter, Pb concentrations were
more than twice as high as during the summer (0.42-1.2 (.im mode, winter: 50 ng/nr':
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summer: 20 ng/m3), and they were also higher when winds blew from the east
(0.42-1.2 (.im mode, east: 60 ng/m3; west: 25 ng/m3). Briiggemann et al. (2009) suggested
that this finding reflected coal burning sources rather than road dust resuspension. Wang
et al. (2006d) used a nine-stage cascade impactor (cut point range: 0.43-11 |_im) 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 (j,m and the 3.3-4.7 |im 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.
3.5.3.3 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 ^m and a slightly smaller secondary peak at
5.6 (mi. 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, average Pb-PM25 concentration was 30 ng/m3
during the day and 92 ng/m3 at night, and average Pb-PMi0 concentration was 59 ng/m3
during the day and 162 ng/m3 at night, to yield PM2.5/PM10 ratios of 0.51 during the day
and 0.57 at night. At the suburban site, average Pb-PM2 5 concentration was 15 ng/m3
during the day and 34 ng/m3 at night, and average Pb-PMi0 concentration was 24 ng/m3
during the day and 42 ng/m3 at night, to yield PM2.5/PMi0 ratios of 0.63 during the day
and 0.81 at night. Rural measurements were only made for Pb-PMi0 and averaged
6 ng/m3 during the day and 5 ng/m3 at night. Goforth et al. (2006) measured TSP and
PM2 5 in rural Georgia and observed a PM2 5 concentration of 6 ng/m3 and a TSP
concentration of 15 ng/m3. Makkonen et al. (2010) measured concentrations of Pb in
PMi, PM2 5, and PMi0 during a spate of wildfires in rural southeastern Finland. They
found that the ratio of PM1/PM10 varied substantially from day to day (examples provided
of 64% on 8/14/07 and 35% on 8/25/07, with PM2 5/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.
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3.5.4
Lead 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 S02 may suggest common industrial sources. Correlation
between Pb and N02 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. Spearman correlations of monitored Pb-TSP concentrations with
concentrations of other criteria pollutants are summarized in Figure 3-28 for 2007-2008
data from 129 monitoring sites, and in Figure 3-29 for 2009 data from 16 monitoring
sites. 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, ranging from 44
non-source-oriented sites for the association of Pb with S02 to 81 sites for the association
of Pb with PMi0; in Figure 3-28, and fewer for each copollutant in Figure 3-29. Each of
these figures illustrates co-pollutant correlations across the U.S. Additionally, seasonal
correlations between Pb and co-pollutants are provided in Figure 3-60 through Figure
3-65 in the Section 3.8, with seasonal co-pollutant measurement data from the literature
(Table 3-31). As evident in each figure, there were considerably fewer source-oriented
sites available for co-located comparisons.
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Source, S02 -
US Overall

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Non-Source, S02 -

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Spearman Correlation Coefficient
Figure 3-28 Correlations of monitored Pb-TSP concentration with copollutant
concentrations, 2007-2008.
Non-Source, S02 -
US Overall


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Spearman Correlation Coefficient
Figure 3-29 Correlations of monitored Pb-TSP concentration with co-pollutant
concentrations, 2009.
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Overall, Pb was most strongly associated with PM25, PM10 and N02 (median R = 0.38 to
0.41), with positive Spearman correlation coefficients observed at nearly all sites.
However, Pb was just as strongly associated with CO in fall and winter (median p = 0.48
to 0.58). Such correlations may suggest common sources affecting the pollutants. Overall
correlation coefficients between Pb and S02 and between Pb and CO were also positive
at most sites, but associations were generally weaker (median p = 0.29 for CO, 0.17 for
S02). The poorest associations were observed between Pb and 03 (median p = 0.00).
Although the overall associations of Pb concentration with PM10 and PM2 5
concentrations were similar, the association with PMi0 was stronger in the spring and the
association with PM2 5 stronger in summer and fall. The strongest associations between
Pb and other criteria pollutants were observed in fall and winter, and the weakest in
summer.
The relationship between Pb and other species in PM2 5 is explored in Figure 3-30, which
describes data from 3 years of CSN results using Pearson R. These data provide a
national perspective on relationships between the various bulk and elemental species
monitored in the CSN network. The strongest association was with Zn (median R = 0.51).
K, Cu, and Br concentrations also exhibited moderately strong associations with Pb
concentrations (median R = 0.40 to 0.41). Such correlations may suggest some common
sources affecting the pollutants. For example, as described in Section 3.2.2.1, piston-
engine aircraft emit Pb as PbBr2 so this source may explain the covariation in Pb and Br
concentrations at the CSN sites. Other species more useful as diagnostic indicators of
crustal, general combustion, industrial emission, and coal combustion processes exhibited
weaker, but still remarkable associations with Pb, including Fe (median R = 0.34), EC
(median R = 0.32), crustal elements (median R = 0.32), Mn (median R = 0.32), and OC
(median R = 0.30). Summer associations between Pb and other species tended to be
weaker than in other seasons, with a Correlation Coefficient greater than R = 0.35
observed only for Zn (median R = 0.37). The weakest associations were with non-volatile
N03", CI, As, Na+2, Hg, Na, and Ni (median R = -0.02 to 0.10).
A few recent studies have used speciation techniques to characterize Pb and other
components of PMi0, PM2 5, 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.
Wojas and Almquist (2007) used ICPMS to characterize trace metals in PM2 5, PM10, and
TSP samples obtained in Oxford, OH. They observed that Pb correlation varied
substantially with other elements across size distributions. Correlations were very high
between Pb and several PMi0 metals, and only Pb was only highly correlated with Mg in
PM2 5. These results suggested that Pb and copollutants emanated from a variety of
sources including road dust and fuel combustion (Table 3-10). Variations in the relative
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proportions of Pb-containing compounds may account for the difference in Pb solubility
in aerosols (von Schneidemesser et al. 2010; Tan et al.. 2006; Fernandez Espinosa and
Ternero-Rodriguez. 2004).
Murphy et al. (2008) studied weekly patterns of metals and other aerosol components
using data collected from 2000 to 2006 at Interagency Monitoring of Protected Visual
Environments (IMPROVE) sites that are described further in Section 3.4.2.2. The authors
concluded that Pb concentrations were impacted by piston aircraft emissions, particularly
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 research station,
Switzerland, about 5% of analyzed aerosol particles in PMi 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, Mg, Al, K, Fe, Zn, Mo, Ag, Ba). 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.
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Table 3-10 Correlations between Pb and copollutants, measured in TSP, PM10,
and PM2.5
Element3
TSP
PM10
PM2.5
Al
0.54
-0.21
0.20
As
0.14
-0.24
-0.03
Ca
0.4
0.96
-0.52
Cd
0.95
0.93
0.20
Co
0.59
-0.19
0.47
Cr
0.04
0.99
0.20
Cu
0.85
0.84
-0.74
Fe
0.43
0.99
-0.51
K
0.81
0.90
NA
Mg
0.43
0.97
0.80
Mn
0.71
0.97
NA
Mo
0.69
0.99
0.04
Ni
0.08
0.99
-0.67
Sb
0.81
0.85
-0.84
Si
0.36
0.99
0.26
V
0.28
-0.17
-0.43
Zn
0.92
0.99
0.27
aHigh correlations (r > 0.7) are shown in bold italics
Source: Wojas and Almquist (2007)
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—
0 O Ol- -
	1 1 I--
	1 003 O O
As
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--H 1
Zh	H
CI
nvol N03-
-
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---1 1 1-
1 H
	» "X" O O000
0 O
Q	»0	Q	'O	Q
Q	O	Q	:
Note: "nvol" = non-volatile, "vol" = volatile, and organic carbon (OC) samples were blank-adjusted.
Figure 3-30 Correlations of monitored lead-PM2.5 concentration with
copollutant concentrations, 2007-2009.
3.5.5	Background Lead Concentrations
1	To inform NAAQS decisions, EPA has historically estimated risk due to pollutant
2	concentrations above background concentrations. In previous NAAQS reviews, a specific
3	definition of background concentrations was used and referred to as policy relevant
4	background (PRB). In those previous reviews, PRB concentrations were defined by EPA
5	as those concentrations that would occur in the U.S. in the absence of anthropogenic
6	emissions in continental North America (CNA), defined here as the U.S., Canada, and
7	Mexico. For this document, we have focused on the sum of those background
8	concentrations from natural sources everywhere in the world and from anthropogenic
9	sources outside CNA. Background concentrations so defined facilitate separation of
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pollution that can be controlled by U.S. regulations or through international agreements
with other countries from those that are judged to be generally uncontrollable by the U.S.
Over time, consideration of potentially broader ranging international agreements may
lead to alternative determinations of which Pb source contributions should be considered
by EPA as part of background.
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 [see pp 1-14, (U.S. EPA.
1986a)I. A 1980 National Academy of Sciences report estimated that average natural
background levels of airborne Pb might range from 0.02 to 0.5 ng/m3 (NAS Committee
on Lead in the Human Environment. 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 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.
Chemistry-transport models are not available for estimating background concentrations
of airborne Pb is problematic because chemistry-transport model calculations are not
available as they are for 03 or for PM2 5 and they would have to include the entire size
range for particle-bound Pb. The only data that might be relevant are from the IMPROVE
network, but only in the PM2 5 size fraction. Pb in this size fraction is most amenable to
long range transport because particles in the PM2 5 size fraction have a much longer
atmospheric lifetime compared to larger size fractions. It is impossible to obtain reliable
estimates of background concentrations solely on the basis of measurements of PM2 5,
PM10-2 5, or PM10. It is preferable to quantify contributions from both background and
non-background sources by using compositional data in techniques such as source
apportionment modeling.
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 and Lassen Volcanic Park
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from 1988 to 2000. Geometric mean 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 or that these contributions were
preferentially scrubbed out during transport. Large enrichments in 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 in a NOAA aircraft
elevated more than 2 km AGL. 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 PM2 5, 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 CD (U.S. EPA. 2004)
and 2009 PM ISA (U.S. EPA. 2009)1. Data obtained at four eastern IMPROVE sites
(Moosehorn NWR, ME; Acadia NP, ME; Swanquarter, NC; Cape Romain NWR, 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. It is assumed that there is no fractionation between size ranges in
this approach. 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 ppm (Reuer and Weiss. 2002). There is
substantial variation with location depending on composition, in particular on the
abundances of U and 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 m
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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).
Concentrations of the Pb content of soils can be used with estimates of the crustal
component of PM25, PM10.2.5 (which is mainly crustal), PM10, and TSP produced by wind
erosion of natural surfaces to estimate contributions to Pb concentrations in these size
fractions. U.S. annual average PM10 concentrations in some arid counties most affected
by windblown dust in the western U.S. are ~ 20 |ig/nr\ If it is assumed that these levels
of PM10 are entirely due to natural wind erosion without any anthropogenic contribution,
an estimate of ~ 0.3 ng/m3 for the contribution of wind erosion on natural surfaces to Pb
in PMio 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. These estimates exceed estimates of natural background presented in the
1986 AQCD (U.S. EPA. 1986a) and the National Academy of Sciences Report (NAS
Committee on Lead in the Human Environment. 1980) by a factor of 2 to 20. The more
recent estimate still indicates that background airborne Pb concentrations are well below
current ambient concentrations.
3.6 Ambient Lead 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 NAWQA
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survey, maxima concentrations in surface waters, sediments, and fish tissues were
30 |ig/L, 12,000 mg/kg, and 23 mg/kg, respectively, compared with median values of
0.50 |ig/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.
3.6.1 Soils
Several studies suggest that soil can act as a reservoir for contemporaneous and historical
Pb emissions. In a recent review of soil data collected from 90 U.S. cities, Mielke et al.
(2010a') cited studies, some of which were 35 years old but many from the last 15 years,
reporting that median soil Pb concentrations ranged from 16 to 189 mg/kg. Soil Pb was
thought to originate from present-day sources, such as industry, debrided paint, and
piston-engine aircraft emissions, as well as historic sources, such as on-road gasoline
emissions, as described in Section 3.2. 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 Eco-SSL guidance document, the
U.S. EPA (U.S. EPA. 2007d. 2003^ augmented these data with observations from an
additional 13 studies conducted between 1982 and 1997, most of them limited to one
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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 at 2 m depths. Aelion et al. (2009) compared Pb soil levels in urban and two
rural locations in the southeastern U.S. (specific sites were not disclosed). The urban and
one of the rural areas had significantly high prevalence of mental retardation or
developmental disabilities (MMDD), while the other urban area did not. Aelion et al.
(2009)	found that soil Pb concentration in the rural area without prevalence of MMDD
ranged from 2.1-53 mg/kg dry basis, while in the rural area with high prevalence of
MMDD, soil Pb ranged from 1.6-140 mg/kg; in the urban area with high MMDD, soil Pb
ranged from 2.4-288 mg/kg. Biasioli et al. (2006) contrasted urban and rural soils 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.
In North American forest soils, Pb concentrations have decreased substantially since the
phase out of leaded motor vehicle gasoline. Evans et al. (2005) observed Pb
concentrations ranging from 60 to 200 mg/kg in Vermont 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 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 m
in the northeastern U.S. concentrations 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).
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 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 of 66-493 mg/kg outside
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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 3.5.4) 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 from surrounding towns
with those from the Bunker Hill Superfund remediation site in Idaho. Median background
soil Pb concentrations was 48 mg/kg, while the median soil Pb concentration at Bunker
Hill was 245 mg/kg.
Recent studies of brownfield soils have shown variable Pb concentrations. Van
Herwijnen et al. (2007) measured soils near a defunct Zn smelter in Avonmouth, UK 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. 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 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 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 had a smoothed "bull's eye" pattern, which suggested
that the Pb in soil is continually resuspended and deposited within the urban area so that
smooth air and soil concentration gradients emanating from the city center could be
created over time. 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 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 be smoothed overtime from resuspension and deposition with contributions
from historic sources of on-road gasoline (Section 3.2.2.6) and Pb paint (Section 3.2.2.7).
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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 in Figure 3-31. Brown et al. (2008) also measured soil Pb concentration along three
transects of Lubbock, TX and observed that soil Pb decreased with increasing distance
from the city center, which was the oldest part of the city.
Median Soil Pb (mg/kg)
3-99
100-199
200 - 299
1300 - 399
400 - 499
500 - 599
600 - 699
700 - 799
H 800 - 899
¦¦ 900 - 999
H 1000-1768
Source: Reprinted with permission of Elsevier Publishing, Mielke et al. (2007a)
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.
Figure 3-31 Map of median Pb content in soil in New Orleans.
Mielke et al. (2008) compared soil Pb concentrations 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 in New Orleans, Mielke et al. (2007b) observed median
and maxima deposition values of roughly 25,000 and 265,000 |ag/m\ respectively.
Median and maxima surface soil Pb concentrations were observed to be 1,000 and
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20,000 mg/kg, respectively. Clark et al. (2006) performed isotopic analysis on urban
garden soils in an area of Boston, MA with no large industrial sources of Pb and
estimated that 60% of the soil Pb could be attributed to historic Pb on-road gasoline
emissions in an urban area of the city while 14% was attributed to historic Pb on-road
gasoline emissions in a suburban area. The remainder of the Pb was attributed to paint
degradation. However, the Clark et al. (2006) model assumed that the only two sources of
soil Pb were paint and historic Pb on-road gasoline emissions. So, the percentages
provided are upper limits of the contribution of those two sources. Additional discussion
of historic sources of Pb is provided in Section 3.2.2.7. Isotope 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 based on a stratified random sampling design
that accounted for land use factors. They compared soil metals within 100 m 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, Wu et al.
(2010) observed that soil Pb concentration 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 m buffer, and length of local
road within a 20 m buffer in which the sample was obtained were significant predictors
of Pb. Home age within 30 m of a soil sample and road length within 3,000 m 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 3.2.2.6) were not considered in
this study. Amato et al. (2009) observed that deposited PM10 onto roadways, measured as
dust samples, in Barcelona, Spain was differentially enriched with Pb. Pb concentration
in PM10 was highest at ring roads (229 mg/kg) and in the city center (225 mg/kg),
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followed by demolition and construction sites (177 mg/kg) and near a harbor
(100 mg/kg). Joshi et al. (2009) also 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.
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.. 2006b; Weiss et al.. 2006). Caravanos et al. (2006b) used GIS to
examined 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
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 m 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 (i/nr). 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).
Soil Pb variability depends on the strength and prevalence of nearby sources. Griffith et
al. (2002) investigated spatial autocorrelation of soil Pb concentration at three sites: urban
Syracuse, NY, rural Geul River, The Netherlands, and an abandoned Pb Superfund site in
Murray, UT. In both Syracuse and Geul River, the soil Pb concentrations were not
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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 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.
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 m of the sand, and 6.4% was
recovered at a depth of 0.3-0.45 m. Pb oxides were observed to be the dominant species
in the contaminated sand. Berthelot et al. (2008) studied soil Pb concentrations in
grounds used for testing military tanks and munitions and measured soil Pb levels to
range from 250 to 2,000 mg/kg.
3.6.2 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
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
preindustrial 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 3-32. 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
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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 (Landers et al.. 2008). 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
3-32 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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Source: danders et al.. 20081
Note: (DENA = Denali, GAAR = Gates of the Arctic, GLAC = Glacier, MORA = Mount Ranier, NOAT = Noatak, OLYM = Olympic,
ROMO = Rocky Mountain, SEKI = Sequoia and Kings Canyon)
Figure 3-32 WACAP data for Pb concentration in sediment at eight National
Parks and/or Preserves.
In a survey of 35 reservoirs and lakes in 16 continental U.S. states, Van Metre et al.
(2006) collected data from sediment cores extending back as far as the early 1800s, and
up to 2001. For most locations, they were able to match at least three bodies of water in
rural (designated as 'reference"), light urban, and dense urban settings. In reference
bodies of water, the median sediment Pb concentration corresponding to the 1990s was
48 mg/kg. It was 56 mg/kg in sediments from light urban bodies, and 214 mg/kg in dense
urban ones (Mahler et al.. 2006). Using the most distant past sediment records, Mahler et
al. (2006) provided approximations of concentrations attributable to anthropogenic
inputs. The median of these values for the 1990s were 28, 22, and 194 mg/kg in the
reference, light urban, and dense urban bodies of water, respectively.
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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. Table
3-11 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 sediment concentrations, and the highest average Pb sediment
concentrations were reported for the city of New Orleans. Figure 3-33 and Figure 3-34
illustrate 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 3-32, the data also illustrated that Pb concentrations in
sediment have declined in the U.S. since 1975 (Figure 3-34), prior to the phase-out of on-
road leaded gasoline.
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Table 3-11 Sediment concentrations in various cities, prior to 2005
City
Avg Pb Concentration (mg/kg)a
Pb Concentration Range (mg/kg)a
Baltimore, MD

1-10,900
Miami, FL
275
25-1,612
Mt. Pleasant, Ml
320
100-840
New Orleans, LA
784
31.7-5,195
New Orleans, LA (rural outskirts)
11
4.8-17.3
St. Louis, MO
427
35-1,860
Syracuse, NY
80
20-800
aDry weight basis,
Source: Reprinted with permission of the American Chemical Society, (Cobb et al.. 2006).
X X
Downstream
100	200	300	400	500	600
River km above Apalachicola Bay, FL
—Pb in streambed-sediment and reservoir-core samples
X Pb background in streambed-sediment and baseline reservoir-core samples
Source: Reprinted with permission of the American Chemical Society, Callenderand Rice (2000).
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.
Figure 3-33 Sediment core data (1992-1994) for the lakes and reservoirs along
the Apalachicola, Chattahoochee, and Flint River Basin (ACF),
which feeds from north of the Atlanta, GA metropolitan area into
the Gulf of Mexico at Apalachicola Bay in the Florida panhandle.
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160
—~—1975-1980

«¦-1980-1985
20
O
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Downstream
o
o
100
200
300
400
500
600
700
800
River km above Apalachicola Bay, FL
Source: Reprinted with permission of the American Chemical Society, Callenderand Rice (20001.
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.
Figure 3-34 Sediment core data (1975-1995) for the lakes and reservoirs along
the Apalachicola, Chattahoochee, and Flint River Basin (ACF),
which feeds from north of the Atlanta, GA metropolitan area into
the Gulf of Mexico at Apalachicola Bay in the Florida panhandle.
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
throughout the city on September 18, 2005 and analyzed for Pb by Presley et al. (2006)
were below the limit of detection. 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. However, several studies
noted no appreciable increase in Pb concentration within Lake Pontchartrain soils and
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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, iron
oxide-rich, and magnesium 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.
3.6.3 Rain
The only network that provides a long-term record of precipitation chemistry across the
U.S. is the National Trend Network operated by the National Atmospheric Deposition
Program, a cooperative effort between EPA and other federal as well as state, tribal, and
local government agencies, educational institutions, private companies, and
non-governmental agencies. Precipitation is monitored for pH, conductance, and major
cations and anions. A separate network, the Mercury Deposition Network, is operated by
the same organization and monitors Hg concentrations in rain and a separate network.
Neither of these networks includes Pb monitoring.
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 j^ig/L for
1989-90, 1990-91 and 2002-03, 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) (Shotvk and Krachler. 2010).
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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
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 3-35). 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%). However, similarly to the situation in the U.S., the greatest
reductions occurred by the late 1990s and only relatively minor reductions have occurred
thereafter; Figure 3-35.
Birkenes
Karvatn
Jergul/Karasjok
Osen
Lista
Source: Reprinted with permission of Pergamon Press, Berg et al. (20081
Figure 3-35 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
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decline in mean concentration from 4.92 ng/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
decline in annual total U.K. emissions of Pb to the atmosphere over each of these time
periods. Since the outright ban on the use of leaded on-road gasoline in 2000, however,
the ratio of Pb concentrations in rainwater to U.K. Pb emissions (metric tons) appears to
have stabilized to a near-constant value of 0.009 j^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.
3.6.4 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; Figure 3-36 (Landers et al.. 2008).
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
snow samples of 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
and these were again 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.
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Estimated contributions to the Pb in Lambert Glacier basin snow were -1% from rock
and soil dust (based on A1 concentrations) and -4.6% from volcanoes (based on the
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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DENA NOAT MORA NOCA OLYM SEKI GAAR ROMO GLAC
Source: danders et al.. 20081
(DENA = Denali, GAAR = Gates of the Arctic, GLAC = Glacier, MORA = Mount Ranier, NOCA = North Cascades, NOAT = Noatak,
OLYM = Olympic, ROMO = Rocky Mountain, SEKI = Sequoia and Kings Canyon)
Figure 3-36 Box plots illustrating Pb concentration in snow melt at nine
National Parks and Preserves.
1	Lee et al. (2008b) collected 42 snow samples during the period autumn 2004-summer
2	2005 from a 2.1 m snow pit at a high-altitude site on the northeast slope of Mount
3	Everest, Himalayas. Pb concentrations ranged from 5-530 pg/g with a mean value of 77
4	Pg/g- The mean value is clearly higher than the Hur et al. (2007) value for Antarctica but
5	is substantially lower than a mean concentration of 573 pg/g for snow from Mont Blanc,
6	France [1990-1991; Lee et al. (2008b)]. The mean Pb concentration for Mount Everest
7	snow was lower during the monsoon (28 pg/g) compared with the non-monsoon periods
8	(137 pg/g). From calculated enrichment factors (Pb/Alsnow:Pb/Alcrust), anthropogenic
9	inputs of Pb were partly important but soil and rock dust also contributed. The low Pb
10	concentrations during monsoon periods are thought to be attributable to low levels of
11	atmospheric loadings of crustal dusts. K. Lee et al. (2008b) noted that their conclusions
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differ from those in Kang et al. (2007). who stated that anthropogenic 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.
3.6.5 Natural Waters
Monitoring data for streams, rivers and lakes are summarized in periodic national
assessments of surface waters are carried out periodically by EPA, and they include
measurement of major biological and chemical stressors. 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 (Landers et al.. 2008); Figure 3-37. 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) and contemporary snow (Johnson and Parnell, as described in
Section 3.6.4). 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 the groundwaters indicate
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1	that the removal process is very effective. Shotyk and Krachler (2010) speculate that even
2	at these very low Pb concentrations, much if not most of the Pb is likely to be colloidal,
3	as suggested by the 2006 Pb AQCD (U.S. EPA. 2006b). Finally, Shotyk et al. (2010)
4	suggest that the pristine groundwaters from Simcoe County, Canada, provide a useful
5	reference level against which other water samples can be compared.

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NOAT
Source: (Landers et al.. 2008)
Note: (DENA = Denali, GLAC = Glacier, MORA = Mount Ranier, NOAT = Noatak, OLYM = Olympic)
Figure 3-37 Boxplots of Pb concentration in surface waters measured at five
National Parks and Preserves.
6	Although Pb concentrations in Kawagama Lake water were approaching "natural
7	values," the 2u6Pb/2"7Pb ratios for the samples that had the lowest dissolved Pb
8	concentrations of 10, 10 and 6 ng/L were 1.16, 1.15 and 1.16, respectively. These values
9	are inconsistent with those expected for natural Pb (the clay fraction from the lake
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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 3-12.
The range of Pb levels in saltwater 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 salt water was reported by
Leland and Kuwabara (1985) and 0.01 to 27 |ag Pb/L by 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 (U.S. EPA. 2008b).
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Table 3-12 Pb concentrations from stream food webs in mining-disturbed areas
of Missouri and the western United States
Area
Total Pb in water (jjg/L)
Dissolved Pb (jjg/L)
Animas River. CO (Besser et al., 2001):
Reference Streams
<1.8
<0.2
Mining-disturbed areas
0.9-8.6
<0.1-6.9
Boulder River. MT (Faraa et al.. 2007):
Reference Streams
0.4 (colloidal)
0.3-0.4
Mining-disturbed areas
0.1-44
0.1-2
Coeur d'Alene River. ID (Clark. 2003: Faraa et al.. 1998):
Reference Streams
2-20
0.01-2
Mining-disturbed areas
6-2000
2-50
New Lead Belt. MO (Besser et al.. 2007: Brumbauqh et al., 2007):
Reference Streams

<0.01-1.6
Mining-disturbed areas

0.02-1.7
Adapted from: Mager (2012)
3.6.6 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 (Landers et al.. 2008)
have shown some Pb storage in lichens. Median Pb concentrations ranged from
0.3 mg/kg in Noatak National Park to 5 mg/kg in Glacier National Park, with substantial
variation in the Glacier and Olympic National Park samples; Figure 3-38. 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.
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—
i=
(D
1=
o
O
NOAT DENA GAAR MORA OLYM SEKI
GU\C
Source: danders et al.. 20081
Note: (DENA = Denali, GAAR = Gates of the Arctic, GLAC = Glacier, MORA = Mount Ranier, NOAT = Noatak, OLYM = Olympic,
SEKI = Sequoia and Kings Canyon)
Figure 3-38 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 2u6Pb/2u7Pb ratio for precipitation and mosses collected in
northeast Scotland. A study in the Vosges mountains 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 2u6Pb/2u7Pb 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
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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].
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. (2008b)
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 3-39. Both Pb concentrations and Pb isotope ratios declined with distance from the
smelter (Aznar et al.. 2008a; Aznar et al.. 2008b).
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Trends
1
0.8"
Regional
pollution
0.6-
0.4-
0.2-
Smelter
emissions
0
1950 1960 1970 1980 1990 2000

800 ¦




Humus

700 ¦



600 ¦


<0



o
5
500 ¦
<
~ (11)
c
400 -


c
--



8
£3
0l
300 ¦
200 -



100 ¦
~ (41)

0 ¦
A (29)
n (21J
2002
Sapwood-h ea rtwood
boundary
Source: Reprinted with permission of Elsevier Publishing, Aznar et al. (2008b)
Figure 3-39 Trends in regional pollution near a copper smelter in Canada and
Pb concentrations at the boundary of heartwood trees within
roughly 75 km of the smelter.
3.6.7 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
—i	1	1	1—
1950 1960 1970 1980
—I	1	
1990 2000
Tree rings
Distance class (km)
~-0-10
3- 10-20
6-20-40
>40
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isotopic ratios of Pb in Mytilus 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
re suspension 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).
3.6.8 Vertebrate Populations
Pb concentrations in fish fillet and liver were measured through the WACAP study in
eight National Parks and Preserves (Landers et al.. 2008). For fish fillet, Pb
concentrations ranged from 0.0033-0.30 mg/kg, with a median of 0.016 mg/kg. Liver
stores were several times higher, with Pb concentrations ranging from 0.011-0.97 mg/kg
and a median of 0.096 mg/kg. Pb concentrations in moose meat and liver were also
measured at the Denali National Park and Preserve as part of WACAP (Landers et al..
2008). Moose meat Pb concentrations ranged from 0.021-0.23 mg/kg with a median of
0.037 mg/kg. Pb concentrations in moose liver ranged from 0.025-0.11 mg/kg with a
median of 0.053 mg/kg. Boxplots of measured Pb concentrations in fish fillet and liver
are shown in Figure 3-40, and boxplots of measured Pb concentrations for moose meat
and liver are shown in Figure 3-41. 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). Still, the WACAP findings suggest some Pb accumulation in fish and
moose in these remote locations.
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E
¦-	-¦
1=
o
1=
CD
o
i=
O
O
Oj
1.0 H
0.S -
0.6 -
0.4 -
0.2 -
0.0 -
<
<
LL
LCl
o
o
<
<
I—
1—
z

-Z.
2
<
<
3
5
LXl
cc
<
<
>
>
LU
ID
<
<
O
o
O
o
_l
_l
lZj
Ci



&



z
o
o












jl|
¦li
.*!
O"
Jl;
¦ii

01
£

01
OI
01
iZ
_l
lZ
_l
iZ
_l
iZ
_l
IZ
_l
iZ
_l
O O
LL CC
LU LU
CO CO
(DENA = Denali, GAAR = Gates of the Arctic, GLAC = Glacier, MORA = Mount Ranier, NOAT = Noatak, OLYM = Olympic, ROMO =
Rocky Mountain, SEKI = Sequoia and Kings Canyon)
Source: (Landers et al.. 2008)
Figure 3-40 Boxplots of Pb concentration in fish fillet and liver measured at
eight National Parks and Preserves.
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1=
o
1=
CD
o
i=
O
O
Oj
o
r-4
kllT"
O
Liver
Meat
Source: (Landers et al.. 2008)
Figure 3-41 Boxplots of Pb concentration in moose meat and liver measured
at Denali National Park and Preserve.
3.7	Summary and Conculusions
3.7.1	Sources of Atmospheric Lead
1	The 2006 Pb AQCD (U.S. EPA. 2006b) documented the decline in ambient air Pb
2	emissions following the ban on alkyl-Pb additives for on-road gasoline. Pb emissions
3	declined by 98% from 1970 to 1995 and then by an additional 76% from 1995 to 2008, at
4	which time national Pb emissions were 970 tons/year. As was the case for the 2008
5	NAAQS review, piston-engine aircraft emissions currently comprise the largest share
6	(56%) of total atmospheric Pb emissions nationally; the 2008 NEI (U.S. EPA. 2011a)
7	estimated that 550 tons of Pb were emitted from piston-engine aircraft. Other sources of
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ambient air Pb, in approximate order of importance with regard to national totals, include
metals processing, fossil fuel combustion, other industrial sources, roadway related
sources, and historic Pb.
3.7.2 Fate and Transport of Lead
The atmosphere is the main environmental transport pathway for Pb, and on a global
scale atmospheric Pb is primarily associated with fine PM. On a global scale, Pb
associated with 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. Because Pb in fine
PM is typically fairly soluble, wet deposition is more important for fine Pb. In contrast,
Pb associated with coarse PM is usually insoluble, and removed by dry deposition.
However, local deposition fluxes are much higher near local industrial sources, and a
substantial amount of emitted Pb is deposited near sources, leading to high soil Pb
concentrations. Deposition is not an ultimate sink for Pb because particles are potentially
resuspended and redeposited many times before reaching a site where further transport is
unlikely, especially for dry deposition. Resuspension modeling has suggested that larger
particles (10-100 |_im) become more readily resuspended compared with smaller particles,
because smaller particles must overcome stronger adhesion forces to be resuspended by
air movement and lift forces are proportional to particle size to the approximate power of
1.5.
In water, Pb is transported as free ions, soluble chelates, or on surfaces of iron and
organic rich colloids. In 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 a disproportionate amount of Pb is removed by
runoff 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.
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. In anoxic sediments binding 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
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associated with OM or iron and manganese 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.
3.7.3 Ambient Lead Monitoring
Since the publication of the 2006 Pb AQCD for Pb (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 existing measurements, our
understanding of possible alternatives to existing Pb-TSP sampling technology, and our
understanding of particle size ranges of Pb particles occurring in different types of
locations have changed little in that time.
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. With the December
2010 completion of action on regulatorily required Pb monitoring, one-year of Pb-TSP
monitoring is also required near 15 specific airports to gather additional information on
the likelihood of NAAQS exceedances near airports due to combustion of leaded aviation
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gasoline. Non-source-oriented monitoring 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, PM10, or PM2 5 using high-volume or low-
volume samplers.
3.7.4 Ambient Air Lead Concentrations
Ambient air Pb concentrations have declined drastically over the period 1980-2010. The
median annual concentrations for all monitors have dropped by 97% from 0.87 |ig/nr' in
1980 to 0.03 (.ig/ni1 in 2010. While the sharpest drop in Pb concentration occurred during
1980-1990, a declining trend was observed between 1990 and 2010. Slightly smaller
reductions were observable among source-oriented Pb concentration (83%) and
non-source-oriented Pb data (91%) for 2000-2010. For source-oriented monitoring over
the period 2008-2010, the 3-month rolling average was measured to be above the level of
the NAAQS in twenty counties across the U.S.
Pb concentrations, seasonal variations, inter-monitor correlations, and wind data were
analyzed for six counties: Los Angeles County, CA, Hillsborough County, FL, Cook
County, IL, Jefferson County, MO, Cuyahoga County, OH, and Sullivan County, TN.
These sites were selected for analysis because they contained a mix of source-oriented
and non-source-oriented monitors in urban areas. Spatial and temporal variability of Pb
concentrations in each county were commonly high. Meteorology, distance from sources
with respect to the monitors, and source strength all appeared to influence the level of
concentration variability across time and space.
Pb concentrations exhibit varying degrees of association with other criteria pollutant
concentrations. Overall, Pb was moderately associated with PM2 5, PMi0 and N02, with
positive Spearman correlation coefficients observed at nearly all sites. However, Pb was
just as strongly associated with CO in fall and winter The poorest associations were
observed between Pb and 03. Among trace metals, the strongest association was with Zn.
Br, Cu, and K concentrations also exhibited moderate associations with Pb
concentrations. Such correlations may suggest some common sources affecting the
pollutants.
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3.7.5
Ambient Lead 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.
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3.8 Chapter 3 Appendix (Supplemental Material)
3.8.1 Variability across the U.S.
Table 3-13 Distribution of 1-month average Pb-TSP concentrations (jjg/m3)
nationwide, source-oriented monitors, 2008-2010
Year Sea" State/
son county
Stat
Cn°au"^ Site ID
name
N mo
means
N
sites
Mean
Min
1
5
10
25
50
75
90
95
99
max
Nationwide statistics
2008-2010


2,318

0.202
0.000
0.003
0.006
0.010
0.029
0.063
0.217
0.578
0.856
1.576
4.440
2008


548

0.318
0.004
0.004
0.013
0.024
0.050
0.110
0.348
0.841
1.240
2.557
4.440
2009


629

0.212
0.002
0.004
0.008
0.013
0.038
0.084
0.256
0.611
0.856
1.357
2.438
2010


1141

0.141
0.000
0.002
0.005
0.008
0.018
0.045
0.136
0.408
0.625
1.233
1.828
Winter


554

0.202
0.000
0.002
0.006
0.008
0.026
0.055
0.184
0.502
0.883
2.438
3.103
Spring


579

0.239
0.000
0.003
0.007
0.012
0.034
0.070
0.272
0.738
0.977
1.905
3.123
Summer


601

0.186
0.001
0.003
0.006
0.010
0.030
0.066
0.212
0.559
0.755
1.233
4.440
Fall


584

0.184
0.000
0.004
0.007
0.011
0.026
0.064
0.206
0.505
0.758
1.178
4.225
Nationwide statistics, pooled by site
2008-2010



111
0.161
0.002
0.003
0.008
0.013
0.031
0.056
0.177
0.441
0.687
0.997
1.275
2008



47
0.323
0.007
0.007
0.022
0.028
0.055
0.148
0.419
0.890
1.205
1.540
1.540
2009



54
0.214
0.007
0.007
0.013
0.018
0.043
0.090
0.343
0.669
0.849
0.921
0.921
2010



101
0.140
0.002
0.003
0.005
0.013
0.030
0.052
0.165
0.392
0.586
0.888
1.185
Winter



108
0.156
0.000
0.003
0.006
0.009
0.021
0.048
0.160
0.475
0.879
1.130
1.488
Spring



110
0.185
0.002
0.002
0.010
0.015
0.027
0.057
0.210
0.568
0.921
1.189
1.548
Summer



111
0.148
0.002
0.003
0.006
0.012
0.025
0.050
0.153
0.430
0.696
0.882
1.031
Fall



110
0.152
0.002
0.004
0.009
0.013
0.034
0.062
0.168
0.421
0.616
1.081
1.189
Statistics for individual counties (2008-2010)
04013
AZ
Maricopa
6
1
0.0218
0.009
0.009
0.009
0.009
0.014
0.021
0.028
0.038
0.038
0.038
0.038
06025
CA
Imperial
33
1
0.016
0.004
0.004
0.006
0.009
0.011
0.015
0.019
0.025
0.032
0.035
0.035
06037
CA
Los
Angeles
224
8
0.0098
0.000
0.000
0.000
0.002
0.006
0.010
0.012
0.017
0.020
0.038
0.044
06065
CA
Riverside
72
2
0.0077
0.000
0.000
0.003
0.004
0.006
0.008
0.010
0.010
0.012
0.014
0.014
06071
CA
San Bern-
ardino
71
2
0.0091
0.001
0.001
0.003
0.004
0.007
0.010
0.012
0.014
0.014
0.022
0.022
08005
CO
Arapahoe
9
1
0.0120
0.004
0.004
0.004
0.004
0.007
0.012
0.016
0.018
0.018
0.018
0.018
08031
CO
Denver
12
1
0.006
0.003
0.003
0.003
0.004
0.005
0.005
0.006
0.008
0.008
0.008
0.008
13089
GA
DeKalb
10
1
0.003
0.002
0.002
0.002
0.002
0.003
0.003
0.004
0.005
0.006
0.006
0.006
17031
IL
Cook
288

0.020
0.010
0.010
0.010
0.010
0.012
0.016
0.025
0.034
0.040
0.060
0.070
17117
IL
Macoupin
24
1
0.0101
0.010
0.010
0.010
0.010
0.010
0.010
0.010
0.010
0.010
0.012
0.012
17119
IL
Madison
36
1
0.019
0.010
0.010
0.010
0.010
0.012
0.016
0.020
0.032
0.053
0.066
0.066
17143
IL
Peoria
36
1
0.011
0.010
0.010
0.010
0.010
0.010
0.010
0.010
0.013
0.013
0.014
0.014
17163
IL
Saint
Clair
36
1
0.0206
0.010
0.010
0.010
0.012
0.014
0.018
0.026
0.032
0.038
0.054
0.054
18089
IN
Lake
36
1
0.015
0.005
0.005
0.005
0.005
0.008
0.014
0.019
0.030
0.033
0.049
0.049
18097
IN
Marion
35
1
0.006
0.002
0.002
0.002
0.003
0.004
0.005
0.008
0.010
0.012
0.013
0.013
18163
IN
Vander-
33
2
0.0045
0.001
0.001
0.001
0.002
0.003
0.004
0.005
0.006
0.010
0.010
0.010
February 2012
3-143
Draft - Do Not Cite or Quote

-------
Year ?e0an"
son
State/
County
Stat
County
name
Site ID
N mo
means
N
sites
Mean
Min
1
5
10
25
50
75
90
95
99
max

25025
MA
Suffolk

31
2
0.009
0.004
0.004
0.004
0.005
0.007
0.008
0.010
0.013
0.016
0.020
0.020

26081
Ml
Kent

12
1
0.005
0.003
0.003
0.003
0.003
0.005
0.005
0.006
0.008
0.008
0.008
0.008

26163
Ml
Vtoyne

36
2
0.011
0.003
0.003
0.003
0.004
0.005
0.009
0.015
0.021
0.023
0.032
0.032

27017
MN
Carlton

12
1
0.000
0.000
0.000
0.000
0.000
0.000
0.000
0.000
0.000
0.000
0.000
0.000

27037
MN
Dakota

118
5
0.004
0.000
0.000
0.000
0.000
0.000
0.002
0.005
0.008
0.010
0.017
0.036

27053
MN
Hennepin

126
4
0.0032
0.000
0.000
0.000
0.000
0.000
0.002
0.005
0.006
0.008
0.010
0.044

27075
MN
Lake

10
1
0.000
0.000
0.000
0.000
0.000
0.000
0.000
0.000
0.000
0.000
0.000
0.000

27123
MN
Ramsey

71
3
0.006
0.000
0.000
0.000
0.000
0.002
0.004
0.008
0.013
0.020
0.028
0.028

27137
MN
Saint
Louis

72
2
0.0015
0.000
0.000
0.000
0.000
0.000
0.000
0.002
0.004
0.006
0.010
0.010

27163
MN
\Nash-
ington

72
3
0.0016
0.000
0.000
0.000
0.000
0.000
0.000
0.003
0.004
0.005
0.006
0.006

29097
MO
Jasper

12
1
0.013
0.007
0.007
0.007
0.007
0.009
0.012
0.017
0.018
0.019
0.019
0.019

29187
MO
Saint
Francois

24
2
0.0327
0.008
0.008
0.009
0.009
0.018
0.032
0.039
0.054
0.080
0.089
0.089

29189
MO
Saint
Louis

33
1
0.0230
0.005
0.005
0.005
0.005
0.006
0.008
0.050
0.050
0.050
0.066
0.066

36047
NY
Kings

24
1
0.013
0.010
0.010
0.010
0.010
0.011
0.012
0.014
0.018
0.020
0.020
0.020

39017
OH
Butler

34
1
0.006
0.002
0.002
0.003
0.004
0.004
0.005
0.007
0.008
0.009
0.009
0.009

39029
OH
Columbia
na

107
3
0.0155
0.004
0.004
0.006
0.007
0.008
0.011
0.018
0.027
0.034
0.065
0.136

39035
OH
Cuyahoga

107
3
0.0143
0.004
0.004
0.006
0.007
0.009
0.012
0.017
0.024
0.030
0.041
0.041

39049
OH
Franklin

36
1
0.009
0.004
0.004
0.005
0.005
0.007
0.009
0.011
0.013
0.014
0.016
0.016

39143
OH
Sandusky

12
1
0.0048
0.003
0.003
0.003
0.003
0.004
0.005
0.006
0.006
0.007
0.007
0.007

39167
OH
Vtoshingt
on

54
2
0.0048
0.002
0.002
0.002
0.003
0.003
0.005
0.006
0.007
0.008
0.010
0.010

40115
OK
Ottawa

16
2
0.012
0.003
0.003
0.003
0.005
0.006
0.013
0.017
0.021
0.025
0.025
0.025

42003
PA
Allegheny

36
1
0.0105
0.000
0.000
0.000
0.000
0.004
0.009
0.015
0.019
0.024
0.053
0.053

42021
PA
Cambria

23
1
0.046
0.040
0.040
0.040
0.040
0.040
0.040
0.044
0.054
0.058
0.128
0.128

42045
PA
Delaware

20
1
0.0432
0.040
0.040
0.040
0.040
0.040
0.043
0.046
0.047
0.048
0.048
0.048

42101
PA
Philadelp
hia

24
1
0.0210
0.011
0.011
0.011
0.012
0.014
0.020
0.027
0.033
0.033
0.039
0.039

42129
PA
V\festmore
land

24
1
0.0419
0.037
0.037
0.040
0.040
0.040
0.040
0.042
0.050
0.050
0.053
0.053

48061
TX
Cameron

35
1
0.0041
0.002
0.002
0.003
0.003
0.003
0.004
0.005
0.006
0.007
0.009
0.009

48141
TX
El Paso

68
3
0.021
0.014
0.014
0.014
0.014
0.015
0.017
0.019
0.029
0.056
0.087
0.087

48201
TX
Harris

32
1
0.005
0.003
0.003
0.003
0.004
0.004
0.005
0.006
0.007
0.008
0.010
0.010

48479
TX
V\febb

29
1
0.013
0.004
0.004
0.005
0.006
0.008
0.011
0.018
0.026
0.028
0.035
0.035

49035
UT
Salt Lake

12
1
0.0173
0.003
0.003
0.003
0.006
0.009
0.011
0.024
0.040
0.043
0.043
0.043

51087
VA
Henrico

7
1
0.007
0.003
0.003
0.003
0.003
0.003
0.004
0.005
0.024
0.024
0.024
0.024
Statistics for individual sites where overall average monthly mean > national 90th percentile (2008-2010)




11090003
32

0.525
0.054
0.054
0.083
0.164
0.252
0.402
0.798
1.053
1.117
1.277
1.277




060371405
36

0.671
0.100
0.100
0.188
0.235
0.285
0.359
0.771
2.086
2.501
2.880
2.880




290930016
36

0.670
0.166
0.166
0.186
0.219
0.330
0.466
0.726
0.974
2.435
4.225
4.225




290930021
36

0.681
0.082
0.082
0.084
0.095
0.194
0.650
0.879
1.437
2.438
2.557
2.557




290990004
36

0.997
0.256
0.256
0.307
0.408
0.598
0.918
1.236
1.690
1.905
2.416
2.416




290990015
21

1.275
0.340
0.340
0.421
0.646
0.756
1.118
1.349
2.440
3.103
3.123
3.123




2909900203
31

0.687
0.191
0.191
0.195
0.297
0.368
0.620
0.808
1.111
1.280
2.220
2.220




29099002f
21

0.719
0.084
0.084
0.141
0.359
0.572
0.666
0.876
1.164
1.168
1.553
1.553




2909900223
31

0.441
0.140
0.140
0.171
0.208
0.303
0.409
0.599
0.683
0.754
0.861
0.861




290999001a
24

0.850
0.186
0.186
0.208
0.319
0.449
0.845
1.071
1.382
1.558
1.623
1.623




2909990053
24

0.986
0.155
0.155
0.250
0.330
0.558
0.864
1.487
1.802
1.828
1.985
1.985
February 2012
3-144
Draft - Do Not Cite or Quote

-------
Year ?e0an"
son
State/ „ .
_ . Stat
County
County
name
Site ID
N mo
means
N
sites
Mean
Min
1 6 10
26
60
76
90
96
99 max



480850009s
36

0.601
0.137
0.137 0.138 0.185
0.420
0.579
0.757
1.101
1.178
1.564 1.564
aSites listed in the bottom six rows of the table fall in the upper 90th percentile of the data pooled by site.
Table 3-14 Distribution of 1-month average Pb-TSP concentrations (jjg/m3)
nationwide, non-source-oriented monitors, 2008-2010
State/
v	 Seaso _ .
Year Count
n
y
Stat
e
County
name
Site
ID
N mo
means
N
sites
Mean
Min
1
5
10
25
50
75
90
95
99
max
Nationwide statistics
2008-2010



2290

0.0120
0.000
0.000
0.000
0.002
0.004
0.010
0.015
0.026
0.040
0.052
0.136
2008



685

0.0126
0.000
0.000
0.000
0.002
0.005
0.010
0.015
0.029
0.040
0.052
0.066
2009



768

0.0114
0.000
0.000
0.000
0.002
0.004
0.010
0.014
0.023
0.040
0.048
0.128
2010



837

0.0120
0.000
0.000
0.000
0.000
0.004
0.009
0.016
0.026
0.036
0.054
0.136
Winter



556

0.0109
0.000
0.000
0.000
0.001
0.004
0.008
0.013
0.022
0.038
0.056
0.087
Spring



574

0.0122
0.000
0.000
0.000
0.002
0.004
0.009
0.015
0.028
0.040
0.052
0.128
Summer



584

0.0119
0.000
0.000
0.000
0.002
0.005
0.010
0.016
0.026
0.040
0.050
0.057
Fall



576

0.0129
0.000
0.000
0.000
0.002
0.005
0.010
0.016
0.026
0.040
0.053
0.136
Nationwide statistics, pooled by site
2008-2010




88
0.0120
0.000
0.000
0.001
0.002
0.005
0.011
0.016
0.024
0.033
0.046
0.046
2008




59
0.0125
0.001
0.001
0.002
0.003
0.006
0.010
0.016
0.024
0.043
0.051
0.051
2009




66
0.0116
0.000
0.000
0.001
0.002
0.004
0.010
0.014
0.024
0.032
0.050
0.050
2010




73
0.0119
0.000
0.000
0.001
0.001
0.005
0.010
0.018
0.023
0.028
0.046
0.046
Winter




88
0.0115
0.000
0.000
0.001
0.001
0.004
0.009
0.016
0.025
0.038
0.048
0.048
Spring




86
0.0119
0.000
0.000
0.001
0.002
0.004
0.009
0.016
0.027
0.032
0.059
0.059
Summer




88
0.0117
0.000
0.000
0.000
0.001
0.005
0.010
0.016
0.026
0.034
0.043
0.043
Fall




88
0.0130
0.000
0.000
0.001
0.003
0.005
0.011
0.017
0.028
0.031
0.054
0.054
Statistics for individual counties (2008-2010)
04013
AZ
Maricopa

6
1
0.0218
0.009
0.009
0.009
0.009
0.014
0.021
0.028
0.038
0.038
0.038
0.038
06025
CA
Imperial

33
1
0.0162
0.004
0.004
0.006
0.009
0.011
0.015
0.019
0.025
0.032
0.035
0.035
06037
CA
Los
Angeles

224
8
0.0098
0.000
0.000
0.000
0.002
0.006
0.010
0.012
0.017
0.020
0.038
0.044
06065
CA
Riverside

72
2
0.0077
0.000
0.000
0.003
0.004
0.006
0.008
0.010
0.010
0.012
0.014
0.014
06071
CA
San Bern-
ardino

71
2
0.0091
0.001
0.001
0.003
0.004
0.007
0.010
0.012
0.014
0.014
0.022
0.022
08005
CO
Arapahoe

9
1
0.0120
0.004
0.004
0.004
0.004
0.007
0.012
0.016
0.018
0.018
0.018
0.018
08031
CO
Denver

12
1
0.0056
0.003
0.003
0.003
0.004
0.005
0.005
0.006
0.008
0.008
0.008
0.008
13089
GA
DeKalb

10
1
0.0033
0.002
0.002
0.002
0.002
0.003
0.003
0.004
0.005
0.006
0.006
0.006
17031
IL
Cook

288

0.0195
0.010
0.010
0.010
0.010
0.012
0.016
0.025
0.034
0.040
0.060
0.070
17117
IL
Macoupin

24
1
0.0101
0.010
0.010
0.010
0.010
0.010
0.010
0.010
0.010
0.010
0.012
0.012
17119
IL
Madison

36
1
0.0188
0.010
0.010
0.010
0.010
0.012
0.016
0.020
0.032
0.053
0.066
0.066
17143
IL
Peoria

36
1
0.0105
0.010
0.010
0.010
0.010
0.010
0.010
0.010
0.013
0.013
0.014
0.014
17163
IL
Saint Clair

36
1
0.0206
0.010
0.010
0.010
0.012
0.014
0.018
0.026
0.032
0.038
0.054
0.054
18089
IN
Lake

36
1
0.0150
0.005
0.005
0.005
0.005
0.008
0.014
0.019
0.030
0.033
0.049
0.049
18097
IN
Marion

35
1
0.0058
0.002
0.002
0.002
0.003
0.004
0.005
0.008
0.010
0.012
0.013
0.013
18163
IN
Vander-
burgh

33
2
0.0045
0.001
0.001
0.001
0.002
0.003
0.004
0.005
0.006
0.010
0.010
0.010
25025
MA
Suffolk

31
2
0.0087
0.004
0.004
0.004
0.005
0.007
0.008
0.010
0.013
0.016
0.020
0.020
26081
Ml
Kent

12
1
0.0053
0.003
0.003
0.003
0.003
0.005
0.005
0.006
0.008
0.008
0.008
0.008
February 2012
3-145
Draft - Do Not Cite or Quote

-------
State/
.. Seaso „
Year _ Count
n
y
Stat
e
County Site
name ID
N mo
means
N
sites
Mean
Min
1
5
10
25
50
75
90
95
99
max
26163
Ml
Vtoyne
36
2
0.0112
0.003
0.003
0.003
0.004
0.005
0.009
0.015
0.021
0.023
0.032
0.032
27017
MN
Carlton
12
1
0.0000
0.000
0.000
0.000
0.000
0.000
0.000
0.000
0.000
0.000
0.000
0.000
27037
MN
Dakota
118
5
0.0035
0.000
0.000
0.000
0.000
0.000
0.002
0.005
0.008
0.010
0.017
0.036
27053
MN
Hennepin
126
4
0.0032
0.000
0.000
0.000
0.000
0.000
0.002
0.005
0.006
0.008
0.010
0.044
27075
MN
Lake
10
1
0.0000
0.000
0.000
0.000
0.000
0.000
0.000
0.000
0.000
0.000
0.000
0.000
27123
MN
Ramsey
71
3
0.0062
0.000
0.000
0.000
0.000
0.002
0.004
0.008
0.013
0.020
0.028
0.028
27137
MN
Saint Louis
72
2
0.0015
0.000
0.000
0.000
0.000
0.000
0.000
0.002
0.004
0.006
0.010
0.010
27163
MN
Vtoshingto
n
72
3
0.0016
0.000
0.000
0.000
0.000
0.000
0.000
0.003
0.004
0.005
0.006
0.006
29097
MO
Jasper
12
1
0.0125
0.007
0.007
0.007
0.007
0.009
0.012
0.017
0.018
0.019
0.019
0.019
29187
MO
Saint
Francois
24
2
0.0327
0.008
0.008
0.009
0.009
0.018
0.032
0.039
0.054
0.080
0.089
0.089
29189
MO
Saint Louis
33
1
0.0230
0.005
0.005
0.005
0.005
0.006
0.008
0.050
0.050
0.050
0.066
0.066
36047
NY
Kings
24
1
0.0131
0.010
0.010
0.010
0.010
0.011
0.012
0.014
0.018
0.020
0.020
0.020
39017
OH
Butler
34
1
0.0055
0.002
0.002
0.003
0.004
0.004
0.005
0.007
0.008
0.009
0.009
0.009
39029
OH
Columbian
a
107
3
0.0155
0.004
0.004
0.006
0.007
0.008
0.011
0.018
0.027
0.034
0.065
0.136
39035
OH
Cuyahoga
107
3
0.0143
0.004
0.004
0.006
0.007
0.009
0.012
0.017
0.024
0.030
0.041
0.041
39049
OH
Franklin
36
1
0.0092
0.004
0.004
0.005
0.005
0.007
0.009
0.011
0.013
0.014
0.016
0.016
39143
OH
Sandusky
12
1
0.0048
0.003
0.003
0.003
0.003
0.004
0.005
0.006
0.006
0.007
0.007
0.007
39167
OH
Vtoshingto
n
54
2
0.0048
0.002
0.002
0.002
0.003
0.003
0.005
0.006
0.007
0.008
0.010
0.010
40115
OK
Ottawa
16
2
0.0124
0.003
0.003
0.003
0.005
0.006
0.013
0.017
0.021
0.025
0.025
0.025
42003
PA
Allegheny
36
1
0.0105
0.000
0.000
0.000
0.000
0.004
0.009
0.015
0.019
0.024
0.053
0.053
42021
PA
Cambria
23
1
0.0463
0.040
0.040
0.040
0.040
0.040
0.040
0.044
0.054
0.058
0.128
0.128
42045
PA
Delaware
20
1
0.0432
0.040
0.040
0.040
0.040
0.040
0.043
0.046
0.047
0.048
0.048
0.048
42101
PA
Philadelphi
a
24
1
0.0210
0.011
0.011
0.011
0.012
0.014
0.020
0.027
0.033
0.033
0.039
0.039
42129
PA
V\fest-
moreland
24
1
0.0419
0.037
0.037
0.040
0.040
0.040
0.040
0.042
0.050
0.050
0.053
0.053
48061
TX
Cameron
35
1
0.0041
0.002
0.002
0.003
0.003
0.003
0.004
0.005
0.006
0.007
0.009
0.009
48141
TX
El Paso
68
3
0.0206
0.014
0.014
0.014
0.014
0.015
0.017
0.019
0.029
0.056
0.087
0.087
48201
TX
Harris
32
1
0.0053
0.003
0.003
0.003
0.004
0.004
0.005
0.006
0.007
0.008
0.010
0.010
48479
TX
V\febb
29
1
0.0134
0.004
0.004
0.005
0.006
0.008
0.011
0.018
0.026
0.028
0.035
0.035
49035
UT
Salt Lake
12
1
0.0173
0.003
0.003
0.003
0.006
0.009
0.011
0.024
0.040
0.043
0.043
0.043
51087
VA
Henrico
7
1
0.0066
0.003
0.003
0.003
0.003
0.003
0.004
0.005
0.024
0.024
0.024
0.024
Statistics for individual sites where overall average monthly mean > national 90th percentile (2008-2010)


170310022
36

0.0330
0.012
0.012
0.014
0.016
0.020
0.033
0.040
0.056
0.062
0.070
0.070


170310026
36

0.0282
0.014
0.014
0.014
0.018
0.020
0.028
0.034
0.044
0.048
0.052
0.052


170316003
36

0.0249
0.012
0.012
0.014
0.018
0.020
0.026
0.031
0.033
0.038
0.040
0.040


2918700063
12

0.0383
0.009
0.009
0.009
0.015
0.024
0.035
0.042
0.080
0.089
0.089
0.089


2918700073
12

0.0271
0.008
0.008
0.008
0.009
0.013
0.026
0.035
0.052
0.054
0.054
0.054


420210808a
23

0.0463
0.040
0.040
0.040
0.040
0.040
0.040
0.044
0.054
0.058
0.128
0.128


4204500023
20

0.0432
0.040
0.040
0.040
0.040
0.040
0.043
0.046
0.047
0.048
0.048
0.048


4212900073
24

0.0419
0.037
0.037
0.040
0.040
0.040
0.040
0.042
0.050
0.050
0.053
0.053


481410002a
23

0.0236
0.016
0.016
0.016
0.016
0.017
0.018
0.021
0.033
0.056
0.087
0.087
'Sites listed in the bottom six rows of the table fall in the upper 90th percentile of the data pooled by site.
February 2012
3-146
Draft - Do Not Cite or Quote

-------
Table 3-15 Distribution of 3-month moving average Pb-TSP concentrations
(jjg/m3) nationwide, source-oriented monitors, 2008-2010
Year
Season
State/
County
State
County name
Site ID
N monthly
means
N
sites
Mean
Min
1
5
10
25
50
75
90
95
99
max
Nationwide statistics3
2008-
2010





2,112

0.2134
0.000
0.004
0.010
0.014
0.035
0.079
0.250
0.600
0.881
1.555
2.889
2008





537

0.3225
0.005
0.006
0.016
0.028
0.056
0.129
0.385
0.900
1.197
2.452
2.889
2009





600

0.2177
0.004
0.005
0.011
0.016
0.040
0.090
0.292
0.622
0.799
1.217
2.070
2010





975

0.1507
0.000
0.002
0.008
0.012
0.024
0.052
0.173
0.436
0.694
1.055
1.375

Winter




443

0.2366
0.003
0.004
0.011
0.014
0.040
0.083
0.272
0.647
0.963
2.070
2.621

Spring




535

0.2376
0.000
0.004
0.011
0.014
0.035
0.078
0.323
0.642
0.999
2.017
2.889

Summer




572

0.2022
0.002
0.003
0.009
0.015
0.034
0.077
0.240
0.580
0.869
1.261
2.163

Fall




562

0.1835
0.002
0.004
0.009
0.013
0.033
0.078
0.220
0.521
0.714
1.186
2.456
Nationwide statistics, pooled by site
2008-
2010






106
0.1671
0.002
0.003
0.012
0.015
0.030
0.059
0.173
0.577
0.717
1.009
1.316
2008






47
0.3300
0.007
0.007
0.024
0.029
0.056
0.154
0.461
0.814
1.284
1.639
1.639
2009






54
0.2203
0.007
0.007
0.013
0.019
0.042
0.086
0.311
0.632
0.840
0.886
0.886
2010






96
0.1415
0.002
0.002
0.008
0.013
0.027
0.053
0.163
0.407
0.619
1.110
1.110

Winter





104
0.1700
0.003
0.004
0.011
0.013
0.025
0.055
0.171
0.522
0.827
1.097
1.324

Spring





101
0.1004
0.001
0.002
0.013
0.016
0.028
0.060
0.186
0.502
0.874
1.231
1.740
Summer 106 0.1597 0.002 0.004 0.010 0.015 0.028 0.058 0.174 0.520 0.788 0.080 1.104

Fall



105
0.1538 ! 0.002
0.004
0.009 ! 0.012 10.032 J 0.066
0.170
0.462 10.630 10.960
1.161
Statistics for individual counties (2008-2010)


01109
AL
Pike

25
1
0.5771
0.223
0.223
0.247
0.256
0.302
0.574
0.719
1.088
1.178
1.210
1.210


06037
CA
Los Angeles
131
4

0.2521
0.023
0.023
0.036
0.041
0.055
0.078
0.237
0.543
0.832
2.452
2.489


12057
FL
Hillsborough
79
3

0.1940
0.011
0.011
0.015
0.037
0.063
0.110
0.249
0.423
0.582
1.770
1.770


13015
GA
Bartow

11
1
0.0125
0.009
0.009
0.009
0.009
0.011
0.013
0.014
0.015
0.016
0.016
0.016


13215
GA
Muscogee
12
1

0.0367
0.014
0.014
0.014
0.020
0.022
0.031
0.052
0.066
0.070
0.070
0.070


17031
IL
Cook

9
1
0.1364
0.068
0.068
0.068
0.068
0.109
0.135
0.150
0.241
0.241
0.241
0.241


17115
IL
Macon

10
1
0.0806
0.048
0.048
0.048
0.052
0.067
0.080
0.088
0.117
0.123
0.123
0.123


17119
IL
Madison

36
1
0.1346
0.027
0.027
0.035
0.036
0.063
0.113
0.207
0.283
0.341
0.416
0.416


17143
IL
Peoria

20
2
0.0121
0.010
0.010
0.010
0.010
0.011
0.012
0.014
0.015
0.016
0.016
0.016


17195
IL
Whiteside
10
1

0.0191
0.012
0.012
0.012
0.014
0.016
0.019
0.022
0.025
0.025
0.025
0.025


17201
IL
Wnnebago
9
1

0.0356
0.019
0.019
0.019
0.019
0.021
0.027
0.057
0.063
0.063
0.063
0.063


18035
IN
Delaware
57
2

0.2866
0.053
0.053
0.059
0.073
0.090
0.159
0.246
0.495
1.867
2.163
2.163


18089
IN
Lake

46
2
0.0305
0.007
0.007
0.011
0.012
0.016
0.027
0.036
0.040
0.057
0.129
0.129


18097
IN
Marion

66
2
0.0198
0.005
0.005
0.006
0.007
0.011
0.014
0.025
0.036
0.043
0.079
0.079


18127
IN
Porter

10
1
0.0131
0.007
0.007
0.007
0.007
0.007
0.013
0.017
0.020
0.022
0.022
0.022


19155
IA
Pottawattamie
12
1

0.1581
0.034
0.034
0.034
0.067
0.113
0.153
0.220
0.246
0.263
0.263
0.263


20169
KS
Saline

9
1
0.2286
0.096
0.096
0.096
0.096
0.107
0.231
0.324
0.421
0.421
0.421
0.421
21151 KY Madison 10 1 0.0212 0.013 0.013 0.013 0.014 0.015 0.017 0.024 0.037 0.040 0.040 0.049


26067
Ml
Ionia

10
1
0.1980
0.106
0.106
0.106
0.110
0.128
0.212
0.259
0.273
0.284
0.284
0.284


27003
MN
Anoka

10
1
0.0161
0.006
0.006
0.006
0.008
0.010
0.013
0.022
0.029
0.031
0.031
0.031


27037
MN
Dakota

36
1
0.2026
0.068
0.068
0.072
0.088
0.104
0.216
0.248
0.357
0.415
0.429
0.429


27145
MN
Stearns

10
1
0.0032
0.000
0.000
0.000
0.001
0.002
0.004
0.004
0.005
0.005
0.005
0.005


29093
MO
Iron

158
6
0.3465
0.010
0.011
0.019
0.022
0.033
0.142
0.549
0.901
1.167
2.076
2.456


29099
MO
Jefferson
423
19

0.4925
0.023
0.033
0.050
0.071
0.187
0.385
0.723
0.989
1.186
2.017
2.889


29179
MO
Reynolds
40
4

0.0397
0.012
0.012
0.014
0.015
0.017
0.031
0.057
0.087
0.089
0.100
0.100
February 2012
3-147
Draft - Do Not Cite or Quote

-------
Year
Season
State/
County
State
County name
Site ID
N monthly
means
N
sites
Mean
Min
1
5
10
25
50
75
90
95
99
max


31053
NE
Dodge

7
1
0.0474
0.019
0.019
0.019
0.019
0.020
0.060
0.067
0.072
0.072
0.072
0.072


31127
NE
Nemaha

6
1
0.0447
0.019
0.019
0.019
0.019
0.024
0.032
0.075
0.087
0.087
0.087
0.087


36071
NY
Orange

99
3
0.0271
0.003
0.003
0.004
0.005
0.007
0.027
0.037
0.068
0.075
0.086
0.086


39035
OH
Cuyahoga
70
3

0.0905
0.006
0.006
0.010
0.011
0.021
0.050
0.122
0.221
0.287
0.531
0.531


39051
OH
Fulton

30
1
0.1609
0.025
0.025
0.027
0.046
0.054
0.092
0.254
0.354
0.453
0.567
0.567


39091
OH
Logan

100
4
0.0499
0.004
0.004
0.004
0.006
0.033
0.047
0.072
0.090
0.095
0.100
0.100


39101
OH
Marion

8
1
0.0379
0.032
0.032
0.032
0.032
0.034
0.037
0.042
0.047
0.047
0.047
0.047
39151 OH Stark 9 1 o.oiso 0.015 0.015 0.015 0.015 0.016 0.018 0.019 0.023 0.023 0.023 0.023


39155
OH
Trumbull

6
1
0.0080
0.005
0.005
0.005
0.005
0.006
0.008
0.010
0.011
0.011
0.011
0.011


40121
OK
Pittsburg

9
1
0.0021
0.002
0.002
0.002
0.002
0.002
0.002
0.002
0.003
0.003
0.003
0.003


41071
OR
Yamhill

10
1
0.0166
0.009
0.009
0.009
0.011
0.013
0.016
0.019
0.026
0.027
0.027
0.027


42003
PA
Allegheny
20
2

0.0414
0.009
0.009
0.011
0.012
0.017
0.030
0.054
0.099
0.120
0.138
0.138


42007
PA
Beaver

41
3
0.1160
0.043
0.043
0.052
0.056
0.083
0.114
0.159
0.170
0.187
0.206
0.206


42011
PA
Berks

105
6
0.0995
0.038
0.039
0.041
0.045
0.051
0.078
0.145
0.183
0.197
0.242
0.251


42045
PA
Delaware
10
1

0.0447
0.043
0.043
0.043
0.043
0.043
0.045
0.046
0.047
0.047
0.047
0.047


42055
PA
Franklin

7
1
0.0447
0.043
0.043
0.043
0.043
0.043
0.045
0.046
0.046
0.046
0.046
0.046


42063
PA
Indiana

10
1
0.0447
0.043
0.043
0.043
0.043
0.043
0.044
0.046
0.049
0.049
0.049
0.049


42079
PA
Luzerne

6
1
0.1078
0.084
0.084
0.084
0.084
0.085
0.103
0.135
0.137
0.137
0.137
0.137


42129
PA
Westmoreland
10
1

0.0434
0.041
0.041
0.041
0.042
0.042
0.044
0.044
0.046
0.046
0.046
0.046


47093
TN
Knox

44
2
0.0165
0.007
0.007
0.009
0.009
0.012
0.016
0.020
0.023
0.027
0.035
0.035


47163
TN
Sullivan

118
4
0.0554
0.030
0.030
0.033
0.035
0.039
0.045
0.060
0.100
0.125
0.134
0.168


48085
TX
Collin

108
3
0.3101
0.048
0.051
0.070
0.085
0.120
0.217
0.469
0.682
0.753
1.189
1.262


51770
VA
Roanoke City
10
1

0.0466
0.013
0.013
0.013
0.016
0.019
0.026
0.097
0.108
0.109
0.109
0.109


55117
Wl
Sheboygan
10
1

0.0897
0.012
0.012
0.012
0.034
0.058
0.076
0.126
0.164
0.170
0.170
0.170


72013
PR
Arecibo

10
1
0.1725
0.059
0.059
0.059
0.068
0.129
0.194
0.213
0.241
0.245
0.245
0.245
Statistics for individual sites where overall average monthly mean > national 90th percentile
2008-2010)
| 111 1011090003 |25 I |o.5771 lo.223 lo.223 lo.247 10.256 lo.302 lo.574 |o.719 11.088 |l.178 |l.210 |l.210
I 111 1060371405 136 I |o.7174 lo.188 (0.188 |o.234 jo.237 |o.309 !0.476 |o.791 (2.178 (2.452 (2.489 (2.489





290930016
36

0.6682
0.207
0.207
0.258
0.313
0.418
0.543
0.634
1.167
2.076
2.456
2.456





290930021
36

0.6950
0.173
0.173
0.192
0.218
0.346
0.689
0.954
1.214
1.275
1.937
1.937





290990004
36

1.0090
0.640
0.640
0.655
0.699
0.775
0.913
1.081
1.555
2.011
2.017
2.017





290990015"
21

1.3162
0.612
0.612
0.632
0.743
0.921
1.074
1.258
2.621
2.634
2.889
2.889





290990020"
29

0.6680
0.452
0.452
0.471
0.482
0.555
0.651
0.754
0.891
0.943
0.989
0.989





290990021"
21

0.7317
0.429
0.429
0.435
0.507
0.547
0.685
0.900
0.999
1.013
1.141
1.141





290999001"
22

0.8413
0.587
0.587
0.592
0.600
0.699
0.845
0.963
1.061
1.100
1.204
1.204





290999005"
22

0.9875
0.612
0.612
0.630
0.644
0.783
0.995
1.220
1.271
1.278
1.375
1.375





480850009"
36

0.6068
0.196
0.196
0.268
0.335
0.469
0.585
0.704
0.965
1.189
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.
bSites listed in the bottom six rows of the table fall in the upper 90th percentile of the data pooled by site.
February 2012
3-148
Draft - Do Not Cite or Quote

-------
Table 3-16 Distribution of 3-month moving average Pb-TSP concentrations
(jjg/m3) nationwide, non-source-oriented monitors, 2008-2010
Year
Season
State/
County
State
County name
Site ID
N monthly
means
N
sites
Mean
Min
1
5
10
25
50
75
90
95
99
max
Nationwide statistics
2008-
2010





2,164

0.0120
0.000
0.000
0.001
0.002
0.005
0.010
0.015
0.025
0.037
0.048
0.073
2008





663

0.0130
0.000
0.000
0.001
0.002
0.005
0.011
0.016
0.027
0.040
0.050
0.055
2009





727

0.0114
0.000
0.000
0.001
0.002
0.004
0.009
0.014
0.024
0.038
0.043
0.073
2010





774

0.0118
0.000
0.000
0.001
0.001
0.005
0.010
0.016
0.025
0.035
0.047
0.057

winter




494

0.0113
0.000
0.001
0.001
0.002
0.005
0.009
0.014
0.023
0.037
0.050
0.055

spring




548

0.0119
0.000
0.000
0.001
0.002
0.005
0.009
0.015
0.025
0.036
0.050
0.073

summer




565

0.0121
0.000
0.000
0.001
0.002
0.005
0.010
0.016
0.026
0.037
0.046
0.053

fall




557

0.0126
0.000
0.000
0.001
0.002
0.005
0.011
0.017
0.027
0.037
0.048
0.057
Nationwide statistics, pooled by site
2008-
2010






86
0.0120
0.000
0.000
0.001
0.002
0.005
0.010
0.016
0.024
0.034
0.046
0.046
2008






59
0.0127
0.001
0.001
0.002
0.003
0.005
0.011
0.016
0.024
0.043
0.050
0.050
2009






65
0.0117
0.001
0.001
0.001
0.003
0.004
0.010
0.014
0.026
0.031
0.049
0.049
2010






71
0.0118
0.000
0.000
0.001
0.001
0.005
0.010
0.017
0.022
0.028
0.045
0.045

Winter





84
0.0118
0.000
0.000
0.001
0.002
0.005
0.010
0.015
0.025
0.036
0.048
0.048

Spring





83
0.0118
0.000
0.000
0.001
0.002
0.004
0.010
0.015
0.025
0.034
0.059
0.059
Summer 86 o.ons o.ooo o.ooo 0.001 0.002 0.005 0.009 0.010 0.023 0.037 0.043 0.043
| Fall |


! 86 i 0.0126
0.000
0.000
0.001
0.002
0.005 0.011 10.016 10.026 0.030
0.046
0.046
Statistics for individual counties (2008-2010)


06025
CA
Imperial

31
1
0.0165
0.007
0.007
0.008
0.011
0.013
0.017
0.021
0.023
0.023
0.023
0.023


06037
CA
Los Angeles
218
8
0.0100
0.000
0.000
0.002
0.004
0.006
0.009
0.013
0.016
0.020
0.028
0.035
0.035


06065
CA
Riverside
72
2
0.0078
0.002
0.002
0.004
0.005
0.007
0.008
0.010
0.011
0.011
0.011
0.011



06071
CA
San Bernardino
69
2
0.0091
0.003
0.003
0.005
0.006
0.007
0.009
0.011
0.013
0.014
0.017
0.017



08005
CO
Arapahoe
7
1
0.0126
0.011
0.011
0.011
0.011
0.011
0.013
0.014
0.014
0.014
0.014
0.014



08031
CO
Denver

10
1
0.0054
0.004
0.004
0.004
0.004
0.005
0.006
0.006
0.006
0.006
0.006
0.006


13089
GA
DeKalb

8
1
0.0035
0.003
0.003
0.003
0.003
0.003
0.004
0.004
0.004
0.004
0.004
0.004


17031
IL
Cook

287
8
0.0196
0.010
0.010
0.010
0.010
0.012
0.017
0.025
0.033
0.038
0.047
0.051


17117
IL
Macoupin
24
1
0.0101
0.010
0.010
0.010
0.010
0.010
0.010
0.010
0.011
0.011
0.011
0.011
0.070


17119
IL
Madison

36
1
0.0188
0.010
0.010
0.010
0.011
0.014
0.016
0.022
0.036
0.036
0.039
0.039


17143
IL
Peoria

36
1
0.0105
0.010
0.010
0.010
0.010
0.010
0.010
0.011
0.012
0.012
0.013
0.013


17163
IL
Saint Clair
36
1
0.0204
0.012
0.012
0.012
0.014
0.016
0.020
0.024
0.029
0.033
0.036
0.036
0.014


18089
IN
Lake

36
1
0.0149
0.007
0.007
0.007
0.007
0.010
0.014
0.018
0.024
0.032
0.037
0.037


18097
IN
Marion

33
1
0.0056
0.003
0.003
0.003
0.003
0.004
0.005
0.007
0.009
0.010
0.011
0.011


18163
IN
Vanderburgh
31
2
0.0047
0.002
0.002
0.003
0.003
0.004
0.005
0.005
0.006
0.007
0.007
0.007
0.013


25025
MA
Suffolk

24
2
0.0093
0.005
0.005
0.006
0.006
0.008
0.009
0.011
0.013
0.015
0.016
0.016


26081
Ml
Kent

10
1
0.0055
0.004
0.004
0.004
0.005
0.005
0.006
0.006
0.006
0.006
0.006
0.006
26163 Ml \\Nayne 32 2 0.0119 0.004 0.004 0.004 0.005 0.005 0.012 0.017 0.021 0.023 0.024 0.024


27017
MN
Carlton

10
1
0.0000
0.000
0.000
0.000
0.000
0.000
0.000
0.000
0.000
0.000
0.000
0.000


27037
MN
Dakota

112
5
0.0036
0.000
0.000
0.001
0.001
0.001
0.003
0.005
0.007
0.012
0.013
0.015


27053
MN
Hennepin
124
4
0.0033
0.000
0.001
0.001
0.001
0.002
0.003
0.004
0.006
0.006
0.015
0.016
0.036


27075
MN
Lake

8
1
0.0000
0.000
0.000
0.000
0.000
0.000
0.000
0.000
0.000
0.000
0.000
0.000


27123
MN
Ramsey

65
3
0.0061
0.001
0.001
0.001
0.001
0.002
0.005
0.008
0.014
0.016
0.017
0.017


27137
MN
Saint Louis
70
2
0.0016
0.000
0.000
0.000
0.000
0.001
0.001
0.002
0.004
0.004
0.005
0.005
0.028


27163
MN
Washington
70
3
0.0017
0.000
0.000
0.000
0.000
0.001
0.001
0.003
0.004
0.004
0.005
0.005

February 2012
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Draft - Do Not Cite or Quote

-------
Year
Season
State/
County
State
County name
Site ID
N monthly
means
N
sites
Mean
Min
1
5
10
25
50
75
90
95
99
max


29097
MO
Jasper

10
1
0.0135
0.009
0.009
0.009
0.011
0.012
0.014
0.015
0.016
0.017
0.017
0.017


29187
MO
Saint Francois
21
2
0.0337
0.011
0.011
0.012
0.012
0.027
0.035
0.042
0.048
0.053
0.054
0.054
0.019


29189
MO
Saint Louis
33
1
0.0243
0.005
0.005
0.005
0.006
0.007
0.008
0.050
0.050
0.050
0.055
0.055



36047
NY
Kings

24
1
0.0131
0.011
0.011
0.011
0.011
0.012
0.013
0.014
0.016
0.018
0.019
0.019


39017
OH
Butler

30
1
0.0055
0.003
0.003
0.004
0.004
0.005
0.006
0.006
0.007
0.007
0.008
0.008


39029
OH
Columbiana
105
3
0.0148
0.005
0.005
0.007
0.008
0.010
0.013
0.017
0.021
0.028
0.054
0.057
0.009


39035
OH
Cuyahoga
105
3
0.0144
0.005
0.006
0.006
0.008
0.010
0.013
0.018
0.023
0.027
0.033
0.035

39049 OH Franklin 36 1 0.0092 0.005 0.005 0.005 0.005 0.008 0.010 0.011 0.011 0.012 0.012 0.012


39143
OH
Sandusky
10
1
0.0052
0.004
0.004
0.004
0.004
0.005
0.005
0.006
0.006
0.006
0.006
0.006
0.016


39167
OH
Washington
48
2
0.0047
0.002
0.002
0.002
0.003
0.004
0.004
0.006
0.007
0.007
0.008
0.008



40115
OK
Ottawa

12
2
0.0128
0.005
0.005
0.005
0.006
0.010
0.014
0.016
0.018
0.019
0.019
0.019


42003
PA
Allegheny
36
1
0.0101
0.000
0.000
0.000
0.000
0.007
0.012
0.014
0.016
0.018
0.025
0.025
0.025


42021
PA
Cambria

23
1
0.0459
0.040
0.040
0.040
0.040
0.040
0.041
0.046
0.069
0.070
0.073
0.073


42045
PA
Delaware
14
1
0.0427
0.040
0.040
0.040
0.040
0.040
0.042
0.045
0.046
0.047
0.047
0.047
0.128


42101
PA
Philadelphia
22
1
0.0214
0.013
0.013
0.014
0.014
0.018
0.022
0.025
0.029
0.029
0.030
0.030



42129
PA
Westmoreland
24
1
0.0417
0.037
0.037
0.040
0.040
0.040
0.041
0.043
0.046
0.047
0.048
0.048



48061
TX
Cameron
33
1
0.0042
0.002
0.002
0.003
0.003
0.004
0.004
0.005
0.005
0.006
0.006
0.006



48141
TX
El Paso

56
3
0.0212
0.014
0.014
0.014
0.015
0.016
0.018
0.023
0.038
0.040
0.040
0.040


48201
TX
Harris

30
1
0.0051
0.004
0.004
0.004
0.004
0.005
0.005
0.006
0.006
0.007
0.007
0.007


48479
TX
Webb

23
1
0.0121
0.006
0.006
0.007
0.007
0.008
0.010
0.016
0.021
0.022
0.026
0.026


49035
UT
Salt Lake
10
1
0.0145
0.007
0.007
0.007
0.007
0.008
0.011
0.016
0.032
0.036
0.036
0.036



48479
TX
Webb

29
1
0.0134
0.004
0.004
0.005
0.006
0.008
0.011
0.018
0.026
0.028
0.035
0.035


49035
UT
Salt Lake
12
1
0.0173
0.003
0.003
0.003
0.006
0.009
0.011
0.024
0.040
0.043
0.043
0.043



51087
VA
Henrico

7
1
0.0066
0.003
0.003
0.003
0.003
0.003
0.004
0.005
0.024
0.024
0.024
0.024
Statistics for individual sites where overall average monthly mean > national 90th percentile
2008-2010)





170310022
36

0.0335
0.016
0.016
0.018
0.026
0.028
0.032
0.038
0.047
0.048
0.051
0.051





170310026
36

0.0281
0.018
0.018
0.019
0.022
0.023
0.026
0.032
0.038
0.043
0.046
0.046
170316003 36 0.0245 0.015 0.015 0.015 0.017 0.020 0.025 0.028 0.031 0.035 0.030 0.030





291870006"
10

0.0412
0.017
0.017
0.017
0.026
0.035
0.043
0.048
0.054
0.054
0.054
0.054





291870007"
11

0.0268
0.011
0.011
0.011
0.012
0.012
0.028
0.035
0.036
0.041
0.041
0.041





291892003"
33

0.0243
0.005
0.005
0.005
0.006
0.007
0.008
0.050
0.050
0.050
0.055
0.055





420210808"
23

0.0459
0.040
0.040
0.040
0.040
0.040
0.041
0.046
0.069
0.070
0.073
0.073





420450002"
14

0.0427
0.040
0.040
0.040
0.040
0.040
0.042
0.045
0.046
0.047
0.047
0.047





421290007"
24

0.0417
0.037
0.037
0.040
0.040
0.040
0.041
0.043
0.046
0.047
0.048
0.048
directly compared to the Pb NAAQS for determination of compliance with the Pb NAAQS.
bSites listed in the bottom six rows of the table fall in the upper 90th percentile of the data pooled by site.
Table 3-17 Distribution of annual 1-month site maxima TSP Pb concentrations
(jjg/m3) nationwide, source-oriented monitors, 2008-2010
Year
Site ID - year
N (sites)
Mean
Min
1
6
10
26
60
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
February 2012
3-150
Draft - Do Not Cite or Quote

-------
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











1290930016-2008	4.2252
1290930021-2008	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

1.5528












290999001 a-2009

1.6228












290999001 a-2010

1.5576












290999005a-2009

1.9850












290999005a-2010

1.8278












480850009a-2008

1.5640











'Sites listed in the bottom eight rows of the table fall in the upper 90th percentile of the data pooled by site.
Table 3-18 Distribution of annual 1-month site maxima TSP Pb concentrations
(jjg/m3) nationwide, non-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

88
0.0284
0.000
0.000
0.004
0.006
0.010
0.020
0.041
0.057
0.070
0.136
0.136
2008

59
0.0232
0.004
0.004
0.005
0.006
0.010
0.016
0.033
0.053
0.058
0.066
0.066
2009

66
0.0210
0.003
0.003
0.005
0.006
0.008
0.014
0.026
0.040
0.056
0.128
0.128
2010

73
0.0233
0.000
0.000
0.002
0.004
0.008
0.015
0.029
0.049
0.065
0.136
0.136
Annual site max 1-month means >
= national 90th percentile (2008-2010)









170310022-2009

0.0700












170310022-2010

0.0620












171193007-2008

0.0660












291870006a-2010

0.0894












291892003a-2008

0.0660












390290019a-2010

0.1360












390290022a-2010

0.0652












420210808a-2008

0.0583












420210808a-2009

0.1280












481410002a-2010

0.0870












481410033a-2009

0.0570











aSites listed in the bottom eight rows of the table fall in the upper 90th percentile of the data pooled by site.
February 2012
3-151
Draft - Do Not Cite or Quote

-------
Table 3-19 Distribution of annual 3-month site maxima Pb-TSP concentrations
(jjg/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

106
0.3605
0.003
0.005
0.016
0.023
0.047
0.109
0.378
1.204
1.937
2.489
2.889
2007

47
0.5831
0.009
0.009
0.038
0.043
0.085
0.242
0.815
2.017
2.456
2.889
2.889
2008

54
0.3611
0.012
0.012
0.017
0.035
0.060
0.121
0.467
1.079
1.258
2.070
2.070
2009

96
0.2112
0.003
0.003
0.011
0.021
0.046
0.091
0.262
0.630
0.865
1.375
1.375
Annual site max 3-month means >= national 90th percentile (2008-2010)

011090003-2008

1.2100












060371405-2008

2.4890












120571066-2008

1.7700












180350009-2008

2.1630











290930016b-2008	2.4560

290930016b-2009

2.0700












290930021 b-2009

1.9370












290990004b-2008

2.0170












290990015b-2008

2.8890












290999001 b-2009

1.2040












290999005b-2009

1.2580












290999005b-2010

1.3750












480850009b-2008

1.2620











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 the table fall in the upper 90th percentile of the data pooled by site.
Table 3-20 Distribution of annual 3-month site maxima Pb-TSP concentrations
(jjg/m3) nationwide, non-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

86
0.0198
0.000
0.000
0.002
0.004
0.007
0.015
0.028
0.044
0.051
0.073
0.073
2008

59
0.0176
0.002
0.002
0.004
0.005
0.007
0.014
0.024
0.039
0.048
0.055
0.055
2009

65
0.0162
0.002
0.002
0.003
0.004
0.006
0.013
0.021
0.038
0.041
0.073
0.073
2010

71
0.0171
0.000
0.000
0.001
0.002
0.006
0.013
0.024
0.037
0.047
0.057
0.057
Annual site max 3-month means >= national 90th percentile (2008-2010)

170310022-2008

0.0480












170310022-2009

0.0470











February 2012
3-152
Draft - Do Not Cite or Quote

-------
170310022-2010
170310026b-2008	0.0460

291870006b-2010

0.0540












291892003b-2008

0.0550












390290019b-2010

0.0570












390290022b-2010

0.0440












420210808b-2008

0.0490












420210808b-2009

0.0730












420450002b-2010

0.0470












421290007b-2008

0.0480











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 NMQS for determination of compliance with the Pb NAAQS.
b Sites listed in the bottom nine rows of the table fall in the upper 90th percentile of the data pooled by site.
February 2012
3-153
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Table 3-21 One-month average Pb-TSP for individual	county	concentrations
nationwide (jjg/m ), source-oriented monitors,	2008-2010
Stcoucode State	County name	N monthly means N sites Mean Min	1 5	10	25	50	75	90	95	99	max
Statistics for individual counties (2008-2010)
01109	AL	Pike	32	1	05252	0 054	0054	0 083	0 164	0 252	0-402	0798	1 053	1 117	1277	1277
06037	CA	LosAngeles	131	4	02380	0 018	0019	0 026	0034	0 047	0085	0 246	0602	0 905	2501	2880
12057	FL	Hillsborough	81	3	01755	0 007	0007	0 017	0020	0 053	0 104	0 187	0530	0 567	1 007	1 007
13015	GA	Bartow	12	1	0 0128	0 007	0007	0 007	0008	0 008	0014	0 016	0017	0 019	0019	0 019
13215	GA	Muscogee	12	1	0 0361	0 004	0004	0 004	0010	0 013	0027	0 043	0058	°-140	°-140	°-140
17031	IL	Cook	11	1	01515	0 028	0028	0 028	0028	0 050	0074	0 196	0-304	0-580	0-580	0-580
17115	il	Macon	12	1	0 0800	0 018	0018	0 018	0025	0 035	0074	0 118	0144	0 168	0 168	0 168
17119	il	Madison	36	1	01367	0 018	0018	0 022	0024	0 037	0068	0 175	0304	0 363	0836	0 836
17143	IL	Peoria	24	2	0 0119	0 010	0 010	0 010	0 010	0 010	0 010	0 012	0 016	0 023	0 0 24	0 024
17195	il	Whiteside	12	1	o.om	o"oio	o"oio	o"oio	0*012	0*012	0*015	0*024	0*036	0*040	0m!	0.040
17201	IL	Winnebago	11	1	0 0339	0 010	0010	0 010	0014	0 020	0024	0 032	0050	0 118	0118	0118
18035	IN	Delaware	59	2	02746	0 034	0034	0 040	0049	0 080	0 128	0 241	0427	1011	4440	4440
18089	IN	Lake	57	3	0 0309	0 004	0004	0 007	0008	0 012	0020	0 035	0052	0 079	0298	0 298
18097	IN	Marion	70	2	0 0195	0 003	0003	0 005	0005	0 008	0012	0 025	0046	0 050	0 125	0 125
18127	IN	Porter	12	1	0 0125	0 004	0004	0 004	0005	0 007	0009	0 021	0024	0 026	0026	0 026
19155	ia	Pottawattamie	12	1	0.1536	0*025	0*025	0*025	0*026	0*063	oTe4	0*257	0*276	0^82	0*282	0.282
20169	KS	Saline	11	1	0-2020	0 043	0043	0 043	0044	0 083	0 133	0-320	0457	0488	0488	0488
21019	KY	Boyd	7	1	0 0042	0 002	0002	0 002	0002	0 004	0004	0 004	0007	0 007	0007	0 007
21151	KY	Madison	12	1	0 0255	0 004	0004	0 004	0008	0 013	0017	0 022	0032	0 121	0121	0121
26067	Ml	Ionia	12	1	01781	0 016	0016	0 016	0023	0 054	0 169	0 279	0361	0414	0414	0414
27003	MN	Anoka	12	1	0 0157	0 003	0003	0 003	0005	0 007	0011	0 021	0022	0 054	0054	0 054
27037	MN	Dakota	36	1	0.1966	0*037	0*037	0*048	0*058	0*084	5T37	0*259	0*424	0*572	0738	0.738
27145	MN	Stearns	12	1	0 0028	0 000	0000	0 000	0000	0 000	0003	0 005	0006	0 008	0008	0 008
29093	MO	Iron	171	7	03388	0 007	0008	0 014	0018	0 033	0093	0 518	0-850	1 110	2557	4 225
29099	MO	Jefferson	453	19	04795	0 011	0015	0 033	0048	0 141	0336	0 659	1118	1451	2220	3 123
29179	MO	Reynolds	48	4	0 0428	0 007	0007	0 008	0011	0 017	0027	0 060	0087	0 099	0268	0 268
31053	NE	Dodge	9	1	0 0515	0 005	0005	0 005	0005	0 021	0031	0 053	0 149	0 149	0 149	0 149
31127	ne	Nemaha	8	1	0.0476	0*008	0*008	0*008	0*008	0*010	0*024	0*049	0*206	0*206	0*206	0.206
36071	NY	Orange	105	3	0 0281	0 001	0001	0 003	0004	0 006	0018	0 044	0063	0 081	°-101	0134
39035	OH	Cuyahoga	72	3	0 0941	0 004	0004	0 007	0008	0 014	0038	0 121	0-210	°-400	0719	0719
39051	OH	Fulton	34	1	01462	0 009	0009	0 009	0026	0 057	0091	0 170	0420	0490	0510	0510
39091	OH	Logan	102	4	0 0480	0 003	0003	0 004	0005	0 020	0042	0 070	0090	°-100	0 120	0 170
39101	OH	Marion	10	1	0 0358	0 025	0025	0 025	0026	0 027	0033	0 041	0054	0 066	0066	0 066
39151	OH	Stark	11	1	0.0175	0*008	0*008	0*008	0*009	0*010	o!oiS	0*024	0*025	0*028	0*028	0.028
39155	OH	Trumbull	8	1	0 0075	0 004	0004	0 004	0004	0 005	0007	0 008	0017	0 017	0017	0 017
40121	OK	Pittsburg	11	1	0 0023	0 002	0002	0 002	0002	0 002	0002	0 003	0003	0 003	0003	0 003
41071	OR	Yamhill	12	1	0 0157	0 006	0006	0 006	0007	0 008	0016	0 020	0025	0 037	0037	0 037
42003	PA	Allegheny	24	2	0 0369	0 006	0006	0 006	0006	0 010	0017	0 040	0 121	0144	0 149	0 149
42007	PA	Beaver	54	3	01130	0 042	0042	0 044	0047	0 068	0096	0 128	0 198	0 272	0286	0 286
42011	PA	Berks	117	6	0 0989	0 034	0035	0 038	0042	0 048	0066	0 119	0-200	0 295	0347	0 348
42045	PA	Delaware	12	1	0 0452	0 043	0043	0 043	0043	0 043	0045	0 047	0048	0 048	0048	0 048
42055	PA	Franklin	11	1	0 0449	0 042	0042	0 042	0043	0 043	0045	0 047	0047	0 047	0047	0 047
42063	PA	Indiana	12	1	0 0454	0 042	0042	0 042	0043	0 043	0044	0 046	0047	0 058	0058	0 058
42073	PA	Lawrence	8	1	0 0438	0 042	0042	0 042	0042	0 043	0044	0 045	0046	0 046	0046	0 046
42079	PA	Luzerne	10	1	0 0953	0 043	0043	0 043	0044	0 045	0071	0 102	0215	0268	0268	0 268
42129	PA	Westmoreland	12	1	0 0439	0 041	0041	0 041	0041	0 043	0044	0 045	0046	0 047	0047	0 047
47093	TN	Knox	48	2	0 0165	0 002	0002	0 005	0006	0 008	0012	0 019	0032	0 038	0063	0 063
February 2012
3-154
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Stcoucode State County name N monthly means N sites Mean Min 1 5 10 25 50 75 90 95 99 max
47163 TN Sullivan 120	4 0 0534 0 021 0023 0 030 0032 0 037 0045 0 059 0083 0 124 0 145 0 156
48085
TX
Collin
108
3
0.3062
0.007
0.028
0.040
0.052
0.104
0.189
0.438
0.717
0.904
1.178
1.564
48375
TX
Potter
6
1
0.0044
0.004
0.004
0.004
0.004
0.004
0.004
0.005
0.006
0.006
0.006
0.006
51770
VA
Roanoke City
12
1
0.0412
0.005
0.005
0.005
0.008
0.010
0.015
0.035
0.054
0.272
0.272
0.272
55117
Wl
Sheboygan
12
1
0.0802
0.001
0.001
0.001
0.003
0.007
0.054
0.136
0.182
0.279
0.279
0.279
72013
PR
Arecibo
12
1
0.1774
0.038
0.038
0.038
0.064
0.102
0.178
0.264
0.290
0.310
0.310
0.310
Table 3-22 One-month average Pb-TSP for individual county concentrations
nationwide (jjg/m3), non-source-oriented monitors, 2008-2010
Stcou
code
State
County name
N
monthly
means
N
sites
Mean
Min
1
5
10
25
50
75
90
95
99
max
Statistics for individual counties (2008-2010)
04013
AZ
Maricopa
33

0.0218
0.009
0.009
0.010
0.011
0.013
0.019
0.029
0.036
0.041
0.041
0.041
06025
CA
Imperial
117
5
0.0107
0.000
0.000
0.000
0.002
0.006
0.010
0.015
0.020
0.024
0.032
0.038
06037
CA
Los Angeles
6

0.0218
0.009
0.009
0.009
0.009
0.014
0.021
0.028
0.038
0.038
0.038
0.038
06065
CA
Riverside
33

0.0162
0.004
0.004
0.006
0.009
0.011
0.015
0.019
0.025
0.032
0.035
0.035
06071
CA
San Bernardino
224
8
0.0098
0.000
0.000
0.000
0.002
0.006
0.010
0.012
0.017
0.020
0.038
0.044
08005
CO
Arapahoe
72
2
0.0077
0.000
0.000
0.003
0.004
0.006
0.008
0.010
0.010
0.012
0.014
0.014
08031
CO
Denver
71
2
0.0091
0.001
0.001
0.003
0.004
0.007
0.010
0.012
0.014
0.014
0.022
0.022
13089
GA
DeKalb
9

0.0120
0.004
0.004
0.004
0.004
0.007
0.012
0.016
0.018
0.018
0.018
0.018
17031
IL
Cook
12

0.0056
0.003
0.003
0.003
0.004
0.005
0.005
0.006
0.008
0.008
0.008
0.008
17117
IL
Macoupin
10

0.0033
0.002
0.002
0.002
0.002
0.003
0.003
0.004
0.005
0.006
0.006
0.006
17119
IL
Madison
288
8
0.0195
0.010
0.010
0.010
0.010
0.012
0.016
0.025
0.034
0.040
0.060
0.070
17143
IL
Peoria
24

0.0101
0.010
0.010
0.010
0.010
0.010
0.010
0.010
0.010
0.010
0.012
0.012
17163
IL
Saint Clair
36

0.0188
0.010
0.010
0.010
0.010
0.012
0.016
0.020
0.032
0.053
0.066
0.066
18089
IN
Lake
36

0.0105
0.010
0.010
0.010
0.010
0.010
0.010
0.010
0.013
0.013
0.014
0.014
18097
IN
Marion
36

0.0206
0.010
0.010
0.010
0.012
0.014
0.018
0.026
0.032
0.038
0.054
0.054
18163
IN
Vanderburgh
36

0.0150
0.005
0.005
0.005
0.005
0.008
0.014
0.019
0.030
0.033
0.049
0.049
25025
MA
Suffolk
35

0.0058
0.002
0.002
0.002
0.003
0.004
0.005
0.008
0.010
0.012
0.013
0.013
26081
Ml
Kent
33
2
0.0045
0.001
0.001
0.001
0.002
0.003
0.004
0.005
0.006
0.010
0.010
0.010
26163
Ml
Wayne
31
2
0.0087
0.004
0.004
0.004
0.005
0.007
0.008
0.010
0.013
0.016
0.020
0.020
27017
MN
Carlton
12

0.0053
0.003
0.003
0.003
0.003
0.005
0.005
0.006
0.008
0.008
0.008
0.008
27037
MN
Dakota
36
2
0.0112
0.003
0.003
0.003
0.004
0.005
0.009
0.015
0.021
0.023
0.032
0.032
27053
MN
Hennepin
12

0.0000
0.000
0.000
0.000
0.000
0.000
0.000
0.000
0.000
0.000
0.000
0.000
27075
MN
Lake
118
5
0.0035
0.000
0.000
0.000
0.000
0.000
0.002
0.005
0.008
0.010
0.017
0.036
27123
MN
Ramsey
126
4
0.0032
0.000
0.000
0.000
0.000
0.000
0.002
0.005
0.006
0.008
0.010
0.044
27137
MN
Saint Louis
10

0.0000
0.000
0.000
0.000
0.000
0.000
0.000
0.000
0.000
0.000
0.000
0.000
27163
MN
Washington
71
3
0.0062
0.000
0.000
0.000
0.000
0.002
0.004
0.008
0.013
0.020
0.028
0.028
29097
MO
Jasper
72
2
0.0015
0.000
0.000
0.000
0.000
0.000
0.000
0.002
0.004
0.006
0.010
0.010
29187
MO
Saint Francois
72
3
0.0016
0.000
0.000
0.000
0.000
0.000
0.000
0.003
0.004
0.005
0.006
0.006
29189
MO
Saint Louis
12

0.0125
0.007
0.007
0.007
0.007
0.009
0.012
0.017
0.018
0.019
0.019
0.019
36047
NY
Kings
24
2
0.0327
0.008
0.008
0.009
0.009
0.018
0.032
0.039
0.054
0.080
0.089
0.089
39017
OH
Butler
33

0.0230
0.005
0.005
0.005
0.005
0.006
0.008
0.050
0.050
0.050
0.066
0.066
February 2012	3-155	Draft - Do Not Cite or Quote

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Stcou
code
State
County name
N
monthly
means
N
sites
Mean
Min
1
5
10
25
50
75
90
95
99
max
39029
OH
Columbiana
24
1
0.0131
0.010
0.010
0.010
0.010
0.011
0.012
0.014
0.018
0.020
0.020
0.020
39035
OH
Cuyahoga
34
1
0.0055
0.002
0.002
0.003
0.004
0.004
0.005
0.007
0.008
0.009
0.009
0.009
39049
OH
Franklin
107
3
0.0155
0.004
0.004
0.006
0.007
0.008
0.011
0.018
0.027
0.034
0.065
0.136
39143
OH
Sandusky
107
3
0.0143
0.004
0.004
0.006
0.007
0.009
0.012
0.017
0.024
0.030
0.041
0.041
39167
OH
Washington
36
1
0.0092
0.004
0.004
0.005
0.005
0.007
0.009
0.011
0.013
0.014
0.016
0.016
40115
OK
Ottawa
12
1
0.0048
0.003
0.003
0.003
0.003
0.004
0.005
0.006
0.006
0.007
0.007
0.007
42003
PA
Allegheny
54
2
0.0048
0.002
0.002
0.002
0.003
0.003
0.005
0.006
0.007
0.008
0.010
0.010
42021
PA
Cambria
16
2
0.0124
0.003
0.003
0.003
0.005
0.006
0.013
0.017
0.021
0.025
0.025
0.025
42045
PA
Delaware
36
1
0.0105
0.000
0.000
0.000
0.000
0.004
0.009
0.015
0.019
0.024
0.053
0.053
42101
PA
Philadelphia
23
1
0.0463
0.040
0.040
0.040
0.040
0.040
0.040
0.044
0.054
0.058
0.128
0.128
42129
PA
Westmoreland
20
1
0.0432
0.040
0.040
0.040
0.040
0.040
0.043
0.046
0.047
0.048
0.048
0.048
48061
TX
Cameron
24
1
0.0210
0.011
0.011
0.011
0.012
0.014
0.020
0.027
0.033
0.033
0.039
0.039
48141
TX
El Paso
24
1
0.0419
0.037
0.037
0.040
0.040
0.040
0.040
0.042
0.050
0.050
0.053
0.053
48201
TX
Harris
35
1
0.0041
0.002
0.002
0.003
0.003
0.003
0.004
0.005
0.006
0.007
0.009
0.009
48479
TX
Webb
68
3
0.0206
0.014
0.014
0.014
0.014
0.015
0.017
0.019
0.029
0.056
0.087
0.087
49035
UT
Salt Lake
32
1
0.0053
0.003
0.003
0.003
0.004
0.004
0.005
0.006
0.007
0.008
0.010
0.010
51087
VA
Henrico
29
1
0.0134
0.004
0.004
0.005
0.006
0.008
0.011
0.018
0.026
0.028
0.035
0.035
Table 3-23 Three-month moving average Pb-TSP for individual county
concentrations (jjg/m3) nationwide, source-oriented monitors, 2008-
2010
Stcou
code
State
County name
N
monthly
means
N
sites
Mean
Min
1
5
10
25
50
75
90
95
99
max
Statistics for individual counties (2008-2010)a
01109
AL
Pike
25

0.5771
0.223
0.223
0.247
0.256
0.302
0.574
0.719
1.088
1.178
1.210
1.210
06037
CA
Los Angeles
131
4
0.2521
0.023
0.023
0.036
0.041
0.055
0.078
0.237
0.543
0.832
2.452
2.489
12057
FL
Hillsborough
79
3
0.1940
0.011
0.011
0.015
0.037
0.063
0.110
0.249
0.423
0.582
1.770
1.770
13015
GA
Bartow
11

0.0125
0.009
0.009
0.009
0.009
0.011
0.013
0.014
0.015
0.016
0.016
0.016
13215
GA
Muscogee
12

0.0367
0.014
0.014
0.014
0.020
0.022
0.031
0.052
0.066
0.070
0.070
0.070
17031
IL
Cook
9

0.1364
0.068
0.068
0.068
0.068
0.109
0.135
0.150
0.241
0.241
0.241
0.241
17115
IL
Macon
10

0.0806
0.048
0.048
0.048
0.052
0.067
0.080
0.088
0.117
0.123
0.123
0.123
17119
IL
Madison
36

0.1346
0.027
0.027
0.035
0.036
0.063
0.113
0.207
0.283
0.341
0.416
0.416
17143
IL
Peoria
20
2
0.0121
0.010
0.010
0.010
0.010
0.011
0.012
0.014
0.015
0.016
0.016
0.016
17195
IL
Whiteside
10

0.0191
0.012
0.012
0.012
0.014
0.016
0.019
0.022
0.025
0.025
0.025
0.025
17201
IL
Winnebago
9

0.0356
0.019
0.019
0.019
0.019
0.021
0.027
0.057
0.063
0.063
0.063
0.063
18035
IN
Delaware
57
2
0.2866
0.053
0.053
0.059
0.073
0.090
0.159
0.246
0.495
1.867
2.163
2.163
18089
IN
Lake
46
2
0.0305
0.007
0.007
0.011
0.012
0.016
0.027
0.036
0.040
0.057
0.129
0.129
18097
IN
Marion
66
2
0.0198
0.005
0.005
0.006
0.007
0.011
0.014
0.025
0.036
0.043
0.079
0.079
18127
IN
Porter
10

0.0131
0.007
0.007
0.007
0.007
0.007
0.013
0.017
0.020
0.022
0.022
0.022
19155
IA
Pottawattamie
12

0.1581
0.034
0.034
0.034
0.067
0.113
0.153
0.220
0.246
0.263
0.263
0.263
20169
KS
Saline
9

0.2286
0.096
0.096
0.096
0.096
0.107
0.231
0.324
0.421
0.421
0.421
0.421
21151
KY
Madison
10

0.0212
0.013
0.013
0.013
0.014
0.015
0.017
0.024
0.037
0.049
0.049
0.049
February 2012
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Stcou
code
State
County name
N
monthly
means
N
sites
Mean
Min
1
5
10
25
50
75
90
95
99
max
26067
Ml
Ionia
10

0.1980
0.106
0.106
0.106
0.110
0.128
0.212
0.259
0.273
0.284
0.284
0.284
27003
MN
Anoka
10

0.0161
0.006
0.006
0.006
0.008
0.010
0.013
0.022
0.029
0.031
0.031
0.031
27037
MN
Dakota
36

0.2026
0.068
0.068
0.072
0.088
0.104
0.216
0.248
0.357
0.415
0.429
0.429
27145
MN
Stearns
10

0.0032
0.000
0.000
0.000
0.001
0.002
0.004
0.004
0.005
0.005
0.005
0.005
29093
MO
Iron
158
6
0.3465
0.010
0.011
0.019
0.022
0.033
0.142
0.549
0.901
1.167
2.076
2.456
29099
MO
Jefferson
423
19
0.4925
0.023
0.033
0.050
0.071
0.187
0.385
0.723
0.989
1.186
2.017
2.889
29179
MO
Reynolds
40
4
0.0397
0.012
0.012
0.014
0.015
0.017
0.031
0.057
0.087
0.089
0.100
0.100
31053
NE
Dodge
7

0.0474
0.019
0.019
0.019
0.019
0.020
0.060
0.067
0.072
0.072
0.072
0.072
31127
NE
Nemaha
6

0.0447
0.019
0.019
0.019
0.019
0.024
0.032
0.075
0.087
0.087
0.087
0.087
36071
NY
Orange
99
3
0.0271
0.003
0.003
0.004
0.005
0.007
0.027
0.037
0.068
0.075
0.086
0.086
39035
OH
Cuyahoga
70
3
0.0905
0.006
0.006
0.010
0.011
0.021
0.050
0.122
0.221
0.287
0.531
0.531
39051
OH
Fulton
30

0.1609
0.025
0.025
0.027
0.046
0.054
0.092
0.254
0.354
0.453
0.567
0.567
39091
OH
Logan
100
4
0.0499
0.004
0.004
0.004
0.006
0.033
0.047
0.072
0.090
0.095
0.100
0.100
39101
OH
Marion
8

0.0379
0.032
0.032
0.032
0.032
0.034
0.037
0.042
0.047
0.047
0.047
0.047
39151
OH
Stark
9

0.0180
0.015
0.015
0.015
0.015
0.016
0.018
0.019
0.023
0.023
0.023
0.023
39155
OH
Trumbull
6

0.0080
0.005
0.005
0.005
0.005
0.006
0.008
0.010
0.011
0.011
0.011
0.011
40121
OK
Pittsburg
9

0.0021
0.002
0.002
0.002
0.002
0.002
0.002
0.002
0.003
0.003
0.003
0.003
41071
OR
Yamhill
10

0.0166
0.009
0.009
0.009
0.011
0.013
0.016
0.019
0.026
0.027
0.027
0.027
42003
PA
Allegheny
20
2
0.0414
0.009
0.009
0.011
0.012
0.017
0.030
0.054
0.099
0.120
0.138
0.138
42007
PA
Beaver
41
3
0.1160
0.043
0.043
0.052
0.056
0.083
0.114
0.159
0.170
0.187
0.206
0.206
42011
PA
Berks
105
6
0.0995
0.038
0.039
0.041
0.045
0.051
0.078
0.145
0.183
0.197
0.242
0.251
42045
PA
Delaware
10

0.0447
0.043
0.043
0.043
0.043
0.043
0.045
0.046
0.047
0.047
0.047
0.047
42055
PA
Franklin
7

0.0447
0.043
0.043
0.043
0.043
0.043
0.045
0.046
0.046
0.046
0.046
0.046
42063
PA
Indiana
10

0.0447
0.043
0.043
0.043
0.043
0.043
0.044
0.046
0.049
0.049
0.049
0.049
42079
PA
Luzerne
6

0.1078
0.084
0.084
0.084
0.084
0.085
0.103
0.135
0.137
0.137
0.137
0.137
42129
PA
Westmoreland
10

0.0434
0.041
0.041
0.041
0.042
0.042
0.044
0.044
0.046
0.046
0.046
0.046
47093
TN
Knox
44
2
0.0165
0.007
0.007
0.009
0.009
0.012
0.016
0.020
0.023
0.027
0.035
0.035
47163
TN
Sullivan
118
4
0.0554
0.030
0.030
0.033
0.035
0.039
0.045
0.060
0.100
0.125
0.134
0.168
48085
TX
Collin
108
3
0.3101
0.048
0.051
0.070
0.085
0.120
0.217
0.469
0.682
0.753
1.189
1.262
51770
VA
Roanoke City
10

0.0466
0.013
0.013
0.013
0.016
0.019
0.026
0.097
0.108
0.109
0.109
0.109
55117
Wl
Sheboygan
10

0.0897
0.012
0.012
0.012
0.034
0.058
0.076
0.126
0.164
0.170
0.170
0.170
72013
PR
Arecibo
10

0.1725
0.059
0.059
0.059
0.068
0.129
0.194
0.213
0.241
0.245
0.245
0.245
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.
February 2012
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Table 3-24 Three-month moving average Pb-TSP for individual county
concentrations (jjg/m3) nationwide, non-source-oriented monitors,
2008-2010
Stcou
code
State
County name
N
monthly
means
N
sites
Mean
Min
1
5
10
25
50
75
90
95
99
max
Statistics for individual counties (2008-2010)a
06025
CA
Imperial
31

0.0165
0.007
0.007
0.008
0.011
0.013
0.017
0.021
0.023
0.023
0.023
0.023
06037
CA
Los Angeles
218
8
0.0100
0.000
0.000
0.002
0.004
0.006
0.009
0.013
0.016
0.020
0.028
0.035
06065
CA
Riverside
72
2
0.0078
0.002
0.002
0.004
0.005
0.007
0.008
0.010
0.011
0.011
0.011
0.011
06071
CA
San Bernardino
69
2
0.0091
0.003
0.003
0.005
0.006
0.007
0.009
0.011
0.013
0.014
0.017
0.017
08005
CO
Arapahoe
7

0.0126
0.011
0.011
0.011
0.011
0.011
0.013
0.014
0.014
0.014
0.014
0.014
08031
CO
Denver
10

0.0054
0.004
0.004
0.004
0.004
0.005
0.006
0.006
0.006
0.006
0.006
0.006
13089
GA
DeKalb
8

0.0035
0.003
0.003
0.003
0.003
0.003
0.004
0.004
0.004
0.004
0.004
0.004
17031
IL
Cook
287
8
0.0196
0.010
0.010
0.010
0.010
0.012
0.017
0.025
0.033
0.038
0.047
0.051
17117
IL
Macoupin
24

0.0101
0.010
0.010
0.010
0.010
0.010
0.010
0.010
0.011
0.011
0.011
0.011
17119
IL
Madison
36

0.0188
0.010
0.010
0.010
0.011
0.014
0.016
0.022
0.036
0.036
0.039
0.039
17143
IL
Peoria
36

0.0105
0.010
0.010
0.010
0.010
0.010
0.010
0.011
0.012
0.012
0.013
0.013
17163
IL
Saint Clair
36

0.0204
0.012
0.012
0.012
0.014
0.016
0.020
0.024
0.029
0.033
0.036
0.036
18089
IN
Lake
36

0.0149
0.007
0.007
0.007
0.007
0.010
0.014
0.018
0.024
0.032
0.037
0.037
18097
IN
Marion
33

0.0056
0.003
0.003
0.003
0.003
0.004
0.005
0.007
0.009
0.010
0.011
0.011
18163
IN
Vanderburgh
31
2
0.0047
0.002
0.002
0.003
0.003
0.004
0.005
0.005
0.006
0.007
0.007
0.007
25025
MA
Suffolk
24
2
0.0093
0.005
0.005
0.006
0.006
0.008
0.009
0.011
0.013
0.015
0.016
0.016
26081
Ml
Kent
10

0.0055
0.004
0.004
0.004
0.005
0.005
0.006
0.006
0.006
0.006
0.006
0.006
26163
Ml
Wayne
32
2
0.0119
0.004
0.004
0.004
0.005
0.005
0.012
0.017
0.021
0.023
0.024
0.024
27017
MN
Carlton
10

0.0000
0.000
0.000
0.000
0.000
0.000
0.000
0.000
0.000
0.000
0.000
0.000
27037
MN
Dakota
112
5
0.0036
0.000
0.000
0.001
0.001
0.001
0.003
0.005
0.007
0.012
0.013
0.015
27053
MN
Hennepin
124
4
0.0033
0.000
0.001
0.001
0.001
0.002
0.003
0.004
0.006
0.006
0.015
0.016
27075
MN
Lake
8

0.0000
0.000
0.000
0.000
0.000
0.000
0.000
0.000
0.000
0.000
0.000
0.000
27123
MN
Ramsey
65
3
0.0061
0.001
0.001
0.001
0.001
0.002
0.005
0.008
0.014
0.016
0.017
0.017
27137
MN
Saint Louis
70
2
0.0016
0.000
0.000
0.000
0.000
0.001
0.001
0.002
0.004
0.004
0.005
0.005
27163
MN
Washington
70
3
0.0017
0.000
0.000
0.000
0.000
0.001
0.001
0.003
0.004
0.004
0.005
0.005
29097
MO
Jasper
10

0.0135
0.009
0.009
0.009
0.011
0.012
0.014
0.015
0.016
0.017
0.017
0.017
29187
MO
Saint Francois
21
2
0.0337
0.011
0.011
0.012
0.012
0.027
0.035
0.042
0.048
0.053
0.054
0.054
29189
MO
Saint Louis
33

0.0243
0.005
0.005
0.005
0.006
0.007
0.008
0.050
0.050
0.050
0.055
0.055
36047
NY
Kings
24

0.0131
0.011
0.011
0.011
0.011
0.012
0.013
0.014
0.016
0.018
0.019
0.019
39017
OH
Butler
30

0.0055
0.003
0.003
0.004
0.004
0.005
0.006
0.006
0.007
0.007
0.008
0.008
39029
OH
Columbiana
105
3
0.0148
0.005
0.005
0.007
0.008
0.010
0.013
0.017
0.021
0.028
0.054
0.057
39035
OH
Cuyahoga
105
3
0.0144
0.005
0.006
0.006
0.008
0.010
0.013
0.018
0.023
0.027
0.033
0.035
39049
OH
Franklin
36

0.0092
0.005
0.005
0.005
0.005
0.008
0.010
0.011
0.011
0.012
0.012
0.012
39143
OH
Sandusky
10

0.0052
0.004
0.004
0.004
0.004
0.005
0.005
0.006
0.006
0.006
0.006
0.006
39167
OH
Washington
48
2
0.0047
0.002
0.002
0.002
0.003
0.004
0.004
0.006
0.007
0.007
0.008
0.008
40115
OK
Ottawa
12
2
0.0128
0.005
0.005
0.005
0.006
0.010
0.014
0.016
0.018
0.019
0.019
0.019
42003
PA
Allegheny
36

0.0101
0.000
0.000
0.000
0.000
0.007
0.012
0.014
0.016
0.018
0.025
0.025
42021
PA
Cambria
23

0.0459
0.040
0.040
0.040
0.040
0.040
0.041
0.046
0.069
0.070
0.073
0.073
42045
PA
Delaware
14

0.0427
0.040
0.040
0.040
0.040
0.040
0.042
0.045
0.046
0.047
0.047
0.047
42101
PA
Philadelphia
22

0.0214
0.013
0.013
0.014
0.014
0.018
0.022
0.025
0.029
0.029
0.030
0.030
42129
PA
Westmoreland
24

0.0417
0.037
0.037
0.040
0.040
0.040
0.041
0.043
0.046
0.047
0.048
0.048
February 2012
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1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
Stcou
code
State
County name
N
monthly
means
N
sites
Mean
Min
1
5
10
25
50
75
90
95
99
max
48061
TX
Cameron
33
1
0.0042
0.002
0.002
0.003
0.003
0.004
0.004
0.005
0.005
0.006
0.006
0.006
48141
TX
El Paso
56
3
0.0212
0.014
0.014
0.014
0.015
0.016
0.018
0.023
0.038
0.040
0.040
0.040
48201
TX
Harris
30
1
0.0051
0.004
0.004
0.004
0.004
0.005
0.005
0.006
0.006
0.007
0.007
0.007
48479
TX
Webb
23
1
0.0121
0.006
0.006
0.007
0.007
0.008
0.010
0.016
0.021
0.022
0.026
0.026
49035
UT
Salt Lake
10
1
0.0145
0.007
0.007
0.007
0.007
0.008
0.011
0.016
0.032
0.036
0.036
0.036
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.
3.8.2 Intra-urban Variability
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 3A-1
where X,, and represent the observed hourly concentrations for time period i at sites j
and k, and p is the number of paired hourly observations. A COD of 0 indicates there are
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 for the years 2008-2010 are near
February 2012
3-159
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1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
32
33
34
35
36
37
38
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
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 3-42 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/yr ("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/yr (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/yr (U.S. EPA. 2008c). Location of the non-source-oriented monitors
varied. Monitor F was positioned on a roof top 60 meters away from a 4-lane arterial road
and 100 m from of a railroad. Monitor G was located on a rooftop approximately 20 m
from an 8-lane arterial road, and monitor H was positioned at the curbside of a four-lane
road roughly 650 m north of that road's junction with 1-405. Monitor I was sited in a
parking lot roughly 80 m from a four-lane road, and monitor J was located approximately
130 m south of a 4-lane highway. Figure 3-43 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 3-42 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 3-44. The average annual concentration at monitor A
was 0.074 |ig/m3. The 95th percentile exceeded the level of the NAAQS in the spring
(0.16 |ig/m3) and summer (0.18 (ig/m3). Monitor C reported the highest concentrations in
Los Angeles County, with a year-round mean of 0.68 |ig/nr\ 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 |ig/nr\ and the 75th percentile of year-round data
exceeded the level of the NAAQS; in spring, the 70th percentile exceeded 0.15 |ig/nr\
February 2012
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1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
Monitor B reported the lowest values among the source-oriented monitors with an
average annual concentration of 0.013 |ig/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 |ig/m3. and the 95th percentile
of concentration was 0.17 |ig/nr\
The non-source-oriented monitors located at sites F-J all recorded low concentrations,
with average values ranging from 0.004 to 0.018 |ig/m3 (Figure 3-44). The highest
average year-round concentrations were recorded at site F. The 95th percentiles at these
sites ranged from 0.01 to 0.04 |ig/nr\ 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 3-25), 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.
February 2012
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Legend
Q TSP Source Monitors
•	TSP Non-source Monitors
•	City-based Population Center
•	County-based Population Center
Interstates
	Major Highways
Bodies o( Water
Urban Areas
Los Angetes County, CA
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 3-42 Pb TSP monitor and source locations within Los Angeles County,
CA (06-037), 2007-2009.
February 2012
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¦Sbrten ¦ L<» i" :n i- -" |'L NI'l.
CM

-" it.i
Source: NRCS (2011V
Note: Clockwise from top left: January, April, July, and October. Note that the wind percentages vary from month to month.
Figure 3-43 Wind roses for Los Angeles County, CA, from meteorological data
at the Los Angeles International Airport, 1961-1990.
February 2012
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Site
A
B
C
D
E
F
G
H
I
J
SITE ID
06-037-
1404
06-037-
1602
06-037-
1405
06-037-
1406
06-037-
1403
06-037-
1103
06-037-
1301
06-037-
4002
06-037-
4004
06-037-
5005
MEAN
0.074
0.013
0.68
0.12
0.068
0.018
0.015
0.0083
0.0087
0.0040
SD
0.040
0.017
1.0
0.092
0.052
0.011
0.012
0.0068
0.0069
0.0064
OBS
66
112
617
242
128
121
108
120
117
109
% BELOW
MDL
0
75
0
0
0
45
65
85
85
95
Source
orientation
Source
Source
Source
Source
Source
Non-
source
Non-source
Non-
source
Non-source
Non-
source
o
u
M
2.9 -
2.8	-
2.7	-
2.e
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 -
o.e
0.5 -
0.4 -
0.3 -
0.2 -
0.1
0.0
A

B
D
it
it
i+iii
H
Y 1 2 3 4 Y1 2 3 4 Y1 2 34 Y 1 2 3 4 Y1 2 3 4 Y1 2 34 Y 1 2 3 4 Y1 2 3 4 Y1 2 34 Y1234
season
Figure 3-44 Box plots of annual and seasonal Pb TSP concentrations (jjg/m )
from source-oriented and non-source-oriented monitors within
Los Angeles County, CA (06-037), 2007-2009.
February 2012
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Table 3-25 Comparisons between Pb TSP concentrations from source-oriented
and non-source-oriented monitors within Los Angeles County, CA
(06-037), 2007-2009



A
B
c
D
E
F
G
H
I
J



Source
Source
Source
Source
Source
Non-
Source
Non-
Source
Non-
Source
Non-
Source
Non-
Source
A
Source
R
1.00
-0.04
0.14
0.10
0.17
0.03
0.00
-0.08
-0.07
-0.27


P
1.00
0.16
0.10
0.08
0.27
-0.15
0.00
0.14
-0.02
-0.09


P90
0.00
0.08
0.49
0.10
0.10
0.08
0.06
0.08
0.08
0.08


COD
0.00
0.63
0.64
0.31
0.34
0.57
0.57
0.79
0.77
0.85
B
Source
R

1.00
0.06
0.17
-0.06
0.33
0.29
0.40
0.22
0.20


P

1.00
0.05
0.05
0.07
0.74
0.12
0.28
0.11
0.10


P90

0.00
3.59
0.25
0.10
0.02
0.02
0.01
0.02
0.02


COD

0.00
0.96
0.84
0.71
0.46
0.48
0.61
0.60
0.81
C
Source
R


1.00
0.59
0.08
0.12
0.24
0.28
0.18
0.08


P


1.00
0.57
0.03
-0.08
0.26
0.28
0.20
0.13


P90


0.00
1.76
2.14
3.59
4.22
3.59
3.59
3.92


COD


0.00
0.68
0.77
0.95
0.96
0.98
0.98
0.99
D
Source
R



1.00
0.18
0.33
0.09
0.32
0.20
0.03


P



1.00
0.12
0.17
0.11
0.24
0.21
0.07


P90



0.00
0.17
0.24
0.25
0.25
0.25
0.25


COD



0.00
0.42
0.78
0.80
0.89
0.89
0.95
E
Source
R




1.00
0.05
0.07
0.00
0.09
-0.07


P




1.00
0.13
0.06
0.24
0.07
0.18


P90




0.00
0.10
0.10
0.11
0.11
0.11


COD




0.00
0.61
0.64
0.78
0.79
0.90
F
Non-Source
R





1.00
0.10
0.43
0.34
0.21


P





1.00
0.02
0.19
0.09
0.09


P90





0.00
0.02
0.02
0.02
0.02


COD





0.00
0.39
0.61
0.58
0.82
G
Non-Source
R






1.00
0.71
0.55
0.54


P






1.00
0.65
0.39
0.38


P90






0.00
0.01
0.02
0.02


COD






0.00
0.54
0.61
0.85
H
Non-Source
R







1.00
0.60
0.51


P







1.00
0.51
0.40


P90







0.00
0.01
0.01


COD







0.00
0.55
0.77
1
Non-Source
R








1.00
0.29


P








1.00
0.37


P90








0.00
0.01


COD








0.00
0.78
J
Non-Source
R









1.00


P









1.00


P90









0.00


COD









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).
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1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
Figure 3-45 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
m 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 m southwest
of the same facility. Monitor C was located next to a two-lane road in Pinellas Park, FL.
Figure 3-46 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 3-47. The average annual concentration at monitor A was 0.15 |ig/m3.
and the 95th percentile was 0.70 |ig/nr\ 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 |ig/m3 and a 95th
percentile of 1.9 |ig/nr\ 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 (ig/m3. This is likely related to
its location next to a quiet road in a small city.
Intersampler correlations, shown in Table 3-26, 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.
February 2012
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plampa-r'Sl. Motersburg
Legend
o rsPSout« Monaw»
•	TSP t4sri-«>jrc« Mentors
4 City-fcased Population Ccnler
•	Cflw*H»K4 PopoWwo Cwl«<
	lnM»il*cs
	J*ajo< Wghvrtys
¦i Boo-m qI WM"
UttMfiAMM
PrnellM ar
-------
Source: NRCS (2011V
Note: Clockwise from top left: January, April, July, and October. Note that wind percentages vary from month to month.
Figure 3-46 Wind roses for Hillsborough/Pinellas Counties, FL, obtained from
meteorological data at Tampa International Airport, 1961-1990.
February 2012
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Site
A
B
C
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
OBS
154
155
58
% BELOW MDL
20
5
95
Source orientation
Source
Source
Non-source
M
c
O
c

-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
Table 3-26 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
B
c



Source
Source
Non-source
A
Source
R
1.00
-0.09



P
1.00
-0.08



P90
0.00
1.20
0.50


COD
0.00
0.71
1.00
B
Source
R

1.00



P

1.00



P90

0.00
2.20


COD

0.00
1.00
C
Non-source
R


1.00


P


1.00


P90


0.00


COD


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 3-48 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 I-90/I-94. Monitor B was located on the northern roadside
of 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 m west of
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 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 m north of 1-90 and 300 m northwest of the
junction between 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 school in a
residential neighborhood on the south side of Chicago, roughly 100 m south of a rail line
and over 300 m west of the closest arterial road. Although not designated as a source
February 2012
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1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
32
33
34
35
36
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. 1-94 curves around this site and was approximately 700 m from the monitor
to the east and around to the north. Figure 3-49 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 3-50 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 |ig/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 |ig/m3. with
a median of 0.02 (ig/m3 and a maximum concentration of 0.31 (ig/m3. In winter, the 95th
percentile of data was 0.14 |ig/nr\ 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 3.2.2.5. 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 |ig/m3 and 0.024 |ig/m3. respectively. The median
concentrations for monitors A and B were 0.02 |ig/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 m from the closest interstate and 570 m from the closest runway at O'Hare
International Airport. The average concentration at this site was 0.012 |ig/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 |ig/nr\ It is possible that the difference between Pb
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
February 2012
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1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
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 3-27). 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
uncorrected 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. 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.
February 2012
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i Chicago
Legend
• City-based Population Centei
9 County-based Population Centei
	Major Highways
Cook County. IL
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 3-48 Pb TSP Monitor locations within Cook County, IL (17-031),
2007-2009.
Febraaiy 2012
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Source: NRCS (20111
Note: Clockwise from the top left: January, April, July, and October. Note that the wind percentages vary from month to month.
Figure 3-49 Wind roses for Cook County, IL, obtained from meteorological
data at O'Hare International Airport, 1961-1990.
February 2012
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Site
A
B
C
D
E
F
G
H
SITE ID
17-031-0026
17-031-6003
17-031-0022
17-031-3103
17-031-0001
17-031-0052
17-031-3301
17-031-4201
MEAN
0.030
0.024
0.031
0.012
0.013
0.017
0.017
0.011
SD
0.020
0.013
0.036
0.0062
0.0078
0.0098
0.0097
0.0031
OBS
179
175
177
168
177
175
171
168
% BELOW
MDL
15
25
25
85
75
55
50
95
Source
orientation
Source
Source
Source
Source
Non-source
Non-
source
Non-
source
Non-source
M
=L
c
O
c
01
U
c
o
(J
0.14
0.13
0.12
0.11
0.10
0.09
0.08
0.07
0.06
0.05
0.04
0.03
0.02
0.01
0.00
B
D
H

—i—i—i—i——i—i—i—i—i—
Y 1 2 3 4 Y 1 2 3 4 Y1234 Y1234 Y1234 Y1234 Y1234 Y1234
season
Figure 3-50 Box plots of annual and seasonal Pb TSP concentrations (jjg/m3)
from source-oriented and non-source-oriented monitors within
Cook County, IL (17-031), 2007-2009.
February 2012
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Table 3-27 Correlations between Pb TSP concentrations from source-oriented
and non-source-oriented monitors within Cook County, IL (17-031),
2007-2009



A
B
c
D
E
F
G
H



Source
Source
Source
Source
Non-
Source
Non-
Source
Non-
Source
Non-
Source
A
Source
R
1.00
0.32
0.00
0.05
0.17
0.39
0.34
0.06


P
1.00
0.26
-0.01
0.08
0.06
0.32
0.18
0.06


P90
0.00
0.03
0.06
0.04
0.04
0.03
0.03
0.04


COD
0.00
0.29
0.38
0.43
0.41
0.36
0.36
0.45
B
Source
R

1.00
0.14
0.07
0.54
0.52
0.60
0.06


P

1.00
0.05
0.10
0.32
0.46
0.35
-0.01


P90

0.00
0.04
0.03
0.03
0.02
0.02
0.03


COD

0.00
0.33
0.36
0.34
0.29
0.30
0.40
C
Source
R


1.00
0.01
0.24
0.05
0.19
-0.04


P


1.00
0.04
0.16
0.10
0.17
0.06


P90


0.00
0.05
0.05
0.04
0.05
0.05


COD


0.00
0.40
0.39
0.35
0.35
0.42
D
Source
R



1.00
0.18
0.12
0.08
0.19


P



1.00
0.21
0.37
0.07
0.53


P90



0.00
0.01
0.01
0.02
0.01


COD



0.00
0.19
0.24
0.28
0.15
E
Non-Source
R




1.00
0.42
0.65
-0.01


P




1.00
0.36
0.40
0.07


P90




0.00
0.02
0.01
0.01


COD




0.00
0.24
0.24
0.20
F
Non-Source
R





1.00
0.54
0.42


P





1.00
0.41
0.45


P90





0.00
0.01
0.02


COD





0.00
0.24
0.26
G
Non-Source
R






1.00
0.01


P






1.00
0.05


P90






0.00
0.02


COD






0.00
0.27
H
Non-Source
R







1.00


P







1.00


P90







0.00


COD







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 3-51 illustrates Pb monitor locations with Jefferson County, MO. Ten source-
oriented monitors surrounded the Doe Run primary Pb smelter in Herculaneum, MO on
February 2012
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1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
32
33
34
35
36
37
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 m west of
the nearest building. Monitors A, B, C, D, F, G, and H were all located approximately
200 m west of the facility. Monitors D, E, and H were situated alongside service roads to
the facility. Monitor I was sited 100 m north of the smelter, and monitor J was located
approximately 600 m 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/yr (U.S. EPA. 2008f). Figure 3-52 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 3-53 illustrates the seasonal distribution of concentrations at monitors A-J in
Jefferson County. The annual average concentrations ranged from 0.18 to 1.36 (.ig/nr1
across the monitors. The maximum concentration was measured at monitor C to be
21.6 |ig/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 emissions measured at the closer monitors were
due to either fugitive emissions from the plant or, for the case where ground equipment or
vehicles are operated nearby, that 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 3-28).
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 m of each other. For the highest correlation
(R = 0 .96, p = 0.96, [for monitors E and F]), monitor F is 250 m 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 southeasterly
February 2012
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winds did not earn pollutants to these sites concurrently with the signal recorded by the
other monitors.
Legend
o T&PSowrca
Crty-tosafl P&p-uiaiion Center
m co«jioty-tMiM4 Population Cantor
1
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 3-51 Pb TSP Monitor locations within Jefferson County, MO (29-099),
2007-2009.
Febraaiy 2012
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Source: NRCS (20111
Note: Clockwise from top left: January, April, July, and October. Note wind percentages vary from month to month.
Figure 3-52 Wind roses for Jefferson County, MO, obtained from
meteorological data at St. Louis/Lambert International Airport,
1961-1990.
February 2012
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Site
A
B
C
D
E
F
G
H
I
J
SITE ID
29-099-
0022
29-099-
0024
29-099-
0015
29-099-
0023
29-099-
0004
29-099-
0020
29-099-
0021
29-099-
0005
29-099-
0011
29-099-
0013
MEAN
0.43
0.36
1.36
0.39
1.12
0.69
0.75
0.29
0.34
0.18
SD
0.54
0.49
1.97
0.54
1.67
1.01
1.25
0.59
0.85
0.33
OBS
622
209
1E3
632
1E3
575
953
351
366
177
% BELOW
MDL
0
5
0
0
5
0
5
25
5
15
Source
orientation
Source
Source
Source
Source
Source
Source
Source
Source
Source
Source
M
=1
iTj
U
7.0
6.5
6.0
5.5 -
5.0 -
4.5 -
4.0 -
3.5
3.0
a)
u
o 2.5 -
2.0 -
1.5 -
1.0 -
0.5
B
mil
t	1—i	—i—i	1—i—i	—i—i	1—i	1——i	r
D
i—i——i—i—i—
H
mil
lilii
t		1	1	1—I	1		1	1	1	1	1		1	1	1	r
ll,
a!
Y1 2 3 4 Y1 2 3 4 Y 1 2 3 4 Y1 2 3 4 Y1 2 3 4 Y1 2 3 4 Y1 2 3 4 Y 1 2 3 4 Y 1 2 3 4 Y1234
season
Figure 3-53 Box plots of annual and seasonal Pb TSP concentrations (jjg/m3)
from source-oriented and non-source-oriented monitors within
Jefferson County, MO (29-099), 2007-2009.
February 2012
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Table 3-28 Correlations between Pb TSP concentrations from source-oriented
and non-source-oriented monitors within Jefferson County, MO
(29-099), 2007-2009


A
B
c
D
E
F
G
H
I
J


Source
Source
Source
Source
Source
Source
Source
Source
Source
Source
A Source
R
1.00
0.66
0.80
0.84
0.60
0.65
0.33
0.32
0.07
0.05

P
1.00
0.59
0.80
0.83
0.57
0.64
0.33
0.35
0.07
0.05

P90
0.00
0.71
1.55
0.42
1.93
1.14
1.41
0.74
0.92
0.78

COD
0.00
0.46
0.48
0.30
0.55
0.45
0.57
0.64
0.67
0.69
B Source
R

1.00
0.54
0.40
0.15
0.15
0.08
0.16
0.11
0.01

P

1.00
0.53
0.43
0.10
0.14
0.07
0.22
0.10
0.09

P90

0.00
1.86
0.87
2.77
1.96
2.08
0.94
1.04
0.91

COD

0.00
0.58
0.51
0.69
0.62
0.68
0.68
0.65
0.65
C Source
R


1.00
0.86
0.56
0.72
0.28
0.32
-0.03
-0.03

P


1.00
0.86
0.59
0.72
0.26
0.27
-0.04
0.04

P90


0.00
1.56
2.26
1.26
2.94
2.65
3.18
2.60

COD


0.00
0.50
0.50
0.46
0.60
0.74
0.73
0.73
D Source
R



1.00
0.70
0.80
0.41
0.48
0.17
0.10

P



1.00
0.71
0.80
0.41
0.56
0.14
0.18

P90



0.00
1.83
1.02
1.38
0.76
0.88
0.70

COD



0.00
0.50
0.36
0.53
0.61
0.63
0.66
E Source
R




1.00
0.96
0.57
0.53
0.09
0.14

P




1.00
0.96
0.54
0.46
0.06
0.16

P90




0.00
0.86
2.16
2.50
3.09
2.57

COD




0.00
0.35
0.49
0.66
0.70
0.72
F Source
R





1.00
0.56
0.56
0.12
0.20

P





1.00
0.56
0.54
0.10
0.19

P90





0.00
1.13
1.51
1.74
1.40

COD





0.00
0.47
0.63
0.65
0.70
G Source
R






1.00
0.85
0.36
0.34

P






1.00
0.87
0.28
0.38

P90






0.00
1.53
2.10
2.08

COD






0.00
0.61
0.63
0.66
H Source
R







1.00
0.24
0.33

P







1.00
0.20
0.30

P90







0.00
0.89
0.56

COD







0.00
0.67
0.65
1 Source
R








1.00
0.87

P








1.00
0.79

P90








0.00
0.62
February 2012




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1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29


A
B
C
D
E
F
G H
1
J

COD







0.00
0.48
J Source
R








1.00

P








1.00

P90








0.00

COD








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 3-54 illustrates Pb monitor locations with 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. 20082). Additionally, monitor B was
located 30 m 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 m
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 m
southeast of 1-490. Monitors D and E were designated as non-source-oriented monitors,
although monitor D was just 600 m further from the Victory White Metal facility than
was monitor C. Monitor D was sited on a residential street located 50 m north of 1-490.
Monitor E was located on the rooftop of a building within 20 m of a four-lane arterial
road. Figure 3-55 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 3-56 illustrates the seasonal distribution of Pb concentration data at the five
monitoring sites. The influence of southern winds, along with close proximity to a
potentially-emitting facility, could have caused the elevated concentrations observed at
monitor B (average: 0.10 |ig/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 (ig/m3. Concentration data from all other monitors were below the level of the
NAAQS. For monitor A, the average concentration was 0.025 |ig/m3. and the median
reached 0.04 |ig/m3 during the summer. Maximum concentration at this monitor was
0.07 |ig/m3. Concentrations at monitor C averaged 0.017 (ig/m3, and those at monitors D
February 2012
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1
2
3
4
5
6
7
8
9
10
11
12
and E averaged 0.014 |ig/m3 and 0.013 |ig/nr\ respectively. Maximum concentrations
reached 0.04 |ig/m3 at all three monitors.
The level of spatial variability is illustrated by the intersampler correlations presented in
Table 3-29. 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.
February 2012
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.Cleveland
Legend
O TSP Source Monitors
•	TSP Non-source Monitors
•	City-based Population Center
•	County-based Population Center
	Interstates
	Major Highways
Bodies of Water
Urban Areas
Cuyahoga County, OH

Note: Top: view of al! 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 1-490.
Figure 3-54 Pb TSP Monitor locations within Cuyahoga County, OH (39-035),
2007-2009.
Febraaiy 2012
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Source: NRCS (20111
Note: Clockwise from top left: Jan, April, July, and October. Note wind percentages vary from month to month.
Figure 3-55 Wind roses for Cuyahoga County, OH, obtained from
meteorological data at Cleveland/Hopkins International Airport,
1961-90.
February 2012
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Site
A
B
C
D
E
SITE ID
39-035-0050
39-035-0049
39-035-0061
39-035-0038
39-035-0042
MEAN
0.025
0.10
0.017
0.014
0.013
SD
0.018
0.060
0.010
0.0072
0.0076
OBS
36
36
36
35
36
% BELOW MDL
20
0
30
45
45
Source orientation
Source
Source
Source
Non-source
Non-source
M
3
c
o
a)
u
o
u
25 i
24
23 i
22
21
20 i
19
18 i
17
16
15
14
13 i
12
11 i
10
09
08
07
06 i
05
04 H
03
02
01
00
A
-1—i—i—i—i—
Y 1 2 3 4
B
-1—i—i—i—i—
Y 1 2 3 4
ill
~i	1	1	r
D

Y 1 2 3 4 Y 1 2 3 4
season
n	1—i	r
Y 1 2 3 4
Figure 3-56 Box plots of annual and seasonal Pb TSP concentrations (jjg/m )
from source-oriented and non-source-oriented monitors within
Cuyahoga County, OH (39-035), 2007-2009.
February 2012
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1
2
3
4
5
6
7
8
9
10
11
12
13
Table 3-29 Correlations between Pb TSP concentrations from source-oriented
and non-source-oriented monitors within Cuyahoga County, OH
(39-035), 2007-2009



A
B
c
D
E



Source
Source
Source
Non-Source
Non-Source
A
Source
R
1.00
-0.06
0.21
0.17
0.24


P
1.00
-0.13
0.24
0.19
0.21


P90
0.00
0.18
0.05
0.04
0.05


COD
0.00
0.64
0.33
0.35
0.37
B
Source
R

1.00
0.26
0.43
0.11


P

1.00
0.31
0.24
0.34


P90

0.00
0.18
0.19
0.19


COD

0.00
0.69
0.71
0.73
C
Source
R


1.00
0.74
0.51


P


1.00
0.77
0.67


P90


0.00
0.01
0.01


COD


0.00
0.17
0.18
D
Non-Source
R
P



1.00
1.00
0.37
0.67


P90



0.00
0.01


COD



0.00
0.17
E
Non-Source
R




1.00


P




1.00


P90




0.00


COD




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 3-57 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. 200810. Monitors A and C are positioned along the facility's
service road and are approximately 100 m and 200 m away from the facility, respectively.
Monitor A is directly next to the road, and monitor C is roughly 15 m from the road.
Monitor B is located in the facility's parking lot roughly 50 m 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 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 3-58 displays
seasonal wind roses for Sullivan County. During winter and spring, the predominant
February 2012
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1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
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 3-59 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 |ig/m3. 0.051 |ig/m3. and 0.059 |ig/m3. respectively. Median
concentrations were 0.08 |ig/m3. 0.03 (ig/m3, and 0.04 |ig/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 (ig/m3, 0.26 |ig/m3. and 0.43 (ig/m3 for monitors A, B, and C. It
was surprising that 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 C 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 3-30). 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.
February 2012
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Bristol
Legend
o TSP Source Monitors
•	City-based Population Center
•	County-based Population Center
Interstates
	Major Highways
Bodies of Water
Urban Areas
Sullivan County, TN
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 3-57 Pb TSP Monitor locations within Sullivan County, TN (47-163),
2008-20102007-2009.
February 2012
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Source: NRCS (20111
Note: Clockwise from top left: January, April, July, and October. Note that the wind percentages vary from month to month.
Figure 3-58 Wind roses for Sullivan County, TN, obtained from meteorological
data at Bristol/Tri City Airport, 1961-90.
February 2012
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Site
SITE ID
MEAN
SD
OBS
% BELOW MDL
Source orientation
Figure 3-59
47-163-3001
0.11
0.11
334
Source
M
i
c
0
0.44 "
0.42 "
0.40 -
0.38 "
0.36 -
0.34 -
0.32 "
0.30 -
0.28 "
0.26 "
0.24 "
0.22 "
0.20
0.18
0.16
0.14
0.12
0.10
0.08
0.06
0.04
0.02
0.00
47-163-3002
0.051
0.036
362
Source
B
47-163-3003
0.059
0.047
345
Source
n—i—i—i—i——i—i—i—i—i——i—i—i—i—r
Y 1 2 3 4 Y 1 2 3 4 Y1234
season
Box plots of annual and seasonal Pb TSP concentrations (jjg/m3)
from source-oriented monitors within Sullivan County, TN
(47-163), 2007-2009.
February 2012
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Table 3-30 Correlations between Pb TSP concentrations from source-oriented
monitors within Sullivan County, TN (47-163), 2007-2009



A
B
c



Source
Source
Source
A
Source
R
1.00
0.06
0.14


P
1.00
-0.04
0.13


P90
0.00
0.21
0.19


COD
0.00
0.47
0.43
B
Source
R

1.00
0.31


P

1.00
0.45


P90

0.00
0.06


COD

0.00
0.23
C
Source
R


1.00


P


1.00


P90


0.00


COD


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).
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3.8.3 Lead Concentration in a Multipollutant Context
Source, S02
Non-Source, S02
Source, PMJ5
Non-Source, PM2,s
Source, PM.a
Non-Source, PM,0
Source, 0$
Non-Source, 03
Source, N02
Non-Source, N0S
Source, CO
Non-Source, CO
US Wintoif
<> qoooooqdi
o o
oo
qo m m ffiomn ex® o m o ocxxs #
o	o o o
O O OO CD CMMomt
o o	o
o oo<
o
oo® cm o o
o o
®	omom
o o
o cd m mmxmammosm m o o#
-1.0
-0.8	OJ	0.5
Spearman Correlation Coefficient
1.0
Source, SO?
Non-Source, S02
Source, PlVfe.s
Non-Source, PM2,5
Source, PM,0
Non-Source, PM.a
Source, O,
Non-Source, 03
Source, NOz
Non-Source, N02
Source, CO
Non-Source, CO
US Spring
QD
o oo
o
o
oo
o	o
mmm oo
o o o
O® 0 
o o o
O ® O® (SKSIiKlSHII) CISC ® 0
o o	o
MMQBD GOO O OOD
O O
ooo ooffl) OMnoso
o o
oowamonoono o 00$
-1.0
-0.5	0.0	0.5
Spearman Correlation Coefficient
1.0
Note: Top: winter; Bottom: spring.
Figure 3-60 Seasonal correlations of monitored Pb-TSP concentration with
copollutant concentrations, 2007-2008.
February 2012
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Source, SO?
Non-Source, S02
Source, PMj5
Non-Source, PM2,$
Source, PM«
Non-Source, PM«j $
Source, 03
Non-Source, O3
Source, NO?
Non-Source, N02 ^
Source, CO
Non-Source, CO 4
US Summer
o 00
® o om
o o©
o o QfflonaiDoot® amo am
00
o o®«a»®«i«ioci)®o0o o
-1.0
¦0.5	0.0	0.5
Spearman Correlation Coefficient
1,0
Source, SOz J US Fill
Non-Source, S02
Source, PM25
Non-Source, PM2,S
Source, PM,o
Non-Source, PMt0
Source, O3
Non-Source, O3
Source, NOz
Non-Source, NO?
Source, CO
Non-Source, CO
o o
o 00 coodoo o ® o oo o go
aoo o
o 000
00 •
OOO OOD OQ
OO	OO
ooa»o®coi» 000 moonymmmom oo o
000
>O0 0K>
O O
OO
omoQXMMMaMmmmm
0 o
o m oammoommmmmm
1,0
-0.5	0.0	0.5
Spearman Correlation Coefficient
1.0
Note: Top: summer; Bottom: fall.
Figure 3-61 Seasonal correlations of monitored Pb-TSP concentration with
copollutant concentrations, 2007-2008.
February 2012
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-------
Non-Source, S02 ! "	O O O	O
Source, PM25 ! "	0
Non-Source, PM2 5 »-	O	O	O O O	O
Non-Source, PM10 1 -	O OOO	O <>
Non-Source, 03 < - O	O	 -	O	O
Non-Source, CO . -	O	O 00	O
	1	1	1	
-1.0	-0.5	0.0	0.5	1.0
Spearman Correlation Coefficient
US Winter
OOO	o
o
O	OOO	o
O OOO	o 
-------
Non-Source, SO2	"
Source, PM2.5	-
Non-Source, PM2 5	-
Non-Source, PM,o	1
Non-Source, 03	-
Non-Source, N02	-
Non-Source, CO	4
-1.0	-0.5	0.0	0.5	1.0
Non-Source, SO2
Source, PM25
Non-Source, PM2s
Non-Source, PM10
Non-Source, O3
Non-Source, NO2
Non-Source, CO -	O O	O	O
	1	1	1	
-1.0	-0.5	0.0	0.5	1.0
Spearman Correlation Coefficient
Note: Top: summer: Bottom: fall.
Figure 3-63 Seasonal correlations of monitored Pb-TSP concentration with
copollutant concentrations, 2009.
US Summer
o o o	o
o
o o o	o
o	o o
O	O OBD O
o o
o	oo o	o
Spearman Correlation Coefficient
US Fall


8
o
o

o
8
o
o
o
o
o
o
o
o

o o oo o o o

o o
o
o o o
i	1	r
.0	-0.5	0.0	0.5	1.
Spearman Correlation Coefficient
Februaiy 2012
3-196
Draft - Do Not Cite or Quote

-------
Zn -
K -
Br -
Cu -
Fe -
N03- -
S -
oc -
EC	-
Crystal	—
by*.	-
Se	-
S04-2	-
NH4+	-
K+	-
Ca	-
Si	-
Ti	-
ft	-
V -
Cr	-
Ni	-
nvol N03-	—
CI	-
Na	-
Hg	-
vol N03-	-
Cd	-
Na+2	-
Mg	-
As	-
COCO OI--
O	OH' h -
O O O I	
O I	
o O I	
O O OOO ij	
o o o I-
O O I—
O O I	
o o o I'-
ll OOI	
O ID ODCDh-	
O O I- -
-	OOO
-HO
00 oo
o o o 0:01- -
00 o o or 1	
o o oa	
° ~
--I I I-
- H O O	O
	10 o
	10 o
Zn	-
Cu	-
K	-
Fe	-
Crnstal	—
Br	-
IVh	-
Ti	-
Ca	-
S	-
S04-2	-
Si	-
ft	-
EC	-
Se	-
OC	-
N03-	-
NH4+	-
K+	-
Cr	-
V	-
vol N03-	-
As	-
Ni	-
Mg	—
Cd	-
Hg	-
Na	-
CI	-
Na+2	-
nvol N03-	—
	~
	c
~	
	£
h	r
~	H
3-
~	
	C
ol	~
O,	[—
H	^
~	
-J	
n-	
~	
^	
- -to o 00
~J H	C
	d
~	
o h	~
Ol	
~	H
=!	h
~ ~	CO ~
~	_
m---
3	
~I	
-c
	c
.n. ini 1	
3	011
~	1 o
-c
O IIDI	
	CZ
o 0x11	r
----c
I	1 o o
D	-\ O
I-	i'H"iHi
ZD-	-<
~-	' o
~	
Top: winter; bottom: spring. Note: "nvol" = non-volatile, "vol" = volatile, and organic carbon (OC) samples were blank-adjusted.
Figure 3-64 Seasonal correlations of monitored Pb-PM2.5 concentration with
copollutant concentrations, 2007-2009.
February 2012
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Draft - Do Not Cite or Quote

-------
Zn	-
Br	-
Cu	-
K	-
S	-
S04-2	~
Ca	-
lUfrl	-
Crystal	—
Fe	-
Se	-
N03-	-
Ti	-
EC	-
OC	-
A	~
Si	-
Mg	—
K+	-
v -
NH4+	-
Na	-
Cr	-
Hg	"
Cd	-
Na+2	-
vol N03-	-
Ni	-
CI	-
As	—
nvol N03-	—
O 'IE' (TO- -
- H oo
-HO O
O O O O
3	H
=F	
Y —
---¦C
~	H
Zn -
K -
Cu -
Br -
OC -
EC -
Fe -
vol N03- -
Crystal —
Mn -
Ca -
K+ -
S -
N03- -
S04-2 -
NH4+ -
Si -
Ti -
A -
Se -
Cr -
V -
nvol N03- —
Mg -
Na -
Na+2 -
As -
Ni -
Hg "
CI -
Cd -
O OO O I	
COOK -
O 'HI O
O O OOO I- -
~ O OO I	
HZD
O OO
O OO	OI	
O	'HI CO!	
O OO	I	
O O	O Ol	
III I- -
I
O O I	
O i I--
O I	
OOO I- -
	1000
-HO (
Top: summer; bottom: fall. Note: "nvol" = non-volatile, "vol" = volatile, and organic carbon (OC) samples were blank-adjusted.
Figure 3-65 Seasonal correlations of monitored Pb-PM2.5 concentration with
copollutant concentrations, 2007-2009.
February 2012
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Table 3-31 Copollutant exposures for various trace metal studies

Adgate et al. (2007)
Riedikeretal. (2003)
Pekey et al.
(2010)

Molnaretal. (2007)

l-R (med)3-"
Personal
(median)0
Vehicle
(range)0
Roadside
(range)0
I -near industry
(range)3
l-R
(median)
l-School
3,b (median)3
l-Pre-School
(median)3
Location
Minnesota

New Jersey

Kocaeli, Turkey
Stockholm, Sweden

PM2.5


24,000
31,579
24,400-29,800



Pb
1.5
3.2
2-3
4-6
34-85
2.8
2.5
1.7
S
272.1
351.6
905-1592
1416-2231
435-489
330
290
220
Ca
85.0
174.1
31-44
18-40
309-452
70
110
58
Al
23.3
58.6


53-60



Na
20.6
31.9






Fe
43.1
78.6
307-332
82-163
44-58
57
100
71
Mg
16.3
27.5






K
38.4
47.5
6-75
23-57
160-215
120
96
67
Ti
0.8
1.4
9-10
6-10
29-39
8.0
13
8.7
Zn
6.5
9.6
5-10
14-17
51-88
14
17
11
Cu
1.-0.15
4.9
18-32
8-16
21-58
9.3
1.7
2.1
Ni
2.4
1.8
0
0
2-3
0.99
1.0
0.72
Mn
0.21
2.3
3-4
3
28-32
2.2
2.5
2.1
Sb
0.12
0.30






Cd
0.12
0.14
4-6
4-7




V
0.05
0.16
1
1
3-5
2.5
2.7
1.8
La
0.00
0.11






Cs
0.00
0.00






Th
0.00
0.00






Sc
0.00
0.01






>
CO
0.07
0.08






Co
0.02
0.07






Cr
1.2
2.6
2
1
3-8
<1.1
1.3
1.1
Si


198-464
338-672
387-401



CI


7-32
3-9




Se


1
1-2




Rb 1 1
Sr


5-28
1




As


1
1
1-2



Mo
Br





2.1
1.3
1.3
al: Indoor; Units: ng/m3
bR: Residential; Units: ng/m3
Units: ng/m3
1
2
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CHAPTER 4 EXPOSURE, TOXICOKINETICS,
AND BIOMARKERS
4.1 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 6). Blood Pb and bone Pb biomarkers (discussed in Section 4.3, Section 4.4,
Section 4.5, and Section 4.6), 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
(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 (ig/m3 for non-source-oriented monitors for the years 2000-2004 and
were 0.10-0.22 (.ig/nr1 for source-oriented monitors during that time period, while blood
Pb levels reached a median of 1.70 (ig/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.
4.1.1 Pathways for Lead Exposure
Pathways of Pb exposure are difficult to assess because Pb has multiple sources in the
environment and passes through various environmental media. These issues are described
in detail in Sections 3.2 and 3.3. 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-
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related Pb exposures also include inhalation and ingestion of Pb-contaminated food,
water or other materials including 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.
Particle size of Pb-PM is relevant to transport through various media leading to exposure.
The inhalability of airborne particles is a gradually decreasing function of particle size.
Inhalability criteria established from experimental data obtained at wind speeds of 1-8
m/s describe PM inhalability of 77% for particles <10 |a,m (dae, aerodynamic diameter).
Inhalability of particles ranging in size from 40 to 100 (.un dae is 50%; above 100 |_im.
inhalability data are lacking (Soderholm. 1989; AC'GIH. 1985). Of the particles that are
not inhaled, their settling to surfaces makes them available for subsequent ingestion. The
main pathway for Pb ingestion by children is by hand to mouth (Lanphearet 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 (j,m, 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. Differences in the size
distribution results may be related to differences in the soil between the two locations
and/or to differences between the analytical methods used to measure size distribution;
Duggan and Inskip (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. Similar studies focusing on particle size distributions of ingestion of
house dust are lacking. Ingestion of house dust has been reported to be the major source
of lead 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.
The complicated nature of Pb exposure is illustrated Figure 4-1, 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 the Pb
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passed through the air compartment. When animals consume plant material exposed to
Pb that has at some point passed through the air compartment, and when human diet
includes animals and/or plants 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 difficult}' of
constructing Pb exposure histories, most studies of Pb exposure through air, water, and
soil can be informative to this review. Figure 4-1 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.
^ Newly Emitted Pb
Historically Emitted Pb
OUTDOOR SOIL
\and dust
NATURAL WATERS
AND SEDIMENTS
C>/ Paint
AIR
SOIL
WATER
PLANT
EXPOSURE,
ANIMAL
EXPOSUREl
v	/
HUMAN
EXPOSURE
AIR
SOIL
WATER
•"jEWELRr"
COSMETICS
--TOYS etc.^
(occupation
rinking WaterN
PIPES )
AIR
SOIL
WATER
Note: The Venn diagram is used to illustrate the passage of Pb through multiple environmental media compartments through which
exposure can occur.
Figure 4-1 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
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environment, location, and risk factors (described in more detail in 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) (Clavton et al.. 1999) as well as in Arizona (O'Rourke et al.. 1999) and
Maryland (Egeghv et al.. 2005). Results from NHEXAS indicate that personal exposure
concentrations of Pb are higher than indoor or outdoor concentrations of Pb (Table 4-1).
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.
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Table 4-1 Estimates of Pb measurements for EPA Region 5 from the NHEXAS
study
Medium3
N
Percentage
measurable13
(CLs)c
Mean (CLs)c
50th (CLs)c
90th (CLs)c
Personal air (ng/m3)d
167
81.6 (71.3; 92.0)
26.83 (17.60; 36.06)
13.01 (11.13; 18.13)
57.20 (31.18; 85.10)
ndoor air (ng/m3)d
213
49.8 (37.2; 62.3)
14.37 (8.76; 19.98)
6.61 (4.99
8.15)
18.50(12.69; 30.31)
Outdoor air (ng/m3)d
87
73.8 (56.3; 91.3)
11.32 (8.16; 14.47)
8.50(7.14; 10.35)
20.36(12.60; 34.91)
Surface dust (ng/cm2)
245
92.1 (87.4; 96.8)
514.43 (-336.6; 1365.5)
5.96 (3.37; 10.94)
84.23 (26.52; 442.63)
Surface dust (mg/kg)
244
92.1 (87.4; 96.8)
463.09 (188.15; 738.04)
120.12 (83.85; 160.59)
698.92 (411.84; 1,062.8)
Window sill dust (ng/cm2)
239
95.8 (92.5; 99.0)
1,822.6 (481.49; 3,163.6)
16.76 (10.44; 39.41)
439.73 (106.34; 4,436.2)
Wndow sill dust (mg/kg)
239
95.8 (92.5; 99.0)
954.07 (506.70; 1,401.4)
191.43(140.4
3; 256.65)
1,842.8 (1,151.3; 2,782.5)
Standing tap water (pg/L)
444
98.8 (97.6; 100.0)
3.92 (3.06; 4.79)
1.92(1.49
2.74)
9.34 (7.87; 12.35)
Flushed tap water (pg/L)
443
78.7 (70.7; 86.7)
0.84 (0.60; 1.07)
0.33 (0.23
0.49)
1.85 (1.21; 3.04)
Solid food (pg/kg)
159
100.0 (100.0; 100.0)
10.47 (6.87; 14.07)
6.88 (6.44
8.04)
14.88(10.78; 19.08)
Beverages (pg/kg)
160
91.5 (85.2; 97.8)
1.42 (1.13; 1.72)
0.99 (0.84
1.21)
2.47 (2.06; 3.59)
Food+Beverages (pg/kg)
156
100.0 (100.0; 100.0)
4.48 (2.94; 6.02)
3.10(2.66
3.52)
6.37 (4.89; 8.00)
Food intake (pg/day)
159
100.0 (100.0; 100.0)
7.96 (4.25; 11.68)
4.56 (3.68
5.36)
12.61 (9.27; 16.38)
Beverage intake (pg/day)
160
91.5 (85.2; 97.8)
2.15 (1.66; 2.64)
1.41 (1.18
1.60) 4.45 (3.15; 5.65)
Food+Beverage intake (pg/day)
156
100.0 (100.0; 100.0)
10.20 (6.52; 13.89)
6.40 (5.21
7.78)
16.05(13.31; 18.85)
Blood (pg/dL)
165
94.2 (88.2; 100.0)
2.18 (1.78; 2.58)
1.61 (1.41
2.17)
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 measurable, 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 measurable is the percentage of the target population of residents (or households) estimated to have Pb levels above limit of detection (LOD).
cThe 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)
Several studies have used a combination of measured values and default model values to
represent exposures and determine their relative contributions to blood Pb. 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. They point out that ambient
air Pb concentrations decrease with distance from the smelter while IEUBK is fairly
insensitive to changes in ambient air concentration. They input the average ambient air
Pb concentration of 0.81 |ag/m3 and assumed that the indoor air Pb concentration was
30% of the outdoor Pb. In the absence of indoor air Pb samples, the default assumption
was adopted for the Cornelis et al. (2006) study. Similarly, Carrizales et al. (2006)
analyzed exposures to children living near a copper smelter in San Luis Potosi, Mexico.
They employed the IEUBK default options for assignment of Pb dust concentration as
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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. Input soil and dust Pb levels were obtained
from an average of 28 sites. Using defaults for all IEUBK values, the soil/dust Pb
ingestion pathway has been predicted to contribute 63-75% (depending of a child's age)
of blood Pb, whereas the air Pb inhalation pathway contributed <1% (SRC. 2007). This
analysis did not specifically estimated 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. In the 1986 Pb AQCD, blood Pb-air Pb
slopes were examined that included both direct (inhalation) and indirect (via soil, dust,
etc.) air Pb contributions, resulting in the conclusion that slopes for the total air Pb
contribution were roughly twice that of the slope due to inhaled air Pb alone. This "recent
air" Pb was described in the 2007 Pb Risk Assessment ("U.S. EPA. 2007f) to include
those pathways involving Pb that is or has recently been in the outdoor ambient air,
including inhalation and ingestion of indoor dust Pb derived from recent ambient air
(e.g., air Pb that has penetrated into the residence recently and loaded indoor dust).
Beyond the direct inhalation pathway (outdoor and indoor ambient air Pb), Appendix I of
the 2007 Pb Risk Assessment (U.S. EPA. 2007f) provides estimates of the contribution of
various other pathways (e.g., diet, drinking water, outdoor soil ingestion) as well as those
involving recent air Pb to blood Pb. Estimates of recent air Pb contributions were highly
dependent on the exposure scenario evaluated. For the General Urban Case Study in that
assessment, "current conditions" at the time of the risk assessment (0.14 |_ig/m3, mean of
maximum quarterly average concentration of Pb in TSP) were modeled as remaining
constant throughout the 7 years of the exposure modeled in the biokinetic model. A
median concurrent blood Pb concentration of 1.8-2.8 (ig/dL (depending on the exposure
and risk model) was estimated as the average of the results at 75 and 81 months of age.
The recent air Pb ingestion pathways contributed a median of 12-28% (depending on the
exposure model) to blood Pb beyond the 0.5-0.6% attributable to inhalation of recent
airborne Pb. In simulations for the 95th percentile of current conditions (0.87 |ag/m3.
95th percentile of maximum quarterly average concentration of Pb in TSP), median
children's blood Pb increased to 2.0-3.1 (ig/dL and the contribution of recent air Pb
ingestion pathways increased to 22-37% of blood Pb. Results of other exposure scenarios
for which greater or lesser contributions of recent air Pb to blood Pb were predicted are
available in Appendix I of the 2007 Pb Risk Assessment (U.S. EPA. 2007f).
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4.1.2 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 4.1.3. Blood Pb sampling is
described in detail in Section 4.3.2.
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 3.4, as are ambient air
Pb monitoring techniques. Factors affecting collection of ambient air Pb samples are
described in detail in Section 3.4. 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
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 lead 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, is 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 3.5. 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
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higher geometric mean collection (vacuum 1: 94.3 jj.g/ft2; vacuum 2: 23.5 jj.g/ft2)
compared with dust wipes (5.6 jj.g/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 et al.. 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. 2007s). 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.
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).
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 et al. 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 Liov. 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.
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4.1.3 Exposure Studies
4.1.3.1 Airborne Lead Exposure
1	Limited personal exposure monitoring data for airborne Pb were available for the 2006
2	Pb AQCD (U.S. EPA. 2006b). As described above, the NHEXAS study showed personal
3	air Pb concentrations to be significantly higher than indoor or outdoor air Pb
4	concentrations (Clavton et al.. 1999). Indoor air Pb concentration was moderately
5	correlated with floor dust and residential yard soil Pb concentration (Rabinowitz et al..
6	1985). Egeghy et al. ("2005) performed multivariate fixed effects analysis of the
7	NHEXAS-Maryland data and found that Pb levels measured in indoor air were
8	significantly associated with log-transformed outdoor air Pb levels, ambient temperature,
9	number of hours in which windows were open, homes built before 1950, and frequency
10	of fireplace usage (Table 4-2).
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Table 4-2 Estimates of fixed effects multivariate modeling of Pb levels
measured during the NHEXAS-MD study

Pb in Indoor Air
Pb in Dust
Dermal Pb
Blood Pb
Fixed Effect
pa p-value
Pa
p-value
Pa
p-value
Pa
p-value
Intercept
-0.50 0.0051
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




-0.12
0.088
Home age (<1950)
0.25 0.025
0.96
0.029




Fireplace (frequency of use)
0.11 0.045
0.46
0.0054




Pb concentration in soilb

0.27
0.037




Interior Pb paint chipping/peeling (yes)

0.43
0.091




Cement at primary entryway (yes)

1.97
0.0064




Indoor pesticide usage last 6 mo (yes)

-0.78
0.0003




Electrostatic air filter usage (yes)

-0.91
0.062




Sex of participants (male)



0.41
0.0012
0.43
<0.0001
Ethnic minority participants (yes)



0.41
0.0063


Washing hands after lawn mowing (no)



1.04
0.0010


Gasoline power- equipment usage (yes)



0.61
0.0072


Bathing or showering activities (yes)



-0.43
0.019


Dust level indoors (scale: 1-3)



0.22
0.019


Residing near commercial areas (yes)



0.32
0.0087


Age of participants (yr)





0.02
<0.0001
Number cigarettes smoked (count)





0.03
<0.0001
Burning wood or trash (days)





0.58
0.0099
Showering frequency (avg # days)





-0.29
0.0064
VNfork outside home (yes)





-0.26
<0.0001
Health status (good)





0.23
0.0009
Adherence to high fiber diet (yes)





-0.15
0.040
Gas or charcoal grill usage (yes)





-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 measured personal exposure to particle-bound Pb along with
other trace metals. 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
4-3). Another study of indoor and outdoor air concentrations of Pb was carried out by
Molnar et al. (2007). PM2 5 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-PM2 5 suggest an outdoor Pb-PM2 5 net
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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 roadwear on days
with dry surfaces. Pekey et al. (2010) measured indoor and outdoor trace element
composition of PM25 and PMi0 in Kocaeli, an industrial region of Turkey, and found that
average airborne Pb concentrations were higher outdoors than indoors for both PM2 5 and
PMio 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 PM25 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.
Domestic wood burning is a potential source of Pb compounds (Section 3.2.2.5). 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, but that the
differences were not significant (indoor concentration: 6.0 (ig/m3 vs. 4.3 (ig/m3, p = 0.26;
personal exposure: 4.6 (ig/m3 vs. 3.0 (ig/m3, p = 0.06).
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 1.4xl0"4 hr"1 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 (.un for activity vs. 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
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detached from the carpet by walking motion, but that smaller particles are more easily
resuspended once detached. Hunt and Johnson (In Press) studied the duration and spatial
extent of resuspension of 0.3-5.0 (.un 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. If Pb
adsorption onto particles varies with sources, size selectivity of resuspension processes
could lead to enhanced Pb exposure from one source relative to another.
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 and N02 in three outdoor microenvironments (residence,
school, and preschool) and found that Pb and other trace metals were generally well
correlated with PM2 5 (r = 0.72-0.85), but Pb was not always well-correlated with N02 (r
= 0.24-0.75). In the case where Pb and N02 were well-correlated, it is possible that the Pb
was traffic related from resuspended pulverized wheel weights or impurities in unleaded
on-road gasoline. For the other two sites where the correlation between Pb and N02 was
low, it is possible that they were less affected by traffic. Table 3-26 in the Appendix to
Chapter 3 illustrates that Pb concentrations in the four studies summarized there are
typically well below the level of the NAAQS. The higher personal air concentratoins
occurred in a heavily industrialized area of Kocaeli, Turkey with an incinerator and
several industrial facilities including metal processing, cement, petroleum refining,
agriculture processing. Otherwise, concentrations were all between 0.002 and
0.006 |ig/nr\ 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 Ca and Fe. In the industrial area of Kocaeli, Pb comprised a greater proportion of
the PM compared with other areas.
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Table 4-3
Comparison of personal, indoor, and outdoor Pb-PM measurements
from several studies
Study
Location
Pb Metric
Sampling
Period
Personal Pb
Indoor Pb
Outdoor Pb
Clayton et al. (1999)
IL, IN, Ml, MN,
OH, Wl
Med. Pb-PMso
(ng/m3)
July, 1995-
May, 1997
13
6.6
8.5
Adgate et al. (2007)
Minneapolis-
St. Paul, MN
Avg. Pb-PM25
(ng/m3)
Spring,
Summer, Fall,
1999
6.2
3.4
2.0
Molnar et al. 12007)
Stockholm, Sweden
Avg. Pb-PM2.5
(ng/m3)
December,
2003-July, 2004

Homes: 3.4
Schools: 2.5
Preschools: 1.8
Homes: 4.5
Schools: 4.6
Preschools: 2.6
Tovalin-Ahumada et
al. 12007)
Mexico City, Mexico
Med. Pb-PM2.5
(ng/m3)
April-May, 2002

26
56

Puebla, Mexico
Med. Pb-PM2.5
(ng/m3)
April-May, 2002

4
4
Pekey et al. (2010)
Kocaeli, Turkey
Avg. Pb-PM2.5
(ng/m3)
May-June,
2006,
December,
2006-
January 2007

Summer: 34
Winter: 85
Summer: 47
Wnter: 72
Avg. Pb-PM,o (ng/m3)
May-June,
2006,
December,
2006-
January 2007

Summer: 57
Wnter: 125
Summer: 78
Wnter: 159
Rasmussen et al.
(2007)
Windsor, Ontario,
Canada
Med. Pb-PM2.5
(mg/kg)
April, 2004
311
124
221
4.1.3.2 Exposure to Lead in Soil and Dust
1	The 2006 Pb AQCD (U.S. EPA. 2006b') lists indoor Pb dust infiltrated from outdoors as a
2	potential source of exposure to Pb soil and dust. Thus, outdoor soil Pb may present an
3	inhalation exposure if resuspended indoors or an ingestion exposure during hand-to-
4	mouth contact. A detailed description of studies of outdoor soil Pb concentration is
5	provided in Section 3.6.1. Indoor measurements can reflect infiltrated Pb as well as Pb
6	dust derived from debrided paint, consumer products, or soil that has been transported
7	into the home via foot traffic. Table 4-4 presents indoor dust Pb concentrations for
8	2006-2011 observational studies in which indoor dust Pb was measured.
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Table 4-4 Measurements of indoor dust Pb concentration from 2006-2011
studies
Reference
Study Location
Metric (units)
Sample Site
Indoor Pb
Concentration
Caravanos et al.
(2006a)
New York City, New York
VNfeekly dust loading
(pg/m)
Glass plate next to open window
of academic building
Median: 52
Khoder et al. (2010)
Giza, Egypt (extensive leaded gasoline use;
industrial area)
VNfeekly dust loading
(pg/m2)
Glass plate in second-floor living
room of apartments
Median: 408
Brattin and Griffin
(2011)
Eureka, Utah near Eureka Mills Superfund
Site
Dust concentration
¦ (mg/kg)
Indoor home site (not specified)
160-2000
Denver, CO, nearVBI70 Superfund Site
Indoor home site (not specified)
11-660
East Helena, MT, near East Helena
Superfund Site
Indoor home site (not specified)
68-1000
Yu et al. (2006)
Syracuse, New York
Dust concentration
range (mg/kg)
Floor
Range: 209-1770
Turner and
Simmonds (2006)
Birmingham, Plymouth, and 2 rural sites, UK
Dust concentration
(mg/kg)
Floor
Median: 178



Smooth floor
Median: 1.7
Avg.: 4.4
Gaitens et al. (2009)
U.S. (nationwide)
Dust loading (pg/m2)
Rough floor
Median: 5.6
Avg.: 16
Smooth windowsill
Median: 2.5
Avg.: 190



Rough windowsill
Median: 55
Avg.: 480


Dust concentration
(pg/m2)
Central perimeter
Avg.: 107
Wilson et al. (2007)
Milwaukee, Wisconsin
Entry
Avg.: 140


Wndow
Avg.: 151
Zota et al. (2011)
Ottawa County, Oklahoma (area
surrounding the Tar Creek Superfund Site)
Dust concentration
(mg/kg)
Indoor (site not specified)
Avg.: 109
Median: 63
Max.: 881

Rural towns, Idaho
Dust concentration
Vacuum
Median: 120
Max: 830
Spalinger et al.
(mg/kg)
Floor
Median: 95
Max: 1,300
(2007)
Bunker Hill, Idaho Superfund site
Dust concentration
Vacuum
Median: 470
Max: 2,000

(mg/kg)
Floor
Median: 290
Max: 4,600
1	Several studies suggested the infiltration of Pb dust into buildings. For example,
2	Caravanos et al. (2006a) collected dust on glass plates at an interior location near an open
3	window, a sheltered exterior location, and an open exterior location for a two-year period
4	in Manhattan, NY. Median weekly dust loading was reported to be 52 (ig/m2 for the
5	indoor site, 153 (ig/m2 for the unsheltered outdoor site, and 347 |_ig/m2 for the sheltered
6	outdoor site. This paper demonstrated the likely role of outdoor Pb in influencing indoor
7	dust Pb loading and indicated that under quiescent conditions (e.g., no cleaning) near an
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open second-story window, the indoor dust Pb level might exceed EPA's hazard level for
interior floor dust of 430 (.ig/nr (40 jxg/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/nr 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. 2002a. 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 (.un 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 this size fraction.
Residual Pb dust contamination following cleaning activities has been documented. For
instance, Hunt et al. (2008) performed tests where a test soil prepared by drying,
grinding, and sieving Pb-contaminated yard soil samples from Herculaneum, MO was
tracked onto a tile test surface and then repeatedly cleaned with a moistened wipe and/or
vacuumed until visual inspection of the tiles uncovered no surface discoloration. The
authors then used wet wipe samples to collect residual soil and estimate Pb deposition
and concentration. After the first walk, tile Pb dust loading was 2,670 |_ig/nr: after the
first vacuuming, it decreased to 398 (ig/m2. After multiple walks, tile Pb dust loading was
7,100 (.ig/nr. and it decreased to 1,400 (.ig/nr after multiple vacuuming. Scanning
electron microscopy (SEM) of the wipe samples revealed that most of the residual dust
particles were in the range of 1-3 |_im in area equivalent diameter. This result indicates
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that Pb-bearing fine particles are not completely captured by home 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
obtained dust from vacuuming carpet samples and found that total Pb concentration in
carpet ranged from 209 to 1,770 mg/kg dust.
Pb dust on floors, windowsills, and other accessible surfaces are potential exposure
sources to small children who use touch to explore their environments. Gaitens et al.
(2009) used National Health and Nutrition Examination Survey (NHANES) data from
1999 through 2004 to examine Pb in dust in homes of children ages 12-60 months. The
median value of Pb dust loading on floors was reported to be 1.7 (ig/m2 (mean:
4.4 (ig/m2), with floors that were not smooth and cleanable having a median Pb dust
loading value of 5.6 (ig/m2 (mean: 16 (.ig/nr). Floor Pb dust loading value was modeled
against several survey covariates and was significantly associated (p <0.05) with floor
surface condition, windowsill Pb dust loading, race and ethnicity, poverty-to-income
ratio, year of home construction, presence of smokers in the home, and year of survey. It
was nearly significantly associated (p = 0.056) with renovations made to pre-1950 homes.
Median Pb dust loading on smooth windowsills was 25 |_ig/m2 (mean: 190 (.ig/nr). When
windowsills were not smooth, the median Pb dust loading was 55 (ig/m2 (mean:
480 (ig/m2). Windowsill Pb dust level was also significantly associated (p <0.05) with
race and ethnicity, year of home construction, window surface condition, presence of
smokers in the home, deterioration of indoor paint, and year of survey. Sill surface was
nearly significantly 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 associated with the log-transform of soil Pb concentration, cement content in
the home entryway, indoor pesticide use, frequency of fireplace usage, number of hours
in which windows were open, and homes built before 1950 (Table 4-2). Wilson et al.
(2007)	studied Pb dust samples from homes in Milwaukee, WI 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.
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
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, while the other was hand-scraped. Immediately after
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sanding, Pb dust samples ranged from <3 to 28,000 mg/kg at the sanded house. Pb dust
samples from the scraped house ranged from 7 to 1,200 mg/kg. Pb in dust or paint
samples was not quantified.
Dust Pb concentrations have also been reported for homes in the vicinity of historic and
active metals mining and smelting sources. Near an active smelter in Port Pirie, Australia,
median hand dust Pb loadings increased with age among a cohort of fourteen children
followed overage 0-36 months (2-5 months: 54 (ig/m2, 6-9 months: 173 (ig/m2, 10-15
months: 424 (.ig/nr. >15 months: 336 (.ig/nr) (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-PM25 concentration of 1 ng/m3. Dust Pb
concentrations were found to increase significantly with 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 formerly a Pb and Zn smelter 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 a maximum of 830 mg/kg, and Pb concentration in floor dust
had a median of 95 mg/kg and a maximum of 1,300 mg/kg. The median Pb dust loading
rate was measured to be 40 |_ig/m2 per day. Among the background homes, median
vacuum and floor mat Pb dust samples were 3 and 2.5 times higher, respectively, when
comparing homes built before 1960 with those built after 1960. Deposition rate of Pb
dust was 5 times higher in the older homes. 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 with a maximum of 2,000 mg/kg and a median of 290 mg/kg
with a maximum of 4,600 mg/kg, respectively. The median Pb loading rate for indoor
dust in houses in these towns was 300 (ig/m2 per day, and the maximum Pb dust loading
rate was 51,000 (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. (2011b) reported
median soil Pb concentration of 558 mg/kg on playground soils at eleven New Orleans
daycare 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)
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reported on the concentration and size distribution of wipe samples on the hands of 368
pre-school children from eleven schools in London, UK. Hand Pb residue (PbH) values
were modeled as linear and power functions of Pb dust (PbD) from this study to obtain
PbH = 16.1 + 0.0064PbD or PbH = 0.393PbD° 533. 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. Measurements of the size distribution produced geometric
mean diameters of 4.5 (.un and 1.5 |_im depending on the analysis method. The hand wipe
samples were effectively limited to particles smaller than 10 |_im. even when the soil itself
exhibited a much larger particle size distribution. However, these studies focused on
playground lead exposure. Que Hee et al. (1985) performed repeated wipe sampling of
hands after rubbing in a reference dust sample and in reporting results for a "small adult
hand" observed that loose particles up to 246 |_im in diameter adhered, which is similar to
the upper limit of 200-300 |_im observed by Yamamoto et al. (2006). If a similar
relationship holds for household dust, then exposure to Pb in dust by hand-to-mouth
activity may be influenced by the dust particle size distribution and the relationship of Pb
content to particle size fraction.
4.1.3.3 Dietary Lead 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 4-2, 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. The source of the
Pb in each case is unclear. 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 and speciated for Pb isotopes. 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 y old children, and 20% for
25-30 y old women. Imported candy contributed 10% of Pb dietary intake for 2-6 y old
children and 9% for 25-30 y old women. Isotopic data from Manton et al. (2005)
suggested that, with the exception of children age 0-12 mos, house dust is a large
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contributor to dietary Pb. The pattern of certain Pb-isotope ratios observed in the diet of
children 0-12 mos are suggested to derive from Ca salts in limestone that may have been
used in dietary supplements in baby formula. The contribution of ambient Pb aerosols to
dietary Pb samples was not statistically significant for this urban exposure study.
Noodles	—
BF(carrot)	—
BF(oatmeal)	—
Cheese	—
Taco	—
Cereal	—
Egg'n'chs	—
Peas	—
Milk(2%)	-
BF(gmbean)	—
Cabbage	—
BF(lamb)	—
Beans	—
Rice	—
BF(applej)	—
Coffee	—
Strawberries	—
Squash	—
Potato	—
Chicken	—
Biscuits	—
BF(cobbler)	-
BF(beef)	-



	H
H 1 1
H--I 1 H


1 1 1



Oh







l-IO-H



w



fflo



Eh



u



CD-h



m



i-OH






o
h —-1 1 |H


1
1
1
1
1
1
1
Ch


0.00	0.02	0.04	0.06	0.08	0.10	0.12	0.14
Concentration (mg/kg)
Source Data: (FDA. 2008)
Note: from the 2008 FDA Total Diet Study. "BF" denotes baby food.
Figure 4-2 Market basket survey results for Pb concentration in foods.
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Drinking Water
Differences in sources and transport of drinking water may cause variation in water Pb
levels. 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). This level (i.e., 8.5
ng/L) is significantly less than the Maximum Contaminant Level for Pb of 15 (ig/L. Pb in
drinking water supplies can derive from atmospheric deposition onto surface waters, as
described in Section 3.3.1.2, 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 engineered 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. Edwards and Dudi (2004) observed a red-brown
particle-bound Pb in Washington, DC 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.
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(V04)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, V043, 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.
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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
Herrera (2010) observed Pb oxide corrosion scales occurring after using acidic alum as a
disinfection agent. High Pb concentrations in Washington, DC drinking water 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 overtime. When adding fluorosilicic acid and
ammonia, the Pb concentration spiked to 900 ppb and increased further over time.
Similarly, Lasheen et al. (2008) observed leaching from pipes in Egypt. In this study, the
authors tested polyvinyl chloride (PVC), polypropylene (PP), and galvanized iron pipes
and observed leaching from both the PVC and PP pipes when exposed to an acid of pH =
6, with PVC having greatest amount of leaching. Exposure to basic solutions actually
resulted in reduction of Pb concentration in the drinking water.
Miranda et al. (2007a) modeled blood Pb levels among children living in Wayne County,
NC as a function of household age, use of chloramines and other covariates. Blood Pb
levels were significantly associated with the year the home was built (p <0.001), use of
chloramines (p <0.001), and the interaction between these two variables (p <0.001).
When year in which the home was built was broken into categories for the independent
variables and interaction terms, Miranda et al. (2007a) found that significance increased
with the age of the home. However, 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 lead 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 chloramines disinfection is attributed to
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the formation of the Pb(II) mineral hydroxypyromorphite (Pb5(P04)30H) 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)02
reduction by an intermediate species from decomposition of monochloramine; and
increasing redox potential which decreases the thermodynamic driving force for
reduction. They suggest that the contact time of monochloramine with Pb02 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 j^ig/L (for relatively high
loadings of Pb02 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 plants 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
Cichorium intybus [chicory], Cynodon dactylon [Bermuda grass], Amaranthus blitoides
[matweed or mat amaranth], and Silybum marianum [milk thistle]) 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, measured as the percent of Pb mass in the shoots compared to the Pb mass within
the entire plant, 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) compiled bioaccumulation data for plant groupings from
various references; these data are reproduced in Table 4-5. Based on this review, grasses
had the highest uptake, followed by leafy vegetables and root crops grown in sandy soils;
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these references also suggested high transfer from roots to shoots among root crops, with
shoots having roughly four times higher Pb bioaccumulation than roots. This is consistent
with the 2006 Pb AQCD (2006b). which reported that Pb deposition onto leafy
vegetables accounted for most Pb.
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 m from the
smelter had concentrations that were 10-20 times lower, which is consistent with findings
described in Section 3.3 that deposition of Pb containing material drops off with distance
from a source. Pb contamination of crops can occur through Pb emissions during aerial
application of fertilizers and pesticides. In 2009, the U.S. Federal Aviation
Administration (FAA) recorded 960,000 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 42% of these flight-hours involved piston
engine aircraft utilizing leaded fuel ("FAA. 2009).
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 |im and 50-100 |im (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 change the acidity of the soil and hence the
bioaccessibility of the Pb within 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).
There is some evidence that Pb in 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% prior to its application. Chen et al. (2008a) 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. (2008b) 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 P and Zn indicators of fertilizer. In four
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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) 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., some imported
crops are produced in countries that still use Pb antiknock agents in on-road gasoline. 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. Likewise, it is possible that resuspended Pb that originated
from legacy mobile and industrial sources could deposit on crops.
The extent to which soil Pb contributes Pb to agricultural crops varies with soil
characteristics. For example, Jin et al. (2005) tested 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.
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Table 4-5 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
Plant Compartment
Soil
n
GM
GSD
AM
SD
Min
Max
All


210
2.0%
14
63%
290%
0.015%
2,500%
Cereals
Grain
All
9
1.0%
3.6
1.8%
1.6%
0.19%
4.8%
Straw
All
4
2.3%
3.5
3.8%
4.0%
0.51%
9.6%
Maize
Grain
All
9
0.12%
2.3
0.17%
0.14%
0.052%
0.38%
Straw
All
3
0.28%
6.6
0.85%
1.3%
0.060%
2.3%
Rice
Grain
All
2


2.2%
1.4%
1.2%
3.2%


All
31
8.0%
13
210%
610%
0.32%
2,500%
Leafy Vegetables

Sand
4
7.3%
1.5
7.8%
3.3%
4.9%
11%

Loam
3
82%
1.0
82%
3.5%
79%
86%


Clay
7
2.8%
4.1
5.1%
4.8%
0.41%
12%
Non-Leafy Vegetables
Fruits
All
5
1.5%
26
78%
170%
0.15%
390%
Shoots
All
2


0.88%
0.42%
0.58%
1.17%


All
17
0.53%
12
34%
120%
0.046%
490%

Pods
Sand
3
0.27%
3.2
0.42%
0.34%
0.065%
0.89%
Legumes
Loam
5
0.14%
4.4
0.42%
0.34%
0.065%
0.89%


Clay
4
0.080%
1.0
0.33%
0.47%
0.046%
1.0%

Shoots
All
1


0.080%





All
27
1.5%
16
41%
98%
0.024%
330%
Root Crops
Roots
Sand
5
6.4%
1.6
7.0%
3.4%
4.2%
12%

Loam
5
2.3%
4.7
0.50%
0.68%
0.024%
1.7%

Shoots
All
12
6.3%
15
250%
570%
0.30%
16%


All
30
0.15%
7.4
9.1%
48%
0.015%
260%
Tubers
Tubers
Sand
5
0.64%
3.5
1.2%
1.6%
0.16%
3.9%


Loam
17
0.052%
2.4
0.073%
0.062%
0.015%
0.23%
Fruits
Fruits
All
5
0.77%
2.6
1.0%
0.60%
0.15%
1.7%
Leaves
All
1


25%



Grasses

All
17
31%
1.8
36%
22%
11%
100%
Natural Pastures

All
34
92%
4.8
23%
29%
0.22%
100%
Leguminous Fodder All 1 1.6%


All
20
0.43%
4.7
1.1%
1.4%
0.052%
4.8%
All Cereals

Sand
5
0.61%
5.3
1.3%
1.3%
0.052%
3.2%

Loam
8
0.17%
3.9
0.53%
1.1%
0.059%
3.2%


Clay
6
0.90%
4.0
1.8%
1.8%
0.22%
4.8%
Pastures/Grasses

All
51
14%
4.2
27%
27%
0.22%
100%


All
24
2.5%
12
130%
420%
0.060%
1,600%
Fodder

Sand
4
4.5%
2.3
5.6%
4.0%
1.6%
11%


Clay
4
0.82%
5.7
2.7%
4.6%
0.16%
9.6%
Source: Reprinted with permission of Elsevier Publishers, Vandenhove et al. 12009).
1	Findings from Pb uptake studies have implications for urban gardening if urban soils may
2	be contaminated with Pb, as described in Section 4.1.3.2. For instance, Clark et al. (2006)
3	tested the soil in 103 urban gardens in two Boston neighborhoods. They found that Pb-
4	based paint contributed 40-80% of Pb in the urban garden soil samples, with the rest
5	coming from historical gasoline emissions. Furthermore, Clark et al. (2006) estimated
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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.
Game
Game meat consumption also may pose 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.. 2009a)l. Jankovska et al. (2010) studied Pb accumulation in sheep
infected with tapeworm and found that sheep harboring the parasites contained
significantly less Pb compared with their uninfected counterparts (p < 0.05).
Potential Pb exposure through consumption of animals exposed to or killed with Pb shot
has also been well documented (Hunt et al.. 2009; Tsuii et al.. 2009; Tsuii et al.. 2008;
Hunt et al.. 2006). For example, Martinez-Hare 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
Accumulation 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 accumulation in fish. Several studies have documented the
potential for human 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 Pb concentrations in blood
varied with respect to species of fish, but Pb concentrations were higher in fish in areas
close to mining activities. Similarly, Besser et al. (2008) observed higher levels of blood
Pb in fish close to mining activities in southeastern Missouri. In a related study of fish
species in the same region of Missouri, blood Pb levels in fish were found to be
significantly higher in sites within 10 km downstream of active Pb-Zn mines (p <0.01)
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compared with fish located further from the mines (Schmitt et al.. 2007a). and elevated
blood Pb levels in fish 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.
Breast Milk
Studies breastfeeding women suggest that infants may be exposed to Pb in breast milk.
Ettinger et al. (2004a') observed among Mexico City women studied in 1994-1995 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 j^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 j^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 Pb contaminated city in Australia,
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 lead.
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).
4.1.3.4 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.
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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, reported to have 75% of particulate mass
below 2.5 |_im diameter, was not reported; year of the study was also not reported.
Operations involving PM 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 4.5.1. Nwajei and Iwegbue (2007)
measured Pb contamination in sawdust; such contamination has been reported to occur
when trees are grown in soil contaminated with Pb (Andrews et al.. 1989). Sawdust
samples from fifteen locations in Nigerian sawmills were reported to have Pb
concentrations ranging from 2.0 to 250 mg/kg.
4.1.3.5 Exposure to Lead from Consumer Products
Pb is present in varying amounts in several consumer products including alternative
medicines, candies, cosmetics, pottery, tobacco, toys, and vitamins (Table 4-6). 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).
Table 4-6 Pb content in various consumer products.
Product
Category
Product
Location of
Purchase
Pb Content (units)
Reference
Alternative and
Traditional
Medicines
C/'ssus quadrangularis, Caulophyllum thalictroides, Turnera diffusa, Centella
asiatica, Hoodia gordonii, Sutherlandia frutescens, Curcuma longa,
fucoxanthin, Euterpe oleracea
(dietary supplements claimed to be from Hoodia gordonii}
U.S.
(Mississippi)8
Not detected (N.D.)
-4.21 mg/kg
Avula et al.
(2010)
Maiva syivestris
Turkey
1.1-2.0 mg/kg
Higsomnez et al.
(2009)
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Product

Location of


Category
Product
Purchase
Pb Content (units)
Reference

Yugmijihwang-tang, Bojungigki-tang, Sibjeondaebo-tang, Kuibi-tang,
Korea
7.9><10"6 to 2.5><10"5 mg/kg body
Kim et al. (2009a)

Ojeogsan

weight/day


Lemongrass, licorice, holy basil, cloves, ginger
India
Average:
Lemongrass & 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
Naithani and
Kakkar (2006)

B-Success 28, Operation Sweep, Aloe Vera Plus Bitter Aloes,
Nigeria
925-27,000 pg
Obi et al. (2006)

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
Sharma et al.
(2009)

Berberis (6. aristata, B. chitria, B. asiatica, B. lyceum), Daruharidra
India
Berberis:
Roots: 3.1-24.7 mg/kg
Stems: 8.0-23.8 mg/kg
Daruharidra:
16.9-49.8 mg/kg
Srivastava et al.
(2006)

Greta powder
U.S.
(California)
770,000 ppm
CDC (2002)

Tamarind Candy
U.S.
Product: 0.15-3.61 mg/kg
Lynch et al.


(Oklahoma)
Stems: 0.36-2.5 mg/kg
Wrappers: 459-27,125 mg/kg
(2000)
Candy
Tamarind Candy
U.S.
(California)
Product: 0.2-0.3 mg/kg
Stems: 400 mg/kg
Wrappers: 16,000-
21,000 mg/kg
CDC (2002)

Lipsticks
U.S.
Average: 1.07 mg/kg
Hepp et al.
Cosmetics



(2009)
Eye Shadows
Nigeria
N.D.-55 mg/kg
Omolaoye et al.
(2010a)
Pottery
Foods prepared in Pb-glazed pottery
Mexico
N.D.-3,100 mg/kg
Villalobos et al.
(2009)

Smokeless Tobacco
U.K.
0.15-1.56 mg/kg
McNeill et al.
Tobacco



(2006)
Cigarette Tobacco (210Pb concentrations)
Pakistan
Activity conc.: 7-20 Bq/kg
Tahirand
Alaamer (2008)

Red and yellow painted toy vehicles and tracks
Brazil
500-6,000 mg/kg
Godoi et al.
(2009)

535 PVC and non-PVC toys from day care centers
U.S. (Nevada)
PVC: avg. 325 mg/kg
Non-PVC: avg. 89 mg/kg
Yellow: 216 mg/kg
Greenway and
Gerstenberger
(2010)
Toys


Non-yellow: 94 mg/kg

Soft plastic toys
India
Average (by city): 21-280 mg/kg
Kumar and
Pastore (2007)

Toy necklace
U.S.
388,000 mg/kg
Meyer et al.
(2008)

Soft plastic toys
Nigeria
2.5-1,445 mg/kg
Omolaoye et al.
(2010b)
Average:	Mindaketal.
Young children: 2.9 pg/day (2008)
Older children: 1.8 pg/day
Pregnant and lactating women:
4.9 pg/day
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.
Vitamins for young children, older children, and pregnant or lactating	U.S.
women
Vitamins
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4.2
Kinetics
This section summarizes the empirical bases for our understanding of 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 4.3 (and Section 4.2.2.1), Pb biokinetics is described from the context of model
predictions.
4.2.1 Absorption
The focus of the following sections within absorption is on inhalation and ingestion
because these are the major exposure routes of Pb in humans. The 2006 Pb AQCD 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 (Pan et al.. 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 lead 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.
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. 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
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be extracted into a physiological or physiological-like solution (e.g., gastric juice or
solution similar to gastric juice).
4.2.1.1 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 Rov. 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 (tl/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.
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 (.un 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
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was calculated to be 16 ng for the PM2 5_io 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 ultrafine-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 |im 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 bioavailable 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 bioavailable 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, UK, 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 PbCr04 comprised the largest fractions, 45% and
21% respectively, of Pb in the <38 |_im fraction. These forms tend to be less
bioaccessibility 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). Mechanistic models
provide one means for integrating these variable 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 (Section 4.3 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 |am 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). Empirical estimates of blood Pb - air Pb slopes for various populations,
derived from epidemiological studies, are summarized in Section 4.5.1.
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Organic Lead
Alkyl Pb compounds can exist in ambient air as vapors. Inhaled tetraalkyl Pb vapor is
nearly completely absorbed following deposition in the respiratory tract. As reported in
the 2006 Pb AQCD, 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 organic Pb is relevant to some aviation fuel exposures
(e.g., persons exposed to leaded gasoline used piston-engine aircraft). Mahaffey (1977)
estimated that 40% of inhaled Pb in urban air (largely attributed to combustion of
gasoline containing tetraethyllead) is bioavailable to adults. Chamberlain et al. (1975)
suggested that 35% of inhaled combustion products of tetraethyl 203Pb fuel 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
engine exhaust from fuel containing tetraethyllead at a concentration of 0.001 (.ig/m3 for a
period of months could produce a 1 (ig/dL increment in blood Pb.
4.2.1.2 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 and
iron status, pregnancy) and physicochemical characteristics of the Pb-bearing material
ingested (e.g., particle size, mineralogy, solubility). Pb absorption in humans may be a
capacity-limited process, in which case the percentage of ingested Pb that is absorbed
may decrease with increasing rate of Pb intake. Numerous observations of nonlinear
relationships between blood Pb concentration and Pb intake in humans provide support
for the likely existence of a saturable absorption mechanism or some other capacity-
limited process in the distribution of Pb in humans (Sherlock and Ouinn. 1986; Sherlock
et al.. 1984; Pocock et al.. 1983; Sherlock et al.. 1982). While evidence for capacity-
limited processes at the level of the intestinal epithelium is compelling, the dose at which
absorption becomes appreciably limited in humans is not known.
In adults, estimates of absorption of ingested water-soluble Pb compounds (e.g., Pb
chloride, Pb nitrate, Pb-acetate) range from 3 to 10% in fed subjects (Maddaloni et al..
1998; Watson et al.. 1986; James et al.. 1985; Heard and Chamberlain. 1982; Rabinowitz
et al.. 1980). The absence of food in the GI tract increases absorption of water-soluble Pb
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in adults. Reported estimates of soluble Pb absorption range from 26 to 70% in fasted
adults (Maddaloni et al.. 1998; James et al.. 1985; Blake et al.. 1983; Heard and
Chamberlain. 1982; Rabinowitz et al.. 1980). Reported fed:fasted ratios for soluble Pb
absorption in adults range from 0.04 to 0.2 (James et al.. 1985; Blake et al.. 1983; Heard
and Chamberlain. 1982; Rabinowitz et al.. 1980).
Limited evidence demonstrates 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 (Ziealer et al.. 1978; Alexander et al.. 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 et al.. 1978; Pounds et al.. 1978; Forbes and
Reina. 1972). The mechanisms for an apparent age difference in GI absorption of Pb have
not been completely elucidated and may include both physiological and dietary factors
(Mushak. 1991). Eating breakfast was shown to be significant predictor of blood Pb
concentrations in 1,344 children 3-5 year s old (Liu et al. 2011b). Blood Pb
concentrations were lower in children who regularly ate breakfast compared to children
who did not eat breakfast, and the difference persisted after controlling for nutritional
variables (blood iron, calcium, copper, magnesium, zinc). 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 been reported for
adults (Maddaloni et al.. 1998; James et al.. 1985).
Nutritional deficiencies have also been linked to Pb absorption in the GI tract,
particularly in children. Children who are iron-deficient have higher blood Pb
concentrations than similarly exposed iron-replete children, suggesting that iron
deficiency may result in higher Pb absorption or, possibly, other changes in Pb
biokinetics that contribute to altered blood Pb concentrations (Schell et al.. 2004; Marcus
and Schwartz. 1987; Mahaffev and Annest. 1986). Studies conducted in animal models
have provided direct evidence for interactions between iron deficiency and increased Pb
absorption, perhaps by enhancing binding of Pb to iron-binding proteins in the intestine
(Bannon et al.. 2003; Morrison and Ouarterman. 1987; Barton et al.. 1978b). An analysis
of data from a sample 448 woman (age 20-55 years) did not find a significant association
between iron body stores (indicated from serum ferritin concentration) and blood Pb
concentrations, although depleted irons stores (serum ferritin of <12 (ig/L) was associated
with higher blood concentrations of cadmium, cobalt and manganese higher (Meltzer et
al.. 2010).The effects of iron nutritional status on blood Pb include changes in blood Pb
concentrations in association with genetic variation in genes involved in iron metabolism.
For example, genetic variants in the hemochromatosis (HFE) and transferrin genes are
associated with higher blood Pb concentrations in children (Hopkins et al. 2008). In
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contrast, HFE gene variants are associated with lower bone and blood Pb levels in elderly
men (Wright et al.. 2004V
Several studies have suggested that dietary calcium 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
calcium carbonate (0.2 g calcium carbonate) than when the Pb was ingested without
additional calcium (Blake and Mann. 1983; Heard and Chamberlain. 1982). A similar
effect of calcium occurs in rats (Barton et al.. 1978a). Similarly, an inverse relationship
was observed between dietary calcium intake and blood Pb concentration in children,
suggesting that children who are calcium-deficient may absorb more Pb than calcium-
replete children (Elias et al.. 2007; Schell et al.. 2004; Mahaffev et al.. 1986; Ziegleret
al.. 1978). These observations suggest that calcium and Pb share and may compete for
common binding and transport mechanisms in the small intestine which are regulated in
response to dietary calcium and calcium body stores (Fullmer and Rosen. 1990; Bronner
et al.. 1986). However, animal studies have also shown that multiple aspects of Pb
toxicokinetics are affected by calcium nutritional status. For example, feeding rats a
calcium 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 calcium nutrition and blood Pb that have been
observed in human populations may not be solely attributable to effects of calcium
nutrition on Pb absorption. Other potential mechanisms by which calcium 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 (>100 (j,m) tends to be
lower than smaller particles (Healv et al.. 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
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humans (Juhasz et al. 2009; U.S. EPA. 2007b; Casteel et al.. 2006; Casteel et al.. 1997;
Weis and Lavcllc. 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/AFpb.acetate)-
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 49% (± 29[SD]), median is 51%, and 5th to 95th percentile range is 12 to -89%.
RBA estimates for soils collected from 8 firing ranges were approximately 100%
(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 (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 (14-40%, n = 3), as did
a sample from a mine tailings pile (RBA = 6%), 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 4-3 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 that 100% (Smith
et al.. 2009a. 2008; Freeman et al.. 1996; Freeman et al.. 1994; Freeman et al.. 1992).
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a;
Group
Source: Casteel et al. (20061.
Note: based on results from juvenile swine assays.
Figure 4-3 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 |im) for 1 hour. The
assay was tested for predicting in vivo RBA of 18 soil-like test materials that were
assayed in a juvenile swine assay (Casteel et al.. 2006). A regression model relating
IVBA and RBA was derived based on these data (Equation 4-1):
RBA = (0.878 X IVBA) - 0.028
Equation 4-1
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
assessment (U.S. EPA. 2007e). A review of soil Pb RBA estimates made using the IVBA
assay described above and Equation 4-1 identified 270 estimates of Pb RBA in soils
obtained from 11 hazardous waste sites. The mean for the site-wide RBA estimates (n =
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11 sites) was 57% (SD 15), median was 63%, and 5th to 95th percentile range was 34 to
71%.
Equation 4-1 cannot be reliably extrapolated to other in vitro assays that have been
developed for estimating Pb bioaccessibility without validation against in vivo RBA
measurements made on the same test materials. Comparisons of outcomes among
different in vitro assays applied to the same soil test materials have found considerable
variability in IVBA estimates (Juhasz etal.. 2011; Smith et al.. 2011; Saikat et al.. 2007;
Van de Wiele et al.. 2007). This variability has been attributed to differences in assay
conditions, including pH, liquid:soil ratios, inclusion or absence of food material, and
differences in methods used to separate dissolved and particle-bound Pb
(e.g., centrifugation versus filtration). Smith et al. (2011) found that algorithms for
predicting RBA based on two different IVBA assays did not yield similar predictions of
RBA when applied to the same material. Given the dependence of IVBA outcomes on
assay conditions, in vitro assays used to predict in vivo RBA should be evaluated against
in vivo RBA estimates to quantitatively assess uncertainty in RBA predictions (U.S.
EPA. 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 72% (95% CI: 44, 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 75%
(Drexler and Brattin. 2007). which corresponds to a predicted RBA of 63%, based on
Equation 4-1. A review of indoor Pb RBA estimates made using the IVBA assay and
Equation 4-1 identified 100 estimates of Pb RBA in dusts obtained from two hazardous
waste sites. Mean Pb RBAs for the Herculaneum site were 47% (SD 7, 10 samples) for
indoor dust and 69% (SD 3, 12 samples) for soil. At the Omaha site, mean Pb RBAs were
73% (SD 10, 90 samples) for indoor dust and 70% (SD 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 64.8% (SD 8.2, age: 52.5 to 77.2 months).
The above results, and the IVBA assays used in studies of interior dust, have not been
evaluated against in vivo RBA estimates for dust samples. Although, expectations are
that a validated IVBA methodology for soil would perform well for predicting RBA of
interior dust, this validation has not actually been experimentally confirmed. Factors that
may affect in vitro predictions of RBA of interior dust Pb could include particle size
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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 etal.. 2011; Lu et al.. 2011; Smith et al.. 2011;
Roussel et al.. 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 4.1.3.2). 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.
4.2.2 Distribution
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; Qldereid 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.
4.2.2.1 Blood
Blood comprises -1% of total Pb body burden. Pb in blood is found primarily (>99%) in
the RBCs (Smith et al.. 2002; Manton et al.. 2001; Bergdahl et al.. 1999; Bergdahl et al..
1998; Hernandez-Avila et al.. 1998; Bergdahl et al.. 1997a; Schutz et al. 1996). 8-
aminolevulinic acid dehydratase (ALAD) is the primary binding ligand for Pb in
erythrocytes (Bergdahl et al. 1998; Xie et al.. 1998; Bergdahl et al.. 1997a; Sakai et al..
1982). Two other Pb-binding proteins have been identified in the RBC, a 45 kDa protein
(Kmax 700 (ig/dL; Kd 5.5 (ig/L) and a smaller protein(s) having a molecular weight
<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
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for Pb (Beredahl 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 (Beredahl et al.. 1998;
Beredahl etal.. 1996; Sakai et al.. 1982). Pb binding to ALAD is saturable; the binding
capacity was estimated to be -850 (ig/dL RBCs (or -40 (ig/dL whole blood) and the
apparent dissociation constant has been estimated to be -1.5 |_ig/L (Beredahl et al.. 1998).
Binding to fetal hemoglobin within the RBC has a higher affinity for Pb than adult
hemoglobin. This suggests that erythrocytes of neonates should be able to store more Pb
than infants (Simon et al. 2007). 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 (Kane et al.. 2009; Jin et al.. 2008; Barbosa et al.. 2006b; Smith et al.. 2002;
Manton et al.. 2001; Beredahl et al.. 1999; Beredahl et al. 1998; Beredahl et al.. 1997b;
DeSilva. 1981; Rentschler et al. In Press). An example of this is shown in Figure 4-4.
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 et al. In Press).
Typically, at blood Pb concentrations <100 (ig/dL, only a small fraction (<1%) of blood
Pb is found in plasma (Marcus. 1985; Manton and Cook. 1984; DeSilva. 1981). However,
as previously noted, plasma Pb may be the more biologically labile and toxicologically
active 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; Qng 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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T3
O)
=L
.Q
CL
CO
E
CO
ro
CL
0	20	40	60	80 100
Blood Pb (Mg/dL)
Source: Adapted with permission of Elsevier Publishing and the Finland Institute of Occupational Health, Bergdahl et al. (1999;
1997b).
Figure 4-4 Plot of blood and plasma Pb concentrations measured in adults
and children.
1	As shown in Figure 4-4, the limited binding capacity of Pb binding proteins in RBCs
2	produces a curvilinear relationship between blood and plasma Pb concentration. The
3	limited binding capacity of RBC binding proteins also confers, or at least contributes, to a
4	curvilinear relationship between Pb intake and blood Pb concentration. A curvilinear
5	relationship between Pb intake and blood Pb concentration has been observed in children
6	(Sherlock and Quinn. 1986; Lacev et al.. 1985; Ryu et al.. 1983). As shown in Figure 4-5,
7	the relationship becomes pseudo-linear at relatively low daily Pb intakes
8	(i.e., <10 (ig/day/kg) and at blood Pb concentrations <25 (ig/dL.
2.0
1.5
1.0
0.5
0.0
oAdults •Children
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_i
T3
CD
=L
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100
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300
400
Pb Intake (|jg/day)
Source: Adapted with permission of Taylor & Francis Publishing, Sherlock and Quinn (19861.
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).
Figure 4-5 Relationship between Pb intake and blood Pb concentration in
infants (n = 105, age 13 weeks, formula-fed).
Figure 4-6 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 4.3 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'Flahertv. 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 this 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 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 -
1) 0.40 -
3
n
0.30 -
ra
ra 0.20 H
Q.
0.10 -
0.00
~i	1	1	r
10 20 30 40
Blood Pb (|ig/dL)
50
	Blood
Pasma
D) 40
0.6 3
0.4 £
era
0	100	200	300
Intake (tig/day)
Simulation based on ICRP Pb biokinetics model CLeaaett. 19931.
400
Figure 4-6 Simulation of quasi-steady state blood and plasma Pb
concentrations in a child (age 4 years) associated with varying Pb
ingestion rates.
Studies conducted in swine provide additional evidence in support of RBC binding
kinetics influencing distribution of Pb to tissues. In these studies, the relationship
between the ingested dose of Pb and tissue Pb concentrations (e.g., liver, kidney, bone)
was linear, whereas, the relationship between dose and blood Pb was curvilinear with the
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slope decreasing as the dose increased (Casteel et al.. 2006). Saturable binding of Pb to
RBC proteins also contributes to a curvilinear relationship between urinary Pb excretion
and plasma Pb concentration (Section 4.2.3) (Besser et al.. 2008; Bergdahl etal.. 1997b).
4.2.2.2 Bone
The dominant compartment for Pb in the body is in bone. In human adults, 94% of the
total body burden of Pb is found in the bones, whereas bone Pb accounts for 73% of the
body burden in 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: Leaaett. 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" 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 et al. 1991). Pb forms highly stable complexes with phosphate and can
replace calcium in the calcium-phosphate salt, hydroxyapatite, which comprises the
primary crystalline matrix of bone (Meirer et al.. 2011: Bros et al.. 1986: Mivake. 1986:
Verbeeck et al.. 1981). Several intracellular kinetic pools of Pb have been described in
isolated cultures of osteoblasts and osteoclasts which appear to reflect physiological
compartmentalization within the cell, including membranes, mitochondria, soluble
intracellular binding proteins, mineralized Pb (i.e., hydroxyapatite) and inclusion bodies
(Long et al.. 1990: Pounds and Rosen. 1986: Rosen. 1983). Approximately 70-80% of Pb
taken up into isolated primary cultures of osteoblasts or osteocytes is associated with
mitochondria and mineralized Pb (Pounds et al.. 1991).
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
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.. 2009a: Barry and Connolly.
1981: Barry. 1975: Gross et al.. 1975: Schroeder and Tipton. 1968). The cortical and
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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'Flahertv. 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.
A key factor affecting Pb uptake into bone is the fraction of bone surface in trabecular
and cortical bone adjacent to active bone marrow. Of the total bone surface against red
marrow, 76% is trabecular and 24% is cortical endosteal (Salmon et al.. 1999). The
fraction of total bone marrow that is red and active decreases from 100% at birth to about
32% in adulthood (C'ristv. 1981). However, bone marrow has much lower Pb
concentrations than bone matrix (SkerfVing et al.. 1983).
4.2.2.3 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. 1975; Gross et al. 1975). Presumably, the Pb
in these soft tissues (i.e., kidney, liver, and brain) exists predominantly bound to protein.
High affinity cytosolic Pb-binding proteins have been identified in rat kidney and brain
(DuVal and Fowler. 1989; Fowler. 1989). The Pb-binding proteins in rat are cleavage
products of a2\i globulin, a member of the protein superfamily known as retinol-binding
proteins that are generally observed only in male rats (Fowler and DuVal. 1991). Other
high-affinity Pb-binding proteins (Kd -14 nM) have been isolated in human kidney, two
of which have been identified as a 5 kDa peptide, thymosin 4 and a 9 kDa peptide, acyl-
CoA binding protein (Smith et al.. 1998). Pb also binds to metallothionein, but does not
appear to be a significant inducer of the protein in comparison with the inducers Cd and
Zn (Waalkcs and Klaassen. 1985; Eaton et al.. 1980).
The liver and kidneys rapidly accumulate systemic Pb (ti/2=0.21 and 0.41 hours,
respectively), which amounts to 10-15% and 15-20% of intravenously injected Pb,
respectively (Leggett. 1993). A linear relationship in dose-tissue Pb concentrations for
kidney and liver has been demonstrated in swine, dogs, and rats (Smith et al.. 2008;
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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.
4.2.2.4 Fetus
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 (Amaral et al..
2010; kordas et al. 2009; Patel and Prabhu. 2009; Carbone etal.. 1998; Gover. 1990;
Graziano et al.. 1990). In a study of mothers having blood Pb levels of <14 (ig/dL, the
ratio of cord blood Pb to maternal blood Pb decreased with decreasing maternal blood Pb
(r=0.82) (Carbone etal.. 1998). In addition, the similarity of isotopic ratios in maternal
blood and in blood and urine of newly-born infants provide further evidence for placental
transfer of Pb to the fetus (Gulson et al.. 1999).
Transplacental transfer of Pb may be facilitated by an increase in the plasma/blood Pb
concentration ratio during pregnancy (Montenegro et al.. 2008; Lamadrid-Figueroa et al..
2006). Maternal-to-fetal transfer of Pb appears to be related partly to the mobilization of
Pb from the maternal skeleton. Evidence for transfer of maternal bone Pb to the fetus has
been provided by stable Pb isotope studies in cynomolgus monkeys exposed during
pregnancy. Approximately 7-39% of the maternal Pb burden transferred to the fetus was
derived from the maternal skeleton, with the remainder derived from contemporaneous
exposure (O'Flahertv. 1998; Franklin et al.. 1997).
4.2.2.5 Organic Lead
Information on the distribution of Pb in humans following exposures to organic Pb is
extremely limited. However, as reported in the 2006 Pb AQCD, 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
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tetraalkyl/trialkyl Pb. The subsequent rise in blood radioactivity, however, probably
represents water-soluble inorganic Pb and trialkyl and dialkyl Pb compounds that were
formed from the metabolic conversion of the volatile parent compounds (Heard et al..
1979).
Alkyl Pb compounds undergo oxidative dealkylation catalyzed by cytochrome P450 in
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).
4.2.3 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. 1961a. 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; Hursh et al.. 1969; Hursh and Suomela.
1968).
Approximately 30% of intravenously injected Pb in humans (40-50% in beagles and
baboons) is excreted via urine and feces during the first 20 days following administration
(Leggett. 1993). The kinetics of urinary excretion following a single dose of Pb is similar
to that of blood (Chamberlain et al.. 1978). likely due to the fact that Pb in urine derives
largely from Pb in plasma. Evidence for this is the observation that urinary Pb excretion
is strongly correlated with the rate of glomerular filtration of Pb (Araki et al.. 1986) and
plasma Pb concentration (Bergdahl et al. 1997b; Rentschler et al. In Press)
(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 4.2.2.1 and demonstrated in Figure 4-6). 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 4.3.7).
Estimates of urinary filtration of Pb from serum (or plasma) range from 13-22 L/day,
with a mean of 18 L/day (Araki et al. 1986; Manton and Cook. 1984; Manton and
Malloy. 1983; Chamberlain et al.. 1978). which corresponds to half-time for transfer of
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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
re-absorption of Pb in the glomerular filtrate (Araki et al.. 1986); although, net tubular
secretion of Pb has been demonstrated in animals (Victory et al.. 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 et al.. 1990; Berger et
al.. 1990; Roster et al.. 1989; Manton and Mallov. 1983; Ryu et al.. 1983; Chamberlain et
al.. 1978; Rabinowitz et al.. 1973). consistent with a plasma Pb to blood Pb concentration
ratio of -0.005-0.01 L/day (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 fecakurinary 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; Hursh et al.. 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 X\i2 = 10 days), and accounts for 70% of the total plasma clearance (O'Flahertv.
1995).
Organic Lead
Pb absorbed after inhalation of tetraethyl and tetramethyl Pb is excreted in exhaled air,
urine, and feces (Heard et al.. 1979). Fecakurinary 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
Duvdii. 1995; Zhang et al.. 1994; Turlakiewicz and Chmielnicka. 1985).
4.3 Lead 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
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(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 5%
of an individual's total body Pb burden resides in blood. Furthermore, blood consists of
several subcompartments. More than 90% of Pb in whole blood is bound to red cell
proteins such as hemoglobin, with the balance in plasma. From a toxicological
perspective, the unbound fraction 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. 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
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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 we have for synthesizing our 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 our 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
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,
we have limited the discussion 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 lead 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 lead
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 lead risk assessment, it has been applied in Pb risk assessment (Abrahams et
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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 lead in the respiratory tract. These types of
simulations can only be approximated with the U.S. EPA IEUBK Model for lead 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 lead into a single absorption term representing the combined processes of
gastrointestinal and respiratory tract absorption of inhaled lead; 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).
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.
4.3.1 Bone Lead 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.
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Bone Pb measurements are typically expressed in units of Pb/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 (Thcppcang 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 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).
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) 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 jj.g/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 jj.g/g bone mineral,
respectively) compared to measurements made with single detectors (8-12 jj.g/g and
14-15 jj.g/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; Todd et al.. 2001; Aro et al.. 2000; Todd et al.. 2000)1. Todd
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et al. (2000) provided a detailed discussion of factors that influence the variability and
measurement uncertainty, including repositioning, sample measurement duration,
overlying tissue, operator expertise, detector resolution, and changes to measurement
process over time. Some of these aspects were also discussed by Hu et al. (1995). From
their cadaver and in vivo measurements, Todd et al. (2000) concluded that the uncertainty
in an individual measurement was an underestimate of the standard deviation of replicate
measurements, suggesting a methodological deficiency probably shared by most current
109Cd-based K-shell XRF Pb measurement systems. In examining the reproducibility of
the bone Pb measurements over a 4lA month period, Todd et al. found the average
difference between the XRF results from short term and longer term measurements was
1.2 |_ig/g. indicating only a small amount of variability in the XRF results over a sustained
period of time.
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 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 demonstrated by Bellinger
et al. (1994a). 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).
4.3.2 Blood Lead 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
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and approximately 0.1 (ig/dL for flameless AAS measurements (Flesial 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.
1996V While the technique has been applied to assessing Pb exposures in adults, ICP-MS
has not received widespread use in epidemiologic studies.
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 document reports that many studies have shown that, with similar
exposures to Pb, individuals with the ALAD-2 allele have higher blood Pb levels than
those without (Kim et al.. 2004; Perez-Bravo et al.. 2004; Bergdahl et al.. 1997b; Smith et
al.. 1995a; Wetmur. 1994; Wetmur et al.. 1991a; Astrin et al. 1987). More recent meta
analyses provide further support for ALAD2 carriers having higher blood Pb levels than
ALAD 1-1 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 ALAD 1 compared to ALAD2 (Wetmur. 1994; Wetmur et al..
1991b). measurements of Pb binding affinity to ALAD1 and ALAD2 (i.e., Pb2+
displacement of Zn2+ binding to recombinant ALAD 1 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
ALAD 1 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 (Miyaki 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
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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).
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 8-ALA into brain
and renal tubular re-absorption of peptides. The PRPT2*2 polymorphism was associated
with increased blood Pb concentrations in a sample of 116 children of Mexican-
American/Hispanic (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) between individual blood
Pb concentrations measured after 6-12 months of age (Schnaas et al.. 2000; Dietrich et
al.. 1993b; McMichael et al.. 1988; Otto et al.. 1985a; 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 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 4-7. 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 4-7, 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 over time.
The rate of change can provide information about the magnitude of change in exposure,
but not necessarily about the time history of the change (Figure 4-7, 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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T3
~0)
T3
ra
CD
_l
T3
O
O
CQ
25
20
Child B
10-
Child A
**~~~~~~~
0 12 24 36 48 60 72 84
Age (months)
;o
~5>
T3
CO
cu
—I
T3
O
_o
CQ
40
Child C
Child D
0
0 12 24 36 48 60
Age (months)
72 84
Note: Child A and Child B have a relatively constant basal Pb intake (|jg/day/kg body weight) from birth; Child B experiences 1 -year
elevated intake beginning at age 24 months (upper panel). 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 |jg/dL, respectively), yet the exposure scenarios for these samples are very different.
As shown in the example of Child C and Child D, two samples can provide information about the magnitude of change in exposure,
but not necessarily the temporal history of the change (lower panel). Simulation based on ICRP Pb biokinetics model fLegqett.
19931.
Figure 4-7 Simulation of temporal relationships between Pb exposure and
blood Pb concentration in children.
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4.3.3
Urine Lead 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 4.2.3)
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 (Fukiii 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 et al.. 1990).
Urinary Pb concentration reflects, mainly, the concentration of Pb in the blood. As such,
urinary concentrations by reflect both recent and past exposures to Pb (see Section 4.3.5).
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 4.3.6 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 (jj.g/dg creatinine) and individual
Pb intakes (|_ig/da\ ) 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
4-8).
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10
-	- Urine
—	Body
8
-- 15 CD
6
10
4
(Q
2
0
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Age (year)
	Urine
n o)
25 30 35 40 45 50 55 60 65 70
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). 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 CLeaaett. 19931.
Figure 4-8 Simulation of relationship between urinary Pb excretion and body
burden in adults.
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4.3.4
Lead in Other Biomarkers
There was extensive discussion in the 2006 Pb AQCD 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.
4.3.4.1 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; Rabinow itz
et al.. 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.. 1995V 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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4.3.4.2
Hair
The 2006 Pb AQCD 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
(Wilhelm 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; Drasch et al.. 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.
4.3.4.3 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; Brodeuret al.. 1983; P'an. 1981);
however, more recent assessments have shown relatively weak or inconsistent
associations (2011; 2010; Costa de Almeida et al. 2009; Barbosa et al.. 2006c; Nriaguet
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. (2006c)
found a significant but relatively weak correlation (log [blood PB] versus log [saliva 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 salivary and blood Pb concentrations in children
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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.
4.3.4.4 Serum 5-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; Hem berg 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 (Opleret al.. 2008; Opler et al.. 2004).
4.3.5 Relationship between Lead in Blood and Lead in Bone
The kinetics of elimination of Pb from the body reflects the existence of fast and slow
pools of Pb in the body. The dominant phase of Pb kinetics in 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 have found that bone Pb
stores can contribute 40-70% to blood Pb (Smith et al.. 1996; Gulson et al.. 1995;
Manton. 1985). Bone Pb burdens in adults are slowly lost by diffusion (heteroionic
exchange) as well as by resorption (O'Flahertv. 1995). Half-times for the release of Pb in
bone are dependent on age and intensity of exposure. Bone compartments are much more
labile in infants and children than in adults as reflected by half-times for movement of Pb
from bone into the plasma (e.g., cortical ty2 = 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
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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.
The apparent slow clearance of Pb from the blood over months and years following the
cessation or reduction in exposures reflects elimination of Pb stores in bone. Longer fast-
phase elimination half-times (60-120 days) were reported for five adults with lead
poisoning following roughly a month to 12 years of exposure and relatively high blood
Pb concentrations of 70-110 (.ig/dL (Rentschler et al.. In Press). A slower phase becomes
evident with longer observation periods following a decrease in exposure. Nilsson et al.
(1991) reported in tri-exponential decay in the blood Pb concentrations of 14 individuals
having a median occupational exposure period of 26 years. Representing 22% of blood
Pb, the fast compartment had a clearance half time of 34 day. 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.
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 lead 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 Pb dose. Pb in adult bone can serve to maintain blood Pb levels long after
external exposure has ceased (Fleming etal.. 1997; Inskip et al.. 1996; Smith et al.. 1996;
Kehoe. 1987; O'Flahertv et al.. 1982). even for exposures that occurred during childhood
(McNeill et al.. 2000). The more widespread use of in vivo XRF Pb measurements in
bone and indirect measurements of bone processes with stable Pb isotopes have enhanced
the use of bone Pb as a biomarker of Pb body burden.
Several studies have found a stronger relationship between patella Pb and blood Pb than
tibia Pb and blood Pb (Park et al.. 2009a; Hu et al.. 1998; Hernandez-Avilaetal.. 1996;
Hu et al.. 1996b). 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
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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.
4.3.5.1 Children
As mentioned in Section 4.2.2.2, 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 4-9 shows a 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 (jig 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 cessation of 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).
Two alternative blood Pb metrics depicted in Figure 4-9 include the time-averaged and
time-integrated blood Pb concentrations. Both the time-averaged and time-integrated
blood Pb metrics display rates of change in response to the exposure event that more
closely approximate the slower kinetics of bone Pb and body burden, than the kinetics of
blood Pb concentration, with notable differences. The time-averaged blood Pb
concentration increases during the exposure event and decays following the event,
consistent with the changing body burden. The time-integrated blood Pb concentration
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(conceptually identical to cumulative blood lead index [CBLI] used in epidemiologic
studies) 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 exposure, 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 cessation of
exposure.
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 is illustrated in the
hypothetical simulations of an exposure event experienced by a child (Figure 4-10). This
pattern is similar for adults.
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upper
middle
10
5
CT>
~ 6
n
Q.
T3
O 4
CD
2


, - 	Blood

*
s
/
/
1 /
/\ — -Bone
[\ 	Body
> \
.1 \
/ N\

j /
/
* ' \
\ ¦
> s
\ * ^
\ ¦ —— ,

7
	
*• —


3.0
2.5
aj
o
20 2
CD
O
1 5 03
i.o Q
1.0
0.5
0.0
4	6
Age (year)
10
Blood
¦ Bone
4	6
Age (year)
lower
— Blood
— -Bone
2.5 03
4 6
Age (year)
Note; Blood Pb concentration is thought to parallel body burden more closely in children than in adults, due to more rapid turnover of
bone and bone-Pb stores in children (upper panel). Nevertheless, the time-averaged blood Pb concentration more closely tracks the
pattern of change in body burden (middle panel). The time-integrated blood Pb concentration increases overtime (lower panel).
Simulation based on ICRP Pb biokinetics model (Leggett. 19931.
Figure 4-9 Simulation of relationship between blood Pb concentration and
body burden in children, with a constant Pb intake from age
2 to 5.
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50
20
T3
40 -
- - Blood
—Absorption
O
c
0
0
0 2 4 6	8 10
Age (year)
Note: The simulations include a 3-year period of elevated Pb intake during ages 2-5 years. The time-integrated blood Pb
concentration closely parallels cumulative Pb absorption. Simulation based on ICRP Pb biokinetics model CLeaaett. 19931.
Figure 4-10 Simulation of relationship between time-integrated blood Pb
concentration and cumulative Pb absorption in children.
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 4-11 which
depicts a hypothetical simulation of an exposure event consisting of a 20-year period of
daily ingestion of Pb in an adult. The exposure event shown in the simulations gives rise
to a relatively rapid increase in blood Pb concentration, to a new quasi-steady state,
achieved in -75-100 days (i.e., approximately 3-4 times the blood elimination half-life).
In contrast, the body burden continues to increase during this period. Following cessation
of the exposure, blood Pb concentration declines relatively rapidly compared to the
slower decline in body burden. Careful examination of the simulation shown in Figure
4-11 reveals that the accumulation and elimination phases of blood Pb kinetics are not
4.3.5.2 Adults
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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. In
the adult simulation shown in Figure 4-11, the initial phase of elimination (the first
5 years following cessation of exposure at 50 years) has a half-time of approximately
14 years; however, the half-time increases to approximately 60 years during the period
5-30 years after cessation of exposure. These model predictions are consistent with the
slow elimination of Pb from blood and elimination half-times of several decades for bone
Pb (e.g., 16-98 years) that have been estimated from observations made on Pb workers
(Wilker et al.. 2011; Fleming et al.. 1997; Gerhardsson et al.. 1995). Based on this
hypothetical simulation, a blood Pb concentration measured 1 year following cessation of
a period of increased Pb uptake would show little or no appreciable change from prior to
the exposure event, whereas, the body burden would remain elevated. These simulations
in Figure 4-11 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.
One important potential implication of the profoundly different kinetics of Pb in blood
and bone is that, for a constant Pb exposure, Pb in bone will increase with increasing
duration of exposure and, therefore, with age. In contrast, blood Pb concentration will
achieve a quasi-steady state. As a result, the relationship between blood Pb and bone Pb
will diverge with increasing exposure duration and age. This divergence can impart
different degrees of age-confounding when either blood Pb or bone Pb is used as an
internal dose metric in dose-response models. In a review of epidemiologic studies that
evaluated the associations between blood Pb, bone Pb and cognitive function, the effects
of bone Pb were more pronounced 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; Kim et al. 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
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1	did not change with over the same period. Therefore, although older individuals tended to
2	have higher bone Pb levels, the 3-year temporal trend for individuals was a loss of Pb
3	from the more labile Pb stores in trabecular bone. Nie et al. (2011) observed that
4	longitudinal observations of blood and bone Pb in elderly adults did not show a
5	significant age effect on the association between blood Pb and bone Pb (patella and tibia),
6	when the sample population (n=776) was stratified into age tertiles (mean age 62, 69 or
7	77 years). Nie et al. (2009) did find that regressed function bone Pb and appeared to level
8	off at bone Pb levels >20 (.ig/g bone mineral.
upper
middle
10
_ 8 -
r 6 -
s 4
2
	Blood
— • Bone
	Body
20
1S
10
25 30 35 40 45 50 55 60 65 70
Age (year)
DO
Q
03
O
~_
5 •=¦
- -B ood
Bone
25 30 35 40 45 50 55
Age (year)
60 65
lower
350
20
300
250
200
x
100
	Blood
—¦-Bone
	Body
50
25 30 35 40 45 50 55 60 65 70
Age {year)
Note: A constant baseline intake results in a quasi-steady state blood Pb concentration and body burden (upper panel). A change in
Pb uptake gives rise to a relatively rapid change in blood Pb, to a new quasi-steady state, and a slower change in body burden. The
long-term time-averaged blood Pb concentration more closely tracks the slower pattern of change in body burden (middle panel).
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 deggett. 19931.
Figure 4-11 Simulation of relationship between blood Pb concentration, bone
Pb and body burden in adults.
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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 jxg/g bone mineral per (ig-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 over time 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.. 2009a).
Mobilization of Lead from Bone in Adulthood
In addition to changes in exposure (e.g., declines in exposure discussed in prior section),
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 et al.. 1997; Lagerkvist et al.. 1996; Schuhmacher et al.. 1996;
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Rothenberg etal.. 1994a). The U-shaped relationship reflects the relatively higher impact
of hemodilution in the second trimester versus the rate of bone Pb resorption
accompanying calcium 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. (1998a) 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.
(2011b) 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 calcium intake was 850 mg per day showed lower amounts of bone resorption
during late pregnancy and postpartum than those whose intake was 560 mg calcium per
day (Manton et al.. 2003). 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 (Ettinner 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: OR =
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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-Roio 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:
Gulson et al.. 1998b). 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).
The Pb content in some bones (i.e., mid femur and pelvic bone) plateau at middle age and
then decreases at older ages (Drasch et al.. 1987). This decrease is most pronounced in
females and may be due to osteoporosis and release of Pb from resorbed bone to blood
(Gulson et al.. 2002). Two studies indicate that the endogenous release rate in
postmenopausal women ranges from 0.13-0.14 (ig/dL in blood per jj.g/g bone and is
nearly double the rate found in premenopausal women (0.07-0.08 (ig/dL per jj.g/g bone)
(Popovic et al.. 2005: Garrido Latorre et al.. 2003). An analysis of data on blood lead
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 lead 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) has also been observed in elderly males (Nie et al.. 2009).
Studies of the effect of hormone replacement therapy on bone Pb mobilization have
yielded conflicting results (Popovic et al. 2005: Berkowitz et al.. 2004: Garrido Latorre
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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).
4.3.6 Relationship between Lead in Blood and Lead in Soft Tissues
Figure 4-12 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.
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„	Blood
¦ i
	Soft Tissue
0.6 c/>
o
4	6
Age (year)
10
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— 6 -
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	Blood
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	Soft Tissue -
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1.6 o
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(Q
0.4
~i	1	1	1	1	1	1	1	1- 0.0
25 30 35 40 45 50 55 60 65 70
Age (year)
Note: Simulation based on ICRP Pb biokinetics model CLegqett. 19931
Figure 4-12
Simulation of blood and soft tissue (including brain) Pb in
children and adults who experience a period of increased Pb
intake.
Information on Pb levels in human brain are limited to autopsy data and the simulation of
brain Pb shown in Figure 4-13 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
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1	(Leggett. 1993). This slow elimination rate is reflected in the slower elimination phase
2	kinetics in shown Figure 4-13. Although in this model, brain Pb to blood Pb transfer half-
3	times are assumed to be the same in children and adults, uptake kinetics are assumed to
4	be faster during infancy and childhood, which achieves a higher fraction of the soft tissue
5	burden in brain, consistent with higher brain/body mass relationships. The uptake half
6	times predicted by Leggett (1993) vary from 0.9 to 3.7 days, depending on age. Brain Pb
7	kinetics represented in the simulations are simple outcomes of modeling assumptions and
8	cannot currently be verified with available observations in humans.
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4	6
Age (year)
10
Blood
Brain
25 30 35 40 45 50 55 60 65 70
Age (year)
o
ff
oo
3
5'
-o
¦c
(Q
Note: Simulation based on ICRP Pb biokinetics model (Leqqett. 1993).
Figure 4-13 Simulation of blood and brain Pb in children and adults who
experience a period of increased Pb intake.
1	Urinary filtering and excretion of Pb is associated with plasma Pb concentrations. Given
2	the curvilinear relationship between blood Pb and plasma Pb, a secondary expectation is
3	for a curvilinear relationship between blood Pb and urinary Pb excretion that may
4	become evident only at relatively high blood Pb concentrations (e.g., >25 (ig/dL). Figure
5	4-14 shows these relationships predicted from the model. In this case, the exposure
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scenario shown is for an adult (age 40 years) at a quasi-steady state blood Pb
concentration; the same relationships hold for children. 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 4-14, 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 4-14, lower panel).
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Urine
— Plasma
> 60 -
0.10
0.05
10 20 30 40
Blood Pb (|ig/dL)
0.20 »
U)
3
u
0.15 g
f
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>-
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8
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— Bone 	Urine 	Blood


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s

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/
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-- 8
-- 6
-- 4
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i	1	1	1	1	1	r
25 30 35 40 45 50 55 60 65 70
Age (year)
DO
o
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(D
¦o
O"
3
(Q
Note: Lower panel: Simulation of blood Pb, bone Pb and urinary excretion of Pb in an adult who experiences a period of increased
Pb intake. Simulation based on ICRP Pb biokinetics model CLeaaett. 19931.
Figure 4-14 Relationship between Pb in urine and Pb in blood.
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4.4
Observational Studies of Lead Biomarker Levels
4.4.1
February 2012
Lead in Blood
Overall, trends in blood Pb levels have been decreasing among U.S. residents over the
past 20 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 1976-2008, with the most dramatic declines
coincident with the phase out of leaded gasoline (Pirkle et al. 1998; Brodv et al.. 1994;
Pirkle et al.. 1994; Schw artz and Pitcher. 1989). The temporal trend for the period
1988-2008 is shown in Figure 4-15. Summary statistics from the most recent publically
available data (1999-2008) are presented in Table 4-7 (CDC. 2011). The geometric mean
Pb concentration among children 1-5 years of age, based on the sample collected during
the period 2007-2008, was 1.51 (ig/dL (95% CI: 1.37, 1.66), which was a slight increase
from the previous year (1.46 (ig/dL, 95% CI: 1.36, 1.57). Figure 4-16 uses NHANES data
to illustrate temporal trends in the distribution of blood Pb levels among U.S. children
aged 12-60 months. The median blood Pb in this age group was 1.4 (ig/dL with a 95th
percentile value of 4.1 ^ig/dL in 2007-2008 (NCHS. 2010). For 2005-2008, 95% of
childhood blood Pb levels were less than 5 j^ig/dL. The geometric mean blood Pb
concentration among adults > 20 years of age was 1.38 (ig/dL (95% CI: 1.31, 1.46) for
the sample collected during the period 2007-2008 (CDC. 2011). Based on these same
data, the geometric mean for all males (aged > 1 y) was 1.47 (ig/dL (95% CI: 1.39, 1.56),
and for females (aged > 1 y) was 1.11 (ig/dL (95% CI: 1.06, 1.16).
There has been a steep decline in mean blood Pb levels from 1975 through 2010 among
all birth cohorts from 1975 to 2010; Figure 4-17. 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 over
time are nearly parallel among the cohorts born before 1930. This 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.
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1	When race/ethnicity groups were compared for years 1999-2004, blood Pb levels in
2	children were highest in the ethnicity category non-Hispanic black (GM 2.8, 95% CI: 2.5,
3	3.0) compared to the categories Mexican-American (GM 1.9, 95% CI: 1.7, 2.0) and
4	non-Hispanic white (GM 1.7, 95% CI: 1.6, 1.8) (Jones et al.. 2009a). Figure 4-18
5	demonstrates the change in percent of children (aged 1-5 years) with various blood Pb
6	levels by race/ethnicity between the survey during 1988-1991 and that during 1999-2004.
7	When these data for children aged 1-5 years were aggregated for all survey years from
8	1988 to 2004, residence in older housing, poverty, age, and being non-Hispanic black
9	were significant predictors of higher Pb levels (Jones et al.. 2009a).
5
1
Children 1-5 yrs
Adults > 20 yrs
88-91 91-94 99-00 01-02 03-04 05-06 07-08
Survey Period
Note: Shown are geometric means and 95% CIs based on data from NHANES III Phase 1 (Brodv et al.. 1994: Pirkle et al.. 19941:
NHANES III Phase 2 (Pirkle etal.. 1998): and NHANES IV (CDC. 2011). Data for adults during the period 1988-1994 are forages
20-49 years, and > 20 years for the period 1999-2008.
Figure 4-15 Temporal trend in blood Pb concentration.
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Table 4-7 Blood Pb concentrations in the U.S. population.
Survey Stratum
Period
Geometric Mean (|jg/dL)
95% Confidence Interval
Number of Subjects

1999-2000
1.66
1.60
1.72
7970

2001-2002
1.45
1.39
1.51
8,945
All
2003-2004
1.43
1.36
1.50
8,373

2005-2006
1.29
1.23
1.36
8,407

2007-2008
1.27
1.21
1.34
8,266

1999-2000
2.23
1.96
2.53
723

2001-2002
1.70
1.55
1.87 898
1-5 yr
2003-2004
1.77
1.60
1.95
911

2005-2006
1.46
1.36
1.57 968

2007-2008
1.51
1.37
1.66
817

1999-2000
1.51
1.36
1.66
905

2001-2002
1.25
1.14
1.36
1,044
6-11 yr
2003-2004
1.25
1.12
1.39
856

2005-2006
1.02
0.95
1.01
934

2007-2008
0.99
0.91
1.07
1,011

1999-2000
1.10
1.04
1.17
2,135

2001-2002
0.94
0.90
0.99
2,231
12-19 yr
2003-2004
0.95
O
oo
oo
1.02
2,081

2005-2006
0.80
0.75
0.85
1,996

2007-2008
0.80
0.74
0.86
1,074

1999-2000
1.75
1.68
1.81
4,207

2001-2002
1.56
1.49
1.62
4,772
>20 yr
2003-2004
1.52
1.45
1.60
4,525

2005-2006
1.41
1.34
1.48 4,509

2007-2008
1.38
1.31
1.46
5,364

1999-2000
2.01
1.93
2.09
3,913

2001-2002
1.78
1.71
1.86 4,339
Males
2003-2004
1.69
1.62
1.75
4,132

2005-2006
1.52
1.42
1.62
4,092

2007-2008
1.47
1.39
1.56
4,147

1999-2000
1.37
1.32
1.43
4,057

2001-2002
1.19
1.14
1.25
4,606
Females
2003-2004
1.22
1.14
1.31
4,241

2005-2006
1.11
1.05
1.17
4,315

2007-2008
1.11
1.06
1.16
4,119

1999-2000
1.83
1.75
1.91
2,742

2001-2002
1.46
1.34
1.60
2,268
Mexican - Americans
2003-2004
1.55
1.43
1.69
2,085

2005-2006
1.29
1.21
1.38
2,236

2007-2008
1.25
1.15
1.36
1,712

1999-2000
1.87
1.75
2.00
1,842

2001-2002
1.65
1.52
1.80 2,219
Non-Hispanic blacks
2003-2004
1.69
1.52
1.89 2,293

2005-2006
1.39
1.26
1.53
2,193

2007-2008
1.39
1.30
1.48 1,746
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Survey Stratum	Period	Geometric Mean (jjg/dL)	95% Confidence Interval	Number of Subjects
1999-2000	162	1.55, 1.69	2J16	
2001-2002	143	1.37, 1.48	^806	
Non-Hispanic whites	2003-2004	137	1.32, 1.43	3/478	
2005-2006	128	1.19, 1.37	^310	
	2007-2008	124	1.16, 1.33	3;461	
Source: Adapted from data from the NHANES (CDC. 2011).
Age strata correspond to the NHANES study design.	
70
50 -
S 40
£
f
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a
~~r~
s
a
ab
—r~
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8
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8
pS
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8
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Source: Adapted from data from the NHANES CNCHS. 20101
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Note: Top: all data. Bottom: data for subjects having blood Pb levels less than 15 |jg/dL.
Figure 4-16 Box plots of blood Pb levels among U.S. children (1-5 years old)
from the NHANES survey, 1988-2008.
o
CM
LO
O) ¦
-Q
CL
-o
O
-2 LO
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0) oo
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d)
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70 i
60
c 50
10
1999 - 2004
50
30
10 -
<1 1 - < 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 4-18 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).
Several studies have shown seasonal variation in blood Pb concentrations in children
[e.g., (Gulson et al.. 2008; Laidlaw et al.. 2005; Haley and Talbot. 2004; Johnson et al..
1996)1. Seasonal variation in blood Pb concentration was evident in Australian children
(n=107) with a group mean blood Pb level of 2.57 (ig/dL when repeated blood Pb
measurements were made over a 5-year period, with lower levels in summer compared
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with winter (Gulson et al. 2008). A cross-sectional study conducted in New York State
from 1995-1998 demonstrated seasonality, with the greatest percent of one and two year
old children with blood Pb > 10 (ig/dL (n=262,687) occurring in August and the lowest
percent in March and April (Halev and Talbot. 2004). Meteorological factors appear to
contribute to blood Pb seasonality. Laidlaw et al. ("2005) analyzed the temporal
relationships between child blood Pb concentrations and various atmospheric variables in
three cites (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 number of blood Pb
measurements included in the analyses were as follows: Indianapolis, 15,969; Syracuse,
14,457 (Johnson and Bretsch. 2002; Johnson et al. 1996); New Orleans, 2,295 (Mielkeet
al.. 2007a). The temporal variation in blood Pb concentrations in each city were predicted
by multivariate regression models that included the following significant variables: PMi0,
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 <0.00001).
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 y old children living within 3.5 km of a Mexican smelter at 0, 6, 12, and 60 months
after environmental intervention took place. 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 (ig/dL for the 598 initial participants, 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 |_ig/dL for the
remaining 232 participants, 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 (ig/dL. It should be noted that these studies are suggestive but not conclusive in
showing that exposure to sources elevates blood Pb because these studies do not control
for factors such as non-ambient in-home exposures. For this reason, Newhook et al.
(2003) point out that they utilize World Health Organization guidelines on fenceline
concentrations in lieu of blood Pb levels in the vicinity of industrial sources to quantify
exposures related to the sources. Lanphear et al. (1998) noted that the probability of
children having blood Pb > 10 j^ig/dL increases both with exterior soil Pb content and
interior Pb dust loading. Mielke et al. (2011a) noted significant increases in percentages
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of children younger than 7 y old with blood Pb level >10 (ig/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, 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 (Simon et al.. 2007). Median
blood Pb level among the children was 1.9 j^ig/dL at 2 months and increased to
13.6 (ig/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%
and 60% of the variance being explained by these variables, respectively; geometric
mean 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, blood Pb levels of 15 infants aged 6-12 months
were compared with cord blood Pb levels, resulting in significantly lower blood Pb levels
later in life (2.24 (ig/dL vs. 4.87 (ig/dL) (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 game had an average blood Pb concentration of
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 (Levesaue et al.. 2003). The geometric mean cord blood Pb level was
0.19 (imol/L, with a range of 0.01-1.31 (imol/L; the Canadian level of concern for cord
blood Pb is 0.48 |_imol/L. The authors contrasted the finding that 7% of Inuit newborns
had cord blood Pb concentration > 0.48 (imol/L 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 3.2.2.1 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 m
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of the airports having significantly increased blood Pb levels with residential proximity to
airports ((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.
Trends in blood Pb levels have been accompanied by changes in Pb isotope ratios for
blood Pb. Isotopic ratios, described in Sections 3.2 and 3.3 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 4-19).
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
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 m-
6 y. 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.
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1
2
3
4
5
6
7
8
9
10
11
12
13
14
source
O Adelaide
yv BK Adult
^ females
~ BK children
A Broken Hill
V Hobart
<| PbinCa
[> Port Pi rie
+
Sydney
children
n 17.20
R Sq Linear = 0.53
16.40-
1990 1992 1994 1996 1998 2000
(a)	Year	(b)
Source: Reprinted with permission of Academic Press, Gulson et al. (2006aJ
12.5—
>

>
10.0—

7.5-


o
m
> A
n
Q.
3 ^ <
& A 
-------
Table 4-8 Epidemiologic studies that provide bone Pb measurements for non-
occupational^ exposed populations
Reference Study Methods
Prior Pb Exposure
Bone Pb	Distribution of
biomarker Bone Pb Cone, (ng/g) Bone Pb (ng/g)
Bandeen-
Roche et al.
Cohort: Baltimore Memory Study cohort
Age (yrs):50-70
N:1140
Location: Baltimore, MD
Study Period:2001-2005	
Cumulative
Tibia
Mean+SD
Tibia: 18.8 ±11.6
Not reported
Bellinger etal.
(1994a)
Cohort: Not reported
Age (yrs): 5-8 (recruited); 19-20 (follow-up)
N: 79
Location: Boston, MA
Study Period: 1989-1990
Cumulative
Tibia
Patella
Mean (Range):
Tibia: 5.4 (3-16)
Patella: 9.2 (4-18)
High exposure: >24
Low exposure: <8.7
Cheng et al.
(2001)
Cohort: 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
Cumulative
Tibia	Mean+SD	Lowest quintile: Tibia:
Patella Tibia:	8.5
Normotensive: 20.27 + 11.55	Patella: 12.0
Borderline hypertension:
23.46 + 15.02	Highest quintile:
Definite hypertension: 22.69 +	Tibia: 36.0
14.71	Patella: 53.0
Patella:
Normotensive: 28.95 + 18.01
Borderline hypertension:
33.73 ±21.76
Definite hypertension: 32.72 ±
19.55
Coon et al.
Cohort: 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)
Cumulative
Tibia
Calcaneus
Mean±SD:
Tibia: 12.5 ±7.8
Calcaneus: 20.5 ± 10.2
Tibia
Q1
Q2
Q3
Q4
0-5.91
5.92-10.40
10.41-15.50
>15.51
Calcaneus
Q1
Q2
Q3
Q4
0-11.70
11.71-19.07
19.08-25.28
>25.29
Elmarsafawy
et al. 12006)
Cohort:Normative Aging Study
Age (yrs): Not reported
N: 471 elderly males
Location: Greater Boston area, MA
Study Period: 6/1991-12/1994
Not reported
Tibia
Patella
Mean±SD:
Tibia: 21.6 ±1 2.0
Patella: 31.7 ±18.3
Not reported
Glass et al.
Cohort:Baltimore Memory Study
Age (yrs): Mean: 59.4; Range: 50-70
N: 1,001
Location: Baltimore, MD
Study Period: 2001-2005
Cumulative (lifetime)
Tibia
Mean±SD:
Tibia: 18.8 ±11.1
NPH Scale:
Lowest fertile: Mean
Tibia level: 16.3 ±
11.0
Middle fertile: Mean
Tibia level: 19.3 ±
10.7
Highest fertile: Mean
Tibia level: 20.3 ±
11.4
Hsieh et al.
(2009b)
Cohort:Not reported
Age (yrs): Mean: Control: 46.06
N: 18 controls
Location: Not reported
Study Period: Not reported
Control group for
occupational exposure
group
Tibia	Mean±SD
Patella Tibia Control: 18.51 ±22.40
Patella Control: 7.14 ±9.81
Not reported
Hu et al.
(1996b)
[As reported in
Navas-Acien
etal., (2008)1
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
Cumulative
Tibia
Patella
Mean±SD:
Tibia: 21.8 ±12.1
Patella: 32.1 ±18.7
Range:
Tibia: <1-96
Patella: 1-142
Figures 1 and 2 show
both types of bone Pb
levels increasing with
age
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Reference Study Methods
Prior Pb Exposure
Bone Pb	Distribution of
biomarker Bone Pb Cone, (ng/g) Bone Pb (ng/g)
Jain et al.
(2007)
Cohort:VA-Normative Aging Study
Age (yrs): Not reported
N: 837 males
Location: Greater Boston, MA
Study Period: 9/1/1991-12/31/2001
Not reported
Tibia	Mean + SD
Patella Tibia:
Non-Cases: 21.4+ 13.6
Cases: 24.2 +15.9
Patella:
Non-cases: 30.6+19.7
Cases: 36.8 + 20.8
Range:
Tibia:
Noncases:
-3-126
Cases: -5-75
Patella:
Noncases: -10-165
Cases: 5-101
Kamel et al.
(2002): Kamel
etal. (2005);
Kamel et al.
Mean + SD (Range):
Tibia:
Non-cases:
Tertile 1:10.2 ±3.8 (-
3-15)
Tertile 2:19.1+2.3
(16-23)
Tertile 3: 35.5+14.4
(24-126)
Cases:
Tertile 1:10.1 +5.3 (-
5-15)
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)
Tertife 1:15.3+4.3(5-
19)
Tertile 2: 25.7+ 3.8
(21-33)
Tertile 3: 53.3 + 17.3
(35-101)
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
Cumulative
Control group for
occupational exposure
group
Tibia	Mean+SE
Patella Tibia Controls: 11.1 + 1.6
Patella Controls: 16.7 + 2.0
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)
Khalil et al.
(2009b)
Cohort:1982 Pb Occupational Study
Age (yrs): Control mean: 55
N: 51 controls
Location: Eastern Pennsylvania
Study Period: 1982-2004	
Control group for
occupational exposure
group
Tibia
Median (IQR)
Tibia Control: 12 (
1-32)
Not reported
Korrick et al.
(1999) (As
reported in
from Navas-
Acien et al.,
Cohort: Nurses' Health Study
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
Nonoccupational^ exposed
Tibia	Mean + SD
Patella Tibia:
Combined: 13.3 + 9.0
Cases: 13.0 + 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
Patella:
10th percentile: 6
90th percentile: 31
Lee et al.
(2001a) [As
reported in
Navas-Acien
etal., (2008)1
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
Control group for
occupational exposure
group
Tibia	Mean + SD
Tibia Controls: 5.8 + 7.0
Range
Tibia Controls:
-11-27
Not reported
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Reference Study Methods
Prior Pb Exposure
Bone Pb	Distribution of
biomarker Bone Pb Cone, (ng/g) Bone Pb (ng/g)
Martin et al.
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)
Cumulative (lifetime)
Tibia
Mean + SD
Tibia: 18.8 ±12.4
Tibia IQR: 11.9-24.8
Needleman et
al. (2002)
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: 15.8 ±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
Not reported
Tibia	Mean ± SD	Table 4 of paper
Tibia Cases (ppm):	distributes bone Pb
All subjects: 11.0 ± 32.7	by > 25 or <25 for
African American: 9.0 ± 33.6	race, two parental
White: 20 ± 27.5	figures, and parent
Tibia Controls (ppm):	occupation
All subjects: 1.5 ±32.1
African American: -1.4 ± 31.9
White: 3.5 ±32.6
Osterberg et
al. (1997) [As
reported in
Shih et al.,
Cohort: Not reported
Age (yrs): Median: 41.5
N: 19 male controls
Location: Not reported
Study Period: Not reported
Control group for
occupational exposure
group
Finger bone
Median (range)
Finger Bone Controls:
4 (-19-18)
Park et al.
Not reported
Cohort: 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)
Not reported
Tibia
Patella
Median (IQR)
Tibia: 19.0 (11-28)
Patella: 23.0 (15-34)
Estimated Patella8:16.3
(10.4-25.8)
Park et al.
(2009b)
Median (IQR) for No.
of metabolic
abnormalities:
Tibia:
18.5(10.5-23)
19(11-28)
19(12-26)
Patella:
22 (13.5-32)
25(16-36)
20(15-32)
Estimated Patella:
16.3(10.8-24.8)
17.1 (11-29.3)
15.1 (9.4-22.1)
Cohort: Normative Aging Study
Age (yrs): Mean: 67.3 ± 7.2
N: 613 males
Location: Greater Boston, MA
Study Period: 8/1991 - 12/1995
Not reported
Tibia
Patella
Median (IQR)
Tibia: 19 (14-27)
Patella: 26 (18-37)
Table 1 of paper
distributes tibia and
patella Pb by
genotype; Table 2 of
paper distributes tibia
and patella Pb by
number of gene
variants
Park et al.
(2010)
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
Cumulative (chronic
exposure)
Tibia
Patella
Mean±SD
Tibia: 22.5 ±14.2
Patella: 32.5 ± 20.4
Tibia IQR: 15
Patella IQR: 21
Table 2 of paper
provides age-
adjusted mean bone
Pb levels (age, race,
education, smoking
[pack-yr],
occupational noise,
noise notch, BMI,
hypertension,
diabetes)
Payton et al. Cohort: VA Normative Aging Study cohort
(1QQ81
Age (yrs): Mean: 66.8
N: 141 males
Location: Boston, MA
Study Period: 4/1993-3/1994
Not reported
Tibia
Patella
Mean ± SD
Tibia: 22.5 ±12.2
Patella: 31.7 ±19.2
Not reported
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Reference Study Methods
Prior Pb Exposure
Bone Pb	Distribution of
biomarker Bone Pb Cone, (ng/g) Bone Pb (ng/g)
Peters et al.
(2007)
Cohort: Normative Aging Study cohort
Age (yrs): Mean: 66.9
N: 513 male cases
Location: Boston, MA
Study Period: 1991-1996	
Cumulative
Tibia
Patella
Mean ± SD
Tibia: 21.5 ±13.4
Patella: 31.5 ±19.3
Not reported
Rajan et al.
(2007)
Cohort: VA Normative Aging Study Cohort
Age (yrs): Mean: 67.5 (at bone scan)
N: 1075 males
Location: Boston, MA
Study Period: 1991-2002
Not reported
Tibia
Patella
Mean ± SD
Tibia: 22.1 ±13.8
Patella: 31.4 ±19.6
Not reported
Rajan et al. Cohort: VA Normative Aging Study Cohort
nnnQ\	13 13 1
Age (yrs): > 45
N: 720 males
Location: Boston, MA
Study Period: 1993-2001
Current and cumulative
Tibia
Patella
Mean ± SD
ALAD1-1
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
Not reported
Rhodes et al.
Cohort: VA Normative Aging Study Cohort
Age (yrs): Mean: 67.1
N: 526 males
Location: Boston, MA
Study Period: 1/1/1991-12/31/1995
Not reported
Tibia
Patella
Mean ± SD
Tibia: 21.9 ±13.5
Patella: 32.1 ±19.8
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)
Roels et al.
Cohort: Not reported
Age (yrs): 30-60
N: 68 males
Location: Belgium
Study Period: Not reported
Control group for
occupational exposure
group
Rothenberg et
al. (2002b), as
reported in
Navas-Acien
et al. (2008)
Tibia	Geometric Mean (Range) Not reported
Tibia Controls:
Normotensive: 21.7 (<15.2-
69.3)
Hypertensive: 20.2 (<15.2-
52.9)
Total: 21.4 (<15.2-69.3)
Cohort: 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
Not reported
Tibia	Mean ± SD
Calcaneus Tibia: 8.0 ±11.4
Calcaneus: 10.7 ± 11.9
Tibia quartiles:
Q1
Q2
Q3
Q4
¦33.7-0.9
1.0-8.0
8.1-16.1
16.2-42.5
Calcaneus quartiles:
Q1
Q2
Q3
Q4
¦30.6-3.0
3.1-10.0
10.1-18.7
18.8-49.0
Shih et al.,
Cohort: Baltimore Memory Study cohort
Age (yrs): Mean: 59.39
N: 985
Location: Baltimore, MD
Study Period: Not reported
Not reported
Tibia
Mean ± SD:
Tibia: 18.7 ±11.2
Not reported
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Reference Study Methods
Bone Pb	Distribution of
Prior Pb Exposure biomarker Bone Pb Cone, (ng/g) Bone Pb (ng/g)
Stokes et al.
(1998V as
reported in
Shih et al.
(2007)
Cohort: Not reported
Age (yrs): 19-29 (in 1994); Mean ± SD:
Cases: 24.3 ± 3.18 Control: 24.2 ± 3.02
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)	Tibia
Environmental (resided near
Pb smelter during childhood)
Mean (Range):
Tibia Cases: 4.6 (-28.9-37)
Tibia Controls: 0.6 (-46.4-
17.4)
Van
Wijngaarden
et al. 12009)
Tibia
No. of Cases:
<1 |jg/g: 31.5%
1-5 |jg/g: 24.4%
5-10 pg/g: 22.3%
>10 pg/g: 21.8%
No. of Controls:
<1 pg/g: 50.4%
1-5 pg/g: 25.6%
5-10 pg/g: 19.4%
>10 pg/g: 4.7%
Mean ± SD Tibia
concentration by age
group
19-21
22-24
25-27
28-30
Cases:
1.47	±8.35
4.48	+ 7.45
4.82 ± 8.92
6.64 ± 9.53
Controls:
19-21
22-24
25-27
28-30
1.27 ±6.60
-0.61 ±6.19
0.60 ± 8.60
1.74 ±6.42
Cohort: Not reported
Age (yrs): Mean: 61.5
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
Wfesserman et
al. (2003)
Cohort: Yugoslavia Prospective Study of
Environmental Pb Exposure
Age (yrs): 10-12
N: 167 children
Location: Kosovska, Mitrovica, Kosovo,
Yugoslavia; Pristina, Kosovo, Yugoslavia
Study Period: 5/1985-12/1986
(mother's enrollment); 1986-1999 (follow-up
through age 12 yr); Tibia Pb measured 11-
13 yr old
Cumulative (lifetime)
Environmental
(Pb smelter, refinery, battery
plant)
Tibia	Mean ± SD:
Tibia
Pristina: 1.36 ± 6.5
Mitrovica: 39.09 ± 24.55
Tibia quartiles:
Q1
Q2
Q3
Q4
¦14.4-1.85
1.85-10.5
10.5-35
35-193.5
Table 3 of paper
distributes tibia Pb by
sex, ethnicity,
address at birth
relative to factory, and
maternal education
VNfeisskopf et
al. (2004). as
reported in
Shih et al.
(2007)
Cohort: Normative Aging Study
Age (yrs): Mean ± SD: 67.4 ± 6.6
N: 466 males
Location: Boston, MA
Study Period: 1991-2002
Environmental
Tibia
Patella
Median (IQR)
Tibia: 19 (12,26)
Patella: 23 (15, 35)
Tibia IQR: 14
Patella IQR: 20
Table 3 of paper
shows mean Pb
levels across categor-
ical variables (yr of
education, smoking
status, computer
experience, first
language English)
VNfeisskopf et
al. (2007b)
Cohort: 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
Not reported
Tibia	Median (IQR)
Patella Tibia
Lowest quintile: 13(9-17)
Highest quintile: 41 (38-59)
Patella
Lowest quintile: 9 (5-15)
Highest quintile: 63 (43-86)
Not reported
VNfeisskopf et
al. (2007a)
Cohort: VA Normative Aging Study cohort
Age (yrs): Mean: 68.7
N: 1,089 males
Location: Boston, MA
Study Period: 1993-2001
Concurrent and cumulative
Tibia
Patella
Median (IQR)
Tibia: 20 (13-28)
Patella: 25 (17-37)
Table 1 of paper
shows distribution of
Pb biomarkers by
categories of covar-
iates (age, education,
smoking status, alco-
hol intake, physical
activity, computer
experience, first
language English)
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Reference Study Methods
Prior Pb Exposure
Bone Pb	Distribution of
biomarker Bone Pb Cone, (ng/g) Bone Pb (ng/g)
VNfeisskopf et
al. 12009)
Cohort: Normative Aging Study; (95% white)
Age (yrs): Mean + SD (at Patella baseline);
Tertile 1:65.2 ±7.1;
Tertiie 2: 66.5 ±6.5
Tertile 3: 70.2 ±7.2
N: 868 males
Location: Greater Boston area, MA
Study Period: 1991-1999
Cumulative
Tibia
Patella
Mean ± SD
Tibia: 21.8 ±13.6
Patella: 31.2 ±19.4
Patella tertiles:
<22
22-35
>35
VNfeisskopf et
al. (2010)
Cohort: 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)
Cumulative
Tibia
Patella
Mean ± SD:
Tibia: 10.7 ±12.1
Patella: 13.6 ±15.9
Wfeuve et al. Cohort: VA Normative Aging Study cohort
'-1	Age (yrs): > 45
N: 720 males
Location: Boston, MA
Study Period: 1991 (measuring
bone Pb levels)
End date not reported
Tibia quartiles:
Q1
Q2
Q3
Q4
<3.1
3.5-9.6
10.0-17.0
>17.3
Patella quartiles:
Q1
Q2
Q3
Q4
<2.7
3.5-11.0
11.3-20.9
>20.9
Cumulative
Tibia	Median (1st-3rd quartile):
Patella Tibia: 19 (13-28)
Patella: 27 (18-39)
Table 1 of paper
shows distribution of
mean Pb biomarker
levels by
characteristics of
participants (age,
education, computer
experience, smoking
status, alcohol
consumption, tertile of
calcium intake, tertile
of physical activity,
diabetes)
Wfeuve et al.
Cohort: Nurses' Health Study cohort
Age (yrs): 47-74
N: 587 females
Location: Boston, MA
Study Period: 1995-2005	
Recent and cumulative
Tibia
Patella
Mean ± SD:
Tibia: 10.5 ±9.7
Patella: 12.6 ±11.6
Not reported
Wright et al.
(2003b), as
reported in
Shih et al.
(2007)
Cohort: Normative Aging Study
Age (yrs): Mean ± SD: 68.2 ± 6.9
N: 736 males
Location: Boston, MA
Study Period: 1991-1997
Environmental
Tibia
Patella
Mean ± SD:
Tibia: 22.4 ±15.3
Patella: 29.5 ±21.2
Tibia:
Difference in mean
from Lowest-highest
quartile: 34.2
Patella:
Difference in mean
from lowest-highest
_2uartilei47^_^_
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Table 4-9 Epidemiologic studies that provide bone Pb measurements for
occupationally exposed populations
Reference
Study Methods
Prior Pb Exposure
Bone Pb Bone Pb Concentration
biomarker	(H9'9)
Distribution of Bone
Pb (ng/g)
Bleecker et al.
(1997), as
reported in
Shih et al.
(2007)
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
Occupational
(Pb smelter workers)
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
Not reported
Bleecker et al.
(2007a)
Cohort: Not reported
Age (yrs): Mean: 39.7
N: 61
Location: Northern Canada
Study Period: Not Reported
Occupational (primary Pb Tibia
smelter workers)
Mean:
Tibia: 38.6
Not reported
Caffo et al. Cohort: Not reported	Cumulative
(2008)	Age (yrs): Mean: 60.39	Occupational
N: 513 males	(Former organolead
Location: Delaware and New Jersey U.S. manufacturing workers)
Study Period: 1994-1997
(Phase 1 recruitment);
2001-2003 (Phase 2 recruitment)
Tibia
Mean ± SD:
Peak Tibia: 23.99 ±18.46
Not reported
Dorsey et al. Cohort: Not reported
(2006)	Age (yrs): Mean: 43.4
N: 652
Location: Korea
Study Period: 10/24/1997-8/19/1999
(enrolled)
Occupational (Pb workers) Tibia
Patella
Mean ± SD:
Tibia: 33.5 ±43.4
Patella: 75.1 ±101.1
Not reported
Glenn et al.
(2003V as
reported in
Navas-Acien
et al. (2008)
Cohort: Not reported
Age (yrs): 40-70; Mean: 55.8 (baseline)
N: 496 males
Location: Eastern U.S.
Study Period: 6/1994-6/1996 (enrolled);
6/1998 (follow-up period ended)
Occupational
(Chemical manufacturing
facility; inorganic and
organic Pb)
Tibia	Mean ± SD:
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
Not reported
Glenn et al. Cohort: Not reported
(2006)	y\ge (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
Cumulative and recent
Occupational (Pb-using
facilities)
Tibia
Mean ± SD:
Tibia: 38.4 ±42.9
Tibia-V\fomen:
Visit 1:28.2±19.7
Visit 2: 22.8±20.9
Tibia-Men:
Visit 1: 41.7±47.6
Visit 2: 37.1 ±48.1
Not reported
Hanninen et
al. (1998). as
reported in
Shih et al.
(2007)
Cohort: Not reported
Age (yrs): Mean±SD:
Male: 43; Female: 48
BPb (max) < 2.4 pmol/L: 41.7 ±9.3
BPb (max) >2.4 pmol/L: 46.6 ±6.2
N: 54; (43 males, 11 females)
Location: Helsinki, Finland
Study Period: Not reported
Occupational (Pb acid
battery factory workers)
Tibia
Calcaneus
Mean±SD:
Tibia:
BPb (max) < 2.4 pmol/L:
19.8 ±13.7
BPb (max) >2.4 pmol/L:
35.3 ±16.6
Calcaneus:
BPb (max) < 2.4 pmol/L:
78.6 ±62.4
BPb (max) >2.4 pmol/L:
100.4 ±43.1
Not reported
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Bone Pb
Bone Pb Concentration
Distribution of Bone
Reference
Study Methods
Prior Pb Exposure
biomarker
(Hg/g)
Pb (ng/g)
Hsieh et al.
Cohort: Not reported
Occupational
Tibia
Mean + SD
Not reported
2009(2009b)
Age (yrs): Mean:
(Pb paint factory workers)
Patella
Tibia


Cases: 45.71


Case: 61.55+ 30.21


Controls: 46.06


Control: 18.51 + 22.40


N: 22 cases; 18 controls





Location: Not Reported


Patella


Study Period: Not reported


Case: 66.29+ 19.48




Control: 7.14+ 9.81

Kamel et al.
(2002): Kamel
etal. (2005);
Kamel et al.
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,
Study Period: 1993-1996
Cumulative	Tibia
Occupational (Pb fumes, Patella
dust, or particles)
MA)
Mean + SE
Tibia
Cases: 14.9 ± 1.6
Controls: 11.1 + 1.6
Patella
Cases: 20.5 + 2.1
Controls: 16.7 + 2.0
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)
Khalil et al.
(2009b)
Cohort: 1982 Pb 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
Cases: 57 (20-86)
Controls: 12 (-8-32)
Not reported
Osterberg et
al. (1997). as
reported in
Shih et al.
(2007)
Cohort: Not reported
Age (yrs): Median: 41.5
N: 38 male cases; 19 male controls
Location: Not reported
Study Period: Not Reported
Occupational (secondary
Pb smelter-inorganic Pb)
Finger bone
Median
Finger Bone:
High Cases: 32
Low cases: 16
Control: 4
Not reported
Range
Finger Bone:
High Cases: 17-101
Low cases: -7-49
	Control:-19-18	
Roels et al. Cohort: Not reported	Occupational (Pb smelter Tibia	Geometric Mean (Range) Not reported
(1994)	Age (yrs): 30-60	workers)
N: 76 male cases; 68 male controls	Mean case exposure: 18 yr	Tibia Cases:
Location: Belgium	(range: 6 to 36 yr)	Normotensive: 64.0 (19.6-
Study Period: Not Reported	167.1)
Hypertensive: 69.0 (21.7-
162.3)
Total: 65.8 (19.6-167.1)
Tibia Controls:
Normotensive: 21.7 (<15.2-
69.3)
Hypertensive: 20.2 (<15.2-
52.9)
	Total: 21.4 (<15.2-69.3)	
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Reference
Study Methods
Prior Pb Exposure
Bone Pb Bone Pb Concentration
biomarker	(H9'9)
Distribution of Bone
Pb (ng/g)
Schwartz
(2000c) et al..
as reported in
Shih et al.,
(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.
Study Period: 6/1994-10/1997 (enrolled);
Completed 2-4 annual follow-up visits;
Tibia Pb taken in 3rd year
Occupational
(tetraethyl and tetramethyl
Pb manufacturing facility)
Tibia
Mean + SD
Current Tibia: Cases: 14.4 +
9.3
Peak Tibia:
Cases: 22.6 +16.5
Not reported
Schwartz et
Cohort: Not reported
Occupational
Tibia
Mean + SD Not reported
al. (2000b), as
Age (yrs): 41.7-73.7 (Combined)
(former organolead

Tibia:
reported in
Mean + SD:
manufacturing workers)

Combined: 14.4 + 9.3
Navas-Acien
et al. (2008)
Combined: 57.6 + 7.6


Hypertensive: 15.4 + 9.1
Hypertensive: 60.2 + 6.9


Nonhypertensive: 14.0 + 9.3

Nonhypertensive: 56.6 + 7.5
N: 543 males


Range Tibia:
Combined: -1.6-52

Location: Eastern U.S.



Study Period: 1995 (recruited);




1996-1997 (Tibia Pb




taken during the 3rd yr)



Schwartz et
al. (2001):
Lee et al.
(2001a)
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
Occupational (battery Tibia
manufacturing, secondary
smelting, Pb oxide
manufacturing, car radiator
manufacturing)
Mean + SD
Tibia
Cases: 37.1 + 40.3
Control: 5.8+ 7.0
Range:
Tibia
Cases: -7-338
Controls:-11-27
Not reported
Schwartz et
Cohort: Not reported
Occupational (current and Tibia
Mean + SD
Tibia:
al. (2005)
Age (yrs): Mean at 1st visit: 41.4
former Pb workers)
Tibia: 38.4+ 43
25th percentile at V1:

N: 576


14.4

Location: South Korea


75th percentile at V1:

Study Period: 10/1997-6/2001


47.1
Stewart et al.
(1999V as
reported in
Shih et al.,
(2007)
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
Occupational
(tetraethyl and tetramethyl
Pb manufacturing facility)
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
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
15-19.99
20-24.99
25-29.99
70 (12.9)
87(16.1)
79 (14.6)
55(10.2)
;30:137 (26.1)
Stewart et al. Cohort: Not reported
(2006)	Age (yrs): Mean: 56.1
N: 532 males
Location: Eastern U.S.
Study Period: 1994-1997; 2001-2003
Cumulative
Occupational
(Organolead workers - not
occupationally exposed to
Pb at time of enrollment)
Tibia
Mean + SD
Current Tibia: 14.5 + 9.6
Peak Tibia: 23.9 ± 18.3
Not reported
Wfeaver et al. Cohort: Not reported
(2008)	Age (yrs): Mean + SD: 43.3 + 9.8
N: 652
Location: South Korea
Study Period: 12/1999-6/2001
Occupational
(Current and former Pb
workers; plants produced
Pb batteries, Pb oxide, Pb
crystal, or radiators, or were
secondary Pb smelters)
Patella
Mean+SD
Patella: 37.5 + 41.8
Not reported
4.4.3 Lead in Urine
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5
6
Table 4-10 Urine Pb concentrations in the U.S. population
Survey Stratum
Period
Geometric Mean (|jg/g CRf
95% Confidence Interval
Number of Subjects

1999-2000
0.721
0.700, 0.742
2,465

2001-2002
0.639
0.603, 0.677
2,689
All
2003-2004
0.632
0.603, 0.662
2,558

2005-2006
0.546
0.502, 0.573
2,576

2007-2008
0.515
0.483, 0.549
2,627

1999-2000
1.170
0.975, 1.41
340

2001-2002
0.918
0.841, 1.00
368
6-11 yr
2003-2004
0.926
0.812, 1.06
290

2005-2006
0.628
0.563, 0.701
355

2007-2008
0.644
0.543, 0.763
394

1999-2000
0.496
0.460, 0.535
719

2001-2002
0.404
0.380, 0.428
762
12-19 yr
2003-2004
0.432
0.404, 0.461
725

2005-2006
0.363
0.333, 0.395
701

2007-2008
0.301
0.270, 0.336
376

1999-2000
0.720
0.683, 0.758
1,406

2001-2002
0.658
0.617, 0.703
1,559
>20 yr
2003-2004
0.641
0.606, 0.679
1,543

2005-2006
0.573
0.548, 0.600
1,520

2007-2008
0.546
0.513, 0.580
1,857

1999-2000
0.720
0.679, 0.763
1,227

2001-2002
0.639
0.607, 0.673
1,334
Males
2003-2004
0.615
0.588, 0.644
1,281

2005-2006
0.551
0.522, 0.582
1,271

2007-2008
0.502
0.471,0.534
1,327

1999-2000
0.722
0.681,0.765
1,238

2001-2002
0.639
0.594, 0.688
1,355
Females
2003-2004
0.648
0.601,0.698
1,277

2005-2006
0.541
0.507, 0.577
1,305

2007-2008
0.527
0.489, 0.568
1,300

1999-2000
0.940
0.876, 1.01
884

2001-2002
0.810
0.731,0.898
682
Mexican - Americans
2003-2004
0.755
0.681,0.838
618

2005-2006
0.686
0.638, 0.737
652

2007-2008
0.614
0.521,0.722
515

1999-2000
0.722
0.659, 0.790
568
Non-Hispanic blacks
2001-2002
0.644
0.559, 0.742
667

2003-2004
0.609
0.529, 0.701
723
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Urine Pb concentrations in the U.S. general population have been monitored in the
NHANES. Data from the most recent survey (CDC. 2011) are shown in Table 4-10. The
geometric mean for the entire sample for the period 2007-2008 (n = 2,627) was 0.52 jj.g/g
creatinine (95% CI: 0.48, 0.55). The geometric means for males (n = 1,327) and females
(n = 1,300) were 0.50 jj.g/g creatinine (95% CI: 0.47, 0.53) and 0.53 jj.g/g creatinine (95%
CI: 0.49, 0.57), respectively.

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7
8
9
10
11
Survey Stratum
Period
Geometric Mean (|jg/g CR)a
95% Confidence Interval
Number of Subjects

2005-2006
0.483
0.459, 0.508
692

2007-2008
0.452
0.414, 0.492
589

1999-2000
0.696
0.668, 0.725
822

2001-2002
0.615
0.579, 0.654
1,132
Non-Hispanic whites
2003-2004
0.623
0.592, 0.655
1,074

2005-2006
0.541
0.500, 0.585
1,041

2007-2008
0.506
0.466, 0.550
1,095
aValues are |jg Pb/g creatinine (CR)
Source: Based on data from the NHANES (CDC. 2011)
4.4.4 Lead in Teeth
The influence of historical Pb exposures was recently studied by Robbins et al. (2010V
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); Figure 4-20.
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 <0.0001) and for this study (p <0.0002).
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19
20
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'A.
-A
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0)
2 25-

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31
32
authors did not clarify if average or median values were presented, nor did they adjust for
potentially confounding factors.
4.5 Empirical Models of Lead Exposure-Blood Lead 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 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 the
relationship between human exposure and tissue Pb levels, parameters which are
expected to vary spatially and temporally. Thus, extrapolation of regression models to
other spatial or temporal contexts, which is often necessary for regulatory applications of
the models, can be problematic.
4.5.1 Air Lead-Blood Lead Relationships
The 1986 Pb AQCD (U.S. EPA. 1986a) described epidemiological studies of
relationships between air Pb and blood Pb. Of the studies examined, the total blood Pb-air
Pb slope (when considering both direct and indirect exposures derived from air) was
estimated to be approximately double the slope estimated from the direct contribution due
to inhaled air alone (U.S. EPA. 1986a). Much of the pertinent earlier literature (e.g., prior
to 1984) was summarized by Brunekreef (1984). Based on the studies available at that
time that considered multiple air-related Pb exposure pathways in the aggregate, 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 (ig/dL) and high air Pb levels (much greater
than 10 (ig/m3) prevailed relative to lower environmental exposures which showed lower
blood Pb and air Pb concentrations (<30 (ig/dL and <3 |ig/m3). For those environmental
exposures, it was concluded that the relationship between blood Pb and air Pb "for direct
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36
37
38
inhalation appears to be approximately linear in the range of normal ambient exposures
(0.1-2.0 (ig/m3)" (pp 1-98 of the 1986 Pb AQCD). Based on meta-analysis of 18 studies of
urban or industrial-urban populations, Brunekreef (1984) estimated the blood Pb-air Pb slope
for children to be 0.3485 ln||_ig/dL blood Pb] per ln| (.ig/ni3 air Pb] (R2 = 0.69; Figure 4-21).
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 study (n=96). The 1.5 (ig/m3 value
is the median of the air Pb concentrations that match the 96 blood Pb concentrations in
Figure 3 of Brunekreef et al. (1984), taken from the Appendix to the same paper. When the
analysis was limited to children whose blood Pb concentrations were <20 (ig/dL, 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, n=43).
Newer studies that provide estimates for the blood Pb-air Pb slope are described below. In
some studies, the blood Pb-air Pb relationship was described with a non-linear regression
function, in which the blood Pb-air Pb slope varied with air Pb concentration. In Table 4-11,
slopes corresponding to the central estimate of the air Pb concentrations are provided, which
are considered to be the best estimate of the slope from 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
slopes; therefore, Figure 4-21 depicts the relationship between the blood Pb-air Pb slope as a
function of air Pb concentration for the range of air Pb concentrations evaluated in each
study (the central estimate is also shown). Figure 4-21 provides a more informative picture
of the extent to which slope estimates vary (and overlap) within and between studies. In the
Schnaas et al. (2004) analysis, the effect of air Pb on blood Pb may have been
underestimated due to inclusion of location and SES terms in their 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. With the exception of Ranft et al. (2008). all studies included in Table
4-11 and Figure 4-22 include a blood Pb-air Pb relationship that reflects all air-related
pathways of exposure. The Ranft et al. (2008)study includes a separate term for soil Pb, so
the blood Pb-air Pb slope presented for that study underestimates the slope that would reflect
all air-related pathways, since soil Pb encompasses deposited ambient air Pb. The Ranft et al.
(2008) model is log-linear, with the natural logarithm of blood Pb being a function of linear
increase in air Pb. This results an upward curvature of the blood Pb-air Pb relationship. By
comparison, log-log models predict an increase in the blood Pb-air Pb slope with decreasing
air Pb concentration, whereas linear models predict a constant blood Pb-air Pb slope across
all air Pb concentrations.
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Table 4-11 Summary of estimated slopes for blood Pb to air Pb relationships
in humans
Reference
Study Methods
Model Description
Blood Pb-Air
Pb Slopea
Children Populations
Brunekreefetal. (1984)
Location: Various countries
Years: 1974-1983
Subjects: Children (varying age ranges; n>190,000)
Analysis: Meta analysis of 18 studies
Model: Log-Log
Blood Pb: 5-41 pg/dL
(mean range for studies)
AirPb: 0.1-24 pg/m3
(mean range for studies)
All children:
4.6 (1.5)c
Children <20
[jg/dL: 4.8 (0.54)d
Hayes et al. (1994)
Location: Chicago, IL
Years: 1974-1988
Subjects: 0.5-6 yr (n = 9,604)
Analysis: Regression of quarterly median blood Pb
and quarterly mean air Pb
Model: Log-Log
Blood Pb: 12-30 pg/dL
(annual median range)
AirPb: 0.05-1.2 pg/m3
(annual mean range)
8.2 (0.62)e
Hilts etal. (2003)
Location: Trail, BC
Years: 1989-2001
Subjects: 0.5-6 yr (Estimated n = 220-460, based on 292-536 blood
Pb measurements/yrwith 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 pg/dL
(annual geometric mean range)
AirPb: 0.03-1.1 pg/m3
(annual geometric mean range)
6.5 (0.48)'
Ranft et al. (2008)
Location: Germany
Years: 1983-2000
Subjects: 6-11 yr (n = 843)
Analysis: Pooled regression 5 cross-sectional studies
Model: multivariate Log-Linear
Blood Pb: 2.2-13.6 pg/dL
(5th-95th percentile)
AirPb: 0.03-0.47 pg/m3
(5th-95th percentile)
3.2 (0.1)8
Schnaas et al. (2004)
Location: Mexico City
Years: 1987-2002
Subjects: 0.5-10 yr (n = 321)
Analysis: Regression of lifetime blood Pb from longitudinal
blood Pb measurements and annual average air Pb data
Model: Log-Log
Blood Pb: 5-12 pg/dL
(annual GM range)
AirPb: 0.07-2.8 pg/m3
(annual mean range in yr of birth)
2.2 (0.4)h
Schwartz and Pitcher
(1989). U.S. EPA (1986a)
Location: Chicago, IL
Years: 1976-1980
Subjects: Black children, 0-5 yr (n = 5,476)
Analysis: Chicago blood Pb screening, gasoline consumption data,
and Pb concentrations in gasoline
Model: Linear
Blood Pb: 18-27 pg/dL
(mean range)
AirPb: 0.36-1.22 pg/m3
(annual maximum quarterly mean)b
8.6 (0.75)'

Location: Mumbai, India
Years: 1984-1996
Subjects: 6-10 yr (n = 544)
Analysis: Regression of blood Pb and air Pb data
Model: Linear
Blood Pb: 8.6-14.4 pg/dL

Tripathi et al. (2001)
(regional GM range)
AirPb: 0.11-1.18 pg/m3
(regional GM range)
3.6 (0.45)1
Adult Populations
Rodrigues et al. (2010)
Location: New England, U.S.
Years: 1994-1995
Subjects: Adult bridge painters (n=84,1 female)
Analysis: Regression analysis of blood Pb and air Pb data (personal
monitors) collected during work performing various job-related tasks
Model: Log-log
Blood Pb: 16.1 pg/dL
(GM, 1,7 GSD)
Air Pb: 58 pg/m3
(GM, 2.8 GSD)
0.01 (58)k
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Mixed Child-Adult Populations
Location: U.S.	Model: Linear
Years: 1976-1980	Blood Pb: 11-18 pg/dL
Subjects: 0.5-74 yr (n = 9,987)	(mean range)	9.3(0.75)'
Analysis: NHANES blood Pb, gasoline consumption data and Pb	AirPb: 0.36-1.22 pg/rn3
concentrations in gasoline	(annual maximum quarterly mean)
Schwartz and Pitcher
(1989V U.S. EPA (1986a)
aSlope is predicted change in blood Pb (|jg/dL per |jg/m3) evaluated at ± 0.01 |jg/m3 from central estimate of air Pb for the study (shown in parentheses)
"Based on data for U.S. (1986 Pb AQCD)
c In(PbB) = In(PbA) x 0.3485 +2.853
d In(PbB) = In(PbA) x 0.2159 + 2.620
0 In(PbB) = In(PbA) x 0.24 +3.17
fPbB= PbAx6.5
gPbB =1.5 x EXP(0.9361 x (PbA-0.1 )/0.44), where 1.5 |jg/dL is the background PbB, and 0.1 |jg/m3 is the median PbAforthe study; model also adjusted for
soil Pb concentration, which may reduce estimated slope
h In(PbB) = Ln(PbA) x 0.213 +1.615 for the 1987 cohort, see text for more study details.
'PbB = PbAx 8.6
'PbB =PbAx 3.6
k In(PbB) = In(PbA) x 0.05 +2.12
' PbB = PbAx 9.63
GM, geometric mean; GSD, geometric standard deviation; PbB, blood Pb concentration (|jg/dL); PbA, air lead concentration (|jg/m3)
60
All children
<20 Mg/dL
50
^ 40
O)
=L
;? 30
Q_
"D
O
-9 20
m
0
5
10
15
20
25
AirPb (|jg/m3)
Data provided from Brunekreef (19841.
Note: The regression model is: (ln[|jg/dL blood Pb] = 0.3485 ln[|jg/m3 air Pb] + 2.85) for all children (n=96 subject groups) and
(ln[|jg/dL blood Pb] = 0.2159 ln[|jg/m3 air Pb] + 2.62) when the sample was restricted to populations that had blood Pb
concentrations <20 |jg/dL (n=44 subject groups).
Figure 4-21 Predicted relationship between air Pb and blood Pb based on a
meta analysis of 18 studies.
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0.0
0.5
1.0
AirPb (jjg/m3)
1.5
ABrunekreef (1984)
<20 |jg/dL
AHayes (1994)
~	Hilts (2003)
•	Ranft(2008)
¦ Schnaas(2004)
OSchwartz (1989)
Chicago
OSchwartz (1989)
U.S.
~ Tri path i (2001)
2.0
Note: Slopes are calculated for a change in air Pb (±0.01 pg/m ) over ranges of air Pb concentrations reported in each study (lines).
The air Pb axis is truncated at 2 pg/m3; the actual range for the Brunkreefe et al. (19841 study was 0.1-6.4 pg/dL per pg/m3 and for
the Schnaas et al. (2004) study was 0.08-2.8 pg/dL per pg/m3. The slope axis has been truncated at 40; the actual range for the
Hayes et al. (19941 study was 5-56 pg/dL per pg/m3 (the high end of the range was estimated for the minimum annual average air
Pb of 0.05 pg/m3). The two estimates for Schwartz and Pitcher (19891 represent data for U.S. and Chicago. Models are log-log (solid
lines), log-linear (dashed line), and linear (dotted lines). Symbols show the slope at the central estimate of air Pb (e.g., mean or
median reported for each study).
Figure 4-22 Blood Pb - air Pb slope (|jg/dL per |jg/m3) predicted from various
epidemiologic studies (links available in Table 4-11).
4.5.1.1 Children
Hilts et al. (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 4-12. Based
on these data, blood Pb decreased 6.5 (ig/dL per 1 |ag/nr 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). Several uncertainties apply to 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
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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 4-12 are likely to overestimate the influence of air, dust, or soil Pb on blood Pb
concentrations at this site.
Table 4-12
Environmental Pb levels and blood Pb levels in children in Trail,

British Columbia



Date
1996
1999
2001
Regression Coefficient
(|jg/dL per |jg/m3)
Blood Pb (|jg/dL)
11.5
5.9
4.7
NA
Air Pb (|jg/m3)
1.1
0.3
0.03
6.5 + 0.52 (R2=0.99, p=0.050)
Soil Pb (mg/kg)
844
756
750
0.069 ±0.008 (R2=0.99, p=0.069)
Interior Dust Pb (mg/kg)
758
583
580
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.
Source: Data from Hilts et al. (2003).
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. Pb was measured in PMi0 samples collected 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. Air and soil Pb concentrations were assigned to each participant by
spatial interpolation from the sampling grid data to each home residence. The 5th-95th
percentile ranges were 0.025-0.465 |ag 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 95th percentile value. In a multivariate linear regression model (R2 = 0.586) 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
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Pb varied depending on the sampling data and ranged from 1.41 to 2.89, with most recent
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 j^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
geometric mean of blood Pb and air Pb among the cohort of children studied were
extracted from Figure 1 in Ranft et al. (2008) for each of the five study years. The
extracted values of the geometric mean of blood Pb and air Pb were used in regressions
employing linear and log-log fits. The linear model obtained was: PbB = (13.65/PbA) +
2.96 (R2 = 0.92); i.e., the linear regression produced a constant slope of 14 (ig/dL per
(ig/m3. The log-log model was: ln(PbB) = (0.48xln(PbA) + 2.61 (R2 = 0.91), resulting in
an inverse curve for dPbB/dPbA vs. PbA with a slope of 22 (ig/dL per (ig/m3 at PbA =
0.1 (ig/m3.
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 in PM10 (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
and were assigned the corresponding annual average air Pb concentrations for their year
of birth and appropriate air monitoring zones. Associations between lifetime (across the
first 10 years of life) blood Pb concentration, air Pb concentration for year of birth and
other variables (e.g., age, year of birth, family use of glazed pottery) were evaluated
using multivariate regression models. The largest slope occurred in the cohort born in
1987, who experienced the largest decline in air Pb (from 2.8 to <0.1 (.ig/m3): the
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predicted slope for this group of children was 0.213 (95% CI: 0.114-0.312) In [jj.g/dL
blood] per ln|(.ig/m3 air]. This slope corresponds to an increase of 2.2 (ig/dL blood Pb per
(ig/m3 at the median annual air Pb concentration of 0.4 (ig/m3 estimated over the years of
the study. Slopes for other birth cohorts ranged from -0.003 (1992) to 0.166 (1988) with a
median of 0.153 for all six cohorts. This median slope corresponds to an increase of
1.7 (ig/dL blood Pb per (ig/m3 at the median annual air Pb concentration of 0.4 (ig/m3
estimated over the years of the study. Considering all 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 and a log-linear model as employeed
by the authors. The linear model obtained was: PbB = (2.50/PbA) + 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 j^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: ln(PbB) = (0.26xln(PbA)) + 2.20 (R2 = 0.94),
resulting in an inverse curve for dPbB/dPbA vs. PbA, with a slope of 4.5 (ig/dL per
(ig/m3 at PbA = 0.4 (ig/m3. The log-linear model as employed by the authors was:
ln(PbB) = (0.32xPbA) + 1.73 (R2 = 0.77); and described the data least well. The log-
linear model produced an exponential curve of dPbB/dPbA vs. PbA, with a slope of
2.04 (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).
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. children 6 months to 74 years of age (n = 9,996) during
1976-1980, and the City of Chicago blood Pb screening program provided approximately
7,000 blood Pb measurements in black children during 1976-1980. Gasoline Pb
consumption was estimated as the product of monthly gasoline sales in the U.S. and
quarterly estimates of Pb concentrations in gasoline reported to U.S. EPA. Based on the
NHANES blood Pb data for white children, the regression coefficient was 2.14 (ig/dL
blood per 100 metric tons of gasoline Pb/day (SE=0.19, p=0.0000); results for black
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children were essentially identical. Based on the Chicago blood Pb data the regression
coefficient was 16.12 (|ig/dL per 1,000 metric tons gasoline Pb/quarter (SE=1.37,
p=0.0001), which is roughly equivalent to 1.79 (ig/dL blood per 100 metric tons of
gasoline Pb/day (the value cited in Schwartz and Pitcher (1989) is 1.97 (ig/dL blood per
100 metric tons of gasoline Pb/day). U.S. EPA (1986a) reported data on gasoline Pb
consumption (sales) and ambient Pb levels in the U.S. during the period 1976-1984
(Table 4-13). 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).
Table 4-13 U.S. gasoline Pb consumption and air Pb levels
Date
Total Gasoline Pb
(103 metric tons/yr)
Total Gasoline Pb
(102 metric tons/day)a
Air Pb
(Ug/m3)
1976
171.4
4.70
1.22
1977
168.9
4.63
1.20
1978
153
4.19
1.13
1979
129
3.53
0.74
1980
78.8
2.16
0.66
1981
60.7
1.66
0.51
1982
59.9
1.64
0.53
1983
52.3
1.43
0.40
1984
46
1.26
0.36
The linear regression coefficient is 0.23 |jg/mJ air per 100 metric tons/day (SE= 0.020, R/= 0.95, p <0.0001).
Conversion factor is 10/365 days/year.
Source: U.S. EPA(1986a).
Tripathi et al. (2001) reported child blood Pb and air Pb trends for the city and suburbs of
Mumbai, India over the period 1984-1996. 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, residential soil, interior dust, and blood for three sampling
periods are summarized in Table 4-14. Based on these data, blood Pb increased 3.6 (ig/dL
per 1 (ig/m3 air Pb (based on linear regression with air or soil 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. The regression estimates were based on group mean blood Pb
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estimates for the 13 sampling locations, rather than on the individual blood Pb estimates,
which included repeated measures on an unreported fraction of the sample.
Table 4-14 Air Pb levels and blood Pb levels in children in Mumbai, India


Blood Pb (jig/dL)

Air Pb (jig/m3)
Location
N
GM
GSD
N
GM
GSD
Borivilli
12
10.4
1.67
10
0.32
1.51
Byculla
117
11.0
1.99
30
0.99
1.73
Deonar
46
9.5
2.29
93
0.11
3.21
Goregaon
21
9.1
1.30
24
0.35
1.77
Govandi
20
8.9
1.42
10
0.10
1.52
Joqeshwari
20
8.6
1.32
24
0.11
2.47
Khar
17
9.0
1.53
22
0.18
3.15
Parel
168
10.4
1.91
37
0.44
1.48
Sion
34
9.6
1.49
96
0.39
1.75
Thans(SS)
37
12.0
1.86
4
1.18
1.04
Vile Parle
19
9.1
1.46
7
0.37
1.34
Colaba
12
9.2
1.86
9
0.14
1.63
Vakola
21
L
C
^r
1.64
7
1.12
1.12
The linear regression coefficient is 3.62 |jg/dL blood per |jg/mJ air (SE= 0.61, R/= 0.76, p <0.001).
GM, geometric mean; GSD, geometric standard deviation; N, number of subjects.
Source: Data are from Tripathi et al. (2001).
3	Hayes et al. (1994) analyzed data collected as part of the Chicago, IL blood Pb screening
4	program for the period 1974-1988, following the phase-out of leaded gasoline. The data
5	included 9,604 blood Pb measurements in children (age: 6 months to 6 years) and
6	quarterly average air Pb concentrations measured at 12 monitoring stations in Cook
7	County, IL. Quarterly median blood Pb levels declined in association with quarterly mean
8	air Pb concentrations. The regression model predicted a slope of 0.24 In [jj.g/dL blood]
9	per ln||_ig/m3 air], as illustrated in Figure 4-23. This slope corresponds to an increase of
10	8.2 (ig/dL blood Pb per (ig/m3 at the average annual mean air Pb concentration of
11	0.62 (ig/m3.
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_i
T3
O)
=L
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.Q
CL
"D
O
O
GO
0.0 0.2 0.4 0.6 0.8 1.0 1.2 1.4 1.6
AirPb (|jg/m3)
Modified from Hayes et al. (19941.
Note: The regression model is: (ln[|jg/dL blood Pb] = 0.24 ln[|jg/m3 air Pb] + 3.17).
Figure 4-23 Predicted relationship between air Pb and blood Pb based on data
from Chicago, IL in children age 0-5 y (1974-1988).
4.5.1.2 Adults
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 (1 female) who were observed
during a 2-week period in 1994 or 1995. Subjects wore personal inhalable PM samplers
designed to capture PM smaller than 100 |_im. while performing various job-related tasks.
The geometric mean air Pb concentration for the 2-week period was 58 (.ig/rn1 (GSD 2.8),
with a maximum daily value of 210 |ag/nr. These Pb concentrations were reported to
have been corrected by the National Institute for Occupational Safety and Health
(NIOSH) respirator protection factors, which were not reported by the authors. Hand
wipe samples were collected at the mid-shift break and at the end of the shift (after the
subjects had reportedly cleaned up for the day; GM = 793 (j.g, GSD 3.7). Blood Pb
samples were collected at the beginning of the 2-week period (GM =16.1 (ig/dL, GSD
1.7). Associations between exposure variables and blood Pb concentrations were
explored with multivariate regression models (Table 4-15). When the model excluded
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hand-wipe data (not all participants who wore the personal air samplers agreed to provide
hand-wipes), the regression coefficient for the relationship between ln[blood Pb
concentration (|ig/dL)] and ln[air Pb (|_ig/m3)| was 0.11 (SE = 0.05, p = 0.03). This slope
corresponds to a 1.3-fold increase in blood Pb concentration for a 10-fold increase in air
Pb concentration and a slope of 0.009 (ig/dL per |_ig/m3. at the average occupational air
Pb concentration for the study (58 (.ig/ni3): non-occupational exposures were not included
in the slope calculation. 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
1.12-fold increase in blood Pb concentration per 10-fold increase in air Pb concentration
and a slope of 0.02 (ig/dL per (ig/m3, at the average occupational air Pb concentration for
the study (58 |_ig/m3).
Table 4-15 Significant predictors of blood Pb concentration in bridge painters
Parameters
Blood Pb (Air Only)
P(SE) p-value
Blood Pb (Air and Hand Wipe)
P(SE) p-value
Intercept
1.90 (0.24)
<0.0001
2.12(0.44)
0.0007
Time of blood Pb (end vs start of study)
0.16(0.04)
<0.0001
-0.31 (0.11)
0.005
Mean air Pb (pg/mJ)
0.11 (0.05)
0.03
0.05 (0.07)
0.45
Hand wipe at break (pg Pb)
—

0.007 (0.06)
0.91
Hand wipe at break * time of blood Pb
—

0.07 (0.01)
<0.0001
Months on bridge painting crews
0.001 (0.0004)
0.03
0.001 (0.0006)
0.04
Education
< High school
> High school
0.38 (0.10)
Reference
0.0002
0.29 (0.13)
Reference
0.03
Respirator fit test
No
Yes
-0.14(0.14)
Reference
0.32
-0.13(0.21)
Reference
0.53
Respirator fit test * time of blood Pb
No
Yes
0.18(0.06)
Reference
0.003
0.17(0.07)
Reference
0.01
Smoke on site
No
Yes
0.14 (0.09) Reference
0.14
0.15(0.10)
Reference
0.14
Smoke on site * time of blood Pb
No
Yes
-0.15(0.05)
Reference
0.002
-0.11 (0.04)
Reference
0.009
Personal hygiene index
Low
High
0.27 (0.11)
Reference
0.02
0.29 (0.12)
Reference
0.02
Site-level variables
Containment facility
Poor
Good
-0.59 (0.18)
Reference
0.001
-0.57 (0.22)
Reference
0.01
Air Pb, hand wipe, and blood Pb levels are natural log-transformed.
Blood Pb concentration in units of |jg/dL.
Source: Data from Rodrigues et al. (2010).
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4.5.2
Environmental Lead-Blood Lead 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 et al.. 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 4-2
PbB = E x AF x USF
Equation 4-3
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 ti/2 for elimination of Pb in blood).
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 j^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. 2006^.
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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; Bomschein 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 4-16). The
regression coefficient for the relationship between ln[blood Pb concentration (|_ig/dL)|
and ln[floor dust Pb concentration (jj.g/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 4-16 Linear model relating environmental Pb exposure and blood Pb
concentration in children
Variables
Overall p-value
Levelsa
Estimate (SE)
p-Value
Intercept
0.172

-0.517 (0.373)
0.172
Age (in yr)
< 0.001
Age
Age 2
Age 3
Age 4
2.620 (0.628)
-1.353 (0.354)
0.273 (0.083)
-0.019(0.007)
<	0.001
<	0.001
0.002
0.008
Yr of construction
0.014
Intercept for missing
1990—present
1978-1989
1960-1977
-0.121 (0.052)
-0.198 (0.058)
-0.196 (0.060)
-0.174 (0.056)
0.024
0.001
0.002
0.003


1950-1959
-0.207 (0.065)
0.003


1940-1949
Before 1940
-0.012 (0.072)
0.000
0.870
PIR
< 0.001
Intercept for missing
Slope
0.053 (0.065
-0.053 (0.012)
0.420
< 0.001
Race/ethnicity
< 0.001
Non-Hispanic white
Non-Hispanic black
Hispanic
Other
0.000
0.247 (0.035
-0.035 (0.030)
0.128 (0.070)
< 0.001
0.251
0.073
Country of birth
0.002
Missing
U.S.b
Mexico
-0.077 (0.219)
0.000
0.353 (0.097)
0.728
< 0.001


Elsewhere
0.154 (0.121
0.209
Floor surface/condition x log floor
PbD
< 0.001
Intercept for missing
Not smooth and cleanable
0.178 (0.094)
0.386 (0.089)
0.065
< 0.001

Smooth and cleanable or carpeted
0.205 (0.032)
< 0.001
Floor surface/condition

Not smooth and cleanable
0.023 (0.015)
0.124
x log floor PbD)2

Smooth and cleanable or carpeted
0.027 (0.008)
0.001
Floor surface/condition

Uncarpeted not smooth and
cleanable
Smooth and cleanable or carpeted
-0.020 (0.014)
0.159
x (log floor PbD)3

-0.009 (0.004)
0.012
Log windowsill PbD
0.002
Intercept for missing
Slope
0.053 (0.040
0.041 (0.011
0.186
< 0.001


Intercept for missing
Mobile home or trailer
-0.064 (0.097
0.127 (0.067)
0.511
0.066
Home-apartment type
< 0.001
One family house, detached
One family house, attached
-0.025 (0.046)
0.000
0.596


Apartment (1-9 units)
Apartment (> 10 units)
0.069 (0.060)
-0.133 (0.056)
0.256
0.022
Anyone smoke inside the home
0.015
Missing
Yes
No
0.138 (0.140)
0.100 (0.040)
0.000
0.331
0.015
Log cotinine concentration (ng/dL)
0.004
Intercept for missing
Slope
-0.150 (0.063)
0.039 (0.012)
0.023
0.002
Window, cabinet, or wall renovation
in a pre-1978 home
0.045
Missing
Yes
No
-0.008 (0.061)
0.097 (0.047)
0.000
0.896
0.045
'Children: n = 2,155 (age 10-60 months); R* = 40%
'includes the 50 states and the District of Columbia
Source: Dixon et al. (2009).
1	Mielke et al. (2007a') analyzed blood Pb and soil Pb concentration data collected as part
2	of a blood Pb screening program in New Orleans (2000-2005). The data set included
3	55,551 blood Pb measurements for children 0-6 years of age and 5,467 soil Pb
4	measurements. Blood Pb and soil Pb concentrations were matched at the level of census
5	tracts. The association between blood Pb concentration and soil Pb concentration was
6	evaluated using non-parametric permutation methods. The resulting best-fit model
7	(R2=0.528) was:
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PbB = 2.038 +(0.172 xPbS05)
Equation 4-4
where PbB is the median blood Pb concentration and PbS is the median soil Pb
concentration. 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
4-24).
10
Blood=2.038+0.172 x Soil0 5
	i	i	i	i	i	i	i	i	i	i	i	i	i	i	i	i	i	i	i	1
0	400 800 1200 1600 2000
Soil Pb (ppm)
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 CMielke et al.. 2007a1.
Figure 4-24 Predicted relationship between soil Pb concentration and blood
Pb concentration in children based on data collected in the New
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 et al.. 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 Pb" 5), 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
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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 j^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.
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, a former Pb mining and
smelting site, was provided in the 2006 Pb AQCD. 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 4-17. 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 4-17). 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 y old children living near one of four industrial sites was given as PbB =
0.24381n(PbS} + 0.2758.
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Table 4-17 General linear model relating blood Pb concentration in children and
environmental Pb levels—Bunker Hill Superfund Site
Parameter
Coefficient
P-value
Standardized
Coefficient
Intercept
-0.1801
0.7916
0.00000
Age (yr)
-0.4075
<0.0001
-0.2497
ln(interior dust Pb); (mg/kg)
0.7288
<0.0001
0.1515
Infyard soil Pb); (mg/kg)
0.2555
0.0002
0.0777
GM soil Pb within 200 ft of residence (mg/kg)
0.0008
<0.0001
0.1380
GM community soil Pb (mg/kg)
0.0018
<0.0001
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., hand-to-mouth activity,
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[jj.g/g soil Pb] and a
slope of 0.45 (ig/dL blood Pb per ln||_ig/ft2| dust Pb loading.
4.6 Biokinetic Models of Lead Exposure-Blood Lead
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. They are also critical for assessing the
potential impacts of public health programs directed at mitigation of Pb exposure or of
remediation of contaminated sites.
However, they 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 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, large uncertainty can be expected to remain about many, or even
most, 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 is 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'Flahcrtv. 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).
4.7 Summary and Conclusions
4.7.1 Exposure
Exposure data considered in this assessment build upon the conclusions of the 2006 Pb
AQCD (2006b). which found air Pb concentrations in the U.S. and associated biomarkers
of exposure to Pb have decreased substantially following the ban 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). 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 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 3.4. Additionally, high spatial variability of the Pb concentrations in
various media also can contribute to exposure error, as described in the 2009 PM ISA
("U.S. EPA. 2009). Models, such as the Integrated Exposure Uptake Biokinetic (IEUBK)
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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 4.1 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 shoe tracking and hand-to-mouth
contact, which is the main pathway of childhood exposure to Pb. In general, soil Pb
concentrations tended to be higher within inner-city communities compared with
neighborhoods surrounding the city. Recent data by Yamamoto et al. (2006) have shown
that the size distribution of particles collected on children's hands have a mode around
40 (.un with the upper tail of the distribution extending to 200-300 |_im. Infiltration of Pb
dust into indoor environments has been demonstrated, and Pb dust has been shown to
persist in indoor environments even after repeated cleanings. Measurements of particle-
bound Pb exposures reported in this assessment have shown that personal exposure
measurements for Pb concentration are typically higher than indoor or outdoor ambient
Pb concentrations. These findings may be related to local resuspension with body
movement.
4.7.2 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 toxicologically 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 (ig/day/kg) 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
accumulates in bone regions undergoing the most active calcification at the time of
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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 andWittmers. 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 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.
4.7.3 Lead Biomarkers
Overall, trends in blood Pb levels have been decreasing among U.S. children and adults
over the past 20 years (Section 4.4). The median blood Pb level for the entire U.S.
population is 1.2 (ig/dL and the 95th percentile blood Pb level was 3.7 (ig/dL, based on
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the 2007-2008 NHANES data (NC'HS. 2010). Among children aged 1-5 years, the
median and 95th percentiles were slightly higher at 1.4 (ig/dL and 4.1 |ig/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).
4.7.4 Air Lead-Blood Lead Relationships
The 1986 Pb AQCD described epidemiological studies of relationships between air Pb
and blood Pb. Much of the pertinent earlier literature described in the 1986 Pb AQCD
was drawn from a meta-analysis by Brunekreef (1984). Based on the studies available at
that time that considered multiple air-related Pb exposure pathways in the aggregate, 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 (ig/dL) and high air Pb levels
(much greater than 10 |ig/m3) prevailed relative to lower environmental exposures which
showed lower blood Pb and air Pb concentrations (<30 (ig/dL and <3 |ig/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 (ig/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 2-7). The range
of estimates from these studies is 2-9 (ig/dL per (ig/m3, which encompasses the estimate
from the Brunekreef (1984) meta-analysis of (3-6 (ig/dL per (.ig/ni3). 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 across all 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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Wetmur. JG. (1994). Influence of the common human delta-aminolevulinate dehydratase polymorphism on lead
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CHAPTER 5 INTEGRATED HEALTH EFFECTS
OF LEAD EXPOSURE
5.1 Introduction
This chapter reviews, summarizes, and integrates the evidence for the broad spectrum of
health effects associated with exposure to Pb. The chapter begins (Section 5.2) 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 for particular endpoints. Subsequent
sections comprise evaluations of the epidemiologic and toxicological evidence for the
health effects of Pb exposure on major outcome categories such as nervous system effects
(Section 5.3), cardiovascular effects (Section 5.4), renal effects (Section 5.5), immune
effects (Section 5.6), effects on heme synthesis and red blood cell function (Section 5.7),
and reproductive and developmental effects (Section 5.8). Section 5.9 provides reviews
of 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 5.9.1), gastrointestinal system (Section 5.9.2), endocrine system (Section 5.9.3),
bone and teeth (Section 5.9.4), ocular health (Section 5.9.5), and respiratory system
(Section 5.9.6). Chapter 5 concludes with a discussion of the evidence for Pb effects on
cancer (Section 5.10).
Individual sections for major outcome categories (e.g., nervous system, cardiovascular,
renal) begin with a brief summary of conclusions from the 2006 Pb AQCD followed by
an evaluation of recent evidence that is intended to build upon evidence from previous
reviews. Within each of these sections, results are organized by endpoint (e.g., cognitive
function, behavior, neurodegenerative diseases) then by specific scientific discipline
(i.e., epidemiology, toxicology). Sections for each of the major outcome categories
(e.g., nervous system, cardiovascular, renal effects) conclude with an integrated summary
of the evaluation of evidence and a conclusion regarding causality. Based upon the
framework (described in the Preamble), a determination of causality was made for a
broad outcome category (i.e., nervous system effects) by evaluating the coherence of
evidence across disciplines and across a spectrum of related endpoints. Each discussion
leading up to the causal determination characterizes the evidence on which the causal
judgment was based, including the strength of evidence for the individual endpoints
within the major outcome category.
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5.2 Modes of Action
5.2.1 Introduction
The diverse health effects of Pb are dependent on multiple factors, including the
concentration and duration of exposure, the particular Pb compounds constituting the
exposure, and which tissues are affected. 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 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 5-1). This section draws information from all of the subsequent
health effects sections in Chapter 5, 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 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 older
studies (which represent the current state of the science) on the possible modes of action
of Pb.
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Altered Ion
Status
(5.2,2)
Exposure
Protein
Binding
(5,2,3)
Cell Death
Genotoxicity
(5.2,?)
Endocrine
Disruption
(5.2,6)
Inflammation
{5.2.5)
Oxidative
Stress
(5.2,4)
Note: The subsections where these MOAs are discussed are indicated in parentheses.
Figure 5-1 Schematic representation of the relationships between the
various MOAs by which Pb exerts its health effects.
5.2.2 Altered Ion Status
Physiologically-relevant metal ions (e.g., Ca, 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 as structural elements in the proper maintenance
of tertiary protein conformations (Garza et al.. 2006). It is through disruption of these
biological functions that Pb effects its negative actions, ultimately interfering with such
tightly regulated processes as cell signaling, intracellular ion homeostasis, ion transport,
energy metabolism, and enzymatic function.
5.2.2.1 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
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.
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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 can exist in a microenvironment that is high
in both Ca2+ and Pb concentrations, potentially increasing their relative susceptibility to
Pb-induced toxicity (Long et al.. 1992). A series of studies from the laboratory of Long,
Dowd, and Rosen have indicated that exposure of cultured osteoblastic bone cells to Pb
disrupts intracellular Ca2+ levels (|Ca2 |,). Exposure of osteoblasts to 1, 5, or 25 (.iM Pb
for 40-300 minutes resulted in prolonged increases in |Ca2 | of 36, 50 and 120% over
baseline, respectively (Schanne et al.. 1997; Schanne etal.. 1989). Long et al. (1992)
observed that exposure of osteoblasts to either 400 ng parathyroid hormone (PTH)/mL
culture for 1 hour or 25 |_iM Pb for 20 hours increased |Ca2 |,. Pretreatment of Pb-
exposed cells with PTH increased | Ca2 | above concentrations observed in either single
exposure, indicating that Pb may disrupt the ability of bone cells to respond to normal
hormonal control. A similar additive increase in |Ca2 | was also observed when bone
cells were co-treated with epidermal growth factor (EGF) plus Pb, versus Pb alone (Long
and Rosen. 1992). Pb-induced increases in | Ca2 | were blocked by a protein kinase C
(PKC) inhibitor, indicating that PKC activation may serve as the mechanism by which Pb
perturbs |Ca2 |, (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
(Ouintanar-Escorza et al.. 2010; Ouintanar-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 | was approximately 2.5-fold higher than that seen in nonexposed
workers (mean [SD] blood Pb level: 9.9 [2] (.ig/dL) (Ouintanar-Escorza et al.. 2007). The
increase in |Ca2 | was associated with higher osmotic fragility and modifications in
erythrocyte shape. In a separate investigation, erythrocytes from 10 healthy volunteers
were exposed to Pb at concentrations of 0.2 to 6.0 (.iM for 24 or 120 hours, concentration-
related increases in | Ca2 | were observed across all concentrations for both durations of
exposure (Ouintanar-Escorza et al.. 2010). Subsequent exposures of erythrocytes to either
0.4 or 4.0 (.iM Pb [corresponding to 10 or 80 (ig/dL in exposed workers (Ouintanar-
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 | 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 | in vitro,
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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 | in human platelets. The observed increase in Ca2+ levels was attributed to the
increased influx of external Ca2+, possibly through receptor-operated Ca2+ channels. In
mouse splenic lymphocytes, 1 (.iM Pb was the lowest exposure concentration found to
increase [Ca2+]i, with incubation periods of 10 minutes or greater (Li et al.. 2008c). These
increases in Ca2+ appeared to be reversible as | Ca2 | 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 in lymphocytes. In rat
tail arteries exposed to 1.2 (.iM Pb-acetate for 1 hour, intracellular stores of Ca2+ increased
over controls, possibly through increased transmembrane influx of Ca2+ (Piccinini et al..
1977).
Exposure of the microsomal fraction of rat brain cells to as little as 0.25 (.iM Pb for
2	minutes resulted in increased release of Ca2+ into the media (Pentvala et al.. 2010).
Further, Pb exposure also decreased the activity of the microsomal Ca2+-ATPase, thus
decreasing the sequestration of Ca2+ into microsomes. The results of this study suggest
that 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 | 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 | 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.
5.2.2.2 Disruption of Ion Transport Mechanisms
As described above, deregulation of Ca2+ homeostasis results in negative effects in
multiple organ systems. Under normal conditions in the life cycle of most cells, cytosolic
concentrations of free Ca2+ fluctuate around 100-200 nM and Ca2+that has entered the
cell must be removed in order to maintain normal homeostatic concentrations (Carafoli.
2005) . An important component of the maintenance of Ca2+ homeostasis is
transmembrane transport of Ca ions via Ca2+-ATPase and voltage-sensitive gates
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(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 the effects of Pb exposure on Na+-K+-ATPase, Ca2+-
ATPase, and Mg2+-ATPases 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] j^ig/dL) (kharoubi et
al.. 2008a) and in erythrocytes of rats exposed to 0.2% Pb in drinking water for 5 weeks
(mean [SD] blood Pb level: 97.56 [11.8] j^ig/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 synaptosomes isolated from rats
exposed to 200 ppm Pb in drinking water for 3 months (blood Pb level: 37.8 (ig/dL)
(Rafalowska et al.. 1996) or 15 mg/kg Pb injected (i.p.) for 7 days (blood Pb level:
112.5 (.ig/dL) (Struzvnska et al.. 1997a). Inhibition of Na -k -ATPase activity was also
observed in primary cerebellar granule neuronal cultures obtained from rats pre- and
postnatally exposed to Pb (0.1 % Pb-acetate in dams drinking water, resulting in blood Pb
level of 4 (.ig/dL) (Baranowska-Bosiacka et al. 201 la). The activity of Ca -ATPase in
the sarcoplasmic reticulum of rabbits exposed to 0.01 (.iM Pb was similarly decreased
(Hechtenberg and Beversmann. 1991). The inhibitory effect of Pb was diminished in the
presence of high MgATP concentrations. The activity of generic ATPase was reported to
be altered in the testes of rats exposed to 300 ppm Pb-acetate gestationally, and in
drinking water after weaning to the age of 6, 8, 10, or 12 weeks (Liu et al.. 2008). In
pregnant rats fed a Pb-depleted (20 ± 5 j^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 (Ederetal.. 1990). However, the offspring
(exposed via placental and lactational transfer of Pb) of Pb-depleted rats displayed
decreased activities in both enzymes compared with offspring of rats with higher Pb
exposures. A similar increase in the Na+-K+-ATPase activity was observed in rats
exposed (i.p.) to 20 mg/kg Pb for 14 consecutive days (Jehan and Motlag. 1995). Co-
exposure of Pb with zinc and copper 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).
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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. (1999b)
observed that l-5xl04 (.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.. 1984V Further study by this group (Calderon-Salinas et al..
1999a) demonstrated that Pb can noncompetitively block the transport of Ca2+ by
inhibiting the activity of Ca2+-Mg2+-ATPase at concentrations of 1-5x103 (.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 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 j^ig/dL Pb was the lowest 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.
In a study investigating ATPase activities in occupationally-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] - 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 magnesium
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 cord blood Ca2+ pump activities.
Newborn hair Pb (geometric mean [5th, 95th percentile]: 1.1 [0.1-8.0] jJ.g/g) was
negatively and significantly associated with both maternal and cord blood Ca2+ pump
activities. In a population of 81 newborns, Huel et al. (2008) found that newborn hair and
cord blood Pb levels (mean [SD] newborn hair Pb and blood Pb levels: 1.22 [1.41] jj.g/g
and 3.54 [1.72] (ig/dL) were negatively associated with Ca2+-ATPase activity in plasma
membranes of erythrocytes isolated from cord blood; newborn hair Pb levels were more
strongly associated with cord Ca2+ pump activity than were cord blood Pb levels.
Pb has also been shown to disrupt cation transport mechanisms through direct action on
voltage-sensitive cation channels. Audesirk and Audesirk (1993. 1991) demonstrated that
extracellular free Pb inhibits the action of multiple voltage-sensitive Ca2+ channels, with
free Pb IC50 (half maximal inhibitory concentration) values of 0.7 (.iM for L-type
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channels and 1.3 |_iM for T-type channels in neuroblastoma cells, 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 Ca2+ channels,
demonstrating an IC50 value of 0.3 (.iM in adrenal chromaffin cells. The observed
disruption of the 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, intracellular Pb was observed to enhance Ca2+ currents through attenuation of
the Ca2+ dependent deactivation of Ca2+ channels at an EC50 value of 200 (.iM. 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 PC 12 cell lines) through store-operated Ca2+
channels (Chin et al.. 2009; Chang et al.. 2008b). In particular, the Orail-STIMl complex
was shown to be critical in the entry of Pb ions into cells, and increased Pb permeation
was directly related to decreased | Ca2 | concentrations at exposure concentrations as low
as 0.1 (iM.
Pb also disrupts the action of Ca2+-dependent potassium channels. Alvarez et al. (1986)
observed that Pb promoted the efflux of potassium from inside-out erythrocyte vesicles in
a concentration-dependent manner at concentrations of 1-300 (.iM. either through action
on a Mg modulatory site or through direct interaction with the Ca2+ binding site. Fehlau
et al. (1989) also demonstrated Pb-induced activation of the potassium channel in
erythrocytes. However, Pb only activated the potassium channels at concentrations below
10 (.iM: higher concentrations of Pb completely inhibited channel activity, indicating the
modulation of potassium permeability is due to alterations in channel gating. Silken et al.
(2001) observed that Pb activated potassium channels in erythrocytes from the marine
teleost Scorpaena porcus in a concentration-dependent manner after a 20-minute
incubation; minor loss of potassium was seen at Pb concentrations of 1-2 (.iM. whereas
exposure to 20-50 (.iM Pb resulted in approximately 70% potassium loss. Competitive
and inhibitory binding assays suggest that Pb directly activates potassium 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-sensitive channels in in vitro experiments (Cooper and Manalis.
1984; Suszkiw et al.. 1984). However, in these same experiments, concentrations of Pb as
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5 (.iM and higher 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 (1995. 1993) reported that
Pb exposure induced the release of norepinephrine (NE) from bovine adrenal chromaffin
cells, and was considerably more potent at doing so than was Ca2+ (K0 5 of 4.6xl03 (.iM for
Pb versus 2.4 (.iM for Ca2+). Activation of PKC was observed to enhance the Pb-induced
release of NE. Westerrink 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 3xl04 (.iM in permeabilized cells. The suppression of Ca2+-
induced evoked release of neurotransmitters combined with the ability of Pb to enhance
spontaneous releases could result in higher noise in the synaptic transmission of nerve
impulses in Pb-exposed animals. In rats exposed to Pb at concentrations of 0.1-1.0% in
drinking water beginning at gestational day (GD)15-16 and continuing to 120 days
postnatal, decreases in total potassium-stimulated hippocampal GABA release were seen
at exposure levels of 0.1-0.5% (range of mean [SD] blood Pb levels: 26.8 [1.3] - 61.8
[2.9] (.ig/dL) (Laslev and Gilbert. 2002). Maximal effects were observed at 0.2% Pb in
drinking water, but effects were less evident at 0.5%, and were absent at 1.0%. In the
absence of Ca2+, potassium-induced GABA release was increased with the two highest
exposure concentrations, suggesting a Pb-induced enhancement of evoked release of
GABA. The authors suggest that this pattern of response indicates that Pb is a potent
suppressor of evoked release at low concentrations, but a Ca2+ mimic in regard to
independently evoking exocytosis and 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-sensitive Ca2+ channels.
5.2.2.3 Displacement of Metal Ions and Perturbed Protein
Function
The binding of metal ions to proteins causes specific changes in protein shape, and the
specific cellular function of many proteins may be altered by conformational changes
(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 et al.
2006). The coordination chemistry that normally regulates metal-protein binding makes
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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 | 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 the cation, as well as
diacylglycerol and phospholipids, for proper cellular activity (Garza etal.. 2006).
Markovac and Goldstein (1988b) observed that, in the absence of Ca2+, exposure to
picomolar 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+ (5.5xl0~5 (.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 at
picomolar concentrations in adrenal chromaffin cells incubated with 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 greater
than one nanomolar.
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. This suggests that Pb redistributes PKC at micromolar
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 Ca nor
diacylglycerol was increased due to exposure indicates that Pb-induced activation of PKC
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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
picomolar 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 micromolar concentrations (Murakami et al.. 1993; Lison et al.. 1990).
Calmodulin is another important protein essential for proper Ca2+-dependent cell
signaling. Calmodulin contains an "EF-hand" Ca2+ binding domain, and 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 Rung. 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
incubation (minimal activation at 0.1 (.iM. EC50 = 0.5-1.0 (iM), 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 (iM). whereas Richardt et al. (1986) observed that Pb was slightly more potent
than was Ca2+ at binding calmodulin (IC50 =11 and 26 |_iM. respectively). Both studies
indicated that Pb was much more effective at binding calmodulin than was any other
metal cation investigated (e.g., mercury, cadmium, iron). 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 4-5.5xl0"4 |_iM (threshold = lxlO"4 (.iM) and 0.45-
0.5 |_iM (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 4.3xl0"4 (.iM [threshold = 3x10~4 (.iM
versus EC50 1.2xl0"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 5xl0"5 (.iM. Pb activated calcineurin, a phosphatase with
widespread distribution in the brain and immune system, at threshold concentrations as
low as 2x10~5 |_iM in the presence of Ca2+ (incubation time = 30 minutes), but inhibited its
activity at concentrations greater than 2x10~4 (.iM (Kern and Audesirk. 2000). Thus,
picomolar 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)
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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 (Dowd et al.. 1994). Pb was found to bind to osteocalcin
more than 1000-times more tightly than was Ca2+ (Kd = 1,6xl0"2 (.iM versus 7 (.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
group also found that Pb bound osteocalcin approximately 10,000-times more tightly than
did Ca2+ (Kd = 0.085 (.iM versus 1.25 x 103 (.iM. 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 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 nanomolar range (Bouton et al. 2001; Qsterode and U1 berth. 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
zinc-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 exposure to Pb results in inhibition of
the enzyme in the erythrocytes of Pb-exposed workers and adolescents (blood Pb levels
>10 (.ig/dL) (Ahamed et al.. 2006; Ademuviwa et al.. 2005b). in human erythrocytes
exposed to Pb for 60 minutes (K = 7xl04 (.iM) (Simons. 1995). and in 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.. 2005). Additional experiments indicated that lower concentrations of zinc
result in greater inhibition of enzyme activity by Pb, suggesting a competitive inhibition
between zinc and Pb at a single site (Simons. 1995).
Zinc-binding domains are also found in transcription factors and proteins necessary for
gene expression, including GATA proteins and transcription factors TFIIIA, Spl, and
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Erg-1 (Ghering et al.. 2005; Huang et al.. 2004; Rcddv and Zawia. 2000; Hanas et al..
1999; Zawia et al.. 1998). Pb was found to form tight complexes with the cysteine
residues in GATA proteins (Pb stability constant ((3pb) = 6.4 x 109M_1 for single zinc
fingers and (32Pb2 = 6.4 x 1019 M~2), and was able to displace bound zinc 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 concentration = lxlO6 (iM). Pb at a minimum
concentration of 10 (.iM also binds to the zinc domain of TFIIIA, inhibiting its ability to
bind DNA at concentrations (Huang et al.. 2004; Hanas et al. 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.
Pb exposure modulated the DNA-binding profiles of the transcription factors Sp 1 and
Erg-1 in rat pups exposed to 0.2% Pb-acetate via lactation, resulting in a shift in DNA-
binding toward early development (i.e., the first week following birth) (Reddv and Zawia.
2000; Zaw ia et al.. 1998). The shifts in Spl DNA-binding profiles were shown to be
associated with abnormal expression of genes related to myelin formation
(Section 5.2.7.5). Further mechanistic research utilizing a synthetic peptide containing a
zinc finger motif demonstrated that Pb can bind the histidine and cysteine residues of the
zinc finger motif, thus displacing zinc 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 zinc
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 Sp 1 (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 of rats exposed to 0.1% Pb in drinking water for 9 months (mean [SD] blood Pb
level: 7.01 [1.64] (ig/dL) and in human erythrocyte lysates exposed to 0.1 (.iM Pb for as
little as 5 minutes (Baranow ska-Bosiacka et al. 2009). and cGMP phosphodiesterase at
picomolar concentrations in homogenized bovine retinas (Srivastava et al.. 1995). Pb was
also reported to inhibit pyrimidine 5'-nucleotidase through competitive inhibition of
magnesium binding, resulting in conformational changes and improper amino acid
positioning in the active site (Bitto et al.. 2006).
In summary, Pb has the ability 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
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1	levels of 6.5 (ig/dL (exposure 4 weeks), in suckling rats exposed to 0.2% Pb via lactation,
2	and in cell-free and cellular in vitro experiments conducted at exposure concentrations
3	ranging from micromoles to picomoles. These alterations in protein function have
4	implications for numerous cellular and physiological processes, including cell signaling,
5	growth and differentiation, gene expression, energy metabolism, and biosynthetic
6	pathways. Table 5-1 provides a list of enzymes and proteins whose function may be
7	perturbed by Pb exposure.
Table 5-1 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

Aminolevulinic acid dehydratase
1
Zn; Simons CI9951

Ferrochelatase
1
Fe (2Fe-2S Cluster); Crooks et al. (20101

Superoxide dismutase
It
Mn. Cu. Zn. Fe: Antonvuk et al. (20091.
Borgstahl et al. (19921

Catalase
It
Fe (Heme); Putnam et al. (20001

Glutathione peroxidase
It
Se; Rotruck et al. (19731
Enzymes
Guanylate cyclase
1
Fe (Hemel: Boerriater and Burnett (20091
cGMP phosphodiesterase
1
Mg, Zn; Ke (20041

NAD synthase
1
Mg; Hara et al. (20031

NAD(P)H oxidase
t
Ca; Leseney et al. (19991

Pyrimidine 5'-nucleotidase
1
Mg, Ca; Bitto et al. (20061. Amici et al. (19971.
Paglia and Valentine (19751

Erythrocyte
phosphoribosyltransferase
1
Mg (Mn, Ca, Co, Ni, Zn); Deng et al. (20101.
Arnold and Kelley (19781

ATPase
It
Ca, Mg, Na-K; Technische Universitat
Braunschweig (20111
Ion Channels/
Transport
Mitochondrial transmembrane pore
t
Ca; He et al. (20001
Calcium-dependent potassium
channel
t
Ca: Silkin et al. (20011. Alvarez et al. (19861
Signal Transduction
Protein kinase C
It
Ca; Garza et al. (20061
Calmodulin
t
Ca; Garza et al. (20061
Pb Binding
Metallothionein
t
Zn, Cu; Yu et al. (2009)
DNA Binding
GATA transcriptional factors
1
Zn; Hanas et al. (19991. Huang et al. (20041
a| indicates increased activity; J, indicates decreased activity; indicates activity can be alternatively increased or decreased.
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5.2.2.4
Mitochondrial Abnormality
Alterations in mitochondrial function, including disruptions in ion transport,
ultrastructural changes, altered energy metabolism, and perturbed enzyme activities due
to Pb intoxication are well documented in the scientific literature. Exposure of rats to Pb
in feed (1% Pb for 4, 6, 8, 10, 12, or 20 weeks) or drinking water (300 ppm for 8 weeks,
500 ppm for 7 months, or 1% for 9 months) resulted in gross ultrastructural changes in
renal tubules 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.. 2010d; Marchlewicz et al.. 2009; Navarro-Moreno et
al.. 2009; Gover. 1968; Gover et al.. 1968).
Transmembrane mitochondrial ion transport mechanisms are 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 0.1% Pb in drinking water throughout
gestation and lactation (Baranowska-Bosiacka et al.. 201 la), astroglia incubated with 0.1
or 1.0 (iM Pb for 14 days (Legare et al.. 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 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; Lidskv and Schneider. 2003). He
et al. (2000) additionally observed cytochrome c release from mitochondria, and caspase-
9 and -3 activation following exposure of 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 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 1% Pb in feed for 6 weeks, the rate of oxygen uptake
during ADP-activated (state 3) respiration was lower compared to controls (Gover et al..
1968). The rate of ATP formation in exposed mitochondria was observed to be
approximately 50% that of control mitochondria. A decrease in state 3 respiration and
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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 (Fow ler et al.. 1980V Pb-induced decreases in ATP and
adenylate energy charge (AEC) were observed concurrently with increases in ADP,
AMP, and adenosine in rats exposed to 1% 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 0.1% Pb
in drinking water throughout gestation and lactation (Baranowska-Bosiacka et al..
2011a). One (.iM Pb (48 hours) was the lowest concentration observed to decrease
cellular ATP levels in differentiated PC-12 cells, and these changes were correlated with
a Pb-induced decrease in the expression of the 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 coupled and uncoupled
respiratory oxygen use in mitochondria, and that Pb prevented pyruvate, but not malate,
uptake. Mitochondrial levels of ATP were diminished after 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.. 2011). Contrary to
the above findings, Rafalowska et al. (1996) reported that, although ATP levels did
decrease, chronic exposure to Pb did not inhibit oxidative phosphorylation in the
synaptosomes of rats exposed to 200 ppm Pb in water for 3 months. Similar effects with
regard to the activity of the mitochondrial oxidative chain were observed in rats injected
with 15 mg/kg Pb 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 to 30 (.ig/dL Pb (Grabowska and Guminska. 1996; Dowd et
al.. 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 controls exposed to lower concentrations of Pb (blood Pb levels: 82.1 versus
39.9 (ig/dL), and suggested that Pb activated anaerobic glycolysis. In 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
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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
1% 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
is inhibited by exposure to Pb in human and rat erythrocytes (see above for concentration
and duration) (Baranowska-Bosiacka et al. 2009).
Disruptions in erythrocyte energy metabolism have been observed in workers
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). 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 increased blood Pb level. The blood Pb level associated with
50% inhibition of NAD synthetase, which requires a magnesium cation for activity (Hara
et al.. 2003). was 43 (ig/dL.
Altered Heme Synthesis
Exposure to Pb is known to inhibit two key steps in the synthesis of heme:
porphobilinogen synthase (i.e., S-aminolevulinic acid dehydratase), a cytoplasmic
enzyme requiring zinc for enzymatic activity that condenses two molecules of
aminolevulinic acid into porphobilinogen, and ferrochelatase, a mitochondrial iron-sulfur
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 for thiol groups at its 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 zinc 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 (average blood Pb levels ranging from 21.92 to 53.63 (ig/dL) (Mohammad et al.
2008; Counter et al.. 2007; Patil et al.. 2006b; Ademuviwa et al.. 2005b) and animals
(blood Pb level: 24.7 (.ig/dL) (Rendon-Ramirez et al.. 2007).
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5.2.3
Protein Binding
Evidence indicates that Pb binds to proteins within cells through interactions with side
group moieties (e.g., thiol residues) and can potentially disrupt cellular function
(Sections 5.2.2.3 and 5.2.2.4). 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): Pb intoxication in children autopsy cases
was observed to induce the formation of intranuclear inclusion bodies in the liver and
kidney. Since that time, further research has been conducted into characterizing the
composition of intranuclear inclusion bodies and identifying specific Pb-binding proteins.
5.2.3.1 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 1% 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 (Navarro-Moreno et al.. 2009; Qskarsson
and Fowler. 1985; Fowler et al.. 1980; McLachlin et al.. 1980; Choie and Richter. 1972;
Goveret al.. 1970a; Gove ret al.. 1970b). 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).
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 kDa protein that migrates as a single band on acrylamide 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 1% Pb-acetate in feed or 0.75% 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 formation of
intranuclear inclusion bodies required protein synthesis (McLachlin et al. 1980; Choie et
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al.. 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 0.2-1.0% 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.
5.2.3.2 Cytosolic Lead Binding Proteins
Numerous studies have also identified cytosolic Pb-binding proteins. Two binding
proteins, with molecular weights of 11.5 and 63 kDa, were identified by (Qskarsson et
al.. 1982) in the kidney postmitochondrial cytosolic fraction after injection with 50 mg
Pb. 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 zinc cation to ALAD (Goering and Fowler. 1985. 1984).
Cadmium and zinc, 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 cadmium 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 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 et al.. 1998; Fowler and DuVal. 1991).
Cytosolic Pb-binding proteins distinct from kidney proteins have also been identified in
the brain of exposed rats and human brain homogenates exposed in vitro (Ouintanilla-
Vega et al.. 1995; DuVal and Fowler. 1989; Goering etal.. 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 zinc to ALAD. Cytosolic Pb-binding
proteins have been shown to be high in glutamic acid, aspartic acid, and cysteine residues
(Fowler etal.. 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).
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5.2.3.3
Erythrocytic Lead 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 Q'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, increases after
exposure in both exposed humans (blood Pb = 30-75 j^ig/dL) and rats (Pb exposure = 25
mM in drinking water) (Boudene et al.. 1984; Fuiita et al.. 1982; Fuiita et al.. 1981).
ALAD is a polymorphic gene with three isoforms: 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 et al.. 1995a; Wetmur. 1994; Wetmur et
al.. 1991a; Astrin et al. 1987V Some newer 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. 2006V
The ALAD-2 protein binds Pb more tightly than the ALAD-lform: in 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 etal.. 1997a). This higher affinity for Pb in ALAD-2
carriers may sequester Pb and prevent its bioavailability for reaction 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 (Schw artz et al.. 1997b). decreased levels of zinc protoporphyrin
(Scinicariello et al.. 2007; Kim et al. 2004). lower cortical bone Pb (Smith et al.. 1995b).
and lower amounts of 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
ALAD-1 gene (blood Pb level = 6.2 ±4.1 (ig/dL), and that ALAD-2 carriers
demonstrated decreased renal function at lower patellar Pb concentrations than those
observed to decrease renal function in 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 among individuals with ALAD 1-1 or ALAD 1-2/2-2 genotypes a significant
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increase in the amount of Pb found in the plasma (0.44 (ig/L versus 0.89 (ig/L,
respectively) and in the % plasma/blood ratio (0.48% versus 1.45%, respectively). This
potentially suggests that individuals with the ALAD 1-2/2-2 genotype 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 kiloDalton (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 likely
would 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.
Studies also observed the presence of an inducible, low-molecular weight (approximately
10 kDa) Pb-binding protein in workers occupationally exposed to Pb (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 that exhibited
toxicity at low blood Pb concentrations were observed to have 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 confirmed by Lolin and
O'Gorman (1988) and Church et al. (1993a. 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 2 Pb workers (blood Pb >160 j^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 worker. The reported protein was rich in cysteine residues
and tentatively identified as metallothionein.
5.2.3.4 Metallothionein
Metallothionein is a low-molecular weight metal-binding protein, most often zinc or
copper, which is rich in cysteine residues and plays an important role in the protection
against heavy metal toxicity, trace element homeostasis, and scavenging free radicals (Yu
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et al.. 2009). Exposure to Pb-acetate induces the production of Pb- and Zn-
metallothionein in mice exposed via i.p. or intravenous (i.v.) injection at 30 mg/kg
(Maitani et al.. 1986). in mice exposed via i.p. injection at 300 (.imol/kg (Yu et al.. 2009).
or in rats exposed via i.p. injection at 24 |_imol/l OOg (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 zinc before injection with Pb resulted in Pb and zinc co-eluting
with zinc-thionein, and that zinc-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
zinc-thionein decreased binding of Pb to liver ALAD and was able to donate zinc 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 zinc-metallothionein reduced Pb-induced
membrane leakage and loss of potassium 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 the wild-type mice (Qu et al.. 2002). 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 mitigation of Pb-induced toxic effects, and that
those with polymorphisms in metallothionein coding genes 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
carriers of the normal allele (24.17 and 21.27 versus 17.03 (ig/dL), and were more
susceptible to the effects of Pb on systolic BP and serum renal function parameters.
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5.2.4
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 (02~), hydroxyl radical (OH*), and hydrogen
peroxide (H202), which can attack and denature functional/structural molecules and,
thereby, promote tissue damage, cytotoxicity, and dysfunction. Pb 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.. 1992bV The origin of ROS
produced after Pb exposure is likely a multipathway process, resulting from oxidation of
S-aminolevulinic acid (ALA), membrane and lipid oxidation, nicotinamide adenine
dinucleotide phosphate (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 proteins, such as superoxide dismutase
(SOD), catalase (CAT), and glutathione peroxidase (GPx). Interestingly, Pb exposure in
many species of plants, invertebrates, and vertebrates discussed in the Ecological Effects
of Pb results in upregulation of antioxidant enzymes and increased lipid peroxidation
(Chapter 7). Oxidative stress is a common mode of action for a number of other metals
(e.g., Cd, Mn, As, Co, Cr) that are often found with Pb and by which possible interactions
with Pb have been suggested to occur (Jomova and Valko. 2011; Jomova etal. 2011;
Matovic etal.. 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.
5.2.4.1 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 S-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 S-ALA synthetase, 8-
ALAD, and ferrochelatase. Pb has been shown to be able to disrupt the zinc ions requisite
for the activity of S-ALAD, the rate limiting step in heme synthesis, leading to enzyme
inhibition at picomolar concentrations (Simons. 1995). Additionally, low blood Pb levels
(mean: 7 (ig/dL) have been found to inhibit the activity of 5-ALAD in humans, and the
lowest blood Pb level observed to be associated with lower S-ALAD activity in these
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studies was 5 (.ig/dL (Ahamed et al.. 2005; Sakai and Morita. 1996). A significant
negative correlation (r = -0.6) was found between blood Pb levels in adolescents (range
of blood Pb levels: 4-20 j^ig/dL) and blood S-ALAD activity in adolescents (Ahamedet
al.. 2006). This inhibition of S-ALAD results in the accumulation of 8-ALA in blood and
urine, where 8-ALA undergoes tautomerization and autoxidation. Oxidized S-ALA
generates ROS through reduction of ferricytochrome c and electron transfer from oxyHb,
metHb, and other ferric and ferrous complexes (Hermes-Lima et al.. 1991; Monteiro et
al.. 1991). The autoxidation of 8-ALA produces 02, OHv H202, and an ALA radical
(Monteiro etal.. 1989; Monteiro et al.. 1986).
5.2.4.2 Membrane and Lipid Peroxidation
A large number of studies in humans and experimental animals have found that exposure
to Pb can lead to membrane and lipid peroxidation. It is possible that ROS produced from
8-ALA oxidation, as described above, interacts with and disrupts membrane lipids
(Oteiza et 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; Quintan et al. 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 H removal easier
(Yiin and Lin. 1995). After Pb exposure (4-12 j^ig/dL in vitro, 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. In the absence of Fe2+,
Pb has not been shown to promote lipid peroxidation; however, it may accelerate
peroxidation by H202 ro„inian etai.. i9ss\ 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 (AA, 20:4) as a percentage of total fatty acids,
and decreased the relative proportion of shorter chain fatty acids (linoleic acid, 18:2)
(Law ton 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 AA. 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
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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. Occupational Pb exposure resulting in elevated blood Pb
levels (means >8 j^ig/dL) 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 a blood MDA levels to be positively correlated (r = 0.7) with blood Pb
levels ranging between 4-20 (ig/dL. Similar results have been shown after Pb exposure in
animal studies (Abdel Moneim et al.. 201 lb; Pandva etal.. 2010; Dogru et al.. 2008; Yu
et al.. 2008; Adcgbcsan and Adenuga. 2007; Lee et al.. 2005). Enhanced lipid
peroxidation has been found in Pb treated (50 jj.g, 1-4 hours) rat brain homogenates
(Rehman etal.. 1995). rat proximal tubular cells (0.5-1 (.iM, 12 hours) (Wang et al..
2011b). 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.
Interestingly, many species of plants, invertebrates, and vertebrates exhibit increased
lipid peroxidation with Pb exposure (Sections 7.2.4, 7.3.5, and 7.3.12). The increase in
lipid peroxidation following Pb exposure observed across species and kingdoms
demonstrate an evolutionarily conserved oxidative response following Pb exposure.
5.2.4.3 NAD(P)H Oxidase Activation
NAD(P)H oxidase is a membrane bound enzyme that requires Ca2+ in order to catalyze
the production of 02~ 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
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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.
5.2.4.4 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 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
protein that requires iron ions to function (Putnam et al. 2000). SOD exists in multiple
isoforms that require copper and zinc (SOD1 and SOD3) (Antonvuk et al.. 2009) or
manganese (SOD2) (Borgstahl etal.. 1992). A number of studies have found decreased
activity of these enzymes (Pandva et al.. 2010; Ergurhan-Ilhan et al.. 2008; Mohammad
et al.. 2008; Yu et al.. 2008; Patil et al.. 2006a; Patil et al.. 2006b; Conterato et al.. In
Press), whereas others have observed increased activity following Pb exposure (Ahamed
et al.. 2008; Lee et al. 2005). Pb exposure (500 ppm, 1, 4, and 8 weeks) in rats showed
that organ SOD and CAT responded differently depending on the dose and tissue
investigated. Activity of SOD and CAT varied based on the brain region analyzed and
time of exposure (Bennet et al.. 2007). Another study found that the brain had
consistently decreased SOD activity, irrespective of 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 SOD and CAT protein 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 and
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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 (Pandva et al.. 2010; Pillai et al.. 2010; Ahamed et
al.. 2009; Ahamed et al.. 2008; Flora et al.. 2007; Ahamed et al.. 2005; Chettv et al..
2005; Nakagawa. 1991. 1989). Long-term Pb exposure elicits 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 (Daggett et al.. 1998; Corongiu and
Milia. 1982; Hsu. 1981; Conterato et al.. In Press). However, other studies have found
that long-term Pb exposure resulting in mean blood Pb levels between 6.6 and 22 j^ig/dL,
causes the depletion of GSH (Mohammad et al.. 2008; Lee et al.. 2005; Ercal et al..
1996). Thus, the time of exposure is important to consider when measuring GSH levels.
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; Ercal et al.. 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;
Sandhir et al.. 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).
GSH is used as a cofactor for peroxide reduction and detoxification of xenobiotics by the
enzymes GPx and GST. GPx requires selenium for peroxide decomposition (Rotruck et
al.. 1973). whereas GST functions via a sulfhydryl group. By reducing the uptake of
selenium, depleting cellular GSH, and disrupting protein thiols, evidence indicates that
Pb decreases the activity of GPx and GST (Pillai et al.. 2010; Yu et al.. 2008; Lee et al..
2005; Nakagawa. 1991; Schrauzer. 1987). Similar to other antioxidant enzymes,
compensatory upregulation of these enzymes is described after Pb exposure in animals
and in Pb-exposed workers (painters with a mean blood Pb level of 5.4 j^ig/dL) (Bokara et
al.. 2009; Ergurhan-Ilhan et al.. 2008; Conterato et al.. 2007; Daggett et al.. 1998;
Conterato et al.. In Press). 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).
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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 GSHto 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) with lower blood Pb levels
(Lee et al.. 2006a'). In this study of nonoccupationally-exposed individuals, a
concentration-dependent relationship was observed at blood Pb levels <7 (ig/dL.
5.2.4.5 Nitric Oxide Signaling
'NO, also known as endothelium-derived relaxing factor, is a potent endogenous
signaling molecule involved in vasodilation. Short- and long-term Pb exposure in animals
decreases the biologically active 'NO, not through reduction in 'NO-production capacity
(Vaziri and Ding. 2001; Vaziri et al.. 1999b). but as a result of inactivation and
sequestration of'NO by ROS (Malvezzi et al. 2001; Vaziri et al.. 1999a). Endogenous
'NO can interact with ROS, specifically 02, 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 et al.. 1999a). '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 immune
cells (Pineda-Zavaleta et al.. 2004; Lee et al.. 2001b; Tian and Lawrence. 1995). 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 zinc 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 et al. 1999b; Gonick et al.. 1997). A direct
inhibitory action of Pb on NOS enzymatic activity has been rejected (Vaziri et al..
1999b). 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 aheterodimer requiring
one molecule of heme for enzymatic activity (Boerrigter and Burnett. 2009V 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 for 30 days and 100 ppm
for 12 months [resulting in a mean blood Pb level of 29.4 j^ig/dL|) (Marques et al.. 2001;
Khalil-Manesh et al.. 1993b). In addition, Pb exposure downregulated the protein
abundance of sGC in vascular tissue (Farmand 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).
5.2.5 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., selenium) accompany this
immunomodulation (Chettv et al.. 2005).
Traditional immune-mediated inflammation can be seen with bronchial
hyperresponsiveness, asthma, and respiratory infections 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 spheres implanted in the brains of rats produced neutrophil-driven
inflammation with apoptosis and indications of neurodegeneration (kibavashi et al..
2010). Pb also induces renal tubulointerstitial inflammation (mean blood Pb level of
18 (.ig/dL or 100 ppm exposure for 14 weeks) (Rodriguez-Iturbe et al.. 2005; Ramesh et
al.. 2001). which 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 23-27 (ig/dL,) (Bravo
et al.. 2007). These events could be in response to the oxidative environment arising from
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Pb exposure, since Pb-induced inflammation and NFkB activation can be ameliorated by
antioxidant therapy (Rodriguez-Iturbe et al. 2004).
Inflammation can be mediated by the production of chemical messengers such as
prostaglandins (PG). Pb exposure has been reported to increase 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 h) (Chettv et al.. 2005; Flohe et al. 2002;
Knowles and Donaldson. 1997; Lee and Battles. 1994). Dietary Pb supplementation 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 NFkB and activator protein-1 (AP-1), as well as the AP-1
component c-Jun (korasln and El-Kadi. 2008; korashv 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, STAT3 or 5, or
CREB (Gao et al.. 2007)
5.2.5.1 Cytokine Production
There are three modes by which Pb affects 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 et al.. 1999; Dentener et al.. 1989). This
can result in local tissue damage during the course of immune responses affecting such
targets as the liver. Second, when Pb acts on dendritic cells, it skews the ratio of IL-
12/IL-10 such that T-derived lymphocyte helper (Th)l responses are suppressed and Th2
responses are promoted (Chen et al. 2004; Miller et al.. 1998). Third, when acquired
immune responses occur following exposure to Pb, Thl lymphocyte production of
cytokines is suppressed (e.g., IFN-y) (Lvnes et al.. 2006; Heo et al. 1996). In contrast,
Tli2 cytokines such as IL-4, IL-5, and IL-6 are elevated (Gao et al.. 2007; Kim and
Lawrence. 2000). The combination of these three modes of cytokine changes induced by
Pb creates a hyperinflammatory state among innate immune cells and acquired immunity
is skewed toward Th2 responses.
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Iavicoli et al. (2006a) 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 is also altered by Pb exposure to cells (1 (.iM. 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-1|3
and TNF-a were observed in the hippocampus after Pb exposure (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
(Struzvnska 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 5.6.3) 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
of children in grades 5 and 6 in Taiwan, investigators did not group children by blood Pb
levels but by potential for Pb exposures due to age of home and location of residence
(Hsiao etal.. 2011). 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). 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
healthy and allergic children living near the oil refinery may have been influenced by
differences in their blood Pb levels.
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Evidence of association between blood Pb levels and cytokine levels in
nonoccupationally-exposed adults was equivocal. Among 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 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 in log of IL-6 (95%
CI: 0, 55%). 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 were 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. Among adults in Italy,
blood Pb levels were not correlated with either Th2 or Thl cytokine levels in men
(Boscolo et al.. 1999) or 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; Yucesov et al.. 1997a). The exception was a study of male foundry
workers, pottery workers, and unexposed workers by Valentino et al. (2007). Multiple
regression analyses were performed with age, BMI, smoking, and alcohol consumption
included as covariates. Although information on concentration-response relationships
was not provided, higher blood Pb level was associated with higher IL-10 and TNF-a.
Levels of IL-2, IL-6, and IL-10 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 j^ig/dL) blood Pb level
groups, with unexposed 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
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et al. (1997a) 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. However, 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 lymphocyte cytokine production.
5.2.6 Endocrine Disruption
5.2.6.1 Hypothalamic-Pituitary-Gonadal Axis
Pb is a potent endocrine disrupting chemical found to be associated with reproductive and
developmental effects in both male and female animal models. 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 has been shown to reduce serum levels of follicle-
stimulating hormone (FSH), luteinizing hormone (LH), testosterone, and estradiol
(Biswas and Ghosh. 2006; Rubio et al.. 2006V 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; Ron is et al.. 1998b; Foster. 1992; Sokol and Berman.
1991). Increases in 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 essential in maintaining gonadotropin-stimulated steroidogenesis,
which results in the formation of testosterone and estradiol. Prenatal Pb exposure
(resulting in 3 |_ig/dL blood Pb) 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
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al.. 1994). Pb may also interfere with release of pituitary hormones through interference
with cation-dependent secondary messenger systems, which 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) 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 et al.. 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 is 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 et al.. 1998a). Direct generation of ROS in epididymal spermatozoa was
observed after Pb exposure in rats (i.p. 20 or 50 ppm, 6 weeks) (Hsu et al.. 1998b). In
addition, lipid peroxidation has been observed in Pb-exposed rats (i.p. 0.025 ppm, 15
days) (Pandva et al.. In Press). Lipid peroxidation in the seminal plasma was significantly
increased in a group of Pb-exposed workers with high blood Pb levels (>40 j^ig/dL)
(Kasperczvk 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 increased 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) and gonadal cells (50 ppm Pb exposure) (kolesarova et al..
2010) show 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 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.
Decreased growth after Pb exposure could be the result of Pb-induced decreased GH
levels (Berry et al.. 2002; Camoratto et al.. 1993; Huseman etal.. 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 5.2xl0"5 |_iM.
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30 minutes) (Lau et al.. 1991). GH secretion may also be altered from decreased
testosterone, a result of Pb exposure.
5.2.6.2 Hypothalamic-Pituitary-Thyroid Axis
The effects of Pb on the hypothalamic-pituitary-thyroid (HPT) axis are mixed. Pb
exposure impacts a variety of players in the thyroid hormone system. A number of human
studies (blood Pb levels >7.3 (ig/dL) have shown that elevated blood Pb are 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 (Swamp et al..
2007). whereas mice and chickens manifested decreased serum T3 concentrations after Pb
exposure accompanied by increased lipid peroxidation (Chaurasia et al.. 1998; Chaurasia
and Kar. 1997). Decreased serum T3 and increased lipid peroxidation were both 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.
5.2.7 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, 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
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or Pb in studies using this form of inorganic Pb, only studies utilizing other forms of
inorganic Pb (e.g., Pb nitrate, acetate) are discussed below.
5.2.7.1 DNA Damage
A number of studies in human populations have observed associations between 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..
2010; Minozzo et al.. 2010; Shaik and Jamil. 2009; Danadevi et al.. 2003; Hengstler et
al.. 2003; Pal us et al.. 2003; Fracasso et al.. 2002; de Restrepo et al.. 2000). It is
important to note that occupationally-exposed adults have very high blood Pb levels, and
in one study (de Restrepo et al.. 2000) the association between blood Pb level and DNA
damage was observed only in workers with blood Pb levels greater than 120 (ig/dL. Also,
the studies were 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 j^ig/dL [range: 28.2-65.5 j^ig/dL|) and Minozzo et al. (2010)
(mean [SD]: 59.43 (ig/dL [28.34]) observed no correlation. Lastly, Pb-exposed workers
are also potentially exposed to other genotoxic materials, making it difficult to rule out
confounding co-exposures. However, Hengstler et al. (2003) examined workers exposed
to Pb, cadmium, and cobalt 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/nr1) 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
cadmium in air and blood and cobalt in air indicating lack of confounding. Two studies
were found that investigated Pb-induced DNA damage resulting from nonoccupational
exposures. Mendez-Gomez (2008) observed that children living at close and intermediate
distances to a Pb smelter had mean (range) blood Pb levels of 19.5 (11.3-49.2) and 28.6
(11.4-47.5) (ig/dL, respectively, compared to blood Pb level of 4.6 (0.1-8.7) (ig/dL for
children living distant to the smelter. DNA damage was increased 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
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of DNA damage compared to control children who lived further away from the mining
facility (mean [range] blood Pb level: 8.3 [3.0-25.0] (ig/dL).
In mice given 0.7 to 89.6 mg/kg Pb nitrate 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 concentration (Devi et al. 2000). The three highest concentrations had
responses which were similar in magnitude and were actually lower than the responses to
lower concentrations tested. In mice exposed to Pb (blood Pb level of 0.68 (ig/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
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. 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 Pb-exposed rats (blood Pb levels of 19.5 and
21.9 (.ig/dL) compared to control rats (Nava-Hemandez 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 0.5 or 1% Pb nitrate in drinking water for 60 days. Interestingly, although the
results were not statistically significant and 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 to 500 (.iM Pb-acetate for 20-25 hours or 100 (.iM for 0.5-4 hours (Hartwig et al..
1990; Snvder and Lachmann. 1989). Pb nitrate, administered to lymphoma cells at 1000-
10,000 (.iM for 6 hours, did not result in any DNA-protein crosslinks (Costa et al.. 1996).
Pb-acetate was observed by Wozniak and Blasiak (2003) to result in DNA single and
double strand breaks in primary human lymphocytes exposed 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 |aM. 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 to 2,100-3,300 (.iM Pb nitrate for 2 hours. Although there was a concentration-
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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 exposure of human endothelial cells to 1-1,000 |iM
Pb nitrate for 24 hours resulted in a concentration-dependent increase in DNA double
strand breaks.
Studies in animal cell lines 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 (jM, 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 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-acetate did not induce single strand DNA breaks in He La cells exposed to 500 (j,M for
20-25 hours (Hartwig et al.. 1990). However, exposure to both Pb-acetate and UV light
resulted in increased persistence of UV-induced strand breaks, compared to 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 to 100 |iM
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). Interestingly, in contrast to the above studies, exposure of lung carcinoma cells to
100, 300, or 500 |iM 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 |iM Pb-acetate (a
nonmutagenic concentration of Pb-acetate), indicating an inhibition of DNA repair.
Treatment of Chinese hamster ovary cells to 0.5-500 |iM Pb-acetate resulted in a
concentration-dependent accumulation of apurinic/apyrimidinic site incision activity,
indicating that DNA repair was diminished (McNeill et al.. 2007).
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5.2.7.2 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 a number of other heavy metals and organic contaminants. Higher
blood Pb level (range 1.6-5.2 j^ig/dL) was associated with greater HPRT mutation
frequency 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 four studies using human cell cultures. Ye
(1993) exposed human keratinocytes to 0.1-100 (iM/mL Pb-acetate for 2-24 hours. This
study did not measure HPRT mutations directly, but rather measured the amount of
tritium 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 Chinese hamster V79 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 assay) exposed to 1,700 (.iM for 5 days (Rov and Rossman. 1992).
Pb-acetate was observed to be nonmutagenic (HPRT assay) in Chinese hamster ovary
cells exposed to 5 (j,M for 6 hours (McNeill et al.. 2007). The observation of mutagenicity
in the second 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 (Rov 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 (.iM. and was
higher than that observed at higher 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
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with Pb nitrate, it was not concentration-dependent (Zelikoffet al.. 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. 1996).
5.2.7.3 Clastogenicity
Clastogenicity is the ability of a compound to induce chromosomal damage, and is
commonly observed as sister chromatid exchange, micronuclei formation, or incidence of
chromosomal aberrations (i.e., breaks or gaps in chromosomes). The potential for Pb to
be clastogenic has been investigated in numerous studies as described below.
Sister Chromatid Exchange
An association between blood Pb levels (means: 10.48 - 86.9 j^ig/dL) and sister chromatid
exchange (SCE) was observed in a number of occupational studies (Wiwanitkit et al..
2008; Diivdu et al.. 2005; Pal us et al.. 2003; Diivdu et al.. 2001; Pinto et al. 2000;
Bilban. 1998; Anwar and Kamal. 1988; Huang et al.. 1988). However, there are
numerous methodological issues that limit firm conclusions from being drawn. Most
notably, occupational co-exposures to other genotoxic materials were possible, although
some studies excluded workers with exposures to known mutagens (Pinto et al.. 2000;
Huang et al. 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 et al. 2003; Diivdu et al.. 2001; Pinto et al..
2000). 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 (Raiah and Ahuia.
1996; Raiah and Ahuia. 1995). However, these two studies may have suffered from
limited statistical power to observe an effect as they included very small numbers of Pb-
exposed workers. 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 exposed to up to 100 mg/kg Pb-acetate i.p., Pb induced SCEs with 50 and
100 mg/kg (Fahmv. 1999). Pb nitrate, also administered i.p. and induced the formation of
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SCEs in a concentration-dependent manner (10-40 mg/kg) in the bone marrow of
exposed mice (Dhiretal.. 1993). Nayak et al. (1989b) exposed pregnant mice to 100-
200 mg/kg Pb nitrate via i.v. injection and observed an increase in SCEs in dams at 150
and 200 mg/kg; no SCEs were observed in the fetuses. Tapisso et al. (2009) exposed rats
to 21.5 mg/kg Pb-acetate (l/10th the LD50) via i.p. injection on alternating days for 11 or
21 days, for a total of 5 or 10 exposures. Induction of SCEs 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 three of these studies utilized an injection route of
exposure that may not be relevant to routes of exposure in the general population
(e.g., air, drinking water exposure).
Few studies were found that investigated SCE formation in human cell lines due to Pb
exposure. Statistically significant, concentration-dependent increases in SCEs were
observed in human lymphocytes obtained from a single donor when incubated with 1,5,
10, or 50 (.iM Pb nitrate (Ustundag and Duvdii. 2007). Melatonin and N-acetylcysteine
were reported to ameliorate these effects, indicating Pb may induce SCEs 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.. In Press).
Studies investigating SCE in rodent cells were more equivocal than those in human cells.
Pb sulfate, acetate, and nitrate were found not to induce SCE in Chinese hamster V79
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 SCEs. 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 SCEs in Chinese hamster ovary cells. Lin et al. (1994) also
observed a concentration-dependent increase in SCEs in Chinese hamster cells exposed
to 3-30 |_iM Pb nitrate for 2 hours.
Micronucleus Formation
Pb-induced micronucleus formation was observed in numerous occupational studies
(Grover et al.. 2010; Khan et al.. 2010b; Minozzo et al.. 2010; Shaik and Jamil. 2009;
Minozzo et al.. 2004; Pal us et al.. 2003; Vaglenov et al.. 2001; Pinto et al.. 2000; Bilban.
1998; Vaglenov et al. 1998). The workers in the occupational studies generally had high
blood Pb levels (>20 j^ig/dL) making comparisons to the general population difficult,
although 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; Pal us et al.. 2003;
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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.
Only one study was found that investigated micronucleus formation in a nonworker
population; Mielzynska et al. (2006) 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 140, 250, or 500 g/kg
Pb-acetate once per week for 10 weeks resulted in statistically significantly increased
numbers of micronucleated polychromatic erythrocytes (PCEs) compared to controls.
Similarly, Alghazal et al. (2008b) exposed rats to 100 ppm Pb-acetate daily for 125 days
and observed statistically significant increases in micronucleated PCEs in both sexes.
Tapisso et al. (2009) exposed rats to 21.5 mg/kg Pb-acetate (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 that 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) exposed mice to 10
or 20 mg/kg Pb nitrate i.p. and observed a concentration-dependent increase in
micronuclei, whereas Jagetia and Aruna (1998) observed an increase in micronuclei in
mice exposed to 0.625-80 mg/kg Pb nitrate i.p., though the increase was not
concentration-dependent. Mice exposed to 1 g/L Pb-acetate 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 to either 1,5, 10, or 50 |_iM Pb nitrate or 3 or 5 ppm
Pb chloride (Ustundag and Duvdii. 2007; Turkez et al.. In Press). Gastaldo et al. (2007)
also observed a dose-dependent increase in micronuclei in human endothelial cells
exposed to 1-1,000 (.iM Pb nitrate for 24 hours. Two 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).
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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
(Grover et al.. 2010; Shaik and Jamil. 2009; Pinto et al.. 2000; Bilban. 1998; De et al..
1995; Huang et al.. 1988). Methodological limitations outlined in previous sections,
including potential for occupational co-exposure to genotoxic substances and generally
high blood Pb levels (>20 (ig/dL) that limit the relevance of findings to the general
population, also pertain to the present findings. 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
(Grove ret 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. Although blood Pb levels were
statistically significantly higher in children living in the Pb-contaminated area than in
children living in the less contaminated area, they were generally comparable (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) exposed mice to 25-
400 mg/kg Pb-acetate i.p., either as a single exposure 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
exposure regimens. Pb nitrate was also observed to produce concentration-dependent
chromosomal damage in mice exposed i.p. to a single exposure of 5, 10, or 20 mg/kg
(Phir et al.. 1992b). In a similar experiment, Dhir et al. (1990) exposed mice to 10, 20, or
40 mg/kg Pb nitrate 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) exposed pregnant mice to 100-200 mg/kg Pb nitrate 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 exposure. In a
similar experiment, low levels of chromosomal aberrations were observed in dams and
fetuses injected with 12.5-75 mg/kg Pb nitrate, but there was no concentration-dependent
response reported and few cells were analyzed (Navak etal.. 1989a). In rats given
2.5 mg/100 g Pb-acetate 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 (Chakrabortv et al..
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1987). The above studies all suffer from the use of a route of exposure that may not be
relevant to human environmental exposures. However, studies utilizing drinking water or
dietary exposures also observed increases in chromosomal damage. Aboul-Ela (2002)
exposed mice to 200 or 400 mg/kg Pb-acetate by gavage for 5 days and reported that
chromosomal damage was present in the bone marrow cells and spermatocytes of animals
exposed to both exposure 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 0.5%
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 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 (Wise et al.. 1994). 3-30 (.iM for 2 hours (Lin
etal.. 1994). or 0.05-1 (.iM 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 (iM).
5.2.7.4 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.
DNA Methylation
A single i.v. injection of 75 (imol/kg Pb nitrate resulted in global hypomethylation of
hepatic DNA in rats (Kanduc et al.. 1991). The observed hypomethylation in the liver
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was associated with an increase in cell proliferation. A few additional studies in humans
observed that higher bone Pb levels were associated with lower DNA methylation
patterns in adults and cord blood of newborns (Wright et al.. 2010; Pilsner et al.. 2009).
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 discernable 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 potential 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 that Pb
reduced mitogenesis in Pb-exposed workers (mean blood Pb levels: 35.4 j^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 exposed to 100 |_iM/kg via i.v. injection (Nakaiima et
al.. 1995; Coni et al.. 1992; Lcdda-Columbano et al.. 1992; Columbano et al.. 1990;
Columbano et al.. 1987). Shinozuka et al. (1996) observed that 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 and NF-kB, TNF-a, and iNOS in
liver cells in exposed animals at 12.5 mg/kg and 100 (iM/kg, respectively (Cheng et al..
2002; Menegazzi et al.. 1997). In the only study that examined Pb exposure via
inhalation, exposure to 10,000 (.iM Pb-acetate for 4 weeks resulted in increased cellular
proliferation in the lungs (Fortoul et al. 2005).
A great amount of 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) (Hciman 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, but did increase the mitotic index (250-1,000 (.iM for 24 hours) (Wise et al.. 2005).
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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 |iM 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, chloride, oxide, and 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). Iavicoli 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 mesenchymal
stem cells when administered at 20-100 (j,M 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 (J.M 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 (j,M, had a mitotic index equal to that in the untreated control
cells.
5.2.7.5 Gene Expression
Two animal studies have investigated the ability of Pb to alter gene expression in regard
to phase I and II metabolizing enzymes. Suzuki et al. (1996) exposed rats to 100 |ig/kg
Pb-acetate or nitrate 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 i.v. exposure to 20, 50, or 100 |imol/kg Pb nitrate
selectively inhibited CYP1A2 levels. Pb was shown not to inhibit CYP1A2 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 10-100
(jM Pb nitrate 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 500 |ig/L Pb-acetate for 72 hours resulted in the
up-regulation of multiple genes associated with cytochrome P450 activity, glutathione
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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 0.2% Pb-acetate via lactation from postnatal day
(PND)l-20, the expression of proteolipid protein, a major structural constituent of
myelin, was statistically significantly elevated 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 induction of 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 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 decreased in their livers. Induction of ABCC5 was observed to attenuate the toxicity
of Cd, but not Hg or As, in developing embryos, the attenuation of Pb-induced toxicity
was not investigated. However, these findings indicate that increased expression of
ABCC5 due to heavy metal exposure may play a part in cellular defense mechanisms.
5.2.7.6 Apoptosis
Occupational exposure to Pb and induction of apoptosis was investigated in a few studies.
One study directly reported that exposure to Pb increased apoptosis compared to
nonexposed controls (Minozzo et al.. 2010). whereas the others reported that two early
indicators of apoptosis, karyorrhexis and karyolysis, were elevated in exposed workers
(Grove ret al.. 2010; Khan et al.. 2010b). Pb nitrate was also observed to induce apoptosis
in the liver of exposed animals (Columbano et al.. 1996; Nakaiima et al.. 1995).
Apoptosis was observed in rat fibroblasts exposed to Pb-acetate and rat alveolar
macrophages exposed to Pb nitrate (lavicoli et al.. 2001; Shabani and Rabbani. 2000).
Observation of Pb-induced apoptosis may represent the dysregulation of genetically-
controlled cell processes and tissue homeostasis.
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5.2.8
Summary
The diverse health effects of Pb are mediated through multiple, interconnected modes of
action. Each of the modes of action discussed has the potential to contribute to the
development of a number of Pb-induced health effects (Table 5-2). While this section
draws from older literature as well as newer lines of evidence, the inclusion of new
evidence does not qualitatively change the conclusions regarding individual modes of
action. However, the new evidence strengthens the conclusions. Evidence for the
majority of these modes of action is observed at low blood Pb levels in humans, between
2 and 17 (ig/dL, with supporting evidence from in vitro assays. As many of these studies
are examining adults with likely higher past than current Pb exposures, uncertainty exists
as to the Pb exposure level, duration, and timing leading to these blood Pb levels
associated with these modes of action. These observable effect levels 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 concentrations represent the threshold
of the effect. The observable effect levels in humans, reported in Table 5-2, are drawn
from the available data; and, do not imply that theses modes of action are not acting at
lower exposure levels or that these doses represent the threshold of the effect. Also, the
data presented in this table does not inform regarding the exposure frequency and
duration required to elicit a particular MO A.
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Table 5-2 MOAs, their related health effects, and information on concentrations
eliciting the MOAs
Mode of Action
Concentrations or doses (Conditions)"
[Related Health Effects
(ISA section)]
Blood Pb
Dose
Altered Ion Status
[All Health Effects of Pb]
3.5 pg/dL
(mean in cord blood; association with cord blood Ca2t-
ATPase pump activity)
Huel et al. (2008)
0.00005 pM free Pb2t
(In vitro; 30 minutes; calmodulin activation assay)
Kern et al. (2000)
Protein Binding
[Renal (5.5), Heme Synthesis and RBC
Function (5.7)]
17.0 pg/dL
(concurrent mean in adult workers with wildtype
metallothionein expression; increased BP susceptibility)
Chen et al. (2010a)
50 pM Pb glutamate
(In vitro; 24 hours; increased nuclear protein in
neurological cell)
Klann and Shelton (1989)
Oxidative Stress
[All Heath Effects of Pb]
5.4 pg/dL
(concurrent mean in adult male workers; decreased CAT
activity in blood)
Conterato et al. (In Press)
0.1 pM Pb-acetate
(In vitro; 48 hours; decreased cellular GSH in
neuroblastoma cells)
Chetty et al. (2005)
Inflammation
[Nervous System (5.3), Cardiovascular
(5.4), Renal (5.5), Immune (5.6),
Respiratory (5.6.4), Hepatic (5.9.1)]
2.5 pg/dL
(concurrent minimum in adult males; increased serum
TNF-a and blood WBC count)
Kim et al. (2007)
0.01 pM Pb-acetate
(In vitro; 48 hours; increased cellular PGE2 in
neuroblastoma cells)
Chetty et al. (2005)
Endocrine Disruption
[Reproductive and Developmental Effects
(5.8), Endocrine System 5.9.3), Bone and
Teeth (5.9.4)]
1.7 pg/dL
(concurrent minimum in women with both ovaries
removed; increased serum FSH)
Krieg (2007)
10 pMPb nitrate
(In vitro; 30 minutes; displaced GHRH binding to rat
pituitary receptors)
Lau et al. (1991)
Cell Death/Genotoxicity
[Cancer (5.10), Reproductive and
Developmental Effects (5.8), Bone and
Teeth (5.9.4)]
3.3 pg/dL
(concurrent median in adult women; increased rate of
HPRT mutation frequency)
Van Larebeke et al. (2004)
0.03 pM 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 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.
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 5-1). Pb will interfere with endogenous Ca2+ homeostasis,
necessary as a cell signal carrier mediating normal cellular functions. |Ca2 | 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-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 PKC or calmodulin.
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Disruption of ion transport not only leads to altered Ca2+ homeostasis, it can also result in
perturbed neurotransmitter function. Pb has been shown to be able to displace metal ions,
such as Zn, Mg, and Ca2+, from proteins due to the flexible coordination number of Pb
and multiple ligand binding ability, leading to abnormal conformational changes to
proteins and altered protein function. Evidence for this metal ion displacement and
protein perturbation has been shown at picomolar 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 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. 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 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
S-ALA, NAD(P)H oxidase activation, membrane and lipid peroxidation, and antioxidant
enzyme depletion. Through the inhibition of S-ALAD due to displacement of Zn,
accumulated S-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 will also sequester and 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 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.
In a number of organ systems Pb-induced oxidative stress is accompanied by
misregulated inflammation. Pb exposure will modulate inflammatory cell function,
production of pro-inflammatory cytokines and metabolites, inflammatory chemical
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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 decreased production of related cytokines.
Evidence indicates that Pb is a potent endocrine disrupting chemical. Pb will disrupt the
HPG axis evidenced 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 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 and require
further investigation.
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
have been associated with these effects in experimental animals and cells. Results vary on
the effect of Pb on DNA repair activity, however 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 j^ig/dL), 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 DNA methylation, mitogenesis, and
gene 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 cation (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
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(Pentvala et al.. 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; Hengstler et al.. 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.
5.3 Nervous System Effects
5.3.1 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 studies conducted in diverse
populations and after adjusting for potential confounding by socioeconomic status (SES),
parental intelligence, and caregiving environment. This association was substantiated in a
pooled analysis of children, 5 to 10 years of age, 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).
Associations between blood Pb levels and decrements in intelligence quotient (IQ),
mental development, memory, and other specific indices of cognitive function in children
ages 6 months to 17 years were most strongly indicated in children with population mean
blood Pb levels (measured at various lifestages) in the range of 5-10 (ig/dL; however,
several results indicated associations in groups of children (ages 2-10 years) with mean
blood Pb levels in the range of 3-5 (.ig/dL (Bellinger. 2008; Can field. 2008; Hornung.
2008; Roio-Tellez. 2008). Based on fewer available studies, the 2006 Pb AQCD
described consistent associations of blood Pb levels with behavioral outcomes, including
inattention, and antisocial and delinquent behavior assessed in children ages 6 to 13 years
(U.S. EPA. 2006b).
Toxicological studies provided coherence with similarly consistent findings for Pb-
induced impairments in learning and behavior in rodents and monkeys (U.S. EPA.
2006b). In contrast with studies in children, Pb exposure was not found consistently to
affect memory of animals. Effects on learning were largely demonstrated as poorer
performance on Morris water maze and discrimination reversal tasks, and effects on
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behavior were largely demonstrated as distractibility, impulsivity, and insensitivity to
changes in reinforcement. These effects in animals were predominantly found with Pb
exposures that resulted in blood Pb levels between 20-50 j^ig/dL: however, some studies
observed learning and behavioral impairments in animals with steady-state blood Pb
levels of 10-15 (.ig/dL (Corv-Slechta. 1994; Altmann et al.. 1993; Rice and Karpinski.
1988; Gilbert and Rice. 1987). Toxicological studies also provided biological plausibility
by characterizing modes of action for Pb-induced nervous system effects. In particular,
toxicological evidence for Pb exposure interfering with neuronal metabolism at the
cellular and histological level (e.g., synaptic architecture during development,
neurotransmitter release, glia, neurite outgrowth, the blood brain barrier, and oxidative
stress), provided biological plausibility for blood Pb levels in children being causally
associated with deficits in multiple functional domains such as cognitive function, motor
function, memory, mood, and behavior. Additional biological plausibility was provided
by associations observed of childhood blood Pb levels with changes indicative of
neuronal damage and altered brain physiology assessed in young adults using magnetic
resonance imaging techniques (Yuan et al. 2006; Cecil et al. 2005; Meng et al.. 2005;
Trope et al.. 2001).
A common finding across several epidemiologic studies of children was a supralinear
concentration-response relationship between blood Pb level and neurocognitive deficits,
i.e., a larger decrement in neurocognitive function per unit increase in blood Pb levels in
children in the lower range of the population blood Pb level distribution (kordas et al..
2006; Schnaas et al.. 2006; Tellez-Roio et al.. 2006; Lanphear et al.. 2005; Rothenberg
and Rothenberg. 2005; Bellinger and Needleman. 2003; Can field et al.. 2003a). Most of
these epidemiologic results were based on the analysis of concurrent blood Pb levels and
a cut-point of 10 (ig/dL to define lower and higher blood Pb levels. These findings were
corroborated by findings from analyses of the pooled cohort data indicating that a
nonlinear relationship fit the data better than a linear relationship did (Lanphear et al..
2005; Rothenberg and Rothenberg. 2005). Consistent with epidemiologic findings,
toxicological studies observed nonlinear Pb concentration-response relationships for
outcomes such as neuronal activation (Lewis and Pitts. 2004). neurogenesis (Gilbert et
al.. 2005). and retinal responses (Fox et al.. 1991; Fox and Farber. 1988; Fox and Chu.
1988).
Another area of focus included the comparison of various lifestages of Pb exposure in
terms of risk of neurodevelopmental deficits. Toxicological studies clearly demonstrated
that in utero with or without early postnatal exposure to Pb was an especially sensitive
window for Pb-induced neurodevelopmental effects. Nonetheless, not all endpoints in
animal toxicology studies had a single defined window of sensitivity but instead were
shown to be affected by exposures at multiple periods during the lifespan of the
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organism. Epidemiologic studies observed cognitive deficits in children ranging from 2 to
10 years of age 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 and behavioral deficits in children and adolescents
(Wasserman et al.. 2003; Greene and Emhart. 1993; Bellinger et al.. 1991; Needleman et
al.. 1979). also pointing to an effect of cumulative childhood Pb exposure. Among studies
of children (ages 2-10 years) that examined blood Pb levels measured at multiple
lifestages, several found that concurrent blood Pb was associated with an equal or larger
decrement in IQ (Chen et al.. 2005; Lanphear et al.. 2005; Wasserman et al.. 1994;
Dietrich et al.. 1993b; Bellinger et al.. 1992). A common limitation of epidemiologic
studies of children was the high correlation among blood Pb levels at different ages,
making it difficult to ascertain which developmental periods of Pb exposure were
associated with the greatest risk of neurodevelopmental decrements (Lanphear et al..
2005). The issue of persistence of the neurodevelopmental effects of Pb exposure also
was considered, with some evidence suggesting that the associations of biomarkers of
early childhood Pb exposure (e.g., deciduous tooth, blood at age 2 or 6 years) with
neurodevelopmental outcomes persisted into adolescence and young adulthood (Ris et al..
2004; Tong et al.. 1996; Needleman et al.. 1990). Some studies in rats and monkeys also
demonstrated that the effects of in utero and early postnatal Pb exposures on
neurodevelopmental outcomes persisted into adulthood (Corv-Slechta. 1994; Altmann et
al.. 1993; Rice and Karpinski. 1988; Gilbert and Rice. 1987).
In epidemiologic studies of adults, a range of nervous system effects (e.g., impaired
memory, attention, reaction time, visuomotor tasks and reasoning, alterations in visual or
brainstem evoked potentials, postural sway) were mostly clearly indicated in Pb-exposed
workers with blood Pb levels in the range of 14 to 40 (ig/dL (Iwata et al.. 2005; Bleecker
et al.. 1997; Baker et al.. 1979; Cantarow and Trumper. 1944). In the limited literature
examining nonoccupationally-exposed adults, the weight of evidence supported
associations of bone Pb levels with cognitive function (Weisskopf et al.. 2004; Wright et
al.. 2003b) but not concurrent blood Pb levels (Krieg et al.. 2005; Nordberg et al.. 2000;
Pavton et al.. 1998; Muldoon et al.. 1996). These findings suggested that rather than
recent exposures, past or cumulative Pb exposures contributed to cognitive deficits in
nonoccupationally-exposed adults. With regards to neurodegenerative diseases, whereas a
few toxicological studies demonstrated Pb-induced amyloid plaques commonly
associated with Alzheimer's disease pathophysiology (Basha et al.. 2005; Zawia and
Basha. 2005). epidemiologic studies did not indicate that Pb exposure was associated
with Alzheimer's Disease in adults. Pb biomarker (blood or bone) levels were
inconsistently associated with amyotrophic lateral sclerosis (ALS) in adults in the general
population; however, some case-control studies found that history of occupational Pb
exposure was more prevalent among ALS cases than controls (Kamel et al.. 2002;
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Chancellor et al.. 1993). Associations were reported for essential tremor and symptoms of
anxiety and depression, 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 weight of evidence continued to be derived from
associations observed of Pb exposure and blood Pb levels in young animals and children,
respectively, with cognitive function decrements, inattention, and impulsivity. Expanding
upon the previous body of evidence, several recent studies in children found similar
associations with lower population mean (or quantile) blood Pb levels (1-5 |_ig/dL).
Whereas previous evidence was inconsistent, several new studies in children reported
associations between concurrent blood Pb levels and attention deficit hyperactivity
disorder (ADHD). Recent studies in adults focused primarily on cognitive function
decrements but also provided additional evidence for Pb-associated mood disorders,
ALS, Parkinson's Disease, and essential tremor. New toxicological studies expanded
evidence for the effects of prenatal and postnatal Pb exposure on learning, memory, and
attention and provided insight into the contribution of social stress to this paradigm. New
or expanded areas of toxicological research related to Pb exposure included mood
disorders, neurofibrillary tangle formation, and adult dementia after early life Pb
exposures. Historically important areas of toxicological research were further expanded
with recent findings for Pb-induced effects on neurotransmitters, synapses, glia, neurite
outgrowth, the blood brain barrier, and oxidative stress. The data detailed in the
subsequent sections continue to enhance the understanding of nervous system effects
associated with Pb exposure.
5.3.2 Cognitive Function and Learning
5.3.2.1 Epidemiologic Studies of Cognitive Function in Children
Epidemiologic studies have assessed global cognitive function most frequently by full-
scale IQ (FSIQ) and its verbal and performance subscale components in children ages 3
to 17 years and by the Bayley Scales of Infant Development in children ages 6 months to
3 years. These indices have strong psychometric properties and are among the most
rigorously standardized measures. A large body of evidence also comprises associations
of blood Pb levels with specific cognitive abilities, including memory and learning,
executive function, language, and visuospatial processing. These specific indices of
cognitive function are reflected in global measures of intelligence and also are more
comparable to tests of learning and memory in animals. Fewer studies have examined
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academic performance and achievement; however, these outcomes may provide
information on the impact of Pb exposure on life success. The epidemiologic evidence for
each of these categories of outcomes was evaluated separately in order of increasing
weight of evidence. Emphasis was placed on prospective studies with repeated
assessments of blood Pb levels and cognitive function and on studies examining blood Pb
levels similar to those measured in contemporary U.S children (i.e., less than 5 (ig/dL),
and younger children whose blood Pb levels are less influenced by higher past Pb
exposures.
Many factors have been shown to influence the intellectual abilities of children, including
parental SES, parental education, parental IQ, quality and stability of the caregiving
environment, 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 association with both blood Pb level
and outcome, these other risk factors may potentially confound (i.e., bias due to an
association with blood Pb level and causal association with the outcome) the associations
observed between blood Pb level and indices of cognitive function. In the evaluation of
the effect of Pb independent from the effects of the other variables, greater weight was
given to studies that accounted for potential confounding in the study design or in
statistical analyses. A detailed discussion of the collective weight of evidence for the
independent associations between Pb exposure and nervous system effects in relation to
the adequacy of control for confounding by other risk factors is located in Section 5.3.11.
Full-scale IQ in Children
Several longitudinal cohort studies were initiated in the 1980s in order to address
limitations of cross-sectional studies, including establishing a temporal association
between blood Pb levels and cognitive outcomes, examining the persistence of cognitive
deficits to older ages, and comparing risk estimates among blood Pb levels measured at
different lifestages. Moreover, cooperation among investigators to adopt similar
assessment protocols facilitated pooled and meta-analyses and comparison of results
across populations that differed in the range of blood Pb levels, race/ethnicity, and SES.
Individual cohort studies in diverse populations were consistent in demonstrating
associations between higher blood Pb and lower FSIQ in populations of school-aged
children with mean blood Pb levels in the range of 5 to 10 (ig/dL (Schnaas et al.. 2006;
Bellinger and Needleman. 2003; Can field et al.. 2003a; Wasserman et al.. 1997; Dietrich
etal.. 1993a; Baghurst et al.. 1992; Bellinger et al.. 1987) (Figure 5-2 and Table 5-3). In
analyses restricted to children in the lower range of the blood Pb distribution (e.g., <
10 (ig/dL), associations were observed in groups of children with mean blood Pb levels
3-4 (.ig/dL (Bellinger. 2008; Can field. 2008; Hornung. 2008). Across cohort studies,
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1	associations were observed with concurrent, prenatal, early childhood (e.g., age 2 years),
2	and childhood average blood Pb levels (e.g., 0-8 year average) (Figure 5-2 and Table
3	5-3). These findings were substantiated in a pooled analysis of seven prospective studies
4	(Boston, MA; Cincinnati, OH; Rochester, NY; Cleveland, OH; Mexico City, Mexico;
5	Port Pirie, Australia; and Kosovo, Yugoslavia) by Lanphear et al. (2005) as well as
6	multiple meta-analyses that included both prospective and cross-sectional studies
7	(Pocock et al.. 1994; Schwartz. 1994; Needleman and Gatsonis. 1990).
Study
Blood Pb Timing
Blood Pb Mean
(SD) (pg/dL)
Blood Pb10th-90th
percentile (pg/dL)
FSIQ a
Kim et al. (2009)
Concurrent
1.73(0.80)
<1-2.8
8-11
Chiodo et al. (2007)
Concurrent
5.0(3.0)
2.1-8.7
7
Chiodo et al. (2004)
Concurrent
5.4(3.3)
2.3-9.5
7.5
Juskoetal. (2011)
Concurrent
5.0(3.3)
1.9-9.0
6
Canfield et al. (2003)
Concurrent
5.8(4.1)
2.1-10.7
5
Bellinger et al. (1992)
2yr
6.5(4.9)
2.2-12.3
10
Kordasetal. (2011)
Prenatal (cord)
6.6(3.3)
1.2-2.4
4
Lanphear et al. (2005)
Concurrent
6.9(1.2)
3.4-25.4
4.8-10
Minetal. (2009)
Minetal. (2009)
Minetal. (2009)
Concurrent
4yr
4yr
7.0(4.1)
7.0(4.1)
7.0(4.1)
w w w
o o o
4
9
11
Schnaas et al. (2006)
Prenatal (maternal) 7.8 (geometric)
3.2-19.1
6-10
Kordasetal. (2011)
Concurrent
8.7(4.4)
1.4-2.7
4
Roy et al. (2011)
Concurrent
11.4(5.4)
5.8-18.4
3-7
Dietrich et al. (1993)
Concurrent
11.8(6.3) (Age 5)
5.5-19.7
6.5
Tong etal. (1996)
Oto 11-13 yr avg
14.0(1.2) geometric
12.7-18.1
11-3
Wasserman et al. (1997) 0 to 7 yr avg
16.2 (geometric)
6.0-44.2
7
Baghurst etal. (1992)
0 to 8 yr avg
16.6 (mean of 25th-50th) 10-27.5
7-8
Change in full scale IQ per 1 [jg/dL increase in the 10th to 90th percentile
of blood Pb level (95% CI)
Note: Studies generally are presented in ascending order of mean blood Pb level. 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 |jg/dL increase in blood Pb level within the 10th to 90th percentile interval. The percentiles are estimated using
various methods and are only approximate values. Effect estimates are assumed to be linear within the 10th to 90th percentile
interval of blood Pb level. The various tests used to measure FSIQ are scored on a similar scale (approximately 40-160). Black
diamonds, blue circles, orange triangle, and gray squares represent associations with concurrent, earlier childhood, prenatal, and
lifetime average blood Pb levels, respectively.
Figure 5-2 Associations of blood Pb levels with full-scale IQ (FSIQ) among
children.
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Table 5-3 Additional characteristics and quantitative results for studies
represented in Figure 5-2
Study
Population/
Location
Blood Pb Data
(Hg/dL)
Statistical Analysis
FSIQ Assessment8
Effect Estimate
(96% Cl)b c
Kim et al. 279 children in Seoul,
(2009a) Seongnam, Ulsan, and
Yeoncheon, Korea, ages
8-11 yr in April-December
2007
Concurrent
Mean (SD): 1.73 (0.80)
10th-90th: < 1-2.8
Log linear regression model
adjusted for age, sex, maternal
education, paternal
education, yearly income, maternal
smoking during pregnancy, indirect
smoking after birth, birth weight,
maternal age
KEDI-WISC at ages 8-11 yr -0.62 (-1.05, -0.19)
Chiodo et al. 506 African-American
(2007)	children in Detroit, Ml area
followed from birth
(1989-1991) to age 7 yr.
Large proportions of
children with prenatal
exposure to cocaine or
marijuana
Concurrent
Mean (SD): 5.0 (3.0)
10th-90th: 2.1-8.7
Regression model adjusted for WISC-III at age 7 yr
prenatal alcohol and drug use, SES,
Symptom Checklist (component of
HOME score), maternal IQ
-0.19 (-0.30, -0.C
Chiodo et al. 246 African-American
(2004)	children in Detroit, Ml area
followed from birth (not
reported) to age 7.5 yr.
Large proportions of
children with prenatal
exposure to cocaine or
marijuana
Concurrent
Mean (SD): 5.4 (3.3)
10th-90th: 2.3-9.5
Log linear regression model
adjusted for SES, education,
number of children <18 yr, HOME
score, maternal vocabulary test
score, sex, parity, family
environment scale
WISC-III at age 7.5 yr
-0.22 (-0.38, -0.05)"
Jusko et al. 194 children in Rochester,
(2011)	NY followed from age 6 mo
(1994-1995) to age 6 yr.
Concurrent
Mean (SD): 5.0 (3.3)
10th-90th: 1.9-9.0
Linear regression model adjusted
for sex, birth weight, transferrin
saturation, maternal race, maternal
IQ, maternal education, HOME
score, family income, and maternal
prenatal smoking
WPPSI-R at age 6 yr
-0.66 (-1.26, -0.06)
Canfield et al.
(2003a)
172 children in Rochester,
NY born 1994-1995
followed from age 6 mo to
age 5 yr.
Concurrent
Mean (SD): 5.8 (4.1)
10th-90th: 2.1-10.7
Linear regression model adjusted
for sex, maternal race, prenatal
smoking, maternal education, child
iron status, household income,
maternal IQ, HOME score, birth
weight
Stanford-Binet at age 5 yr
-0.61 (-0.99, -0.24)
Bellinger etal.
148 children in the Boston,
MA area followed from birth
(1979-1981) to age 15-17
yr.
Early childhood (age 2
yr)
Mean (SD): 6.5 (4.9)
10th-90th: 2.2-12.3
Linear regression model adjusted
for HOME score (age 10 and 5 yr),
child stress, race, maternal IQ,
SES, sex, birth order, marital status
WISC-Ratage10yr
-0.58 (-0.99, -0.18)
Lanphear et 1,333 children pooled from
al. (2005) Boston, Cincinnati,
Cleveland, Mexico City, Port
Pirie, Rochester, and
Yugoslavia cohorts
Concurrent
Mean (SD): 6.9 (1.2)
10th-90th: 3.4-25.4
Log linear regression model
adjusted for HOME score, birth
weight, maternal IQ, maternal
education
FSIQ measured at ages 5-10 yr -0.25 (-0.34, -0.15)
Min et al. 267 primarily African-
(2009)	American children in the
Cleveland, OH area
followed from birth
(1994-1996) to age 11 yr.
Children were exposed
prenatally to multiple drugs.
Age 4 yr
Mean (range): 7.0
(1.3-23.8)
10th-90th: 3.0-12.1
Linear regression model adjusted
for HOME score, caregiver's
vocabulary test, sex, parity,
maternal marital status, head
circumference at birth
WISC-R at age 4 yr (concurrent)
WISC-R at age 9 yr
WISC-R at age 11 yr
-0.50 (-0.89, -0.11)
-0.41 (-0.78, -0.04)
-0.54 (-0.91,-0.17)
Schnaas et al.
150 children in Mexico City,
Mexico followed from birth
(1987-1992) to age 10 yr.
Prenatal (maternal
28-36 weeks)
Geometric mean (95%
CI): 7.8 (2.5-24.5)
10th-90th: 3.2-19.1
Log linear mixed effects regression
model adjusted for sex, SES,
maternal IQ, HOME score, birth
weight, postnatal blood Pb, random
slope for subject
McCarthy GCI at ages 6-10 yr -0.44 (-0.73, -0.15)
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10
Study
Population/
Location
Blood Pb Data
(Hg/dL)
Statistical Analysis
FSIQ Assessment8
Effect Estimate
(96% Cl)b c
Kordas et al. 186 children in Mexico City,
(2011)	Mexico followed from birth
(1994-1995) to age 4 yr
Prenatal (cord)
Mean (SD): 6.6 (3.3)
10th-90th: 1.2-2.4
Concurrent
Mean (SD): 8.7 (4.4)
10th-90th: 1.4-2.7
Linear regression model adjusted
for sex, birth weight, gestational
age, maternal age, years of
schooling, IQ, smoking status,
marital status, crowding in home,
type of floor in home
McCarthy GCI at age 4 yr
Prenatal
Concurrent
-0.20 (-0.79, 0.39)
-0.60 (-0.99, -0.21)
Roy et al.
(2011)
717 children in Chennai,
India ages 3-7 yr
Concurrent
Mean (SD): 11.4 (5.4)
10th-90th: 5.8-18.4
Log linear model adjusted for
mid-arm circumference, age, sex,
family income, parental education,
parental IQ, family size
Binet-Kamat at ages 3-7 yr
-0.39 (-0.65, -0.13)
Dietrich et al.
(1993a)
253 children in Cincinnati,
OH followed from birth
(1979-1985) to age 20-23
yr.
Concurrent NR
Age 5 yr
Mean (SD): 11.8(6.3)
10th-90th: 5.5-19.7
Linear regression model adjusted
for HOME score, maternal IQ, birth
weight, birth length, child sex,
maternal cigarette consumption
during pregnancy
WISC-R at age 6.5 yr
-0.33 (-0.60, -0.06)
Tong et al. 375 children in Port Pirie,
(1996)	Australia followed from birth
(1979-1982) to age 11-13
yr.
Lifetime avg
Geometric mean
(GSD): 14.0(1.2)
10th-90th: 12.7-18.1
Regression model adjusted for sex,
age, school grade, parental
occupational prestige, HOME score,
maternal IQ, family functioning
score, parental smoking, marital
status, parental education, maternal
age, birth weight, birth order,
feeding method, breastfeeding
duration, family size, life events,
prolonged absences from school
WISC-R at age 11-13 yr
-0.12 (-0.24,
0.003)
Wfesserman et 290 children in Kosovo,
al. (1997) Yugoslavia followed from
birth (1985-1986) to age
10-12 yr.
Lifetime avg
Geometric mean: 16.2
10th-90th: 6.0-44.2
Generalized estimating equations
with log-transformed blood Pb
adjusted for age, sex, sibship size,
birth weight, language spoken in
home, HOME score, maternal age,
maternal education, maternal
Raven score
WISC-III at ages 10-12 yr
-0.20 (-0.28, -0.11)
Bag hurst et al.
494 children in Port Pirie,
Australia followed from birth
(1979-1982) to age 11-13
yr.
Lifetime avg
Mean of 25-50th: 16.6
10th-90th: 10-27.5
Log linear regression model
adjusted for sex, birth weight, birth
order, feeding method,
breastfeeding duration, parental
education, maternal age, parental
smoking, SES, quality of home
environment, maternal IQ, parents
living together
WISC-R at age 7-8 yr
-0.20 (-0.40, 0.10)
aWISC = Wechsler Intelligence Scale for Children, WPPSI = Wechsler Preschool and Primary Scale of Intelligence, GCI = General Cognitive Index
bEffect estimates are standardized to a 1 |jg/dL increase in blood Pb level within the 10th to 90th percentile interval. Effect estimates are assumed to be
linear within the 10th to 90th percentile interval of blood Pb level. The percentiles are estimated using various methods and are only approximate values.
c95% CI was constructed using a standard error that was estimated for a p-value of 0.01. Authors specified a p-value of <0.01.
The analysis pooling data from seven prospective studies included 1,333 children
ages 5-10 years of age with a median (5th-95th percentile) concurrent blood Pb level of
9.7 (ig/dL (2.5-33.2 (ig/dL) (Lanphear et al.. 2005). In multivariate models that adjusted
for study site, maternal IQ, Home Observation for the Measurement of Environment
(HOME) inventory (assessment of physical environment, parental responsivity, learning
stimulation, emotional climate, and family interactions), birth weight, and maternal
education, higher concurrent, peak, average lifetime, and early childhood blood Pb levels
were associated with lower FSIQ measured at age 5-10 years, with the largest decrement
in FSIQ estimated for concurrent blood Pb level (-0.25 points [95% CI: -0.34, -0.15] per
1 (ig/dL increase in blood Pb level in the 10th to 90th percentile interval [2.2-12.3 (ig/dL]
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of blood Pb level)1. Various models were investigated to characterize the shape of the
blood Pb concentration-response relationship. Consistent with findings from several
individual cohort studies, Lanphear et al. ("2005) found that a log-linear model best fit the
data, with a greater decrease in FSIQ estimated for an increase in concurrent blood Pb
from 2.4-10 (ig/dL (3.9 points [95% CI: 5.3, 2.4]) than an increase from 10 to 20 (ig/dL
(1.9 points [95% CI: 2.6, 1.2]). Among children with concurrent blood Pb less than
10 (ig/dL, the median blood Pb level was 4.2 (ig/dL (Hornung. 2008).
A key additional strength of the pooled analysis by Lanphear et al. (2005) was the
extensive examination of the potential for confounding by several factors related to SES
and the caregiving environment. Variables such as HOME score, birth weight, maternal
IQ, and maternal education were statistically significantly associated with FSIQ and were
included in the final model with blood Pb level. While a smaller decrement in FSIQ was
estimated for concurrent blood Pb level in this adjusted model than in the unadjusted
model (-0.43 points [95% CI: -0.53, -0.33] per 1 (ig/dL increase in blood Pb level in the
10th to 90th percentile interval of blood Pb level), the adjusted blood Pb effect estimate
was nonetheless statistically significant. HOME score was not available in the Rochester
study; however, exclusion of data from that cohort resulted in a 3% less negative effect
estimate, indicating the lack of a strong influence of HOME score alone on blood Pb
level-IQ associations. Other variables such as child sex, tobacco exposure during
pregnancy, alcohol use during pregnancy, maternal age at delivery, marital status, and
birth order were not statistically significantly associated with FSIQ and did not alter the
effect estimate for concurrent blood Pb level. The individual study populations
represented a wide range of SES, maternal education, and cultural backgrounds.
Sensitivity analyses, in which one study was successively excluded, revealed that no
single study was responsible for driving the results. Per 1 (ig/dL increase in blood Pb
level in the 10th to 90th percentile interval of blood Pb level, effect estimates excluding
one study at a time ranged between -0.22 and -0.27, indicating the robustness of the
concurrent blood Pb level effect estimate despite between-study variability in the
distributions of potential confounding factors.
The small number of studies published since the 2006 Pb AQCD continued to
demonstrate associations between higher blood Pb level (primarily concurrent) and lower
FSIQ in children between ages 3 and 11 years (Figure 5-2 and Table 5-3). Similar to
studies reviewed in the 2006 Pb AQCD, most recent studies demonstrated associations
between blood Pb level and lower FSIQ in populations with mean blood Pb level between
5 to 10 (ig/dL. New results from the prospective cohorts were limited. Mazumdar et al.
'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 ng/dL increase in blood Pb level
within the estimated 10th to 90th percentile interval. Effect estimates are assumed to be linear within the 10th to
90th percentile interval of blood Pb level.
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(2011) reported on the follow-up of the Boston cohort to age 28-30 years (55 out of the
original 249 enrolled at birth). Blood Pb levels measured at age 6 months, 4 years,
10 years, and levels averaged over childhood were associated with decrements in FSIQ in
adults. The effect estimates were similar in magnitude for all childhood blood Pb
measures, except for 6 month blood Pb level, which was associated with a smaller FSIQ
decrement. These findings indicated that the effect of childhood Pb exposures may persist
to adulthood.
Jusko et al. (2011; 2008) affirmed the findings in the Rochester cohort previously
reported by Canfield et al. (2003a). who examined the cohort at age 5 years (Figure 5-2
and Table 5-3). Jusko et al. (2011) examined the Rochester cohort at age 6 years and,
similar to Canfield et al. (2003a). found that a 1 (ig/dL increase in concurrent blood Pb
level was associated with a 0.66-point decrease (95% CI: -1.26, -0.06) in FSIQ. Model
covariates were the same as those analyzed by Canfield et al. (2003a): sex, race, family
income, maternal education, race, prenatal smoking, birth weight, transferrin saturation,
maternal IQ, and HOME score. At both age 5 and 6 years, higher peak and lifetime
average blood Pb levels also were associated with lower FSIQ, and incremental increases
in blood Pb levels tended to be associated with larger FSIQ decrements in analyses
restricted to children with blood Pb levels less than 10 (.ig/dL (Jusko et al.. 2008; Canfield
et al.. 2003a). Canfield et al. (2003a) provided additional information on the extent of
confounding by SES- and caregiving-related variables. The effect estimate in the
covariate-adjusted model was 40% smaller than it was in the model with concurrent
blood Pb level alone; however, the association with concurrent blood Pb level remained
statistically significant.
Additional evidence recently 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 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 concurrent blood Pb levels between ages 1 and 10 years. In contrast, Kordas et al.
(2011) found that concurrent blood Pb level was associated with a larger decrement in
FSIQ at age 4 years than was cord blood Pb level. Children in the latter study were born
between 1994 and 1995 and at age 4 years had a mean (SD) blood Pb level of
8.7 (4.4) (ig/dL. In Schnaas et al. (2006). the geometric mean (95% CI) blood Pb level at
age 4 years was 10.3 (4.2, 20.5) (ig/dL. It is not clear whether different temporal patterns
of Pb exposure may have contributed to the contrasting associations for prenatal and
concurrent blood Pb levels in the two studies.
Surkan et al. (2007) examined children from urban Boston, Massachusetts and rural
Farmington, Maine participating in atrial designed to assess the effect of amalgam dental
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fillings on child neurodevelopment. Prior to placement of amalgam fillings, blood Pb
levels were measured, and children were administered an extensive battery of
neuropsychological tests including tests of memory, learning, visual-motor ability,
reading, reaction time. A strength of the study was that the range of concurrent blood Pb
level in this study, 1-10 (ig/dL, was lower than that examined in previous studies. Thus, it
provided information on associations between blood Pb level and IQ decrements within a
lower range of blood Pb levels. However, consistent with several previous studies,
Surkan et al. (2007) found lower FSIQ in the group of children with concurrent blood Pb
levels of 5 to 10 (ig/dL. For example, adjusting for age, race/ethnicity, birth weight, SES,
and primary caregiver IQ, children with blood Pb levels 5-10 (ig/dL had lower FSIQ
scores (-6.04 points [95% CI: -10.7, -1.36]) compared with children who had levels
1-2 |_ig/dL (referent group). Children with blood Pb levels 2-5 (ig/dL did not have FSIQ
scores (-0.12 [95% CI: -3.30, 3.06]) that differed from children with blood Pb levels
1-2 (ig/dL. Another strength of this study was the analysis of potential confounding by a
larger list of variables than that included in the final models. Although HOME score was
not examined, several variables such as SES, caregiver education, parenting stress,
marital status of caregiver, and maternal utilization of prenatal or annual health care were
not significantly (p > 0.20) associated with child IQ either independently or in a model
with blood Pb level (Surkan et al.. 2007).
Other recent studies found blood Pb-associated lower FSIQ in populations of children in
Asia with lower blood Pb levels (Kim et al.. 2009b; Zailina et al.. 2008) than were
previously examined. Kim et al. (2009^) examined children ages 8 to 11 years in Korea
(born mid- to late-1990s) with a mean (range) concurrent blood Pb level of 1.73
(0.42-4.91) (ig/dL. Children were tested using the Korean Educational Development
Institute-WISC, which assesses vocabulary, arithmetic, picture arrangement, and block
design. In a log-linear regression analysis adjusted for age, sex, maternal and paternal
education, yearly income, prenatal smoking, postnatal environmental tobacco smoke
exposure, birth weight, and maternal age at birth, a 1 (ig/dL higher concurrent blood Pb
level was associated with a 0.64-point lower (95% CI: -1.05, -0.19) FSIQ within the 10th-
90th percentile interval of blood Pb level (< 1 to 2.8 (ig/dL). The potential confounders
examined in this study were not found to have a strong influence on the findings as a
similar magnitude of effect was estimated in a model that included only blood Pb level
(-0.73 points [95% CI: -1.19, -0.27] per 1 (ig/dL increase in blood Pb level in the 10th to
90th percentile interval). Although several important SES-related confounders were
considered, there was no direct assessment of the home environment and the primary
caregiver IQ in this study, which are notable limitations.
Kim et al. (2009^) also examined effect modification of the blood Pb-FSIQ relationship
by concurrent blood manganese (Mn) levels. The mean (range) blood Mn level was
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14.3 (5.3-29.02) fig/dL. Blood Pb and Mn levels were not correlated (r = 0.03. p = 0.64).
To examine effect modification, children were divided into two groups: blood Mn level
above and below the median (14 Lig/dI .). Multivariate linear regression models predicting
FSIQ and verbal IQ (VIQ) used concurrent blood Pb level as the predictor variable in the
low and high Mn groups. As illustrated in Figure 5-3, the associations of concurrent
blood Pb level with FSIQ and VIQ was larger in magnitude and statistically significant
for children in the high Mn group compared with children m the low Mn group (-0.65
points [95% CI: -1.26, -0.05] in high Mn group versus -0.50 points [95% CI: -1.23, 0.22]
in low Mn group per 1 ug/dL increase in blood Pb level in the 10th to 90th percentile
interval). However, higher blood Pb level was associated with lower FSIQ in children in
the low Mn group.
i-.: l..i
um: i uLmiiiiiimimiifm
LUIJI
[ I.
0.0	0.2	04
Blood Lead Concentration
0.0 0.2 0.4
Blood Lead Concentration
Source; Reprinted with permission of Elsevier Science, Kim et al. (2009b).
Note: High and low Mn refer to levels above and below the median of 14 pg/dL, respectively.
Figure 5-3 Effect modification of the association between concurrent blood
Pb level and FSIQ by blood Mn level.
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Higher concurrent blood Pb level also was associated with a lower IQ score (as assessed
by the McCarthy General Cognitive Index) among children ages 6-8 years in Malaysia
with a mean blood Pb level of 3.69 j^ig/dL (Zailina et al.. 2008). The concentration-
response function was not reported; however, in a multivariate model, blood Pb level and
not potential confounding factors such as paternal education, maternal education,
household income, number of siblings, and birth order was reported to be significantly
associated with FSIQ. As with Kim et al. (2009b). a notable limitation was the lack of
consideration of confounding by the caregiving environment.
Pb has long been known to impact dopaminergic neurons, inhibit depolarization-evoked
neurotransmitter release, and stimulate spontaneous neurotransmitter release
(Section 5.3.8.8). Further, dopaminergic activity is a key process in regulating cognitive
function. Thus, variants in dopamine-related genes have the potential to act as effect
modifiers of Pb-associated neurodevelopmental effects. Recent studies in children that
examined effect modification by variants in dopamine genes produced contrasting results.
In a study of children at ages 2 and 4 years, Kordas et al. (2011) did not find differences
in association between concurrent blood Pb level and FSIQ by the Taq Al/Al dopamine
receptor 2 (DRD2) genotype, which is associated with reduced dopamine receptor
density. Roy et al. (2011) examined children of similar age (3-7 years) in Chennai, India
and found that among children with the Taq Al/Al genotype, a 1 (ig/dL higher blood Pb
level was associated with a 0.84-point lower (95% CI: -1.66, -0.01) FSIQ within the 10th-
90th percentile interval of blood Pb level (5.8-18.3 (ig/dL). The same increment in blood
Pb level was associated with a 0.36-point lower (95% CI: -0.76, -0.04) FSIQ in children
with the Taq A2/A2 higher receptor density genotype. Concurrent blood Pb level was
associated with FSIQ in both studies in analyses including all subjects. However, there
were many differences between study populations that may have contributed to
differences in effect modification by the DRD2 variant. Compared with the group in
India, the group in Mexico had a lower mean blood Pb level and lower mean FSIQ score.
Additionally, children in the Mexico City group with the Taq Al/Al genotype had a
higher mean FSIQ score, whereas FSIQ scores were similar between Taq genotypes in
the children in India.
Other recent studies examined children in the U.S.; however, in these studies, there were
high proportions of children with prenatal exposure to alcohol, tobacco, marijuana, or
cocaine that may limit the generalizability of findings. Chiodo et al. (2007) examined a
population of African American children (age 7 years, born 1989-1991) in the Detroit,
MI area with relatively low blood Pb level (mean [SD]: 5.0 [3.0]) but high prevalence of
prenatal exposure to cocaine (38%) or marijuana (35%). In a detailed analysis of potential
confounding factors, investigators found that cocaine exposure did not meet the criterion
for inclusion in the model (i.e. ,p>0 .10 for association with FSIQ). Prenatal marijuana
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exposure was associated with FSIQ (p < 0.10) and was included in a multivariate model
along with caretaker education, SES, HOME score, symptom checklist, number of
children in the home, maternal IQ, sex, and prenatal cigarette smoking. In this
multivariate model, higher blood Pb level was associated with lower FSIQ (Figure 5-2
and Table 5-3). These findings supported an independent association of blood Pb level
despite the high prevalence of prenatal drug exposure in the study population.
Min and colleagues (2009) followed children in Cleveland, OH from age 4 to 11 years
who were exposed prenatally to alcohol (77%), cigarette smoking (61%), marijuana
(31%), or cocaine (51%). The study population was primarily African-American (86%)
and of low SES (98%); 39% of mothers had not finished high school, and 86% were
unmarried at the time of enrollment. The mean (range) blood Pb level, measured at age
4 years, was 7.0 (ig/dL (1.3-23.8). Several lines of evidence indicated that prenatal drug
exposure did not heavily influence the association between blood Pb level and FSIQ.
Prenatal alcohol exposure was correlated with blood Pb level but was not statistically
significantly associated with IQ (p > 0.10) and did not change the blood Pb level effect
estimate by more than 10%. Prenatal cocaine exposure was associated with FSIQ at age
9 years; however, blood Pb level at age 4 years remained associated with a decrement in
FSIQ at age 9 years after adjusting for cocaine exposure and several SES-related factors
(Table 5-3). Based on the examination of interaction terms, associations between blood
Pb level and FSIQ were not statistically significantly different between drug-exposed and
-unexposed children.
Similar to previous studies with repeated assessments of cognitive function over time,
Min et al. (2009) found that the association between blood Pb level at age 4 years and
FSIQ persisted to older ages. Higher blood Pb level at age 4 years was associated with
similar magnitudes of decrements in FSIQ at ages 4, 9 and 11 years (Figure 5-2 and
Table 5-3). Researchers also examined the shape of the concentration-response function
using a restricted cubic spline function. Although the cubic spline term did not attain
statistical significance (p = 0.19), qualitative analysis indicated that the association
between blood Pb level and FSIQ decrements persisted and was greater at lower levels
blood Pb (< 7 (ig/dL). These findings were consistent with those from the pooled analysis
(Lanphear et al.. 2005) and other individual studies (Tellez-Roio et al.. 2006).
Bayley Scales of Infant Development
The Bayley Scales of Infant Development are the most widely used tests of infant
intelligence. The Mental Development Index (MDI) is statistically analogous to IQ,
i.e., both scores have a population-standardized mean of 100 and standard deviation of
15. While MDI assesses general cognitive function in infants, it is important to note that
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MDI scores are not necessarily correlated with IQ scores measured later in childhood.
Associations between blood Pb levels and MDI scores were found in most of the
prospective cohort studies in children between age 6 months and 3 years and mean blood
Pb levels 5-10 (ig/dL (Table 5-4). Recent studies examined children who were born in the
mid-1990s to mid-2000s and continued to demonstrate associations between higher cord
and concurrent blood Pb level and lower MDI scores (Table 5-4). Studies that found
associations with concurrent blood Pb levels also tended to find associations with
prenatal cord or maternal blood Pb levels. Recent studies provided new information on
effect modification by nutritional status, maternal self-esteem, co-exposure to Mn, and
genetic variants. While studies adjusted for multiple SES-related variables including
maternal IQ and education, most did not consider confounding by the caregiving
environment. Concurrent and cord blood Pb levels were associated with MDI, adjusting
for HOME score in diverse cohorts (Solon et al. 2008; Wasserman et al. 1992; Bellinger
et al.. 1987). In the Cleveland cohort, associations of cord and postnatal blood Pb levels
with MDI at ages 6 months to 2 years became null after adjusting for covariates including
HOME score (Ernhart et al.. 1988; Emhart et al.. 1987). In the analysis of the Cincinnati
cohort, HOME score was not significantly associated with blood Pb level or MDI and
thus, did not meet the criteria for model inclusion (Dietrich et al. 1987a). Collectively,
evidence does not indicate that Pb-associated MDI decrements are driven by confounding
by quality of the caregiving environment.
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Table 5-4 Associations of blood Pb level with Bayley MDI in children ages
6 months to 3 years3
Study
Population/
Location
Blood Pb Levels
(Hg/dL)
Statistical Analysis
Cognitive Index
Effect Estimate
(96% Cl)b
Jedrychowski et
al. (2009b)
444 children born
2001-2004 followed
prenatally to age 36
mo
Krakow, Poland
Prenatal (cord blood)
Geometric mean
(range): 1.29 (0.44-5)
10th-90th: 1.2-1.3
Log linear regression model adjusted for	at age 12 mo
maternal education, birth order, prenatal	at aae 24 mo
smoking, sex, and within-subject MDI
correlation	a' a9e 36 mo
-1.9 (-3.8, 0.12)
-2.6 (-5.0, -0.21)
-2.3 (-4.3, -0.30)
Claus Henn et
al.
455 children born
1997-2000 followed
prenatally to age 36
mo
Mexico City
12 month
Mean (SD): 5.1 (2.6)
10th-90th: 2.5-8.4
Linear mixed effects regression adjusted
for sex, hemoglobin, gestational age,
maternal IQ, maternal education, blood
Pb-blood Mn interaction
assessed ages 12 to 36 mo
12 mo blood Mn < 2.0 pg/dL
12 mo blood Mn 2.0-2.8 pg/dL
12 mo Mn > 2.8 pg/dL
-3.0 (-5.22, -0.78)
-0.07 (-0.39, 0.25)
-2.2 (0, 4.44)
Tellez-Rojo et
al. 12006)
294 children born
1997-1999 followed
prenatally to age 24
mo
Mexico City
Same cohort as above
Concurrent
12 month
Mean (SD): 4.7 (2.9)
10th-90th: 1.9-8.2
24 month
Mean (SD): 5.3 (4.1)
10th-90th: 2.1-10.7
Log linear regression model adjusted for at age 12 mo
sex, age, birth weight, maternal IQ, at age 24 mo
cohort
-0.26 (-0.79, 0.26)
-0.89 (-1.32, -0.46)
Dietrich et al.
(1993a)
96 to 302 children born
1979-1984 followed
prenatally to age 6 mo
Cincinnati, OH
Prenatal (cord)
Mean (SD): 6.3 (4.5)
10th-90th: 2.3-11.7
Neonatal (10 day)
Mean (SD): 4.6 (2.8)
10th-90th: 1.9-8.1
3 month
Mean (SD): 5.9 (3.4)
10th-90th: 2.6-10.1
Linear regression model adjusted for
birth weight, gestation, maternal age,
race, sex, SES
assessed at 6 mo
Cord blood Pb
Neonatal blood Pb
3-month blood Pb
-0.66 (
-3.49 (
-0.48 |
¦1.4, 0.07)
¦6.0, 0.96)
¦1.0, 0.05)
Bellinger etal.
(1987)
249 children born
1979-1981 followed
prenatally to age 24
mo
Prenatal (cord)
Low: < 3
Medium: 6-7
High: >10
Regression model adjusted for maternal
age, race, maternal IQ, maternal
education, number of years of smoking,
number of alcohol drinks per week in 3rd
trimester of pregnancy, SES, HOME,
sex, birthweight, gestational age, birth
order
assessed ages 6 to 24 mo
High vs. low cord blood
High vs. medium cord blood
-4.8 (-7.3, -2.3)
-3.8 (-6.3, -1.3)
Surkan et al.
309 children ages
12-36 mo during
1996-2001 or
2004-2005
Mexico City, Mexico
Concurrent
Mean (SD): 6.4 (4.3)
10th-90th: 2.0-12.4
Linear mixed effects regression model
adjusted for sex, maternal age, maternal
IQ, maternal education, parity, alcohol
consumption, smoking, cohort, maternal
self-esteem
assessed ages 12 to 36 mo
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)
Pilsner etal. 255 children age 24 Prenatal (cord blood) Linear regression model adjusted for at age 24 mo	-0.73 (-1.2,-0.23)
(2010)	mo born 1994-1995 Mean (SD): 6.7 (3.6) maternal age, maternal IQ, marital
Mexico City, Mexico 10th-90th: 3.5-10.5 sta'us> gestational age
inadequate folate intake, MTHFR
genotype
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Study
Population/
Location
Blood Pb Levels
(Hg/dL)
Statistical Analysis
Cognitive Index
Effect Estimate
(96% Cl)b
assessed at age 24 mo


Prenatal 1st trimester
-0.64 i
[-1.3, -0.03)
Prenatal 3rd trimester
-0.38 i
[-1.0, 0.24)
Cord blood
-0.06 i
[-0.82, 0.70)
Concurrent
-0.23 i
[-0.92, 0.45)
Hu et al. (
146 children born
1997-1999 followed
prenatally to age 24
mo
Mexico City, Mexico
Prenatal (maternal
blood Pb) in 1st
trimester
Mean (SD): 7.1 (5.1)
10th-90th: 2.5-13.1
3rd trimester
Mean (SD): 6.9 (4.2)
10th-90th: 2.8-12.1
Cord blood
Mean (SD): 6.2 (3.9)
10th-90th: 2.5-11.0
Concurrent
Mean (SD): 4.8 (3.7)
10th-90th: 1.6-9.1
Log linear regression model adjusted for
sex, maternal age, current weight,
height-for-age Z score, maternal IQ,
concurrent blood Pb (in models
examining blood Pb at other lifestages)
Solon et al.
502 children ages 6-35
mo, born 1997-2004
Visayas, Philippines
Concurrent
Mean (SD): 7.1 (7.7)
10th-90th: 1.6-14.9
Two-stage linear regression model to
account for determinants of blood Pb
(sex, roof materials, water source,
breastfed for > 4 months) and cognitive
function (HOME score, maternal
education, maternal smoking, born
premature, region of residence)
assessed ages 6 to 35 mo
-3.32 (-5.02, -1.60)
Vimpani et al.
592 children followed
prenatally to age 24
mo
Port Pirie, Australia
20% subjects had
24-month blood Pb
levels > 30
Linear regression model adjusted for
maternal age, paternal education,
maternal education, paternal workplace,
maternal workplace, parental
relationship, maternal prenatal marital
status, child birth rank, mouthing activity,
oxygen use at birth, apgar score,
neonatal jaundice, size for gestational
age, maternal IQ
at age 24 mo
Maternal avg prenatal blood Pb -0.64
Cord blood Pb	0.03
6 mo blood Pb	-0.40, p < 0.05
24 mo blood Pb	-0.06
Lifetime avg	-0.31, p < 0.05
Wfesserman et
al. (
392 children followed
prenatally to age 24
mo
Kosovska Mitrovica
and Pristina,
Yugoslavia
Concurrent Means:
35.5 (K. Mitrovica)
8.4 (Pristina)
Log linear regression model adjusted for
sex, birth order, birth weight, ethnic
group, HOME, maternal education,
maternal age, maternal IQ
at age 24 mo
Cord blood Pb
6 mo blood Pb
12 mo blood Pb
18 mo blood Pb
24 mo blood Pb
Per tripling blood Pb
-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
MDI = Mental Development Index, MTHFR = methylenetetrahydrofolate reductase
aStudies are presented in order of increasing population mean blood Pb level.
bExcept where noted, effect estimates are standardized to a 1 |jg/dL increase in blood Pb level within the 10th to 90th percentile interval of blood Pb level.
In a population of children in Krakow, Poland with lower cord blood Pb levels (median
1.23 (ig/dL, 95% CI: 1.24, 1.34 (ig/dL) than those previously examined, increasing cord
blood Pb level was associated with similar magnitudes of decrease in MDI at ages 12, 24,
and 36 months (Jedrvchowski et al.. 2009b) (Table 5-4). Consistent with the hypothesis
that the developing male central nervous system may be more vulnerable than that of
females to environmental insults resulting in later behavioral problems (Moffitt et al..
2001). investigators estimated a larger in MDI per unit increase in cord blood Pb level
among the 233 males than among the 223 females (Figure 5-4). Although mean cord
blood Pb levels were similar between males (1.35 (ig/dL) and females (1.41 (ig/dL), the
mean 36-month MDI score was lower among males than among females (101 and 105,
respectively, p = 0.0001).
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Boys
Girls
m
o -
03
o
o
CO o
co 2
co
in
o
o
CT>
Beta coeff = -10.6,
p = 0.013
Beta coeff. = -4.9,
P = 0.244
-.5
-r—
.5
1 -.5	0
Pb level {log transformed)
-i—
.5
95% CI
Fitted values
Graphs by gender
Source: Reprinted with permission of Elsevier Science, Jedrychowski et al. (2009a).
Figure 5-4 Regression of fitted MDI score at 36 months on log-transformed
cord blood Pb level by sex.
Solon et al. (2008) examined children in the Philippines, ages 6 to 35 months. A key
strength of this study was the adjustment for HOME score as well as maternal education,
maternal smoking, premature birth, region of residence, and years of schooling of child.
Not only did investigators find an association between higher concurrent blood Pb level
and lower MDI score (Table 5-4), but they also found effect modification by red blood
cell folate levels. Among children with folate levels less than or equal to 230 (ig/mL,
blood Pb level had a marginal association with lowering MDI scores in the range of 0.80
to 2.44 points. Among children with higher folate levels, blood Pb level was not
estimated to have a negative marginal impact. These findings from Solon et al. (2008)
indicated that children with folate deficiencies may be at increased risk of Pb-associated
decreases in cognitive function. The results were consistent with observations that higher
folate level is associated with lower blood Pb level since folate improves Pb excretion by
inhibiting the binding of Pb to blood elements.
Multiple studies in different Mexico City mother-child pair cohorts recently reported
associations between blood Pb levels (e.g., maternal, cord blood, or child postnatal) and
decrements in Bayley MDI in children between age 12 and 36 months (Claus Henn et al.;
Pilsner et al.. 2010; Surkan et al.. 2008; Hu et al.. 2006; Tellez-Roio et al.. 2006). Recent
studies extended findings from Tellez-Rojo et al. (2006) with follow-up of the same
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cohort to age 3 years and observations of effect modification by maternal self-esteem
(Surkan et al. 2008) and blood Mn levels (Claus Henn et al.).
Claus Henn et al. examined 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. Among subjects with age 12-month blood Mn levels less than 2.0 (ig/dL, a
1 (ig/dL higher age 12-month blood Pb level was associated with a 3.0-point (95% CI: -
5.2, -0.78) lower MDI. Among subjects with blood Mn levels greater than 2.8 (ig/dL, a
1 (ig/dL higher age 12-month blood Pb level was associated with a 2.2-point (-4.4, 0)
lower MDI. Interactions were not found using age 24-month blood Mn and Pb levels.
Effects estimates were similar in magnitude before and after adjustment for sex,
gestational age, hemoglobin, maternal IQ, maternal education, and visit. Findings in this
Mexico City cohort added to those from Kim et al. (2009b) in older children that
associations between blood Pb level and cognitive function may be influenced by Mn co-
exposure, and results from both studies indicated that blood Pb level was associated with
decrements in cognitive function with both lower and higher levels of blood Mn.
Biological plausibility for the interactive effects of Pb and Mn exposure is provided by
observations that Mn has similar modes of action and cellular targets as does Pb,
i.e., altering Ca+2 metabolism, inducing oxidative damage to neuronal cells, diminishing
dopamine transmission. Claus Henn et al. also indicated that Pb-Mn co-exposure may
affect cognitive function in children as young as 12 to 36 months of age.
Surkan et al. (2008) stratified data by the level of maternal self-esteem as reported by
mothers. Higher age 24-month blood Pb level was associated with lower MDI score
among children with mothers in the lowest three quartiles of self-esteem but not among
children with mothers in the highest quartile of self-esteem (Table 5-4). Model covariates
included cohort, sex, maternal IQ, maternal age, maternal education, parity, maternal
smoking, maternal alcohol consumption, and maternal self-esteem. While HOME score
was not examined, maternal self-esteem may incorporate many of the same aspects as
HOME score, including mother-child interactions, educational interactions, and dietary
practices. These findings indicated that higher maternal psychosocial functioning
(e.g., lower stress, anxiety, and depression and higher self-esteem) may contribute to
better caregiving, which in turn may improve neuropsychological functioning of the
child. The biological plausibility of these findings in children is well-supported by
observations in animals unrelated to Pb that environmental enrichment reverses the
effects of early stress experiences on reactions such as depressed behavior, HPA
activation, and immunosuppressant (Laviola et al.. 2008; Laviola et al.. 2004; Morelev-
Fletcher et al.. 2003; Francis et al.. 2002). With specific regards to Pb exposure, social
isolation or enrichment has been found to exacerbate or protect against, respectively, Pb
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exposure-induced learning impairments in animals (Section 5.3.2.2). It is worth
mentioning in this context that the potential programming effects of stress on childhood
health outcomes may occur at an even more fundamental level, i.e., through epigenetic
programming (Dolinov and Jirtle. 2008V
Hu et al. (2006) compared associations among prenatal blood Pb levels measured at
different stages of gestation among 146 mother-child pairs meeting the following criteria:
born at 37 weeks or greater gestational age, at least one valid Pb measurement during
pregnancy, complete information on maternal age and IQ, and measurement of child
blood Pb level at 24 months when the 24-month MDI was ascertained. Among blood Pb
levels measured at various lifestages, 1st trimester maternal blood Pb level (whole or
plasma Pb) was associated with larger decreases in subsequent 24-month MDI scores
compared with maternal 3rd trimester, cord, and child concurrent blood Pb (Table 5-4).
These results were adjusted for sex, 24-month blood Pb level, height-for-age Z score,
weight, maternal age, and maternal IQ. HOME score was not included in the final model;
however, investigators examined the potential for confounding by a larger list of
unspecified variables. Prenatal Pb exposure effects also were indicated in a study
conducted in Mexico City that examined FSIQ repeatedly during ages 6-10 years
(Schnaas et al.. 2006). Third trimester (weeks 28-36) maternal blood Pb level, was
associated with larger decreases in FSIQ than were maternal blood Pb level at
12-20 weeks, maternal blood Pb level at delivery, or concurrent child blood Pb levels
measured between ages 6 and 10 years.
In another recent study in Mexico City, Mexico, investigators found higher cord blood Pb
level to be associated with a lower MDI score in children at age 24 months (-0.73 points
[95% CI: -1.2, -0.23] in score per 1 (ig/dL increase in cord blood Pb level in the 10th-
90th percentile interval [3.5-10.5 j^ig/dL|) (Pilsner et al.. 2010). Investigators additionally
examined effect modification by variants in the methylenetetrahydrofolate reductase
(MTHFR) gene. The MTHFR enzyme is involved in folate metabolism, which, in turn, is
involved in homocysteine methylation to the amino acid methionine. The transfer of
methyl groups that results from folate metabolism is important for biological processes
including Phase II detoxification reactions and epigenetic regulation of gene expression.
The MTHFR gene has common functional variants, including the C677T SNP, which
produces an enzyme with lower metabolic activity and is associated with lower serum
folate levels (Kordas et al.. 2009). Lower folate levels have been associated with higher
blood Pb levels. Although Pilsner et al. (2010) found that both cord blood Pb levels and
the MTHFR C677T allele were associated with lower child MDI score at age 24 months,
they did not find a statistically significant interaction between blood Pb level and the
MTHFR 677T allele. Results from stratified analyses were not reported, thus differences
in the magnitude of association between genotypes could not be compared.
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Specific Indices of Cognitive Function in Children
In addition to indices of global cognitive function, blood Pb levels also have been
associated with specific cognitive abilities, including memory, learning, visuospatial
processing, and other executive functions such as planning and problem solving in
children and adolescents (kordas et al.. 2006; Tellez-Roio et al.. 2006; Chiodo et al..
2004; Ris et al.. 2004; Can field et al.. 2003b; Lanphear et al.. 2000; Bellinger and Stiles.
1993; Dietrich et al.. 1992; Bellinger et al.. 1991; Dietrich et al.. 1991; Needleman et al..
1979) (Figure 5-5 and Table 5-5). Studies often found associations with multiple indices
of cognitive function, and because several tests are interrelated, it is difficult to attribute
the effects of Pb exposure to a specific domain of cognitive function. Evidence that blood
Pb levels are associated with a spectrum of cognitive indices provides coherence for the
associations observed between blood Pb levels and FSIQ, a global measure of cognitive
function. Furthermore, these tests of learning and memory in humans are homologous to
tests in animals, and compared with that for FSIQ, evidence for these specific indices
may improve understanding of the coherence between findings in humans and animals
(Rice. 1996). These complex cognitive functions have been described to be mediated by
the actions of neurotransmitters dopamine and glutamate in the hippocampus, prefrontal
cortex, and ventral striatum of the brain. In toxicological studies, Pb exposure has been
shown to disrupt the function of these systems, which provides biological plausibility for
associations observed between blood Pb levels and deficits in a range of executive
functions in children.
Studies published since the 2006 Pb AQCD added to the evidence for associations
between higher blood Pb level (concurrent most frequently examined) and lower
performance in specific indices of cognitive function (Figure 5-5 and Table 5-5). The
weight of evidence indicated that decrements in specific cognitive indices were
associated with concurrent blood Pb level and in populations with mean blood Pb levels
between 5-10 (ig/dL. Most studies considered potential confounding by multiple SES-
related variables, and many found blood Pb-associated decrements in cognitive function
in both unadjusted and covariate-adjusted models (Palaniappan et al.. 2011; Chiodo et al.
2007; Froehlich et al.. 2007). Many studies did not examine potential confounding by
HOME score. However, studies that did (Min et al.. 2009; Solon et al.. 2008; Froehlich et
al.. 2007; Can field et al. 2004; Can field et al. 2003b) and did not adjust for HOME score
(krieg et al.. 2010; Jedrvchowski et al.. 2008; Surkan et al.. 2007; Lanphear et al. 2000)
both found associations between blood Pb level and decrements in specific cognitive
indices.. Several studies found associations after adjusting for different potential
confounding factors, including parental IQ or education and HOME score. The set of
covariates varied among the specific cognitive endpoints, and HOME score was not
associated with every cognitive index. For example, only household income remained
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significantly associated with total errors in the spatial working memory task and was
included in the final model (Froehlich et al. 2007). Solon et al. (2008) found that in a
multivariable model with blood Pb level and several potential confounders, HOME score
remained significantly associated with verbal IQ in 5 year-old children, whereas maternal
education, maternal smoking, and premature birth did not. Combined, these findings
support the independent association of blood Pb level with specific cognitive indices and
indicate that the potential confounding variables vary among endpoints and study
populations.
Several analyses of the Rochester cohort at ages 4 and 5 years indicated associations of
blood Pb level with various indices of cognitive function, including visual and working
memory, planning, and problem solving (Froehlich et al.. 2007; Canfield et al. 2004;
Can field et al.. 2003b'). In particular, the associations observed with tests of memory and
rule learning and reversal provided strong coherence with observations of Pb-induced
impaired performance in animals in homologous tests of the Morris Water Maze, the
delayed spatial alternation, and discrimination reversal learning (Section 5.3.2.2).
Compared with other prospective cohort studies, findings from the Rochester cohort
demonstrated associations between blood Pb level and decrements in executive functions
in younger children with a lower mean blood Pb level (6.5 and 6.0 (ig/dL at ages 4 and
5 years, respectively). An additional strength of these studies was the detailed analysis of
potential confounding by various demographic and SES-related variables (including
HOME score) and other environmental exposures. Covariates were selected for specific
models based on their association with a particular test (p < 0.20), and several models
adjusted for maternal IQ, maternal education, and HOME score.
Canfield et al. (2003b) reported associations of concurrent blood Pb level with various
parameters of the Shape School tasks, which test the ability of children to recall correctly
shapes and colors that are attributed to specific cartoon characters. Concurrent blood Pb
level was associated with color knowledge, shape knowledge, and number of correct
responses at age 4 years (Figure 5-5 and Table 5-5). At age 5 years, higher lifetime
average blood Pb level (area under the curve calculation using repeat measurements
between age 6 months and 5 years) was associated with poorer performance on multiple
tasks related to learning, working and spatial memory, and planning as assessed by the
Cambridge Neuropsychological Testing Automated Battery (CANTAB) in both
unadjusted and covariate-adjusted analyses (Canfield et al.. 2004). These memory tests in
children share homology with the Morris water maze test in animals in that both test the
ability of subjects to use efficient search strategies to identify target locations, retain
spatial information, and access remembered items in working memory. Both measure the
time to complete the task and the number of errors made. At age 4 and 5 years,
associations of concurrent blood Pb level with several indices of executive function were
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robust to additional adjustment for attentiveness and child FSIQ, indicating that higher
blood Pb level is associated with specific aspects of associative learning that are not
represented by attentiveness and general knowledge base.
Study	Blood Pb Mean 10th-90th percentile
(SD) (pq/dL) interval ([jg/dL)
Choetal. (2010)	1.9(0.67)	1.2-2.8
Lanphearetal. (2000) 1.9(geometric mean) 1.74-2.06
Kreig etal. (2010)
1.95(0.16)
Chiodo etal. (2007)	5.0(3.0)
Chiodo etal. (2004)	5.4(3.3)
Froehlich etal. (2007)	6.1 (4.9)
Canfield etal. (2004)	7.2(3.6)
Canfield etal. (2003b)	6.5
Solon etal. (2008)
Minetal. (2009)
1.69-2.19
2.1-8.87
2.3-9.5
1.9-11.7
3.5-11.8
7.7(7.7)
7.1(5.1)
Data not available
1.6-14.9
7.0(range: 1.3-23.8) 3.0-12.1
Palaniappan etal. (2011) 11.5(5.3)
6.0-18.3
Outcome
Color-word score3
DigitSpan
Math score
DigitSpan
Reading score
Math score
Arithmetic13
Objectassembly3
Spatialspanb
Perseverativeerrorsb
Spatial memory3
Rule learning3
Spatialspan3
Nontargeterrorsc
Total errors0
Proportion correct
Colorknowledge
Verbal IQ
Performance IQ
Verbal comprehension, age 9
Verbalcomprehension.age 11
Math score,age9
Math score, age 11
Reading score,age9
Reading score, age 11
Visuomotorcomposite
-0.6
-0.4
-0.2
0.2
Change in standardized test score per 1 |jg/dl_ increase in blood Pb level
within the 10th-90th percentile interval (95% CI)
aStandard error was estimated from p-value.
Sufficient data were not provided to calculate 95% CIs.
cDirection of effect estimate was changed to indicate that a negative estimate represents poorer performance and a positive
estimate represents improved performance. Black diamonds, gray squares, and blue circles represent associations with concurrent,
lifetime average, and earlier childhood blood Pb levels
Note: Regression coefficients were standardized to their standard deviation to facilitate comparisons among tests with different
scales. Studies generally are presented in order of increasing mean blood Pb level. Effect estimates are standardized to a 1 pg/dL
increase in blood Pb level within the 10th-90th percentile interval. The percentiles are estimated using various methods and are only
approximate values. Effect estimates are assumed to be linear within the 10th to 90th percentile interval of blood Pb level.
Figure 5-5 Standardized regression coefficients describing the associations
of blood Pb levels with specific indices of cognitive function in
children.
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Table 5-5 Additional characteristics and quantitative results for studies
represented in Figure 5-5
Study
Population/
Location
Blood Pb Levels
(Hg/dL)
Statistical Analysis
Effect Estimate
Cognitive Index (96% Cl)a
Cho et al.
(2010)
667 children ages 8-11 yrin
2008
Five Korean cities
Concurrent
Mean (SD): 1.9
(0.67)
10th-90th: 1.2-2.8
Log linear regression model
adjusted for age, sex, parental
education, maternal IQ, child IQ,
birth weight, urinary cotinine
Color-word scorec 0 (-0.09, 0.08)"
using Stroop test at ages 8-11 yr
Lanphear et
al. (2000)
4,852 children ages
6-16 years
U.S. NHANES III
(1991-1994)
Concurrent
Geometric mean: 1.9
(95% CI: 1.70, 2.10)
10th-90th: 1.74, 2.06
Linear regression model adjusted
for sex, race/ethnicity, poverty index
ratio, reference adult education,
serum ferritin levels, serum cotinine
levels
Digit Span (WISC-R) -0.02 (-0.03, -0.01)
Math Score (WRAT-R) -0.007 (-0.01, -0.001)
at ages 6-16 yr
Krieg et al. 842 children ages 12-16 yr
(2M) U.S. NHANES III
(1991-1994)
Concurrent
Mean (95% CI): 1.95
(1.63, 2.27)
10th-90th: 1.69-2.19
Log linear regression model
adjusted for sex, caregiver
education, family income, race-
ethnicity, test language
Block design (WISC-R)
Digit span (WISC-R)
Reading score (WRAT-R)
Math score (WRAT-R)
assessed at ages 12-16 yr
-0.10 (-0.18, -0.02)
-0.08 (-0.13, -0.04)
-0.11 (-0.16, -0.05)
-0.07 (-0.12, -0.01)
Chiodo et al.
506 African-American
children in Detroit, Ml area
followed from birth
(1989-1991) to age 7 yr.
Large proportions of
children with prenatal
exposure to cocaine or
marijuana
Concurrent
Mean (SD): 5.0 (3.0)
10th-90th: 2.1-8.7
Regression model adjusted for
SES, caretaker education, maternal
IQ (all outcomes)
Plus: age, HOME, Symptom
Checklist (arithmetic and mazes)
Plus: sex, prenatal cigarettes/day,
prenatal marijuana use (arithmetic)
Arithmetic
Mazes
Object assembly
WISC-III at age 7 yr
-0.07, p > 0.05c
-0.11 (-0.18, -0.04)
-0.13 (-0.21,-0.05)
Chiodo et al. 246 African-American
(2004) children in Detroit, Ml area
followed from birth (not
reported) to age 7.5 yr.
Large proportions of
children with prenatal
exposure to cocaine or
marijuana
Concurrent
Mean (SD): 5.4 (3.3)
10th-90th: 2.3-9.5
Log linear regression model
adjusted for SES, maternal
education, sex (all outcomes)
Family Environmental Symptom
(family functioning), number of
children < 18 years, caregiver
vocabulary, prenatal alcohol use
(Perseverative errors)
Prenatal cocaine use (conceptual
level responses)
HOME score, Symptom Checklist,
caregiver age (Spatial span)
Corsi Spatial Span
WCST Perseverative errors
WCST Conceptual Level
Responses
at age 7.5 yr
-0.10, p > 0.05cd
-0.49, p > 0.05c,d
-0.77 (-1.5, 0)b
Froehlich et
al. (2007)
174 children born
1994-1995 followed from
birth to age 5 yr
Rochester, NY
Concurrent
Mean (SD): 6.1 (4.9)
10th-90th: 1.9-11.7
Linear regression model adjusted
for income (spatial memory); NICU,
sex (rule learning); HOME score,
maternal IQ, race (spatial span); or
maternal IQ, transferrin saturation
(problem solving)
Spatial memory
Rule learning and reversal
Spatial span
Problem solving
using CANTAB at age 5 yr
-0.02 (-0.05. 0.01)"
-0.03 (-0.06, -0.001)"
-0.007 (-0.01,0)"
-0.04 (-0.09, 0.01)"
Canfield et 174 children 174 children
al. (2004) born 1994-1995 followed
from birth to age 5 yr
Rochester, NY
Lifetime avg
Mean (SD): 7.2 (3.6)
10th-90th: 3.5-11.8
Linear regression model adjusted
for maternal IQ, duration of
breastfeeding, household income,
maternal ethnicity, first prenatal visit
(spatial span length), NICU
admission, sex, spatial span length,
spatial working memory problem
(total nontarget errors, total errors),
age at testing (nontarget errors),
Maximum spatial span length
Total nontarget errors
Total errors
Using CANTAB at age 5 yr
-0.17 (-0.25, 0.59)"
-0.09 (-0.11,-0.06)"
0.04 (-0.07, 0.007)"
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Study
Population/
Location
Blood Pb Levels
(Hg/dL)
Statistical Analysis
Cognitive Index
Effect Estimate
(96% Cl)a
Canfield et 170 children born
al. (2003b) 1994-1995 followed from
age 6 mo to age 4 yr
Rochester, NY
Concurrent
Mean: 6.5
10th-90th: data not
available
Linear mixed effects model adjusted
for:
Sex, gestational age, birth order,
maternal IQ, maternal education,
prenatal smoking (correct
responses)
Age, birth order, maternal IQ,
maternal education, household
income, marital status, HOME
score, prenatal smoking (color
knowledge)
Age, maternal IQ, household
income, marital status, HOME
score, race (shape knowledge)
Proportion correct responses
Color knowledge
Shape knowledge
using Shape School Task at
ages 4 and 4.5 yr
-0.017 (-0.027, 0.007)
-0.074 (-0.101,-0.047)
-0.057 (-0.094, -0.020)
Solon et al. 502 children ages 6-35 mo
(2™) 377 children 3-5 yr,
2003-2004
Visayas, Philippines
Concurrent
Mean (SD): 7.1 (7.7)
10th-90th: 1.6-14.9
Two-stage linear regression model
to account for determinants of blood
Pb (sex, roof materials, water
source, breastfed for > 4 months)
and cognitive function (HOME
score, maternal education, maternal
smoking, born premature, region of
residence)
Verbal IQ
Performance IQ
using WPPSI-III at ages 3-5 yr
-0.34 (-0.63, -0.05)"
-0.45 (-2.37, 1.48)"
Min et al. 267 primarily African-
(2009) American children in the
Cleveland, OH area
followed from birth
(1994-1996) to age 11 yr.
Children were exposed
prenatally to multiple drugs.
Age 4 yr
Mean (range): 7.0
(1.3-23.8)
10th-90th: 3.0-12.1
Linear regression model adjusted
for HOME score, caregiver's
vocabulary test, sex, parity,
maternal marital status, head
circumference at birth
Verbal comprehension age 9 yr
Verbal comprehension age 11 yr
Perceptual reasoning age 9 yr
Perceptual reasoning age 11 yr
Math at age 9 yr
Math at age 11 yr
Reading at age 9 yr
Reading at age 9 yr
Using WISC-R and WJTA
-0.028 (-0.055, 0)
-0.044 (-0.072, -0.015)
-0.033 (-0.064, -0.003)
-0.045 (-0.074, -0.016)
-0.028 (-0.057, 0.002)
-0.033 (-0.062, -0.004)
-0.039 (-0.069, -0.009)
-0.043 (-0.073, -0.014)
Palaniappan 815 children ages 3-7 years
et al. (2011) in 2003-2006 in Chennai,
India
Concurrent
Mean (SD): 11.5(5.3)
10th-90th: 6.0-18.3
Generalized estimating equation
adjusted for sex, hemoglobin level,
maternal education, parental
education, average monthly income,
clustering at school and class level
Visuomotor composite
Using WRAVMA at ages 3-7 yr
-0.017 (-0.029, -0.005)
Studies not included in figure due to unavailability of blood Pb-cognitive function continuous effect estimates
Jedrychowski 452 children followed from
et al. (2008) birth (2001-2003) to age 6
mo
Prenatal (cord)
Mean (95% CI): 1.42
(1.35, 1.48)
10th-90th: 1.36-1.46
Logistic regression model adjusted
for maternal education, parity, sex
No maternal IQ
Verbal recognition memory
score <52.2
Using FTII at 6 mo
OR (95% CI)
1.47 (1.07, 2.01)
Surkan et al. 534 children ages 6-10 yr
(2QQZ)	Boston, MA and
Farmington, ME
Concurrent
Mean (SD): 2.2 (1.6)
Analysis of covariance adjusted for
child IQ, caregiver IQ, age, SES,
race, birth weight
WIAT at age 6-10 yr
Reading score
Math score
Visual memory score
Perseveration errors
Stroop color-word interference
blood Pb level 5-10
pg/dLvs. 1-2 pg/dLe
-5.20 (-9.45, -0.95)
-7.6, -0.43)
-11.9,-1.1)
-14.6, -3.7)d
1.6, 3.1)
-4.02
-6.47
-9.19
0.75 i
Kordas et al. 294 children with blood Pb Concurrent
(2006)	levels < 10 pg/dL	Range: 2.1-10.0
Linear regression model adjusted
for sex, age, hemoglobin, family
possessions, forgetting homework,
house ownership, crowding,
maternal education, birth order,
family structure, arsenic exposure,
tester, school
Visual memory span
<	10 correct vs. > 10 correct:
0.90
Stimulus discrimination
<	20 correct vs. 20 correct
OR (95% CI)
0.90 (0.74, 1.10)
0.85 (0.63, 1.13)
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Study
Population/
Location
Blood Pb Levels
(Hg/dL)
Statistical Analysis
Cognitive Index
Effect Estimate
(96% Cl)a
Bellinger and
Stiles (1993)
148 children in the
Boston, MA area followed
from birth (1979-1981) to
age 10 yr
57 month and
concurrent blood Pb
Exact levels NR,
mean reported to be
Linear regression model adjusted
for HOME score, family stress, race,
marital status
Perseverative errors
57 month blood Pb	-0.05 (
Concurrent blood Pb	-0.05 (
Percent conceptual responses
57 month blood Pb	-0.93 (
Concurrent blood Pb	NR
Using WCST at age 10 yr
¦0.09, -0.01)"
¦0.11,0.01)"
¦2.0, 0.13)
Bellinger et al. 79 children in the Boston, Deciduous tooth Pb Regression model adjusted for
(1994a) MA area followed from
birth (1979-1981) to age
19-20 yr
Q1
Q2
Q3
Q4
2.9-5.9 pg/g Parental IQ; sf> SES, current drug st color-word test
use, current alcohol use, current
6.0-8.7 pg/g j||jcjt drUg use! maternal education,
8.8-19.8 pg/g maternal age, birth order
19.9-51.8 pg/g	WCST Number Correct
Mean (SE)
Q1
Q2
Q3
Q4
Q1
Q2
Q3
Q4
103 (8.1),
116(7.6),
127 (8.4),
125 (7.7)
70.5 (2.9),
72.7 (2.8),
74.5(3.1),
71.4(2.9)
WISC = Wechsler Intelligence Scale for Children, WRAT = Wide Range Achievement Test, WCST = Wisconsin Card Sorting Test, CANTAB = Cambridge
Neuropsychological Testing Automated Battery, WPPSI = Wechsler Preschool and Primary Scale of Intelligence, WRAVMA = Wide Range of Visual Motor
Ability, FTII = Fagan Test of Infant Intelligence, WIAT = Wechsler Individual Achievement Test
aEffect estimates are transformed to a z-score and standardized to a 1 pg/dL increase in blood Pb level in the 10th-90th percentile interval.
b95% CI was constructed using a standard error that was estimated from a p-value.
"Sufficient data were not provided to calculate 95% CI.
dThe direction of the effect estimate was changed such that a negative estimate represents poorer performance and a positive estimate represents better
performance.
eEffect estimates compare test performance of children in higher blood Pb groups to children in lowest blood Pb group.
A recent analysis of the Rochester cohort age 5 years extended previous findings with
observations that higher concurrent blood Pb level also was associated with lower
performance on many of the same tests executive function as assessed by CANTAB
(Froehlich et al.. 2007) (Figure 5-5 and Table 5-5). In addition to spatial memory,
Froehlich et al. (2007) found associations with rule learning and reversal tasks. Pb
exposure of animals consistently has been found to induce impairments in homologous
tests of cognitive flexibility, including delayed spatial alternation tests (Section 5.3.2.2).
Both of these tests assess the ability of children and animals to complete a task according
to a change in rules or reinforcement. Impaired performance in both children and animals
is indicated by increased response errors, decreased percent of correct responses, and
perseverance (i.e., animals repeatedly pressing the same lever without moving between
the two locations or children repeatedly selecting the wrong target despite a change in
rules or reinforcement. Some effect estimates describing the blood Pb-cognitive function
association were attenuated after adjustment for covariates (e.g., spatial span and
planning tasks), whereas others were magnified (e.g., total trials in rule learning, total
errors in spatial working memory) (Froehlich et al.. 2007). Further, despite attenuation,
several associations did not change in statistical significance (Figure 5-5 and Table 5-5).
Froehlich et al. (2007) also added to previous findings in the Rochester cohort by
indicating that associations were modified by sex and dopamine receptor (DRD4) genetic
variants. An increase in blood Pb level was associated with larger decrements in rule
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learning and reversal and spatial memory in children with the DRD4 exon III 7-repeat
microsatellite (assessed using a blood Pb-DRD4-7 interaction term, p = 0.042). The
biological plausibility for these findings is provided by observations that the DRD4-7
variant is associated with reduced dopamine-induced signaling in downstream pathways
(e.g., cyclic AMP), that Pb exposure is found to impact dopaminergic activity
(Section 5.3.8.8), and that dopamine is a key neurotransmitter that regulates cognitive
processes. Associations of concurrent blood Pb level with rule learning and reversal also
were greater in boys. Additional support for these observed interactions was provided by
observations with children with the DRD4-7 variant and boys had poorer performance on
rule learning and reversal tasks.
Several other studies found associations between higher blood Pb level and poorer
performance on rule learning and reversal tasks in children and young adults (Surkan et
al.. 2007; Kordas et al.. 2006; Chiodo et al.. 2004; Bellinger et al.. 1994a; Bellinger and
Stiles. 1993). Across studies, the age range of subjects was 6 to 20 years, and most
studies examined concurrent blood Pb levels. In the Boston cohort, age 4.5 year blood Pb
levels were associated with poorer cognitive flexibility at age 10 years (Bellinger and
Stiles. 1993). and deciduous tooth Pb (measured at ages 6-7 years) was associated with
poorer cognitive flexibility at ages 19-20 years (Bellinger et al. 1994a).
Blood Pb-associated decrements in executive function in children with mean blood Pb
levels < 5 were not as clearly demonstrated. Among children in five Korean studies, ages
8-11 years, blood Pb level was not associated with performance a color-word association
test. In the study of children (ages 6-10 years) participating in NECAT, Surkan et al.
(2007) provided information on the blood Pb-cognitive function concentration-response
relationship. Children with concurrent blood Pb levels 3-4 (ig/dL had lower scores
compared with children with blood Pb levels 1-2 j^ig/dL; however, only a few
associations were statistically significant (e.g., digit span, memory, motor speed). In
analyses adjusted for caregiver IQ, age, SES, race, and birth weight, children with
concurrent blood Pb levels of 5 to 10 (ig/dL had lower performance in a number of
executive functioning domains such as working memory, visuospatial skills, cognitive
flexibility, and ability to formulate, test, and adapt hypotheses (Figure 5-5 and Table 5-5).
Compared with children with blood Pb levels of 1-2 (ig/dL, children with blood Pb
levels 5-10 (ig/dL had lower scores on the reading (5.2 points [95% CI: 0.95, 9.45]) and
math (4.0 points [95% CI: 0.43, 7.6]) composites of the Wechsler Individual
Achievement Test. These decrements were robust to adjustment for child FSIQ. Thus,
these findings combined with those from the Rochester cohort (Canfield et al.. 2004;
Can field et al.. 2003b) provide evidence that higher blood Pb level is associated with
decrements in specific cognitive functions that may not be represented by FSIQ, a test of
global function.
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In other studies of populations with mean concurrent or cord blood Pb level < 3 |_ig/dL.
the contributions of likely higher past Pb exposures cannot be excluded. Also, these
studies did not have as extensive consideration for potential confounding. In infants in
Krakow, Poland with a mean cord blood Pb level of 1.42 (ig/dL, higher cord blood Pb
level was associated with a lower score on a visual recognition memory test
(Jedrvchowski et al. 2008) (Table 5-5). The cord blood Pb levels may have been
influenced by past Pb exposures of the mother mobilized from bone stores to the blood.
While investigators adjusted for maternal education, parity, and sex, they did not adjust
for other SES-related variables such as maternal IQ or HOME score.
The influence of past Pb exposures also is uncertain in the NHANES analyses, which
included adolescent subjects born in the 1970s. Krieg et al. (2010) examined an older
subset of children (ages 12-16 years) from NHANES III (1988-1994) who were
previously examined in Lanphear et al. (2000). Among all ages and older adolescents,
higher concurrent blood Pb level was associated with lower scores on block design, digit
span, reading score, and arithmetic tests (Figure 5-5 and Table 5-5) (Krieg et al.. 2010;
Lanphear et al.. 2000). In analyses of quartiles of blood Pb level Lanphear et al. (2000)
additionally found a monotonic decrease in math and reading scores across increasing
quartiles in covariate-adjusted models. Per unit increase in blood Pb level, the decrease in
arithmetic and reading scores were larger among children with blood Pb levels less than
2.5 (ig/dL than among all subjects, with statistically significant associations found for
children with blood Pb levels less than 5 (ig/dL. While the NHANES results were
adjusted for sex, race/ethnicity, test language, and multiple SES-related variables such as
family income and caregiver education, information on HOME score was unavailable.
Krieg et al. (2010) provided additional information on effect modification by vitamin D
receptor (VDR) variants. Although there were not differences in blood Pb levels among
the various haplotypes of VDR, various polymorphisms and haplotypes modified the
association between blood Pb level and a range of neurocognitive tests. The VDR
regulates calcium absorption and metabolism, and effect modification by VDR variants is
consistent with the well-established mode of action of Pb in mimicking calcium in shared
transport and metabolic pathways. However, several inconsistencies were observed by
Krieg et al. (2010) in that a particular variant was associated with a lower Pb-associated
decrement in some indices but a greater Pb-associated decrement in other indices. For
example, among children ages 12-16 years, the VDR rs2239185 CC genotype was
associated with the largest blood Pb-associated decrease in digit span score and reading
score. The decreases in digit span score (95% CI) per 1 j^ig/dL increase in concurrent
blood Pb level were -1.5 points (-2.2, -0.71) forthe CC genotype and -0.26 points (-0.99,
0.46) for the TT genotype. Conversely, the TT genotype was associated with the greatest
Pb-associated decrease in arithmetic score. The decreases (95% CIs) in math score per
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1 jj.g/dL increase in concurrent blood Pb level were -8.4 points (-11.5, -5.2) for the TT
genotype and -2.7 points (-10.1, 4.6) for the CC genotype. Effect modification by VDR
rs731236 was more consistent across cognitive tests, with larger blood Pb-associated
decrements in cognitive performance found in children with the CC genotype. It is
important to recognize that not all of these VDR variants have been linked to functional
changes.
Other recent studies conducted in diverse populations of children also found associations
between higher blood Pb level and decrements in learning, memory, and visuospatial
skills; however, they had limited or no consideration of potential confounding by SES-
related variables, which limits the implications of their findings (Nelson and Espy. 2009;
Counter et al.. 2008; Min et al.. 2007; Vega-Dienstmaier et al.. 2006).
Academic Performance in Children
As described in the preceding sections, a preponderance of evidence demonstrates Pb-
associated deficits in FSIQ and specific indices of cognitive function, including math,
reading, and vocabulary skills. Deficits in global cognitive function and the
aforementioned skills can lead to poorer academic achievement and school performance,
which may be more objective measures of abilities and skills and have important
implications for success later in life. Aptitude tests are used to predict future performance
of an individual on a task or test. Achievement tests and school performance, in
comparison, assess the actual knowledge of an individual in subject areas the individual
has studied and measure the acquired knowledge of that subject. Studies reviewed in the
2006 Pb AQCD consistently demonstrated associations of Pb biomarkers with measures
of academic achievement and performance including scores on math or vocabulary tests,
class rank, teacher assessment of academic functioning, and high school completion.
Several studies found that blood or dentin Pb levels measured at an early age (ages
2-8 years) were associated with academic performance at older ages (ages 8-18 years),
suggesting the effect of early exposure to Pb may be persistent (Lcviton et al.. 1993;
Bellinger et al.. 1992; Needleman et al.. 1990). Several studies found associations with
concurrent blood Pb levels (Kordas et al.. 2006; Wang et al.. 2002a; Al-Saleh et al.. 2001;
Lanphear et al.. 2000). Among recent studies, academic performance was examined less
frequently; however, findings were consistent with the extant body of evidence.
The longitudinal study by Bellinger et al. (1992) was particularly noteworthy for
examining associations of blood Pb level at several ages from the prenatal period to age
10 years with Kaufman Test of Educational Achievement (KTEA) scores at age 10 years
in the Boston cohort. While blood Pb levels at several ages were associated with lower
academic achievement scores when in a model alone, only blood Pb level at age 2 years
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showed a statistically significant association with lower predicted academic achievement
after adjusting for HOME score, child stress, maternal age, race, maternal IQ, SES, sex,
birth order, marital status, and number of residence changes. These covariates were
selected because they were associated with blood Pb level and KTEA score, changed the
blood Pb effect estimate by more than 10%, or because they were recognized to be
important predictors of cognitive development (e.g., maternal IQ, birth order, sex). Age
2 year blood Pb level was associated with a similar magnitude of decrease in the KTEA
battery composite score in univariate and multivariate models. Additionally, the
association was robust to adjustment for FSIQ, indicating that blood Pb levels may be
associated with poorer performance on academic tasks not reflected in FSIQ.
Miranda et al. (2007b) and (2009) linked blood Pb surveillance data collected between
ages 0 and 5 years with end-of-grade (EOG) testing data for 8,600 fourth grade children
in seven of the largest counties in North Carolina and 57,678 children in the entire state,
respectively. A key strength of these analyses was the availability of individual-level data
on blood Pb level and achievement score on a large number of children representative of
the North Carolina fourth grade population, which provided large numbers of children
with lower blood Pb levels (2-5 (ig/dL). Thus, in this study, there was greater power to
discern differences in achievement scores among children in the lower range of blood Pb
levels. However, due to the records-based research design, investigators had a smaller set
of available potential confounding variables than those typically considered in analyses
of neurodevelopmental outcomes. In addition to adjusting for sex, race, school-type,
school district, and age of blood Pb measurement, investigators adjusted for participation
in a free or reduced lunch program as a measure of SES, and parental education as a
measure of parental IQ. Miranda et al. (2007b) additionally adjusted for daily use of a
computer as a measure of a stimulating home environment. Higher early childhood blood
Pb levels were associated with lower reading and math EOG scores (Figure 5-6 and 5-7),
and in both analyses, children with a blood Pb level of 2 (ig/dL had lower EOG scores (p
< 0.05) compared with children with a blood Pb level of 1 (ig/dL. Further, across deciles
of blood Pb level, the decrease in EOG score generally was monotonic (Figure 5-6). In
the statewide dataset, compared with children with an earlier (measured at some point
between birth and age 5 years) blood Pb level of 1 j^ig/dL, children with an earlier blood
Pb level of 2 j^ig/dL had a 0.30-point lower (95% CI: -0.58, -0.01) reading EOG score
(Miranda et al.. 2009).
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-0.5
-1.0
-1.5
-2.0
-a
-2.5
-3.0
-3.5
~ Model 1: linear BLL
~ Model 3: dummy for each BLL (1 to 210)
-4.0
2
3
6
9
4
5
7
8
*10
Blood lead levels (ug/dL)
Source: Reprinted with permission of Elsevier Science, Miranda et al. (2007b).
Note: These results illustrate the decrease from a baseline score of 262.6 for a hypothetical referent white female, screened at
2 years of age, living in Wake County, NC, with parents with a high school education, not enrolled in the school lunch program, and
who does not use a computer every day (i.e., model covariates = zero).
Figure 5-6 Associations between childhood blood Pb levels and fourth grade
End-of-Grade (EOG) math scores.
While the linear regression analyses indicated fractional-point decreases in EOG scores
per 1 (ig/dL increases in blood Pb level, Miranda et al. (2009) additionally used quantile
regression to discern differential effects in various segments of the EOG distribution
(i.e., what is the 10th percentile of EOG scores conditioned on early childhood blood Pb
levels). Compared with linear regression, quantile regression is more robust in response
to outliers and predicts outcomes at the top and bottom tails of the outcome distribution
rather than at the mean. With increasing blood Pb level, the lower tail of the EOG
distribution was stretched out more so than were the middle or upper portions of the
distribution. For example, an increase in blood Pb level from 1 to 5 (ig/dL was associated
with a greater decrease in EOG score among children in the 5th percentile of EOG than in
children in the 95th percentile of EOG score (Figure 5-7). These findings indicated that
children residing at the lowest performance regions of the EOG score distribution may be
more affected by Pb exposure. Similarly, using quantile regression, Miranda et al. (2009)
showed that while cumulative social risk (lower parental education, being enrolled in a
school lunch program) had a greater negative association with academic achievement in
these children, blood Pb level was independently associated with EOG score decrements
that were as large as 1 to 2 points.
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Cummulative Deficit: Decrease in EOG scores by multiple risk factors
SI 10* 15* KS 15% m m 4« 45* 50* 551 601 65% 70k 75% » S3* »* »
WRCn program, p wmm cornpietea college)
*12
)• Effect of inc ¦	1 toSpg/dL
Quarrtile
Source: Reprinted with permission of Elsevier Science, Miranda et al. (2009)
Note: Baseline score calculated for a hypothetical referent individual with a blood Pb level of 1 |jg/dL, parents completed college,
and not enrolled in the school lunch program (i.e., model covariates = zero).
Figure 5-7 Greater reduction in EOG achievement test scores with
increasing blood Pb level in lower percentiles of the test score
distribution.
Similar to Miranda et al. (2009). Chandramouli et al. (2009) observed associations
between early childhood blood Pb levels (age 30 months) and later academic
performance (Standard Assessment Tests [SAT] at age 7 years) among participants of the
Avon Longitudinal Study of Parents and Children conducted in the U.K. However, in this
study, decrements in SAT score were most clearly indicated in children with blood Pb
levels of 5 to 10 (ig/dL. Compared with children with blood Pb levels 0-2 (ig/dL, children
with blood Pb levels 5-10 (ig/dL scored 0.51 points lower (95% CI: -0.82, -0.32) on the
reading test and 0.49 points lower (95% CI: -0.78, -0.31) on the math test, adjusting for
maternal education, home ownership, maternal smoking, home facilities score, parental
SES, family adversity index, and parenting attitudes. Children with blood Pb levels
2-5 |_ig/dL did not consistently have lower scores compared with children with blood Pb
levels 0-2 (ig/dL. While these aforementioned recent studies found associations for early
childhood blood Pb levels, unlike the longitudinal assessment by Bellinger et al. (1992).
they did not have blood Pb measurements available at other lifestages. Therefore, recent
studies could not provide information on associations with blood Pb levels at other
lifestages.
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5.3.2.2
Toxicological Studies of Cognition, Memory and Learning
As described in Section 5.3.2.1, epidemiologic studies in children have consistently
demonstrated associations between blood Pb levels and decrements in a range of
executive functions related to memory, reasoning, problem solving, and verbal skills.
Studies in animals also demonstrate Pb-induced decrements in various indices of
executive function, and several tests conducted in animals share homology with tests
conducted in humans.
Data on neurocognition and learning as well as other nervous system endpoints with
concentration-response data are shown in Figure 5-8 and accompanying Table 5-6. These
data demonstrate that numerous studies in animal models have demonstrated that Pb
exposure produces a spectrum of changes in the central nervous system, including
deficits in CNS development and plasticity and altered homeostasis of mediators of
cognitive function that manifest with deficits in memory, learning, and cognition. It is
thought that early life Pb exposure can permanently alter CNS development and other
pathways that contribute to these aforementioned memory and learning deficits, but the
animal toxicology data show that multiple sensitive developmental windows for Pb
exposure exist (Rice and Gilbert. 1990). As illustrated in Figure 5-8 and Table 5-6, Pb-
induced impairments in learning primarily previously were observed in animals with
blood Pb levels > 30 (ig/dL. Several new studies added to the evidence for impaired
learning and memory in animals with lower blood Pb levels, 8-17 (ig/dL (Corv-Slechta et
al.. 2010; Li et al.. 2009c; Niu et al. 2009; Virgolini et al.. 2008a; Stangle et al.. 2007). In
these studies of animals with lower blood Pb levels, Pb exposures began early, during the
gestational or lactation period.
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Behavioral; Neonate; Rat; Female (1)
Behavioral; Neonate ; Rat; Both (2)
Behavioral; Neonate; Rat; Male (3)
Behavioral; Neonate; Rat; Both(4)
Cognition; Neonate; Rat; Both (S)
Cognition;Adult; Rat; Male(6)
Cognition;Adult; Monkey; Both (7)
Cognition;Neonate; Mouse; Bath(8)
Cognition;Neonate; Rat; Both (9)
Cognition;Neonate; Rat; Both (4)
Corti coster one; Neonate; Rat; Female (10)
Corticosteroid;Neonate; Rat; MaJe(ll)
Corticosterone;Adult; Rat; Male(12)
Corticosterone;Adult; Rat; Male(13)
Morphology; Neonate; Rat; Both (14)
Morphology; Adult; Rat;Male (G)
Morphology; Adult; Monkey; Female (IS)
Morphology; Adult; Rat; Male (1G)
Morphology; Neonate; Mouse; Both (8)
Morphology; Neonate; Rat;Both (9)
Motor function; Neonate; Mouse; Male (17)
Motor function; Adult; Rat; Male(12)
Motor function; Neonate; Rat; Both (5)
Motor function; Adult; Rat; Male(13)
Motorfunction; Neonate; Rat; Both (4)
Motor function; Neonate; Rat; Male (3)
Neurophysiology; Neonate; Rat; Both (14)
Neurotransmitter; Neonate; Mouse; Both(18)
Neurotransmitter, Neonate; Mouse; Both(17)
Neurotransmitter; Adult; Rat; Female (13)
Neurotransmitter; Adult; Rat; Male(13)
Neurotransmitter; Neonate; Rat; Female (11)
Oxidative Stress; Adult; Monkey; Female (15)
Oxidative St res; Neonate; Mouse; Both (8)
Physical Development; Adult; Rat; Female(14)
Physical Development; Neonate; Rat; Female(1)
Physical Development; Neonate; Rat; Male (3)
Physical Development; Adult; Rat; Male(G)
Proliferation/diff/survival; Neonate; Rat; Both (9)
Stress-induced cortkcsterone;Neonate; Rat; Female(10)
Stress-induced corticosterone;Adult; Rat; Female(13)
Stress-induced corticosterone;Neonate; Rat; Female(11)
Stress-induced motorfunction; Adult;Rat; Both(13)
Stress-induced neurotransmitter; Neonate; Rat; Female (10)
Stress-induced neurotransmitter; Adult; Rat; Female (13)
Stress-induced neurotransmitter. Adult; Rat; Male(13)
Stress-induced neurotransmitter; Neonate; Rat; Male(ll)
1	10	100	1,000
Blood Lead(ng/dl)
Dosimetric representation reported by blood Pb level. (ID corresponds to Table 5-6.)
Figure 5-8 Nervous system summary array of toxicological outcomes after
Pb exposure.
~	Highest Concentration
~	lowest Cone, with Response
A Highest Cone, with No Response
o Lowest Concentration
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Table 5-6 Summary of findings from neurotoxicological exposure-response
array presented in Figure 5-8
Study ID
in Figure
6-8
Reference
Blood Pb
Level
(Hg/dL)
Outcome
1
Beaudin et al. (2007)
13 & 31
Behavior, neonate: Lactational Pb exposure, offspring deficient in Reward Omission testing.
Physical development; Postnatal Pb exposure (birth to 4 weeks of age): Pb-dependent development of over-
reactivity to reward omission and errors is reversible with chelation treatment.
2
Grant etal. (Grant et al..
1980)
57
Behavior, neonate: chronic Pb exposure (drinking water) to dams and pups, Changed behavior, males.



Behavior, neonate: Pb exposure (oral gavage of pups) during lactational period, Changed emotional behavior,
males and females.
3
Kishi et al. (1983)
59 & 186
Motor function, neonate: Pb exposure (oral gavage of pups) during lactational period, motor function (rotarod
performance) impaired, both sexes.
Physical development; Pb exposure during lactation (oral gavage): Delayed development of righting reflex in
male rats.
Behavior-Pb exposure (oral gavage of pups) during lactation: aversive conditioning affected by Pb exposure,
male and females.
Cognition-Pb exposure (oral gavage of pups) during lactation: Response inhibition impaired, both sexes.
Motor function- Pb exposure (oral gavage of pups) during lactation: Increased motor activity and impaired motor
coordination (rotarod), male and females.
Cognition; Developmental Pb exposure (PND1-PND30): Impaired learning with visual discrimination task,
heightened response to errors, both sexes.
5	Stangle et al. (2007) 13 & 31 	
Motor function; Developmental Pb exposure (PND1-PND30): Alcove latency and response latency significantly
affected by Pb exposure, both sexes.
Cognition-Adult male 21 day Pb exposure: Hyperactivity with Habituation to new cage environment.
Morphology; 21 day Pb exposure to adult males: Marker of neuronal injury-elevated hippocampal glial fibrillary
6	Gong & Evans (1997) 38 & 85 acidic protein (GFAP).
Physical development; Adult male rats (21 day Pb exposure): Neurotoxicity measured with brain glial fibrillary
acidic protein (GFAP).
7	Rice (1990)	32 & 36 Cognition-Chronic Pb exposure from birth: Spatial discrimination reversal task impairment, both sexes.
Cognition-Gestational & lactational Pb exposure: Morris water maze performance impaired.
Morphology; Gestational & lactational Pb exposure: Increased levels of inflammatory cytokines & exocytosis
8	Li et al. (2009c)	80 & 100 related proteins in brains of pups at weaning, both sexes.
Oxidative stress-gestational and lactational Pb exposure: Elevated hippocampal TNF levels in offspring, males
and females.
Cognition- Gestational & lactational Pb exposure: Morris water maze performance impaired.
Morphology: Increased levels of Alzheimer disease-associated proteins in mice with gestational and lactational
9	Li et al. (2010b)	80 & 102 Pb exposure, both sexes.
Proliferation/diff/survival, gestational & lactational Pb exposure: Increased hippocampal expression of P-tau and
amyloid beta in male and female pups.
Corticosterone: Lifetime Pb +/- stress: Correlation between 9-month old's corticosterone level and frontal cortex
dopamine levels in behaviorally tested female offspring.
Stress: Corticosterone-Lifetime Pb plus stress: Affects Fl performance, dopamine and serotonin levels in female
offspring.
Stress: Corticosterone-neurotransmitter-Lifetime Pb exposure in female rats plus stress: Dopamine homeostasis
affected.
Virgolini, Rossi-George, „
Wfeston et al. (2008b)
^ ,,n77 . 33,174 &
Overmann (1977) 226
,n Cory-Slechta et al. man
10 (2M)
Corticosterone: Maternal Pb plus stress: Elevated corticosterone in male offspring with prenatal stress + offspring
stress was further enhanced with Pb exposure.
Stress: Corticosterone-Maternal Pb plus stress: Affects Fl performance.
Neurotransmitter; Gestational and lactational Pb exposure: Induced DA and 5HT changes in rat offspring.
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Study ID
in Figure
6-8
Reference
Blood Pb
Level
(Hg/dL)
Outcome
Stress induced neurotransmitter effects, Maternal Pb plus stress: 5-HT and 5-HIAA (5-HT metabolite), and
dopamine turnover were significantly affected in males.



Corticosterone: Chronic Pb exposure from weaning: Pb exposure alone decreased basal plasma corticosterone
levels at 5 months of age, males.
12
Virgolini et al. (2005)
15 & 27
Motor function: Chronic Pb exposure from weaning: Locomotor activity significantly decreased Fl response rates
& increased post-reinforcement pause period in a concentration-dependent manner, males.
Stress: Corticosterone-Chronic Pb plus stress: Affects neurotransmitters & Fl performance


31
Corticosterone: Maternal Pb exposure (gestation and lactation) +/- stress: Differential basal corticosterone levels
between behavioral and non-behavioral tested rats, females.


11 S/or 31
Stress: Corticosterone-Maternal Pb plus stress: Affects Fl performance, dopamine, serotonin, and NE levels.


Motor function: Maternal Pb +/- stress: Increased locomotor activity (run rate) with Pb and stress exposure.
13
Virgolini, Rossi-George,
Lisek et al. (2008a)
31
Neurotransmitter; Gestational and lactational Pb exposure: Induced NE aberrations in adult rat offspring (both
sexes).
Stress induced motor function: Maternal Pb +/- stress: Increased locomotor activity (run rate) with Pb and stress
exposure.
Stress induced neurotransmitter; Gestational and lactational Pb exposure + stress: Induced HVA (monoamine
neurotransmitter metabolite) and NE aberrations in female adult rat offspring.



Morphology; Gestational Pb exposure: Neurite outgrowth marker PSA-NCAM decreased in rat pups, both sexes.
14
Huetal. (2008b)
4 & 12
Neurophysiology; gestation Pb exposure: decreased hippocampal sialyltransferase activity, both sexes.
Physical development; t-Gestational Pb exposure: Early brain synapse development impaired (hippocampal PSA-
NCAM).
15
Wu et al. (2008a)
19 & 26
Morphology: Elevated expression of Alzheimer's disease-related genes and Tc factors in aged brains of female
monkeys (exposed to Pb as infants).



Oxidative stress: Elevated oxidative DNA damage in aged brains of female monkeys (exposed to Pb as infants).
16
Tavakoli-Nezhad et al.
(2001)
18, 29, & 54
Morphology; 3 to 6 weeks of Postnatal (starting at PND22) Pb exposure in males: Decreased number of
spontaneously active midbrain dopamine neurons.
17
Leasure et al. (2008)
10 & 42
Motor function; Mouse maternal (dam) Pb exposure: Induced decreased rotarod performance in offspring (1 year-
old male offspring).
Neurotransmitter; Mouse maternal (dam) Pb exposure: Affects 1 year old male offspring dopamine homeostasis,
both sexes.
18
Fortune & Lurie (2009)
8 & 43
Neurotransmitter; Mouse maternal (dam) Pb exposure: Affects offspring superior olivary complex (auditory)
neurotransmitters, both sexes.

Memory and Learning - Morris Water Maze
1	Blood Pb levels have been associated with decrements in both verbal (e.g., digit span)
2	and spatial memory in children (e.g., spatial and visual memory span) (Figure 5-5 and
3	Table 5-5). In the 2006 Pb AQCD, the results from different animal studies testing the
4	effect of Pb on memory were mixed with impaired memory observed in animals with
5	blood Pb levels 10-35 (ig/dL but improved memory observed in other studies. Low-dose
6	Pb was not found to affect short-term memory, i.e., recall of a learned task. Mixed results
7	for memory also may be due to the fact that memory tests also test inattention. Memory
8	tests may give incorrect results when opportunities exist for impaired attention to
9	contribute to test results (U.S. EPA. 2006b). In animals, spatial memory is tested most
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commonly using the Morris water maze. The 2006 Pb AQCD reported deficits in the
Morris water maze performance with Pb exposure. New research continued to show Pb-
induced impaired Morris water maze performance. The Morris water maze tests spatial
memory and learning by having a mouse swim and locate or remember the location of a
platform submerged in opaque water. The Morris water maze is homologous to the
spatial memory components of the CANTAB and WISC-R, which test the ability of
subjects to recall correctly a sequence of locations (Froehlich et al.. 2007; Can field et al..
2004; Chiodo et al.. 2004).
Li et al. (2009c') examined Pb exposures from gestation through lactation. During this
period, dams received Pb-acetate dissolved in drinking water (0.1%, 0.5%, and 1% with
corresponding blood Pb levels of 4, 8 and 10 (ig/dL at postnatal day [PND]21).
Beginning at weaning, Pb-exposed pups were subjected to Morris water maze
performance testing. Pb-exposed pups had statistically significant increases in escape
latency and number of crossings of the platform area at 0.5% and 1% Pb-acetate exposure
(blood Pb levels of 8 and 10 j^ig/dL, respectively), indicating impaired memory and
learning (Li et al.. 2009c). The pups in Li et al. (2009c') were not separated by sex. Cao et
al. (2009) found that chronic administration of the supplement melatonin exacerbated Pb-
induced impairments in spatial memory and long-term potentiation. Adult male Wistar
rats that had been exposed to Pb-acetate (0.2%) from birth received melatonin (3 mg/kg)
from weaning via gastric gavage for 60 days. At this point (PND81-90), animals
performed in the Morris water maze, and LTP of the hippocampal dentate gyrus was
measured.
Another study found that various dietary supplements or methioninecholine concomitant
with Pb exposure in weanling males shortened the escape latency of Pb-exposed pups to
resemble more closely the escape latency of control pups (Fan et al.. 2010; Fan et al..
2009a). Zinc and methionine were effective dietary supplements in Fan et al. (2009a);
glycine, taurine, vitamin C, vitamin Bl, tyrosine had no effect on the Pb-associated
Morris water maze results. These data on the effect of early life Pb exposure on learning
and memory in the Morris water maze affirm earlier findings including those by
Kuhlmann et al. (1997) who used maternal Pb diet exposure (gestation and lactation),
continuous Pb exposure (gestation through adulthood) or post-weaning Pb exposure and
only found only significant impairments in the maternal and continuous exposure groups.
Another component of memory is working memory, which is the ability to temporarily
keep information in mind while using the information to perform a related or unrelated
task. The Morris water maze measures working memory in addition to learning. The
2006 Pb AQCD reported that working memory as assessed by the Morris water maze was
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significantly affected in chronic developmentally Pb-exposed (Pb-acetate in feed 10 days
prior to mating through PND 21) female offspring at PND 21 (Jettetal.. 1997V
Working memory also can be measured by testing delayed spatial alternation (DSA),
which specifically tests spatial reversal learning. With DSA, an animal receives rewards
based on alternating responses between two separate levers. In children, tests of spatial
reversal learning includes the rule learning and reversal components of the CANTAB,
Stroop Test, and WCST, and decrements in DSA have been associated with higher blood
Pb levels in children (Figure 5-5 and Table 5-5). Studies detailed in earlier Pb AQCDs
showed that Pb-exposed animals had deficits under DSA testing (Alber and Strupp. 1996;
Rice and Gilbert. 1990; Rice and Karpinski. 1988; Levin et al.. 1987; Levin and
Bowman. 1986). Studies in nonhuman primates showed Pb-induced behavioral
impairment on DSA tasks (Rice and Gilbert. 1990) across many exposure periods,
including early-life and chronic Pb exposures. Specifically, there were multiple lifestages
during which Pb exposure induced deficits in this task. As in children, deficits in DSA
included increased response errors, decreased percent of correct responses, and
perseverance at a task (i.e., repeatedly pressing the same lever without moving between
the two locations). These observations have been consistently made in nonhuman
primates with continuous Pb exposure or juvenile to adult exposure but less consistently
made in rats with juvenile only or juvenile to adult exposure. In fact, some rodent studies
showed increased accuracy in the delayed alternation trials (Corv-Slechta et al.. 1991).
Working memory is a subcategory of executive function or goal-oriented problem
solving. Pb-induced deficits in working memory may be one of many factors that
contribute to associations between blood Pb levels and inattention observed in humans
(Min et al.. 2007; Surkan et al.. 2007; Chiodo et al.. 2004; Schweitzer et al.. 2000; Stiles
and Bellinger. 1993).
Learning - Y Maze
A recent study using the three-branch radial Y-maze showed Pb-induced effects on
learning in rat offspring exposed during lactation and into adulthood (Niu et al.. 2009).
The Y-maze has a light at the end of each branch. The branch with the illuminated light is
a safe area whereas the other two branches are electrified and cause a mild electric shock
when entered. The Y-maze test evaluates learning based on three criteria: 1) learning
days or the number of days required to learn the maze (90% correctly); 2) the error
number (EN) or the number of Y-maze runs required to learn the maze; and 3) the total
reaction time (TRT) or the total amount of time spent in the maze per test day. Wistar
Albino rat pups were exposed to Pb-acetate during lactation until the termination of the
experiment at 12 weeks of age (dam drinking water 300 mg/L, blood Pb level: 17 j^ig/dL
at 6 weeks of age. Rats were evaluated on the three Y-maze endpoints at 6, 8, 10 and
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12 weeks of age.) Pb induced statistically significant impairments in learning, in
particular, at 8, 10 and 12 weeks of age. EN and TRT were significantly affected only at
8 and 10 weeks, respectively. Hippocampal glutamate levels were also significantly
attenuated in Pb-exposed animals, and this is discussed in more detail in Section 5.3.8.4.
Learning - Response Inhibition and Schedule-Controlled Behavior Testing
Response inhibition is another measure of executive function and is measured with
multiple tests of premature responses, decreased pause time between two scheduled
events, and increased perseverance. These tests include Differential Reinforcement of
Low Rates of Responding (DRL), Fixed Interval (FI) testing (Table 5-7), FI with
Extinction (FI-Ext) or Fixed Ratio (FR)-FI, and Signal Detection with Distraction.
Multiple studies from the 2006 Pb AQCD as well as earlier literature showed that early
life Pb exposure contributed to response inhibition across the spectrum of these
aforementioned tests. Monkeys with moderate blood Pb levels (11-13 j^ig/dL) learned the
DRL task more slowly but eventually acquired reinforcement rates equal to that in
controls. Newer data from female rats exposed to Pb (Stannic et al.. 2007) continued to
show animals with premature responses after Pb exposure or response inhibition
decrements. These findings in animals are consistent with observations in children that
blood Pb levels are associated with poorer impulse control (Figure 5-14 and Table 5-9).
In children, impulsivity is commonly assessed by having parents or teachers rate problem
behaviors such as trouble waiting, interrupting others, and responding at inappropriate
times. However, studies also have specifically tested response inhibition assessed using
the continuous performance test, which similar to animal tests, measures reactions to a
stop signal (Section 5.3.3.1).
The 2006 Pb AQCD discussed learning or cognition as measured with schedule-
controlled behaviors including FI and FR operant conditioning and found that FI response
rate was affected differentially with low-level (as low as 11 j^ig/dL) and high-level Pb
(peak levels of 115 j^ig/dL) exposures increasing and decreasing FI response rate,
respectively. This nonlinear response was since further explored in recent work, much of
which also examined the interaction between stress and Pb exposure.
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Table 5-7 Summary of effects of maternal and lifetime Pb exposure on
Fl performance


Maternal Pb"

Lifetime Pbc
Pb (ppm)
Overall rate3
PRPa
Overall rate3
PRPa
0 ppm:
0-PS
No Significant Effect11
No Significant Effect
No Significant Effect
No Significant Effect
o-os
No Significant Effect
*1-23%
No Significant Effect
No Significant Effect
60 ppm:
50-NS
No Significant Effect
No Significant Effect
*t 95%
No Significant Effect
50-PS
No Significant Effect
No Significant Effect
*t 79.2%
*1-42%
50-OS
*t 64.9%
No Significant Effect
*t 74.7%
* |-39.3%
160 ppm:
150-NS
*t 42.4%
*| -30.3%
No Significant Effect
No Significant Effect
150-PS
No Significant Effect
*1-25.7%
*t 90.7%
*| -44.7%
150-OS
*t 59.2%
No Significant Effect
*t 78.5%
No Significant Effect
Note:*Dam blood Pb levels ranged from 5-13 |jg/dL over gestation and lactation; offspring blood Pb ranged from 7-13 |jg/dL from early lifetime points out
to ten months of age. Thus, this study demonstrates that lifetime Pb exposure with or without prenatal stress induced learning deficits in female mice.
Mechanistically, these authors propose that associations of Pb and stress with learning deficits may be related to aberrations in corticosterone and
dopamine.
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; |, represents increase; j, represents decrease. PRP = post-reinforcement pause.
bData from Virgolini et al. (2005). 'Denotes significant effect versus Oppm control (p <0.05).
cData from current study [Rossi-George et al. (2011)1
'Denotes significant effect vs. Oppm control (p <0.05).
Source: Reprinted with permission of Elsevier Science, Rossi-George et al. (20111 (Table 1).
Learning Ability with Stress
The combined paradigm of Pb exposure and stress experienced by a laboratory animal is
now being studied by multiple investigators who are focusing on the common pathway of
HPA axis alteration and altered brain neurotransmitter levels. These studies indicated
greater impairments in learning with Pb when combined with stress. These findings
provide support for the association between higher blood Pb level and lower MDI score
observed in children with mothers with low self-esteem but not children with mothers
with high self-esteem (Surkan et al.. 2008) (Table 5-4). In children, maternal self-esteem
has been linked with the ability to cope with stress and improved maternal-child
interactions. As indicated in Figure 5-8 and Table 5-6, Pb exposure has been shown to
increase corticosterone levels and exacerbate Pb-induced dopamine release and learning
ability. Cory-Slechta and colleagues have conducted multiple investigations in this area.
Most recently, they showed enhanced learning deficits in female rat offspring following
lifetime Pb exposure combined with maternal restraint or prenatal stress (Corv-Slechta et
al.. 2010). This exposure paradigm used dams who were exposed to Pb for 2 weeks prior
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to mating through lactation and pups from a mixed sex litter from their dams during the
aforementioned period and then via drinking water Pb (50 ppm) through the remainder of
their lifetime. The resultant blood Pb levels of dams and pups ranged from 5 to 13 (ig/dL.
In a separate investigation, Pb plus stress-related outcomes were examined in female
offspring of dams who were exposed to Pb from 2 months prior to mating through
lactation, i.e., developmental Pb exposure (2 exposure groups: 50 or 150 ppb Pb-acetate
in drinking water) (Vimolini et al.. 2008a'). Dams were subject to restraint stress at
GD16-17. Compared with controls, marked increases in response rates on FI performance
were found in the Pb plus stress female offspring, whose mean blood Pb level was
11 (ig/dL (50 ppb Pb-acetate). Because these animals did not show effects with maternal
stress or Pb exposure alone, the results indicated a potentiation of effects with Pb and
stress co-exposures.
Similarly, lifetime Pb exposure (50 or 150 ppm, resulting in blood Pb levels of 11-16 and
25-33 (ig/dL, respectively) plus stress (maternal or offspring) also induced FI aberrations
at the post-reinforcement pause (PRP) period in female offspring (Rossi-George et al..
2011) (Table 5-7). Again, the results indicated a potentiation of effects. Within the FI
schedule, the PRP represents timing capacity or proper temporal discrimination. Namely,
the PRP is the period during which the animal must wait or pause before depressing the
lever for a reward. In this case, Pb plus stress exposed animals started responding too
early due to a decreased pause or PRP interval. Aberrant FI performance in infants and
children has been used as a marker for impulsivity. Separately, overall FI response rate, a
hyperactive behavior, was significantly increased with Pb exposure alone and with
maternal or offspring stress at the 50 ppm exposure dose. At 150 ppm Pb, stress
(maternal or offspring) increased FI response rate but Pb alone had no effect on FI.
Biochemical analysis of possible mechanistic contributions to these aberrations revealed
alterations in frontal cortex norepinephrine, reductions in dopamine homeostasis in the
nucleus accumbens, and enhancement of the striatal monoamine system. This study on
the effect of lifetime Pb exposure with or without stress on FI testing itself or during the
PRP component of FI testing further affirms Pb-related learning deficits and provides
possible mechanistic explanations.
Pb exposure over various developmental windows has been shown to affect
corticosterone levels in rodents. These findings indicate that associations of Pb and stress
with learning deficits (FI testing in females) may be related to aberrations in
corticosterone and dopamine. Maternal Pb exposure (150 ppm drinking water from
2 months prior to mating through lactation with restraint stress as detailed above) induced
increased basal corticosterone in female and male offspring at 9 months of age; no
interactions of Pb and stress were observed in this model (Corv-Slechta et al.. 2004). By
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14 months of age, these offspring had reduced corticosterone concentrations versus
control animals, indicating a possible acceleration of age-related decreases in basal
corticosterone levels (Corv-Slechta et al.. 2008). Pb-induced decreases in corticosterone
were enhanced with maternal stress. Postnatal exposure of male rodents to Pb (PND
21-5 months of age) produced animals with significant decrements in baseline
corticosterone; this effect produced a U-shaped concentration-response curve with
significant decrements in basal corticosterone levels in the 50 ppm exposure group versus
control (Virgolini et al.. 2005). In summary, developmental (gestational and lactational)
and post-weaning exposure to Pb induced changes in the HPA axis (corticosterone levels)
in both sexes that were dynamic as the animal aged.
Mechanistic understanding of the cognitive deficits observed with Pb and/or stress
exposure was explored in a recent study. HPA hypofunction following dam Pb exposure
(pup gestational and lactational Pb exposure) with or without maternal stress was
reported (Rossi-George et al.. 2011; Virgolini et al. 2008a). This study used the same
model of developmental Pb exposure as is detailed in the preceding paragraph. Outcomes
were examined in both male and female offspring. At 2 months of age without stress,
basal corticosterone in females was significantly increased with 150 ppm Pb (resulting in
blood Pb level of 32 j^ig/dL at PND21) and 50 ppm Pb (resulting in blood Pb level of
19 (ig/dL) (Figure 5-9). Pb plus stress attenuated the Pb-induced elevations in
corticosterone to baseline levels (Figure 5-9). At age 10 months, Pb and stress accelerated
the age-dependent decrease in corticosterone levels in females. In males, basal
corticosterone levels were not affected significantly by Pb and/or stress at 2 or 10 months
of age (Figure 5-9). These authors also explored the function of the glucocorticoid
negative feedback loop using the dexamethasone suppression test and found that Pb
and/or maternal stress significantly impacted this negative feedback by increasing nuclear
glucocorticoid receptor levels. This negative feedback loop was impacted more at the
lower dose (50 ppm versus 150 ppm Pb-acetate). As summarized in Table 5-7, the results
indicate that lifetime Pb exposure when combined with stress can exacerbate impairments
in learning. The interaction between Pb and stress may be mediated via effects on
corticosterone and dopamine.
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Female
Male
900
600
O)
c
300
Basal
PbxS
Final
50 150	0 50 150
Pb Exposure (ppm)
900-1
600-
300-
Basal
S:PS
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the 50 ppm exposure group only (not in 150 ppm Pb exposure group). On the other hand,
females had concentration-dependent corticosterone responses to Pb exposure in both
exposure models (lifetime Pb exposure and dam Pb exposure). Maternal stress alone also
led to HPA axis negative feedback hypofunction. Pb plus maternal stress enhanced
negative feedback in males and attenuated this effect in females. Pb exposure with or
without maternal stress prolonged the effect of DEX-dependent corticosterone
suppression in males. These data together show that HPA axis alterations could provide a
link between the contribution of Pb and stress to health effects.
Schedule-controlled behavior is often measured using FI or FR testing. Because the FI
animals are regularly handled by laboratory personnel and participate in 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.
Specifically, the corticosterone differences between FI and NFI animals after
developmental Pb exposure (dam-only Pb exposure) have been measured. Virgolini et al.
(2008b) found that basal corticosterone levels were significantly different between FI and
NFI animals. Also, the combination of dam Pb exposure with maternal stress was
explored in FI and NFI animals. At the baseline age of 4-5 months, NFI animals
displayed significant differences from FI animals. Pb exposure with or without stress did
not induce differences in corticosterone levels in FI females. The corticosterone level of
male FIs was affected by Pb and stress exposure (Virgolini et al.. 2008b'). In the FI males,
the 50 ppb Pb exposure group (50Pb) had decreased corticosterone versus control (no Pb
exposure) and the 150 ppb Pb exposure group (150Pb) had elevated corticosterone versus
control. Male NFI animals showed a U shaped concentration-response curve with 50Pb
animals having significantly less corticosterone than did control or 150Pb animals. In the
NFI males, stress did not affect corticosterone levels or interact with the effect of Pb. NFI
females exposed to 150Pb had significantly elevated corticosterone versus control (no Pb
exposure). These data demonstrate that behaviorally trained animals have altered HPA
axis and response to Pb exposure versus animals that are housed under conditions without
daily handling by caregivers.
Virgolini et al. (2008b) also expanded evidence for Pb exposure-stress interactions
through the examination of the effects additional intermittent stress as an adult. The
authors proposed that associations of Pb and stress with learning deficits (FI testing in
females) may be related to aberrations in corticosterone and dopamine. Dam exposure to
Pb (50 or 150 ppm Pb-acetate) followed by intermittent stress (cold, novelty or restraint)
in offspring as adults induced statistically significant changes in FI response rate.
Females were more sensitive to the adult intermittent stressors at the higher dose of Pb
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(150 ppm) with statistically significant increases in FI response rate and decreased PRP,
i.e., increased impulsivity (Figure 5-10). Males were more sensitive (decreased FI
response rate due to decreased run rate) to the restraint stress at the lower Pb dose
(50 ppm). At the higher dose of Pb, males were more sensitive to the cold stress
(increased FI response rate and increased run rate) (Figure 5-11). Corticosterone levels
were examined in this study and showed concentration-dependent correlations with FI
outcomes in females but were independent of dose in males.
Another study examined female rats with lifetime Pb exposure combined with prenatal
stress and found enhanced learning deficits (drinking water 50 ppm Pb-acetate, resulting
in offspring blood Pb levels: 7-13 (.ig/dL) (Corv-Slcchta et al.. 2010). Learning was
evaluated with multiple schedule of repeated learning (RL) and performance testing. RL
was impaired, but performance was not affected with Pb exposure. The Pb-impaired RL
was further enhanced with prenatal stress. There were statistically significant associations
between Pb/stress and corticosterone concentration, dopamine from the frontal cortex,
dopamine turnover in the nucleus accumbens, and total number of responses required to
learn a sequence. Also, Pb-exposed offspring with and without maternal stress exposure
had statistically significant decreases in hippocampal nerve growth factor (NGF) versus
controls. Thus, this study demonstrated that lifetime Pb exposure with or without prenatal
stress induced learning deficits in female mice.
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"l	ISO-i	120-1	.
j diri
0 50 ISO	0 SO 150
V
+	1»T	120-1	*
J MM
0 50 ISO	0 50 ISO
Paztretnforcwricnt Pause Time
0 I	300 -I	126-|
L L II
50 150	0 50 150
|Pb (pfxn)]	[Pb (ppm)]
Source: Reprinted with permission of Elsevier Science, Virgolini, Rossi-George, Weston, et al. (2008b).
Figure 5-10 Changes in Fl performance (Fl overall performance, run rate, PRP)
in female offspring with maternal Pb exposure plus various
stressors (restraint, cold, novelty) in adulthood.
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Restraint
Overall Rate
200 150 -
y i
0 50 150 0
Run Rate
s
Percent of Basal
m
250 ISO ¦
ftl'J
0 50 150 0
Percent of Basal
501 150-i
Percent of Basal
S 1 S
0 50 150
[Pb (ppm)]
0 50 150
[Pb (ppm)]
0 SO 150
[Pb(ppm)]
Source: Reprinted with permission of Elsevier Science, Virgolini, Rossi-George, Weston, et al. (2008b).
Figure 5-11 Changes in Fl performance (Fl overall performance, run rate, PRP)
in male offspring with maternal Pb exposure plus various
stressors (restraint, cold, novelty) in adulthood.
Cognitive Flexibility
Cognitive flexibility is a component of executive function that measures the ability to
reallocate mental resources when situations change (Monsell. 2003). This flexibility is
assessed as the ability to alter behavioral responses according to the context of the
situation and incorporates the function of attention, working memory, and visual
processing. Discrimination reversal learning and concurrent random interval (RI-RI)
scheduling are used to measure cognitive flexibility. The 2006 Pb AQCD reported
discrimination reversal learning deficits in monkeys with blood Pb levels of 11-20 (ig/dL.
Rats also showed similar deficits but the authors attributed the changes to learning-related
problems instead of cognitive flexibility (Garavan et al.. 2000; Hilson and Strupp. 1997).
Interestingly, recent work has shown that N-Methyl-D-aspartic acid or N-Methyl-D-
aspartate (NMDA) receptors and D2-like receptors, two well-characterized targets of Pb,
are involved in discrimination reversal learning (HeroId. 2010). Another test of cognitive
flexibility is called concurrent random interval (RI-RI) scheduling in which depression on
two response levers is reinforced at different frequencies. The 2006 Pb AQCD reported
monkeys with Pb-induced cognitive flexibility impairment under RI-RI (Ncw land et al..
1994). Coherence for these observations of Pb-induced deficits in cognitive flexibility is
provided by consistent evidence in animals and children for associations of Pb with
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individual tests of inattention, memory, and visual processing. Studies in children also
demonstrate associations between blood Pb levels and decrements in specific tests of
cognitive flexibility that are homologous to tests in animals such as discrimination
reversal learning, the WCST, and Stoop test (Figure 5-5 and Table 5-5). While tests in
animals measure the ability to complete a task according to changes in reinforcement,
tests in humans assess the ability of subjects to complete a task according to changes in
rules.
5.3.2.3 Toxicological Studies on the Effects of Chelation
Earlier work in the animal toxicological literature has shown that succimer or chelation
treatment of Pb-exposed lab animals was able to normalize various aberrant Pb-induced
behaviors including activity level, habituation (Gong and Evans. 1997) and forced-swim
immobility (Stewart et al. 1996). A more recent study examined the effect of succimer
treatment on various behavioral and cognitive outcomes in control and neonatally Pb-
exposed female animals (PND 1-30 Pb-acetate exposure, 300 ppm dam through lactation
and either 30 or 300 ppm pup water) by drinking water, generating a moderate Pb (m-Pb)
exposure and a high Pb (h-Pb) exposure group. Pb blood levels at PND52 in the control,
m-Pb, h-Pb, m-Pb+succimer, and h-Pb+succimer were 1.5; 12.6; 31; 2.8; and 8.5 (ig/dL,
respectively. Brain Pb levels at the same time for the same groups were 41 (control),
1,040 (m-Pb), 3,690 (h-Pb); and 196 (m-Pb+succimer) and 1,370 (h-Pb+succimer) ng/g
dry weight. Succimer treatment significantly attenuated the m-Pb induced impaired
learning ability. Effects on arousal that were significantly affected in h-Pb rats were
significantly attenuated with succimer treatment. Succimer treatment in the h-Pb animals
only slightly improved learning ability but did not improve the impaired inhibitory
control (Stangle et al.. 2007). These are important findings because they provide evidence
that certain neurobehavioral or cognitive impairments associated with Pb exposure appear
to be reversible with chelation therapy.
In another study, a 3-week course of Pb-acetate (PND 1-17, dam drinking water) plus or
minus succimer/chelator (PND 31-52) treatment was given to determine if succimer
could alleviate behavioral deficits in rats exposed to Pb for the first 4 weeks of life. Pb-
exposed animals had altered reactivity and increased reward omission and errors. Pb-
exposed animals receiving chelation treatment had normalized reactivity to reward
omission and errors (Beaudin et al.. 2007). Pb-induced behavioral abnormalities were
attenuated with chelation therapy.
The interaction between Pb and methionine choline also has been examined (Fan et al.
2009a'). Pb-exposed rats were supplemented with methionine choline to provide
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information on its effect on Pb disposition in various tissues (blood, bone, brain) and its
contributions to neurocognitive or neurobehavioral changes. As a sulfur source,
methionine is a chelator and a free-radical scavenger. Choline is important for cell
membranes and neurotransmitter synthesis (Zeisel and Blusztain. 1994V In this model,
methionine choline attenuated Pb-induced memory and learning deficits
(Section 5.3.2.2). Exposure of weanling male rats to Pb-acetate in drinking water
(300 mg/L) through PND60 produced a blood Pb level of 60 (ig/dL, bone Pb level of
165 jj.g/g, and brain Pb level of 0.63 jj.g/g. Methionine choline supplementation
significantly attenuated blood and bone Pb levels but produced a nonsignificant
attenuation of brain Pb (0.51 jJ.g/g) in rats that had significant improvements in learning
and memory (Section 5.3.2.2). Also, in another study, the metal chelators DP-109 and
DP-460 were neuroprotective for Pb-related effects in the ALS mouse neurodegenerative
model or Tg(SODl-G93A) model (Petri et al. 2007).
In summary, succimer or chelation treatment appears to be able to restore Pb-dependent
impairments of learning and arousal and be neuroprotective in a concentration-dependent
fashion. In these studies, succimer use was more efficacious at lower doses of Pb
exposure. Chelation did not restore Pb-impaired inhibitory control. Chelation with the
antioxidant supplement affected the disposition of Pb in various tissues, significantly
attenuating blood and bone Pb levels and nonsignificantly attenuating brain Pb.
5.3.2.4 Integrated Summary of Cognitive Function in Children
Results from recent epidemiologic studies and animals studies expand the strong existing
evidence base demonstrating that Pb exposure is associated with impaired cognitive
function in children. A large epidemiologic evidence base demonstrates associations of
higher blood Pb level with lower FSIQ in school-aged children, lower MDI scores in
children ages 6 months to 3 years, and lower scores on tests of specific cognitive
functions in children ages 6 months to age 16 years (Figure 5-2 and Figure 5-5; Table
5-3, Table 5-4, and Table 5-5). Blood Pb level has been associated with a range of
cognitive indices, including memory, verbal and math skills, and cognitive flexibility.
There was no clear indication that blood Pb level was more strongly associated with
performance in a particular domain of cognitive function. Studies in animals clearly
demonstrated impaired performance on Morris Water Maze, the delayed spatial
alternation, and discrimination reversal learning with Pb exposure. These animal
observations provided strong coherence with associations in children with homologous
tests of spatial memory and rule learning and reversal (Figure 5-5 and Table 5-5). Several
studies found that blood or dentin Pb levels measured at an early age (ages 2-8 years)
were associated with poorer academic performance at older ages (ages 8-18 years),
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suggesting the effect of early exposure to Pb may be persistent and have important
implications for success later in life.
The weight of epidemiologic evidence for Pb-associated cognitive function decrements
was provided by several prospective studies conducted in diverse populations that found
associations with blood Pb levels measured during various lifestages (concurrently,
prenatally (cord or maternal), earlier in childhood, and averaged over multiple years) and
had extensive evaluation of potential confounding variables. The association with FSIQ
was substantiated in a pooled analysis of children, 5 to 10 years of age, 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). Several new studies added evidence for associations of FSIQ with concurrent
blood Pb level in populations with lower mean blood Pb levels. Previously, the weight of
evidence supported these associations in populations with mean blood Pb levels in the
range of 5-10 (ig/dL. However, several new studies shifted the weight of evidence to
lower blood Pb levels (primarily concurrent), with populations means in the range of
2-7 (.ig/dL (2011; Kim et al.. 2009b; Min et al . 2009: Znilina et al.. 2008; Chiodo et al..
2007). Studies of specific cognitive indices and academic performance did not
consistently find decrements in performance in children with blood Pb levels < 5 (ig/dL
(C'ho et al.. 2010; Miranda et al. 2010; Chandramouli et al.. 2009; Surkan et al.. 2007).
Several new toxicological studies added to the evidence for impaired learning and
memory in animals with lower blood Pb levels, 8-17 (.ig/dL (Corv-Slechta et al. 2010; Li
et al.. 2009c; Niu et al.. 2009; Virgolini et al. 2008a; Stannic et al.. 2007). A large body
of new evidence from Cory-Slechta and colleagues demonstrated that lifetime Pb
exposure when combined with stress in animals exacerbated impairments in learning
(2011; Corv-Slechta et al.. 2010; Rossi-George et al. 2009; Virgolini et al.. 2008a).
Findings also indicated that these interactions potentially were mediated via effects on
corticosterone and dopamine.
With regards to important lifestages of Pb exposure, the weight of toxicological evidence
demonstrates impaired learning and memory in animals exposed to Pb gestationally with
or without early postnatal exposure. In particular, impairments in learning and memory
observed with lower blood Pb levels (8-17 (ig/dL), were found with Pb exposures that
began during the gestational or lactation period. The prospective epidemiologic studies
found decrements in cognitive function of children in association with concurrent,
prenatal (cord and maternal), early childhood, and cumulative average blood Pb levels.
However, collectively, based on the frequency of examination, the weight of evidence
demonstrates associations between cognitive function decrements and concurrent blood
Pb levels in children ages 4-10 years. Among studies that examined MDI in children ages
6 months to 3 years, several found stronger associations of MDI with prenatal (maternal
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or cord) blood Pb than with postnatal child blood Pb (Hu et al.. 2006; Bellinger et al..
1987; Dietrich et al.. 1987a; Vimpani et al.. 1985). Child concurrent or lifetime average
blood Pb levels also were associated with MDI scores (Claus Henn et al.; Solon et al.
2008; Surkan et al.. 2007; Tellez-Roio et al.. 2006; Wasserman et al.. 1992; Vimpani et
al.. 1985). Thus, both postnatal child and maternal Pb exposures may contribute to lower
cognitive function in young children. It is important to note that in studies that examined
maternal pregnancy or cord blood Pb levels, there is greater uncertainty regarding the
relative contributions of maternal past and recent Pb exposures that contributed to
associations. The influence of past or cumulative maternal Pb exposures was supported
by observations that maternal patella Pb levels were associated with MDI among children
at age 2 years in Mexico City (Gomaa et al.. 2002). Nonetheless, results for prenatal
blood Pb levels and concurrent child blood Pb levels measured during specific postnatal
ages indicate that relatively short-duration Pb exposures (< 1 year) influence the
cognitive development of children between birth and age 3 years. Short-duration Pb
exposures may be especially important during early childhood because processes such as
neurogenesis and synaptic pruning are highly active during the first few years of life
(Rice and Barone. 2000; Landrigan et al.. 1999).
Epidemiologic studies found independent associations of blood Pb level after adjusting
for multiple potential confounding factors, including parental IQ, SES, household
income, and HOME score. In most studies that provided unadjusted and adjusted effect
estimates, blood Pb level was found to be associated with a smaller but statistically
significant decrement in FSIQ after adjusting for potential confounding factors
(Palamappaii et al.. 2011; Kim et al.. 2009b; Chiodo et al.. 2007; Froehlich et al.. 2007;
Lanphear et al.. 2005; Can field et al.. 2003a). Analyses of associations of covariates with
blood Pb level and cognitive function indicated that the potential confounding variables
may vary across populations and endpoints. HOME score was not associated with every
cognitive index. Studies that adjusted for multiple SES-related factors but did not
examine HOME score produced similar magnitudes of associations as did studies that
adjusted for HOME score (Table 5-3, Table 5-4, toTable 5-5). Thus, confounding by
HOME score may be minimized when other correlated variables are accounted for in
analyses. Further, it is uncertain the extent to which HOME score alone may confound
blood Pb-IQ associations as studies have not reported the magnitude of change in the
blood Pb effect estimate with just the addition of HOME score in the model. Thus, while
the caregiving environment can be an important confounder, the overall weight of
evidence indicates that it does not mitigate the strong findings linking higher blood Pb
levels with lower FSIQ in children. Prenatal drug exposure also was not found to have a
large influence on the relationship between blood Pb level and FSIQ (Min et al. 2009;
Chiodo et al. 2007). That Pb exposure induces impairments in tests of learning and
memory in animals further that are directly homologous to tests conducted in children
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further demonstrates that confounding by SES and HOME does not fully account for the
associations between blood Pb levels and cognitive function observed in children.
The evidence from epidemiologic and toxicological studies linking Pb exposure with
decrements in multiple cognitive functional domains is strengthened 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 5.3.8). Pb has been shown to increase the permeability of the blood-brain barrier
and deposit in the target central nervous system. 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 young
children being particularly at increased risk for Pb-associated impairments 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 release in these regions.
Numerous studies also have shown Pb-induced changes in hippocampal function,
including changes in glutamate release, receptor binding, and long-term potentiation.
Thus, several lines of toxicological evidence establish a neuroanatomical and
neurochemical basis for the effect of Pb on cognitive function
5.3.2.5 Epidemiologic Studies of Cognitive Function in Adults
Adults without Occupational Lead Exposures
As described in the preceding section, Pb exposure of animals that begins in gestation
and lasts through the early postnatal period or for a lifetime 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. In contrast, epidemiologic studies have
examined cognitive performance in adults primarily in association with concurrently
measured blood and bone Pb levels. For nonoccupationally-exposed adults, the 2006 Pb
AQCD cited some evidence for associations of cognitive performance with bone Pb
levels but not blood Pb levels (U.S. EPA. 2006b). Studies published since the 2006 Pb
AQCD continued to indicate that bone Pb levels were associated more consistently with
cognitive function in nonoccupationally-exposed adults (Table 5-8). Despite the large
number of available publications, it is important to recognize that several studies are
variants on analyses in the same population (e.g., NHANES or the Normative Aging
Study [NAS]) and should not be considered as all independent assessments of the Pb-
cognitive function relationship. Another point to consider is that although cross-sectional
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and longitudinal studies have been conducted, most examine the associations between
one measurement of cognitive function and one concurrent measurement of blood Pb
level. Such a design has the inherent weakness that temporality cannot be determined,
which limits the causal inference regarding the effects of Pb exposure. When bone Pb
level is used as the exposure measure, this issue is mitigated somewhat because bone Pb
is a biomarker that reflects many cumulative years of exposure. Additionally, with single
measurements of both blood and bone Pb level, it is difficult to characterize the specific
timing, duration, and level of Pb exposure that contributed to the associations observed
with cognitive function. This uncertainty may apply particularly for assessments of blood
Pb levels, which are influenced both by current exposures and cumulative Pb stores in
bone that are mobilized during bone remodeling compared with occupationally-exposed
adults (Sections 4.3 and 4.7.3).
Several analyses of the large prospective Baltimore Memory Study and NAS have
contributed to greater understanding of the relationship between biomarkers of Pb and
cognitive effects in adults. In both cohorts, subjects were evaluated periodically with in-
person clinical assessments and self-administered questionnaires. Both studies were noted
for comparisons of associations between bone and blood Pb levels within the same cohort
and for their repeated assessments of Pb biomarker levels and cognitive function over
time. In particular, the repeated assessments permitted the examination of associations of
Pb biomarkers with changes in cognitive function over time, which allowed investigators
to establish temporality between Pb biomarker levels and subsequent changes in
cognitive function. The Baltimore Memory Study (BMS) included men and women,
50-70 years of age, 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. Of
particular note, 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 Greater Boston area. Subjects primarily were white and at enrollment were
aged 21 to 80 years and had no current or past chronic medical conditions. Studies
differed with respect to the potential confounding variables analyzed. The most notable
difference was the inclusion of age, smoking and alcohol intake as covariates in the NAS
analyses. Despite differences in study population characteristics and potential
confounding variables considered, findings were similar between cohorts.
In the BMS, longitudinal analyses were conducted with repeat cognitive testing of study
subjects at approximately 14-month intervals. Most subjects completed follow-up; 91%
of the original cohort returned for a second round of testing and 83% for a third round
each (Bandeen-Roche et al.. 2009). An interquartile range higher tibia Pb level (12.7
jxg/g) was associated with a 0.019 units per year decrease in eye-hand coordination z-
score, adjusting for age, sex, interviewer, race and SES. Tibia Pb was associated with a
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larger decrease among African Americans than among whites (Table 5-8). Tibia Pb levels
were weakly associated with other time-related decreases in other indices of cognitive
function (e.g., language, processing speed, executive function).
Cross-sectional analyses of the BMS included comparisons of the associations for blood
and bone Pb level. The 991 adults in Shih et al. (2006) had a mean (SD) blood and tibia
Pb levels of 3.46 (2.23) (ig/dL and 18.7 (11.2) jxg/g, respectively, Whereas higher
concurrent blood and bone Pb level both were associated with poorer performance in the
domains of language, processing speed, eye-hand coordination, executive function,
verbal memory and learning, visual memory, and visuoconstruction, associations with
tibia Pb level tended to be larger in magnitude (per SD increase) and statistically
significant (Table 5-8). Tibia Pb levels were associated with worse performance on tests
in all domains in models adjusted for age, sex, testing technician, and presence of the
apolipoprotein (APO)E-s4 allele (potential risk factor for Alzheimer's Disease). The
magnitudes of associations were attenuated with additional adjustment for education,
race, and SES. In these fully-adjusted models, higher tibia Pb levels were associated with
poorer performance in domains except language and processing speed, with a borderline
statistically significant association observed for visuoconstruction. In linear models,
visuoconstruction scores were 0.0044 SDs (95% CI: -0.0091, 0.0003) lower per 1 jxg/g
bone higher tibia Pb level. Analysis of a quadratic term for tibia Pb indicated no evidence
of nonlinearity.
Other cross-sectional analyses indicated effect modification by race and neighborhood
psychosocial hazards. In contrast with the longitudinal results, 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
functioning, and verbal memory and learning among whites than among African
Americans (Bandeen-Roche et al.. 2009) (Table 5-8).
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Table 5-8 Associations of blood and bone Pb levels with cognitive function in
adults9
Study
Population/
Methods
Cognitive Test
Subgroup/ Blood Pb Effect Estimate
Model	(96% Cl)b
Bone Pb Effect Estimate
Cl)b
Bandeen-
Roche et
al. (2009)
1,140 adults, ages 50-70 yr
Baltimore Memory Study
(BMS) cohort
Baltimore, MD
Marginal linear regression
models adjusted for age,
household wealth,
education, race/ethnicity
Longitudinal associations
Eye/hand coordination
African-Americans
White
Cross-sectional associations
Verbal memory/learning
African-Americans
White
NOT EXAMINED
Tibia Mean (SD): 18.8(11.6) pg/g
Per 12.7 pg/g (IQR) increase
-0.032 (-0.052, -0.012) per year
-0.009 (-0.024, 0.006) per year
0.006 (-0.09, 0.10)
-0.076 (-0.15, 0.001)
Executive Function
African-Americans
White
0.009 (-0.05, 0.07)
-0.07 (-0.12,-0.01)
Shih et al. 985 adults, mean age: 59 yr
2QQ6) BMS, Baltimore, MD
Linear regression adjusted
for:
Model A: age, sex,
technician, presence of
APOE-E4 allele
Model B: Model I, years of
education, race/ethnicity,
wealth
Raven's Colored Progressive
Matrices
Language
Eye-hand coordination
Executive functioning
Visuoconstruction
Concurrent blood Pb Mean Tibia Mean (SD): 18.7 (11.2) pg/g
(SD): 3.5 (2.2) pg/dL
Model A
Model B
Model A
Model B
Model A
Model B
Model A
Model B
-0.013 (-0.064, 0.037)
-0.004 (-0.044, 0.036)
-0.024 (-0.0704, 0.022)
-0.017 (-0.057, 0.024)
-0.031 (-0.075, 0.012)
-0.022 (-0.069,0.016)
-0.042 (-0.101,0.017)
-0.031 (-0.084, 0.021)
-0.0896 (-0.146, -0.034)
0.007 (-0.034, 0.047)
-0.0896 (-0.134, -0.046)
-0.034 (-0.067,0)
-0.0896 (-0.123, -0.066)
-0.016 (-0.056, 0.025)
-0.024 (-0.190, 0.141)
-0.049 (-0.101,0.002)
TIBIA Pb STRONGER
Glass et 1,001 adults,
al. (2009) mean ag0 gg yr
BMS, Baltimore, MD
Multilevel hierarchical
regression model adjusted
for age, sex, race/ethnicity,
education, testing
technician, time of day
Raven's Colored Progressive
Matrices
Language
Eye-hand coordination
Executive functioning
Visuoconstruction
NOT EXAMINED
Middle NPH
High NPH
Middle NPH
High NPH
Middle NPH
High NPH
Middle NPH
High NPH
Tibia Pb Mean (SD): 18.8(11.1)
pg/g
0.011 (-0.089, 0.111)c
-0.09 (-0.189, -0.011)c
-0.04 (-0.133, 0.044)c
-0.067 (-0.167, 0.033)c
-0.022 (-0.111, 0.067)c
-0.111 (-0.1998, -0.022)c
-0.003 (-0.155, 0.149)c
-0.007 (-0.189, 0.175)c
VNfeisskopf
etal.
(2007a)
1,089 males,
mean age 68.7 yr;
Normative Aging Study
(NAS), Boston, MA
Linear repeated measures
analysis adjusted for age,
age squared, education,
smoking, alcohol intake, yr
between bone Pb
measurement and first
cognitive test, yr between
cognitive tests
CERAD, Neurobehavioral
Evaluation System, WIAS-R,
MMSE, VMI
Visuospatial, pattern
comparison (+ = poorer
performance)
Executive function verbal
fluency
Short-term memory, word list
NOT EXAMINED
Mean (IQR): Tibia: 20 (15) |jg/g
Patella: 25 (20) pg/g
Estimates per IQR increase:
Tibia: 0.79 (0.40,1.2) overtime
Patella: 0.73 (0.40,1.2) overtime
Tibia: -0.40 (-1.6, 0.80) overtime
Patella: -0.86 (-2.00, 0.30) over
time
Tibia: -0.28 (-1.2, 0.60) overtime
Patella: -0.81 (-1.7, 0.05) overtime
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Population/
Subgroup/
Blood Pb Effect Estimate
Bone Pb Effect Estimate (96%
Study
Methods Cognitive Test
Model
(96% Cl)b
Cl)b
\Nang et
358 males, median ages:

NOT EXAMINED
Median: 19 pg/g (Tibia), 23 pg/g
al. (2007a)
67.2 yr (HFE wild-type)


(Patella)

67.7 yr (HFE variant)


-0.20 (-1.0, 0.70) per year

Normative Aging Study, MMSE
HFE wildtype


Boston, MA
One HFE variant

-1.40 (-3.3, 0.40) per year

Linear regression adjusted
Two HFE variants

-6.3 (-10.4, -2.1) per year

for age, years of education,



nonsmoker, former smoker,




pack-years, nondrinker,




alcohol consumption,




English as first language,




computer experience,




diabetes



Wfeuve et
720 males, ages > 45 yr

Concurrent blood Pb Mean
Median (IQR): Tibia: 19 (15) pg/g;
al. 12006)
NAS, Boston, MA
Linear mixed effects
regression adjusted for

(IQR): 5.2 (3) pg/dL
Patella: 27 (21) pg/g
Tibia

smoking status, alcohol
consumption, calorie MMSE score
ALAD wildtype
-0.013 (-0.053, 0.027)
-0.03 (-0.14,0.07)

adjusted calcium intake,
ALAD2 carrier
-0.087 (-0.180, 0.007)
-0.11 (-0.30,0.06)

regular energy expenditure


Patella

on leisure time physical
activity, diabetes
ALAD wildtype

-0.03 (-0.11,0.04)

ALAD2 carrier

-0.12 (-0.30,0.06)
Rajan et 720 males, ages > 45 yr
al. (2008) Normative Aging Study,
Boston, MA
Linear regression adjusted
for blood Pb main effect,
ALAD genotype, age at
cognitive test, education,
alcohol consumption,
cumulative smoking, English
as first language
CERAD, Neurobehavioral
Evaluation System, WIAS-R
Visuospatial, constructional
praxis
Executive function verbal
fluency
Verbal memory, word recall
Concurrent blood Pb Mean Mean (SD):
(S5!: 5'^ ?!l^dL Tibia: 219 C13'8) ijg/g (alad
wildtype), 4.8 (2.7) pg/dL wiidtype), 21.2 (11.6) pg/g (ALAD2
carriers)
Patella: 29.3 (19.1) pg/g (ALAD
wildtype), 27.9(17.3) pg/g (ALAD2
carriers)
(ALAD2 carriers)
-0.048 (-0.216, 0.120)"
-0.028 (-0.2044, 0.148)"
0.003 (-0.168, 0.174)"
Tibia: -0.216 (-0.419, -0.013)"
Patella: 0.018 (-0.182, 0.218)"
Tibia: -0.089 (-0.292,0.1143)"
Patella:-0.018 (-0.218 0.1082)"
Tibia: 0.064 (-0.127, 0.254)"
Patella: 0.127 (-0.728, 0.989)"
Perceptual speed, mean
latency
-0.168 (-0.392,0.066)"
Tibia:-0.152 (-0.495, 0.191)"
Patella:-0.146 (-0.40, 0.109)"
TIBIA Pb STRONGER
V\feuve et 587 females, ages 47-74 yr
al. (2009) Nurses> Hea!th study^
Boston, MA
Generalized estimating
equations adjusted for age,
age-squared at Pb
assessment, age at
cognitive assessment,
education, husband's
education, alcohol
consumption, smoking
status, physical activity,
aspirin use, ibuprofen use,
use of Vitamin EE
supplements, menopausal
status and postmenopausal
hormone use
Telephone Interview for
Cognitive Status and East
Boston Memory Test
Composite cognitive score
Composite except letter
fluency
Concurrent blood Pb Mean Mean (SD)
(SD): 2.9 (1.9) pg/dL	Tibia pb: 10 5 (9 7) pg/g
-0.015 (-0.068, 0.038)
0.015 (-0.070, 0.101)
Patella Pb: 12.6(11.6) pg/g
Tibia: -0.039 (-0.087,0.0097)
Patella:-0.012 (-0.058, 0.035)
Tibia: -0.049 (-0.097,0)
Patella:-0.035 (-0.081,0.012)
TIBIA Pb STRONGER
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Study
Population/
Methods
Cognitive Test
Subgroup/ Blood Pb Effect Estimate
Model	(96% Cl)b
Bone Pb Effect Estimate
Cl)b
Krieg and
Butler
2,823 adults, ages 20-59 yr,
U.S. 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
Neurobehavioral Evaluation
System 2
Symbol Digit Substitution
(mean total latency, sec)
Serial digit learning score
Ages 20-39 yr
Ages 40-59 yr
Ages 20-39 yr
Ages 40-59 yr
Concurrent Blood mean
(SD): 2.88 (6.91) |jg/dL
-0.097 (-0.422, 0.228)6
-0.290 (-0.601, 0.0207)6
-0.117 (-0.463, 0.228)6
0.401 (-0.193, 0.995)6
NOT EXAMINED
Krieg et al. 2,090 adults, ages 20-59 yr
(2QQ9) 1976 adults, ages > 60 yr
U.S. 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 hrs,
test language (20-59 yr) and
sex, age, education, family
income, race-ethnicity, test
language (> 60 yr)
Neurobehavioral Evaluation
System 2
Symbol Digit Substitution
(mean total latency, sec)
Serial digit learning score
VNford recall
Story recall
Ages 20-59 yr
ALAD GG
ALAD CC/CG
Ages 20-59 yr
ALAD GG
ALAD CC/CG
Ages > 60 yr
ALAD GG
ALAD CC/CG
Ages > 60 yr
ALAD GG
ALAD CC/CG
Concurrent Blood Pb Mean
(SD):
20-59 yr: 2.85 (7.31) pg/dL;
> 60 yr: 4.02 (3.56) pg/dL
-0.132 (-0.358, 0.095)6
-0.526 (-1.118, 0.066)e
-0.022 (-0.526, 0.482)6
0.025 (-0.406, 0.456)e
-0.075 (-0.285, 0.135)6
0.025 (-0.406, 0.456)e
0.085 (-0.0997, 0.271)6
-0.466 (-1.072, 0.139)6
NOT EXAMINED
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Study
Population/
Methods
Cognitive Test
Subgroup/ Blood Pb Effect Estimate
Model	(96% Cl)b
Bone Pb Effect Estimate
Cl)b
Krieg et al. 2,093 adults, ages 20-59 yr
(2212) 1,799 adults, ages >60 yr
U.S. 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 three
hours, test language (20-59
yr) and sex, age, education,
family income, race-
ethnicity, test language (> 60
yr)
Neurobehavioral Evaluation
System 2
Symbol Digit Substitution
(mean total latency, sec)
Serial digit learning score
(+ = poorer performance)
VNford recall
Story recall
Ages 20-59 yr
VDR haplotype
CC
VDR haplotype
CT
VDR haplotype
TC
VDR haplotype 1
Ages 20-59 yr
VDR haplotype
CC
VDR haplotype
CT
VDR haplotype
TC
VDR haplotype 1
Ages > 60 yr
VDR haplotype
CC
VDR haplotype
CT
VDR haplotype
TC
VDR haplotype TT
Ages > 60 yr
VDR haplotype
CC
VDR haplotype
CT
VDR haplotype
TC
VDR haplotype TT
Concurrent blood Pb Mean
(SD): 20-59 yr: 2.85 (7.32)
pg/dL; > 60 yr: 4.02 (3.39)
Mg/dL
-3.916 (-8.638, 0.805)6
0.139 (-0.278, 0.556)e
-0.505 (-1.025, 0.015)e
-0.695 (-0.783, 0.871)6
-2.533 (-4.868, -0.198)6
-0.322 (-0.922, 0.278)6
0.447 (0.542, 0.351)6
0.044 (-0.783, 0.871 )e
-0.766 (-1.817, 0.285)
-0.085 (-0.40, 0.21)
-0.034 (-0.471, 0.403)
-0.095 (-0.895, 0.705)
0.146 (-1.674, 1.966)
0.003 (-.193, 0.20)
0.034 (-0.322, 0.3899)
-0.166 (-0.434, 0.102)
NOT EXAMINED
Gaoetal. 188 adults, mean age 69.2 CERAD, CSID, IU story recall,
(2008) yr	Animal fluency test, IU token
test
Sichuan and Shandong
Provinces, China
ANCOVA adjusted for age,
sex, education, BMI,
APOE £4
Composite cognitive score
Concurrent plasma Pb
Mean (SD): 0.39 (0.63)
Mg/dL
-0.006 (-0.016, 0.004)
NOT EXAMINED
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2
3
4
5
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7
8
9
10
11
12
13
14
15
16
17
Study
Population/
Methods
Cognitive Test
Subgroup/ Blood Pb Effect Estimate
Model	(96% Cl)b
Bone Pb Effect Estimate (
Cl)b
Van
Wijngaarden
et al. (2009)
47 adults, mean age 61.5
yr
Rochester, NY
Linear regression adjusted
for age, gender,
educational level, history
of hypertension
CANTAB and Montreal
Cognitive Assessment
Delayed matching, % correct
NOT EXAMINED
Mean (SD): Tibia: 2.0 (5.2) pg/g;
Calcaneus: 6.1 (8.5) pg/g
Calcaneus
Lowest fertile: 87.56f
Medium fertile: 86.67
Highest Tertile: 80.67, p = 0.027
Tibia
Lowest tertile: 85.42f
Medium tertile: 87.08
Highest tertile: 82.44, p = 0.25
Calcaneus
Lowest tertile: 2.541
Total trials	Medium tertile: 2.61
Highest tertile: 2.72, p = 0.21
Tibia
Lowest tertile: 2.62'
Medium tertile: 2.59
Highest tertile: 2.66
"Studies are presented by cohort then generally in the order of discussion in the text.
'Effect estimates have been standardized to the standard deviation of the cognitive test scores and standardized to an SD or IQR increase in blood or bone Pb level.
"Effect estimates indicate interactions between Pb and category of neighborhood psychosocial hazard (NPH), with the lowest tertile of NPH serving as the reference
group.
dEffect estimates indicate interactions between Pb and ALAD genotype.
"The directions of effect estimates were changed to indicate a negative slope as a decrease in cognitive performance.
'Results refer to mean cognitive function scores among tertiles of bone Pb.
Similar to the BMS, in the NAS cohort, higher tibia Pb levels were associated with
decreases in cognitive performance over time ("Weisskopf et al.. 2007a'). Weisskopf et al.
("2007a') expanded evidence by finding associations with patella Pb levels. Two
measurements of cognitive function, collected approximately 3.5 years apart were
available for 60-70% of participants. Longitudinal analyses were conducted with repeated
measures plus a bone Pb-time interaction term in order to estimate the association
between bone Pb level and decline in cognitive test score overtime. Although bone Pb
levels were associated with increased response latency on a pattern comparison test, they
were associated with fewer errors on the same test. The authors proposed that this 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. In the analysis with patella Pb, Weisskopf et al. (2007a') found a nonlinear
association, with latency times becoming worse over time (i.e., larger values or slower
response time) up to approximately 60 jxg/g patella Pb, but the change over time leveling
off at higher levels (Figure 5-12). Below 60 jxg/g, a 20 jxg/g difference in patella Pb level
was associated with an increase in latency of approximately 0.15 ms.
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o
•q-
o
89 |jg/g bone
mineral) were removed. The estimate is indicated by the solid line and the 95% confidence interval by the dashed lines. Patella Pb
concentrations of all individual subjects are indicated by short vertical lines on the abscissa, (reference = 0 at mean of patella Pb
concentration).
Figure 5-12 Nonlinear association between patella Pb level and the relative
change over time in response latency on the pattern comparison
test.
Longitudinal analysis of the NAS cohort also indicated effect modification by
hemochromatosis (HFE) gene variants (Wang et al.. 2007a). In models adjusted for
age, years of education, nonsmoker, former smoker, pack-years, nondrinker, alcohol
consumption, English as first language, computer experience, and diabetes, an
interquartile range higher tibia Pb level (15(xg/g) was associated with a 0.22 point steeper
annual decline (95% CI: -0.39, -0.05) in MMSE score among men with either the H63D
and C282Y variant. The association was found to be nonlinear, with larger Pb-associated
declines observed at higher tibia Pb levels (Figure 5-13). This difference was comparable
to the difference in MMSE score between men who were 4 years apart in age in their
study sample. Tibia Pb level was associated with a smaller decline in MMSE score in
men with only HFE wildtype alleles (Figure 5-13). HFE variants, H63D and C282Y, are
associated with hemochromatosis, a disease characterized by higher iron body burden.
Higher iron body burden has been linked with lower Pb absorption, thus the apparent
interaction could be related to altered Pb toxicokinetics rather than some other direct
biological interaction. For example, the relation between tibia Pb level and Pb dose at the
biologically relevant site(s) could be shifted to the left among for HFE variant allele
carriers such that for the same tibia Pb level, Pb the dose at the relevant biological site(s)
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may be higher leading to a greater annual decline in MMSE score as indicated in Figure
5-13 (solid curve). This implies that similar decreases in MMSE score would still be
found in wildtype HFE carriers only at higher tibia Pb levels.
o
-0.2
-0.4
-0.6
HFE wildtype
HFE variant allele
-0.8
0
50
Tibia lead biomarker (,ug/g)
Source: Wang et al. (2007a).
Note: The lines indicate curvilinear trends estimated from the penalized spline method. Among HFE wild-types, the association
between tibia Pb and annual cognitive decline was nearly linear, but among variant allele carriers, the association tended to deviate
from linearity (p = 0.08). The model was adjusted for age, years of education, nonsmoker, former smoker, pack-years, nondrinker,
alcohol consumption, English as first language, computer experience, and diabetes.
Figure 5-13 Exploration of nonlinear association of tibia Pb level with annual
rate of cognitive decline, by hemochromatosis (HFE) gene variant.
The 2006 Pb AQCD described associations of blood and tibia Pb levels with poorer
cognitive performance among 141 NAS men (Pavton et al.. 1998). Several new and
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 genetic variants. In contrast with the longitudinal analyses, Weisskopf et al.
(2007a) found that repeat measures of bone Pb levels were not consistently associated
with cognitive function in cross-sectional analyses. In a study of Mini-Mental State
Examination (MMSE) tests scores (test of general cognitive function) among 720 men
45 years of age and older, higher concurrent blood Pb levels were associated with lower
MMSE scores among ALAD2 carriers (Weuve et al.. 2006). A 3 j^ig/dL higher concurrent
blood Pb level (the interquartile range) was associated with a 0.26 point lower mean
MMSE score (95% CI: -0.54, -0.01) among ALAD2 carriers and a 0.04 point lower score
(95% CI: -0.16, -0.07) among noncarriers. A subsequent study did not find a consistent
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direction of modification of the association between blood or bone Pb levels and other
tests of cognitive function by ALAD genotype (Raian et al.. 2008). Nonetheless, Raj an et
al. (2008) found tibia Pb levels to be associated more strongly and consistently with
poorer cognitive performance compared with concurrent blood or patella Pb levels (Table
5-8).
Effect modification by stress was examined in both the NAS and BMS cohorts. In the
NAS cohort, higher bone Pb levels were associated with poorer cognitive function among
individuals with higher individual-level perceived stress (Peters et al.. 2007). Higher tibia
Pb levels also were associated with larger decrements in cognitive performance among
BMS subjects living in neighborhoods with a greater number psychosocial hazards
(e.g., number of violent crimes, emergency calls, off-site liquor licenses) (Glass et al..
2009) (Table 5-8). These observation of effect modification by environmental stress in
adult humans is supported by several observations of Pb-stress interactions in impaired
learning and memory and adult animals with Pb exposures from gestation through post-
weaning and lifetime Pb exposures (Section 5.3.2.2).
Weuve et al. (2009) studied the association of blood and bone Pb levels with cognitive
function in a subset of 587 healthy women in the Boston, MA area participating in the
Nurses' Health Study. Blood and bone Pb levels were measured between the ages of 47
and 74 years, and the mean (SD) blood Pb level in this group was 2.9 (1.9) (ig/dL
measured in samples collected an average of 5 years before cognitive testing. As in the
aforementioned studies of adults, tibia and patella Pb levels were more consistently
associated with cognitive performance than was blood Pb levels (Table 5-8). 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, and a one SD (10 jxg/g bone)
higher tibia Pb level was associated with 0.051-point lower (95% CI: -0.010, -0.003)
standardized composite score. A similar magnitude of decrease was estimated for an
increase in age of 3 years in these women. The magnitude of association was slightly
smaller for an SD unit increase in patella Pb level (-0.033 [95% CI: -0.080, 0.014).
Several studies analyzed data from the U.S.-representative NHANES III (1991-1994)
population of men and women and investigated effect modification by age and genetic
variants. Only blood Pb levels were available and were measured in samples collected
concurrently with cognitive testing. Krieg and Butler (2009) did not find blood Pb level
consistently to be associated with poorer performance on cognitive testing among 2,090
adults 20-59 years of age or among 1,796 adults 60 years of age and older. Because
different types and numbers of tests were conducted in the two age groups, it is difficult
to compare findings between age groups. In the subset of the population with genetic
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analysis, blood Pb-cognitive function associations were not found to be modified by
ALAD genetic variants in the same direction (Krieg et al.. 2009). Among adults ages
20-59 years with the CC and CG ALAD genotype groups combined (i.e., ALAD2
carriers), reaction time improved (i.e., faster reaction time) by 38 ms per 10-fold increase
in concurrent blood Pb level. In contrast, among ALAD2 carriers, higher blood Pb levels
were associated with poorer performance on a symbol-digit substitution task. The
potential direction of effect modification by the ALAD2 genotype is not immediately
clear as the greater affinity of the ALAD2 enzyme subunit for Pb may increase
susceptibility to Pb-associated health effects by increasing blood Pb levels or it may
diminish Pb-associated health effects by decreasing Pb bioavailability by maintaining it
in a sequestered state in the bloodstream. Krieg et al. (2010) did find differences in the
association between concurrent blood Pb level and scores on a symbol-digit substitution
test by the VDR variants, rs731236 and VDR rs2239185, and by the VDR haplotype.
Similar to observations in adolescent NHANES participants (Section 5.3.2.1), results
were not uniform across the various tests. However, blood Pb level generally was
associated with greater decrements in cognitive performance among adults with the CC
genotypes of VDR variants.
Other studies with smaller numbers of subjects generally produced results consistent with
those from the larger studies above. A cross-sectional 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). It
should be noted, though, that Pb in plasma makes up a very small fraction of all Pb in
blood and is a different, and much less used, biomarker than Pb in whole blood. The
relevance of this Pb fraction is not entirely clear. Pb in plasma is not bound to
erythrocytes, as is about 99% of blood Pb. Thus, it has been postulated that plasma Pb
may be more toxicologically active (Chuang et al.. 2001; Hernandez-Avilaetal.. 1998).
In another cross-sectional study of 47 men and women in Rochester, NY (55-67 years of
age), subjects in the higher two tertiles of calcaneal bone (trabecular bone with higher
turnover rate than tibia) Pb level performed worse on delayed matching-to-sample and
paired associated learning tasks (Van Wiingaarden et al.. 2009) (Table 5-8). In analyses
of tibia Pb levels, subjects in the highest tertile of tibia Pb level did not consistently
perform worse on cognitive tests (Table 5-8). The exact calcaneal and tibia Pb levels in
tertiles were not reported.
Adults with Occupational Lead Exposures
The 2006 Pb AQCD concluded that in adults, blood Pb levels were associated with
cognitive function most consistently among those with occupational Pb exposures. These
findings were supported by results from a few recent studies of occupationally-exposed
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adults. Dorsey et al. (2006) followed up on a cohort of Pb-exposed workers in Korea with
a mean age of 43.4 years, on whom patella Pb measurements were made. This group
represented a typically highly-exposed occupational group with an average blood Pb
level of 30.9 (ig/dL. In this cohort, both blood and tibia Pb levels previously were found
to be associated with poorer performance on a battery of neurocognitive tests (Schwartz
et al.. 2005; Schwartz et al.. 2001). Dorsey et al. (2006) found higher patella Pb levels to
be associated with poorer manual dexterity, poorer sensory function, and greater
depression symptoms. These associations for patella Pb levels were not as strong as the
previously reported associations were with either blood or tibia Pb levels in this
occupational cohort.
A follow-up study of the original 1982 Lead Occupational Study was conducted
2001-2004 with 83 of the original 288 Pb-exposed workers and 51 ofthe original 181
controls (khalil et al.. 2009b). Those originally in the exposed workers group had last
worked in a job with Pb exposure from 0.02 to 16 years (median: 6) prior to follow-up
testing. While the follow-up participation was somewhat low, participants did not differ
from nonparticipants on most baseline cognitive tests except for performing slightly
better on aspects of the grooved pegboard test and recall on a paired associates learning
task. This suggests that the follow-up participation was not biased to poor performers. At
follow-up, the former Pb-exposed workers performed worse than did the controls in total
cognitive score and in the spatial and general intelligence domains (p <0.05). They also
performed worse in all other domains (e.g., motor, executive, and memory) although the
differences were not as large. A similar pattern was observed in analyses using tibia Pb
levels measured at the follow-up visit to represent exposure. Weaker associations were
observed with concurrent blood Pb levels (median among the exposed: 12 (ig/dL).
Among the former Pb workers, higher tibia Pb levels were associated with a greater
decrease in total score and scores for spatial and executive domains between baseline and
follow-up. Tibia Pb level were associated inversely with poorer performance in other
domains as well. As in nonoccupationally-exposed adults, the stronger findings for tibia
Pb levels in former Pb-exposed workers indicate stronger effects of higher past Pb
exposures than lower current exposures on cognitive function.
Additional studies aimed to characterize factors that either mediate or modify the
association between Pb biomarkers and cognitive function. A study of 61 current Pb
smelter workers with a mean age of 40 years and blood Pb level of 29.1 (ig/dL found that
both a working lifetime time-weighted integrated blood Pb level (an index of cumulative
exposure) (p = 0.09) and tibia Pb level (p = 0.08) were associated with longer times to
complete the grooved pegboard test (Bleecker et al.. 2007a). Among 112 Pb smelter
workers, working lifetime time-weighted integrated blood Pb level was associated with
poorer performance on attention and digit symbol tasks among those with low cognitive
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reserve (assessed by performance on the Wide Range Achievement Test-R for reading)
(Bleecker et al. 2007b').
Apolipoprotein E is a transport protein for cholesterol and lipoproteins. The gene appears
to regulate synapse formation (connections between neurons) and may be particularly
critical in early childhood. A genetic variant, called the ApoE-s4 allele is a haplotype
between 2 exonic SNPs and is perhaps the most widely studied genetic variant with
respect to increasing risk of neurologic disease. ApoE-s4 carriers are at two-fold
increased risk of developing Alzheimer's disease, although the majority of such
individuals still do not develop the disease. Thus, it is biologically plausible that ApoE-s4
carriers may be biologically susceptible to cognitive dysfunction. A study of
occupationally-exposed adults found that among individuals with at least one ApoE-s4
allele, blood Pb level was associated with poorer performance on digit symbol, pegboard
assembly, and complex reaction time tests (Stewart et al.. 2002).
Studies also indicated that Pb-exposed workers may be at increased risk of motor
dysfunction. Among Pb smelter workers, working lifetime time-weighted integrated
blood Pb level was associated with poorer motor performance (p <0.05). Iwata et al.
("2005) examined the cross-sectional association between blood Pb level and aspects of
postural sway among 121 Pb-exposed workers in Japan with blood Pb levels between 6
and 89 (ig/dL (mean: 40 (ig/dL). In multiple regression analyses adjusted for age, height,
and smoking and drinking status, 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). The authors calculated a benchmark dose level (Budtz-
Jorgensen et al. 2001; NRC. 2000) of 14.3 (.ig/dL from a linear concentration-response
model of their data. A supralinear concentration-response function was found to fit the
data slightly better than was a linear function.
Summary of Cognitive Function in Adults
In summary, among nonoccupationally-exposed adults, there is weak evidence that
cognitive function is associated with concurrently measured blood Pb levels. The
strongest evidence was provided by NHANES analyses, in which concurrent blood Pb
levels were associated with lower cognitive function in particular age and genetic variant
subgroups (Krieg et al.. 2010; Krieg and Butler. 2009; Krieg et al.. 2009). These analyses
did not have bone Pb measures for comparison. It is important to note that because bone
Pb is a major contributor to blood Pb levels, blood Pb level also can reflect to a large
extent longer term exposures, including higher past exposures, especially among adults
without occupational exposures. Thus, in the NHANES analyses of adults, it is difficult
to characterize the relative contributions of recent and past Pb exposures to the
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associations observed between concurrent blood Pb level and cognitive function.
Consistent with the conclusion of the 2006 Pb AQCD, several recent studies found
associations between bone Pb levels and cognitive function in adults (Table 5-8). Much
of the evidence was provided by analyses of two cohorts: BMS and NAS. Recent
longitudinal analyses demonstrated that higher 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.. 2007a'). These findings suggested that
long-term Pb exposure may contribute to ongoing declines in cognitive function in adults.
Recent studies that analyzed both blood and bone Pb levels generally found stronger
associations for bone Pb levels, in particular tibia Pb levels, across the various cognitive
tests that were performed (Table 5-8). The discrepant findings for blood and bone Pb
levels indicate that cumulative Pb exposure that likely included higher past exposures,
may be a better predictor of cognitive function in adults than is blood Pb level. Patella
and tibia Pb levels were examined in the NAS and Nurses' Health Study; however,
evidence did not consistently indicate that tibia Pb levels were more strongly associated
with decreases in cognitive performance (Weuve et al. 2009; Weisskopf et al.. 2007a).
Additional support for the effects of cumulative or past Pb exposure is provided by
analyses of the Boston prospective cohort as adults. Deciduous tooth Pb was associated
with decrements in specific cognitive indices at ages 19-20 years (Bellinger et al. 1994a).
and blood Pb levels measured at age 6 months, 4 years, 10 years, and levels averaged
over childhood were associated with decrements in FSIQ at ages 28-30 years (Mazumdar
et al.. 2011).
Although based on limited examination, there is some indication that certain variants in
HFE, ALAD, or VDR genes modify the association between Pb and cognitive function in
nonoccupationally-exposed adults; however, results were uniform across the various
cognitive tests performed. Aside from identifying populations potentially at increased
risk and elucidating underlying modes of action, such effect modification also serves to
strengthen the basic inference about associations between Pb biomarkers and cognitive
function. Specifically, when effect modification is identified, potential confounding
factors would have to vary by levels of the modifying factor, which is usually unlikely,
particularly when considering genotype. It also is important to bear in mind that it is not
always clear whether the observed effect modification reflects a change in the
toxicokinetics of Pb and therefore a change in dose at the biological site of action or a
direct biological interaction that increases the toxicity of Pb at a particular target organ or
tissue.
In contrast with nonoccupationally-exposed adults, in adults with current occupational Pb
exposures, cognitive function was associated with both blood and bone Pb levels. These
findings indicate that among adults with occupational Pb exposures, both current and
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cumulative exposures affect cognitive function. In the study of former Pb workers, blood
Pb levels (median: 12 (ig/dL) and findings were more similar to those from
nonoccupationally-exposed adults. Among former Pb workers, tibia Pb levels were
associated more strongly with cognitive performance than were blood Pb levels (khalil et
al.. 2009b). Thus, in the absence of higher current Pb exposures, cumulative Pb exposures
may have a greater effect on cognitive function in adults.
5.3.3 Behavioral Effects
5.3.3.1 Epidemiologic Studies of Behavioral Effects in Children
Noncognitive effects of Pb are more complex to study relative to IQ tests. There are
fewer objective tests of behavior, and existing tests do not have as strong psychometric
properties and are less rigorously standardized compared with IQ tests. In several studies,
behaviors are assessed frequently using teacher and/or parent ratings and thus are subject
to greater measurement error. However, domain-specific neuropsychological assessments
are advantageous as they may provide greater insight into the underlying CNS damage
that may be associated with exposures (e.g., structural, neural system, neurotransmitter)
(White et al. 2009). Several epidemiologic studies reviewed in the 2006 Pb AQCD
reported associations between blood Pb levels and a wide range of behavioral effects,
with the weight of evidence supporting associations with inattention and hyperactivity
and smaller bodies of evidence indicating associations with misconduct and delinquent
behaviors and withdrawn and depressive behaviors (Bellinger and Rappaport. 2002;
Needleman et al.. 2002; Dietrich et al.. 2001; Burns et al.. 1999; Wasserman et al.. 1998;
Needleman et al.. 1996; Bellinger et al.. 1994a). Coherence was provided by similar
findings of inattention, impulsivity, and changes in social behavior in Pb-exposed
animals (Section 5.3.3.2). Epidemiologic studies found that blood Pb levels were
associated with decrements in cognitive function and behavioral problems within the
same population of children, which demonstrates the strong relationship between the two
neurodevelopmental domains. The strong relationship between cognitive function and
behavior in children also is demonstrated by the fact that the schedule controlled behavior
tests in animals measure both memory and inattention. Thus, behavioral problems
associated with Pb exposure may contribute to problems with learning, which may
progress to antisocial and delinquent behavior later in life.
Most previous epidemiologic studies found that blood or dentin Pb levels measured at an
early age (e.g., 2-6 years of age) were associated with behavioral problems later in
childhood and early adulthood (e.g., 7-22 years of age). Most studies examined
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associations with blood Pb levels assessed at a single time point; however, even the
prospective studies with serial measurements of blood Pb levels, found a range of
behavioral problems to be associated with both prenatal and early childhood blood Pb
levels (Dietrich et al.. 2001) and lifetime average blood Pb levels (Burns et al.. 1999).
Thus, evidence did not conclusively identify a single lifestage of Pb exposure that was
associated with the greatest risk of behavioral problems and misconduct. Recent studies
strengthened the evidence for associations of blood Pb levels with inattention and
aggression (Figure 5-14 and Table 5-9). Based on a smaller number of studies, new
evidence demonstrates associations of blood Pb levels with ADHD diagnosis and
diagnostic indices as well as with conduct disorders (Table 5-10 and Table 5-11). As with
cognitive function, the epidemiologic evidence for each category of outcomes was
evaluated separately in order of increasing weight of evidence. Emphasis was placed on
prospective studies with repeated assessments of blood Pb levels and behavior, studies
assessing effects relevant to blood Pb levels in contemporary U.S. children (i.e., less than
5 (ig/dL), and studies of younger children whose blood Pb levels are less influenced by
higher past Pb exposures.
Similar to cognitive function, associations between blood Pb levels and behavioral
outcomes may be potentially confounded by factors such as parental SES, parental
education, parental IQ, quality and stability of the caregiving environment, and
nutritional status. Accordingly, in assessing whether blood Pb-behavior associations were
independent of the effects of the other variables, greater weight was given to studies that
accounted for potential confounding in the study design or in statistical analyses.
Inattention and Hyperactivity in Children
Consistent with previous evidence, recent studies provided strong evidence that blood Pb
level were associated with various endpoints related to inattention, hyperactivity, and
impulsivity in children after adjusting for potential confounding by multiple SES-related
variables and co-exposures (Figure 5-14 and Table 5-9). Evidence was equally consistent
for inattention assessed using teacher and parent ratings and objective tests that measure
sustained attention such as the continuous performance test (CPT). The associations
observed with CPT, in particular, provide strong coherence with findings in animals for
Pb-induced impairments in homologous tests of response inhibition in Schedule
Controlled Behavior Tests (Section 5.3.3.2). Both tests measure reactions to stop signals,
i.e., premature responses, reaction time. Most of the recent evidence is derived studies of
non-U.S. children. Whereas previous studies primarily examined associations with blood
Pb levels measured earlier in childhood, recent studies indicated associations with
concurrent blood Pb levels. Further, many earlier studies of inattention and impulsivity
included children with higher blood Pb levels than those observed in contemporary
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children. Recent studies provided new evidence of blood Pb-associated inattention and
impulsivity in populations of children with mean blood Pb levels ranging from 2 to
5 (.ig/dL (C'ho et al.. 2010; Plusqiiellec et al.. 2010; Chiodo et al.. 2007; Plusqiiellec et al..
2007).
Previous results from prospective studies indicated that early childhood Pb exposures
were associated with attention deficits in adolescence and young adulthood (Ris et al..
2004; Bellinger et al. 1994a). Bellinger et al. (1994a) found that compared with young
adults (ages 19-20 years) who had deciduous tooth Pb levels 2.9-5.9 ppm, adults with
deciduous tooth Pb levels >19.9 ppm committed more errors on the CPT and performed
worse on Stroop and WCST, which measure the ability to shift focus and execute a
different task. In the CLS cohort, Ris et al. (2004) found increases in attention as
measured by CPT in association with prenatal maternal, 3-60 month average, and
78 month blood Pb levels in adolescents ages 15-17 years, adjusting for maternal IQ,
SES, HOME, and adolescent marijuana use. Prenatal blood Pb level was associated with
slightly larger increases in attention, and larger associations were estimates for males
(Figure 5-14 and Table 5-9).
Studies in children in Korea with relatively low blood Pb levels (means < 3 j^ig/dL) both
demonstrated associations with measures of inattention (C'ho et al.. 2010; Min et al..
2007). In a population of children ages 8-11 years, Cho et al. (2010) found relationships
of concurrent blood Pb levels with ADHD symptoms (i.e., inattentiveness, hyperactivity,
and total score) rated by teachers and parents, with the association with teacher ADHD
rating attaining statistical significance. In addition to the low blood Pb levels in this study
(mean: 1.9 j^ig/dL [range: 0.53-6.16]), a strength of this study was the comparison of
effect estimates with and without adjustment for potential confounders such as age, sex,
paternal education, maternal IQ, child IQ, city of residence, birth weight, and urinary
cotinine. In multivariate models, effect estimates decreased by 2 to 14%; however,
associations remained statistically significant. Mean ADHD ratings by teacher and
parents were similar (both 9.1); however, parental ratings had greater variability (SD:
11.5 for parents and 8.6 for teachers), which may have contributed to differences in
association. Although higher blood Pb levels were associated with more errors
(responding to a nontarget) on the CPT test, they were not consistently associated other
indicators of inattention on the CPT or Stroop test (Figure 5-14 and Table 5-9). Further,
effect estimates lost statistical significance when urinary cotinine was included in models.
Cho et al. (2010) did not examine potential confounding by HOME score but they did
examine parental history of neuropsychiatry disease (e.g., ADHD, learning disability,
depression, obsessive-compulsive disorder). Mean blood Pb levels were similar in
children with and without parental history of neuropsychiatric disease (1.80 and
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1.93 (ig/dL, respectively, p = 0.32). Thus, the relationship between blood Pb level and
ADHD-related symptoms in the child was not likely confounded by parental
neuropsychiatric disease. Results from this study also indicated that the relationship
between blood Pb level and ADHD-related symptoms was independent of the
relationship between blood Pb level and IQ. Min et al. (2007) examined a population of
children (ages 7-16 years) in Korea with a mean concurrent blood Pb level of 2.9 (ig/dL
and adjusted for maternal score on attention tests. Investigators found that a 1 j^ig/dL
increase in concurrent blood Pb level was associated with a 16.8 ms increase (95% CI: -
1.08, 0.06) in reaction time, a test of attention. Although this study did not consider
confounding by SES and caregiving environment, the findings was consistent with those
from Cho et al. (2010).
Similar results were obtained in a study of children, ages 8-12 years, in Romania, except
that concurrent blood Pb level was associated with similar magnitudes of increase in
inattention, hyperactivity, and impulsivity as assessed by both parents and teachers
(Nicolescu et al.. 2010). Results for parent and teacher ratings were strengthened by
observations that blood Pb levels also were associated with increased false-alarm rates in
responses to stop signals (Figure 5-14 and Table 5-9). Blood Pb levels in this study also
were relatively low (median: 3.7 [95% CI: 1.7, 11.1]), and removing five children with
blood Pb levels at or above 10 (ig/dL had minimal impact on observed associations.
Exposures to aluminum and mercury, other neurotoxic metals, also were examined, and
blood Pb level was associated with the largest, statistically significant increases in
inattention, impulsivity, and hyperactivity.
The aforementioned studies did not consider confounding by the caregiving environment,
i.e., HOME score. Studies that did adjust for HOME score, also found associations of
blood Pb level with inattention and hyperactivity (Chandramouli et al.. 2009; Chiodo et
al.. 2007). In their longitudinal study of children in the U.K., Chandramouli et al. (2009)
found association of higher blood Pb level at age 30 months with hyperreactivity at ages
7 and 8 years; however, this association was observed primarily in children with blood Pb
levels greater than 10 (ig/dL (Table 5-9) and the strongest for ratings given by teachers.
Children with elevated blood Pb levels did not consistently have elevated odds of
inattention as assessed using stop signal tasks. Similar to Cho et al. (2010). this
association was not influenced by the inclusion of child IQ in the model. Therefore, these
findings add support for increasing blood Pb levels having effects on behavior
independent of effects on cognitive function. Chiodo et al. (2007; 2004) found that
concurrent blood Pb level was associated with inattention in 7 year-old children in
Detroit, MI, as assessed using teacher ratings and CPT (Figure 5-14 and Table 5-9).
Potential confounders were selected based on the association of each with a particular
endpoint, thus model covariates varied among endpoints. HOME score was associated
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with inattention teacher ratings (Chiodo et al. 2007) but not with all CPT indices of
inattention (2007; Chiodo et al.. 2004). Blood Pb level was associated with inattention
alone and adjusted for covariates. These findings indicated that HOME score did not
account for all of the association between blood Pb level and behavioral outcomes. As
was discussed in Section 5.3.2.1, this cohort comprised large proportions of children with
prenatal exposure to alcohol and drugs. While this may limit the generalizability of
findings, it is important to note that alcohol and drug use were not associated with every
inattention index or in each study. Blood Pb level remained associated with inattention in
a model that included alcohol and marijuana use along with other demographic and SES-
related variables.
A series of studies conducted in Inuit children living in northern Quebec, Canada
reported associations between blood Pb levels and measures of inattention after extensive
consideration of potential confounding variables (Plusquellec et al.. 2010; Plusquellec et
al.. 2007; Fraser et al.. 2006). Plusquellec et al. (2007) found higher cord blood Pb level
to be associated with higher ratings by investigators for distractibility in completing
tasks, frenetic movement, and duration of off task behaviors at age 11 months (Figure
5-14 and Table 5-9). In the same cohort of children, concurrent blood Pb level but not
cord blood Pb level was associated with impulsivity, irritability, and duration of off task
behavior at ages 4-6 years (Plusquellec et al.. 2010). Fraser et al. (2006) additionally
indicated that at ages 4-6 years, the relationship between concurrent blood Pb level and
motor function (i.e., transversal sway, reaction time) may be mediated by the effects of
blood Pb level on an inattention/impulsivity index. These studies conducted in Inuit
children evaluated confounding by examining associations of demographic, SES, and co-
exposures with outcomes independently and with blood Pb level in models. HOME score
was not associated with inattention or impulsivity. These findings further indicate that
confounding by HOME score cannot fully account for the collective body of evidence for
associations between blood Pb level and measures of inattention and impulsivity.
Consistent with findings in a group of children in Korea (C'ho et al. 2010) and children in
the U.K. (Chandramouli et al.. 2009). Nigg et al. (2008) found that associations of blood
Pb level with a hyperactivity/impulsivity index in a group of U.S. children, ages
8-17 years, were independent of associations with IQ. Whereas other studies assessed
direct and indirect effects of blood Pb level on behavioral outcomes by including IQ in
the model as a covariate, Nigg et al. (2008) specifically used regression-based path
analysis, a more rigorous method to characterize the impact of one variable on the
association of another in the model after controlling for other previous variables. After
adjusting for sex and income, investigators found that concurrent blood Pb level had a
direct association with hyperactivity/impulsivity that was not completely mediated by the
blood Pb-IQ association. Instead, the association between blood Pb level and IQ was
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found to be mediated via the association with hyperactivity/impulsivity. Based on
associations between hyperactivity/impulsivity slower stop signal reaction time and
response variability, the authors concluded that the association between blood Pb level
and hyperactivity/impulsivity was mediated via poorer response on the stop task. It is
important to note that other potential confounders including parental IQ and HOME score
were not examined
Several other recent studies found associations between concurrent blood Pb levels and
inattention; however, their results should be interpreted with caution due to their limited
consideration of confounding by SES-related variables (Li et al.. 2008d). their
examination of children living near Pb sources who likely have limited applicability to
contemporary children in the U.S. general population (Bao et al.. 2009; kordas et al..
2007). or both (Liu et al.. 201 la).
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Study	Mean(SD)	10tti-90th	Blood Pb timing
blood Pb (pg/dl_)	percentile
Niggetal. (2008)	1.04(0.54)	0.5-1.7	Concurrent
Choetal. (2010)	1.9(0.67)	1.2-2.8	Concurrent
Plusquellecetal. (2007) 4.64(3.54) 1.4-8.3 Prenatal
Chiodoetal. (2007) 5.0(3.0) 2.1-8.7 Concurrent
Chiodoetal. (2004) 5.4(3.3) 2.3-9.5 Concurrent
Fraseretal. (2006) 5.3(4.9)	1.4-10.6 Concurrent
Plusquellecetal.(2010) 5.4(5.0)	1.4-10.8 Concurrent
5.4(5.0)	1.4-10.8
Kordasetal. (2007) 11.5(6.1)	5.4-19.2 Concurrent
Outcome
Hyperactivity/lmpu Isivity
ADHD rating scale, teacher
ADHD rating scale, parent
Comission Errors, CPT
Response Time, CPT
Word Reading Score
Distraction
Off task duration
Hyperactivity
Attention problems
Comission Errors, CPT
Omission Errors
Number of errors, CPT
Comission Errors, CPT
Reaction Time
I m pu Isi vity/activity
I mpu Isivity
Off task duration
Off-task passive behavior
Risetal. (2004)	NR	NR Prenatal (maternal) Inattention
3-60 mo. avg
78 mo.
+*-
~
-_y-
-0.5 -0.3 -0.1 0.1 0.3 0.5
Standardized Effect Estimate per 1 |jg/dL increase in
blood Pb level within the 10th-90th percentile interval
Note: Test scores were standardized to their standard deviation to facilitate comparisons among tests with different scales. Studies
generally are presented in order of increasing mean blood Pb level. Effect estimates are standardized to a 1 |jg/dL increase in blood
Pb level within the 10th-90th percentile interval. Black diamonds, orange triangles, and blue circles represent associations with
concurrent, prenatal (maternal), and earlier childhood blood Pb levels, respectively.
Figure 5-14 Associations of blood Pb levels with behavioral indices of
inattention, hyperactivity, and impulsivity.
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Table 5-9 Additional characteristics and quantitative results for studies
presented in Figure 5-14
Study
Blood Pb Levels
Population/Location (|jg/dL)
Statistical Analysis
Outcome
Effect Estimate
(96% Cl)a
Nigg et al.
(2008)
150 children ages 8-17 Concurrent
V- Mean (SD)
Year of birth and location g.^ „r -| Q4
NR (0.53)
12-17 yr: 1.03
(0.54)
Regression-based path analysis
adjusted for sex and income.
No parental IQ or education
Hyperactivity/impulsivity
using Teacher ADHD Rating
Scale and Kiddie Schedule for
Affective Disorders and
Schizophrenia at ages 8-17 yr.
0.21 (0, 0.42)"
Cho et al. 667 children ages 8-11 Concurrent
(2010)	yr in 2008	Mean (SD): 1.9
Five Korean cities (0.67)
10th-90th: 1.2-2.E
Log linear regression model
adjusted for age, sex, parental
education, maternal IQ, child IQ,
city, birth weight, urinary cotinine
ADHD rating scale, teacher
ADHD rating scale, parent
Korean ADHD Rating Scale IV
commission errors, CPT
Response time, CPT
VNford reading score, Stroop
assessed at ages 8-11 yr.
0.042 (0.017, 0.067)
0.010 (-0.013, 0.033)
0.03 (-0.01,0.07)
-0.01 (-0.05, 0.03)
-0.02 (-0.06, 0.02)c
Plusquellecet
al. (2007)
148-164 children ages
11 mo born 1995-2002
Inuit communities
Quebec, Canada
Prenatal (cord)
Mean (SD): 4.6
(3.5)
10th-90th: 1.4-8.3
Log linear regression model
adjusted for delivery complication,
home organization, low birth weight
(distractibility) and sex, maternal
anxiety, prenatal alcohol exposure
(off task duration)
Several more examined
Distraction
Off task duration
using Bayley Behavioral Rating
Scale at age 11 mo
0.06 (0.021,0.098)
0.03 (-0.01, 0.073)
Chiodo et al.
(2007)
506 African-American
children in Detroit, Ml
area followed from birth
(1989-1991) to age 7 yr.
Large proportions of
children with prenatal
exposure to cocaine or
marijuana
Concurrent
Mean (SD): 5.0
(3.0)
10th-90th: 2.1-8.7
Linear regression model adjusted
for child age, sex
Several more examined
Hyperactivity
Attention problems
using PROBS-14 and Conners1
Teacher Rating Scale-39 at age
7 yr
Comission Errors (%)
Omission Errors (%)
0.13(0.03, 0.23)"
0.13 (0.03, 0.23)"
-0.08"
0.18 (0.07, 0.29)c
Chiodo et al.
246 African-American
children in Detroit, Ml
area
followed from birth (not
reported) to age 7.5 yr.
Large proportions of
children with prenatal
exposure to cocaine or
marijuana
Concurrent
Mean (SD): 5.4
(3.3)
10th-90th: 2.3-9.5
Log linear regression adjusted for
SES, maternal vocabulary score
(Number of errors and errors of
commission); Number of children >
18 yr (Number of errors); Child age
(errors of commission);
Education, sex, prenatal cocaine
exposure (reaction time)
Several more examined
Number of errors
Comission Errors
Reaction time
Using CPT at age 7.5 yr
0.35 (0, 0.69)
0.05 (p > 0.05)d
0.25 (p > 0.05)d
Fraser et al.
90 children ages 5-6 yr
born 1993-1996
Inuit communities
Quebec, Canada
Concurrent Mean
(SD): 5.3 (4.9)
10th-90th:
1.4-10.6
Linear regression model adjusted
for binge drinking in pregnancy
Several more examined
Impulsivity/activity
using modified Infant Behavioral
Rating Scale at ages 5-6 yr
0.25 (0.04, 0.46)
Plusquellec et 95-98 children ages 5-6
al. (2010) yr born 1993-1996
Inuit communities
Quebec, Canada
Concurrent Mean Log linear regression model
(SD): 5.4 (5.0) adjusted for birth weight, sex,
10th-90th'	Par'ty' caregiver education
-I 4_.|q g '	(impulsivity) and birth weight, SES,
child blood hemoglobin (off task
duration)
Several more examined
Impulsivity
Off task duration
using modified Infant Behavioral
Rating Scale at ages 5-6 yr.
0.019 (0.001,0.036)
0.02 (0, 0.039)
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Blood Pb Levels
Study	Population/Location	(|jg/dL)	Statistical Analysis
Outcome
Effect Estimate
(96% Cl)a
Kordas et al.
(2007)
168 children ages 6-8 yr.
born 1993-1995
Torreon, Mexico
Residence near metal
foundry
Concurrent
Mean (SD): 11.5
(6.1)
10th-90th:
5.4-19.2
Linear regression model adjusted
for age, sex, SES, home
ownership, crowding in home,
maternal education, family
structure, forgetting homework
Off task passive behavior
using instrument developed by
investigator at ages 6-8 yr
0.034 (0.005, 0.063)
Ris et al.
195 children, ages 15-17 NR
years, born 1979-1985
Linear regression adjusted for sex,
maternal IQ, HOME, SES,
adolescent marijuana use
Several more examined
Inattention using OPT at ages
15-17 yr
Prenatal (maternal)
3-60 mo avg
78 mo
0.16 (0.04, 0.27)
0.11 (0.04, 0.19)
0.12 (0.02, 0.22)
Studies not included in figure because of categorical analysis of Pb biomarker or outcome:
Chandramouli 488 children ages 7-8 yr Age 30 mo
et al. (2009) born 1991-1992	Mean (SDj. NR
Avon, UK
Linear regression model adjusted
for maternal education, home
ownership, maternal smoking,
HOME score, maternal SES, family
adversity index, parenting attitudes
at 6 mo.
Hyperactivity, teacher
using Strengths and Difficulties
Questionnaire at ages 7-8 yr
Selective inattention
using Test of Everyday Attention
for Children at ages 7-8 yr
OR: 0.84 (0.47, 1.52)e
blood Pb 2-5 vs. 0-2 pg/dL
1.25 (0.67, 2.33)e
blood Pb 5-10 vs. 0-2 pg/dL
2.82 (1.08, 7.35)e
blood Pb > 10 vs. 0-2 pg/dL
OR: 0.97 (0.62, 1.52)e
blood Pb 2-5 vs. 0-2 pg/dL
1.01 (0.64, 1.61 )e
blood Pb 5-10 vs. 0-2 pg/dL
0.88 (0.42,1,85)e
blood Pb > 10 vs. 0-2 pg/dL
Nicolescu et
al. (2010)
83 children ages 8-12 yr
born 1995-1999
Bucharest and
Pantelimon, Romania
Concurrent
Median (IQR): 3.7
(2.6)
10th-90th: 2.0-8.5
Log linear regression model
adjusted for city, sex, age,
computer experience, handedness,
eye problems, number of siblings,
parental education, prenatal
smoking, parental
psychological/psychiatric problem
Inattention Parent rating
Inattention Teacher rating
Premature Response
using German version of Test
battery for attention
performance at ages 8-12 yr.
1.3% (-4.0, 6.9)'
4.3% (-0.02, 10.4)
8.3% (-0.02, 19.0)
Bellinger et al. 79 young adults, ages
(1994a) 19-20 yr, born 1970,
Boston, MA area
Deciduous tooth
(age 5-8 yr)
01: 2.9-5.9 ppm
Q2: 6.0-8.7 ppm
Q3: 8.8-19.8 ppm
Q4:
19.9-51.8 ppm
Regression model adjusted for
parental IQ, sex, SES, current drug
use, current alcohol use, current
illicit drug use, maternal education,
maternal age, birth order
Several more examined
Correct Responses on CPT
Q1
Q2
Q3
Q4
98.0(1.0)®
97.6(1.1)
96.9(1.1)
94.6(1.1)
aEffect estimates are standardized to a 1 pg/dL increase in
deviation of the test score to facilitate comparisons among
'Standard error was estimated from the reported p-value.
blood Pb level within the 10th to 90th percentile interval and standardized to the standard
tests that are scored on different scales.
"Direction of effect estimate was changed to indicate an improvement in performance.
dSufficent data were not provided to calculate 95% CIs.
eOdds in higher quantile of blood Pb level compared to that in lowest quantile of blood Pb level.
'Results represent the change in false alarm rate.
'Results represent the mean (SD) score in each quartile.
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Attention Deficit Hyperactivity Disorder in Children
Studies of ADHD were included in the 2006 Pb AQCD; however, previous studies did
not conclusively demonstrate that blood Pb level was associated with ADHD itself.
Importantly, previously studies examining ADHD were fewer in number and limited by
smaller sample sizes (Gittleman and Eskenazi. 1983; David et al.. 1972V Recent studies
consistently reported associations of blood Pb levels with ADHD in children. Coherence
and biological plausibility for these associations are provided by the consistent body of
epidemiologic and toxicological evidence indicating Pb-associated increases in
inattention, hyperactivity, and impulsivity.
Several recent studies reported associations between concurrent blood Pb level and
ADHD diagnosis or diagnostic indices in children between the ages of 8 and 17 years
(Table 5-10). While none of the studies examined the potential for confounding by
HOME score, they did evaluate confounding by several other demographic and SES-
related variables, as well as parental history of psychopathology, including ADHD (Cho
et al.. 2010; Nicolescu et al.. 2010). Recent studies also provided evidence of association
between blood Pb level and ADHD in populations born in the mid- to late-1990s with
relatively low concurrent blood Pb levels (means: 1.9 and 3.7 j^ig/dL) (C'ho et al. 2010;
Nicolescu et al. 2010). However, the temporal trends in Pb exposure in these non-U.S.
populations may have differed from those in U.S. children of the same age.
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Table 5-10 Associations between blood Pb level and ADHD diagnosis or
diagnostic indices in children
Study
Blood Pb Levels
Population/Location ((jg/dL)a
Statistical Analysis
Outcome
Effect Estimate
(96% Cl)b
Nigg et al.
150 children ages
Concurrent
Three-group analysis of covariance
ADHD Dx
F(2,145) = 6.08, p = 0.04
(2008)
8-17 yr.
Mean (SD)
with sex and income.
by two licensed clinicians


Year of birth and
8-11 yr: 1.04
(0.53)
12-17 yr: 1.03
(0.54)
10th-90th: 0.5-1.7
No parental IQ or education


location NR



Braun et al.
4,704 children ages
4-15 yr born
1984-1998
U.S. NHANES
1999-2002
Concurrent
3rd quintile:
1.1-1.3
Logistic regression model adjusted for
postnatal ETS, prenatal ETS, age, sex,
race, childcare attendance, health
insurance coverage, ferritin levels
ADHD Dx or medication use
at age 4-15 yr
1.4(0.4, 3.4) blood Pb
0.8-1.0 vs. <0.8 pg/dLc
2.1(0.7, 6.8), blood Pb
1.1-1.3 vs. <0.8 pg/dLc
2.7 (0.9, 8.4), blood Pb
1.4-2.0 vs. <0.8 pg/dLc
4.1 (1.2,14.0), blood Pb >
2.0 vs. <0.8 pg/dLc
Froehlich et al.
(2007)
2,588 children, ages
8-15 yr born
1986-1996
U.S. NHANES
2001-2004
Concurrent
2nd quartile:
0.9-1.3
Logistic regression model adjusted for
current household ETS exposure, sex,
age, race/ethnicity, income, preschool
attendance, maternal age, birth weight,
and interaction terms for Pb and
prenatal ETS interaction
ADHD Dx
8.1 (3.8, 18.7), blood Pb
level > 2.0 pg/dL plus
prenatal ETS exposure vs.
blood Pb level <0.8 pg/dL
and no prenatal ETS
exposure"
Cho et al.
(2010)
667 children ages
8-11 yr born
1997-2000
Five Korean cities
Concurrent
Mean (SD): 1.9
(0.67)
10th-90th: 1.2-2.8
Log linear regression model adjusted
for age, sex, parental education,
maternal IQ, child IQ, city, birth weight,
urinary cotinine
Total ADHD rating, teacher
Total ADHD rating, parent
using Korean ADHD Rating
Scale IV at ages 8-11 yr.
0.042 (0.017, 0.067)
0.010 (-0.013, 0.033)
Nicolescu et al. 83 children ages 8-12 Concurrent Log linear regression model adjusted
(2010)	yr born 1995-1999 Median (IQR): 3.7 for city, sex, age, computer experience,
Bucharest and	p.6)	handedness, eye problems, number of
Pantelimon, Romania 10th-90th: 1.8-7.1 sibl'n9s> Parer)tal edu,catioj?> Prenatal
smoking, family psychopathology
ADHD score, parent
ADHD score, teacher
using German version of
Conner's scales at ages 8-12
yr.
OR: 1.04(1.00, 1.10)
OR: 1.06(1.00, 1.12)
Roy et al. 756 children ages 3-7 Concurrent Mean Log linear regression model adjusted
(2009a)	yr tested 2005-2006 (SD): 11.4 (5.3) for age, sex, hemoglobin, average
Chennai, India 10th-90th:	montkh|y intc°|?e> Pal\ental educaton,
5 8-18 3	number of other children, clustering in
school and classroom
ADHD index z-score
using Conners'ADHD/DSM-
IV Scales at ages 3-7 yr.
0.002 (0, 0.033)
Chen et al.
(2007)
780 children in TLC
trial followed ages 2-7
yr
Baltimore, MD;
Cincinnati, OH;
Newark, NJ;
Philadelphia, PA
Children underwent
chelation therapy
Concurrent Mean Regression-based path analysis
(SD): 12.0 (5.2) adjusted for city, race, sex, language,
10th-90th'	parental education, parental
6 5-18 7	employment, single parent, age at
blood Pb measurement, caregiver IQ
ADHD index
Using Conners1 Parent Rating
Scale-Revised at age 7 yr.
0.54 (-1.22, 2.30) Direct
0.90 (0.35,1.45) Indirect
aStudies are presented in order of increasing quantile or population mean blood Pb level.
bExcept where noted, effect estimates represent regression coefficients. All effect estimates are standardized to a 1 |jg/dL increase in blood Pb level within
the 10th-90th percentile interval.
cOdds ratio in higher quantile of blood Pb level, with children in the lowest quantile of blood Pb level serving as the reference group.
In addition to finding associations of concurrent blood Pb level with inattention and
hyperactivity, Cho et al. (2010) found a statistically significant relationship with a total
ADHD index as rated by teachers in population of children ages 8-11 years in Korea. In a
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multivariate model that included age, sex, paternal education, maternal IQ, child IQ, city
of residence, birth weight, and urinary cotinine, the effect estimate decreased by 12%;
however, the association remained statistically significant (Table 5-10). Neuropsychiatric
disease (e.g., ADHD, learning disability, depression, obsessive-compulsive disorder) in
parents was not associated with child blood Pb level, and thus, likely did not confound
the relationship between blood Pb level and ADHD index in the child. Results from this
study also indicated that the relationship between blood Pb level and ADHD index was
independent of the relationship between blood Pb level and IQ.
Similar to Cho et al. (2010). Nicolescu et al. (2010) found higher concurrent blood Pb
levels to be associated with greater inattention, impulsivity, and total ADHD score among
children in Romania, ages 8-12 years. As with individual symptoms, only concurrent
blood Pb level, and not aluminum or mercury, was associated with a higher total ADHD
rating (4% increase [95% CI: 0, 10] per 1 (ig/dL increase in concurrent blood Pb level
within the 10th-90th percentile interval [1.8-7.1 |_ig/dL|). The association did not change
substantially in an analysis that excluded the 5 children with blood Pb levels at or above
10 (ig/dL. Adjustment for potential confounding by city, sex, age, computer experience,
handedness, eye problems, number of siblings, parental education, prenatal smoking
exposure, prenatal alcohol exposure, and parental history of psychological or psychiatric
problems resulted in a decrease in statistical significance for blood Pb level associations,
although investigators did not report whether the magnitude of association changed. It is
important to note that model covariates were selected a priori but not all were associated
with both blood Pb level and behavioral endpoint. Only sex, number of siblings, and
maternal education were reported to be significantly associated with both blood Pb level
and total ADHD rating. Parental history of psychological or psychiatric problems was
significantly correlated with parental but not teacher rating of ADHD total score;
however, ORs were fairly similar for ADHD score rated by teachers and parents. Despite
the potential for over-adjustment in this study, higher blood Pb level was associated with
a higher ADHD total score rating.
A recent analysis of data from NHANES 1999-2002 found a relationship between higher
blood Pb level and greater odds of ADHD (parent-report of a diagnosis of ADHD or use
of stimulant medication) among children ages 4-15 years (Braun et al.. 2006). Strengths
of this study included the large sample representative of U.S. children (n = 4704) and the
low concurrent blood Pb levels at which associations were observed. These authors
reported a monotonic increase in ORs from the lowest to highest quintile of blood Pb
level, adjusting for age, race, prenatal smoking exposure, smoker in the home,
preschool/child care attendance, health insurance coverage, and ferritin levels (Figure
5-15). With children in the lowest quintile serving as the reference group (<0.8 (ig/dL),
children in the fifth quintile of concurrent blood Pb level (>2.0 (ig/dL, maximum not
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reported) had the highest prevalence of ADHD (OR: 4.1 [95% CI: 1.2, 14]). A similar
OR was estimated when children with blood Pb levels >5.0 j^ig/dL were excluded from
the highest group. Children in the other three higher quintiles of blood Pb level (range:
0.8-2.0 (ig/dL) also had a higher prevalence of ADHD relative to the reference group
(Table 5-10).
o
<
0.8-1.0
1.1-1.3
1.4-2.0
Quintiles of blood lead concentration (jxg/dL)
Source: Braun et al. (2006)
Note: Adjusted for child age, gender, race/ethnicity, preschool attendance, serum ferritin, prenatal tobacco smoke exposure, smoker
in the household, and insurance status (p for trend = 0.012). Data from ages 4-15 years from NHANES 1999-2002 by quintile of
concurrent blood Pb level.
Figure 5-15 Adjusted odds ratios for Attention Deficit Hyperactivity Disorder
(ADHD) among U.S. children.
In the NHANES 2001-2004 dataset restricted to children ages 8-15 years, Froehlich and
colleagues (2009) found an interaction between prenatal tobacco smoke exposure
(maternal report) and concurrent blood Pb levels. Investigators found ADHD to be
independently associated with prenatal tobacco smoke exposure (OR: 2.4 [95% CI: 1.5,
3.7]) and concurrent blood Pb levels (OR: 2.3 [95% CI: 1.5, 3.8]) in children with blood
Pb levels >1.3 (ig/dL compared with children with blood Pb levels < 0.8 (ig/dL. These
results were adjusted for current household smoke exposure, sex, age, race/ethnicity,
income, preschool attendance, maternal age, and birth weight. As in the younger
NHANES dataset, a similar OR was estimated when children with blood Pb levels >
5.0 (ig/dL were excluded from the highest tertile. The strongest association was observed
in children with both high blood Pb level and prenatal tobacco smoke exposure.
Compared to children in the lowest tertile of blood Pb levels with no exposure to prenatal
tobacco smoke, children in the highest tertile of blood Pb level with exposure to prenatal
tobacco smoke had the highest prevalence of ADHD (OR: 8.1 [95% CI: 3.5, 18.7]).
Although ADHD was associated with low concurrent blood Pb levels (1.3-5 (ig/dL), it is
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important to acknowledge that the adolescents in both NHANES analyses were born in
the late 1980s and may have had higher Pb exposures earlier in childhood that
contributed to the observed associations.
Consistent with findings in the NHANES population, Nigg et al. (2008) found an
association between concurrent blood Pb level and ADHD diagnosis among children
8-17 years. Concurrent blood Pb levels in this population also were low, ranging between
0.40 and 3.47 (ig/dL. However, the examination of adolescents adds uncertainty
regarding the relative contributions of higher past Pb exposures and current exposures to
the observed associations. A key strength of this study over other studies was the
verification of parental reports of ADHD by licensed clinicians using the same diagnostic
criteria. Although the results were adjusted for household income, other potential
confounders including parental IQ and HOME score were not examined.
Roy and colleagues (2009a') examined associations between concurrent blood Pb level
and a range of behavioral problems in 756 children, ages 3-7 years, in Chennai, India. In
this population, the mean blood Pb level was higher than that in most other studies (11.4
[SD: 5.3] (ig/dL); however, this study demonstrated associations of concurrent blood Pb
level with inattention, hyperactivity, and a total ADHD index (assessed by teachers using
Conners' ADHD/Diagnostic and Statistical Manual for Mental Disorders) in younger
children compared with other studies. In generalized estimating equations, higher blood
Pb level was associated with a higher ADHD index score adjusting for age, sex,
hemoglobin, average monthly income, parental education, number of other children, and
clustering at school and classroom levels (Table 5-10). In analyses that did not adjust for
potential confounding variables, mean ADHD scores were similar in the first three
quartiles of blood Pb level and elevated in children with blood Pb levels > 18.71 (ig/dL,
suggesting that in this population, associations may have been driven by children with the
highest blood Pb levels.
Whereas several studies indicated the direct effects of increasing blood Pb level on
inattention and hyperactivity indices, Chen et al. (2007) found stronger indirect effects of
blood Pb level on an ADHD index among children at age 5 years (Table 5-10). Important
limitations of this study include the high blood Pb levels of children at ages 12 to
33 months (20-44 (ig/dL) that made them eligible for a randomized controlled trial of
chelation. Because this study only enrolled children with Pb poisoning, it is difficult to
extend findings to children with lower blood Pb levels. Blood Pb levels of study subjects
remained high at age 5 years (mean: 12.0 (ig/dL) and out the range of the current U.S.
general population of children.
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
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confounding by parental history of ADHD. Given the highly heritable nature of ADHD,
parental history of ADHD is a strong risk factor for ADHD in children (Faraone and
Dovle. 2001). Therefore, the risk associated with parental history of ADHD may mask
the smaller magnitudes of risk associated with environmental exposures, including Pb.
However, little information is available on an association between ADHD in the parent
and blood Pb level in the child, thus, it is uncertain whether parental ADHD confounds
and fully accounts for the associations observed between blood Pb level and ADHD in
children. It should be noted that for parental ADHD to be a confounder, parental Pb
levels would have to be highly associated with ADHD in the parent and with blood Pb
level in the child. Among children in Korea, Cho et al. (2010) found that parental history
of psychopathology (ADHD or several other conditions) was not associated with child
blood Pb level. Among children in Romania, parental history of psychopathology was
associated with a higher ADHD rating in the child as assessed by parents but not teachers
(Nicolescu et al.. 2010). Higher blood Pb level was associated with higher ADHD ratings
in children in models with blood Pb level alone and in models that adjusted for parental
psychopathology plus other covariates. Further, associations of blood Pb level with
ADHD rating assessed by parents and teachers were similar in magnitude (Table 5-10).
While available data are limited, they do not provide strong evidence that parental
psychopathology fully accounts for the associations observed between blood Pb level and
ADHD in children.
Social Misconduct and Delinquent Behavior
The 2006 Pb AQCD described several studies in which higher levels of blood and bone
Pb were associated with higher frequency or risk of misconduct in children or delinquent
behavior in adolescents (U.S. EPA. 2006b). Previous studies primarily indicated
associations with measures of cumulative Pb exposure, including early childhood average
blood Pb level, lifetime average blood Pb level, or bone Pb level. Recent studies added to
the collective body of evidence by demonstrating associations for concurrent blood Pb
level and for a variety of conduct problems including aggression (Chiodo et al.. 2007).
conduct disorder (Braun et al.. 2008). oppositional defiant disorder fNigg et al.. 2008)
and more serious behaviors such as criminal arrests (Wright et al.. 2008) (Table 5-11).
Associations were found with early childhood and concurrent blood Pb levels and were
demonstrated in populations with lower blood Pb levels (means: 1-8 (ig/dL) than those
examined in the 2006 Pb AQCD. Nonetheless, the weight of cumulative evidence
supports associations in populations with mean blood Pb levels between 5 and 11 j^ig/dL.
However, because some studies examined outcomes in adolescents and young adults born
in the 1970s and 1908s, there is greater uncertainty regarding the level, timing,
frequency, and duration of Pb exposure that contributed to the observed associations.
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Several other studies examined children and adolescents born in the 1990s, who likely
did not have as high historical Pb exposures earlier in childhood (Chandramouli et al..
2009; Braun et al.. 2008; Chiodo et al.. 2007).
Table 5-11 Associations between blood Pb level and misconduct and delinquent
behavior in children and young adults9
Study Population/Location
Blood Pb
Levels
(Hg/dL)
Statistical Analysis
Outcome
Effect Estimate
(96% Cl)b
Blood Pb level as a continuous variable
Chiodo et al.
(2007)
506 African-American
children in Detroit, Ml
area followed from birth
(1989-1991) to age 7 yr.
Large proportions of
children with prenatal
exposure to cocaine or
marijuana
Concurrent
Mean (SD):
5.0 (3.0)
10th-90th:
2.1-8.7
Linear regression model adjusted for
HOME, child sex, current marijuana
use (both outcomes)
Caretaker education, alcohol use
during pregnancy, (delinquent behavior
Age, caretaker psychopathology,
maternal IQ (inappropriate behavior)
Several more examined
Delinquent behavior
Inappropriate behavior
using Achenbach Teacher
Report Form at age 7 yr
0.09(0, 0.18)c
0.09 (0, 0.18)c
Dietrich et al. 195 children followed 0-6 yr avg: NR
(2001)	from birth (1979-1985) to
age 15-17 yr
Cincinnati, OH
Linear regression model adjusted for
HOME score, parental IQ, current SES
Delinquent behavior
using the Self-Report of
Delinquent Behavior at ages
15-17 yr.
1.21 (1.08, 1.37)
Wright et al.
250 adults followed from
birth (1979-1985) to age
19-24 yr
Cincinnati, OH
Age 6 yr
Mean (SD):
8.3 (4.8)
10th-90th:
3.9-14.5
Negative binomial regression models
adjusted for maternal IQ, sex, SES,
maternal education
Several more examined
Criminal arrests
Violent arrests
RRs: 1.05(1.00, 1.09)
prenatal
1.05(1.01,1
assessed from county records
at ages 19-24 yr
01 (0.97,1
06 (0.97, 1
08 (1.03, 1
05(1.01, 1
09)	6 yr
05) 0-6 yr avg
15) prenatal
14) 6 yr
10)	0-6 yravg
Chen et al. 780 children participating Concurrent
12007)	in TLC trial followed Mean (SD):
between ages 2-7 yr 12.0 (5.2)
Baltimore, MD;	10th-90th:
Cincinnati, OH; Newark, 6.5-18.7
NJ; Philadelphia, PA
Children underwent
chelation therapy
Regression-based path analysis
adjusted for city, race, sex, language,
parental education, parental
employment, single parent, age at
blood Pb measurement, caregiver IQ
Externalizing behavior, parent
Externalizing behavior, teacher
using Behavior Assessment
System for Children at age 7
yr.
1.024 (0.996,
1.008(1.002,
1.036(1.003,
1.004 (0.998,
1.053) Direct
1.014) Indirect
1.069) Direct
1.010) Indirect
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Study
Population/Location
Blood Pb
Levels
(Hg/dL)
Statistical Analysis
Outcome
Effect Estimate
(96% Cl)b
Blood Pb level as a categorical variable
Braun et al.
(2008)
Children ages 8-15 yr
U.S. NHANES
2001-2004
2nd quartile:
0.8-1.0
Logistic regression with sample
weights applied to produce national
estimates, adjusted for oversampling
of minorities and young children and
adjusted for age, poverty income ratio,
maternal age, sex, race, prenatal ETS,
cotinine, blood Pb levels
Conduct disorder
Using Diagnostic Interview
Schedule for Children-
Caregiver Module at age 8-15
yr
OR: 7.24 (1.06, 49.47), blood
Pb level 0.8-1.0 pg/dL vs.
blood Pb <0.8 pg/dLd
12.37 (2.37,64.56), blood Pb
level 1.1-1.4 pg/dL vs. blood
Pb <0.8 pg AT
8.64(1.87, 40.04), blood Pb
level 1.5-10 pg/dLvs. blood
Pb <0.8 pg/dC
Chandramouli
et al. (2009)
488 children ages 7-8 yr
born 1991-1992
Avon, UK
Age 30 mo
Mean (SD):
NR
Linear regression model adjusted for
maternal education, home ownership,
maternal smoking, HOME score,
maternal SES, family adversity index,
parenting attitudes at 6 mo.
Antisocial activities
using Antisocial Behavior
Interview at ages 7-8 yr
OR: 0.93 (0.47, 1.83)d
blood Pb 2-5 vs. 0-2 pg/dL
1.44 (0.73, 2.84)d
blood Pb 5-10 vs. 0-2 pg/dL
2.90 (1.05, 8.03)d
blood Pb > 10 vs. 0-2 pg/dL
aResults with continuous blood Pb are presented first followed by results with categorical blood Pb. Within each group, studies generally are presented in
order of increasing mean blood Pb level.
bEffect estimates are standardized to a 1 |jg/dL increase in blood Pb level.
c95% CI was estimated from a reported p-value of 0.05.
dOdds in higher quantile of blood Pb level compared to that in lowest quantile of blood Pb level.
The consistency of association of Pb biomarker levels with social misconduct and
delinquent behavior was corroborated in a recent meta-analysis (Marcus et al.. 2010) that
included 19 studies (those reviewed in the 2006 Pb AQCD plus several recent studies)
with a total of 8,561 children and adolescents (mean ages ranging from 3.5 years to
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 associations to considerations of heterogeneity in study design,
definition and assessment method of conduct problems, potential confounding variables
examined, and blood Pb levels. 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 or blood Pb levels; however, similar effect sizes were
estimated for blood and bone Pb levels. Although the authors suggested that hair Pb may
be a better indicator of cumulative Pb exposure compared to bone Pb levels, due to the
high turnover of bone in throughout childhood and into adolescence, an empirical basis
for interpreting hair Pb measurements in terms of body burden or exposure has not been
firmly established (Section 4.3.4.2).
In the meta-analysis, effect sizes did not differ significantly between longitudinal and
cross-sectional studies, among studies that examined different conduct problems
(i.e., opposition defiance, delinquency, externalizing problems), or among studies that
assessed conduct disorders using self-report, teachers report, or criminal records.
Controlling for covariates such as SES, birth weight, parental IQ, and home environment
did not attenuate the relationship between blood Pb level and conduct problems. In
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addition to strengthening the evidence for the independent associations of Pb biomarker
levels with conduct disorders, the results indicate that the lack of adjustment for any
particular covariate, including HOME score, does not warrant limiting inferences from a
particular study.
Studies of children ages 7-8 years found associations between blood Pb level and
misconduct adjusting for HOME score and several other potential confounding variables.
They examined different life stages of blood Pb level and did not clearly indicate
associations between blood Pb level and misconduct with blood Pb levels less than
5 (ig/dL. In their longitudinal study of children in the U.K. who were born in the 1990s,
Chandramouli et al. (2009) found associations of age 30 month blood Pb level not only
with hyperreactivity but also with antisocial behavior at age 7-8 years. Results were
adjusted for maternal education, home ownership, maternal smoking, HOME score,
maternal SES, family adversity index, and parenting attitudes at 6 months. In analyses of
blood Pb level as a categorical variable, greater antisocial activity was most clearly
indicated in children with blood Pb levels greater than 10 (ig/dL (Table 5-11). Chiodo et
al. (2007) examined children, ages 7.5 years in Detroit, Michigan with a mean (SD)
concurrent blood Pb level of 5.0 (3.0) (ig/dL. Higher blood Pb was associated with higher
teacher ratings of delinquent behaviors and inappropriate behaviors, adjusting for HOME
score and other variables. The association with inappropriate behavior additionally
adjusted for caregiver psychopathology. As with cognitive function and inattention
outcomes, prenatal alcohol and drug exposure were not found to influence associations
for blood Pb level. Blood Pb level was associated with behavioral problems in unadjusted
and adjusted analyses, corroborating the independent associations for blood Pb level.
In the U.S. NHANES 2001-2004 dataset of children ages 8-15 years (born in the 1990s),
Braun et al. (2008) analyzed blood Pb as a categorical variable and found higher
prevalence of conduct disorder with concurrent blood Pb levels in the range of 0.8 to 1.0
(ig/dL, the lowest level among all studies examined. Compared with children with blood
Pb levels less than 0.8 (ig/dL, the OR (95% CI) in children with blood Pb levels 0.8-1.0
(ig/dL was 7.24 (1.06, 49.47). Odds ratios also were elevated in children with blood Pb
levels higher than 1.0 (ig/dL (Table 5-11). The wide 95% CIs likely were due to the small
numbers of cases of conduct disorder, as assessed using DSM-IV criteria. For example,
there were 22 cases of conduct disorder among children with blood Pb levels 0.8-1.0
(ig/dL. Poisson regression models showed that children with blood Pb levels 0.8-1.0
(ig/dL had 1.55 (95% CI: 1.09, 2.22) times as many conduct disorder symptoms as did
children with blood Pb levels < 0.8 (ig/dL. Investigators had data available on a limited
set of potential confounders, and poverty income ratio was used to control for potential
confounding by SES.
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While other recent studies supported associations between biomarkers of Pb exposure
and misconduct and delinquent behavior, they examined Pb levels in blood or tooth
samples collected in the 1980s when Pb exposures were much higher (Fergusson et al..
2008; Wright et al.. 2008). Although this evidence may be less informative as to the
relationship between current blood Pb levels and misconduct and delinquent behavior in
adolescents and young adults, it is important to note the consistency of findings among
populations with widely varying blood Pb levels and historical Pb exposures.
In the CLS cohort, prenatal and postnatal blood Pb levels previously were reported to be
associated with self- and parent-reported delinquent and social acts at ages 16-17 years
(Dietrich et al.. 2001). Wright et al. (2008) extended these findings to include
associations of prenatal and postnatal blood Pb levels with criminal arrests (ascertained
from county records) at ages 19-24 years. Mean blood Pb levels were 8.3 (ig/dL (range
1-26) for the prenatal period (maternal blood), 13.4 (ig/dL (range 4-37) for the average
between birth and age 6 years, and 8.3 j^ig/dL (range 2-33) at age 6 years. In models that
adjusted for maternal IQ, sex, SES score, and maternal education, the relative risks (RRs)
for total arrests per 1 (ig/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) for average childhood blood Pb
level, and 1.05 (95% CI: 1.01, 1.09) for blood Pb level at age 6 years. Blood Pb levels
measured during these lifestages also were associated with increased risk of violent
criminal arrests (Table 5-11). Although interactions terms for blood Pb by sex were not
statistically significant, the attributable risk for males was considerably higher for males
(0.85 arrests/year [95% CI: 0.48, 1.47]) than for females (0.18 [95% CI: 0.09, 0.33]).
Results from the two CLS studies suggest that in addition to the prenatal blood Pb levels,
early childhood blood Pb levels may also predict criminal behavior in adulthood.
However, it is important to note that in these CLS studies, concurrent blood Pb levels
were not analyzed. Therefore, these studies do not provide information on the potential
effects of more recent Pb exposures or differences in association between earlier and
more recent blood Pb levels.
A strength of Wright et al. (2008) was the detailed examination of potential confounding
by a large number of variables. All of the examined covariates were weakly correlated
with blood Pb levels (r = 0.24-0.35), thereby minimizing the potential for confounding.
Nonetheless, variables such as maternal IQ, maternal education, and SES were included
in the model because they were associated with arrests in the full multivariate model or
changed the blood Pb level by more than 10%. HOME score was similar between
subjects with and without criminal arrest records and did meet the criteria for inclusion in
the final model.
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In the New Zealand cohort of children born in 1977, dentin Pb levels measured between
ages 6 and 8 years were associated with self-reported and documented violent or property
convictions at ages 14-21 years (Fergusson et al.. 2008). Similar to those from the meta-
analysis, these findings pointed to an effect of cumulative childhood Pb exposure.
Analyses were adjusted for SES, parental criminal record, parental education, and
parental alcoholism, and although effects estimates decreased in adjusted models,
associations remained statistically significant.
Several studies of misconduct and delinquent behavior aimed to characterize whether
associations with biomarkers of Pb exposure were independent of effects on IQ and
educational attainment. Most studies found that associations of Pb biomarkers with
misconduct and delinquent behavior remained statistically significant in a model that
additionally adjusted for child IQ or educational attainment, indicating that Pb exposure
may have a direct effect on misconduct independent of its effect on IQ (Chandramouli et
al.. 2009; Fergusson et al.. 2008; Burns et al.. 1999). It is important to note that because a
decrement in IQ may on the causal pathway to behavioral problems, including both IQ
and behavioral problems in the same model may result in an underestimate of the effect
on behavior. Chen et al. (2007) used path analysis to characterize the direct effects and
indirect effects (mediated through child IQ) of blood Pb level on externalizing problems
(i.e., outbursts of behavior); however, results were inconclusive. A direct effect was
estimated for externalizing problems assessed by teachers and an indirect effect was
estimated for problems assessed by parents. These findings may have limited
applicability to the general population given that the children in the study population had
been referred for chelation therapy at enrollment because of high blood levels, and it is
uncertain whether the observed associations were due to the residual effect of high blood
Pb levels (20-44 (ig/dL) four years earlier.
5.3.3.2 Toxicological Studies of Behavior
Neurobehavioral Changes
The effects of Pb on neurobehavioral changes in animals are well characterized with
various targeted sites including the prefrontal cerebral cortex, cerebellum, and
hippocampus; affected functions include cognition, execution of motor skills, and
memory/behavior. As discussed in earlier Pb AQCDs, young animals are especially
susceptible to the effects of Pb due to the ongoing development of the nervous system
with greater Pb absorption and retention. Pb exposure has been documented to induce
neurobehavioral changes in exposed animals including effects on learning, social
behavior, memory, attention, motor function, locomotor ability and vocalization. At the
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cellular level, Pb impairs axon and dendritic development and contributes to
neurochemical changes in proteins, membranes, redox/antioxidant balance, and
neurotransmission through a multitude of mechanisms, many of which involve the
capability of Pb to mimic calcium. Very early toxicological research on neurobehavioral
endpoints failed to capture the disposition of Pb and its resulting body burden or blood Pb
level and was thus difficult to use in risk assessment.
The 1986 Pb AQCD reported findings of Pb-induced aberrant operant conditioning tasks
in rodents and non-human primates (some with resulting blood Pb levels: 11 to 15 (ig/dL)
as well as hyperactive or inappropriate Pb-mediated responses (U.S. EPA. 1986b). It was
indicated that these effects were possibly of hippocampal origin and showed a curvilinear
response, decreasing at higher Pb doses possibly due to impairment of motor function
(Maetal. 1999; Crofton et al.. 1980). Pb exposure in laboratory animals resulted in
distractibility, reduced to adapt to changes in behavior, impaired ability to inhibit
inappropriate responses, and perseveration (U.S. EPA. 2006b'). Pb has been shown to
impair learning in Fixed Interval tasks (FI) as indicated by premature responses in the
absence of a fixed schedule of reinforcement or reward (Section 5.3.2.2). These findings
in animals are consistent with observations in children that blood Pb levels are associated
with poorer impulse control as rated by parents or teachers or assessed using the
continuous performance test (Figure 5-14 and Table 5-9). Interresponse rates and overall
run rate are the two subcomponents of FI response rate. The 2006 Pb AQCD reported
consistent findings for Pb exposure (resulting in blood Pb levels: 58 to 94 (ig/dL)
affecting FI response rates, by means of decreased interresponse times. Some studies
indicated decreased interresponse times in animals with blood Pb levels of 11 j^ig/dL
(U.S. EPA. 2006b). Discrimination reversal has been shown to be especially sensitive to
Pb exposure. Spatial and non-spatial discrimination reversal or reversal of a previously
learned habit was significantly affected after developmental Pb exposure and was
exacerbated with distracting stimuli. Repeat-acquisition testing revealed that these
deficits were likely not due to sensory or motor impairment at this dose. Together, the
data from the 2006 Pb AQCD showed that social behavior and learning in rodents and
nonhuman primates is significantly affected by Pb exposure that results in blood Pb
levels 15-40 (ig/dL.
In the new literature, gestationally and early postnatally (gestation to PND10, G+P) Pb-
exposed male mice (low and high dose Pb: 10 and 42 (ig/dL blood Pb level at PND10,
respectively) were significantly less active than were control mice, and low dose mice
were significantly less active than were high dose mice, demonstrating a nonlinear
concentration-response relationship (Lcasure et al.. 2008). A similar nonlinear
concentration-response relationship was observed for changes in corticosterone in male
mice exposed post-weaning to Pb (Virgolini et al.. 2005). Activity level of G+P Pb-
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exposed female mice versus controls was unaffected (Lcasure et al.. 2008).
Amphetamine-induced motor activity was monitored in male and female G+P Pb-
exposed mice at 1 year of age. Amphetamine-induced activity of male low and high dose
G+P Pb-exposed offspring was significantly elevated over that in controls; G+P Pb-
exposed females had no change in sensitivity to amphetamine-induced motor activity
(Lcasure et al. 2008).
Herring gull chicks injected with a single i.p. bolus dose of Pb (100 mg/kg Pb-acetate, a
dose created to be similar to that which wild herring gulls are exposed in the wild) on
PND2 were found to have neurobehavioral deficits and learning deficits. Pb-exposed
chicks displayed multiple deficits related to impaired survival skills including decreased
time spent begging the parent for food, decreased accuracy at pecking for food in the
mouth of the parent bird, decreased time spent in the shade (behavioral
thermoregulation), decreased learning in food location, decreased recognition of familiar
individuals (caretaker or sibling), and slower development of motor skills (treadmill test)
versus control birds (Burger and Gochfeld. 2005). The impaired thermoregulation with
Pb exposure agrees with earlier work in Pb-exposed rat pups that also showed impaired
thermoregulatory behavior, i.e., impaired ultrasonic vocalization (Davis. 1982). These
studies in herring gull chicks demonstrate that a single dose of Pb early in life can induce
neurobehavioral deficits that affect survival skills.
Rhesus monkeys exposed to Pb in daily milk formula from PND8 to 1 or 2 years of age
(Pb-acetate/50% glucose in 4 cc of commercial milk formula producing blood Pb of
35-40 (ig/dL) were assessed for tactile defensiveness using the Sensory Processing Scale
for Monkeys, an adaptation of laboratory observational measures of sensory processing
for children (Baranek and Berkson. 1994). Tactile defensiveness in children is defined as
"feelings of discomfort and a desire to escape the situation when certain types of tactile
stimuli are experienced" (Avres. 1964) and is associated with emotional dysregulation,
inattention, and difficult social relations. Other reports have shown that tactile
defensiveness in monkey offspring is affected by prenatal stress alone without Pb
exposure (Schneider et al.. 2008). a factor that has been shown to affect other Pb-related
outcomes.
Attention
Epidemiologic studies consistently have reported associations between higher blood Pb
level and deficits in attention (inattention, distractibility, impulsivity, or ADHD) in
children (Section 5.3.3.1). Animal toxicological studies in the 2006 Pb AQCD detailed
attention deficits in animals undergoing various tests including FR/waiting-for-reward
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testing and tests that employed signal detection with distraction, a test recording
omissions after exposure to an external distraction.
With FR testing, rats are trained on a FR/waiting-for-reward behavioral baseline, learning
to produce food delivery by pressing a lever 50 times. Rodents can earn additional food
by withholding lever presses by waiting; free food is given at increasing time intervals
after completion of the FR. Brockel and Cory-Slechta (1998) exposed male Long-Evans
rats to 0, 50, or 150 ppm Pb-acetate in water from weaning, which produced respective
blood Pb levels of <5, 11, and 29 (ig/dL after 3 months of exposure. After 40 days of
exposure, rats were trained on a FR schedule, and the higher dose animals had
significantly more frequent response rates and resets of the waiting period than did the
low dose group and controls. In the waiting behavior component, wait time was
significantly lower in both treated groups compared to controls. The higher dose group
animals also had an increased number of reinforcers and a higher response to
reinforcement ratio than did low dose and controls. Mechanistic understanding of the
aforementioned FR deficits was studied using the same FR schedule with similar
postweaning dosing of 0, 50, and 150 ppm Pb that yielded respective blood Pb levels of
<5, 10, and 26 (ig/dL after 3 and 7 months of exposure. Administration of dopamine
receptor antagonists attenuated Pb-induced effects on FR schedule testing, suggested a
role for D2 receptors in Pb-induced behavioral impairments.
Another study used a similar postweaning Pb exposure as that in earlier studies (Cory-
Slechta etal.. 1998) to yield blood Pb levels of <5, 16, and 28 (ig/dL and evaluate
sustained attention using testing of signal detection during a distraction (Brockel and
Corv-Slechta. 1999). Rats earned food rewards by discriminating correctly between a
target and distracter light. A 13 second time-out was given for incorrect responses. Pb
exposure produced no deficits in attention with this testing. Further work since the 2006
Pb AQCD affirmed these findings. Testing for signal detection with distraction showed
no effect after postnatal Pb exposure in female rats (Stanale et al.. 2007). The two dose
groups (20 ppm or 300 ppm Pb-acetate in drinking water with lactational, and drinking
water exposure, PND1-PND30) yielded blood Pb levels of 13 j^ig/dL and 31 j^ig/dL
(Stangle et al.. 2007). Another recent study reported that impaired auditory threshold task
related behavioral testing was likely due to inattention in Pb-exposed animals. The
inability of some of the monkeys to engage or focus on the task at hand yielded fewer
available measurements in Pb-exposed animals versus control animals (Laughlin et al..
2009). These rhesus monkeys were exposed to Pb-acetate gestationally (dam drinking
water, 3 months prior to mating) through age 5.5 months (weaning) and had resulting
bone Pb levels at 11 years of 7 and 13 (ig/dL for prenatal and postnatal groups,
respectively, and 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.
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Thus, the inattention literature demonstrates that impulsivity, waiting-for-reward
behavior, and sustained attention may all be affected by Pb with the first two possibly
affected to a greater extent than is sustained attention.
5.3.3.3 Epidemiologic Studies of Mood in Children
A majority of investigation of the effects of Pb on behavior in children has focused on
externalizing behaviors such as inattention, hyperactivity, aggression, and delinquency.
However, studies also have linked biomarkers of Pb exposure with internalizing
behaviors related to emotional dysfunction, including withdrawn behavior, depression,
fearfulness, and anxiety. These behaviors were assessed frequently in school-aged
children using the Child Behavior Checklist, and as with inattention, hyperactivity, and
misconduct, often were rated by parents and/or teachers.
Several earlier results were provided by the prospective cohort studies in Boston,
Cincinnati, Port Pirie, Australia, and Yugoslavia. Collectively, these studies found higher
ratings for internalizing behaviors in association with concurrent blood Pb level as well
as biomarkers of cumulative Pb exposure such as multiyear average blood, tooth, and
bone Pb levels (Burns et al.. 1999; Wasserman et al. 1998; Bellinger et al. 1994b;
Dietrich et al.. 1987b). While several studies demonstrated associations with biomarkers
of long-term, cumulative exposures, results from the Cincinnati cohort provided support
for the effects of shorter-duration exposures. Among CLS infants, blood Pb level at age
3 months and average blood Pb level at age 6 months was associated with less positive
mood in white infants at age 6 months (Dietrich et al.. 1987b). In the Boston cohort,
subjects were examined at age 8 years and 19-20 years. Pb levels measured in deciduous
teeth but not cord blood were associated with a higher rating of internalizing behaviors at
age 8 years but not in adulthood, indicating the lack of persistence of effects of early
exposure (Bellinger et al.. 1994a; Bellinger et al.. 1994b). Wasserman et al. (1998) also
examined blood Pb levels measured at various lifestages in the Yugoslavia cohort and
found that concurrent blood Pb was associated more strongly with anxious-depressed and
withdrawn behaviors at age 3 years than were blood Pb levels measured prenatally or
between ages 6 months and 2 years. Thus, while biomarkers of Pb exposure were
consistently associated with poorer mood and emotional state in children, there was no
clear indication of differences in association among biomarkers measured at various
lifestages. Associations between Pb biomarkers with mood also were consistently
reported in recent studies, with most studies examining concurrent blood Pb level (Liu et
al.. 2011a; Bao et al.. 2009; Rov et al. 2009b; Chiodo et al. 2004). However, several had
limited generalizability to the general population of U.S. children due to the examination
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of children residing near Pb sources or high prevalence of prenatal alcohol or drug
exposure (Liu etal.. 2011a; Bao et al. 2009; Chiodo et al.. 2004V
Associations of Pb biomarkers with mood and emotional state in children were observed
after adjusting for a wide range of potential confounding variables including parental
education and HOME score. Several studies that each adjusted for a different set of
covariates found similar effect estimates in univariate and multivariate models (Burns et
al.. 1999; Wasserman et al.. 1998; Bellinger et al.. 1994b). supporting the robustness of
associations with Pb biomarkers. Associations between blood Pb level and internalizing
behaviors also were observed in studies that adjusted for maternal psychopathology
(Burns et al.. 1999; Wasserman et al.. 1998V Thus, the evidence indicates that
associations observed between Pb biomarkers and internalizing behaviors in children are
not driven by confounding by any particular measured variable.
In the collective body of evidence, associations between blood Pb level and internalizing
behaviors were observed in populations of school-aged children with mean blood Pb
levels between 8 and 28 (ig/dL. Mean blood Pb levels were not related to the age of the
study population. In the limited investigation of populations with mean blood Pb levels of
approximately 5 (ig/dL, results were inconclusive. Chiodo et al. (2004) found an
association with internalizing behaviors in children (age 7 years) in Detroit, in whom the
prevalence of prenatal alcohol and drug use was high. Neither exposure was found to
influence associations with blood Pb level. Another study that examined Inuit children
(age 5 years) in Quebec, Canada, did not find an association between concurrent blood Pb
level and internalizing behaviors (Plusqiiellec et al.. 2010).
A common observation across studies was finding that Pb biomarkers were associated
with multiple indices of neurodevelopmental function, i.e., FSIQ, executive function,
externalizing, and internalizing behaviors, within the same population. Whereas some
studies found stronger associations for externalizing behaviors than for internalizing
behaviors (Plusquellec et al.. 2010; Bellinger et al.. 1994a; Sciarillo etal.. 1992). most
did not find a clear difference in the strength of association (Rov et al.. 2009b; Chiodo et
al.. 2004; Wasserman et al.. 1998; Bellinger et al.. 1994b). In the Port Pirie population,
Burns et al. (1999) found that lifetime average blood Pb levels were associated with
externalizing behaviors more strongly in boys and with internalizing behaviors more
strongly in girls, indicating potential sex-based differences. These findings demonstrating
Pb effects on a wide spectrum of cognitive and behavioral indices are not surprising
given that Pb exposure is linked to changes in the HPA axis and dopaminergic and
GABAergic systems, which are involved in mediating cognitive function, behavior, and
mood. Dietrich et al. (1987b) provided support for Pb exposure affecting a wide range of
neurodevelopmental outcomes by characterizing the direct and indirect effects of Pb
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using structural equations. The associations of prenatal Pb exposure (maternal and infant
day 10 blood) with poorer mood in infants aged 6 months were found to be mediated
through lower birth weight and/or shorter gestation. These results suggested that Pb may
exert its effects by impairing nervous system development. It is well known that the fetal
period is an active period for neuronal differentiation, dendritic branching, and
synaptogenesis. Thus, it is not surprising that prenatal Pb exposure effects on nervous
system development could result in a wide range of neurodevelopmental effects assessed
later in childhood.
5.3.3.4 Epidemiologic Studies of Mood and Psychiatric Effects in
Adults
Examination of the potential effects of Pb exposure on mood and psychiatric outcomes in
adults has received far less attention than that in children and cognitive function in adults.
Nonetheless, evaluation of mood states is an integral part of the neurocognitive test
battery of the World Health Organization (WHO), and it has been suggested 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, several early
studies of Pb-exposed workers (mean blood Pb levels ranging from 23.5 to 64.5 (ig/dL)
found higher prevalence of symptoms related to mood disorders and anxiety among Pb-
exposed workers than unexposed controls (mean blood Pb levels ranging from
15-38 (.ig/dL) (Schwartz et al.. 2005; Maizlish et al.. 1995; Parkinson et al. 1986; Baker
et al.. 1985; Baker et al.. 1984; Litis et al.. 1977).
While comprising a smaller body of evidence, studies of adults without occupational
exposures demonstrated associations of blood and bone Pb level with mood. Analyses of
men ages 48-70 years in the NAS indicated associations of both concurrent blood (mean:
6.3 (ig/dL [SD: 4.16]) and tibia (mean: 21.9 jj.g/g [SD: 13.5]) Pb levels with greater self-
reported symptoms of depression and anxiety (Rhodes et al.. 2003). As bone Pb is a
major contributor to blood Pb levels in adults without current occupational Pb exposure,
associations with both biomarkers may indicate effects of cumulative Pb exposure. In a
subsequent analysis of the same dataset, Raj an et al. (2007) found effect modification of
the associations for patella and tibia Pb levels by ALAD genotype, although results were
not in a consistent direction for a particular genotype. For a majority of the mood
symptoms considered, tibia bone Pb levels were associated with larger ORs among men
with the ALAD 1-1 genotype. In contrast, ORs for associations between patella Pb levels
and several mood symptoms such as depression and positive symptom distress index
were larger among ALAD 1-2/2-2 carriers. In the NAS, inconsistent effect modification
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by ALAD genotype also was observed for associations between tibia Pb levels and
cognitive performance (Raian et al. 2008) (Section 5.3.2.5).
A cross-sectional study of 1,987 adults age 20-39 years participating in NHANES
1999-2004 was the largest study of mood and included both men and women (Bouchard
et al.. 2009V However, only concurrent blood Pb levels were available for analysis.
Investigators assessed major depressive disorder (MDD), panic disorder and generalized
anxiety disorder (GAD) using the WHO Composite International Diagnostic Interview,
which follows criteria defined in DSM. Adults with concurrent blood Pb levels higher
than 0.7 (ig/dL had higher prevalence of all three symptoms. Adults in the highest
quintile of blood Pb level (>2.11 (ig/dL) had the highest OR for MDD (OR: 2.32 [95%
CI: 1.13, 4.75]) and panic disorder (OR: 4.94 [95% CI: 1.32, 18.48]) compared with
adults with blood Pb levels < 0.7 (ig/dL. For all endpoints, ORs were larger in analyses
excluding current smokers. While this study demonstrated associations with relatively
low concurrent blood Pb levels, it is important to note the uncertainty of the magnitude,
timing, frequency, and duration of Pb exposure that contributed to the observed
associations.
In analyses of cohorts in California and New England, Opler et al. (2008; 2004) reported
associations between higher prenatal levels of plasma S-ALA and subsequent diagnosis
of schizophrenia spectrum disorder (ascertained using DSM-IV criteria) in adolescence
and adulthood. In the absence of blood Pb levels, investigators measured S-ALA levels in
stored serum samples to serve as a surrogate for Pb exposure citing previous observations
of a high correlation (0.90) between 8-ALA levels > 9.05 ng/mL and blood Pb levels >
15 (ig/dL. In the California cohort, a 5-ALA level > 9.05 ng/mL was associated with
schizophrenia spectrum disorder with an OR (95% CI) of 2.43 (0.99, 5.96), adjusting for
maternal age at delivery. In analyses combining the California and New England cohorts,
a 8-ALA level >9.05 ng/mL was associated with schizophrenia spectrum disorder with
an OR (95% CI) of 1.92 (1.05, 3.52), adjusting for maternal age and education. A
covariate-adjusted OR was not presented for the New England cohort alone, and it
appeared that the association in the combined cohorts was driven by that observed in the
California cohort. Studies in other populations with direct measurements of blood Pb
levels are warranted to characterize the potential effects of Pb exposure on schizophrenia.
These limited available data indicate associations with higher blood Pb levels (>
15 (ig/dL) in individuals born in the 1950s and 1960s who likely had higher early Pb
exposures. Thus, these findings may have limited relevance to current lower levels of Pb
exposure.
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5.3.3.5 Toxicological Studies of Mood, Emotional, and Psychotic
Changes
As in epidemiologic studies, neurotoxicological studies often focus more on motor,
sensory, behavioral or cognitive outcomes and less on psychological pathologies. As
described in Section 5.3.3.3 above, epidemiologic studies indicating associations of Pb
biomarkers with mood disorders in children and adults have grown in recent years.
Likewise, studies increasingly have found that developmental Pb exposure (gestation and
lactation) affects emotional state and mood disorder-like behavior in adult animal
offspring. Offspring ofWistar rats treated with 10 mg of Pb-acetate daily by gavage
during pregnancy (G) or pregnancy plus lactation (G+L) were tested in the open field test
and the forced swimming test also known as the Porsolt test (de Souza Lisboa et al..
2005). Blood Pb levels in the pups at PND70 were 5-7 (ig/dL. The open field test
monitors activity levels and movements of animals in three dimension and is a measure
of hyperactivity, emotion (grooming or freezing), and exploratory behavior (rearing). In
the forced swim test, animals are placed in a tank and monitored for helplessness, a
response common in animal models of depression. (G+L) Pb-treated male rats had
increased emotionality with the open field test as indicated by decreased rearing and
increased immobility. (G+L) Pb-exposed female offspring had a significantly increased
depressive phenotype in the forced swim test as indicated by increased immobility
(de Souza Lisboa et al. 2005). It is interesting to note that this is one of many Pb-induced
changes that seem to be sex-specific with males showing increased emotionality and
females showing elevated depressive-like symptoms (Figure 5-16). In another study,
anxiety was assessed in offspring after G+L Pb exposure. Dams received Pb-enriched
drinking water (2.84 mg/mL Pb-acetate trihydrate) and produced pups with blood Pb
levels of 698 ng/g at PND25. Using the elevated plus maze test, Pb-exposed animals
displayed no signs of anxiety (Molina et al.. 2011). However, it is important to note the
high concentrations of Pb exposure used in this study. Another study found that female
pups exposed to Pb-acetate (0.2% to pups PND1-30) had increased anxiety at PND30 as
measured by a decreased percentage of entries and time spent in the open arms of the
plus-maze (Fox et al.. 2010).
Depression may seem initially like an unexpected comorbidity for immune inflammatory
dysfunction, but many forms of depression are linked with the same cytokine imbalances
that occur with Pb-induced innate immune dysfunction (Maes; Pace and Miller. 2009).
Some researchers use sickness behavior and its associated malaise as a model for
depression. Sickness behavior is characterized by overall malaise, decreased food intake,
immobility and changes in core body temperature. Dyatlov and Lawrence ("2002)
observed in mice that sickness behavior, which is due to an interaction of the immune
system and the CNS, was potentiated by Pb exposure (resulting in blood Pb level:
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17 (ig/dL) and was correlated with depletions in specific thymic T-cell populations
(Figure 5-16). Pb exposure also potentiated the infection-dependent elevation in IL-1(3, a
cytokine that has been shown to inhibit hippocampal glutamate release in young but not
aged animals. Sickness behavior was induced with Listeria monocytogenes infection; Pb
exposure occurred from birth through lactation and was continued for a brief period after
weaning until the experiment was terminated.
Neuro-
transmitters
NMDA Receptor
Hypofu fiction
Immune
Elevated Cytokines
Sickness Behavior
Structural
Decreased
Hippocampal
Neurogenesis
Behavioral
Humans
i
Epidemiological
Associationsof Pb
With Schizophrenia
Cf More
emotional
Rodents
X \
^ Depression
Note: The blue headings provide four independent possible contributors to development of psychiatric disorders in Pb-exposed
individuals.
Figure 5-16 Animal toxicology evidence of possible Pb-induced contributors
to the development of mood and psychotic disorders.
Schizophrenia is associated with a shortened lifespan in humans as reflected by increased
standardized mortality ratio (McGrath et al.. 2008). An environmental origin of
schizophrenia was proposed years ago (Tsuang. 2000). and prenatal Pb exposure,
assessed using ALAD activity as a biomarker, was linked to schizophrenia in a group of
adolescents and young adults in California (Opler etal.. 2008; Qpler etal.. 2004).
Because of these observations, the animal toxicological field is beginning to explore the
mechanisms that may contribute to schizophrenia development and has proposed two
explanations. These are Pb-induced NMDA receptor (NMDAR) hypofunction and Pb-
induced decreases in hippocampal neurogenesis (Figure 5-16 to Figure 5-18). Pb may
bind a divalent cation site in the NMDAR and allosterically inhibit glycine binding
(Hashemzadeh-Gargari and Guilarte. 1999); human studies of patients with schizophrenia
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have shown aberrations at this site (Covle and Tsai. 2004). These findings are consistent
with the glutamatergic hypothesis of schizophrenia which purports that NMDAR
noncompetitive antagonist use in patients with schizophrenia exacerbates their psychotic
symptoms and that administration of antagonist to non-psychotic subjects can induce a
schizophrenic phenotype. The second mechanistic hypothesis for Pb-associated
schizophrenia induction is decreased hippocampal degenerate gyrus (DG) neurogenesis,
which is seen in patients with schizophrenia (Kcmpcrmann et al. 2008; Re if et al.. 2006),
in animal models of schizophrenia (Maeda et al.. 2007) and in animal models with
developmental Pb exposure (Verina etal.. 2007; Jaako-Movits et al.. 2005) (Figure 5-16
to Figure 5-18). Animal models of schizophrenia (i.e., phencyclidine administration)
show decreased hippocampal DG neurogenesis that can be reversed by treatment with
clozapine, which is often used to treat schizophrenia (Maeda et al.. 2007). These DG
pathways are also NMDAR-dependent. Studies cited in this section are further detailed in
other sections of the ISA. Another study gives insight into pathways involved in Pb-
induced changes in neurogenesis in the developing brain. Exposure of zebrafish embryos
to Pb (50-700 |_iM Pb-acetate in embryo medium from 0 to 6 days post hatch) caused
significant apoptosis (increased TUNEL positive brain cells) and impaired neurogenesis
as shown by decreased levels of gfap and huC in the brain; however two other genes
involved in neurogenesis, crestin and neurogeninl, were unaffected in brains of Pb
exposed embryos (Dou and Zhang. 2011). Thus, toxicological studies indicate that Pb
exposure may result in mood disorders via behavioral, neurochemical, and ultrastructural
changes.
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Human Populations or Cohorts
Rodent
Studies
Pb exposure and Schizophrenia
Associated in Human Cohorts
(Epidemiologic)
NMDA Receptor Decreased Hippocampal
Hypofunction	Neurogenesis
(Neurochemical)	(Structural)
Glutamatergic hypothesis of
schizophrenia
Pb exposure
NMDAR
Antagonism
Pharmacologically-
induced Schizophrenia in
Animal models
Figure 5-17 Schematic representation of the contribution of Pb exposure to
the development of a phenotype consistent with schizophrenia.
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B
a 3000
Con
Source: Reprinted with permission of Elsevier Science, Fox et al. (20101
Note: (A) Control; (B) Pb exposed light micrograph pictures of Brd-U positive cells; (C) Counts of Brd-U positive cells (proliferating
cells); and (D) Volume of dentate gyrus. *p <0.05 vs. control.
Figure 5-18 Neurogenesis (production of new cells) in the rat dentate gyrus
after postnatal Pb exposure.
5.3.3.6 Integrated Summary of Behavior and Mood
Epidemiologic studies in children demonstrate associations of higher blood Pb levels
with a range of behavioral problems, with the weight of evidence demonstrating
associations with inattention and hyperactivity as rated by parents or teachers and as
assessed using objective neuropsychological tests. Previous studies found associations
with early childhood blood or tooth Pb level (i.e., ages 2-6 years), and recent studies
expanded evidence to include associations with concurrent blood Pb level. Recent
epidemiologic studies consistently found associations with inattention and hyperactivity
in children ages 1 to 12 years with mean concurrent blood Pb levels of 2 to 5 (ig/dL (Cho
et al.. 2010; Nicolescn et al.. 2010; Plusquellec et al.. 2010; Chiodo et al.. 2007;
Plusquellec et al.. 2007). similar to those associated with cognitive function decrements.
The epidemiologic findings are strengthened by observations in animals of Pb-induced
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inattention and impulsivity in homologous tests of response inhibition (e.g., DSA, FI,
DR). Tests in children such as CPT, Stroop, WCST share homology with tests in animals
such as DSA, DR, and FR tasks in that they all test the ability to inhibit inappropriate
responses, organize behavior in response to varying stimuli, and learn from the
consequences of previous actions. Inattention and impulsivity in animals are most clearly
indicated with gestational and early postnatal Pb exposures that result in blood Pb levels
of 10 to 40 (ig/dL. These findings in children and animals for Pb-associated dysfunction
in response inhibition and response variability may be explained by observations that Pb
affects dopaminergic neurons of the frontal striatum of the brain by altering dopamine
release and receptor density . The circuitry in this regions is thought to mediate response
inhibition. Whereas previous evidence was inconsistent, several recent epidemiologic
studies indicate associations between higher concurrent blood Pb level and higher
prevalence or incidence of ADHD diagnosis and its contributing diagnostic indices in
children ages 8-17 years (C'ho et al.. 2010; Nicolescu et al.. 2010; Rov et al.. 2009a; Nigg
et al.. 2008; Braun et al.. 2006). The biological plausibility for associations with ADHD
is strongly supported by the large epidemiologic and toxicological evidence base
demonstrating Pb-associated increases in inattention and impulsivity, both of which are
primary symptoms of ADHD. A smaller but equally consistent body of evidence
indicated associations of concurrent and early childhood blood Pb levels with social
misconduct in children and delinquent behaviors in adolescents and young adults
(Chandramouli et al.. 2009; Braun et al.. 2008; Wright et al.. 2008; Chiodo et al.. 2007).
Associations of blood Pb levels with ADHD, misconduct, and delinquency were
observed in populations of children with a wide range of blood Pb levels, 1 to 11 (ig/dL,
all similar in the strength of evidence.
While mood and emotional state have been examined less frequently compared with
inattention and misconduct, several studies found associations of biomarkers of
cumulative Pb exposure (i.e., tooth or childhood average blood Pb) and concurrent blood
Pb levels with parental or teacher reports of withdrawn behavior or depression in children
with mean blood Pb levels 8-28 (ig/dL (Section 5.3.3.3). These findings in children are
supported by a small body of toxicological studies in which prenatal plus lactational Pb
exposure resulted in depression-like behavior in rodents.
Rather than examining externalizing behaviors and criminal behavior, a small body of
studies of behavior in nonoccupationally-exposed adults examined and found
associations of blood (Bouchard et al.. 2009) and tibia (Raj an et al.. 2008) Pb levels with
depression and anxiety symptoms. All of these studies used single assessments of Pb
biomarker levels and outcomes and analyzed associations in a cross-sectional manner.
Thus, there is uncertainty regarding the critical level, timing, frequency, and duration of
Pb exposure associated with mood and psychiatric symptoms. The associations with bone
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Pb levels indicate an effect of cumulative Pb exposure. Concurrent blood Pb levels in
adults reflect both cumulative and recent exposure, and it is uncertain what are the
relative contributions of past versus recent Pb exposures to the observed associations.
Associations of Pb biomarkers with withdrawn behavior and anxiety in children and
adults may by explained by evidence for Pb-induced changes in the HPA axis,
dopaminergic and GABAergic CNS processes. These processed have been shown to
mediate anxiety and depression. While it may seem that Pb exposure affects different
behaviors in children and adults, it is important to acknowledge differences in the
evidence base. Studies of behavior in children and young adults have focused on
inattention and misconduct; studies of older adults did not examine externalizing
behaviors or misconduct. 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.
5.3.4 Sensory Organ Function
5.3.4.1 Epidemiologic Studies of Sensory Organ 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 and poorer
auditory processing in children (U.S. EPA. 2006b). Such evidence is limited largely to
studies described in the 2006 Pb AQCD. The prospective CLS with repeat measurements
of blood Pb prenatally to age 5 years provided information on potentially important
lifestages of exposure. In this cohort with higher blood Pb levels than those in
contemporary U.S. children (lifetime average mean: 17.4 [SD: 8.8] (ig/dL), poorer
auditory processing was associated with higher prenatal (maternal), neonatal (10-day),
early childhood, and lifetime average blood Pb levels, with the strongest associations
observed for neonatal blood Pb level. A 1 j^ig/dL higher neonatal blood Pb level was
associated with a 0.20-point (p < 0.01) and 0.26-point (p< 0.10) lower score on the total
and left ear Filtered Word test (indicative of incorrectly identified filtered or muffled
words), after adjusting for hearing screen, social class, HOME score, birth weight,
gestational age, obstetrical complications, alcohol consumption, and prenatal and
postnatal blood Pb levels (Dietrich et al.. 1992). Overall, the findings pointed to an effect
of early Pb exposure during infancy.
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Large U.S. studies, including NHANES II (Schwartz and Otto. 1987) and the Hispanic
Health and Nutrition Examination Survey (HHANES) also found associations between
higher blood Pb level and lower auditory function; however, their cross-sectional design
precluded comparisons among various lifestages of exposure (Schwartz and Otto. 1991).
In these studies, concurrent blood Pb level (median: 8 (ig/dL) from 6 to 18 (ig/dL was
associated with a 2-db loss in hearing and an increase in the percentage (15%) of children
with a substandard hearing threshold (2,000 Hz). Higher concurrent blood Pb level also
was associated with higher hearing thresholds across several frequencies in a population
of children in Poland with similar blood Pb levels (median: 7.2 (ig/dL [range: 1.9 to 28])
(Osmanetal.. 1999). In the HHANES and Polish studies, associations persisted in
analyses restricted to subjects with concurrent blood Pb levels below 10 (ig/dL.
Mechanistic support for these observations in children was provided by Otto and
colleagues (Otto and Fox. 1993; Otto et al. 1985b). who found associations of blood Pb
level with lower brainstem auditory evoked potentials in children. Rothenberg et al.
(1994b) and Rothenberg et al. (2000) reported similar findings; however, the direction of
association differed between prenatal (maternal) and postnatal (ages 1-4 years) blood Pb
level. Postnatal blood Pb level was associated with lower interpeak intervals in auditory
evoked potentials at age 5-7 years. Prenatal maternal blood Pb level showed abiphasic
relationship, with a lower evoked potentials at blood Pb levels of 1-8 j^ig/dL and higher
evoked potentials at blood Pb levels of 8-30 (ig/dL. Recent studies also aimed to identify
the locus in the auditory system where Pb may exert its effects on auditory function.
Investigation was limited to a population of children with high blood Pb levels (means 33
and 37 (.ig/dL) living in Pb glazing communities in Ecuador (Buchanan et al.. 2011;
Counter et al.. 2011). In these studies, 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.
5.3.4.2 Epidemiologic Studies of Sensory Organ Function in
Adults
Studies of auditory function reviewed in the 2006 Pb AQCD provided consistent
evidence of association between blood Pb levels and changes in auditory evoked
brainstem potentials in occupationally-exposed adults, but less consistent findings for
hearing thresholds (U.S. EPA. 2006b). A few new studies of Pb workers found Pb-
associated increases in hearing thresholds. A recent study provided new evidence in
nonoccupationally-exposed adults for associations of tibia Pb levels with hearing loss.
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Men in the NAS study were free of hearing loss at baseline and had hearing tested
repeatedly (median 5 observations per subject) over a median of 23 years (Park et al.
2010). Higher tibia Pb level, 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. Blood Pb was not examined in this study. In cross-sectional
analyses adjusted for age, race, education, body mass index, pack-years of cigarettes,
diabetes, hypertension and occupational noise (based on a job-exposure estimate from
occupations), and presence of a noise notch (indicative of noise-induced hearing loss),
higher patella bone Pb level (measured within 5 years of hearing test) was associated with
a higher hearing thresholds for frequencies greater than 1 kHz. A 21 jxg/g (IQR) increase
in patella bone 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. Similar, but slightly weaker associations
were found for tibia bone Pb levels.
With new investigation, the weight of evidence indicates that blood Pb levels are
associated increased hearing thresholds in adults with occupational Pb exposures. In a
study of 183 Pb workers with relatively low blood Pb levels, 1 to 18 (ig/dL, higher blood
Pb level was correlated with higher hearing threshold at 4 kHz (Forst et al.. 1997). Other
studies examined workers with much higher blood Pb levels. In a study of 220 Pb battery
workers with higher blood Pb levels (mean: 56.9 (ig/dL [SD: 25.3]) (Wu et al.. 2000).
although hearing impairment was associated with a measure of cumulative Pb exposure
based on years of work and ambient Pb measurements, no association was found with
blood Pb levels at the time of hearing testing in analyses adjusted for age, sex, and
duration of employment. Another cross-sectional study examined 259 steel plant workers
with no parental history of ear-related problems, no congenital abnormalities, no
occupational organic solvent exposure, and hearing loss difference no more than 15 dB
between both ears (Hwang et al.. 2009). The participants had a mean (SD) concurrent
blood Pb level of 54.3 (34.6) (ig/dL. Average noise levels also were measured in work
areas and dichotomized at 80dB. In analyses adjusted for age and work area noise
(dichotomized at 80 dB), workers with blood Pb level > 7 (ig/dL had a statistically
significant higher prevalence (range of ORs: 3.06 to 6.26) of hearing loss at frequencies
of 3, 4, 6, and 8 kHz compared to workers with blood Pb levels < 4 (ig/dL.
A hospital-based case-control study recruited workers referred for hearing testing
(average hearing thresholds above 25 dB) as cases and workers with normal hearing
thresholds who were having occupational health examinations for other reasons as
controls (Chuang et al.. 2007). The 121 cases had a geometric mean blood Pb level of
10.7 (ig/dL, and the 173 controls had a geometric mean blood Pb level of 3.9 (ig/dL based
on measurements in samples collected at the time of the study. In models that adjusted for
age, smoking, alcohol consumption, years of noise exposure, as well as Mn, As, and Se
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levels in blood, higher blood Pb levels were associated with a statistically significant
higher average hearing threshold (0.5-6 kHz).
In summary, together, the findings from NAS (Park et al.. 2010) and studies of Pb-
exposed workers indicate that biomarkers of Pb exposure are associated with lower
auditory function in adults. Evidence for association with concurrent blood Pb levels is
provided studies of adults with current occupational Pb exposures. Because only bone Pb
levels measured after auditory testing were examined in the NAS study, further
investigation is required to characterize the timing, level, frequency, and duration of Pb
exposure contributing to lower auditory function in adults without occupational
exposures.
5.3.4.3 Toxicological Studies of Sensory Organ Function
The 1986 and 2006 Pb AQCDs detailed the effects of Pb in animals on vision, including
the retina and CNS visual processing areas, as well as the auditory system and described
possible or known mechanisms of action where available. Pb exposure effects on sensory
dysfunction and impaired sensory processing have been postulated to contribute to Pb-
associated effects on neurocognition and attention. In support of this hypothesis, Dietrich
et al. (1992) found that higher prenatal (maternal) and postnatal blood Pb levels were
associated with both lower performance on the screening test for auditory processing
disorders and lower cognitive function in 5 year-old CLS children (Section 5.3.4.1). New
research in this area expands upon the extant evidence by exploring sensory function in
animals with lower Pb exposures and blood Pb levels.
Auditory Effects
The 2006 Pb AQCD discussed impaired auditory function in nonhuman primates exposed
to Pb from gestation through age 8-9 years (resulting in blood Pb levels 33-56 (ig/dL
during Pb exposure period). Brainstem auditory evoked responses (BAER), which are
used as a general test to assess neurological auditory function, revealed Pb-related effects
that persisted even after Pb exposure had ceased and blood Pb levels had returned to
baseline. In Pb-exposed animals (birth to age 13 years), half of the pure tone detection
thresholds were outside of the control range at certain frequencies (Rice. 1997). In
concordance with the data from developmentally Pb-exposed laboratory animals,
elevated auditory thresholds, as measured with brainstem auditory evoked response
(BAER), have also been associated with blood Pb levels in children (Rothenberg et al..
2000; Rothenberg et al.. 1994a). In addition to indicating hearing loss, BAER can
indicate impaired synaptic maturation and incomplete neuron axon myelination leading to
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impaired neuronal conduction (Schwartz and Otto. 1991; Gozdzik-Zo 1 nierkie\\ icz and
Moszviiski. 1969. Thus, the findings from Rice (1997) indicated that early life Pb
exposure impaired auditory function. The cochlear nerve in both developing and mature
humans appeared to be especially sensitive to the Pb insult. At low to moderate Pb
exposures, elevated thresholds and increased latencies were observed in brainstem
auditory evoked potentials.
In a recent study, Laughlin et al. (2009) studied rhesus monkeys exposed to Pb-acetate
gestationally through age 5.5 months (dam drinking water, 3 months prior to mating until
weaning, resulting in bone Pb levels at 11 years of 7 and 13 (ig/dL for prenatal and
postnatal groups, respectively and blood Pb levels during Pb exposure of 35 and
46 (ig/dL, respectively) and conducted auditory threshold testing and threshold task
testing at 13 years of age after blood Pb levels had returned to baseline. 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 elevations in auditory thresholds in animals. Auditory threshold task-
related behavioral testing was also impaired in Pb-exposed animals. This study has
multiple limitations which likely contributed to its not finding statistically significant
aberrations. These limitation included 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
and thus had fewer available measurements, differences between the sexes in inattention,
and mixing of the postnatal Pb and prenatal Pb animals into one group (Pb-exposed
animals).
In summary, studies in nonhuman primates have shown lower BAER due to chronic Pb
exposure (gestation through adulthood) that result in blood Pb levels in the range of
33-56 (ig/dL. In concordance with these findings, previous epidemiologic studies
described lower BAER in children with higher prenatal maternal blood Pb levels and
concurrent blood Pb levels. Collectively, these findings provide support for the
associations observed between higher blood Pb levels and increased hearing thresholds
(concurrent blood Pb) and lower auditory processing in children (neonatal)
(Section 5.3.4.1). The animal evidence for long-term Pb exposure effects provide
coherence with findings in adults with occupational Pb exposures for associations
between blood Pb levels and increasing hearing thresholds and in nonoccupationally-
exposed adults, for associations with biomarkers of cumulative Pb exposure (i.e., tibia
and patella Pb level) (Section 5.3.4.2).
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Effects on Vision
In toxicological studies, Pb has been shown to affect multiple aspects of the visual system
from the retina, to the sensory processing areas of the brain, to neurons involved in
vision. The selective action of Pb on retinal rod cells and bipolar cells is well documented
in earlier Pb AQCDs (Fox et al.. 1997; Fox and Sillman. 1979) and research in this area
continues to date. Pb exposure during perinatal development and adulthood has also been
shown to affect the visual cortex (Costa and Fox. 1983) and subcortical neurons (Cline et
al.. 1996). Extensive work in nonhuman primates with various Pb exposure paradigms
(development or lifetime Pb) showed sensory impairment, i.e., dysfunction of spatial and
temporal visual function (Rice. 1998). Environmentally-relevant doses of Pb (10-3 (.iM)
administered to tadpoles inhibited the growth of developing neurons in the subcortical
retinotectal pathway, the main efferent from the retina (Cline et al.. 1996). Functional
tests like electroretinograms (ERGs) show Pb-related aberrations in children, rodents and
nonhuman primates.
The animal toxicological data show that developmental windows and the dose of Pb
contribute to the complex and variable effects of Pb with the retina. Table 5-12
summarizes Pb-related effects in retinal ERG studies. Female rats exposed postnatally to
moderate and high levels of Pb (0.02 or 0.2% Pb-acetate exposure in dam drinking water
from birth through weaning, resulting in weaning blood Pb levels of 19 and 59 (ig/dL,
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) (Table 5-12). Similar results were obtained in multiple studies conducted in in vitro
models (Otto and Fox. 1993; Fox and Farber. 1988; Fox and Chu. 1988). Monkeys
exposed to moderately 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 (ig/dL,
respectively) had persistently increased maximal retinal ERG amplitude (B-wave only,
supernormality) and increased mean ERG latency when assessed 2 years after
termination of Pb exposure when blood Pb levels were <10 (ig/dL (Lilienthal et al.. 1988)
(Table 5-12).
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Table 5-12 Summary of Pb-related retinal ERG studies
Study
Species
Sexa
Pb Exposure
Protocol/Dose
Maximal
Blood Pb
Level (ng/dL)
ERG
Abnormality
Progenitor
cell
proliferation
Retina Cellular
Apo ptosis
Retina Dopamine
Levels
Retina
Cell Layer
Thickness
Fox et al.
(2008)
Long-
Evans
Rat
F
P renatal-P N D10b
DW
Low 27 ppm
Moderate, 55 ppm
High, 109 ppm
12
24
46
Supernormal
Supernormal
Subnormal
Yes
Yes
No
Not affected
Not affected
Yes
Dose-dependent J,
Dose-dependent J,
Dose-dependent J,
t
t
1
Lilienthal et
al, (1988)
Rhesus
Monkey
M
& F
Pre- and Post-natal
(lifetime), DW
350 ppm
600 ppm
-50
-115
Supernormal
Supernormal

-
-
-
Fox et al.
(1997)
Long-
Evans
Rat
F
PND1-PND21
0.02% DW
0.2% DW
19
59
Subnormal
Subnormal
-
Yes
Yes

1
1
Rothenberg
etal.
(2002a)
Human
children
M
& F
Prenatal
1st trimester
> 10.5
Supernormal
-
-
-
-
Guguchkova
et al. (1972)
Human
M
Occupational

Subnormal



-
Otto and Fox
(1993)
Human
M
Occupational

Subnormal



-
aF: Females; M: Males
bPND: postnatal day
M—11 Denotes not measured.
A recent study exposed female Long-Evans rats to low (27 ppm), moderate (55 ppm), and
high (109 ppm) levels of Pb-acetate in drinking water beginning 2 weeks before mating,
throughout gestation, and until PND10 (G+P exposure) (Fox et al.. 2008) (Table 5-12).
Blood Pb levels in G+P Pb-exposed pups in the three groups were 10-12 (ig/dL (lower),
21-24 (ig/dL (moderate), and 40-46 (ig/dL (higher). This developmental window in the
retina of the rat is equivalent to gestational human retinal development. Results of this
rodent study demonstrated persistent supernormal scotopic rod photoreceptor-mediated
ERGs (lower and moderate Pb exposure) similar to the associations observed between
ERG and prenatal maternal blood Pb levels >10.5 (ig/dL in male and female children
(Rothenberg et al.. 2002a). Supernormal scotopic ERGs may be recorded without other
overt opthalogical changes and are rarely seen in the clinical setting (Terziivanov et al..
1983). Lower and moderate levels of G+P Pb exposure 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
G+P Pb exposure (109 ppm, blood Pb level of 46 (ig/dL) or moderate to higher level
postnatal exposure (PND1-21, blood Pb levels 19 and 59 (ig/dL) statistically significantly
decreased the number of rods in central and peripheral retina, induced scotopic ERG
subnormality in adult rats (Table 5-12), and statistically significantly decreased the retinal
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Zn concentration. Pb exposure induced concentration-dependent decreases in adult rat
retinal dopamine synthesis and utilization/release (Fox et al.. 2008). Thus, the dose of Pb
and the exposure window produce complex interactions in the retina that differentially
affect retinal histology and functional tests, i.e., ERG as summarized in Table 5-12.
low intensity	High intensity
I!85-1 stimulus I)	he-si stimulus 21
Note: *p <0.05
Source: Fox et al. (20081
Figure 5-19 Retinal a-wave and b-wave ERG amplitude in adult rodents after
prenatal and early postnatal Pb exposure.
A recent study of adult zebrafish that were exposed as embryos (2 to 24 hours post-
fertilization) to water containing 0.03 |iM PbCl2 showed fish with impaired response to
visual stimulation (visual response to a rotating bar) under low light conditions. These
zebrafish also failed to respond normally to mechanosensory stimulation (0.01 and 0.03
|iM PbCl2), showing a significantly impaired startle response. These data show that low-
level developmental Pb exposure causes sensorimotor deficits in fish (Rice et al.. 2011).
Another study evaluating visual spatial acuity in rhesus monkeys (Laughlin et al.. 2008)
exposed to Pb-acetate postnatally (PND8-26 weeks of age via commercial milk formula,
achieving a target blood Pb of 35-40 (ig/dL) found no effects of Pb exposure on spatial
acuity as assessed with modified the Teller preferential looking paradigm.
Mechanistic understanding of the effect of Pb on the visual system includes its capability
to displace divalent cations, act as an inhibitor of physiological enzymes, regulate cell
proliferation and apoptosis, impair and perturb normal neuroanatomy formation,
i.e., cytoarchitecture in the brain, and affect neurotransmitters. The effects of Pb on the
retina have been shown to by mediated by its capability to act as a cGMP
phosphodiesterase (PDE) inhibitor (Srivastava et al.. 1995; Fox and Farber. 1988). The
drug sildenafil citrate, another cGMP PDE inhibitor, can also cause visual problems
including alterations in scotopic ERGs (Laties and Zrenner. 2002). With postnatal
exposure of animals or in vitro Pb exposure of isolated rods, Pb has been shown to induce
elevated cGMP which contributes to elevated rod calcium concentration (Fox and Katz.
1992) and subsequently to apoptotic cell death in a concentration-dependent fashion. In
s
a.
¦
a>
«> w
> n. 300
I IB Control	•
~ Low Pti	* JL
¦	Moderate Pb	pH
¦	High Pb
fh	II
Low intensity	High intensity
ttost stimulus I)	(tost Stimulus 2}

H :]l" riUinsily
(test stimulus 21
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separate work in mice, low and moderate doses of Pb (27 or 55 ppm Pb-acetate in dam
drinking water from gestation to PND10, resulting in blood Pb levels of 12 and 25 (ig/dL,
respectively), induced greater rod and rod bipolar cell neurogenesis (proliferation) and
greater thickness and cell number of the outer and inner neuroblastic layers of the retina
(Giddabasappa et al.. 2011; Fox et al.. 2008). Rodents with moderate dose G+P Pb
exposure (resulting in blood Pb level of 25 j^ig/dL) had 27-fold greater and prolonged
retinal progenitor cell proliferation (Giddabasappa et al.. 2011); at higher doses of Pb
(109 ppm Pb-acetate, resulting in blood Pb levels of 56 (ig/dL) there was no rod
neurogenesis. Nitric oxide has been shown to regulate retinal progenitor cell proliferation
in chick embryos (Magalhaes et al.. 2006). Thus, these authors postulated that impaired
NO production may contribute to aberrant retinal cell proliferation (Giddabasappa et al.
2011). Pb exposure has been shown to impair NO synthase activity in other organs
(Section 5.2.4.5). 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. In the
aforementioned model of G+P rodent Pb exposure through PND10, Pb decreased
dopamine (DA) synthesis and use in a concentration-dependent manner (Fox et al.. 2008)
(Figure 5-20). These new data provide further insight into retinal changes by showing
increased proliferation of Pb-exposed retinal progenitor cells without changes in
apoptosis in G+P exposed rats (Fox et al.. 2008).
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Dark-adapted
Light-adapted
Control
Low Moderate
GLE
L C TIT)
Low Moderate High
GLE
0.35 _
DOPAC/DA
HVA/DA

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5.3.5
Motor Function
Some studies in children have assessed fine motor function, i.e., response speed,
dexterity, eye-hand coordination, as part of a battery of cognitive function and behavioral
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.
Multiple studies were conducted in the CLS cohort at age 6 years and in adolescence at
ages 12 and 15-17 years. Assessments in younger children indicated associations of
concurrent and lifetime average blood Pb levels but not prenatal blood Pb levels with
poorer visuomotor control and upper limb dexterity (Dietrich et al.. 1993b) and poorer
postural balance (Bhattacharva et al.. 1995V Associations with both fine and gross motor
function were adjusted for HOME score and race. Additional covariates included
maternal IQ, SES, and sex for fine motor functions (Dietrich et al.. 1993b) and height,
BMI, birth weight, bilateral ear infection, and foot area for postural balance
(Bhattacharva et al.. 1995). Blood Pb levels were associated with fine and gross motor
function in unadjusted and adjusted analyses, indicating that bias from the measured
confounders was not driving the observed associations. Subsequent analyses in the CLS
demonstrated associations of earlier childhood average blood Pb levels (0-5 year average
or 78 month average) with poorer fine (Ris et al.. 2004) and gross motor function
(Bhattacharva et al.. 2006) assessed in adolescence. While these findings suggest the
persistence of Pb effects or early, long-term exposure effects, it is important to note that
blood Pb levels measured later in childhood or concurrently with motor function were not
examined.
Studies in other populations found associations of blood Pb level (primarily concurrent)
with poorer fine motor function in children ranging in age from 3 to 16 years
(Palaniappan et al.. 2011; Min et al.. 2007; Wasserman et al.. 2000). Among children in
New England participating in NECAT (described in Section 5.3.2.1) Surkan et al. (2007)
found higher blood Pb levels to be associated with lower FSIQ but better fine motor
function as indicated by faster finger tapping speed. Gross motor function also was
assessed in other populations of children. In the Yugoslavian cohort, lifetime average
blood Pb level was not associated with gross motor function (Wasserman et al. 2000).
Concurrent blood Pb level was associated with greater sway oscillation, alternating arm
movements, and action tremor in a group of Inuit preschool children in Quebec, Canada
(Despres et al.. 2005). Investigators considered potential confounding by several
variables including HOME score, maternal education and nutrient levels. In animal
studies, Pb exposure has shown mixed effects on endurance, balance and coordination as
measured by rotarod performance. G+P Pb-exposed male mice had significantly shorter
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mean latencies to fall from the rotarod compared with controls; females were unaffected.
Further, low dose G+P Pb-exposed male mice had significantly poorer rotarod
performance (i.e., fell off more quickly) than did high dose G+P male mice (resulting
blood Pb levels at PND10: 10 (ig/dL and 42 (ig/dL, respectively), indicative of a
nonlinear concentration-dependent relationship (Leasure et al. 2008). Other rotarod
experiments at higher doses of Pb exposure and at various speeds of rotarod rotation
yielded mixed results (Kishi et al. 1983; Grant et al.. 1980; Overmann. 1977).
The collective body of evidence demonstrates that within the same population of
children, blood Pb levels are associated with decrements in cognitive function, attention,
and fine and gross motor function. Studies found associations with concurrent and
childhood average blood Pb levels mostly in populations with mean levels ranging from
11 to 28 (ig/dL. Despres et al. ("2005) found impaired gross motor function in a children
with a mean blood Pb level of 5 (ig/dL. Min et al. (2007) found impaired fine motor
function in children with a mean concurrent blood Pb level of 2.9 (ig/dL; however, the
results were not adjusted by SES-related variables, motor skills are the result of the
coordination of complex cognitive and physical processes in the cortex, cerebellum,
vestibular systems, and visual system, and biological plausibility for the associations of
blood Pb levels with fine motor skills observed in children is provided by observations
that Pb exposure affects development and function of these systems.
5.3.6 Seizures in Animals
One neurological sign of high dose Pb exposure has been the development of epileptic
form activity or seizures in animals (krishnamoorthv et al.. 1993). However, earlier
studies in the animal toxicological literature exploring the effects of Pb on seizure
activity and threshold showed mixed results. Pb-acetate (250, 500, or 1,000 ppm for 30
days in drinking water to PND60 male Wistar rats, resulting in blood Pb levels of -20,
35, and 42 (ig/dL, respectively) significantly decreased the elapsed time required to
develop the first myoclonic jerk and tonic-clonic seizure (Arrieta et al.. 2005); also, the
dose of pentylenetetrazol (PTZ) required to induce seizures was significantly decreased
across all Pb dose groups. Some studies showed no effect of Pb on kindled animals
(Schwark et al.. 1985; Alfano and Petit. 1981). Other studies showed differential
susceptibility to convulsant-inducing agents in developmentally Pb-exposed rats (Chen
and Chan. 2002). Sprague Dawley rats were exposed to Pb-acetate (0.2% w/v in drinking
water from PND1-25, followed by 25 days with no Pb exposure). Seizures were induced
at PND25 or PND50. At PND25, Pb exposure significantly decreased (PTZ)-, picrotoxin
(PIC)-, and strychnine (STRY)-induced convulsion thresholds, but increased N-methyl-
D-aspartate (NMDA) and 4-aminopyridine (4-AP)-induced convulsion thresholds. At
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PND50, the effects of PTZ, NMDA, and 4-AP remained similar to those at PND25, but
PIC and STRY-dependent convulsion thresholds were reversed and significantly
increased (Chen and Chan. 2002). Chen and Chan ("2002) hypothesized that this
differential effect may be due to selective effects on inhibitory and excitatory
neurotransmission as an effect of age and blood Pb level.
Recent investigation expands on the work by Arrieta et al. (2005) by showing similar
effects in another rodent species, BALB/c mice. Adult male BALB/c mice were exposed
to Pb for 30 days via drinking water (resulting in blood Pb levels in control and 50, 100,
200 and 400 ppm Pb groups of 0.02, 6, 11, 15 and 18 (ig/dL, respectively) (Mesdaghinia
et al.. 2010). Exposure to 50 ppm Pb did not affect PTZ-induced seizure threshold, but all
other doses significantly reduced the thresholds of face and forelimb clonus, myoclonic
twitch, running and bouncing clonus, and tonic hindlimb extension. These studies show
that Pb exposure may modulate seizure activity in animals.
5.3.7 Neurodegenerative Diseases
5.3.7.1 Epidemiologic Studies of Neurodegenerative Diseases in
Adults
The 2006 Pb AQCD described several studies examining associations of blood and bone
Pb levels with neurodegenerative diseases such as Alzheimer's disease and dementia.
Among NAS men, higher bone Pb levels were associated with lower MMSE scores
(Weisskopf et al. 2004; Wright et al.. 2003b). which indirectly pointed to a potential
association with dementia, given that the MMSE is widely used as a screening tool for
dementia. Overall, studies had sufficient limitations (e.g., indirect assessment of
dementia, comparison of Alzheimer's Disease patients and healthy controls, lack of Pb
biomarker data), and findings were inconclusive (U.S. EPA. 2006b). New studies on
dementia are not available to assess further the associations with Pb biomarkers.
Similarly, new studies examining Alzheimer's disease are not available, and as in 2006,
the evidence is still inconclusive regarding the association with Pb biomarkers. In
contrast, there has been additional investigation of ALS, Parkinson's disease (PD), and
essential tremor, which is described below.
Amyotrophic Lateral Sclerosis
Most studies of the association between Pb and ALS have relied on indirect methods of
assessing Pb exposure and overall, have produced inconsistent results. Case-control
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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 et
al.. 1997). Another case-control study 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 (ig/dL (e.g., OR: 14.3 [95% CI: 3.0, 69.3] for blood Pb
levels 3-4 (ig/dL) (Kamel et al.. 2002). In analyses of tibia or patella Pb tertiles, subjects
in the highest two tertiles (>10 jj.g/g patella Pb and > 8 jj.g/g tibia Pb) had higher ALS
prevalence. For example, the ORs (95% CIs) for patella Pb levels 10-20 jj.g/g and tibia Pb
levels 8-14 jj.g/g were 2.1 (0.6, 7.4) and 1.6 (0.5, 5.6), respectively. 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 that the association could be the result of
reverse causality since the half-life of blood Pb is only about 30 days, and blood was
collected from people who already had ALS. If, for example, reduced physical activity
among those with ALS led to more bone turnover, then more Pb would be released from
bones into circulation leading to elevations in blood Pb levels among cases as a result of
effects of the disease.
Since the 2006 Pb AQCD, a few additional studies have been conducted with the same
New England-area case-control study population. Kamel et al. (2005) reported that the
association between blood Pb level and ALS was not modified by the ALAD genotype
(Kamel et al.. 2005). Another report examined survival of ALS among 100 of the original
110 ALS cases (Kamel et al.. 2008). Higher tibia Pb levels were associated with longer
survival time. Findings were similar for patella and blood Pb levels, although they were
associated with smaller increases in survival time. These paradoxical findings raise the
concern that in a case-control study of ALS, the association between bone Pb levels and
ALS may be biased because the case group may comprise more individuals with longer
survival time. Consequently, their bone Pb levels may be higher because they reflect a
longer period of cumulative exposure. On the other hand, the decreased mobility due to
the disease itself would tend to increase bone resorption and lower bone Pb levels over
time. This process might mitigate this effect for bone Pb but would tend to increase blood
Pb levels among cases. However, this was not observed in the one study that had bone
and blood Pb biomarkers (Kamel et al. 2002). Because the strongest findings for survival
were found for tibia Pb, it is unlikely that the findings were biased due to increased
survival of cases.
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
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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 blood Pb levels was associated with ALS with an OR (95%
CI) of 2.6 (1.9, 3.7). Associations did not differ substantially by indicators of bone
turnover but were 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 driving the association between blood Pb level and ALS.
Whether other types of reverse causality are contributing, however, cannot be ruled out.
This study did not have measures of bone Pb and therefore could not assess the
association with biomarkers of cumulative Pb exposure.
In summary, studies have found associations of blood and bone Pb levels with ALS but in
relatively few different cohorts. The case-control design of most studies and issues of
reverse causality and bias due to survival time make it difficult to draw firm conclusions.
Studies in additional cohorts using designs other than case-control comparisons are
needed to address the limitations of the available studies and characterize better the
potential effects of Pb on ALS.
Parkinson's Disease
A few previous studies, some ecological (Rvbicki et al.. 1993; Aauiloniiis and Hartvig.
1986) and some case-control relying on questionnaire data or occupational history
(Gulson et al.. 1999; Gorell et al.. 1997; Tanner et al.. 1989) indicated associations
between exposure to heavy metals, particularly Pb, and risk of PD. Available evidence
was limited and far from conclusive. A recent large case-control study (330 cases, 308
controls) recently reported on associations between biomarkers of Pb and PD in a
population with virtually no occupational exposures to Pb (Weisskopf et al.. 2010).
Subjects in the highest quartile of tibia Pb level (>16.0 jj.g/g) had higher odds of PD
compared to those in the lowest quartile (< 5 jj.g/g) (OR: 1.91 [95% CI: 1.01, 3.60]). In
this study, cases and controls were recruited from several different sources including
movement disorder clinics and community-based cohorts, which could have introduced
some biases. However, when analyses were restricted to cases recruited from movement
disorder clinics and to their spouse, in-law, or friend as controls, the results were even
stronger (OR: 3.21 [95% CI: 1.17, 8.83]). 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.
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Manganese exposure has been associated with Parkinsonian symptoms and could
potentially confound associations between Pb and PD. Weisskopf et al. (2010) did not
adjust for Mn exposure in analyses. However, unlike a setting of occupational exposure
to Pb, general environment exposure to Pb is much less likely to be correlated with
environmental Mn exposure. Thus, the likelihood of any associations with Pb being
confounded by co-occurring Mn exposure is less likely.
Coon et al. (2006) conducted a smaller case-control study of 121 PD patients and 414
age-, sex-, and race-, frequency-matched controls all receiving health care services from
the Henry Ford Health System. Subjects in the highest quartile of both tibia (OR: 1.62
[95% CI: 0.83, 3.17] for levels > 15 jag/g) and calcaneus (OR: 1.50 [95% CI: 0.75, 3.00]
for levels > 25.29 jj.g/g) bone Pb levels had higher odds of PD compared to those in the
lowest quartiles (0-5.91 jj.g/g for tibia and 0-11.70 jj.g/g for calcaneus). The highest OR
for PD was estimated for subjects in the highest quartile of whole-body lifetime exposure
to Pb (> 80.81 jj.g/g) estimated using PBPK modeling), compared to the lowest quartile of
exposure (0-40.04 jj.g/g) (OR: 2.27 [95% CI: 1.13, 4.55] for levels > 80.81 j-ig/g)- These
analyses did not adjust for Mn either; however, in this study it was not clear what the
extent of occupational exposure to Pb was among the participants. Thus, it is uncertain
whether the observed associations were specifically related to Pb exposure or could have
resulted from co-occurring Mn exposure.
In summary, a small number of recent studies expand on previous evidence by finding
associations of bone Pb levels, biomarkers of cumulative Pb exposure, with PD in adults.
Nonetheless, additional investigation is warranted to establish the temporality between Pb
exposure and development of PD and to assess potential confounding by Mn exposure.
Essential Tremor
In a relatively small body of literature, concurrent blood Pb levels have been consistently
associated with essential tremor, although studies have had relatively small sample sizes
and have produced imprecise effect estimates. The 2006 Pb AQCD 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 Louis et al. (2005). the magnitude of association was larger among carriers of
an ALAD2 allele than among adults with only ALAD1 alleles.
Since 2006, Dogu et al. (2007) reported on a case-control study of 105 essential tremor
cases from a movement disorder clinic in Turkey and 105 controls (69 spouses and 36
other relatives living in the same district). After adjusting for age, sex, education,
cigarette smoking, cigarette pack-years, and alcohol use, a 1 (ig/dL higher blood Pb level
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(measured at the time of study recruitment) was associated with essential tremor with an
OR (95% CI) of 4.19 (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, but past
occupational history was not reported. Occupational Pb exposures were less likely in the
New York area population.
In this small body of studies, associations were observed between blood Pb level and
essential tremor in adults. However, due to the case-control design of studies, temporality
between exposure and development of essential tremor cannot be established. Further, the
level, timing, frequency, and duration of Pb exposure associated with PD is uncertain as
all studies examined blood Pb level at the time of study recruitment. Occupational
histories were not reported in these studies. Thus, it is not clear what past exposure to Mn
may have contributed to the associations observed with blood Pb levels.
5.3.7.2 Toxicological Studies of Neurodegenerative Disease
Although epidemiologic studies have provided weak evidence for associations of blood
Pb level and ALS in adults, toxicological studies have found that Pb exposure induces
neurophysiologic changes consistent with ALS. For example, chronic Pb exposure
(Pb-acetate in drinking water at 200 ppm from weaning onward, resulting blood Pb level:
27 (ig/dL) reduced astrocyte reactivity and induced increased survival time in the
superoxide dismutase transgenic (SOD1 Tg) mouse model of severe ALS (Barbeito et al..
2010). In this model, Pb exposure did not significantly increase the onset of the ALS
disease but increased survival time in SOD1 Tg mice (Barbeito et al.. 2010). This finding
does provide biological plausibility for the association observed between blood Pb level
and longer survival time in patients diagnosed with ALS (Kamel et al.. 2008). In mice,
astrocyte of vascular endothelial growth factor (VEGF) also was examined to understand
its possible contribution to Pb effects on increasing survival time in ALS models.
Baseline levels VEGF were elevated in astrocytes from the ventral spinal cord of
untreated SOD1 Tg mice versus untreated nontransgenic animals. VEGF was not induced
in the astrocytes of Pb-treated nontransgenic mice. Pb-exposed SOD1 Tg mice had
significant elevations of astrocyte VEGF versus vehicle-treated SOD 1 Tg animals
(Barbeito et al. 2010). Nontransgenic animals exposed to Pb showed no elevation in
VEGF expression above that in nontransgenic vehicle-treated animals (Barbeito et al.
2010). Other research has suggested that ALS initiation is dependent on motor neuron
function and ALS progression is dependent on astrocyte and microglia function
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(Y amanaka et al.. 2008; Boil lee et al.. 2006). Thus, the aforementioned findings for Pb-
induced effects on astrocytes provide a mechanistic explanation for Pb effects on ALS
progression in animals.
Others have 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, Barbeito et al. (2010) found that
an up-regulation of VEGF production by astrocytes in the Pb-exposed SOD1 Tg mice
was protective against motor neuron death in the SOD1 Tg cells (Barbeito et al.. 2010).
Chronic Pb exposure in a mouse model of ALS was associated with increased survival
time and was correlated with higher spinal cord VEGF levels, making astrocytes less
cytotoxic to surrounding motor neurons (Barbeito et al.. 2010). Also, in another study the
metal chelators DP-109 and DP-460 were neuroprotective in the ALS mouse model or
Tg(SODl-G93A) (Petri et al.. 2007).
Improper activation of microglia and release of inflammatory cytokines and metabolites
can contribute to neurodegeneration (Zhang et al.. 2010b; Oian and Flood. 2008). These
two cell types are known to accumulate or sequester Pb in the nervous system.
Researchers have implicated dysfunctional astrocytes as playing an important role in the
chain of misregulated inflammation leading to neurodegenerative conditions (Barbeito et
al.. 2010; De Kevser et al.. 2008).
Cell Death Pathways
Earlier work has documented that Pb exposure can induce cell death or apoptosis in
various models including rat brain (Tavakoli-Nezhad et al.. 2001). retinal rod cells (He et
al.. 2003; He et al.. 2000). cerebellar neurons (Oberto et al.. 1996). and PC 12 cells
(Sharif! and Mousavi. 2008). These observations indicate that Pb-induced cell apoptosis
may mediate its neurodegenerative effects. A recent study reported that chronic (40 days)
Pb exposure induced hippocampal apoptosis in young (exposure starting at 2-4 weeks of
age) and adult (exposure starting at 12-14 weeks of age) male rats exposed to 500 ppm Pb
by drinking water (resulting in blood Pb levels of 98 (ig/dL); apoptosis was verified by
light and electron microscopy, and increased pro-apoptotic Bax protein levels (Sharif! et
al.. 2010). Another study followed the developmental profile of changes in various
apoptotic factors in specific brain regions of animals exposed to Pb-acetate (0.2% dam
drinking water) during lactation. Male offspring blood Pb level at the end of lactation or
PND20 was 80 (ig/dL. The data showed that hippocampal mRNA for various apoptotic
factors including caspase-3, Bcl-x and Brain-derived neurotrophic factor (BDNF) was
significantly upregulated on PND12, PND15 and PND20. The cortex of these male pups
also showed upregulation of Bcl-x and BDNF on PND 15 and PND20 (C'hao et al.
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2007). The cerebellum did not have elevated apoptotic mRNA levels in this model. This
study showed temporal and regional changes in activation of death protein message levels
in male offspring.
Pb exposure has also been shown to induce apoptosis during spinal cord development in
chicks exposed to 150 or 450 (ig in ovo Pb-acetate exposure at embryonic day 3 or 5 and
visualized six days later; TUNEL positive cells were at significantly higher levels in Pb-
exposed animals and were visualized in all layers of the developing spinal cords (Miiller
et al.. In Press). 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. Another recent study
examined Pb treatment animals (i.p. injections of 350 mg/kg, twice daily, blood Pb levels
at PND14 of 0.15 (ig/dL in controls and 8.10 (ig/dL in Pb-treated animals) during
PND7-14, a known period of sensitivity to NMDA antagonist-dependent neuronal
apoptosis (Ikonomidou et al.. 1999). Liu et al. (2010b) examined apoptotic effects in 30
day-old male rats that were exposed to Pb-acetate lx/d for 6 weeks via intragastric
infusion. Four dose groups: control, low (2), medium (20), and high (200 mg/kg BW) had
blood Pb levels of 1.0 to 7.5 (ig/dL; 4.5 to 11 (ig/dL; 9 to 42 (ig/dL; and 48 to 73 (ig/dL,
respectively. Liu et al. (2010b) reported Pb-induced apoptosis in the brain with
hippocampal XIAP (statistically significant at high dose only) and Smac (statistically
nonsignificant trend) downregulation and associated histopathology showing
hippocampal neuronal apoptosis (TUNEL positive staining, significant at all doses) 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 induced regional-specific changes in
brain apoptotic proteins poly(ADP-ribose) polymerase, Bcl-2, caspase-3) with a greater
effect observed in the hippocampus and cerebellum and a lesser effect observed in the
brainstem and the frontal cortex (Kiran Kumar et al.. 2009)
Collectively, studies have shown that Pb exposure induces neuronal apoptosis in animals
during various developmental windows: early postnatal and adulthood. The new data
continue to show that Pb exposure induced apoptosis in brains of animals.
Lead-Induced Neuronal Plaque Formation
Epidemiologic studies have not provided compelling evidence that Pb exposure is
associated with Alzheimer's Disease in adults (Section 5.3.7.1); however, Pb exposure in
early life has been shown to promote Alzheimer's Disease-like pathologies in the brains
of aged adult animals. Alzheimer's disease is characterized by amyloid-beta peptide (Ab)
accumulation, hyper-phosphorylation of the tau protein, neuronal death and synaptic loss.
In the last decade, the developmental origins of adult health and disease (DoHAD)
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paradigm and the similar Barker hypothesis have indicated that early life exposures can
result in aberrant adult outcomes. Bolin et al. (2006) demonstrated the connection
between developmental exposure to Pb in the rat with early life programming and the
resulting inflammation-associated DNA damage with neurodegenerative loss in the adult
brain. Wu and colleagues (2008a) had similar findings in a study using infantile exposure
to Pb in monkeys. The investigations reinforce the need to directly examine the long-term
effects of developmental exposure to toxicants rather than relying on adult exposure
alone to predict probable health risks from prenatal, neonatal or juvenile exposure
(Dietert and Piepenbrink. 2006). Mechanistically, some of these pathologies have been
associated with changes in the epigenome.
The fetal basis of amyloidogenesis has been examined extensively by the Zawia
laboratory in both rodents and nonhuman primates. Mechanistically, amyloid plaques
originate from the cleavage of the amyloid precursor protein (APP) to Ab, which
comprises the plaque. In rodents exposed to Pb as neonates or as adults, neonatal Pb
exposure induced amyloidogenesis in the aged animal brains; adult exposure to Pb did
not contribute to plaque formation. Basha et al. (2005) exposed male rodents neonatally
via lactation to Pb (PND1-PND20 exposure, dam drinking water Pb-acetate 200 ppm,
resulting in pup PND20 blood Pb level of 46 (ig/dL and cortex 0.41 jj.g/g wet weight of
tissue) and examined cortical APP gene expression over the lifetime. A bimodal response
was observed, with a significant increase in APP expression above that in control animals
first manifesting neonatally and second manifesting in old age (82 weeks of age) (Basha
et al.. 2005). A concomitant bimodal response was observed in specificity protein 1
(Spl), a transcription factor known to be related to APP expression. Ab, the amyloid
plaque constituent, was also significantly elevated in these aged animals developmentally
exposed to Pb. A subset of rodents exposed to Pb only as aged adults (18-20 weeks of
age) was unresponsive in APP or Spl expression or Ab production after Pb exposure,
indicating the developmental lifestage and not adult lifestage as the susceptible period for
Pb-induced amyloidogenesis. The Zawia lab (Wu et al.. 2008a) produced similar findings
for amyloid plaques in the brains of monkeys that were exposed to Pb as infants
(PND1-PND400), i.e., 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 (1992. 1990). After weaning but with continued Pb exposure, the monkeys
had blood Pb levels of 19-26 (ig/dL. However, in old age when amyloid plaques had
manifested, blood Pb levels and brain cortex Pb levels had returned to control or baseline
levels. Together, the rodent and nonhuman primate toxicological studies concur and show
that developmental Pb exposure induced elevations in neuronal plaque proteins in aged
animals.
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Mechanistic understanding of Ab production and elimination after Pb exposure was
explored in human SH-SY5Y neuroblastoma cells exposed to Pb concentrations of 0, 5,
10, 20, and 50 (.iM for 48 hours. These studies showed that Pb affected two separate
pathways to contribute to elevated 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 examined by Behl et al. (2009). The choroid
plexus is capable of removing beta-amyloid peptides from the brain extracellular matrix;
however, Pb was shown to impair this function, possibly via the metalloendopeptidase,
insulin-degrading enzyme (IDE), which metabolizes Ab (Behl et al.. 2009). In another
study, the effect of Pb on transcription factors essential in the regulation of the
developing brain was explored. Pb exposure has been shown to perturb DNA binding of
transcription factors including SP1 at essential sites like zinc finger proteins. In Long-
Evans hooded rat pups exposed to Pb during lactation, these Pb-induced developmental
perturbations of SP1 DNA binding were found to be ameliorated by exogenous zinc
supplementation (Basha etal.. 2003).
An additional study with developmental Pb exposure (gestational plus lactational, dam
drinking water solutions of 0.1%, 0.5% or 1%, blood Pb level 40, 80 and 1,00 (ig/dL)
showed that the hippocampus contained neurofibrillary changes as early as PND21.
These changes manifested with Tau hyper-phosphorylation, and increased tau and beta
amyloid hippocampal protein levels in Pb-exposed offspring (Li et al.. 2010b).
In summary, the multiple recent studies showed that developmental Pb exposure induced
significant increases in neuronal plaque associated proteins, which is the pathology found
in humans with Alzheimer's disease. Adult exposure to Pb did not generate this
neurofibrillary pathology, further demonstrating that early life Pb is a sensitive window
for Pb-induced pathology including Ab-peptide accumulation, activation of Ab-
supporting transcription factors, as well as tau hyperphosphorylation. Epidemiologic
evidence for Pb-associated Alzhemier's disease is weak, and although several study
design limitations were noted (Section 5.3.7.1), the animal evidence indicates that
epidemiologic studies assessing concurrent bone or brain Pb levels or occupational Pb
exposure may not have examined the etiologically relevant exposure period. Notably,
animals were not behaviorally assessed for dementia.
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5.3.8
Modes of Action for Lead Nervous System Effects
5.3.8.1 Effects on Brain Physiology and Activity
A small body of available epidemiologic evidence with small sample sizes demonstrated
associations of Pb biomarkers with electrophysiologic changes in the brains of young
adults as assessed by magnetic resonance imaging (MRI) or spectroscopy (MRS) (Yuan
et al.. 2006; Cecil et al. 2005; Meng et al.. 2005; Trope et al. 2001). By characterizing
underlying mechanisms by which Pb exposure may disrupt brain function, these studies
have provided biologically plausible evidence for the effects of Pb exposure on cognitive
and psychological and behavioral function observed in adults.
Findings such as lower levels of N-acetylaspartate (NAA), creatine (Cr), and choline are
linked to decreased neuronal density and alteration in myelin. Notably, Trope et al.
(2001) and Meng et al. (2005) reported that all subjects had normal MRIs with no
evidence of structural abnormalities. Thus, the clinical significance of the observed
physiological changes was unclear. Additionally, these studies compared subjects with
relatively high childhood blood levels (23-65 j^ig/dL) to those with childhood blood Pb
levels <10 (ig/dL. Therefore, it is unclear whether physiological changes would be
observed in association with lower blood Pb levels. Cecil et al. (2005) and Yuan et al.
(2006) conducted functional MRI in 42 adult (ages 20-23 years) participants from the
CLS cohort during a verb generation language task and found that childhood average
blood Pb level was associated with decreased activation in the left frontal gyrus and left
middle temporal gyrus, regions traditionally associated with semantic language function.
Although these findings were in adults, they were consistent with findings in the same
cohort of subjects that indicated associations of blood Pb level with other indices of
language skills in childhood.
Since the 2006 Pb AQCD, studies examining MRI data were largely limited to CLS
cohort participants as adults (ages 19-24), and recent results continue to support
associations of childhood blood Pb levels with physiological changes in the brain of
adults. These recent studies expanded on previous studies by including larger sample
sizes, aiming to characterize important lifestages of Pb exposures, and evaluating
potential links between changes in brain activity and functional neurodevelopmental
deficits. Whereas previous CLS analyses focused on activity in specific regions of the
brain, Cecil et al. (2011) examined brain metabolites in 159 subjects and found that
childhood average blood Pb levels were associated with lower levels of NAA and Cr in
the basal ganglia and lower levels of choline in white matter. These results were adjusted
for age and FSIQ. A recent analysis of 31 men in the NAS cohort similarly reported an
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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, which 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 CLS young adults found that childhood average blood Pb levels were
associated with altered brain architecture. Brubaker et al. (2009) reported associations of
childhood average blood Pb levels with diffusion parameters that were indicative of less
organization of fibers throughout white matter. Results were adjusted for potential
confounding by variables such as 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 on different neural elements may be related to the stage of myelination
development at various periods of exposure.
Another study of 157 CLS participants provided evidence of region-specific reductions in
adult gray matter volume in association with childhood blood Pb levels (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 traditionally related to
executive functions, mood regulation, and decision-making). Further, investigators found
that fine motor factor scores 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 MRI changes associated with blood Pb levels
may be indicative of decrements in cognitive 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 with changes in cognitive function
(e.g., visuoconstruction, visual memory, eye-hand coordination) (Schwartz et al.. 2007)
and behavior (impulsivity, aggression, violence) (Yang et al.. 2005; Raine et al.. 2000).
In a subsequent comparison of blood Pb levels measured at various lifestages, Brubaker
et al. (2010) found that blood Pb levels at older ages were associated with greater losses
in gray matter volume than were childhood average or maximum blood Pb levels. Both
Cecil et al. (2008) and Brubaker et al. (2010) found that Pb-associated reductions in gray
matter were more pronounced in CLS males than females.
Studies of Pb-workers also found associations of blood and bone Pb levels with changes
in brain structure and physiology, adding support for 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..
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2009b; Jiang et al.. 2008; Bleecker et al.. 2007a; Stewart et al. 2006). 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.. 2007a').
In summary, results in a few different populations indicated associations of blood or bone
Pb levels with changes in brain structure and physiology as assessed by MRI or MRS. A
majority of evidence was derived from the CLS study. Several studies linked these
changes to functional changes in cognitive performance or motor skills. While the
collective findings pointed to the effects of long-term or past Pb exposures, it is important
to recognize that other lifestages of exposure, including recent adult, were not widely
examined.
5.3.8.2 Oxidative Stress
Pb has been shown to induce oxidative stress in multiple animal models, and this
oxidative stress can contribute to DNA damage, which can be measured with the adduct
8-hydroxy-2'-deoxyguanosine (8-oxo-dG). The contribution of reactive oxygen or
nitrogen species to these Pb induced changes was assayed by examining the ratio of
8-oxo-dG to 2-deoxyguanosine (2-dG). 2-dG is product of oxidative cleavage that can
generate 8-oxo-dG from a parent compound forming a DNA adduct during conditions of
nitrosative or oxidative stress. The 8-oxo-dG to 2-dG ratio data from rodent male
offspring were similar to the amyloid data with significant biphasic elevations in
developmentally Pb-exposed animals (0.2% Pb-acetate in dam drinking water from
PND1-20) versus control, nonPb-exposed animals at early (PND5) and late life time
points (80 weeks of age) (Bolin et al. 2006). Activity of the base-excision DNA repair
enzyme oxoguanine glycosylase or Oggl was unaffected by Pb exposure (Bolin et al.
2006). Interestingly, similar findings were reported in a monkey study. The ratio of
8-oxo-dG to 2-dG in the brains of aged monkeys (23 years) after being exposed to Pb as
infants, was significantly elevated above that in controls (Wu et al.. 2008a). Several lines
of evidence indicate that oxidative stress is involved in neurodegenerative pathologies
including Alzheimer's disease; hydrogen peroxide-induced oxidative stress has been
shown to induce intracellular accumulation of amyloid beta-protein (Abeta) in human
neuroblastoma cells (Misonou et al. 2000). Similar to the amyloid findings in Pb-
exposed animals the oxidative stress markers showed no significant changes above
baseline when animals were exposed to Pb only as aged adults (Wu et al.. 2008a; Bolin et
al.. 2006). Thus, the data for biomarkers of oxidative stress concur with the
amyloidogenesis data with both demonstrating kinetically similar biphasic significant
elevations in markers of oxidative stress and amyloidogenesis with early life Pb exposure
and an absence of effect with adult only exposure.
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Because the brain has the highest energy demand and metabolism of any organ, energy
homeostasis is of utmost importance. Pb has been shown to inhibit various enzymes
involved in energy production or glucose metabolism including glyceraldehydes-3
phosphate dehydrogenase, hexokinase, pyruvate kinase, and succinate dehydrogenase
(Verma etal.. 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 (delta omega sub m) and the
functioning of the mitochondria.
To characterize the effect of Pb exposure on these mitochondrial parameters, brains from
the offspring of Pb-exposed mice were collected at PND8 (Baranowska-Bosiacka et al..
201 lb). Cerebellar granular cells were harvested from PND 8 control and Pb-exposed
animals (0.1% Pb-acetate in dam drinking water, resulting in blood Pb levels of 4 (ig/dL
and cerebella Pb levels of 7.2 jj.g/g dry weight). These neuronal cells were cultured for 5
days in vitro, at which point various mitochondrial parameters were measured. With Pb
exposure, reactive oxygen species 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. Pb-exposed crayfish that were placed under hypoxic conditions
adapted to the situation by decreasing their metabolism (Morris et al. 2005) and showed
mitochondrial changes consistent with those observed in granule cells (Baranowska-
Bosiacka et al.. 201 lb). These studies showed impaired mitochondrial function and
energy production in neuronal cells from mice with gestational and lactational Pb
exposure with concomitant increases in mitochondrial and cellular ROS production.
The effects of co-administration of flaxseed oil (FSO) with Pb-acetate on oxidative stress
and neurotoxicity were examined in adult male Wistar albino rats (Abdel Moneim et al.).
Animals were administered Pb-acetate i.p. for 5 days (20mg/kg, resulting in blood Pb
level ~31 |_ig/dL the day after the last Pb injection) or FSO (oral gavage lOOOmg/kg body
weight for 5 days, 1 hour prior to Pb dosing, resulting in blood Pb level -12 (ig/dL).
Administration of the polyunsaturated fatty acid FSO significantly attenuated the blood
Pb level of Pb-exposed animals and control animals, indicating that FSO may alter Pb
toxicokinetics in animals. As would be expected with a lower blood Pb level, FSO+Pb
exposed animals had significant attenuations in Pb-induced histological neuronal damage
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and brain tissue DNA fragmentation. Pb exposure effects on indicators of oxidative stress
included decreased antioxidant GSH levels, elevated lipid peroxidation, and decreased
antioxidant enzyme activity in brain tissue. These Pb-induced effects were attenuated in
FSO+Pb exposed animals. These results indicated that FSO likely attenuates Pb-related
neurotoxicological damage and oxidative stress via its action on altering Pb
toxicokinetics.
Green tea extract (GTE) was shown to be protective against Pb-induced oxidative stress
in the brains of adult male rats (1.5% GTE +/- 0.4% Pb-acetate for 6 weeks) (Hamed et
al.. 2010). GTE+Pb-exposed animals had significantly less brain Pb than did Pb-exposed
animals (1.2 versus 1.9 ppm). Levels of pro-oxidants or antioxidants including lipid
peroxides (LPO), nitric oxides (NO), total antioxidant capacity (TAC), glutathione
(GSH), glutathione-S-transferase (GST), and superoxide dismutase (SOD) were
measured in brain tissue homogenate 24 hours after the termination of Pb exposure.
Comparing the Pb-exposed animals to controls, brain LPO was significantly elevated and
brain GSH, GST and NO were significantly decreased. GTE and Pb co-exposure
significantly attenuated the Pb-related changes. This study reported a positive correlation
between whole blood Pb levels and brain tissue LPO levels and a negative correlation
between whole blood Pb and NO levels. The antioxidant quercetin (Que) rescued chronic
Pb exposure-related (0.2% Pb-acetate in drinking water from birth to PND67, 30 mg/kg
BW Que for one week from PND60-67) impaired synaptic plasticity in adult male and
female Wistar rat dentate gyrus (DG) (Hu et al. 2008a). Pb-related impaired long-term
potentiation, paired-pulse reactions, and input/output functions were significantly
attenuated with Que treatment at PND67. Que treated animals had significantly less
hippocampal Pb than did the Pb-exposed animals and this decreased Pb brain burden
likely contributed to the attenuated response observed with Que exposure.
The vulnerability of the highly energetic brain tissue to stressors and cell death can be
exacerbated with an energy imbalance. Baranowska-Bosiacka et al. (201 lb) showed that
Pb exposure induced energy imbalances in immature rat brains exposed to Pb. Wistar rat
pups (PND15) of both sexes were injected daily for 2 weeks with Pb-acetate (15mg/kg
BW, i.p., resulting in blood Pb levels of 3 and 30 (ig/dL, control and Pb-exposed,
respectively) and thereafter sacrificed for analysis of regional brain purines and
purinergic receptors. 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, enhanced expression of the proinflammatory P2XR receptor was
observed in the glial fraction, indicating the astrocyte pool may be involved in the
pathological changes found in Pb-exposed immature rat brains.
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5.3.8.3 Nitrosative Signaling and Nitrosative Stress
The nitric oxide system is increasingly being recognized as a signaling system in addition
to its more classical role as a marker of cellular stress. In studies of learning and memory
using the Morris water maze, hippocampal changes in NO were noted in Pb-exposed
animals after completion of the test. Pb exposure has been repeatedly shown to increase
the escape latency in Pb-exposed animals (Section 5.3.2.2). Chetty (2001) initially
reported decreased hippocampal neuronal nitric oxide synthase (NOS) with perinatal Pb
exposure. Namely, with repeated swim tests, control animals more quickly found a
submerged platform (i.e., escaped), than did Pb-exposed animals. After either 4 or
8 weeks of Pb exposure to weanling male rats (resulting in blood Pb level of 6.2 (ig/dL),
hippocampal NOS and NO were significantly decreased. 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 B1 decreased hippocampal NOS. Supplementation also
changed Pb-related effects on hippocampal NO. Pb-induced NO increased with taurine
and decreased with vitamin C, tyrosine or glycine. Dietary supplementation with tyrosine,
methionine, or ascorbic acid after 4 weeks of Pb exposure in weanling males (4-week
blood Pb level of 47.6 (ig/dL and 8-week blood Pb level of 8.1 (ig/dL), induced
significant increases in NO. Zinc supplementation in this model had no effect on the NO
system. The investigators concluded various combinations of nutrients significantly
attenuate Pb-related decreases in NO/NOS. Specifically, nutrients prevented (8 weeks Pb
plus concomitant exposure to methionine, zinc, ascorbic acid, and glycine) or restored
(4 weeks Pb exposure followed by 4 weeks nutrient exposure, taurine and thiamine) Pb-
related decrements in NO/NOS concentrations (Fan et al. 2009a).
5.3.8.4 Synaptic Changes
Previous toxicological studies point to an effect of developmental Pb exposure on
synapse development, which mechanistically may contribute to multiple Pb-related
aberrant effects, including changes in long-term potentiation (LTP) and facilitation.
Earlier work has shown that developmental Pb exposure results in altered density of
dendritic hippocampal spines (Kiralv and Jones. 1982; Petit and LcBoutillier. 1979).
aberrant synapse elimination (I ohmnnn and Bonhoeffer. 2008). and abnormal long-term
and short-term plasticity (MacDonald et al. 2006). Newer research using the Drosophila
larval neuromuscular junction model has shown 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
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reduced activity of the plasma membrane calcium 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 low-dose body burden
(100 (.iM) of Pb calculated to be 13-48 (.iM per larvae. 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. 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
calcium 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.
A study by Li et al. (2009c) focused on inflammatory endpoints and synaptic changes
after gestational plus lactational dam drinking water Pb exposure (solutions of 0.1%,
0.5% or 1%, resulting in offspring blood Pb levels of 40, 80 and 100 (ig/dL, respectively
at PND 21). 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 (Li et al.. 2009c'). 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.
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. In heme synthesis, Pb is known to substitute for the
cation zinc in another nucleotidase, pyrimidine 5'-nucleotidase, and thus, the nucleotidase
is used as a biomarker of Pb exposure. Acute exposure (96 hours) of male and female
zebrafish to Pb-acetate (2 (ig/dL) in their water induced significant decreases in ATP
hydrolysis in brain tissue. 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 lb). 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, NTPDasel and 5'-nucleotidase,
were unchanged (Senger et al.. 2006). Thus, Pb has been shown to affect nucleotidase
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activity in the central nervous system of zebrafish, possibly contributing to aberrant
neurotransmission.
Another enzyme important in synaptic transmission at cholinergic junctions in the CNS
and at neuromuscular junctions peripherally is acetylcholinesterase (AChE). After 24
hours of exposure to Pb-acetate (2 (ig/dL water), AChE activity was significantly
inhibited in zebrafish brain tissue. In Pb-exposed fish, AChE activity returned to baseline
by 96 hours and maintained baseline activity after chronic exposure of 30 days. Thus, Pb
was shown also to affect synaptic homeostasis of AChE in the brains of zebrafish
(Richetti et al.. 2010V
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 were exposed to Pb (10 or 100 (.iM solutions in
media) during the period of synaptogenesis (Neal et al.. 2010a'). This exposure induced
the loss of two proteins necessary for presynaptic vesicular release, synaptophysin (Syn)
and synaptobrevin (Syb), without affecting a similar protein synaptotagmin (Syt). This
deficit was found in both GABAergic and glutamatergic neurons. Pb also induced an
increase in number of presynaptic contact sites. But, these sites may be nonfunctional as
they lack the protein receptor complexes necessary for proper vesicular exocytosis.
Another factor involved in growth and signaling of presynaptic neurons is BDNF, which
is synthesized and released by postsynaptic neurons. BDNF is regulated by the NMDAR
and acts in a retrograde fashion, participating in presynaptic maturation. In 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 presynaptic effects. Thus, this cell culture model showed that
Pb-related presynaptic aberrations are controlled by NMDAR-dependent BDNF effects
on synaptic transmission.
Animals exposed to Pb postnatally (Wistar Albino rats, drinking water 300 mg/L
Pb-acetate, resulting in blood Pb levels of 17 (ig/dL at 6 weeks of age), from birth
through lactation and through 12 weeks of age showed decreased learning ability and
decreased hippocampal glutamate at 6, 8, 10 and 12 weeks of age fNiu et al.. 2009) as
well as significant decrements in the hippocampal glutamate synthesis-related enzymes
aspartate aminotransferase and alanine aminotransferase.
5.3.8.5 Blood Brain Barrier
Two barrier systems exist in the body to separate the brain or the central nervous system
from the blood. These two barriers are the blood brain barrier (BBB) and the blood
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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 this barrier.
Pb exposure during various developmental windows has been shown to increase the
permeability of the BBB (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.. 1988V Studies
reviewed in earlier Pb AQCDs have shown that the chemical form of Pb and its
capability to interact with proteins and other blood components affects its capability to
penetrate the BBB (U.S. EPA. 2006b'). Pb also has been shown to compromise the
function of the BCB and decrease the CSF level of transthyretin, a thyroid binding
protein made in the choroid plexus. The choroid plexus and cerebral endothelial cells that
form the BBB and BCS tight junctions have been shown to accumulate Pb more than
other cell types and regions of the CNS do. Hypothyroid status can contribute to impaired
learning and IQ deficits (Lazarus. 2005).
Recent research with 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. Cerebral
vasculature leakiness was ameliorated or resembled controls after iron supplementation.
These weanlings also had significant Pb-induced decreases in the BBB tight junction
protein occludin in the hippocampus, brain cortex, and cerebellum that were rescued to
control levels with iron supplementation (Wang et al.. 2007b'). These data demonstrate
that Pb induced a leaky BBB in weanling rats with associated decreases in the junctional
protein occludin; dietary supplementation with iron ameliorated these Pb-induced
impairments of the BBB in male rats. 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 ZOl
and Z02 (Balbuena et al.. 2011). Because expression of these junctional and scaffold
proteins did not show decrements, it was determined that these protein decrements were
due to post-translational modifications.
A study examined the effects of Pb on transendothelial electrical resistance (TEER), a
marker of BBBB integrity, in an in vitro co-culture system employing endothelial cells
(RBE4 or bovine brain microvascular endothelial cells) and astrocytes (primary Sprague-
Dawley neonatal pup astrocytes, in utero day 21) as the barrier between Pb containing
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exposure media and neurons. After 14 hours of exposure to Pb, TEER was significantly
impaired in a concentration-dependent manner, with the lowest significant effects found
with 1 (.iM Pb-acetate exposure. Thus, Pb exposure was found to contribute to leakiness
of the BBB by decreasing the resistance across the junction.
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 diffiisivity that changes under pathological conditions like
cerebral edema. After 4 weeks of exposure to 500 ppm Pb, the water ADC was
significantly increased in the hippocampus, mesencephalic reticular formation, and
cerebellum but unaffected in other brain areas. After 12 weeks of 500 ppm Pb exposure,
ADC was significantly increased in the corpus callosum and caudate putamen. Exposure
to 50 ppm Pb for 12 weeks increased the ADC values in the cerebellum and
mesencephalic reticular formation. The brain areas with elevated ADC also showed
increased BBB permeability as measured with evans blue albumin complex. These data
show that adult animals with chronic Pb exposure are susceptible to regional edema and
regional increased permeability of the BBB, even with low blood Pb levels (12 (ig/dL)
(Lopez-Larrubia and Cauli. 201IV
5.3.8.6 Cell Adhesion Molecules
Classic cell adhesion molecules including NCAM and the cadherins are junctional or cell
surface proteins that are critical for cell recognition and adhesion. Cell adhesion
molecules, particularly the cadherins, are calcium-dependent and thus interaction from
competing cations like Pb can potentially contribute to nervous system barrier function
disruption, tissue development dysregulation, immune dysfunction, and affect learning
and memory (Prozialeck et al.. 2002).
5.3.8.7 Glial Effects
Astroglia and oligodendroglia are supporting cells in the nervous system that maintain the
extracellular space in the brain and provide support and nutrition to neurons via nutrient
transport, structural support to neurons, and myelination. Glial cells provide immune
surveillance in the brain or contribute to inflammation-mediated pathologies. In the
central nervous system, Pb treatment of Wistar rats 15 mg/kg of Pb-acetate, i.p.) during
early postnatal maturation was observed to produce chronic glial activation with
coexisting features of inflammation and neurodegeneration (Struzvnska et al.. 2007).
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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 (Tiffanv-Castiglioni et al.. 1989) by sequestering Pb. This glial sequestration of Pb
has been shown to decrease brain glutamine concentrations at doses of 0.25 ± 1.0 (.iM
Pb-acetate via Pb-related reduction in glutamine synthetase activity in the astroglia;
astroglia take up glutamate after its release and convert it to glutamine. Pb has been
shown to induce hypomyelination and demyelination (Coria et al.. 1984) mediated
through the oligodendrocytes with younger animals being more susceptible to the effects
of Pb (Tiffanv-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
To determine the contribution of the gliotransmitter serine to Pb-mediated changes in
long-term potentiation (LTP), Sun et al. (2007) exposed pups to Pb-acetate in utero,
lactationally, through PND28 via 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 LTPs, but the magnitude ofLTDs was
restored with supplementation with D-serine (Sun et al.. 2007). which is known to be
regulated by the NMDAR (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 to LTP by D-serine. NMDAR-independent
LTP hippocampal neurotransmission, which was examined in slices of Pb-exposed
mossy-CA3 synapses, was not rescued by exogenous D-serine supplementation. These
data indicate that glial transmission is affected with Pb exposure and that the NMDAR
may also be involved in this aberrant glial transmission.
5.3.8.8 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.
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Monoamine Neurotransmitters and Stress
Combined exposures of maternal stress and Pb exposure can synergistically enhance
behavioral and neurotoxic outcomes in offspring of exposed animals and can sometimes
potentiate an effect that would otherwise be sub-threshold. Virgolini et al. (2008a) found
that CNS effects of developmental Pb exposure (50 or 150 ppm via drinking water,
2 months prior to mating through lactation, resulting in blood Pb levels of 11 (ig/dL and
35 (ig/dL, respectively) were enhanced with combined maternal and offspring stress.
Offspring neurotransmitter concentrations were significantly affected with Pb exposure,
but the most interesting findings were those of potentiated effects, i.e., effects that were
not observed with Pb exposure alone or stress alone. These potentiated effects were only
observed when Pb was combined with stress (maternal [MS] and/or offspring stress
[OS]). Potentiation of serotonin (5HT) levels in females was significant in the frontal
cortex in females and in the nucleus accumbens (NAC) in the male offspring (50 and
150 ppm Pb drinking water exposure) (Corv-Slechta et al. 2009). Regional 5HT levels
were unaffected in offspring with Pb exposure alone. The concentration
of 5-Hydroxyindoleacetic acid (5HIAA), the main metabolite of 5HT, was significantly
increased with Pb exposure alone in the striatum of male offspring with 150 ppb Pb
exposure alone; with the remaining Pb-stress exposure combinations, Pb plus stress
potentiated striatal and frontal cortex 5HIAA in males. Potentiated 5HIAA levels in
females were significant in the NAC at both Pb doses; stress alone also significantly
increased 5HIAA levels in 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
The monoamine neurotransmitters include DA, 5HT, and norepinephrine (NE). Earlier
work has shown that perinatal Pb exposure of rats induced increased tyrosine
hydroxylase, increased DA and increased cerebral cortex catecholamine
neurotransmission (Devi et al.. 2005; Leret et al.. 2002; Bielarczvk et al.. 1996). Earlier
publications detailing important time windows of exposure, duration of exposure, and
dose of Pb indicated varying effects on monoamine transmitters. In more recent work,
these neurotransmitters, among others, have been implicated in Pb effects on auditory
function in the brainstem in various integration centers there 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. Low-level Pb exposure has been associated
with altered processing of auditory temporal signals in animal studies (Lurie et al. 2006;
Finkelstein et al.. 1998). Blood Pb levels for control, very low Pb (VLPb) and low Pb
(LPb) exposure groups were 1.4, 8.0, and 42.2 (ig/dL, respectively. Developmental Pb
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exposure from the formation of breeding pairs to PND21, which is at the end of auditory
development in the mouse, led to significant decreases in immunostaining of LSO and
SOC brainstem sections for monoamine vesicular transporter VMAT2, and for 5HT and
dopamine beta-hydroxylase (DbH), a marker for NE. Statistically significant alterations
in VMAT2 and DbH were found with both VLPb and LPb exposure; however,
decrements in 5HT were statistically significant only with VLPb. Immunostaining for TH
and transporters including VGLUT1, VGAT, VAChAT indicated that they were
unaffected by developmental Pb exposure. These data provide evidence that specific
regions of the brainstem involved in auditory integration are affected by developmental
Pb exposure via effects on the monoamine neurotransmitter system (Fortune and Lurie.
2009). The Pb-induced effects on the auditory portion of the brainstem at the level of the
monoamine system provide possible mechanistic understanding of the animal data
showing Pb-induced impaired auditory processing.
Dopamine
The 2006 Pb AQCD detailed evidence for low-dose Pb-related decreased dopaminergic
cell activity in the substantia nigra and ventral segmental areas. Earlier studies with
moderate- to high-dose postnatal or adult Pb exposure have reported changes in DA
metabolism, as indicated by changes in DA and DOPAC, a DA metabolite. Given these
findings, a recent study measured DA and DOPAC in various brain regions of year-old
male rodents to examine if GLE affected DA metabolism. Low- and high-dose GLE in
male rodents induced significant elevations in DOPAC concentration and the DOPAC to
DA ratio in the forebrain. In the forebrain, DA was significantly decreased in low-dose
GLE males and significantly elevated in high-dose GLE males compared to controls. In
the striatum, DOPAC was significantly elevated with both low- and high-dose GLE
males, but DA concentration was only significantly elevated in high-dose GLE males.
The striatum ratio of DOPAC to DA was not significantly different from that in controls.
These new data expand upon the monoamine literature base in which indicates low
concentration perinatal Pb exposure of rats was found to induce increased sensitivity of
the dopamine receptors (D2 and D3) (Gedeon et al.. 2001; Corv-Slechta et al.. 1992).
produce higher DA levels (Devi et al.. 2005; Leret et al.. 2002). and enhance
catecholamine neurotransmission in the cerebral cortex, cerebellum, and hippocampus
(Devi et al. 2005).
The interaction of DA and the nitric oxide 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 et al.. 1980). 7-nitroinidazole (7-NI), a selective inhibitor of nNOS,
enhanced amphetamine-evoked DA release in the rat striatum (Nowak et al.. 2008).
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Prenatal Pb exposure attenuated the facilitatory effect of 7-NI on DA release in the
striatum. This interaction is ROS-independent; using spin trap measurements, there were
no significant concentration changes in hydroxyl radical with Pb exposure (Nowak et al..
2008). Thus, the neuronal NO system appears to be involved in specific aspects of Pb-
related dopaminergic changes.
Dopamine and Vision
In various experimental animal models, the loss of retinal DA or zinc is associated with
abnormal rod-mediated scotopic ERGs. These effects may explain observations of Pb-
associated retinal effects (Rothenberg et al. 2002a; Lilienthal etal.. 1994; Lilienthal et
al.. 1988; Alexander and Fishman. 1984). In the human and animal toxicological
literature, prenatal maternal blood Pb level (humans), gestational + early postnatal Pb
exposure (rodents, low and moderate dose Pb only), or gestational continuing to lifetime
(non-human primate) Pb exposure has been associated with increased amplitude
(supernormality) of ERGs (Rothenberg et al.. 2002a; Lilienthal etal.. 1994; Lilienthal et
al.. 1988). In the animal toxicological literature, subnormality of the ERGs has been
observed with postnatal Pb exposure (Fox et al. 2008; Otto and Fox. 1993; Fox et al..
1991; Fox and Farber. 1988) and high-dose developmental (prenatal + early postnatal,
rodent) Pb exposure (Fox et al.. 2008). Producing results consistent with observations in
humans (Rothenberg et al. 2002a). Fox et al. (2008) showed that low- (LPb) and
moderate-level (MPb) gestational+early postnatal Pb exposure in the rat, a period
equivalent to the human gestational retinal development period, produced supernormal
retinal ERGs. In children, supernormal ERGs were associated with prenatal Pb maternal
blood Pb levels >10.5 (ig/dL (Rothenberg et al. 2002a). The animal data provide
mechanistic information that may begin to explain these supernormal ERGs that are seen
in children and rodents, i.e., significant increases in retinal neurogenesis and significant
decreases in retinal DA use and dopamine turnover (DOPAC:DA ratio). High-dose
developmental (rodent, gestational+early postnatal to PND10) Pb exposure (HPb)
produced significant subnormal retinal ERGs. Subnormal ERGs also were found in
occupationally-exposed adults (Otto and Fox. 1993; Guguchkova. 1972). Female rats
were exposed to Pb-acetate in drinking water from 2 weeks prior to mating throughout
gestation and lactation until PND 10, a period of developmental exposure that is
equivalent to gestational exposure in humans. Peak blood Pb levels in the offspring at
PND 1-10 were 12, 24, and 46 (ig/dL in the LPb, MPb, and HPb groups, respectively. LPb
and MPb gestational exposure induced increased cellularity or retinal thickness in the
outer nuclear layer, inner nuclear layer and total retina (Lcasure et al.. 2008). In
summary, the retina is affected by low-dose Pb exposure and gestational Pb exposure, as
indicated by concentration-dependent decreases in DA use and turnover. Inverted U-
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shaped Pb concentration-response relationship were reported for retinal endpoints
including ERG and retinal thickness.
NMDA
NMDA receptors (NMDAR) have been shown to contribute to synaptic plasticity, and Pb
exposure at different developmental stages has been shown to contribute to aberrations in
LTP or long term depression (LTD) in the hippocampus via reduced NMDA current,
among other mechanisms (Liu et al.. 2004). The 2006 Pb AQCD indicated that Pb
induced decreases in stimulated glutamate release that affected LTP. Further, the
relationships between Pb exposure and decreased magnitude and increased 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 work examining dietary supplement use, found that Pb-
related decreases in the gene expression and protein level of NMDAR subunit NR1 in
weanling male rats were rescued with methioninecholine co-exposure (Fan et al.. 2010).
Fan et al. (2010) found that Pb-related suppression of the NMDAR subunits NR2A and
NR2B was not rescued with methioninecholine treatment. 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 Raianna. 2006).
Studies continue to show that Pb exposure affects neurogenesis or proliferation of new
cells in the hippocampus. Earlier work by Schneider et al. ("2005) showed that postnatal
Pb exposure (30-35 days starting at PND25, 1,500 ppm Pb-acetate in chow, resulting in
blood Pb level of 20 j^ig/dL) of male Lewis rats induced significant decrements in BrdU
incorporation (proliferation) at PND50-55. Recent publications affirm this original
finding with different sex of animals, dosing and exposure time windows. Postnatal Pb
exposure to Wistar rat pups (0.2% Pb-acetate from PND1-30, 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 dentate
gyrus at PND80 (Fox et al.. 2010) (Figure 5-18). In another study, developmental/lifetime
Pb exposure (1,500 ppm Pb-acetate in chow from 10 days before mating to termination of
experiment at PND50, resulting in blood Pb levels of 0.8 (ig/dL in controls and 26 (ig/dL
in Pb-exposed animals) to female Long-Evans rats induced significant decrements in
hippocampal granule cell neurogenesis or proliferation of new cells in adult rats (Verina
et al.. 2007). outcomes that affect LTP, spatial learning, neuronal outgrowth, and possibly
mood disorders such as schizophrenia. NMDAR mediates the integration of new neurons
into existing neuronal pathways in the adult hippocampal DG, which is important to
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learning and memory. Lifetime Pb exposure (dam Pb-acetate chow exposure 10 days
prior to mating through pregnancy to PND50 or PND78) induced significant decrements
in hippocampal granule cell neurogenesis or proliferation of new cells in adult rats. Also,
Pb-exposed animals had significant decreases in brain volume in the stratum oriens (SO)
region of the hippocampus, specifically significant decreases 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 neuron in the outer portion of the DG. These findings show
that exposure to environmentally-relevant doses of Pb induced significant aberrations in
adult hippocampus granule cell neurogenesis and morphology, providing mechanistic
explanations for Pb-induced neuronal aberrations. Guilarte et al. (2003) demonstrated that
Pb exposure of rats from an enriched environment was associated with reduced learning
impairment, increased expression of hippocampal NMD A receptor subunit 1, and
increased induction of brain derived neurotrophic factor mRNA (Guilarte et al.. 2003).
Glutamate Receptor
Glutamate receptors including the ionotropic NMDAR and the metabotropic glutamate
receptors (mGluR) are known targets of Pb toxicity with recent findings showing a role
for mGluR5 in learning and memory. In vitro (GD18 fetal rat cultures, 100 (.iM. 1 (.iM,
0.01 (.iM PbCl2 in culture media) and in vivo studies (gestational and lactational
Pb-acetate exposure; control, 0.05, 0.2, or 0.5% in dam drinking water, with respective
weanling blood Pb levels of 3, 18, 57, 186 (ig/dL) showed that Pb exposure induced
mGluR5 mRNA and protein decrements in a concentration-dependent manner (Xu et al..
2009c).. The Pb-related attenuation of mGlu5 expression may contribute to the effect of
Pb on LTP and LTD.
5.3.8.9 Neurite Outgrowth
The 2006 Pb AQCD reported that Pb decreased neurite outgrowth at 20|ag/dL and noted
that Pb interfered with neurite outgrowth via protein kinase mediated pathways
(MAPK/ERK); earlier work had documented decreased primary DA neuron outgrowth
with 0.001 (.iM Pb exposure (Lidskv and Schneider. 2004). Recent studies have shown
that exposure of dams to low-dose Pb (resulting in blood Pb level of 4 (ig/dL)
significantly decreased pup hippocampal neurite outgrowth (pup blood Pb level:
12 (ig/dL) and reduced the expression of hippocampal polysialylated neural cell adhesion
molecule (PSA-NCAM), NCAM, and sialytransferase (Hu et al. 2008b). PSA-NCAM is
transiently expressed in newly formed neurons during the period of neurite outgrowth
from embryrogenesis until the early postnatal period and is down-regulated in the adults
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except in areas known to exhibit synaptic plasticity (Seki and Arai. 1993). NCAM is
important for memory formation, plasticity and synapse formation, and early-life Pb
exposure was found to affects its expression in laboratory rats.
5.3.8.10 Epigenetics
Many investigators are beginning to show that environmental chemical exposures are
associated with epigenetic changes. Air pollution exposure is being linked increasingly
with epigenetic changes (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 are mediated by histone modification, DNA methylation, miRNA changes, 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.
Monozygotic twins are often used to study epigenetic changes, and a recent study
identified differential methylation of a specific locus in 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 Pb-
exposed monkeys (aged animals, blood Pb) and mouse brains (fetal cells exposed to Pb in
culture, 0.1 (.iM Pb) (Wu et al.. 2008b). Changes in DNA methyltransferases (Dnmtl,
Dnmt3a) were noted in control primate brains as they aged and these changes were
further exacerbated by Pb exposure (Bihaqi et al.. 2011). Another enzyme involved in
DNA methylation, methyl CpG binding protein 2 MECP2, showed a similar trend as the
Dnmts. Profiles of the histone modifying gene H34mc2 increases with age in control
animals. This age-related increase is significantly attenuated in Pb-exposed animals. The
cerebral cortex tissue used in this experiment was obtained from female primates who
had received 1.5 mg/kg • day Pb-acetate via diet from birth until 400 days of age
(resulting in blood Pb levels 19-26 (ig/dL at age 400 days) (Rice. 1990).
Methyltransferases catalyze biological methylation reactions and are dependent on the
cofactor S-adenosyl methionine (SAM) for this transfer to acceptor molecules. SAM
exposure after gestational and lactational Pb exposure (dam Pb-acetate exposure to
1,500 ppm Pb-acetate followed by 20-22 days of daily 20 mg/kg BW SAM exposure)
improved hippocampal LTP and Morris water maze performance at PND 44-54. Thus,
the impaired cognition and synaptic plasticity induced by developmental Pb exposure
were attenuated with SAM treatment (Cao et al.. 2008).
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5.3.8.11 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 (Ademuviwa 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. Mechanistically, Pb
exposure has been shown to depress the activity of cholesterol-7-a-hydroxylase, an
enzyme involved in bile acid biosynthesis (koiima et al.. 2005); bile acids are the route
by which cholesterol is eliminated from the body. Pb exposure produced significant
increases in brain triglycerides with an 83% increase at 300 ppm and a 108% increase at
400 ppm. At 200 ppm, Pb exposure induced a statistically nonsignificant decrease in
brain triglycerides. Pb exposure across all three dose groups 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 (Abe et al.. 2007). 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 at 0, 200, 300, and 400 ppm
were 7, 41, 61, and 39 (ig/dL, respectively. In summary, based on limited examination,
Pb exposure significantly increased brain cholesterol, triglycerides, and phospholipids as
well as significantly increased plasma free fatty acids. These effects were sometimes
more prominent at lower doses of Pb. Future characterization of molecular and cellular
pathways affected by Pb exposure may bring insight to these Pb-related changes in
phospholipidosis and cholesterogenesis.
5.3.9 Lifestage of Lead Exposure and Neurodevelopmental Deficits
Environmental exposures during critical lifestages can affect key physiological systems
that orchestrate plasticity (Feinberg. 2007). Exposure to environmental toxins during
prenatal and/or early postnatal development may alter the normal course of
morphogenesis and maturation that occurs in utero and early in life, resulting in changes
that affect structure or function of the central nervous system via altered neuronal growth
and/or synaptogenesis/pruning structure (Rice and Barone. 2000; Landrigan et al.. 1999).
Synaptic pruning, which is active throughout early childhood (ages 1-4 years), may
underlie the elevated risk of young children to environmental exposures. MRI studies
have provided understanding of brain development in normally developing children and
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adults, ages 3-30 years (Giedd et al. 2009; Lenroot and Giedd. 2006). Total cerebral
volume was found to peak at age 10.5 and 14.5 years in females and males, respectively.
The volume of the cerebellum was found to peak 2 years after the cerebral volume peaks.
Lateral ventricular volume showed the most inter-individual variation and was found to
increase throughout childhood and adolescence. White matter volume generally increased
throughout childhood and adolescence. Gray matter volume and associated structures
show inverted U-shaped developmental trajectories with peak volumes found in late
childhood or early adolescence. Females generally had gray matter volume peaks 1 to
3 years earlier than did males. Thus, the anatomical development of the brain during
childhood and adolescence is found to be a dynamic process with variation by age, brain
region, and sex. Observations that brain development is active throughout childhood and
in adolescence indicates that Pb exposure may affect neurodevelopment throughout this
period.
The elevated risk of Pb-associated neurodevelopmental deficits in children is well
supported by findings in animals that prenatal and postweaning Pb exposure alters brain
development via changes in synaptic architecture (Section 5.3.8.4) and neuronal
outgrowth (Section 5.3.8.9) and leads to impairments in memory and learning
(Section 5.3.2.2) and emotional and depressive changes postnatally (Section 5.3.3.5).
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 (Baver. 1989).
Hence, when normal development is altered, the early effects may persist into adult life
even in the absence of current exposure, magnifying the public health impact. Supporting
evidence is provided by a few available new toxicological studies that find that Pb
exposure during neonatal development but not in adulthood leads to neurodegenerative
amyloid plaque formation in the brains of aged rodents and monkeys (Section 5.3.7.2).
With repeated assessments of children prenatally to later childhood and early adulthood,
the prospective cohort studies have aimed to distinguish among neurodevelopmental
effects associated with blood Pb levels measured at different periods of development. 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 that the risk of neurodevelopmental decrements is greatest
for blood Pb levels measured at a particular lifestage. In these studies, the identification
of developmental periods when children are at increased risk of Pb-associated
neurodevelopmental decrements has been complicated by the high degree of correlation
in the blood Pb levels of children over time and the confounding of age and peak blood
Pb levels (Lanphcar et al.. 2005; Dietrich et al.. 1993b; Needleman et al.. 1990).
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As described in detail in the 2006 Pb AQCD, several studies with varying lengths of
follow-up demonstrated associations of prenatal blood Pb levels (maternal and umbilical
cord) with neurodevelopmental deficits throughout childhood and into early adulthood
(U.S. EPA. 2006b). Pb exposure during the prenatal lifestage may be associated with
increased risk of neurodevelopmental effects not only because of the nervous system
developmental processes that are active as described above but also because of factors
that result in elevated Pb exposures. Substantial fetal Pb exposure may occur from
mobilization of maternal skeletal Pb stores that may be related to past Pb exposures
(Gulson et al.. 2003; Hu and Hernandez-Avila. 2002). Pb can cross the placenta to affect
the developing fetal nervous system (Rabinowitz. 1988). Maternal and cord blood Pb
levels generally have been shown to be highly correlated, indicating that blood Pb levels
in a newborn infant reflects that of the mother (Schell et al.. 2003).
Prenatal blood Pb levels (maternal and cord) were consistently associated with cognitive
function decrements and behavioral problems assessed between infancy and age 3 years
(Table 5-4 and Table 5-13). Among studies that had blood Pb measurements at both
lifestages, some found stronger associations for prenatal blood Pb levels (Hu et al.. 2006;
Bellinger et al.. 1984). and other found stronger associations for concurrent blood Pb
levels (Wasserman et al.. 1998; Wasserman et al.. 1992). Studies that found associations
with concurrent blood Pb levels also tended to find associations with prenatal cord or
maternal blood Pb levels. Thus, both postnatal child and maternal Pb exposures may
contribute to lower cognitive function in young children. Several studies found that
prenatal or neonatal blood Pb levels were associated with neurodevelopmental
decrements assessed neonatally (within 30 days) or early in infancy (within 3 months),
which indicated that relatively short durations of Pb exposure were associated with
neurodevelopmental decrements (Shen et al.. 1998; Rothenberg etal.. 1989; Dietrich et
al.. 1987a; Emhart et al.. 1986V In the studies of neurodevelopmental effects in infancy,
prenatal blood Pb level may be serving as a surrogate of postnatal blood Pb levels as both
are expected to be highly correlated.
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Table 5-13 Associations of cognitive function and behavioral outcomes with
blood Pb levels measured at different lifestagesa
Study
Population/Location
Blood Pb Levels
(Hg/dL)
Statistical Analysis
Outcome
Effect Estimate
(96% Cl)b
Assessments in children up to age 3 years
Rothenberg et al.
(1989)
42 children followed
prenatally to child age 30
days
Mexico City, Mexico
Maternal week 36
gestation mean (SD):
15.0(6.4)
Maternal at birth mean
(SD): 15.5(5.7)
Regression model adjusted
for smoking, single mother,
problems in pregnancy,
alcohol use in previous
month, use of spinal block,
gravidity, income
Self-quieting ability
(regulation of state)
at age 30 days
Assessed using
Newborn Brazelton
Assessment System
Prenatal:-0.091 (-0.18, 0)
Dietrich et al. (1986
) 305 children followed
prenatally to age 6 mo.
in Cincinnati, OH
Prenatal (maternal)
mean (SD): 8.0 (3.8)
Concurrent mean (SD):
5.9 (3.4)
Log linear regression model
adjusted for birth weight,
gestation, sex
Bayley MDI assessed
at age 6 mo
Prenatal:-0.6 (-1.1,-0.09)
Concurrent: -0.23 (-0.58, 0.12)
Bellinger etal.
(1987)
249 children followed from
birth (1979-1981) to age
36 mo
Boston area, MA
Prenatal (cord blood)
mean (SD): 6.6 (3.2)
Regression and longitudinal
analyses adjusted for the
mother's age, race, IQ,
education, number of years
of cigarette smoking, number
of alcoholic drinks per week
in the third trimester, mean
family social class over the
period of the study, quality of
the care-giving environment,
infant's sex, birth weight,
gestational age, birth order
Bayley MDI assessed
at age 6,12,18, 24
mo
Prenatal: -4.8 (-7.3, -2.3),
blood Pb levels > 15 pg/dLvs.
blood Pb levels <3C
Hu et al. (2006)
146 children born
1997-1999 followed
prenatally to age 24 mo
Mexico City, Mexico
Prenatal (maternal 1st
trimester) mean (range):
7.1 (1.5-43.6)
Early childhood (12 mo)
mean (SD): 5.2 (3.4)
Concurrent mean (SD):
4.8 (3.7)
Log linear regression model
adjusted for concurrent blood
Pb, sex, maternal age,
current weight, height-for-age
Z score, maternal IQ
Bayley MDI assessed
at age 24 mo
Prenatal 1st trimester: -4.1 (-
8.1, -0.17)
Prenatal (avg): -3.5 (-7.7,
0.63)
12 month:-2.4 (-6.2,1.49)
Concurrent: -1.0 (-3.9,1.9)
Gomaa et al. (2002)
197 children followed
prenatally to age 24 mo
Mexico City, Mexico
Prenatal (cord blood)
mean (SD): 6.7 (3.4)
Log linear regression model
adjusted for maternal IQ,
maternal age, sex, parental
education, marital status,
breastfeeding duration, child
hospitalization status
Bayley MDI assessed
at age 24 mo
Prenatal: -2.1 (-3.9, -0.39)
Wfesserman et al.
(1992)
392 children followed
prenatally to age 24 mo
Kosovo, Yugoslavia (K.
Mitrovica, Pristina)
Prenatal (cord blood)
mean (SD): 14.4(10.4)
Concurrent means:
K. Mitrovica: 35.4,
Pristina: 8.5
Log linear regression model
adjusted for sex, birth order,
birth weight, ethnic group,
HOME score, years of
maternal education, maternal
age, maternal intelligence
Bayley MDI assessed
at age 24 mo
Concurrent: 4.1 (-6.2, -2.0)
Prenatal: -3.2 (-7.2, 0.86)
Jedrychowski et al.
(2009b)
444 children born
2001-2004 followed
prenatally to age 36 mo
Krakow, Poland
Prenatal (cord blood)
geometric mean (range):
1.29 (0.44-5)
Linear regression model
adjusted for maternal
education, birth order,
prenatal ETS, sex
Bayley MDI assessed
at age 36 mo
Prenatal: -2.9 (-6.0, -0.76)
Cognitive function assessments at school age
Wfesserman et al.
(1994)
332 children followed
prenatally to age 3-4 yr
Kosovo, Yugoslavia (K.
Mitrovica, Pristina)
Prenatal (cord blood)
mean (SD): 14.4(10.4)
Concurrent means:
K. Mitrovica: 39.9
Pristina: 9.6
Log linear regression model
adjusted for HOME score,
maternal age, maternal
intelligence, maternal
education, language, birth
weight, sex
McCarthy GCI
assessed at age
3-4 yr
Concurrent: 4.1 (-6.2, -2.0)
Prenatal:-3.2 (-6.1,-1.2)
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Study
Population/Location
Blood Pb Levels
(Hg/dL)
Statistical Analysis
Outcome
Effect Estimate
(96% Cl)b
Dietrich et al. (1992)
259 followed from birth
(1979-1984) to age 5 yr
Cincinnati, OH
Prenatal (cord blood)
mean (SD): 8.2 (3.8)
Neonatal (10 days)
mean (SD): 4.8 (3.3)
Concurrent mean (SD):
11.9(6.4)
Linear regression model
adjusted for fetal distress and
growth, perinatal
complications, postnatal
indices of health and
nutritional status,
sociodemographic
characteristics, HOME score
Total FWS
assessed using
KABC at age 5 yr
Neonatal:-0.38, p< 0.01"
Prenatal:-0.26, p< 0.01"
Concurrent: -0.19, p < 0.01d
Lifetime avg: -0.16, p < 0.01'1
Bellinger etal.
(1991)
170 children followed from Early childhood (24 mo)
birth (1979-1981) to age mean (SD): 6.8 (6.3)
57 mo
Boston area, MA
Early childhood tooth
mean (SD): 2.8 (1.7)
Mg/g
Concurrent mean (SD):
6.4(4.1)
Log linear regression model
adjusted for family social
class, maternal IQ, marital
status, preschool attendance,
HOME score, out of home
care, number of residence
changes, recent medication
use, number of adults in
household, sex, race, birth
weight, birth order
McCarthy GCI
assessed at age 57
mo
Early childhood blood:
-3.0 (-6.7, -0.2)
Early childhood tooth: -2.5
(-10.2, 5.2)
Concurrent blood: -2.3 (-6.0,
1.4)
Dietrich et al.
(1993b)
245 children followed from
birth (1979-1984) to age 6
yr
Cincinnati, OH
Prenatal (cord blood)
mean (SD): 8.4 (3.8)
Neonatal (10 days)
mean (SD): 4.8 (3.1)
Concurrent mean (SD):
10.1 (5.6)
Linear regression model
adjusted for obstetric
complications, perinatal
status, sex, social class,
maternal intelligence, quality
of rearing environment,
earlier measures of
neurobehavioral status
B ru i n i n ks-Ose retsky
Test of Motor
Proficiency
assessed at age 6
yr
Concurrent:
-0.18 (-0.26, -0.10)
Neonatal:-0.15 (-0.33, 0.03)
Lifetime avg: -0.11
(-0.19, -0.03)
Prenatal:-0.04 (-0.20, 0.12)
Dietrich et al.
(1993b)
253 children followed from
birth (1979-1985) to age
6.5 yr
Cincinnati, OH
Prenatal (cord blood)
mean (SD): 8.3 (3.7)
Neonatal (10 days)
mean (SD): 5.0 (3.4)
Concurrent mean (SD):
11.8(6.3)
Linear regression model
adjusted for fetal distress and
growth, perinatal
complications, prenatal
maternal substance abuse,
postnatal indices of health
and nutritional status,
sociodemographic
characteristics, maternal IQ,
HOME score
FSIQ assessed
using WISC-R at
age 6.5 yr
Concurrent:
-0.33 (-0.60, -0.06)
Lifetime avg: -0.13 (-0.35, 0.0
Neonatal: -0.03 (-0.42, 0.36)
Prenatal: 0.15 (-0.26, 0.56)
Bag hurst et al.
494 children followed from
birth (1979-1982) to age
11-13 yr
Port Pirie, Australia
Prenatal mean of
second quartile: 7.4
Early childhood (2 yr)
mean of second quartile:
16.6
Lifetime avg mean of
second quartile: 15.7
Log linear regression model
adjusted for sex, birth weight,
birth order, feeding method,
breastfeeding duration,
parental education, maternal
age, parental smoking, SES,
quality of home environment,
maternal IQ, parents living
together
FSIQ assessed
using WISC-R at
age 7-8 yr
Early childhood:
-2.0 (-3.8, -0.21)
Lifetime avg: -1.6 (-3.7, 0.52)
Prenatal: 0.26 (-0.67,1.5)
Lanphear et al.
1,333 children pooled from
Boston, Cincinnati,
Cleveland, Mexico City,
Port Pirie, Rochester, and
Yugoslavia cohorts
Median (5th-95th)
Early childhood: 12.7
(4.0-34.5)
Peak: 18.0 (6.2-47.0)
Lifetime avg: 12.4
(4.1-34.8)
Concurrent: 9.7
(3.5-33.2)
Log linear regression model
adjusted for HOME score,
birth weight, maternal IQ,
maternal education
FSIQ measured at
ages 4.8-10 yr
Concurrent: -
0.23 (-0.32, -0.14)
Peak:-0.20 (-0.29, -0.11)
Lifetime avg:
-0.16 (-0.22,-0.09)
Early childhood:
-0.14 (-0.23, -0.06)
Pocock et al. (1994)
Meta-analysis of 5
prospective (over 1,100
children and 14 cross-
sectional studies (3,499
children)
Early childhood (2 yr)
range in means:
6.8-21.2
Meta-analysis of combining
effect estimates from
individual studies
FSIQ assessed
using various tests
at ages 5-10 yr
Early childhood:
-2.7 (-4.1,-1.2)
Postnatal mean: -1.3 (2.9, 0.37)
Around birth: 0.26 (-1.5, 2.0)
Prenatal: -3.9 (-6.6, -1.4)
Early childhood avg:
0.10 (-3.9, 4.1)
Later childhood avg: 0.17 (-1.4,
1.8)
Schnaas et al.
150 children followed from
prenatally (1987-1992) to
age 6-10 yr
Mexico City, Mexico
Prenatal (maternal 28-36 Log linear mixed effects
wkgestation): 7.8 (2.5,
24.6)
Early childhood avg (1-5
yr) mean (range): 9.8
(2.8-36.4)
Later childhood avg
(6-10 yr): 6.2 (2.2-18.6)
model adjusted for blood Pb
levels at other lifestages, sex,
birth weight, SES, maternal
IQ, First FSIQ measurement
FSIQ assessed
using WISC-R at
ages 6-1 Oyr
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Study
Blood Pb Levels
Population/Location	(|jg/dL)	Statistical Analysis	Outcome
Effect Estimate
(96% Cl)b
Bellinger etal.
148 children followed from
birth (1979-1981) to age
15-17 yr
Boston area, MA
Prenatal: NR
Early childhood (2 yr)
mean (SD): 6.5 (4.9)
Concurrent mean (SD):
2.9 (2.4)
Linear regression model
adjusted for HOME score
(age 10 and 5), child stress,
race, maternal IQ, SES, sex,
birth order, marital status
FSIQ assessed
using WISC-R at
age 10 yr
Early childhood:
-0.68 (-0.99,-0.17)
Prenatal: -0.48 (-5.7, 4.7), blood
Pb >10 |jg/dL vs. <3 pg/dLc
Concurrent: -0.46 (-1.5, 0.56)
Ris et al. (2004)
195 children in followed
from birth (1979-1985) to
age
15-17 yr
Cincinnati, OH
NR
Linear regression model
adjusted for maternal IQ, sex,
and average total HOME
Learning/IQ
composite
assessed using
WISC-III indices at
age 15-17 yr
Prenatal:-0.08 (-0.18, 0.03)
Early childhood, 6.5 yr:
-0.08 (-0.17, 0.003)
Early childhood avg:
-0.03 (-0.18, 0.03)
Behavioral assessments
Wfesserman et al.
379 children followed
prenatally to age 3 yr
Kosovo, Yugoslavia
(K. Mitrovica, Pristina)
Prenatal mean (SD):
16.1 (2.6)
Concurrent mean (SD):
25.8(19.1)
Hierarchical log linear
regression analyses adjusted
for town, sex, ethnicity,
maternal education, HOME
Anxiety/depression
assessed using
Child Behavior
Checklist at 3 yr
Concurrent: 1.46 (0.04,2.S
Prenatal: 1.16 (0.02,2.3)
Leviton et al. (1993)
1,923 children followed
from birth (1979-1980) to
age 8 yr
Boston area, MA
Prenatal blood 2nd
quartile: 4.8-6.3
Early childhood (tooth)
second quartile: 2.0-2.9
Mg/g
Log linear regression model
adjusted for single-parent
family, gestational age <37
wk, mother not a college
graduate, self-identification
as black, only child, daycare
during first 3 yr
Hyperactivity
assessed using
Boston Teacher
Questionnaire at
age 8 yr
Prenatal, girls: 0.26 (-0.69,1.13
Early childhood, girls:
0.10 (-0.92,1.1)
Bellinger etal.
(1994a)
1,782 children followed
from birth (1979-1980) to
age 8 yr
Boston area, MA
Prenatal (cord blood)
mean (SD): 6.8 (3.1)
Early childhood (tooth)
mean (SD): 3.4 (2.4)
ppm
Log linear regression
analyses adjusted for
prepregnant weight, race,
delivery by cesarean section,
marital status, paternal and
maternal education, sex, birth
weight, maternal smoking,
prenatal care beginning after
the first trimester, recipient of
public assistance, number of
children in family, child
currently on medication
Problem behaviors
(t-scores) assessed
using Teacher
Report Form of the
Child Behavior
Profile at age 8 yr
Early childhood:
1.8 (0.49, 3.1)
Prenatal:-0.31 (-1.7,1.07)
Ris et al. (2004)
195 children in followed
from birth (1979-1985) to
age 15-17 yr
Cincinnati, OH
NR
Linear regression model
adjusted for maternal IQ, sex,
and average total HOME
Inattention
composite
assessed using
Continuous
Performance Test
at 15-17 yr
Prenatal: 0.16 (0.04,0.27)
Early childhood, 6.6 yr:
0.12 (0.02,0.22)
Early childhood avg:
0.11 (0.03,0.19)
Dietrich et al. (2001)
195 children followed from
birth (born 1979-1985) to
age
15-17 yr
Cincinnati, OH
NR
Linear regression model
adjusted for birth weight,
HOME score, SES, parental
IQ
Parental report of
delinquent behavior
at 15-17 yr
Prenatal: 0.19 (0.02,0.37)
Early childhood, 6.5 yr:
0.13 (-0.01, 0.27)
Early childhood avg:
0.09 (-0.02, 0.20)
MDI = Mental Developmental Index, ETS = Environmental tobacco smoke, HOME = Home Observation for Measurement of the Environment, GCI =
General Cognitive Index, FWS = Filtered Word Test, KABC = Kaufman Assessment Battery of Children, FSIQ = Full-scale IQ, WISC = Wechsler Intelligence
Scale for Children, NR = Not reported
aStudies are organized by age of neurodevelopmental assessment. Effect estimates are presented in order of increasing magnitude, with statistically
significant results in bold.
bEffect estimates are standardized to a 1 |jg/dL increase in blood Pb level in analyses of blood Pb as a continuous variable.
"Effect estimate 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 provided in order to calculate 95% CI.
1	Prenatal and neonatal (10 days after birth) blood Pb levels also were associated with
2	cognitive function and behavioral outcomes in children examined at school-age (ages
3	4-17 years) (Table 5-13). Concurrent blood Pb levels generally were estimated to have
4	similar or larger magnitudes of effect (Wasserman et al.. 1998; Wasserman et al.. 1994;
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2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
32
33
34
35
36
37
38
Dietrich et al.. 1993a; Dietrich et al.. 1993b; Bellinger et al.. 1992; Dietrich et al.. 1992).
Studies conducted in the Cincinnati cohort examined diverse neurodevelopmental effects
and found that prenatal and neonatal blood Pb levels were associated with impairments in
behavior and auditory processing (Ris et al.. 2004; Dietrich et al.. 2001; Dietrich et al..
1992) but not cognitive function or motor function in children at age 6 years or in
adolescence (Bhattacharva et al.. 2006; Bhattacharva et al.. 1995; Dietrich et al.. 1993a;
Dietrich et al.. 1993b). These findings suggest that the effects of prenatal Pb exposure
may vary among different populations and specific endpoints.
Early childhood blood Pb levels were associated with diverse neurodevelopmental effects
assessed later in childhood and into early adulthood in both recent (Chandramouli et al..
2009; Min et al. 2009; Miranda et al.. 2009) and previous studies that did not compare
various life stages of Pb exposure (Yuan et al.. 2006; Cecil et al.. 2005; Tong et al.. 2000).
In a meta-analysis of results from five cohort studies (Pocock et al. 1994). a larger
decrease in FSIQ was estimated for an increase in peak (around age 2 years) blood Pb
level than for blood Pb level measured around birth or after age 2 years (Table 5-13).
This lag effect may be the result of a toxicological process in which some period of time
is required for past Pb exposure to affect CNS function. Alternatively, associations with
early childhood or peak blood Pb levels may reflect the greater reliability of
neurodevelopmental assessments at later ages when the processes modalities of children
are more highly differentiated. Early testing may lead to false negative results and fail to
identify a child who is at risk for later neurodevelopmental dysfunction. Further, due to
the correlation of blood Pb levels over time, it is difficult to assess whether early
childhood or peak blood Pb levels were serving as surrogates of concurrent or cumulative
blood Pb levels.
As presented in Table 5-13, several studies of school-aged children estimated larger
blood Pb-associated decreases in neurodevelopmental function for concurrent or lifetime
average (range: 5 to 13 years) blood Pb levels than for blood Pb levels at other lifestages.
These findings were substantiated in the analysis pooling data from seven prospective
studies, in which concurrent, peak, average lifetime (5- to 10-year average), and age
2 year blood Pb levels all were negatively associated with IQ measured between ages 5
and 10 years, with the larger magnitude of decrease associated with increases in
concurrent and peak blood Pb levels (Lanphear et al.. 2005). Childhood average blood Pb
levels (Lanphear et al.. 2005; Dietrich et al. 1993a; Dietrich et al.. 1993b) and deciduous
tooth Pb levels (Bellinger et al.. 1994a) have been associated with neurodevelopmental
effects. While the ages of children varied among these studies as did the years over which
blood Pb levels were averaged, the results nonetheless indicate cumulative Pb exposure
over multiple years may contribute to neurodevelopmental effects in children.
Associations with concurrent blood Pb level also were demonstrated consistently in
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1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
32
33
34
35
36
37
studies without comparisons to blood Pb levels during other lifestages (Figures and
Tables in Sections 5.3.2.1 and 5.3.3.1).
Some studies have aimed to improve the characterization of important lifestages of Pb
exposure by examining children in whom blood Pb levels are weakly correlated over time
(i.e., children whose blood Pb level ranking changed overtime) (Hornung et al.. 2009;
Schnaas et al.. 2006; Chen et al. 2005; Tong et al.. 1998; Bellinger et al.. 1990).
Collectively, the results did not conclusively demonstrate stronger findings for concurrent
blood Pb level. However, with the exception of Schnaas et al. (2006). the results did not
preclude an association with concurrent blood Pb level. Schnaas et al. (2006) followed
children in Mexico City prenatally through age 10 years and found prenatal blood Pb
levels (maternal blood at 28-36 weeks of pregnancy) to be weakly correlated with
repeated measures of blood Pb between ages 1 and 10 years (Pearson r < 0.25). In a
mixed effects model that included prenatal and multiple postnatal blood Pb measures,
only prenatal blood Pb level was associated with a decrement in FSIQ (Table 5-13).
Analysis of variance inflation factors indicated a lack of collinearity among the serial
blood Pb measures.
Tong et al. (1998) found that higher early-life blood Pb level was associated with a larger
deficit in IQ (Figure 5-21 and Table 5-14). As part of the Port Pirie, Australia cohort
study, investigators separately examined intellectual attainment in 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. In comparisons of tertiles of change in blood Pb level between
age 2 and 11-13 years, investigators found that FSIQ at ages 2, 4, 7, and 11-13 years did
not significantly differ between children with the largest declines (>16 (ig/dL) in blood
Pb level and children with a smaller decline (<10 (ig/dL). These findings indicated an
influence of higher blood Pb levels early in life despite declines in blood Pb with age and
a persistence of Pb effects. The results do not preclude an independent association with
concurrent blood Pb level.
In several different cohorts of U.S. children, larger decrements in cognitive function were
estimated for concurrent blood Pb levels with consideration of blood Pb levels measured
earlier in childhood (Hornung et al.. 2009; Chen et al.. 2005; Bellinger et al.. 1990)
(Figure 5-21 and Table 5-14). In the Boston cohort, Bellinger et al. (1990) found that at
age 57 months, FSIQ, as assessed by McCarthy GCI, was similar between children with
higher (>10 (ig/dL) and lower (<3 (ig/dL) prenatal cord blood Pb levels. Additionally,
higher concurrent blood Pb level (age 57 months) was associated with the largest decline
in GCI score overtime (score at age 57 months-score at age 24 months) among children
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1
2
3
4
5
6
7
8
9
10
with high prenatal blood Pb levels (>10 (ig/dL), which pointed to a larger decrease in
FSIQ in children with both high early and concurrent blood Pb levels (Figure 5-21 and
Table 5-14). The findings 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 positive home and caregiving environment (e.g., HOME score >52, higher SES,
higher maternal IQ) may also protect against decrements in cognitive function associated
with higher postnatal Pb exposures. Collectively, these results suggest that cognitive
development is not fixed early in childhood and can be affected negatively or positively
by postnatal influences.
Study
Hornung etal. (2009) 6yr.
Age of FSIQ
Assessment
Bellingeretal. (1990) 24and57 mo
Tong etal. (1998)
11 -13yr
Chen etal. (2005) 7yr.
Blood Pb (|jg/dL)
variable examined
0.5 ratio age 6 to 2 yr3
2.0 ratio age 6 to 2 yr3
Concurrent, prenatal <3b
Concurrent, prenatal 3-10b
Concurrent, prenatal a 10b
< 10.2 decline age 2 to 11-13
10.2-16.2 decline age 2 to 11-13
> 16.2 decline age 2 to 11-13
Low 2 yr (< 24.9), Low 7 yr (< 7.2)
Low 2 yr (< 24.9), High7yr(> 7.2)
High 2yr. (> 24.9), Low7yr (< 7.2)
High 2yr. (> 24.9),High 7yr(> 7.2)
<>
-e-
-1	-0.5	0
Change in FSIQ (95% CI)
aValues represent the ratio of blood Pb level at age 6 years to that at age 2 years.
bResults represent the decrease in FSIQ per 1 pg/dL increase in concurrent blood Pb level in children in different groups of prenatal
cord blood Pb levels. FSIQ scores were standardized to their standard deviation. Effect estimates in blue represent associations for
higher prenatal or early childhood blood Pb levels relative to concurrent blood Pb levels.
Note: Effect estimates represent associations between concurrent blood Pb level and cognitive function (standardized to standard
deviation) in children. Studies are presented in order of increasing prenatal level.
Figure 5-21 Associations of cognitive function in children with different
degrees of changes in blood Pb levels over time.
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Table 5-14 Additional characteristics and quantitative results for studies
presented in Figure 5-21

Population/
Blood Pb Levels


Effect Estimate
Study
Location
(Hg/dL)
Statistical Analysis
Outcome
(96% CI)8
Hornung et
462 children followed
Geometric mean
Linear regression model adjusted for
FSIQ assessed using
0.5 ratio of blood Pb level at age
al. (2009)
from birth
(5th-95th):
city, HOME score, birth weight,
WISC-R at age 6 yr
6 to age 2:0 (reference)

(1979-1984) to age 6
Peak: 13.6
maternal IQ, maternal education

2.0 ratio of blood Pb level at age

yr
(4.6-34.4)


6 to age 2 yr: -0.70 (-1.0, -0.40)

Rochester, NY and
Cincinnati, OH
Early childhood:
8.9 (3.0-23.8)
Lifetime mean: 8.5
(3.0-22.1)
Concurrent: 6.0
(1.9-17.9)



Bellinger et 170 children followed
al. (1990) prenatally to age 57
mo
Boston area, MA
NR
Log linear regression adjusted for
HOME score, social class, maternal
IQ, maternal age, sex, ethnicity
Change in McCarthy GCI
score (z-score) between
age 57 and 24 mo
For concurrent blood Pb level
Prenatal <3 pg/dL: -0.16
(-0.43, 0.11)
Prenatal 3-10 pg/dL: -0.14
(-0.57, 0.29)
Prenatal > 10 pg/dL: -0.46
(-0.81,-0.11)
Tong et al.
375 children followed Means: 21.2 (age Log linear regression model adjusted
from birth	2 yr), 7.9 (age for sex, birth weight, birth rank, feeding
(1979-1982) to age 11-13 yr)	style, breastfeeding duration, maternal
11-13 yr	IQ, maternal age, SES, HOME score,
Port Pirie Australia	parental smoking, parents living
together.
ANOVA to assess association of
change in IQ with change in blood Pb
across time intervals
Change in cognitive
function (z-scores) using
Bayiey MDI at age 2 yr,
McCarthy GCI at age 4 yr,
WISC-R at ages 7, and
11-13 yr
<10.2 pg/dL decline: 0.03
(-0.15, 0.21)c
10.2-16.2 pg/dL decline: 0.04
(-0.15, 0.23)c
>16.2 pg/dL decline: -0.01
(-0.20, 0.18)c
Low age 2, Low age 7: 0d
Low age 2, High age 7: -0.27
(-0.48, -0.05)
High age 2, Low age 7:0
(-0.21,0.20)
High age 2, High age 7:-0.28
(-0.47,-0.10)
Chen et al.
780 children
participating in the
TLC trial from age
12-33 mo to age 7 yr
Baltimore, MD;
Cincinnati, OH;
Newark, NJ;
Philadelphia, PA
Children underwent
chelation therapy
Mean (SD):
26.2
Age 2 yr
(5.1)
Age 5 yr
(5.2)
Age 7 yr:
Low age 2 yr:
<24.9
Low age 7 yr: <
6.2
12.0
8.0 (4.0)
Linear regression model adjusted for
city, race, sex, language, parental
education, parental employment, single
parent, age at blood Pb measurement,
caregiver IQ
WISC-III at age 7 yr
aEffect estimates represent the cognitive function score or change in score over time standardized to its standard deviation.
bEffects are estimated for concurrent blood Pb level (continuous variable) in children in different categories of prenatal blood Pb level: <3 |jg/dL, 3-10 |jg/dL,
and > 10 |jg/dL.
"Investigators estimated changes in IQ in groups of children with different degrees of decline in blood Pb levels over the study period: children with
<10.2 |jg/dL decline, children with a 10.2-16.2 |jg/dL decline, and children with >16.2 |jg/dL.
investigators compared IQs among children with different categories of blood Pb level early and later in childhood: low levels at age 2 (<11.4 |jg/dL) and
age 7 (<7.2 |jg/dL), low levels at age 2 (<11.4 |jg/dL) and high levels at age 7 (>7.2 |jg/dL), high levels at age 2 (>11.4 |jg/dL) and low levels at age 7
(<7.2 |jg/dL), and high levels at age 2 (>11.4 |jg/dL) and age 7 (>7.2 |jg/dL). Cutoffs were based on the median blood Pb levels.
1	Pooling the Cincinnati and Rochester cohorts (n = 397), Hornung et al. (2009) created a
2	new indicator of Pb exposure: the ratio of blood Pb level at 6 years of age to that at
3	2 years of age. The greatest decrease in cognitive and behavioral development was
4	observed for children with blood Pb ratios greater than 1 (indicating an increase in blood
5	Pb level from 2 to 6 years of age) (Figure 5-21, Figure 5-22, and Table 5-14). Presumably
6	areas under the curve would be similar among children with blood Pb level ratios of 1,
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greater than 1, and less than 1, indicating that cumulative blood Pb levels would not be
predictive.
As part of the multicenter TLC, Chen et al. (2005) also found higher concurrent blood Pb
level (> median 7.2 (ig/dL) to be associated with lower IQ 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). It is important to note that children participating in TLC had
undergone chelation therapy due to high blood Pb levels (20-44 (ig/dL) at age 12 to
33 months, and the findings may have limited application to the general population of
children currently living in the U.S. In fact, in all of the studies that examined weakly
correlated serial blood Pb measurements, blood Pb levels were higher than those
currently measured in U.S. children. Additionally, in several study populations, children
experienced large decreases in blood Pb levels over time. It is unclear whether these
findings would apply to children in the U.S. who currently are within the same age range
and who would be expected to have smaller decreases in blood Pb levels over time.
18
6-year:2-year
ratio = 1.25
IQ = 83.7
6-year:2-year
ratio = 0.5
IQ = 89.0
16
14
12
o)
	-~	
6-year:2-year
ratio = 1.0
IQ = 85.5
8
5 6
4
2
0
2
3
4
5
6
Age (years)
Source: Hornung et al. (2009)
Note: All three patterns have an identical mean blood Pb level of 10 |jg/dL.
Figure 5-22 Estimated IQ in combined Cincinnati and Rochester cohorts, for
three patterns of blood Pb level levels from 1 through 6 years of
age: peak at 2 years (blue diamonds), peak at 5 years (black
triangles), and constant blood Pb level (white squares).
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To conclude, in the collective body of epidemiologic evidence in children, it is difficult to
ascertain which lifestage of Pb exposure is associated with the greatest risk of Pb-
associated neurodevelopmental effects. Associations have been observed with prenatal,
early-childhood, lifetime average, and concurrent blood Pb levels as well as with
childhood tooth Pb levels. Comparisons among different lifestages of exposure is
complicated further by the fact that blood Pb levels in children, although highly 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, concurrent blood Pb level in children
also may reflect body burden (Section 4.3.4.6). Nonetheless, while the evidence indicates
that prenatal and early-childhood blood Pb levels are associated with neurodevelopmental
deficits, subsequent exposures that are reflected in concurrent, cumulative blood Pb
levels or tooth Pb levels also are demonstrated to contribute to neurodevelopmental
deficits throughout school-age and into adolescence. With additional results from studies
described in Sections 5.3.2.1 and 5.3.3.1, the weight of epidemiologic evidence supports
associations of concurrent blood Pb level with neurodevelopmental effects in children.
These findings are consistent with the understanding that the nervous system continues to
develop throughout childhood.
5.3.10 Examination of the Lead Concentration-Response Relationship
With each successive Pb AQCD and supplement, epidemiologic and toxicological studies
find that progressively lower blood Pb levels are associated with cognitive deficits and
behavioral impairments. For example, among children, such effects were observed in
association with blood Pb levels in the range of 10-15 (ig/dL in the 1986 Addendum and
1990 Supplement and 10 (ig/dL and lower in the 2006 Pb AQCD (U.S. EPA. 2006^.
Furthermore, in the 2006 Pb AQCD, several individual studies, pooled analyses, and
meta-analyses estimated a supralinear blood Pb concentration-response relationship in
children, i.e., greater decrements in cognitive function per incremental increase in blood
Pb level among children in lower strata of blood Pb levels compared with children in
higher strata of blood Pb levels (Figure 5-23 and Table 5-15). As lower concentrations of
Pb exposure are being used experimentally, the toxicological literature also has reported
nonlinear concentration-response relationships for various endpoints, including those
coherent with cognitive function decrements observed in children. Also consistent with
the epidemiologic literature, some toxicological studies have shown larger magnitudes of
effect (absolute effects) in lower Pb exposure groups (relative to control groups) than in
the higher exposure groups.
In epidemiologic studies, a supralinear concentration-response relationship was observed
for concurrent, early childhood, and lifetime average blood Pb levels. Most
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epidemiologic studies used a blood Pb level of 10 (ig/dL to define lower and higher blood
Pb levels. Several of these observations were made in groups of children aged 2 to
10 years, and the mean blood Pb levels in the lower strata of blood Pb levels were in the
range of 3-5 (.ig/dL (Bellinger. 2008; Cnn field. 2008: Hornung. 2008; Roio-Tellez. 2008).
Except for the pooled analysis, lower strata of blood Pb levels comprised at least 55% of
the study population and consequently the concentration-response relationships
calculated for these lower strata likely are not outliers comprised a minimum 55% of the
study population and consequently the concentration-response relationships calculated for
these lower strata are not outliers or unrepresentative of the overall study population.
Using data pooled from seven prospective studies, Lanphear et al. (2005) fit various types
of models to the data and observed that a cubic spline, log-linear model, and piece-wise
linear model all supported a more negative concentration-response relationship at
concurrent blood Pb levels <10 (ig/dL. A linear model was found to be inadequate as the
polynomial terms for concurrent blood Pb were statistically significant. These findings
were corroborated in a separate analysis by Rothenberg and Rothenberg (2005) who
found that the log-linear model fit the relationship between blood Pb level and IQ better
than a linear model did.
A few studies demonstrated larger Pb-associated decreases in cognitive function with
blood Pb levels < 5 (ig/dL. Tellez-Rojo et al. (2006) estimated a larger decrement in IQ
per unit increase in blood Pb level for children (age 2 years) with concurrent blood Pb
levels <5 j^ig/dL compared with children with levels 5-10 (ig/dL, and > 10 (ig/dL (Figure
5-23 and Table 5-15). In an NHANES analysis 1989-1994, 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 compared with children
with levels <5 (ig/dL, <7,5 (ig/dL, <10 (ig/dL, and all subjects. However, as the study
population included adolescents, the oldest of whom were 16 years of age and born
1972-1978, higher Pb exposures earlier in childhood may have contributed to
associations.
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Study
Lanphearetal. (2000)
Blood Pb Timing Outcome
Concurrent
Tellez-Rojoetal. (2006) Concurrent
Bellingerand Needleman
(2003)
Canfield etal. (2003)
Lanphearetal. (2005)
Jusko (2011)
Kordasetal. (2006)
Schwartz (1994)
Lifetime avg
Peak
Early childhood
Blood Pb
stratum (pg/dL)
Reading score All
<10
<7.5
<5
<2.5
Bayley MDI
FSIQ
FSIQ
FSIQ
FSIQ
FSIQ
>10
<10
5-10
<5
>10
<10
All
<10a
>10a
<10a
>7.5a
<7.5a
All
< 10
all
<10
>15
<15
Sample size
4853
4681
4526
4043
2467
90
294
101
193
NR
NR
172
101
1089
244
1230
103
174
96
532
293
NR
NR
~
Change in Cognitive Function Score perl pg/dL
increase in blood Pb level (95% CI)
Note: Studies are presented in order of increasing mean blood Pb level. Strata refer to peak blood Pb level measured in child at any
point during follow up. FSIQ = full-scale IQ, MDI = mental development index. Effect estimates are standardized to a 1 |jg/dL
increase in blood Pb level. Black symbols represent effect estimates among all subjects or in the highest blood Pb stratum. Blue
symbols represent effect estimates in lower blood Pb strata
Figure 5-23 Comparison of associations between blood Pb level and cognitive
function among various blood Pb strata.
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Table 5-15 Additional characteristics and quantitative results for studies
presented in Figure 5-23
Study
Population/Location
Blood Pb Levels
(Hg/dL)
Statistical Analysis
Outcome
Blood Pb
stratum
(Hg/dL)
Effect Estimate
(96% CI)'
Lanphear et al.
4,853 children ages 6-16
yr
NHANES 1988-1994
Concurrent mean
(SE): 1.9(0.1)
Linear regression model
adjusted for sex, race/ethnicity,
poverty index ratio, reference
adult education level, serum
ferritin level, serum cotinine
level
WRAT reading
subtest at ages 6-16
yr
All subjects
<10
<7.5
<5
<2.5
-0.70 (-1.03,-0.37)
-0.89 (-1.52,-0.26)
-1.06 (-1.82,-0.30)
-1.06 (-2.00,-0.12)
-1.28 (-3.20,-0.64)
Tellez-Rojo et
al. 12006)
294 children followed
from birth (1994-1995,
1997-1999) to age 2 yr
Mexico City, Mexico
Concurrent (age 2
yr) mean (SD): 4.28
(2.25)
Linear regression model
adjusted for sex, birth weight,
maternal IQ
Bayley MDI at age 2
yr
>10
<10
5-10
<5
0.07 (-10, 9.2)"
-1.04 (-1.8, -0.30)"
-0.94 (-2.1,0.2)"
-1.71 (-3.0, -0.42)"
Bellinger etal.
Bellinger and
Needleman
148 children followed	Early childhood
from birth (1979-1981) to	(age 2 yr) mean
age 10 yr	(SD): 6.5 (4.9)
Boston area, MA
Linear regression model
adjusted for HOME score (age
10 and 5), child stress, race,
maternal IQ, SES, sex, birth
order, marital status
WISC-Ratage 10 yr
>10
<10
-0.58 (-1.0, -0.2)°
-1.56 (-2.9, -0.2)"
Canfield et al.
(2003a)
172 children born
1994-1995 followed from
infancy to age 3-5 yr
Rochester, NY
Lifetime avg (3 or 5
yr) mean (SD): 7.4
(4.3)
Mixed effects models adjusted
for sex, maternal race, parental
smoking, child iron status,
maternal income, maternal IQ,
HOME score
Stanford-Binet at age All
3 or 5 yr	<10
-0.46 (-0.76, -0.15)
-1.37 (-2.56,-0.17)
Lanphear et al.
1,333 children pooled
from Boston, Cincinnati,
Cleveland, Mexico City,
Port Pirie, Rochester,
and Yugoslavia cohorts
Concurrent
Median (5th-95th)
9.7 (2.5-33.2)
Linear regression model FSIQ measured at
adjusted for HOME score, birth ages 4.8-10 yr
weight, maternal IQ, maternal
education
>10
<10
>7.5
<7.5
-0.13 (-2.3, -0.03)
-0.80 (-1.74,-0.14)
-0.16 (-2.4, -0.08)
-2.94 (-5.16,-0.71)
Jusko et al.
(2011)
194 children in
Rochester, NY followed
from age 6 mo
(1994-1995) to age 6 yr.
Peak
Mean (SD): 11.4
(7.3)
Linear regression model
adjusted for sex, birth weight,
transferrin saturation, maternal
race, maternal IQ, maternal
education, HOME score, family
income, and maternal prenatal
smoking
WPPSI-R at age 6 yr All
<10
-0.19 (-0.46, 0.07)
-1.66 (-3.1,-0.23)
Kordas et al.
602 children in 1st grade
Torreon, Mexico
Concurrent mean
(SD): 11.4(6.1)
Linear regression model
adjusted for sex, age,
hemoglobin, family
possessions, forgetting
homework, house ownership,
crowding, maternal education,
birth order, family structure,
arsenic exposure, tester,
school
Math achievement
test in 1st grade
all
<10
-0.17 (-0.28, -0.06)
-0.42 (-0.92, 0.08)
Schwartz
Meta-analysis of 7
studies with sample sizes
75-579 children
Early childhood (2-3
yr) range in study
means: 6.5-23
Meta-analysis of combining
effect estimates from individual
studies
FSIQ measured at
school-age
Studies with
mean >15
Studies with
mean < 15
-2.32 (-3.10, -1.54)
-3.23 (-5.70, -0.76)
aEffect estimates are standardized to a 1 |jg/dL increase in blood Pb level.
b95% CIs calculated from reported p-value.
1	Several studies also found a nonlinear blood Pb-cognitive function relationship in
2	nonparametric regression analyses using lowess with smoothing parameters or using
3	splines (Min et al.. 2009; Jusko et al.. 2008; Schnaas et al.. 2006). Min et al. (2009)
4	performed a formal test of nonlinearity using nonparametric regression by testing the
5	statistical significance of a restricted cubic spline term for blood Pb level. Although the
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term was not statistically significant, the concurrent blood Pb level-FSIQ curve appeared
to be more negative at blood Pb levels < 7 (ig/dL. In a similar analysis in a Mexico City
cohort, Schnaas et al. (2006) found a more negative blood Pb-FSIQ slope at concurrent
blood Pb levels < 6 (ig/dL and found a test of nonlinearity to be statistically significant.
Among children in India ages 3-7 years, Palaniappan et al. (2011). most associations of
concurrent blood Pb level with various indices of executive function in nonparametric
regression analyses were found to be linear.
Studies of adults have not assessed widely whether the relationship between blood or
bone Pb level and cognitive performance is described better with a linear or nonlinear
function. In the various NHANES analyses, only log-linear models were used to fit the
data (Krieg et al.. 2010; 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; Weisskopf et al..
2007a; Shih et al. 2006). There was some evidence for nonlinearity for some (Figure
5-12 and Figure 5-13) but not all cognitive tests or all subjects in the NAS cohort. For
example, Wang et al. (2007a) found that among NAS men with an HFE variant, there
was a steeper Pb-associated decline in MMSE score at higher tibia Pb levels (20-25 jxg/g,
Figure 5-13). In the BMS cohort, observations of a statistically nonsignificant quadratic
term (Shih et al.. 2006) or spline (Bandeen-Roche et al.. 2009) for tibia Pb indicated that
a linear model adequately fit the relationship between tibia Pb level and various tests of
cognitive performance.
Attenuation of the concentration-response relationships at higher exposure or dose levels
has been reported in the occupational literature, and explanations have included greater
exposure measurement error, competing risks, and saturation of biological mechanisms at
higher levels; larger proportions of at-risk populations at lower exposure levels; and
variations in other risk factors among exposure levels (Stavner et al. 2003). Other
explanations for nonlinearity include differential activity of mechanisms at different
exposure levels, confounding by omitted or misspecified variables, and the lower
incremental effect of Pb due to covarying risk factors such as low SES, poor caregiving
environment, and higher exposure to other environmental factors.
The contribution of these factors to the supralinear relationship between blood Pb levels
and cognitive function in children has not been examined widely in epidemiologic studies
to date. However, in several different populations, higher blood Pb levels have been
measured in potentially at-risk groups such as those with higher poverty, greater exposure
to tobacco smoke, lower parental education, and lower birth weight, which argues against
a larger proportion of at-risk populations at lower blood Pb levels (Lanphear et al.. 2005;
Lanphear et al.. 2000). It has been suggested that in populations of low SES, poorer
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caregiving environment, and greater social stress, the incremental effect of Pb exposure
may be attenuated due to the overwhelmingly larger effects of these other risk factors
(Schwartz. 1994). Several studies found statistically significant associations of these
sociodemographic risk factors with neurocognitive deficits, and Miranda et al. (2009)
found that indicators of SES (i.e., parental education and enrollment in a free/reduced fee
lunch program) accounted for larger decrements in EOG scores than did blood Pb level
(Figure 5-7). Few studies have compared blood Pb level effect estimates among groups in
different sociodemographic strata, and the limited data are mixed. Greater Pb-associated
neurocognitive deficits were reported in low-SES groups by Bellinger et al. (1990). In a
meta-analysis of eight studies, Schwartz (1994) found a smaller decrement in IQ per
1 (ig/dL increase in blood Pb level for studies in disadvantaged populations (-2.7 points
[95% CI: -5.3, -0.07]) than for studies in advantaged populations (-4.5 points [95% CI:
-5.6, -2.8]). It is important to note that blood Pb level is associated with deficits in
cognitive function in both higher and lower SES groups; however, it is unclear what
differences there are between groups in the decrement per unit increase in blood Pb and
whether these differences can explain the observed nonlinear concentration-response
relationship.
Rothenberg and Rothenberg (2005) formally assessed the influence of residual
confounding on the nonlinear blood Pb-FSIQ concentration-response relationship by
comparing model fit between linear and spline transformations (df = 2) of covariates such
as maternal IQ, HOME score, and maternal education. Inclusion of covariates as spline
functions did not significantly improve model fit either with a linear blood Pb term or log
blood Pb term, which indicated that their inclusion as linear functions was adequate.
These findings demonstrate that the improved model fit with log-specification of blood
Pb level was not due to residual confounding by covariates.
Bowers and Beck (2006) postulated that a supralinear slope necessarily will be found
when modeling a relationship between a log-normally distributed variable and a normally
distributed 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. IQ scores generally were not forced into a normal distribution. Four of
the seven studies included in the pooled analysis by Lanphear et al. (2005) did not use
normalized IQ scores, and scores were not normalized in the pooled analysis (Hornung et
al.. 2006). Further, a log-linear model (a linear relationship between IQ and the log of
blood Pb) provided the best fit of the pooled data.
In support of the nonlinear associations between blood Pb levels and cognitive function
observed in children, toxicological studies provided some evidence of nonlinear
relationships between Pb exposure and effects related to impaired learning and memory
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in animals. Multiple studies showed U- or inverse U-shaped curves with lower exposures
of Pb having different or often the opposite effect from higher exposures. Coherent with
epidemiologic findings, 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 (Corv-Slechta. 1994). Consistent with
these findings, Rossi-George et al. (2011) found that 50 ppm gestational plus lactational
Pb exposure when combined with stress increased FI responses of 2-month old rats
whereas 150 ppm Pb exposure with stress did not affect FI responses. It is important to
note that the larger effects observed with lower Pb exposures were less consistently
observed with longer duration exposures (e.g., 8-11 months) (Rossi-George et al.. 2011;
Corv-Slechta. 1990). As discussed in Section 5.3.2.2, FI responses measure learning and
memory by reflecting the ability of animals to learn to respond to one schedule of
reinforcement and change responses according to a change in reinforcement.
Toxicological studies provided additional support for nonlinear relationships between Pb
exposure and neurodevelopmental effects by finding that lower and higher Pb exposures
differentially activate mechanisms underlying such effects. Gilbert et al. (1999) found
reduced LTP with Pb exposures ranging between 0.1 and 0.5% but not 1.0%. LTP is one
indication of synaptic plasticity (Section 5.3.8.4) and is considered to contribute to
learning and memory. Likewise, glutamate release in the hippocampus was reduced in
animals with blood Pb levels 27-40 (ig/dL but not with blood Pb levels of 62-117 (ig/dL
(Laslev and Gilbert. 2002). Glutaminergic neurotransmission via its NMDA receptor has
been implicated in learning and memory (Section 5.3.8.8).
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 toxicological studies. Some studies found that lower Pb
exposures (~ 50 ppm) did not affect or increased DA activity relative to controls and
higher Pb exposure (109-250 ppm) (Leasure et al.. 2008; Virgolini et al.. 2005; Lewis and
Pitts. 2004). However, compared with control and lower Pb exposures, higher Pb
exposures (109 or 150 ppm) were found both to increase and impair DA activity (Leasure
et al.. 2008; Virgolini et al.. 2005). These differential responses of DA may be related to
the diverse actions of DA in different regions of the brain and on a range of CNS effects.
For example, the increased dopamine turnover with 50 ppm Pb exposure may explain the
greater spontaneous and amphetamine-induced motor activity in males induced by
50 ppm GLE (Leasure et al.. 2008).
Lower and higher Pb exposures also were found to differentially affect calcineurin
enzyme activity; activity was inhibited by higher Pb exposure and stimulated by lower Pb
exposure (Kern and Audesirk. 2000). While calcineurin activity has been found to
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modulate learning, LTP, and behavior in animals, studies have found lower calcineurin
activity to be associated with both impaired and improved effects related to learning
(Zeng et al.. 2001). Thus, it is uncertain whether altered calcineurin activity contributes to
the nonlinear relationships observed between Pb exposure and learning. Nonetheless,
results indicated that lower and higher Pb exposures may have different modes of action.
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. Similarly, Lasley and Gilbert (2002) also indicated that at higher
concentrations (i.e., 1%), Pb stimulated gluatamate release by acting as a calcium
mimetic.
While epidemiologic studies have not examined widely the shape of the concentration-
response relationship for other nervous system effects, toxicological studies have found
nonlinear relationships for diverse outcomes. U-shaped Pb concentration-response
relationships were found for rotarod performance, amphetamine-induced motor activity,
and latency to fall from rotarod (Leasure et al.. 2008). Inverted U-shaped Pb
concentration-response relationships were found for histological parameters such as the
numbers of rod photoreceptors and bipolar cells, activity level, and adult body weight
(Leasure et al. 2008) as well as ERG wave amplitudes (Fox et al.. 2008) and
hippocampal neurogenesis (Fox et al.. 2008; Gilbert et al.. 2005). Additional evidence
points to differences in hormonal homeostasis by Pb exposure level. In male mice with
chronic Pb exposure (PND21-9 months of age), basal corticosterone levels were
significantly lower in the 50 ppm exposure group than in the control or 150 ppm Pb
exposure group.
Sensory organ functions in animals also have shown to be differentially affected by lower
versus higher Pb exposure (developmental Pb exposurefrom gestation to PND10, pup
blood Pb levels 12, 24, and 46 (ig/dL. Inverted U-shaped dose-response curves have been
observed for rod photoreceptor numbers or neurogenesis (Giddabasaooa et al.. 2011) and
retinal thickness (Fox et al.. 2010). Thus, these dichotomous histological findings may
give insight to the complex sensory organ findings that vary by exposure window and
exposure dose (Section 5.3.4.3).
The supralinear concentration-response relationship widely documented for Pb is
consistent with the lack of a threshold for Pb-associated neurodevelopmental effects as a
smaller effect estimate would be expected at lower blood Pb levels if a threshold existed.
Schwartz (1994) explicitly assessed evidence for a threshold in data from the Boston
prospective cohort by regressing FSIQ and blood Pb level on potential confounders
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including age, race, maternal IQ, SES, and HOME score and fitting a nonparametric
smoothed curve to the residuals of both regression models (variation in FSIQ or blood Pb
level not explained by covariates). A 7-point decrease in IQ was observed over the range
of blood Pb residuals below 0, which corresponds to the mean blood Pb level in the study
(6.5 (ig/dL). Thus, in the Boston study, the association between blood Pb level and FSIQ
was clearly demonstrated at blood Pb levels below 5 (ig/dL.
An important limitation of previous studies in terms of characterizing the concentration-
response relationship, in particular, identifying whether a threshold exists, was the limited
examination of associations in populations or subgroups with blood Pb levels more
comparable to the current U.S. population mean. While Schwartz (1994) did not find
evidence for a threshold in the Boston study data, the mean blood Pb in that population
was 6.5 (ig/dL, and 56% of subjects had a blood Pb level >5 (ig/dL. Recent studies
indicate a downward shift in the distribution of blood Pb levels. In the various NHANES
analyses of children, a large proportion of children had blood Pb levels <1 j^ig/dL, for
example, 50% of subjects in the 2001-2004 NHANES population (Braun et al.. 2008).
More sensitive quantification methods have improved the detection limits, for example,
from 0.6 |_ig/dL in 1999-2002 NHANES to 0.025 |_ig/dL in 2003-2004 NHANES. In
analyses of children in multiple blood Pb quantiles below 1 (ig/dL, Braun et al. (2008)
found higher odds ratios for conduct disorder and ADHD among children with blood Pb
levels 0.8-1.0 (ig/dL (2nd quartile) compared with children with blood Pb levels 0.2-
0.7 |_ig/dL (1st quartile). Despite the availability of large proportions of subjects at blood
Pb levels below 1 (ig/dL, the ability to discern a threshold for Pb-associated
neurodevelopmental effects is limited due to the large proportions of adolescents in
NHANES analyses who were born the 1970s and whose higher past exposures may have
contributed to associations observed with concurrent blood Pb levels. Nonetheless,
several recent studies reported associations between blood Pb levels and deficits in
cognitive and behavioral endpoints in children ages 8-11 years with mean or quantile
blood Pb levels <2 (.ig/dL (C'ho et al.. 2010; Kim et al.. 2009b; Miranda et al.. 2009). In
comparisons of various quantiles of blood Pb, Miranda et al. (2009) reported lower EOG
scores in children in North Carolina with blood Pb levels of 2 (ig/dL compared with
children with blood Pb levels of 1 (ig/dL. Collectively, these new findings in children, do
not provide evidence for a threshold for the neurodevelopmental effects of Pb in the
range of blood Pb levels examined to date.
It is important to note, however, that the lack of a reference population with blood Pb
levels reflecting pre-industrial Pb exposures limits the ability to identify a threshold.
Estimates of "background" blood Pb levels have been garnered from the analysis of
ancient bones in pre-industrialized societies. These studies suggest that the level ofPb in
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blood in preindustrial humans was approximately 0.016 (.ig/dL (Flegal and Smith. 1992).
approximately 65-fold lower than that currently measured in U.S. populations and lower
than the levels at which neurodevelopmental effects have been observed (1 (ig/dL). Thus,
the current evidence does not preclude the possibility of a threshold for
neurodevelopmental effects in children existing in the large range of blood levels
between 1 j^ig/dL and preindustrial "background" levels.
To conclude, several studies found a supralinear blood Pb-cognitive function
concentration-response relationship in children based on comparisons of effect estimates
in lower and higher strata of blood Pb level and nonparametric regression. Supporting
evidence was provided by some toxicological studies that showed that lower Pb
exposures induced learning and memory impairments in animals compared to control
exposures or higher Pb exposures. While explanations for this supralinear relationship
have not been well characterized by epidemiologic studies, recent toxicological studies
that examined lower concentrations of experimental Pb exposure suggest that
mechanisms may be differentially activated at lower and higher Pb exposures.
Nonlinear concentration-response relationships were found for effects such as LTP,
hippocampal glutamate release, and calcineurin expression. Observations of associations
between blood Pb levels and deficits in cognitive and behavioral endpoints in children
(ages 8-11 years) with mean or range of blood Pb levels < 2 (ig/dL (C'ho et al.. 2010; Kim
et al.. 2009b; Miranda et al.. 2009) do not provide evidence for a threshold for
neurodevelopmental effects of Pb in the range of blood Pb levels examined to date.
5.3.11 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 and behavioral function in children has been the
assessment and control for potential confounding factors. By definition, a confounder is
associated with both the exposure and the outcome and consequently has the potential to
bias the association between the exposure of interest and the outcome. Epidemiologic
studies of Pb biomarkers in children have most commonly examined potential
confounding by quality of the caregiving environment (i.e., HOME score), 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
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income of both parents. A relatively smaller number of studies have considered
confounding by other environmental exposures, measures of parental substance abuse,
and psychopathology. Studies have varied with respect to the number of potential
confounders examined, with some studies considering multiple SES-related variables and
other studies focusing on a smaller set. It is important to note that the extent of
confounding by a particular variable likely is study-specific, i.e., dependent on the
population examined. Thus, the impact of adjustment for specific covariates on the Pb
effect estimate also is likely to be study-specific.
Studies have used various methods to control for confounding, including examining a
homogeneous population with respect to SES, examining populations in which factors are
not correlated, conducting multivariate regression, characterizing the magnitude of
change in the blood Pb level effect estimate with adjustment for a covariate, and
examining associations in different strata of a covariate. While no single method is
without limitation, the consistency of findings among different methodologies
substantiates independent associations of blood Pb level with cognitive function
decrements and behavioral problems in children. The evidence derived from each of
these control strategies is discussed below.
In the Boston Prospective Study, potential confounding by SES was largely controlled for
by study design. The study subjects were generally middle- to upper-middle-class
children with married, college-educated parents. Hence, the potential for confounding by
SES in this study was considerably less compared to other studies examining the same
associations. Yet, in this cohort, blood Pb levels measured prenatally, at age 2 years, and
integrated over ages 2 and 5 years were associated decrements in full-scale IQ and
various measures of executive function (Stiles and Bellinger. 1993); (Bellinger et al..
1992; Bellinger et al. 1990). In some analyses of this cohort, larger effects were
estimated as compared with other studies.
Studies also have demonstrated associations between blood Pb levels and cognitive
function in populations in which blood Pb levels were not associated with SES-related
variables (Factor-Litvak et al. 1999; Bellinger et al. 1987). In the Boston cohort,
parental education, social class, and HOME score were similar among low (<3 (ig/dL),
medium (6-7 (ig/dL), and high (>10 (ig/dL) cord blood Pb level groups. Nonetheless,
adjusting for these and other demographic variables, Bellinger et al. (1987) found that
children in the high cord blood Pb group had a 4.8-point lower Bay ley MDI score at age
24 months than did children in the low cord blood Pb group. 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 one city and were lower in another city. Among all
children, higher blood Pb level was associated with lower FSIQ and specific indices of
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learning and memory and with greater depression and withdrawn behavior (Factor-Litvak
etal.. 1999).
The primary method used by epidemiologic studies to control for potential confounding
has been multivariate regression. This was the main method employed in recent studies
that examined children with blood Pb levels in the range of current U.S. levels. Some
studies selected a set of covariates based on a priori knowledge, whereas others selected
particular 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 variables included in models as potential
confounders. Some included multiple SES-related variables, whereas others analyzed one
or two variables. Regardless of the method used to select model covariates or the number
of covariates included, studies consistently demonstrated associations of higher blood Pb
level with lower cognitive function and greater behavioral problems. This consistency
suggests that confounding by particular variables may be specific to the population
examined and that no single measured variable or set of variables fully accounts for the
associations observed with blood Pb levels and neurodevelopmental effects in children.
The consistency of associations across populations with different SES and across studies
examining different covariates was reinforced in pooled and meta-analyses (Marcus et
al.. 2010; Lanphear et al. 2005; Schwartz. 1994). Pooling data from seven international
longitudinal cohorts, Lanphear et al. ("2005) found similar blood Pb-FSIQ effect estimates
(-2.6 to 8.6% difference) among models, each with one study omitted. These results
indicated that the pooled estimate is relatively stable despite between-study differences in
population characteristics, including SES. In a meta-analysis, Schwartz (1994) found that
individual study blood Pb-IQ effect estimates fell within a relatively narrow range despite
large differences among studies in the correlation between blood Pb level and SES. A
wider range of effect estimates would be expected if omitted SES factors confounded the
blood Pb level association. A recent meta-analysis examined the association between
blood Pb level and conduct problems in older and recent studies of children (Marcus et
al.. 2010). Among the studies included, adjustment for variables such as SES and home
environment did little to attenuate the association between Pb and conduct problems.
Among studies that provided both unadjusted and adjusted effect estimates, most
indicated statistically significant associations of blood Pb level with neurodevelopmental
outcomes in children before and after adjusting for potential confounders. Although
examples exist where blood Pb level was estimated to have a weaker, statistically
nonsignificant effect after adjustment for potential confounders (Tong and Lu. 2001;
Emhart et al.. 1989). most notably in multiple analyses of the Cleveland cohort, studies
also reported stronger blood Pb level associations in covariate-adjusted models (Canfield
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et al.. 2003b; Factor-Litvak et al.. 1999). In recent studies with population mean blood Pb
levels between 5 and 10 (ig/dL, associations between blood Pb levels and cognitive
function (e.g., FSIQ, executive function, learning, and memory) remained statistically
significant after adjusting for covariates (Min et al.. 2009; Chiodo et al.. 2007; Froehlich
et al.. 2007; kordas et al.. 2006; Schnaas et al.. 2006). Although most effect estimates
changed by 20-50% in multivariate models, they remained within the 95% CI of the
unadjusted estimate. In particular, Schnaas et al. (2006) and Froehlich et al. (2007) found
some adjusted effect estimates to be larger in magnitude compared with unadjusted
estimates. Although information was not provided on the magnitude of change in effect
estimates after adjustment for confounding, Chiodo et al. (2007) found statistically
significant associations between blood Pb level and behavioral problems
(e.g., inattention, hyperactivity, social problems, impulsivity) in models with blood Pb
level alone and with covariates. These findings demonstrate that SES-related and
demographic factors may partially account for but do not fully explain the associations
observed between higher blood Pb level and neurodevelopmental impairments.
A challenge to separating the effects of Pb exposure from those related to SES and
quality of the caregiving environment is the high correlation typically observed among
these measures. In such cases, it is difficult to know how much variation in the outcome
to attribute to each of the various risk factors (Needleman and Bellinger. 2001). For
example, the high correlation between blood Pb level and SES may lead to an
underestimation of the Pb effect when SES is added to the model. A reduction in the
magnitude and statistical significance of the Pb effect estimate following adjustment for
some measure of SES may be the result of the misattribution of the variance in outcome
due to Pb to the variance due to SES. SES may be a proxy for Pb exposure rather than a
confounder of the association of interest. This misattribution may be exacerbated when
several correlated variables are included in the same model (i.e., overcontrol).
Instead of being a confounder, Pb may be on the causal pathway of the association
between social class and IQ. Lower social class in urban children is closely linked to
residence in older housing in poor condition that, in turn, increases exposure of children
to environmental Pb and increases their risk of cognitive deficits (Clark et al.. 1985). In
such cases, statistical adjustment for SES will lead to the underestimation of the Pb effect
(Bellinger. 2004a). One extreme example of overcontrol of this nature can be found in the
New Zealand studies where investigators regularly adjusted for residence in older
wooden housing, which is associated with higher exposure to Pb paint and accumulated
dust and soil (Fergusson et al.. 1988a. b). However, it is worth noting that, even in the
models including this variable, Pb remained a statistically significant predictor of
intellectual and academic under-attainment in the Christchurch Health Study. Variables
related to SES have been shown to be effect modifiers in studies of Pb and child
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development (Tong et al.. 2000; Bellinger et al.. 1990). That is, increases in blood Pb
level have been associated with larger decreases in cognitive function among children
from less advantaged (e.g., lower SES, lower HOME score, lower maternal IQ)
household than among those from more advantaged households. Similarly, a larger
decrease in cognitive function in association with higher blood Pb level was found among
children with mothers with lower self-esteem, an indicator of stress and less connection
with the child (Surkan et al.. 2008).
In the collective body of epidemiologic evidence, higher blood Pb levels are consistently
associated with cognitive function decrements and behavioral problems in children, with
the weight of evidence indicating decrements in FSIQ and diverse measures of executive
function and inattention and hyperactivity, respectively. 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 of control for
confounding and adjust for different potential confounders. While no single method is
without limitation, the consistency of findings among different methodologies and sets of
covariates substantiates independent associations of blood Pb level with cognitive
function decrements and behavioral problems in children. In addition, Pb exposure has
been extensively studied in animals that produce blood Pb levels in the range of those
examined in children. Experimental animal studies are not vulnerable to confounding by
such factors as social class and correlated environmental factors. Adding further support
for the independent associations of blood Pb levels with neurodevelopmental outcomes in
children is the coherence of findings in animal studies for Pb-induced impairments in
tests of learning, inattention, and impulsivity, especially tests that are directly
homologous to those in children, i.e., spatial memory, rule learning and reversal, and
response inhibition. Further, an extensive body of toxicological evidence for Pb-induced
changes in brain physiological processes that mediate cognition and behavior, including
changes in neurogenesis, synaptic pruning, and neurotransmitter function in the frontal
lobe and striatum of the brain provides biological plausibility for observations in
children.
5.3.12 Public Health Significance of Associations between Lead Biomarkers
and Neurodevelopmental Effects
As described in Section 5.3.2.1, most studies found that a 1 (ig/dL increase in blood Pb
level was associated with fractional decrements in FSIQ in school-aged children (Figure
5-2 and Table 5-3) and Bayley MDI in infants (Table 5-4). Similarly, a 1 (ig/dL increase
in blood Pb level typically was associated with decreases in specific cognitive abilities
(Figure 5-5 and Table 5-5) and increases in inattention and hyperactivity (Section 5.3.3.1,
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Figure 5-14 and Table 5-9) on the order of less than 1 standard deviation. These findings
have generated discussion of the public significance of blood Pb level-associated changes
in cognitive function and behavior in children, specifically, whether these fractional
changes have consequences on the health and life-success of individuals. The World
Health Organization definition of health is the following: "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. Deficits in health indices or life-success may not be
observable except in aggregate, 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). It should also be noted that these deficits when measured in
children may set affected children on trajectories more prone toward lower educational
attainment and financial well-being. Thus, early deficits in children may have lifetime
consequences.
It has been argued that fractional decrements in IQ points are meaningless given that
these Pb-associated changes are within the 3- to 4-point standard error on 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. It is important to note
that evidence does not indicate that the standard error is nonrandom, i.e., biased in one
direction. In particular, evidence has not indicated that children with higher blood Pb
levels systematically test lower than their true IQ value and 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 bias the association to the null.
The issue of individual-level versus population-level risk also pertains to the relevance of
the magnitude of decrease in cognitive function or increase in behavioral problems per
incremental increase in blood Pb level. Although fractional changes in IQ, memory, or
inattention 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).
Weiss (1990) predicted, on purely statistical grounds, that a downward shift of five points
in mean IQ, if the amount of dispersion in the distribution remained the same, should be
accompanied by a doubling of the numbers of individuals with scores two or more
standard deviations below the mean and a reduction by half of the number of individuals
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with scores two or more standard deviations above the mean. Thus, for an individual
functioning in the low range of the IQ distribution, a Pb-associated decline of several
points might be sufficient to drop that individual into the range associated with increase
risk of educational, vocational, and social failure.
This hypothesis is supported by several observations. 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 EOG scores in lower distribution of EOG scores (Figure 5-7).
Needleman et al. (1982) found that a downward shift in the mean IQ value was associated
not only with a substantial increase in the percentage of individuals achieving very low
scores, but also with substantial decreases in percentages achieving very high scores.
Based on the study by Bellinger et al. (1987) examining associations between blood Pb
level and IQ scores in children, Weiss (1988) discussed the shift in the population
distribution of IQ from a mean of 100 and a standard deviation of 15 to a mean of 95, a
5% reduction. When the mean IQ level is 100, 2.3% of the individuals in a given
population would score above 130. However, with the population distribution shift and
the resulting mean decline in IQ, only 0.99% of the individuals would score above 130.
Weiss (1988) stated that the implication of such a loss transcends the current
circumscribed definitions of risk. In a similar analysis presented in the 2006 Pb AQCD,
using a blood Pb-IQ effect estimate of-0.9 points/|_ig/dL (based on the median of effect
estimates for blood Pb levels <10 (ig/dL), the fraction of the population with an IQ level
less than 80 more than doubles from 9% with a blood Pb level of 0 (ig/dL to 23% with a
blood Pb level of 10 j^ig/dL (Figure 5-24). The proportion with an IQ level below 70, a
level often requiring community support to live (WHO. 1992) increases 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 5-24) (U.S. EPA. 2006b).
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0.10
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Source: U.S. EPA (2006) Pb AQCD
Figure 5-24 Effect of blood Pb level on the proportion of the population with
IQ levels <70 and <80 points.
Evidence also has demonstrated larger changes in neurodevelopmental outcomes in
children with additional risk factors such as lower SES (Tong et al.. 2000; Bellinger et
al.. 1990). mothers with low self-esteem (Surkan et al.. 2008). or co-exposures to
manganese (Kim et al.. 2009a). Moreover, interventions that shift the population mean, in
a beneficial direction, by an amount that is without clinical consequence for an individual
have been shown to produce substantial decreases in the percentage of individuals with
values that are clinically significant (Bellinger. 2007. 2004b).
Also supporting the public health significance of blood Pb level-associated changes in
cognitive function and behavior are observations within the same cohorts that higher
blood Pb level are associated with decrements in IQ and measures of executive function
earlier in age and with lower academic performance, antisocial behavior, or delinquent
behavior assessed later in adolescence or in early adulthood (Chandramouli et al.. 2009;
Wright et al.. 2008; Kordas et al.. 2006; Canfield et al.. 2004; Kordas et al.. 2004;
Canfield et al.. 2003b; Dietrich et al.. 2001; Dietrich et al.. 1993a; Bellinger et al.. 1992).
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Thus, the blood Pb-associated deficits measured in children may set affected children on
trajectories more prone toward lower educational attainment and life success. Studies also
found that higher blood Pb level is associated with measures of inattention and
hyperactivity as rated by teachers or parents and by objective tests and with ADHD
diagnosis or diagnostic indices (Nicolescu et al.. 2010; Rov et al.. 2009a; Nigg et al..
2008). These findings demonstrate that associations of blood Pb level with small changes
in a health index can be markers or indicators of other changes that are likely to have
occurred whose significance is more certain.
5.3.13 Summary and Causal Determination
Recent epidemiologic and toxicological studies substantiated the strong body of evidence
presented in the 2006 Pb AQCD for the role of Pb in nervous system effects in two
domains: cognitive function and behavior in children. Specifically, new studies of
cognitive function focused on indices of learning, memory, and measures of executive
function in children consistently showed Pb-related impairments. Coherence for these
findings in children was provided by extensive toxicological evidence for Pb-induced
impairments in homologous tests of learning and memory in juvenile rats and monkeys.
New epidemiologic studies of behavior focused on and found evidence for Pb-related
effects on inattention and impulsivity in children. These findings were supported by a
large historical evidence base in juvenile animals. Consistent with these observations,
new evidence demonstrated associations with ADHD in children. Recent studies of adults
without current occupational Pb exposures continued to find associations between bone
Pb levels, a biomarker of cumulative Pb exposure, and poorer cognitive function.
Additional toxicological evidence for Pb-induced inhibition of neurotransmitter release,
decline in synaptic plasticity, and decreases in the magnitude of LTP strengthened the
biological plausibility for Pb exposure effects on decrements in both cognitive function
and behavior in children. This section presents a summary of the collective body of
evidence and identifies the new insights provided by recent studies.
5.3.13.1 Cognitive Function in Children
Epidemiologic studies provide robust evidence for higher blood Pb levels being
associated with lower FSIQ in children ages 3-11 years (Figure 5-2 and Table 5-3), with
the strongest evidence demonstrated by the consistency of association in prospective
studies in diverse populations (e.g., varying distributions of blood Pb levels, SES,
parental intelligence, and quality of caregiving) and the persistence of associations after
adjustment for potential confounding by SES, parental intelligence, caregiving
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environment, and other environmental exposures. Recent epidemiologic studies in
children focused on and found associations of concurrent blood Pb levels with specific
indices of cognitive function such as reading and verbal skills, memory, learning, and
visuospatial processing. Several new studies shifted the weight of evidence for
associations with cognitive performance to lower blood Pb levels (primarily concurrent),
with populations means in the range of 2-7 (.ig/dL (2011; Kim et al.. 2009b; Min et al..
2009; Miranda et al.. 2009; Zailina et al.. 2008; Chiodo et al.. 2007). Evidence clearly
indicates that prenatal cord and concurrent blood Pb levels are associated with lower
cognitive function in younger children ages 6 months to 3 years (Table 5-4). Coherence
for findings in children is derived from extensive evidence in animals that gestational and
early postnatal Pb exposures resulted in impaired learning and memory, in particular, in
homologous tests of spatial memory and rule learning and reversal. Reflecting the
tendency to examine higher Pb exposures, toxicological studies most clearly
demonstrated neurodevelopmental effects with blood Pb levels of 20-40 (ig/dL in
animals. However, several new studies added to the evidence for impaired learning and
memory in animals with lower blood Pb levels, 8-17 (.ig/dL (Corv-Slechta et al. 2010; Li
et al.. 2009c; Niu et al.. 2009; Virgolini et al. 2008a; Stannic et al.. 2007).
The large body of evidence indicating associations between blood Pb levels and
decrements in the diverse set of indices related to learning, memory, and other executive
functions provides coherence with findings for FSIQ, a global measure of cognitive
function that reflects the integration of these individual domains. Further, evidence for
effects on various diverse measures of cognitive function provides biological plausibility
for associations observed between blood Pb levels and factors that may be indicators of
life success, including the level of educational attainment and academic performance.
5.3.13.2 Behavior in Children
Epidemiologic studies in children demonstrate associations of higher blood Pb levels
with a range of behavioral problems, with the weight of evidence demonstrating
associations with inattention and hyperactivity as rated by parents or teachers and as
assessed using objective neuropsychological tests. Previous studies found associations
with early childhood blood or tooth Pb level (i.e., ages 2-6 years), and recent studies
expanded evidence to include associations with concurrent blood Pb level. Recent
epidemiologic studies consistently found associations with inattention and hyperactivity
in children ages 1 to 12 years with mean concurrent blood Pb levels of 2 to 5 (ig/dL (Cho
et al.. 2010; Nicolescu et al.. 2010; Pliisaiiellec et al.. 2010; Chiodo et al.. 2007;
Plusqiiellec et al. 2007). similar to those associated with cognitive function decrements.
The epidemiologic findings are strengthened by observations in animals of Pb-induced
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inattention and impulsivity in homologous tests of response inhibition. Such effects in
animals are most clearly indicated with gestational and early postnatal Pb exposures that
result in blood Pb levels of 10 to 40 (ig/dL. Whereas previous evidence was inconsistent,
several recent epidemiologic studies indicate associations between higher concurrent
blood Pb level and higher prevalence or incidence of ADHD diagnosis and its
contributing diagnostic indices in children ages 8-17 years (C'ho et al. 2010; Nicolescu et
al.. 2010; Rov et al.. 2009a; Nigg et al.. 2008; Braun et al. 2006). The biological
plausibility for associations with ADHD is strongly supported by the large epidemiologic
and toxicological evidence base demonstrating Pb-associated increases in inattention and
impulsivity, both of which are primary symptoms of ADHD. A smaller but equally
consistent body of evidence indicated associations of concurrent and early childhood
blood Pb levels with social misconduct in children and delinquent behaviors in
adolescents and young adults (Chandramouli et al.. 2009; Braun et al.. 2008; Wright et
al.. 2008; Chiodo et al.. 2007). Associations of blood Pb levels with ADHD, misconduct,
and delinquency were observed in populations of children with a wide range of blood Pb
levels, 1 to 11 (ig/dL, all similar in the strength of evidence.
While the different behavioral indices are examined, Pb exposure also was found to affect
behavior (decreased ability to escape predators or capture prey) in aquatic and terrestrial
species (Sections 7.2 and 7.3).
5.3.13.3 Other Nervous System Effects in Children
A few new toxicological studies augmented the evidence for Pb-related effects to the
visual system by demonstrating retinal changes in male rodent offspring in association
with lower blood Pb levels (<15 (ig/dL) than previously examined (20 to > 100 (ig/dL)
(Section 5.3.4.3). A small body of epidemiologic evidence together with a large historical
base of toxicological evidence indicated associations of Pb biomarkers or exposure with
impaired auditory function. Associations were found in children ranging from 4 to
19 years in age and with mean concurrent blood Pb levels of 7-12 (ig/dL Section 5.3.4.1).
While mood and emotional state have been examined less frequently compared with
inattention and misconduct, several studies found associations of biomarkers of
cumulative Pb exposure (i.e., tooth or childhood average blood Pb) and concurrent blood
Pb levels with parental or teacher reports of withdrawn behavior or depression in children
with mean blood Pb levels 8-28 (ig/dL (Section 5.3.3.3). These findings in children are
supported by a small body of toxicological studies in which prenatal plus lactational Pb
exposure resulted in depression-like behavior in rodent models. Studies also reported
associations of early childhood average and concurrent blood Pb levels with lower fine
and gross motor function in children ages 3 to 17 years (Section 5.3.5). A common
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observation across studies was finding that biomarkers of Pb exposure were associated
with decrements in multiple neurodevelopmental outcomes, including cognitive function,
externalizing behaviors, internalizing behaviors, and motor function, within the same
population. These findings indicate that Pb exposure affects a broad spectrum of
neurodevelopmental effects in children.
5.3.13.4 Factors that Modify Risk in Children
Several host and environmental factors were examined for their modification of
associations between blood Pb levels and nervous system effects in children. Although
each particular factor was examined in only one to two epidemiologic studies, it is
important to note that several findings are supported by a larger base of toxicological
evidence. Interactions of blood Pb levels with race/ethnicity and SES are not well
characterized. Most investigation focused on sex-based differences. Cumulative
epidemiologic evidence does not conclusively demonstrate increased risk of males or
females for Pb-associated cognitive function decrements. Animal studies continue to
demonstrate differential effects in males and females that vary depending on the
endpoint.
The weight of evidence from animal studies continues to support interactions between
developmental Pb exposure and stress. New animal studies find a potentiating effect of
stress, whereby lower concentration Pb exposures impact behavior and memory with co-
exposures to stress than with Pb exposures alone. In comparison, epidemiologic evidence
for such interactions is sparse; however, a recent study indicated that among children
with a positive social environment, as characterized by maternal self-esteem, blood Pb
level is not associated with a decrease in cognitive function (Surkan et al.. 2008). This
finding was consistent with a previous study in rats, in which Pb-exposed animals reared
in cages with enriched environments (i.e., toys) performed better in a test of spatial
learning and memory than did their Pb-exposed littermates reared in traditional caging,
Studies have examined modification of Pb-associated cognitive function by genes that
affect Pb toxicokinetics and/or function in neurophysiological and neurochemical
processes that mediate cognition. Most investigation has focused on ALAD variants but
has not consistently found to be in a consistent direction. Inconsistencies also have been
observed for VDR and dopamine receptor variants. In addition to host factors, recent
studies suggested that associations between blood Pb levels and cognitive function in
children are greater with co-exposures to environmental tobacco smoke (Froehlich et al..
2007) and manganese (Claus Henn et al.; Kim et al. 2009b)..
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5.3.13.5 Nervous System Effects in Adults
Among adults, associations of blood Pb level with the spectrum of nervous system effects
(e.g., impairments in memory, attention, mood, balance, motor function) were most
consistently observed in occupationally-exposed adults with blood Pb levels > 14 (ig/dL.
Evidence in adults without occupational Pb exposures is derived primarily from cross-
sectional analyses in a few different cohorts, thus the important magnitude, timing,
frequency, and duration of Pb exposures contributing to the observed associations are
uncertain. The weight of evidence demonstrated associations of bone Pb levels and not
blood Pb levels with poorer cognitive performance, including declines in function over
time (Bandeen-Roche et al.. 2009; Weisskopf et al. 2007a'). These findings point to a
stronger effect of cumulative Pb exposure, including likely higher past exposures. One
explanation for the overall weaker body of evidence in adults may be that cognitive
reserve may compensate for the effects of Pb exposure on learning new information.
Compensatory mechanisms may be overwhelmed with age, which may provide an
explanation for more consistent associations observed for higher tibia Pb levels,
representing higher long-term or cumulative Pb exposure.
Based on a smaller body of epidemiologic studies, blood and bone Pb levels were
associated with essential tremor and PD, respectively, in adults (Section 5.3.7.1).
However, in these case-control studies, it is difficult to establish temporality between Pb
exposure and disease. Support for epidemiologic findings for PD is provided by
toxicological evidence for Pb-induced decreased dopaminergic cell activity in the
substantia nigra, which contributes to the primary symptoms of Parkinson's disease.
Whereas evidence for association with Alzheimer's Disease in adults is weak,
developmentalbut not adult-only Pb exposures of monkeys (early postnatal, PND 1-400)
and rats (lactational) has been shown to induce formation of amyloid plaques, pathology
that underlies Alzheimer's Disease (Section 5.3.7.2). Thus, epidemiologic studies that
examined concurrent bone or brain Pb levels or occupational Pb exposure may not have
examined the etiologically relevant exposure period.
Rather than examining externalizing behaviors and criminal behavior, a small body of
studies of behavior in adults examined and found associations of blood (Bouchard et al..
2009) and tibia (Raian et al.. 2008) Pb levels with depression and anxiety symptoms. It is
not surprising that Pb exposure may increase the risk of different nervous system
endpoints in children and adults given the predominance of different neurophysiological
processes operating at different ages, for example, neurogenesis and brain development
in children and neurodegeneration in adults.
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5.3.13.6 Neurophysiological and Neurochemical Changes
Extensive evidence from toxicological studies clearly provides the coherence and
biological plausibility for effects observed in epidemiologic and toxicological studies on
cognitive function, behavior, mood, and neurodegenerative conditions by characterizing
underlying mechanisms (Section 5.3.8). Dopamine plays a key role in cognitive functions
mediated by the prefrontal cortex and also motor functions mediated by the substantia
nigra. Thus, extensive evidence for Pb-induced dopaminergic changes in animals
provides mechanistic support for associations in humans between blood Pb levels and
cognitive deficits and in adults for associations with Parkinson's Disease. Current
toxicological research has been expanded to document that early-life Pb exposure can
contribute to neurodegeneration and neurofibrillary tangle formation in the aged brains of
animals. Pb has been shown to induce complex neurochemical changes in the brain that
differ by region of the brain, neurotransmitter type, age and sex of the organism. These
changes remain aberrant overtime and are dynamic in nature. Pb exposure affects
NMDA receptors, which may explain findings in animals and humans for Pb-associated
symptoms of depression and withdrawn behavior. Recent toxicological studies continue
to document Pb exposure effects on synapse formation, adhesion molecules, and
nitrosative stress. A new study of epigenetics details that Pb exposure affects methylation
patterns in rodent brains, which may provide a mechanism by which prenatal Pb exposure
leads to impaired neurodevelopmental function later in life. Biological plausibility also is
provided by a small body of evidence in young adults in which childhood blood Pb levels
are associated with altered structure and activity in regions of the brain (assessed by MRI
or MRS) that mediate cognitive processes and behavior.
5.3.13.7 Lifestages and Duration of Lead Exposure
Toxicological studies clearly demonstrate that in utero and early postnatal exposure to Pb
results in impaired learning, memory, and behavior. This evidence is well supported by
knowledge that processes such as neurogenesis, synaptogenesis, and synaptic pruning are
most active during this developmental period. In epidemiologic studies reviewed in the
2006 Pb AQCD, biomarkers of prenatal, early life, concurrent, and cumulative Pb
exposures were associated with decrements in neurodevelopmental function in children,
with no clear indication that blood Pb levels measured at a particular lifestage was more
strongly associated with neurodevelopmental effects (Section 5.3.9). Distinguishing
among the effects of Pb exposures at different lifestages is difficult in epidemiologic
studies due to the high correlations among blood Pb levels within children over time.
Recent studies in children primarily examined concurrent blood Pb levels but also found
association with prenatal cord (Pilsner et al.. 2010; Jedrvchowski et al. 2009b;
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Plusqiiellec et al. 2007) and early childhood blood Pb levels (Chandramouli et al.. 2009;
Min et al.. 2009; Miranda et al.. 2009). Thus, while biomarkers of Pb exposure across
lifestages were associated with neurodevelopmental decrements in children, the weight of
cumulative evidence demonstrated associations with concurrent blood Pb levels. In
adults, the weight of evidence supported associations of cumulative Pb exposure
(i.e., tibia Pb) with poorer cognitive performance.
With regards to duration of exposure, several epidemiologic studies indicated that Pb
exposures of less than 1 year, represented by blood Pb levels prenatally (maternal or
cord) and at age 6, 12, or 24 months, were associated with lower cognitive function in
children ages 6 months to 3 years (Table 5-4). Toxicological studies supported these
observations with effects observed with gestational and early postnatal Pb exposures.
5.3.13.8 The Lead Concentration-Response Relationship
In the 2006 Pb AQCD, several individual epidemiologic studies, pooled analyses, and
meta-analyses estimated a supralinear blood Pb concentration-response relationship in
children, i.e., greater decrements in cognitive function per incremental increase in blood
Pb level or more negative slope among children in lower strata of blood Pb levels
compared with children in higher strata of blood Pb levels (Figure 5-23 and Table 5-15).
Although a majority of epidemiologic evidence used a blood Pb level of 10 (ig/dL to
define lower and higher blood Pb levels, some found that among children less than age
10 years with mean concurrent blood Pb levels of 3 (ig/dL, the blood Pb-cognitive
function slope was more negative with blood Pb levels <7.5 (Lanphear et al.. 2005) or
<5 (.ig/dL (Tellez-Roio et al. 2006).
While explanations for this supralinear relationship have not been well characterized by
epidemiologic studies, recent toxicological studies that examined lower concentrations of
experimental Pb exposure suggest that mechanisms may be differentially activated at
lower and higher Pb exposures. Multiple studies showed U- or inverse U-shaped curves
with lower Pb exposures having different levels or often the opposite effect from higher
exposures. Nonlinear concentration-response relationships were found for effects such as
learning (Section 5.3.2.2), adult forebrain dopamine levels (Lcasure et al.. 2008). and
neurogenesis (Fox et al.. 2008). Sensory organ parameters in animals, namely, the
numbers of rod photoreceptors and bipolar cells and ERG wave amplitudes also show
vastly different changes with low versus higher Pb exposure (Table 5-12).
The examination of populations of children with large proportions of subjects at very low
blood Pb levels has improved the ability to discern a threshold for Pb-associated effects
on cognitive function and behavior. Several recent epidemiologic studies reported
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associations between blood Pb levels and deficits in cognitive and behavioral endpoints
in children (ages 8-11 years) with mean or range of blood Pb levels < 2 (ig/dL (C'ho et al..
2010; Kim et al. 2009b; Miranda et al.. 2009). Collectively, these new findings in
children do not provide evidence for a threshold for neurodevelopmental effects of Pb in
the ranges of blood Pb levels examined to date.
5.3.13.9 Evidence that Forms the Basis of the Causal
Determination
In summary, recent findings strengthen epidemiologic and toxicological evidence
indicating that Pb exposure is associated with nervous system effects. The weight of
epidemiologic and toxicological evidence clearly supports associations of higher blood
Pb levels with lower cognitive function in children, i.e., full-scale IQ and various
measures of learning and memory. In epidemiologic studies, these associations were
substantiated in children ages 1 to 11 years and in populations with mean blood Pb levels
between 2 and 7 (ig/dL. Observation of a supralinear concentration-response relationship
and associations with mean (or quantile) blood Pb levels < 2 (ig/dL do not provide
evidence for a threshold for the neurodevelopmental effects of Pb exposure.
Epidemiologic and toxicological evidence clearly demonstrates Pb-associated increases
in behavioral problems, in particular, inattention and impulsivity. Associations are
substantiated in children ages 1 to 12 years with mean concurrent blood Pb levels of 2 to
5 (ig/dL. In animals, the weight of evidence demonstrates effects on cognition and
behavior with prenatal and early postnatal Pb exposures that resulted in blood Pb levels
of 10 to 40 (ig/dL. In epidemiologic studies, associations with cognitive function and
behavior were observed after adjustment for a range of potential confounding variables,
but most commonly, parental IQ, parental education, and other SES-related variables. In
children, the weight of evidence supports cognitive function decrements and behavioral
problems in association with concurrent blood Pb levels. Associations also are observed
with prenatal, early childhood, and childhood average blood Pb levels, thus uncertainty
remains regarding the lifestage of exposure within childhood that is associated with the
greatest risk. The weight of toxicological evidence demonstrates neurodevelopmental
effects with prenatal and early postnatal Pb exposures that can have effects persisting to
adulthood. The biological plausibility for epidemiologic and toxicological findings for
effects on cognitive function and behavior is provided by evidence characterizing
underlying mechanisms, including Pb-induced changes in neurogenesis, synaptogenesis
and synaptic pruning, long term potentiation, and neurotransmitter function. In adults, the
timing, level, frequency, and duration of Pb exposure implicated in nervous system
effects remain uncertain. Among occupationally-exposed adults, a spectrum of nervous
system effects is associated with concurrent blood Pb level (>14 (ig/dL), which reflects
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both current and cumulative exposure. However, in adults without occupational exposure,
cognitive performance is more strongly associated with tibia Pb levels than blood Pb
levels, which indicates an effect of long-term, cumulative Pb exposures. Based most
heavily on cognitive function decrements and inattention in children, the collective body
of evidence integrated across epidemiologic and toxicological studies is sufficient to
conclude that there is a causal relationship between Pb exposures and nervous system
effects.
5.4 Cardiovascular Effects
5.4.1 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 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 5.5). 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 et al.. 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, meta-analysis of human studies
found that each doubling of blood Pb level (between 1 and >40 (ig/dL 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
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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.. 1996b)
or increased hypertension with increasing bone Pb level (Lee et al.. 2001a). Across
studies, over a range of bone Pb concentrations (<1.0 to 96 jxg/g), every 10 jxg/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
jxg/g increase in bone Pb concentration over a range of <1 to 52 jxg/g.
With regard to etiologically-relevant timing of Pb exposure, the weight of toxicological
evidence demonstrated increases in BP after long-term (> 4 weeks) Pb exposure. In
epidemiologic studies, as cardiovascular outcomes were most often examined in cross-
sectional studies with one or a limited number of Pb biomarker measurements,
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, are complicated by the
generally higher past exposures common in this population. Detailed interpretation of Pb
in blood and bone are provided in Sections 4.3 and 4.7.3. 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 during 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
(ig/dL) in adults may be influenced by higher past Pb exposures (Section 4.4.1).
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 (U.S.
EPA. 2006b). 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, diet,
alcohol use, BMI, comorbidities, and smoking. The epidemiologic evidence was
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substantiated with results from several available prospective studies indicating
associations between Pb biomarkers and the incidence of cardiovascular health effects.
5.4.2 Blood Pressure and Hypertension
5.4.2.1 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 employed have been historically established by reference to informed medical
opinion, and as medical knowledge has improved BP cut-points defining hypertension
have been lowered overtime. 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 new studies in the
current review used the same 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, sudden death) is the most
lethal sequela of hypertension (Ingelsson et al.. 2008; Chobanian et al.. 2003; Pastor-
Barriuso et al.. 2003; Prospective Studies Collaboration. 2002; Kannel. 2000a. b; Neaton
et al.. 1995). Several recent general population and occupational cohort and cross-
sectional studies strengthened the evidence that blood and bone Pb level were associated
consistently with measures of BP (Figure 5-25 and Table 5-16) as well as with the
prevalence and incidence of hypertension (Figure 5-26 and Table 5-17). Further, recent
studies expanded evidence by 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 community-based study of older
adults (n = 964, age ranging from 50 to 70 years) in Baltimore, MD. 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
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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. Martin et al. (2006) concluded that Pb in blood has an
acute 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 an acute effect but
may also reflect an effect of cumulative Pb exposure.
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Reference
Population
Pb Distribution" Pb bio marker
Martinet al I2J06)
Glenn e. si 120061
VVcs-.fr e'. a I 1133s)
Srimcai lello =tal 120101
Martinet si 12001:1
Glernetsl (2006)
Peters ecsi (20071
Peters etal f2D0T|
Weaveretal (20051
Scipicariello et al 12010)
Martinet si. (2006)
Zhang et al. (2010)
Adults, Baltimore, MO
Korean Pb Workers
Korear Pb Workers
NHAMES'i] - Whites
NH ANES I '< - Blacks
MHWESIIl- Meucans
Adults, Baltimore, MD
•¦oresr Pb Workers
NASrren. High Stress
N AS men, Lov> Stress
NASrren. High Sties;
"orean Pb Workers
PiHANESHi- '-Vl-ites
NH ANES Ih - Blacks
NHMESIil- Mexicans
WAS it en, HFE Wild-type
"MS rren, HFE H63D
MAS rren, HFE C252"
MAS rren, Artv HFE variant
fiAJ rren, HPE VV11d-t\pe
'.AS rren HFEH63D
f4AS rren, HFE C282'«
ft AS rren, Any HFE variant
2.9(2.0, 4.4)
2T 2HP i, 2S 51
1 6i0 S, 3 Jl
1	410 5 5 6i
2	01,1 D £ ?i
15 7|10 5,11 51
IS 1(12 2, ~6 SI
IS 1112 2, 26 9)
25 9ilB 4 j9 SI
~4 3167 3 82 01
1 bfO S 3 31
1	4(0 6 3 61
2	0110 i 9)
2 912 0 4 41
15 -llJ 5 _J 51
13 112 2"?
IS 114 26
20 114, 27
19(14 2^
26(17 14
2~I19 3~
25 117
26iIS 37
Blood Pb
Blood Pb (concurrent)
Blood Db (longitudinal)
Blood Pb
In 5 load Pb
In Flood Pb
In Dlood Pt
"ibis Pb
'iba °b i historical)
"itis Pb
Tibia Pb
Patella Pt
Patella Pb
In Blood PI:
In Eloou Pb
In Blood Pb
OIodce Pb
"ibia Pb
Tibia Pb
Tibia Pb
Tibi3 Pb
Tibia Pb
Patella Pb
Datella Pb
Patella Pb
Patella °b
SBP

O
O
9
t
i
DBP
PP
Change in BP (mmHg) per 1 [xg/dL increase in
blood PborlO j-ig/g bone Pb
aPb distribution presents the median (IQR) that were estimated from the mean and SD assuming a normal distribution.
bEffect estimates were standardized to 1 |jg/dL blood Pb or 10 jjg/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. For associations of a 1 |jg/dL increase in blood Pb level (closed
circles) or 10 jjg/g increase in bone Pb (open circles) with systolic BP (SBP; blue), diastolic BP (DBP; red), and pulse pressure (PP;
purple) in adults.
Figure 5-25 Concentration-response relationships (95% CI).
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Table 5-16 Additional characteristics and quantitative data for associations of
blood and bone Pb with BP measures for studies presented in
Figure 5-25
Study
Population
/Location
Parameter
Pb Data
Statistical Analysis
Effect Estimate
P (96% CI)
Martin et al.
964 men and women, 50- BP
70 yr, 40% African
American, 55% White, 5%
other, in Baltimore, MD
Concurrent Mean
Blood Pb:
Mean (SD): 3.5 (2.3) pg/dL
African American: 3.4 (2.3)
White: 3.5 (2.4)
Tibia Pb:
Mean (SD): 18.8 (12.4) |jg/g
African American: 21.5 (12.6)
White: 16.7(11.9)
Multiple linear regression base model
adjusted forage, 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
select models that are presented in
Figure 5-25 and tabulated here.)
Blood Pb
SBP: 1.05 (0.53,1.58)
DBP; 0.53 (0.25, 0.81)
Tibia Pb:
SBP: 0.07 (-0.05, 0.14)
DBP: 0.05 (-0.02, 0.08)
mmHg per pg/dL blood Pb
mmHg per pg/g bone Pb
Glenn et al.
575 Pb exposed workers, BP
age 18-65 yr, in South
Korea (10/1997-6/2001)
Blood Pb mean (SD):
Visit 1
Visit 2
Visit 3
Visit 1
Visit 2
Visit 3
20.3 (9.6), Wsmen
20.8(10.8), Women
19.8(10.7), Women
35.0(13.5), Men
36.5(14.2), Men
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 body mass index, sex,
baseline BP lowering medication use,
alcohol consumption, body mass
index, sex, BP lowering medication
use.
Tibia Pb, mean (SD):
Visit 1
Visit 2
Visit 1
Visit 2
28.2(19.7), Women
22.8 (20.9), Women
41.7(47.6), Men
37.1 (48.1), Men
Patella Pb, mean (SD):
Visit 3 49.5 (38.5) Women
Visit 3 87.7 (117.0)
Model 1 (short-term)
Blood Pb concurrent
0.08 (-0.01, 0.16)
Blood Pb (longitudinal)
0.09(0.01,0.16)
Model 4: short and longer-
term)
Blood Pb concurrent
0.10(0.01,0.19)
Blood Pb longitudinal:
0.09(0.01,0.16)
mmHg per 10 pg/dL blood
Pb
V\feaver et al.
652 current and former BP
Pb workers in South
Korea (12/1999-6/2001)
Concurrent Blood Pb:
Mean (SD): 30.9 (16.7)
pg/dL
Concurrent Patella Pb:
Mean (SD): 75.1 (101.1)
pg/g
Linear regression model adjusted for
age, gender, BMI, diabetes,
antihypertensive and analgesic
medication use, Pb job duration, work
status, tobacco and alcohol use
SBP
Patella Pb
0.0059 (-0.008, 0.02)a
Blood Pb
0.1007 (0.02, 0.18)a
mmHg per 1 pg/dL blood Pb
or 1 pg/g patella Pb
Interaction between blood
Pb/patella Pb with ALAD and
vitamin D receptor
polymorphisms not
significant.
Peters et al. 513 elderly men (mean BP
(2007)	67 y) from NAS in Greater
Boston, MA area
Tibia Pb:
mean (SD): 21.5 (13.4) pg/g
Patella Pb:
Mean (SD): 31.5 (19.3) pg/g
Logistic and linear regression models
adjusted for age, age squared,
sodium, potassium, and calcium
intake, family history of hypertension,
BMI, educational level, pack-years of
smoking, alcohol consumption, and
physical activity
SBP
Tibia Pb/ High Stress:
3.57 (0.39, 6.75)
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 tibia Pb
Patella Pb/ Low Stress: NR
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Study
Population
/Location
Parameter
Pb Data
Statistical Analysis
Effect Estimate
P (96% CI)
Scinicariello et
al. (2010)
6,016 NHANES III (1988- BP
1994) participants > 17 yr
Concurrent Blood Pb:
Overall Mean (SE): 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 calcium,
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 In
Blood Pb
Significant interactions with
blood Pb and ALAD
observed in relation to SBP
for non-Hispanic whites and
non-Hispanic blacks
Zhang et al.
(2010a)
619 older adult males
(mean 67 yr) enrolled in
the NAS in Greater
Boston, MA area
PP	Wild type HFE
Tibia Pb: Median (IQR):8(12-
27) pg/g
Patella Pb: Median
(IQR) :26( 17-37) pg/g
C282Y HFE
Tibia Pb: Median (IQR):20
(14-27) pg/g
Patella Pb: Median
(IQR):25(17-37) pg/g
H63D HFE
Tibia Pb: Median
(IQR):19(14-26) pg/g
Patella Pb: Median
(IQR):27(19-37)pg/g
Linear mixed effects regression
models with repeated measurements
adjusted forage; education; alcohol
intake; smoking; daily intakes of
calcium, sodium, and potassium; total
calories; family history of
hypertension; diabetes; height; heart
rate; high-density lipoprotein (HDL);
total cholesterol:HDL ratio; and waist
circumference
PP
mmHg per 13 pg/g Tibia Pb:
Wld 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 pg/g Patella
Pb:
Wld 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)
Perlstein et al.
(2007)
593 predominantly white
men from NAS in Greater
Boston, MA area (1991-
1997)
PP	Blood Pb:
Overall mean (SD): 6.12
(4.03) pg/dL
Mean (SD) quintiles:
Q1
Q2
Q3
Q4
Q5
2.3	(0.8) pg/dL
3.9 (0.3) pg/dL
5.4	(0.5) pg/dL
7.4 (0.6) pg/dL
12.4(4.4) pg/dL
BP association assessed using
spearman correlation coefficients.
PP associationfadjusted 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
Tibia Pb:
Median: 19 pg/g
Mean (SD) quintiles:
PP
4.2 (1.9, 6.5) mmHg higher
in men with tibia Pb > 19
pg/g (median) compared with
men with tibia Pb < pg/g
Blood Pb (mean difference):
Q5
Q4
Q3
Q2
Q1
-1.49 (-4.93, 1.94)
-1.39 (-4.94, 2.15),
-2.56 (-5.78, 0.67)
-4.37 (-7.88,-0.86)
Referent group
Tibia Pb (mean difference):
Q1
Q2
Q3
Q4
Q5
7.4 (3.2) pg/g
14.1(1.4) pg/g
18.9(1.4) pg/g
24.9 (2.2) pg/g
40.9(14) pg/g
Q5
Q4
Q3
Q2
Q1
2.58 (-1.15, 6.33)
2.64 (-0.93, 6.21)
-0.73 (-4.27, 2.82)
-3.02 (-6.48, 0.44)
Referent group
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Study
Population
/Location
Parameter
Pb Data
Statistical Analysis
Effect Estimate
P (96% CI)
Navas-Acien et
al. (2008)
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 |jg/g Tibia Pb
Prospective/SBP
0.33 (-0.44,1.11)
X-sectional SBP
0.26 (0.02, 0.50)
X-sectional DBP
0.02 (-0.15, 0.19)
Hypertension per 10 pg/g
patella Pb
x-Sectional hypertension
OR: 1.04(1.01, 1.07)
Pooled Estimate
hypertension
OR: 1.04(0.96, 1.12)
Yazbecketal.
971 pregnant women, age BP
18-45 yr, in France
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;
cadmium, manganese, and selenium
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.
Elmarsafawy 471 elderly men (mean BP
et al. 12006)b 67 yr) from NAS in
Greater Boston, MA area
Blood Pb:
Mean (SD): 6.6 (4.3) pg/dL
Tibia Pb:
Mean (SD): 21.6 (12.0) pg/g
Patella Pb:
Mean (SD): 31.7 (18.3) pg/g
Linear regression models adjusted for Tibia Pb
age, BMI, family history of
hypertension, history of smoking,
dietary sodium intake, and cumulative
alcohol ingestion
High calcium group (>800
mg/d):
SBP: 0.40.(0.11, 0.70)
Low calcium group (<800
mg/d):
SBP: 0.19 (0.01, 0.37)
mmHg per pg/g tibia Pb
a95% CIs estimated from given p-value.
'References not included in Figure 5-25.
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Reference
Martin etal. (2006)
Peters etal. (2007)
Study Location Population
Baltimore, MD
Boston, MA High Stress
Elmarsafawy et al. (2006) NAS
Yazbecket al. (2009) France
Muntner et al. (2005) NHANES I
High Stress
Low Calcium
High Calcium
Pregnant Women
Non-Hispanic Whites
Non-Hispanic Blacks
Mexican Americans
Scinicarielloetal. (2010) NHANES III Non-Hispanic Whites
age >17 y
mn BLL= 2.99 |ig/dl_
1998-1994
Park etal. (2009)
mn BLL= 3.52 |ag/dL
NHANES I
Non-Hispanic Blacks
Mexican Americans
Overall
White Men
Black Men
White Women
Black Women
Men <50
Men >50
Women <50
Women >50
Blood Pb (|ig/dL)
3.5(2.3)
Tibia Pb (ng/g)
18.8(12.4)
21.5(13.4)
Patella Pb (ng/g)
31.5(19.3)
6.6(4.3)
21.6(12.0)
31.7(18.3)
6.6(4.3)
21.6(12.0)
31.7(18.3)
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)
Comparison
per 2.5 |ig/dl_
per 12.9 |jg/g
per 11.6 ng/g
per 17.1 ng/g
per 1 ng/g
per 1 ng/g
per 1 ng/g
per 1 ng/g
per 1 ng/g
per 1 ng/g
Reference
Q2vQl
Q3vQl
Q4vQl
Reference
Q2vQl
Q3vQl
Q4vQl
Reference
Q2vQl
Q3vQl
Q4vQl
Reference
Q2vQl
Q3vQl
Q4vQl
Reference
Q2vQl
Q3vQl
Q4vQl
ALAD2 v 1 b
Reference
Q2vQl
Q3vQl
Q4vQl
ALAD2 v 1 b -
Reference
Q2vQl
Q3vQl
Q4vQl
ALAD2 v 1 b -
Continuous
per 0.75 |ag/d L

0 12 3 4
Odds Ratio (95% CI)
Note: Studies are categorized by Pb biomarker. Within each category, studies generally are presented in order of discussion in the
text, (a) The outcomes plotted are hypertension prevalence with the exception of Yazbeck et al. (20091 which measured pregnancy
induced hypertension and Peters et al. (20071 which measured hypertension incidence, (b) 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. (d) Effect estimates were standardized to a 10 |jg/g increase in bone Pb.
Figure 5-26 Odds ratios (95% CI) for associations of blood and bone Pb with
hypertension prevalence and incidence.
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Table 5-17 Additional characteristics and quantitative data for associations of
blood and bone Pb with hypertension measures for results
presented in Figure 5-26
Study
Population/
Location
Parameter
Pb Data
Statistical Analysis
Effect Estimate
(96% CI)
Martin et al.
964 men and women,
50-70 y, 40% African
American, 55% White,
5% other, in Baltimore,
MD
Hypertension
(current use of
antihypertensive
medication, mean
SBP > 140 mmHg
or DBP > 90
mmHg)
Blood Pb:
Mean (SD): 3.5 (2.3) pg/dL
Tibia Pb:
Mean (SD): 18.8(12.4) pg/g
Logistic regression models
adjusted forage, sex, BMI,
antihypertensive medication use,
dietary sodium intake, dietary
potassium intake, time of day,
testing technician, and serum
homocysteine
Blood Pb level
OR=1.02 (0.87,1.19)
Tibia Pb
OR=1.24 (1.05, 1.47)
mmHg per pg/dL blood Pb
mmHg per pg/g bone Pb
V\feaver et al. 652 current and former Hypertension
Pb workers in South
Korea (12/1999-
6/2001)
(mean SBP >140
mmHg, DBP >90
mmHg; and/or use
of antihypertensive
medications; or
physician
diagnosis)
Blood Pb:
Mean (SD): 31.9 (14.8) pg/dL
Patella Pb:
Mean (SD): 37.5 (41.8) pg/g
Logistic regression models
adjusted forage, gender, BMI,
diabetes, antihypertensive and
analgesic medication use, Pb job
duration, work status, tobacco
and alcohol use
None ofthePb exposure
metrics examined were
(blood, patella, and In
patella) were significantly
associated with
hypertension (results not
reported)
Peters et al.
(2007)
513 elderly men
(mean 67 y) from NAS
in Greater Boston, MA
area
Hypertension
(mean SBP >140
mmHg, DBP >90
mmHg; or physician
diagnosis)
Tibia Pb:
mean (SD): 21.5 (13.4) pg/g
Patella Pb:
Mean (SD): 31.5 (19.3) pg/g
Cox proportional hazards models Hypertension Incidence
adjusted for age, age squared,
sodium, potassium, and calcium
intake, family history of
hypertension, BMI, educational
level, smoking, alcohol
consumption, baseline SBP and
DBP, and physical activity
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
Elmarsafawy et
al. 12006)
471 elderly men
(mean 67 y) from NAS
in Greater Boston, MA
area
Hypertension
(mean SBP > 160
mmHg, DBP >95
mmHg; and/or
physician diagnosis
with current use of
antihypertensive
medications)
Blood Pb:
Mean (SD): 6.6 (4.3) pg/dL
Tibia Pb:
Mean (SD): 21.6 (12.0) pg/g
Patella Pb:
Mean (SD): 31.7 (18.3) pg/g
Logistic regression models
adjusted forage, BMI, family
history of hypertension, history of
smoking, dietary sodium intake,
and cumulative alcohol ingestion
Low calcium group (<800
mg/d):
Blood Pb: 1.07 (1.00, 1.15)
Tibia Pb: 1.02 (1.00,1.04)
Patella Pb: 1.01 (1.00,
1.03)
High calcium group (>800
mg/d):
Blood Pb: 1.03(0.97, 1.11)
Tibia Pb: 1.01 (0.97,1.04)
Patella Pb: 1.01 (0.99,
1.03)
Per pg/dL blood Pb or pg/g
tibia or patella Pb
Yazbecketal.
971 pregnant women,
age 18-45 y, in France
PIH
(SBP >140 mmHg
or DBP > 90 mmHg
after the 22nd wk of
gestation)
Blood Pb:
PIH group mean (SD): 2.2 (1.4)
pg/dL
No PIH group mean (SD): 1.9
(1.2) pg/dL
Q1
Q2
Q3
Q4
<1.20 pg/dL
1.20-1.70 pg/dL
1.71-2.30 pg/dL
>2.30 pg/dL
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) perl
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)
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Study
Population/
Location
Parameter
Pb Data
Statistical Analysis
Effect Estimate
(96% CI)
Muntneretal. 9,961 NHANES (1999- Hypertension
2002) participants
(current use of
antihypertensive
medication, SBP >
140 mmHg, or DBP
>90 mmHg)
Concurrent Blood Pb:
Overall Mean (CI): 1.64(1.59-
1.68) pg/dL
quartile 1: <1.06 pg/dL, quartile
2:1.06-1.63 pg/dL, quartile 3:
1.63-2.47 pg/dL, and quartile 4:
> 2.47 pg/dL
Multivariable logistic regression Adjusted OR
models adjusted for age, sex,
diabetes mellitus, BMI, cigarette
smoking, alcohol consumption,
high school education, and
health insurance status
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)
Significant trend (p=0.04)
Scinicariello et 6,016 NHANES III
al. (2010)	(1988-1994)
participants > 17 y
Hypertension
(current use of
antihypertensive
medication, SBP >
140 mmHg, or DBP
>90 mmHg)
Concurrent Blood Pb:
Mean (SE): 2.99 (0.09) pg/dL
Q1 0.7-1.4 pg/dL,
Q2 1.5-2.3 pg/dL,
Q3 2.4-3.7 pg/dL,
Q4 3.8-52.9 pg/dL
Non-Hispanic Whites: 2.87
(0.09)
Non-Hispanic Blacks 3.59 (0.20)
Mexican American 3.33 (0.11)
Multivariable logistic regression
model adjusted for race/ethnicity,
age, sex, education, smoking
status, alcohol intake, BMI,
serum creatinine levels, serum
calcium, glycosylated
hemoglobin, and hematocrit
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
Non-Hispanic blacks:
Q1 Reference
02	POR=1.83 (1.08, 3.09)
03	POR=2.38 (1.40, 4.06)
04	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)
Q3POR=1.43 (0.61,3.38)
04P0R=1.27 (0.59, 2.75)
ALAD1-2/2-2: POR= 0.49
(0.08, 3.20)
ALAD-1: Reference group
POR for hypertension with
ALAD2 carriers across
quartiles of blood Pb level
also reported. ALAD2
carriers associated with
hypertension in non-
Hispanic whites.
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Study
Population/
Location
Parameter
Pb Data
Statistical Analysis
Effect Estimate
(96% CI)
Park et al.
(2009a)
12,500 NHANES I
(1988-1994)
participants
Hypertension NHANES III Concurrent Blood
Pb
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)
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
OR perSD (0.75 pg/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),
\Nomen <50 yr: 1.23 (1.04,
1.46),
\Nomen >50 yr:1.09 (0.94,
1.26).
aNot included in Figure 5-26 because OR data were not reported.
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.
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 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 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.
These results indicate that circulating Pb (e.g., blood Pb) may act continuously on
systolic BP and reduction 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 relatively young subjects with a low prevalence of hypertension
suggests that at least one of the biological pathways that influences how systolic BP
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responds to Pb operates over a relatively rapid timeframe and 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. This study was strengthened by the analysis of
associations between changes in blood Pb and changes in BP over time within individual
subjects. Bone Pb level may exert influence on blood Pb levels and consequently on BP
in an aging population with prolonged Pb exposure. Thus, there is strength in the
conclusions drawn from this study that inform 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.
In a separate analysis of the same occupationally exposed group in year three of follow-
up, Weaver et al. (2008) examined cross-sectionally associations of concurrent patella Pb
and blood Pb level with systolic BP, diastolic BP, and hypertension and effect
modification by ALAD 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, while 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 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 a p-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 blood Pb level
range of the study than by a linear function.
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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
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 jj.g/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 none of the
confidence intervals for PP referenced to the lowest quintile of tibia Pb (< 7.4 jj.g/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 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 5-27, 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 [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 [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 with those with lower perceived stress (RR of developing
hypertension among those with high stress: 2.66 [CI: 1.43, 4.95] per SD increase in tibia
Pb and 2.64 [CI: 1.42, 4.92] per SD increase in patella Pb). These results provide
evidence supporting adults with higher stress as a population at increased risk of Pb-
associated cardiovascular effects.
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140 n
135 -
4 O High perceived
o> 130 -
E
E 125 -
• » 	
£ mW f	^ High perceived stress
*	• Low perceived stress
115-
	Trend (high stress)
•		Trend (low stress)
110
-10 0 10	30	50	70
Tibia lead (i-ig/gl
90
Source: Peters et al. (2007]
Figure 5-27 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 risk 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 risk of hypertension, particularly among subjects with low dietary calcium.
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. 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 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 cadmium, manganese, and
selenium. 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
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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). Cadmium
and selenium 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 selenium and Pb concentrations
was not significant, and the putative protection effects of selenium through antioxidative
properties were not found in this study.
Wells et al. (2011b) 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 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 standard deviation (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 95%
lower bound confidence limit on the maternal blood Pb level (estimated from cord blood
Pb levels) that was associated with a 1 SD increase in all blood pressure outcomes was
about 1.4 (ig/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 in the more recent
study an order of magnitude lower.
New 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 5.4.3.4, and chronic kidney disease results are discussed in
Section 5.5.2.2. Blood Pb increased regularly with age (geometric means [95% CIs]:
1.28 (ig/dL [1.23, 1.33] in the 18-39 age group to 2.32 (ig/dL [2.20, 2.44] in the 75 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.
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In the NHANES III 1988-1994 population, Scinicariello et al. (2010) found a gene-
environment interaction between blood Pb level and ALAD genotype in relation to SBP
and DBP in a cross-sectional analysis. These interactions varied across race/ethnicity
strata. The strongest associations were observed among non-Hispanic blacks (Figure
5-25, Table 5-16). A statistically significant interaction was observed between concurrent
blood Pb level and ALAD 1-2/2-2b 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 Q2 association for Mexican Americans)
(Figure 5-26, Table 5-17). In addition, non-Hispanic white ALAD2 carriers in the highest
blood Pb level quartile 3.8-52.9 j^ig/dL) had a significantly higher association with
hypertension compared with ALAD 1 homozygous individuals in the highest quartile of
blood Pb. In the same NHANES population, Park et al. (2009a) 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 5-26, Table 5-17).
Associations also were observed with estimated bone Pb.
5.4.2.2 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
5-28 (results summarized in Table 5-18). An array of studies have provided evidence that
long-term Pb exposure (> 4 weeks), resulting in blood Pb levels below 10 (ig/dL can
result in the onset of hypertension (after a latency period) in experimental animals that
persists long after the cessation of Pb exposure (U.S. EPA. 2006b). Tsao et al. (2000)
presented evidence for increased systolic and diastolic BP in rats with blood Pb levels
relevant to those in humans (mean [SD]: 2.15 [0.92] (ig/dL blood Pb; 140 [7] mmHg
systolic BP, 98 [7] mmHg diastolic BP) compared to untreated controls (mean [SD]: 0.05
[0.05] (ig/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 (ig/dL (e.g., mean [SD]: 5.47 [2.1] (ig/dL blood Pb;
143 [6] mmHg systolic BP, 97 [8] mmHg diastolic BP), with the effect leveling off at
higher blood Pb levels.
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200
ISO
£ 150
140
120
100
0
5
10
15
20
25
30
35
40
Blood Pb Level (ng/dL)
Bravo etal. 2007	Chang etal. 1997 — Chang et a I. 2005	Heydarietal. 2006 — Nakhouletal. 1992
^—Rizzi etal. 2009	Rizzi etal. 2009 ^—Tsaoetal. 2000	Zhangetal. 20C8	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 5-28 Changes in BP after Pb exposure (represented as blood Pb level)
in unanesthetized adult rats across studies.
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Table 5-18 Characteristics of studies of blood Pb with BP measures in animals
presented in Figure 5-28
Reference3
Lifestage;
Sex
Exposure
Duration
Exposure Level;
Route
Mean [SEM]b
Blood Pb
Level (kig/dL)
n
ASBP
(mmHg; lowest
blood Pb level
compared with
control)0
Comments
Fiorim et al. (2011)
Adult; M
7 days
4 pg/100 g followed by
0.05 |jg/100 g daily;
intramuscular
9.98 [1.7]
12
16

Nakhoul et al.
(1992)
Adult; M
8 weeks
100 ppm; drinking water
5.3 [3]
7
28
Spontaneously
hypertensive rat
model
Chang et al. (2005)
Adult; M
8 weeks
2% (20,000 ppm) then
removal and measurements
1-7 mo after; drinking water
Range: 4.5-83
5
13.8

Tsao et al. 12000)
Adult
8 weeks
0.01 -2% (100-20,000
ppm); drinking water
Range of means:
2.15 [0.29]-
85.76 [1.29]a
10
13

Rizzi et al. (2009)
Adult; M
8 weeks
30-90 ppm; drinking water
7.6 [1.3], 19.3
[3.4]
11
13.3

Chang et al. (1997)
Adult; M
8 weeks
0.5% (500 ppm); drinking
water
29.1 [0.6]a
10
58

Heydari et al. (2006)
Adult; M
12 weeks
100 ppm; drinking water
26.8 [2.2]
6
25.8

Bravo et al. 12007)
Adult; M
14 weeks
100 ppm; drinking water
23.7 [1.9]a
12
30

Zhang et al. (2009a)
Adult; M
40 weeks
100 ppm; drinking water
28.4 [1.1]a
8-10
15.3

aStudies are presented in order of increasing duration of exposure.
'Standard deviation converted to SEM.
"Difference in systolic BP (SBP) between group means not within one exposure group.
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 j^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 higher blood Pb levels (Rizzi et al.. 2009).
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Studies found that Pb-induced increased BP persisted long after cessation of Pb exposure.
Bagchi and Preuss ("2005) found that elevated systolic BP was maintained for 210 days
after Pb exposure cessation. 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) (Sharif! et al.. 2004). A study utilizing intra-
arterial pressure measurements found that a single high-dose Pb injection (resulting in
mean [SEM] blood Pb levels of 37 [1.7] (ig/dL) increased systolic arterial pressure after
only 60 minutes (Simoes et al. 2011). These studies suggest that there is the potential for
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 5.4.2.3). 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,
5.4.2.3 Hypertension Modes of Action
The 2006 Pb AQCD 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
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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) (Dursiin et al.. 2005;
Attri et al.. 2003; Gonick et al.. 1997; Vaziri et al.. 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.. 1999b); instead it was found to result from inactivation and sequestration of'NO by
ROS (Malvezzi et al. 2001; Vaziri et al.. 1999a). 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 et al. 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..
2001).	The reduction of the vasodilator 'NO 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 (02) and 02" 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 02~ 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:
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23.7 (ig/dL) were also prevented by treatment with the immunosuppressant,
mycophenolate mofetil (MMF) (mean blood Pb level in MMF-treated animals: 27 j^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 the extract
(Badavi et al.. 2008) and vitamin (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, and Pb-induced decreases in serum total antioxidant capacity, high
density lipoprotein (HDL), and heart glutathione peroxidase (GPx); however, internal
doses of Pb were not measured to clarify the mechanism of action in this study (Roshan
et al.. 2011).
Exposure to Pb can also affect the activity and levels of antioxidant enzymes. Male (c)
and female ( ) 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 (>
100 ppm) and increased cardiac thiobarbituric acid reactive substances (TBARS, measure
of lipid peroxidation) (1,000 ppm). Pb exposure in male rats did not affect the activity of
SOD or production of TBARS, 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 exposure (3 100 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 (Rodrimiez-lturbe 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-
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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, NQOl 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
et al.. 2010; Zhang et al.. 2007a; Hevdari et al.. 2006V 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
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 (.iM, 1 hour) decreased ACh-induced
vasorelaxation, which triggers the release of'NO from the endothelial cell, in isolated rat
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 Cordcllini. 2008). However, there was no
change observed in the relaxation response to ACh by Pb alone (Fiorim etal.. 2011; Rizzi
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et al.. 2009; Grizzo and Cordellini. 2008). Conversely, Skoczynska and Stojek ("2005)
found that Pb exposure (50 ppm; blood Pb level 11.2 j^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 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).
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
Alteration of the adrenergic system from Pb exposure, which can increase peripheral
vascular resistance, and thereby arterial pressure, may be one cause 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 et al.. 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 (Simoes 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).
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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 (Skoczvnska and Stoiek. 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 D1 receptors
were found to abolish Pb-induced aortic contraction (Fazli-Tabaei et al.. 2006; Hevdari 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; Hevdari et al.. 2006). Removal of the endothelium blunted the PHE-
induced contraction. Conversely, short-term Pb exposure (7 days) decreased the
contractile response induced by PHE in rat aortas resulting in a decreased vascular
reactivity (Fiorim etal.. 2011). This decrease may be playing a compensatory role in
attempting to correct the Pb-induced BP elevation. Additionally, Pb blunted the
isoproterenol-induced relaxation, supporting a role for the (3-adrenoceptors (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 the preceding
section (Oxidative Stress Response).
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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 kallkrein activities; and renal Angll positive
cells (Rodrimiez-lturbe et al.. 2005; Sharif! et al.. 2004; Carmignani etal.. 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) exposure of Pb to rats increased the plasma ACE
activity (Fiorim et al.. 2011; Simoes etal.. 2011). and Fiorim et al. (2011) additionally
found this increase to be correlated with the Pb-induced increase in systolic BP.
However, at these short time points there were no changes in the Angll receptors 1 or 2
protein levels or expression. Treatment with the Angll receptor (AT, R) blocker, losartan,
or the ACE inhibitor, enalapril, blocked the Pb-induced systolic BP increase (Simoes et
al.. 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 (1% Pb, 40 days; blood Pb level >240 (ig/dL after
exposure, 12-13 (ig/dL after chelation after 1 year) compared to control rats that
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 ofNa+/K+"ATPase,
which plays a major role in Na+ reabsorption and is regulated by the RAAS (Fiorim et al..
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2011; Simoes 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, were
unchanged in rats with Pb-induced hypertension (Gonick et al.. 1998). However, in vitro
Pb exposure promoted the release of the prostaglandin substrate, 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
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 animals after Pb exposure (Valencia et al. 2001; Watts et
al.. 1995; Piccinini et al. 1977). 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- 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 (Purdv et al.. 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.. 2001; Watts et al..
1995; Piccinini etal.. 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 COX derivatives. Pb exposure (1 ppm, 24 hours) to rat aortic segments
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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. This study suggests 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 studies suggest that Pb-induced vascular reactivity may
depend on the participation of a COX-derived vasoconstrictor, such as prostaglandins,
prostacyclins, or thromboxanes.
In summary, a small number of available recent studies continued to show that Pb
exposure affect 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 been shown to interrupt baseline or endogenous "NO-
mediated vasodilation of vessels via alterations in PKC, sGC, VSMC, endothelial cells,
NADPH oxidase, and Ca+2 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 sGC and COX-2.
5.4.2.4 Summary of Blood Pressure and Hypertension
The 2006 Pb AQCD 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 studies comprising over 40,000 adults (Figure 5-29). 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. Recent epidemiologic studies
supported this conclusion 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 jj.g/g. As these studies were mostly cross-sectional in design and were conducted
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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 with increased BP and hypertension (Sections 4.3 and 4.7.3).
A recent study in an ethnically diverse community-based cohort of women and men aged
50-70 years of age found associations of both blood and tibia Pb levels with BP (Martin
et al.. 2006). This study also affirmed findings from other studies by demonstrating that
with each increase of 1 (ig/dL concurrent blood Pb level, systolic BP increased 1 mmHg
and diastolic BP increased 0.5 mmHg and strengthened the evidence for an independent
association with blood Pb level through the extensive examination of potential
confounding. Additionally, recent epidemiologic studies provided evidence for
associations in an adult cohort between blood Pb 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 blood Pb level of 1.64 j^ig/dL (Muntner et al.. 2005V
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. A new 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 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 mediate shorter-term Pb-associated increases in BP and longer-term Pb-
associated development of hypertension.
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 (Figure 5-28). While the contribution of low concurrent blood Pb levels
to the findings is difficult to ascertain in adults, animal toxicological studies provide
support for low blood Pb level effects with increases in BP observed in groups of animals
with 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. In addition, new animal evidence suggests the potential for increased BP
following short-term (4 weeks) Pb treatment that included injected boluses (Fiorim et al..
2011; Simoes et al.. 2011; Sharif! et al.. 2004). New studies also demonstrated partial
reversibility (not to levels in controls) of Pb-induced elevations in BP following Pb
exposure cessation or chelation.
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POCQCK 9L
KRCMHOUT 85
ORSSAUD 65
WEISS 86
DE KORT 87
LOCKETT 87
PARKINSON 87
RABiNOWITZ 87
ELWCOO (CI 88
ELWOOO [HP) B8
ELWOOD (HP) 88
GARTSIDE [W] 88
GARTSIDE IB) B8
GARTSIDE [Wl 88
GARTSIDE IB) S3
NcRI (FWI 88
NERI 88
GRANDJEAN 89
GRANDJEAN 89
REIMER 89
APOSTOLI 90
APOSTOLI 90
MORRIS 90
MORRIS 90
SHARP [Wl 90
SHARP IBI 90
STAESSEN 90
STAESSEN 90
M01LER 92
HEN5E 93
HENSE 93
MAHESWARAN 93
MENDITTO 9*.
PROCTOR 96
STAESSENIP)96
STAESSEN 
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2
increase in tibia Pb and 1.04 (95% CI: 0.96, 1.12) per 10 (ig/g increase in patella Pb were
reported.
First author, year Mean
Lead
(uala)
Increase in SBP (95%CI)
Increase in DBP (95%CI)
Hypertension RR or OR (95%CI)
Tibia lead
Prospective
Glenn 2006^
Cheng 200113
Glenn 2003'*
38.4 -0.02 (-0.03 to 0.004)
21.9 -
14.7 0.78 (0.24 to 1.31)
Overall: 0.33 (-0.44 to 1.11)
1.15 (0.94 to 1.41)
0.07 (-0.30 to 0.45)
Cross-sectional
Lee 200117	37.2	0.20 (-0.05 to 0.45)
Hu 9615/Cheng 0113 32.1	1.01 (0.01 to 2.02)
Martin18 2006 18.8	0.20 (-0.80 to 1.10)
Schwartz^ 2000 14.4	0.74 (-0.73 to 2.21)
Korrick16 1999 13.3	-
Overall: 0.26 (0.02 to 0.50)
Patella lead
Cheng'3 2001 31.4	-
Hu 961B/Cheng 0113 32.1	0.29 (-0.36 to 0.95)
Korrick'61999 17.3	-
i	1	1	r
-10 12
Increase in SBP (mmHg / year)
-10 12
Increase in DBP (mmHg / year)
0B5 1 12 15
Hypertension RR
-0.02 (-0.20 to 0.17)
0.20 (-0.30 to 0.70)
0.35 (-0.75 to 1.45) -
0.02 (-0.15 to 0.19)
1.05 (1.00 to 1.11)
1.15 (0.97 to 1.35)
1.13 (0.98 to 1.29)
0.90 (0.70 to 1.17) <
1.03	(1.00 to 1.05)
1.04	(1.01 to 1.07)
1.14(1.01 to 1.28)
1.09 (0.98 to 1.22)
1.00 (0.98 to 1.03)
1.04 (0.96 to 1.12)
Increase in SBP (mmHg)
Increase in DBP (mmHg)
Hypertension OR
Source: Reprinted with permission of Elsevier Publishers, Navas-Acien et al. (2008)
In the Normative Aging Study, Hu et al. d996b1 reported the cross-sectional association between bone Pb levels and the
prevalence of hypertension and Cheng et al. (20011 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.
Figure 5-30 Prospective and cross-sectional increase in systolic BP (SBP)
and diastolic BP (DBP) and relative risk of hypertension per
10 pg/g increase in bone Pb levels.
3	A few recent epidemiologic studies also emphasized the potential interaction between
4	measures of long-term Pb exposure, i.e., bone Pb levels, and factors such as chronic
5	stress and HFE genetic variants to moderate or modify the relationship of BP and
6	hypertension with Pb. For example, among NAS men, tibia Pb level was associated with
7	a larger risk of developing hypertension in an originally nonhypertensive group among
8	men with higher self-reported stress (Peters et al.. 2007).
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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
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 (Figure 5-25
and Figure 5-26 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 5-25). PP represents a good predictor of cardiovascular morbidity and mortality
and an indicator of arterial stiffness (Zhang et al. 2010a). The aforementioned genes are
related to iron metabolism and have been linked with differences in Pb distribution in
blood and bone. Park et al. (2009^) provided further evidence of variants in iron
metabolism genes impacting the association of bone Pb levels with QT interval changes
(Table 5-19 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.
5.4.3 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; Re vis et al.. 1981). A
limited number of available epidemiologic studies discussed in the 2006 Pb AQCD
provided evidence of associations of blood Pb level with ischemic heart disease (IHD)
and peripheral artery disease (PAD).
5.4.3.1 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-
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term Pb exposure is associated with a number of these conditions, numerous
toxicological studies have investigated and found an effect of Pb on endothelial
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 in workers versus 7.8 (ig/dL in unexposed controls) (Poreba et 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 (.iM. 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
that 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 the rate limiting enzymes in the
regulation of prostaglandins, cytosolic phospholipase A2 (cPLA2) and COX-2. The
induction of these enzymes was mediated by activation of ERK1/2, MEK1, and MEK2.
Further investigation into 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)
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attenuated Pb-induced PGE2 secretion and activation of cPLA2 and COX. A follow-up to
this study found that Pb treatment (1 (.iM) 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 studies suggest that Pb can
induce pro-inflammatory events in VSMC in the form of increased PGE2 secretion and
cPLA2 and COX-2 expression 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; Ueda et al.. 1997; kaii et al.. 1995; kishimoto etal.. 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 cytotoxicity (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 Pb to ECs induced
GRP78 and GRP94 gene (2-25 |_iM) and protein (GRP78 [5-25 |_iM| and GRP94 [10-
25 (.iM |) expression through activation of the IREl-JNk-AP-1 pathways (Shinkai et al..
2010). This study suggests that the functional damage caused by Pb exposure to EC may
be partly attributed to induction of ER stress.
5.4.3.2 Cholesterol
As blood cholesterol rises so does the risk of coronary heart disease. Early occupational
studies (Ademiiviwa et al.. 2005a; Bener et al.. 2001a; kristal-Boneh et al.. 1999)
examining higher than current adult blood Pb levels 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 j^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
(Menke et al.. 2006). In developing models to predict bone Pb levels, Park et al. (2009a)
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noted that total and HDL cholesterol were selected as 2 of 18 predictors for the bone Pb
level model. Their findings suggested that higher Pb exposure may be associated with
higher total and HDL cholesterols. A recent study reported increased LDL and decreased
HDL in rats treated with Pb (20 ppm, i.p., 3 days/week, 8 weeks) (Roshan et al.. 201IV
The major risk factor that lipids represent for heart disease make relating lipid levels to
Pb exposures an interesting but challenging hypothesis to test.
5.4.3.3 Heart Rate Variability
Pb has been shown not only to affect vascular contractility in animals, but also is
associated with cardiac contractility. Eum et al. (2011) and Park et al. (2009b) followed
up a previous NAS report (Cheng et al.. 1998). which found increasing duration of
corrected QT interval (QTc) with increasing bone Pb levels in men <65 years, but not in
men > 65 years. Eum 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] jj.g/g), and tibia bone (mean [SD]: 21.6 [12.0] jj.g/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 jj.g/g), participants in the highest tertile (> 23 jj.g/g) had a
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. There were no statistically significant associations with patella or blood Pb
levels. These associations 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 j^ig/dL) compared to control subjects (Poreba et al.. 201 la).
Park et al. (2009b) examined whether polymorphisms in genes known to alter iron
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
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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 for both C282Y and H63D had
lower bone Pb levels. The 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 jj.g/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 variant 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 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 >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 LF/HF was linear with log(LFZHF).
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 j^ig/dL) (Reza et
al.. 2008V 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
found that Pb (100 (.iM) exposure, in a concentration-dependent manner, reduced
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myocardial contraction using rat right ventricular strips 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
Ca: content (Lai et al.. 1991; Favalli et al.. 1977; Piccinini et al.. 1977). which could
result in increased cardiac output and hypertension.
5.4.3.4 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. 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 (ig/dL) was 1.92 (95% CI: 1.02, 3.61). These results are consistent with those
from a previous NHANES analysis conducted by Navas-Acien et al. (2004).
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 but 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. 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
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illustrates the limited reliability of spot urine Pb measurements compared to blood Pb
measurements.
5.4.3.5 Ischemic Heart Disease
A few studies discussed in the 2006 Pb AQCD 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) reported on the incidence of IHD
(physician confirmed MI, angina pectoris) among older adult males enrolled in the NAS
and followed during the period of September, 1991 to December, 2001. 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 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
contribute independently to IHD incidence.
IHD, characterized by reduced blood supply to the heart, may result from increased
thrombosis. 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. 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; rat RBCs: 5 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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5.4.3.6
Atherosclerosis
Epidemiologic and toxicological studies provide evidence for increased atherosclerosis
and intimal medial thickening (IMT) due to Pb exposure. The association between 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 (95% CI) of 1.02 (1.00, 1.03). Urine Pb was 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 et al.). 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. Porcba et al. (201 lb) reported increased local arterial stiffness and more
frequent left ventricular diastolic dysfunction in Pb-exposed workers with hypertension
compared to nonexposed controls with hypertension. Occupational exposure to Pb (mean
blood Pb levels: 24 j^ig/dL in workers, 8.3 (ig/dL in nonexposed 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
(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 (.iM, 72 hours) and VSMC elastin expression (50 |_iM. 12 hours), as discussed
above (Section 5.4.3.1) (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.
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Table 5-19 Characteristics and quantitative data for associations of blood and
bone Pb with other CVD measures in epidemiologic studies ordered
as they appear in the text
Study
Population/
Location
Parameter
Pb Data
Statistical Analysis
Effect Estimate (96% CI)'
Heart rate variability
Eum et al. (2011)
600 men free of
electrographic
abnormalities at the
time of baseline ECG
from NAS 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) pg/dL
Baseline Patella Pb:
Mean (SD): 30.3 (17.7) pg/g
Baseline Tibia Pb:
Mean (SD): 21.6 (12.0) pg/g
Repeated measures linear
regression adjusted for 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.
Q1
Q2
Q3
<16 pg/g (n = 191)
16.0-23 pg/g (n = 208)
>23 pg/g (n = 195)
Tibia Pb:
Adjusted 8-year change (95%
CI):
QTc:
Q2 vs. Q1 (reference): 7.49
(1.22,13.75) msec,
Q3 vs. Q1: 7.94 (1.42, 14.45)
msec
p for trend = 0.03
QRSc:
Q2 vs. Q1: 0.52 (-3.60, 4.65)
msec
Q3 vs. Q1: 5.94 (1.66, 10.22)
msec
p for trend = 0.005
No associations with patella or
blood Pb
Park et al.
613 men from NAS in
Greater Boston, MA
area (8/1991 -
12/1995)
QTc interval Baseline Blood Pb:
Median (IQR): 5 (4-7) pg/dL
Baseline Patella Pb:
Median (IQR): 26 (18-37) pg/g
Baseline Tibia Pb:
Median (IQR): 19(14-27) pg/g
Linear regression models
adjusted forage, BMI, smoking
status, serum calcium, and
diabetes.
Per IQR (3 pg/dL) increase in
blood Pb
1.3 (-0.76, 3.36) msec after 8-
year follow up
Per IQR (19 pg/g) increase in
patella Pb
2.64 (0.13, 5.15) msec
Per IQR (13 pg/g) increase in
tibia Pb
2.85 (0.29, 5.40) msec
Park et al. (2006)
413 men from NAS in
Greater Boston, MA
area (11/14/2000-
12/22/2004)
HRV
(SDNN, HF,
HFnorm, LF,
LF norm,
LF/HF)
Baseline Patella Pb (measured
within 6 mo of HRV:
Median (IQR): 23.0 (15-34) pg/g
Estimated8: Median (IQR):
16.3(10.4-25.8) pg/g
Baseline Tibia Pb:
Median (IQR): 19.0(11-28) pg/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 p-biockers,
calcium channel blockers,
and/or ACE inhibitors
Tibia Pb: Model 2
Change (95%CI)
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 17 pg/g tibia Pb
Patella Pb: Model 2 Change
(95%CI)
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) (%)
Per 15.4 pg/g patella Pb
Peripheral artery disease
Effect estimates were more
pronounced among those with
metabolic syndrome.
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Study
Population/
Location
Parameter
Pb Data
Statistical Analysis
Effect Estimate (96% CI)'
Muntneretal. (2005) 9,961 NHANES(1999-
2002) participants
PAD
Range Concurrent Blood Pb:
Q1
Q2
Q3
Q4
<1.06 pg/dL,
1.06-1.63 pg/dL
1.63-2.47 pg/dL
>2.47 pg/dL
Logistic regression models
adjusted for age, race/ethnicity,
sex, diabetes mellitus, BMI,
cigarette smoking, alcohol
consumption, high school
education, health insurance
status
OR (95% CI):
1.00 (Reference),
1.00 (0.45, 2.22),
1.21 (0.66, 2.23),
1.92 (1.02, 3.61)
Navas-Acien et al.
790 participants, age > PAD	Concurrent urinary Pb:
S™HANES	Mean (10th-90th %): 0.79 pg/L
(iyyy-2uuu)	(0 2-2 3)
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
Model 1
Model 2
Model 3
1.17(0.81, 1.69)
1.17 (0.78, 1.76)
0.89 (0.45, 1.78)
Per IQR increase in urinary Pb
Array of metals in urine also
evaluated.
Ischemic Heart Disease
Jain et al. 12007)
837 men from NAS in
Greater Boston, MA
area (1991-2001)
IHD
(Ml or angina
pectoris)
Baseline Blood Pb Mean (SD):
Non-cases 6.2 (4.3) pg/dL;
Cases 7.0(3.8) pg/dL
Baseline Patella Pb Mean (SD):
Non-cases 30.6 (19.7) pg/dL;
Cases 36.8(20.8) pg/dL
Baseline Tibia Pb Mean (SD):
Non-Cases 21.4 (13.6) pg/g;
Cases 24.2 (15.9) pg/g
Cases:
Blood Pb range: 1.0 to 20.0 pg/dL
Patella Pb range: 5.0 to 101 pg/g
Tibia Pb range: -5 to 75 pg/g
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 pg/dL
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)
Per 1 SD increase in Pb
biomarker
Estimated patella Pb accounts for declining trend in patella Pb levels between analysis of bone Pb and HRV.
bHeart-rate-corrected QT interval calculated by Bazett's formula
cIVCD, intraventricular conduction defect; AVCD, atrioventricular conduction defect
5.4.3.7 Summary of Vascular Effects and Cardiotoxicity
There are a limited number of studies that investigate the associations between Pb
biomarkers and cardiovascular effects other than BP or hypertension (Table 5-19). As
presented in Table 5-19, 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, 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. 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 other vascular effects and cardiotoxicity. A recent study provided evidence for the
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interaction between biomarkers of Pb exposure and the HFE C282Y and HMOX-1 L
variant on the prolonged QT interval in nonoccupationally-exposed older men (Park et
al.. 2009b). Also, in the NAS population, bone Pb levels were associated with larger
decreases in HRV parameters among subjects 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.
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 (Eum et al.
2011). 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-Acicn
et al.. 2004). In addition, in the NAS cohort of older adult men, blood Pb (> 5 (ig/dL) and
patella Pb levels were associated with increased morbidity from IHD (Jain et al. 2007).
A recent study involving both human and toxicological studies elucidated mechanisms
for observed Pb-mediated arterial IMT, an early event in Pb-induced atherogenesis
(Zeller et al.. 2010). 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. Overall, the relatively
available few studies provide support for associations between Pb biomarkers and other
cardiovascular conditions, yet further research is warranted to understand these
relationships.
5.4.4 Cardiovascular Function and Blood Pressure in Children
5.4.4.1 Introduction
The study of cardiovascular function effects in relation to blood Pb levels in children
potentially offers unique information on several topics. First, by examining endpoints
predictive of future cardiovascular pathology, these studies may offer information on the
potential cardiovascular effects of Pb exposure in an understudied population. Second,
examination of cardiovascular changes that are antecedent to increased BP and changes
in other CVD-related endpoints at later lifestages may inform uncertainties in regards to
the time course of cardiovascular changes associated with Pb exposure. Finally, these
studies address gaps in knowledge regarding Pb exposure effects in populations with
mean blood Pb levels in the range of < 5 to 10 (ig/dL.
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An advantage of examining the literature about the association between cardiovascular
effects and blood Pb levels in children is that the blood Pb levels of children in new
studies may better reflect their relatively recent Pb exposure and its effect on CVD than
blood Pb levels do in adults because of the uncertainties related to the effects of earlier
higher Pb exposure levels. A possible disadvantage of studying the associations of blood
Pb levels and cardiovascular effects in children is the lower prevalence of such conditions
in children. For example, the prevalence for hypertension in children (9 to 10 years old)
ranges from to 2 to 5 percent (Daniels. 2011; Steinthorsdottir et al. 201IV The
prevalence of hypertension increases with age. More than half of people aged 60 to 69
years have hypertension (Chobanian et al. 2003). Further, 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 lower disease prevalence may
make it more difficult to find a relationship between blood Pb level and cardiovascular
effects in child studies. Although studies in children may be limited with regard to study
power in assessing associations with present CVD pathology, they provide for the
examination of early function changes that are associated with subsequent CVD.
The limited numbers of studies published on children examined endpoints such as total
peripheral resistance (TPR), BP, and autonomic nervous system activation. These new
and earlier studies are presented in Table 5-20. Multiple 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 et al.. 2011; Gump et al..
2009; Gump et al.. 2007; Gump et al.. 2005). Zhang et al. (In Press) examined children in
Mexico City born from 1994 to 2003, when Pb was being taken out of gasoline in
Mexico (Martinez et al.. 2007). The geometric mean cord and concurrent blood Pb levels
of the children in the Mexico City cohort were 4.67 and 2.56 (ig/dL.
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Table 5-20 Studies of Children Cardiovascular Endpoints and Pb Biomarkers
Ordered as They Appear in the Text
Study
Population/
Location
Parameters
Blood Pb Dataa
Statistical Analysis
Effect Estimates/Results
Gumpetal. 122 children age	SBP, TPR (total
(2005)	9.5 yr in Oswego,	peripheral
NY (born at a single	vascular
hospital in New	resistance)
York from 1991-94)
Cord blood Pb:
GM (GSD):
2.56 |jg/dL (1.16)
Childhood (mean age of
measurement: 2.6 yr)
blood Pb: GM (GSD):
4.06 |jg/dL (1.14)
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 pg/dL increase in childhood blood
Pb level, 0.088 (95% CI: 0.023, 0.153)
dyne-s/cm5 change in TPR
Per 1 pg/dL increase in cord blood Pb
level, 12.16 (95% CI: 2.44, 21.88)
mmHg higher SBP
Gumpetal. 122 children age SBP, TPR
12007)	9.5 yr in
Oswego, NY
Childhood (mean age of
measurement: 2.6 yr)
blood Pb:
GM (GSD):
4.06 pg/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 pg/dL, SES not
significantly associated with SBP
Gumpetal.
122 children age Salivary Cortisol Cord blood Pb:
®;5yrin MV	GM (GSD):
Oswego, NY	2.56 pg/dL (1.16)
Childhood (mean age of
measurement: 2.6 yr)
blood Pb: GM (GSD):
4.06 pg/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. 12007)
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, 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
Gumpetal.
(2011)
140 children
ages 9-11 yr
SBP, TPR, HRV
(heart rate
variability) in
response to acute
stress (mirror
tracing task)
Concurrent blood Pb:
GM: 1.01 pg/dL
Quartiles:
0.14-0.68 pg/dL
0.69-0.93 pg/dL
0.94-1.20 pg/dL
1.21-3.76 pg/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: 01: 5.30, 02:7.33,03:7.07,
04:7.23, p for trend = 0.31
Change in TPR (%) across quartiles:
01:2.91,02:8.18,03:9.55, Q4: 9.51,
pfor 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
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Study
Population/
Location
Parameters
Blood Pb Dataa
Statistical Analysis
Effect Estimates/Results
Zhang et al. (In
Press)
457 mother child SBP
pairs in a birth
cohort, born 1994 to
2003 in Mexico City.
Children were
evaluated
2008-2010 at ages
7-15 yr
Cord blood Pb:
GM (GSD): 4.67 pg/dL
(1.18) (N=323)
Concurrent blood Pb:
GM (GSD):
2.56 |jg/dL (1.16) (N=367)
Maternal post-partum
bone Pb:
Median (IQR):
Tibia Pb: 9.3 (3.3,
16.1) pg/g
Patella Pb: 11.6(4.5,
19.9) pg/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 pg/g) increase
in maternal tibia Pb was associated with
a 2.11 (95% CI: 0.69, 3.52) mmHg
increase in SBP
IQR (16 pg/g) increase in maternal
patella Pb was associated with a 0.87
(95% CI: -0.75, 2.49) mmHg increase in
SBP
IQR (4 pg/dL) increase in cord blood Pb
was associated with a 0.75 (95% CI: -
1.13, 2.63) mmHg increase in SBP
Factor-Litvak et 260 children ages SBP
al. (1999; 1996) 5.5 years old in K.
Mitrovica and
Pristina, Yugoslavia
Concurrent blood Pb range:
4.1 to 76.4 pg/dL
Linear regression analysis.
Potential confounders
considered: gender, maternal
education, birth weight, HOME
score, and BMI.
Per 1 pg/dL increase in concurrent blood
Pb level, 0.05 (95% CI:-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) Young adults age BP
19-29 years, born
1965-1975, male
and female; half of
the subjects had
grown up around an
active Pb smelter in
Silver Valley, Idaho
While the concurrent mean
blood Pb level was 3.15
pg/dL for the highest bone
Pb category (> 10 pg/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 pg/dL among
participants with bone Pb
level >10 pg/g.
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 pg/g to >10 pg/g)
were tested together.
Group in highest quartile of tibia Pb level
(>10 pg/g) had 4.26 (95% CI: 1.36,
7.16) mmHg higher SBP and 2.80 (95%
CI: 0.35, 5.25) mmHg higher DBP
compared to the lowest tibia Pb group (<
1 pg/g).
aBlood Pb data are estimates of geometric mean (GM) and geometric standard deviation (GSD) using the arithmetic mean and SD.
5.4.4.2 Cardiovascular Functioning in Children
The relationship between cardiovascular functioning (TPR, BP, stroke volume, and
cardiac output,) and blood Pb levels was examined by Gump et al. (2007; 2005) in a
cohort 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 5-31. Testing blood Pb with
linear, quadratic, and cubic terms did not produce significantly different Pb-TPR
associations which the authors suggested showed effects that 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. Biological plausibility for these observations in children is
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provided by observations that Pb exposures increases TRP in toxicological studies and by
mechanistic evidence indicating that Pb-induced changes in SNS activity may mediate
such effects (Section 5.4.2.3). Additionally, higher blood Pb level measured at age 2.6
years was associated with a smaller stroke volume and cardiac output responses to acute
stress at age 9.5 years. 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.
Blood Lead level (ng/dL)
Source: Reprinted with permission of Elsevier (Gump et al., 2005)
Figure 5-31 Children's adjusted total peripheral resistance (dyn-s/cm5)
responses to acute stress tasks, as a function of 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 to that used in the previous studies to
examine the 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). This result is consistent with those of
Gump et al. (2005). Also, these new findings provided evidence of associations with
concurrent blood Pb levels and with lower blood Pb levels (Gump et al.. 2011) than were
previously examined in Gump et al. (2005).
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Studies examining HRV in adults and animal toxicology are discussed in Section 5.4.3.3.
In Gump et al. (2011). cardiac autonomic regulation decreased in a
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 (Gump et al. 2007; Gump et al.. 2005). In this new
study, Gump et al. (2011) found the same but for concurrent blood Pb level and at lower
blood Pb levels.
5.4.4.3 Blood Pressure in Children
Zhang et al. (In Press) conducted a longitudinal study that examined changes in BP in
323 girls and boys aged 7 to 15 years old in a Mexico City cohort for a relationship 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 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 of 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 jj.g/g (IQR: 3.3,
16.1 jj.g/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 Bagbv. 2005; te Velde et al.
2004; Barker et al.. 1989). 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. The relationship between birth weight and Pb
biomarkers is discussed in Section 5.8.7.
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.
(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 blood
Pb levels between 1.21 and 3.76 j^ig/dL had a 7.23 mmHg change, and children with
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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. 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 5.4.2.1). 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 young adults (ages 19-29 years) in the
highest category of bone Pb levels (greater than 10 jJ-g/g), in whom the mean concurrent
blood Pb level was 65 (ig/dL, systolic BP was 4.26 mmHg higher compared with young
adults with bone Pb levels < 1 (ig/dL.
The pathogenesis of CVD has been hypothesized to begin in childhood (kapuku et al..
2006). Early markers observable in youth include increased blood pressure during stress,
reduced heart rate variability, 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 et al.
(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 of 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 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 j^ig/dL) and increased arterial
systolic BP after weaning (Grizzo and Cordellini. 2008) suggesting a role for childhood
Pb exposure leading to adult disease.
5.4.4.4 Summary of Child Cardiovascular Studies
Studies have reported antecedent cardiovascular changes such as TPR responses to acute
stress tasks as a function of childhood blood Pb levels. Also, a study reported associations
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with acute stress-induced autonomic and cardiovascular dysregulation responses.
Biomarkers of prenatal Pb exposure (maternal post-partum patella and tibia) were related
to later higher BP. This may be related to intrauterine growth restriction and subsequent
accelerated weight gain in childhood and 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 of age of onset of puberty, and cross-
sectional design. Some of these uncertainties result in random error, however, and result
in the attenuation of observed associations rather than the generation of spurious
associations. These study findings indicate that in children with mean blood Pb levels in
the range of < 5 to 10 (ig/dL, increasing blood Pb level may be associated with small
increases in BP and changes in the cardiovascular system that may be related to later
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
considering the lower prevalence and strength of compensatory mechanisms in children.
There is uncertainty in the shape of the concentration-response relationship to
cardiovascular endpoints at lower blood Pb levels since most studies modeled a linear
relationship. Several studies that compared linear and nonlinear relationships between
blood Pb level and decrements in cognitive function found a better fit for the nonlinear
relationship (See Section 5.3.10).
Cardiovascular endpoints other than baseline BP may be more sensitive outcomes by
which to measure Pb-associated cardiovascular effects in very 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 outcomes such as TPR related to acute stress. It
suggests 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
cardiovascular outcome in children is an important factor to consider in the design of
studies. Rather than using indicators of already present cardiovascular problems, such as
BP, informative evaluation of cardiovascular changes that are antecedent to increased BP
and changes in other CVD-related endpoints at later lifestages may reduce uncertainties
regarding the time course of cardiovascular changes associated with early Pb exposure
where the relevance and persistence of these endpoints has been shown to be associated
with future pathology.
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Overall this body of evidence, based on different cohorts, locations, and study designs,
begins to form a literature base that indicates there is 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 blood pressure increases are more prevalent in
older adults than in children, BP increases have been related to higher blood Pb level in
older studies of children and young adults (Gerr et al.. 2002; Factor-Litvak et al.. 1999;
Factor-Litvak et al.. 1996). The newer children studies provide information in
populations with mean blood Pb levels in the range of < 5 to 10 (ig/dL for BP and
antecedents for CVD such as increases in TPR and changes in cardiac autonomic
regulation.
5.4.5 Mortality
The 2006 Pb AQCD stated that collectively the then 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 the association of
adult concurrent blood Pb with all cause and cause-specific mortality (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 Pb exposure in their past than
during the study period. Using NHANES II (1976-1980) data, Lustberg and Silbergeld
(2002) found significant increases in all-cause, circulatory and cancer mortality,
comparing adults with blood Pb levels (measured 12-16 years before ascertainment of
vital status) of 20-29 (ig/dL to those with blood Pb levels less than 10 (ig/dL. Using
NHANES III data, Schober et al. (2006) found significant increased all-cause mortality,
cardiovascular, and cancer mortality comparing adults with blood Pb levels (measured a
median of 8.8 years before ascertainment of vital status) from 5-9 (ig/dL and above
10 (ig/dL to those with blood Pb levels less than 5 (ig/dL.
Several new 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 other earlier analyses by
considering a more extensive number of potential confounding factors and by
characterizing concentration-response relationships. A few longitudinal prospective
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studies in different cohorts conducted by different researchers with different methods in
different parts of the U.S. provide coherence with 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 (Figure 5-32).
The authors reported that there were no interactions between blood Pb and other adjusted
variables. In the previous NHANES III analysis of the association of blood Pb with
mortality, Schober et al. (2006) included participants greater than 40 years of age (N =
9686) 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)
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 and
generally statistically nonsignificant higher HRs 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 ail-cause mortality (95% CI)
Hazard ratio of cardiovascular mortality (95% CI)
<60
! * 1 25
n441

1 M (1 08
1 58'
FMce r


Nun Hisp me white
1 3? >1 0^
1.60)
N n Hispnrtit. biark
1 23 <0 99
1.52)
f/oxt^an American
1 17 «C 86
1.60)
anomenupai s»a ctiks



1 41 ii 11
1.78)
Fern k
1 ?4 ,1 00
1.54)
P p iiih ipat saf
1 0?(0 54
1.95r~
P V-4 iTPrif
1 ?4 n 00
154)
Residence


Rural
1 28 it 05
1.54)
Urban
1 42 i1 18
1.72)
Smoking


Never
I 21 *0 93
1 58/
former
1 61 H.33
1 94)
Cut rent
1 34 f0 96
1 87i
Body mass index (kg/m2)


<25
1 51 * 1 10
1 96)
>=25
1 \ \ 03
1 S?)
Hypertension


No
1 31 1 08
1,58)
Yes
1 32*1 00
1.60)
Diabetes


No
1 37 i J
158)
Yes
1 12 iO 73
1.71}
Estimated glomerular filtration


rate (ml/min/1.73m2)


< 60
1.44(1.01 -
• 2.06)
>»• 80
1.32(1.12-
• 1.56)
Overall
1.34(1,16
1.54)
0.5
?4	3??'
! 49 .1 12	1 9^
1 49 (t 12	1 99*
1 13 m3 79	1 &V
1 55 ?^
1 41 -t01
1 /5 p *9
1 Oh;
,2 56.
2 3
1 5' it 10	2 £4)
a 49	? a ^
105(0 54	2 04)'
? or «\ 32	Tin
1 04 sU 94	1
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levels also have been found for blood Pb-cognitive function relationships in children
(Section 5.3.10).
2.0 i
w
0)
c
=• 1.5
(0
cc
CO
N
fl5
X
1.0
0.7
		All cause
		Myocardial infarction
	Stroke
		Cancer
" 10
"t/T
OJ
si
c
o
TO
3
Q.
O
CL
"5 O
0)
U>
re
4*
c
5
o
0
Q.
0
Blood Lead, |Jg/dL
Source: Reprinted with permission of Lippincott Williams & Wilkins, Menke et al. (20061
Note: A histogram of blood Pb levels is superimposed in the background and displayed on the right axis.
Figure 5-33 Multivariate-adjusted relative hazard (left axis) of mortality
associated with blood Pb level between 1 pg/dL and 10 pg/dL.
Schober et al. (2006) examined proportional hazard assumptions, tested for linear trend
across blood Pb tertiles, and evaluated log-transformed continuous blood Pb level as
a 5-knot cubic spline (position of knots not reported). A statistically significant increasing
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 5-34. 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). Dashed lines are 95% CIs. The hazard
ratio was fixed at 1.0 for the referent blood Pb level of 1.5 (ig/dL. In this study, hazard
ratios were less than one in the lower range of blood Pb levels. Despite differences in the
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age groups included, follow-up time, categorization of blood Pb levels, and differences in
hazard ratio across the blood Pb range, results reported by Menke et al. (2006) and
Schober et al. (2006) are consistent in finding associations between higher blood Pb and
increased CVD mortality.
2.0
1.8
1.6
"2	1.4
co
J	1.2
a>	1.0
jo 0.8
a>
oc 0.6
0.4
0.2
0.0
Source: Schober et al. (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 5-34 Relative risk of all cause mortality for different blood Pb levels
compared with referent level of 1.5 |jg/dL (12.5th percentile).
In addition to the NHANES analyses described above, studies of older adult males
(Weisskopf et al.. 2009) and older adult females (Khalil. 2010: Khalil et al.. 2009a) 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 death 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


\
\

\
-	
" ~7
/

/

i
¦ i i i i i i 			 i i i i
1.0 2.0 3.0 4.0 5.0 6.0 7.0 8.0 9.0 10.0
Blood lead (ug/dL)
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21
to the NHANES analyses, baseline blood Pb was not significantly related to
cardiovascular mortality in the NAS study. 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. The youngest subjects at baseline in the Weisskopf
et al. (2009) study were approximately 50-55 years old, compared to the youngest in the
Menke et al. (2006) and Schober et al. (2006) 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 variable,
which can influence these results, were determined by an iterative best fit procedure.
Concentration-response relationships shown in Figure 5-35 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 jj.g/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.
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All-cause
All cardiovascular
MIIMIimiMIIIIIIIIIII HIIIIM I I
iiimiiniiniiiiiii mil nun i. i
0
50	100
Patella lead, pg/g
0
50	100
Patella lead, pg/g
Ischemic Heart Disease
0
50	100
Patella lead, pg/g
Source: Reprinted with permission of Lippincott Williams & Wilkins, Weisskopf et al. (2009)
Note: The reference logHR = 0 at the mean of patella Pb concentration. The estimates are indicated by the solid line and the 95%
pointwise 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.
Figure 5-35 Associations between patella bone Pb level and the log of HR
(logHR) for all-cause, cardiovascular, and ischemic heart disease.
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.. 2009a). This
prospective cohort (N = 533) enrolled female volunteers 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 comparing women with blood Pb
levels >8 j^ig/dL to those with blood Pb levels <8 j^ig/dL was significantly increased (HR:
1.59 [95% CI: 1.02, 2.49]). 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]) HR was increased 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
5-36. 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).
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*4" -
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36
37
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 provide confirmation of 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. The Khalil et al. (2010; 2009a) SOF provides
supporting results in a different study cohort consisting of white females aged 65-87
years. Further, a number of prior studies have found association between accumulated Pb
reflected in bone Pb measurements and higher CVD morbidity (Sections 5.4.2.1 and
5.4.3). 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, with few exceptions
associations between higher levels of Pb biomarkers and higher risk of mortality were
consistently observed (Figure 5-37 and Table 5-21). In studies that examined mortality
from specific CVD causes, MI, stroke, and IHD mortality, which are causes related to
higher BP and hypertension were all significantly elevated with higher Pb biomarker
levels.
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Reference
Menkeetal. (2008)
n = 13,946
mn age-44
mn blood °b=2,58
Schober et al. (2006)
n = 975?
age >40y
Luslberf & Sllbergeld
120021, n = 4,l»
mn agc54 y
mn blood Pb=14,0
Weisskopf et al. (2009)
n = 868 men
mn age=67.3 y
mn blood Pb 5.7
Khalil et at. (2009)
ii = 533 women
mn age=?0 y
mr» blood Pb=5.3
Outcome
All Cause
CVD
Ml
Stroke
All Cause
CVD
All Cause
CVD
All Cause
CVD
IHD
Alt Cause
CVD
IHD
All Cause
CVD
IHD
All Cause
CVD
CHD
Stroke
Study
NHANES I
N BANES II
NAS
SOF
1986-1988
Pb Biomtrfcer Comparison
Groups
Blood Pb (iig/dl)
&3 63 vs. <1.9 i
1.94-3 62 vs <1.93
>3 63 vs. <1.93
1.94-3 62 v» <1 93
>3.63 vs <1 93
1.94-3 bi Vi il.93
S3.63 vs. Sl.93
1.94-3.62 vs. Sl.93
>10 vs. < 5
5-9 vs.« 5
210 vs. < 5
5-9 vs. <5
20-29 vs. < 10
10-19 vs. <10
20-29 vs. <10
10-19 vs. <10
>6 vs. < 4
4-6 vs. < 4
>6 vs. <4	—
4-6 vs. <4	—
>6 vs. <4	-
4-6 vs. <4	-
Tibia Pb (ng/g)
Tertite 3 vs. 1 (NR|
fertile 3 vs. 1 (NR)
Tertite 3 vs. 1 (NR)
Patella Pb (m/gt
>	35 vs. < 22
22-3S vs. < 22
>35 vs. <22
22-35 vs. <22
>	35 vs. < 22
22-35 vs. <22
a 8 vs. < 8
>8 vs. <8
i 8 vs. < 8
4 8 vs. < 8	-
ro~

0.0 1.0 2.0 3.0 4.0 5.0 §.0 7.0 §.0 9.0 10.0
Hazard Ratio <95* CI)
Note: Studies are presented in order of strength of study design and follow the order of discussion in the preceding 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).
Figure 5-37 Hazard ratios for associations of blood Pb (closed markers) and
bone Pb (open markers) with all-cause mortality (black diamonds)
and cardiovascular mortality (blue circles).
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Table 5-21 Additional characteristics and quantitative data for associations of
blood and bone Pb with CVD mortality for studies presented in
Figure 5-37
Study
Population/
Location
Parameter
Pb Data
Statistical Analysis
Hazard Ratio or SMR
(95% CI)
Menke et al.
13,946 adult
participants of
NHANES III,
> 17 yr (1988-1994)
All cause and cause-specific
mortality
(through 2000)
CVD:ICD-9 390-434; ICD-10
100-199), Ml (ICD-9 410-414 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 pg/dL
Tertiles:
<1.93 pg/dL,
1.94-3.62 pg/dL,
> 3.63 pg/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: urban
residence, cigarette smoking, alcohol
consumption, education, physical
activity, household income,
menopausal status, BMI, CRP, total
cholesterol, diabetes mellitus, Model
3: hypertension, GFR category
All-cause (3rd vs. 1st
fertile):
1.25(1.04,1.51)
CVD (3rd vs. 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)
Schober et al.
9,686 adult
participants of
NHANES III,
>40 yr
All cause and cause-specific
mortality
Ordered categorical
blood Pb level,
measured a median
of 8.55 yr prior to
death
<5 pg/dL
5-9 pg/dL
>10 pg/dL
Survey-design adjusted Cox
proportional hazard adjusted for sex,
age, race/ethnicity, smoking, education
level
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 (3rd vs. 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)
Lustberg and
Silbergeld (2002)
4,190 adult
participants of
NHANES III, yr
(1976-1980)
All cause and cause-specific
mortality
Categorical blood
Pb level
Mean: 14.0(5.1)
Median: 13 pg/dL
<10 pg/dL
(Reference)
10-19pg/dL
20-29 pg/dL
Crude mortality 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)
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Study
Population/
Location
Parameter
Pb Data
Statistical Analysis
Hazard Ratio or SMR
(95% CI)
VNfeisskopf et al. 868 men, >55 yr,
95% white, from
NAS in Greater
Boston area, MA
All cause and cause-specific
mortality
Pb biomarkers
collected an
average of 8.9
years before death
Blood Pb:
Mean (SD): 5.6
(3.4) pg/dL
Patella Pb:
Mean (SD): 31.2
(19.4) pg/g
Tertiles:
<22 pg/g,
22-35 |jg/g,
>35 |jg/g
Tibia Pb:
Mean (SD): 21.8
(13.6) pg/g
Cox proportional hazard regression
analysis adjusted for age, 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 (3rd vs. 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 (3rd vs. 1st):
1.30 (0.54, 3.17)
Khalil et al.
(2009a)
533 women, 65-87
yr, from Study of
Osteoporotic
Fractures cohort in
Baltimore, MDand
Monongahela
Valley, PA
All cause and cause-specific
mortality
Blood Pb measured Cox proportional hazards regression > 8 pg/dL vs. <8 pg/dL
an average 12 (SD; analysis adjusted for age, clinic, BMI,
3) yr before death: education, smoking, alcohol intake,
All ran
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Study
Population/
Location
Parameter
Pb Data
Statistical Analysis
Hazard Ratio or SMR
(95% CI)
aNeuberger et al. Residents at or near Cause-specific mortality
(2009)	Tar Creek
Superfund site,
Ottawa County, OK
(exposed pop.
5,852, unexposed
pop. 16,210)
Not reported
Standardized mortality ratio (SMR)
based on 2000 U.S. Census data
Heart disease:
Both sexes:
114.1 (113.1,115.2)
Men:
118(116.4,119.6)
Women:
111 (109.5,112.5)
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.
(2007)
933 male Pb
smelter workers
from Sardinia, Italy
(1973-2003)
All cause and cause-specific Not reported
mortality
SMR
All cause: 56 (46, 68)
CVD: 37 (25, 55)
'These references not included in Figure 5-37 because they reported standardized mortality ratios.
5.4.6 Air Lead-Particulate Matter Studies
5.4.6.1 Cardiovascular Morbidity
A relatively small number of studies used Pb measured in PM10 and PM2 5 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 3.5.3). Moreover, data illustrating the relationships of Pb-
PMio and Pb-PM2 5 with blood Pb levels are lacking. A few available studies exposed
rats, dogs, or humans to concentrated ambient 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, exposure to Pb-containing CAPS
resulted in various changes related to increased vasoconstriction (Urch et al.. 2004;
Wellenius et al.. 2003; Batalha et al.. 2002). Whereas studies of Pb biomarkers primarily
found cardiovascular effects with indicators of long-term Pb exposure, studies of Pb-
containing CAPS provide evidence for cardiovascular effects with short-term exposure
(2-6 hours over multiple days) It is important to note that Urch et al. (2004) estimated the
Pb effect on brachial artery diameter based only on the ambient concentrations of Pb but
did not directly expose their young (mean age: 35 years, SD: 10), healthy adult
participants to Pb isolated from CAPS.
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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 PIH assessed at
delivery (Chen et al.. 2006c). In contrast, epidemiologic studies provide weak evidence
for an association between short-term changes (daily average) in ambient air
concentrations of Pb- PM25 and cardiovascular morbidity in adults. 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 zinc. 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 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 (Samat et al..
2008) or lack of association (Andersen et al.. 2007) specifically to Pb.
5.4.6.2 Mortality
Time-series epidemiologic studies of ambient air Pb- PM2 5 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- PM25. 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 5.4.6.1 above and also are relevant
to the interpretation of these findings for mortality.
5.4.7 Summary and Causal Determination
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
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incidence of hypertension (U.S. EPA. 2006b). This conclusion was substantiated by the
coherence observed between epidemiologic and toxicological findings. The large
evidence base comprising epidemiologic studies conducted in diverse populations using
different designs clearly demonstrated a positive association between blood Pb level and
BP. Meta-analysis of these studies found that each doubling of concurrent blood Pb level
(between 1 and >40 (ig/dL) was associated with a 1 mmHg increase in systolic BP and a
0.6 mmHg increase in diastolic BP (N aw rot et al.. 2002). In addition, most of the
reviewed studies using bone Pb levels also showed an association with BP, pointing to an
effect of cumulative Pb exposure. Similarly, toxicological studies provided evidence that
long-term exposure to Pb (> 4 weeks) results in increased BP in experimental animals
that persists long after the cessation of Pb exposure. Also, animal toxicological studies
provided mechanistic evidence to support the biological plausibility of Pb-induced
hypertension, including Pb-induced oxidative stress, activation of RAAS, altered
sympathetic activity, and vasomodulator imbalance. Studies in the 2006 Pb AQCD also
found associations between Pb biomarkers and other cardiovascular diseases such as
IHD, cerebrovascular disease, peripheral vascular disease, and cardiovascular disease
related mortality; however, the available evidence was limited.
Building on the strong body of evidence presented in the 2006 Pb AQCD, recent
epidemiologic and toxicological studies strengthened the evidence that long-term Pb
exposure is associated with cardiovascular effects in adults with the largest body of
evidence demonstrating associations of Pb with increased BP and hypertension. Recent
epidemiologic studies addressed past uncertainties, including the potential for
confounding. 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). These findings are consistent with those of other studies,
demonstrating that with each increase of 1 (ig/dL concurrent blood Pb level, systolic BP
increased 1 mmHg and diastolic BP increased 0.5 mmHg. 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) where the geometric mean blood Pb level of the
population was 1.64 (ig/dL (Muntner et al. 2005). Despite the extensive evidence for
relatively low concurrent blood Pb levels, as these cardiovascular outcomes were most
often examined in adults that have been exposed to higher levels of Pb earlier in life,
uncertainty remains concerning the Pb exposure level, timing, frequency, and duration
contributing to the observed associations. A new 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 bone Pb level with
hypertension (Glenn et al.. 2006). The results indicated that different mechanisms may
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mediate shorter-term Pb-associated increases in BP and longer-term Pb-associated
development of hypertension.
Collectively, all animal toxicological studies providing blood Pb level and BP
measurements reported increases in BP with increasing blood Pb level (Figure 5-29).
Importantly, an animal toxicological study provided support for increased BP following
long-term Pb exposure resulting in low blood Pb levels with mean blood Pb levels as low
as 2 (ig/dL (Tsao et al.. 2000). However, a majority of studies examined Pb exposures
that resulted in mean blood Pb levels >10 (ig/dL. Thus, the weight of the evidence
demonstrated such effects in animals with blood Pb levels >10 (ig/dL. New studies also
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 epidemiologic evidence for a relationship between Pb exposure and cardiovascular
disease is strengthened by observations of cardiovascular effects in association with both
blood and bone Pb level after adjusting for multiple potential confounding factors,
including age, sex, BMI, antihypertensive medication, SES, race/ethnicity, alcohol
consumption, serum total cholesterol, smoking, educational level, diabetes, and measures
of renal function. The epidemiologic evidence also was substantiated by several available
prospective studies that found associations between biomarkers of Pb and cardiovascular
health endpoints. These studies inform the temporality of these relationship between
biomarkers of Pb exposure and cardiovascular morbidity (e.g., HRV, IHD, BP,
hypertension) (Eumetal.. 2011; Jain et al.. 2007; Peters et al.. 2007; Glenn et al. 2006)
and mortality (khalil et al.. 2009a; Weisskopf et al.. 2009; Menke et al. 2006; Schober et
al.. 2006).
Epidemiologic studies continued to demonstrate a relationship between bone Pb, which is
a metric of cumulative Pb exposure, and BP in adults. Studies that examined both blood
and bone Pb levels did not conclusively demonstrate a stronger association for either
blood or bone Pb (Perlstein et al.. 2007; Glenn et al.. 2006; Martin et al.. 2006). In a
recent meta-analysis, Navas-Acien et al. (2008) found that studies passing the detailed
inclusion criteria all showed a relationship between higher bone Pb levels and higher BP.
Also, all but one that characterized hypertension showed higher relative risks or odds
ratios associated with higher bone Pb levels. Recent epidemiologic studies also
emphasized the interaction between bone Pb levels and factors that modify the
association with BP or hypertension, such as race/ethnicity, chronic stress and metabolic
syndrome. 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). Also, bone Pb level was associated with larger decreases in HRV (which has
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been associated with increased cardiovascular events) among adults with metabolic
syndrome, also associated with increased risk of cardiovascular events (Park et al.. 2006).
The extensive epidemiologic evidence for bone Pb levels in concert with the animal
toxicological evidence for long-term Pb exposure, support an effect of long-term,
cumulative Pb exposure on cardiovascular morbidity.
Recent epidemiologic studies of adults investigated the interaction of Pb biomarkers with
genetic variants in associations with cardiovascular effects. 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 (Figure 5-26 for results)
(Scinicariello et al.. 2010). Additionally, bone Pb concentration was associated with
larger increases in PP, which represents as a good predictor of cardiovascular morbidity
and mortality and an indicator of arterial stiffness, among adults with the HFE H63D
and/or C282Y variant (Zhang et al.. 2010a) (Figure 5-26 for results). Park et al. (2009b)
provided further evidence of HFE and transferrin gene variants, related to iron
metabolism, impacting the associations of bone Pb levels with cardiovascular effects,
evaluated by QT interval changes.
Epidemiologic and toxicological evidence indicates that Pb exposure not only increases
BP and hypertension, but can contribute to the development of other cardiovascular
diseases in adults. However, fewer studies have been published compared to studies of
BP and hypertension. Both recent epidemiologic and toxicological studies provide
evidence in adults for blood Pb-associated increased atherosclerosis, thrombosis, IHD,
PAD, arrhythmia, and cardiac contractility in populations with mean blood Pb levels
>2.5 (ig/dL (Table 5-19). Further, animal toxicological evidence continued to build on the
evidence supporting the mechanisms leading to these cardiovascular system responses, as
well as Pb-induced changes in BP and hypertension. Enhanced understanding of Pb-
induced oxidative stress including 'NO inactivation, endothelial dysfunction leading to
altered vascular reactivity, activation of the RAAS, and vasomodulator imbalance
provides biological plausibility for the consistent associations observed between higher
blood and bone Pb levels and greater cardiovascular effects.
Several studies in children reported associations of childhood (concurrent and measured
at an average age of 2.6 years) blood Pb levels with cardiovascular changes such as TPR
and autonomic and cardiovascular dysregulation in response to acute stress tasks
measured between ages 5 to 15 years (Gump et al.. 2011; 2009; Gump et al.. 2007). Also,
maternal bone and cord blood Pb levels, biomarkers of prenatal Pb exposure, were related
to higher BP in children (Gump et al.. 2005; Zhang et al. In Press). These study findings
of children add evidence that Pb exposure is associated with small increases in BP, and
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that changes in the cardiovascular system precedent to later CVD are potentially
impacted by early-life Pb exposure.
New evidence extended the potential continuum of Pb-related cardiovascular effects by
demonstrating associations between Pb biomarkers and both cardiovascular and all-cause
mortality with follow-up periods ranging between 8 and 12 years. All-cause mortality
was increased with increasing blood Pb level. A recent analysis of the NHANES III
sample reported associations of adult blood Pb level with cardiovascular mortality, with
stronger associations observed with myocardial infarction and stroke mortality (Menkeet
al.. 2006). These findings were supported by a community-based cohort of women age
65-87 years, in which higher effect estimates were observed for mortality from
cardiovascular disease and coronary heart disease (khalil et al. 2009a'). 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 and IHD with hazard ratios of 5.6 and 8.4,
respectively.
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
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. 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 coronary disease, stroke, PAD, and cardiac failure
flngelsson et al.. 2008; Chobanian et al.. 2003; Pastor-Barriuso et al.. 2003; Prospective
Studies Collaboration. 2002; Kannel. 2000a. b; Neaton et al. 1995V
In summary, new studies evaluated in the current review supported and expanded upon
the strong body of evidence presented in the 2006 Pb AQCD that Pb exposure is
associated with cardiovascular health effects. The weight of both epidemiologic and
toxicological evidence continues to support a consistent relationship between Pb
exposure and increased BP or hypertension development in adults. The epidemiologic
evidence is strengthened by several prospective studies that find associations between
biomarkers of Pb and BP and hypertension and by effect estimates that are observed after
adjustment for multiple potential confounding factors. The weight of epidemiologic
evidence supported associations in adults with mean concurrent blood Pb levels less than
5 (ig/dL. As these outcomes in epidemiologic studies were most often observed in adults
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with likely higher past than current Pb exposures, uncertainty exists as to the Pb exposure
level, timing, frequency, and duration contributing to the observed associations. Recent
epidemiologic studies found that bone Pb level, a metric of cumulative exposure, is
strongly related to hypertension risk in adults with mean bone Pb levels greater than
20 jj.g/g. However, uncertainties also exist as to the specific Pb exposure conditions that
contributed to the associations. The weight of animal evidence also demonstrates an
increase in BP after long-term (i.e., greater than 4 weeks) exposure to Pb. Whereas the
majority of studies examined and found increases in BP in animals with mean blood Pb
levels greater than 10 (ig/dL, a recent studies found elevated BP in animals with a mean
blood Pb level of 2 (ig/dL. By demonstrating Pb-induced oxidative stress including 'NO
inactivation, endothelial dysfunction leading to altered vascular reactivity, activation of
the RAAS, and vasomodulator imbalance, toxicological studies have characterized the
modes of action of Pb and provided biological plausibility for the consistent associations
observed in epidemiologic studies between blood and bone Pb and cardiovascular effects.
These associations of Pb with cardiovascular morbidity observed in both epidemiologic
and toxicological studies support recent epidemiologic findings of increased Pb-
associated cardiovascular mortality. Collectively, the evidence integrated across
epidemiologic and toxicological studies as well as across the spectrum of other
cardiovascular endpoints examined is sufficient to conclude that there is a causal
relationship between Pb exposures and cardiovascular health effects.
5.5 Renal Effects
5.5.1 Introduction
This section summarizes key findings with regard to effects of Pb on the kidney in animal
toxicology and epidemiologic studies. Findings summarized across epidemiologic and
toxicological studies indicate that chronic 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, altered glomerular filtration rate (GFR),
chronic kidney disease (CKD) development, and altered 'NO homeostasis with ensuing
elevated BP. As several of these outcomes 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.
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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 damages the
kidneys and its vasculature and systemic hypertension ensues with effects on the
cardiovascular and renal systems (Section 5.4). 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 cause 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 examining 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.
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 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 4.3.5. 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.
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5.5.1.1 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
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 metabolized 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
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 etal.. 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
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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. Unfortunately, they are insensitive for detection of early kidney damage,
as 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 Fi alpha, and thromboxane B2); and cytotoxicity (e.g., N-acetyl-(3-D-
glucosaminidase [NAG]) (Cardenas etal. 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 molecule-1
(Kim-1), developed in AKI research, may prove useful for chronic nephrotoxicant work
as well (Ferguson et al.. 2008; Devaraian. 2007).
5.5.2 Nephrotoxicity and Renal Pathology
5.5.2.1 Toxicology
Figure 5-38 and Table 5-22 presents the recent animal toxicological data for studies
investigating the effects of Pb (as blood Pb level) on various measures of kidney health
and function. 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 j^ig/dL) than those in the current U.S. general
population. As indicated in Figure 5-38 and Table 5-22, toxicological information on
renal dysfunction with blood Pb levels <10 (ig/dL) generally is not available.
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Dysfunction in kidney function measures, including urinary flow, ALP, microalbumin,
and NAG, was observed at blood Pb concentrations above 20 ug/'dL ( Wang et al..
2010d).
o Highest Concentration	~ Lowest Cone, with Response
A Highest Cone, with No Response o Lowest Concentration
Biomarker(Wanget a 1.2010)
Inflammation (Roncal et al. 2007)
Oxidative Stress - Redox (Masso-Gonzalez et al.
2009)
Oxidative Stress - Redox (Navarro-Moreno et al.
2009)
Oxidative Stress - Redox (Wanget al. 2010)
Morphology (Roncal et al. 2007)
Morphology (Masso-Gonzalez etal. 2009)
Morphology (Navarro-Moreno eta 1.2009)
Morphology (Wanget al. 2010)
Kidney function (Ademuyiwa etal. 2009)
Kidneyfunction (Navarro-Morenoetal. 2009)
Kidney function (Roncal et al, 2007)
Kidneyfunction (Wanget al. 2010)
1	10	100
Blood Lead Level (ng/dL)
Figure 5-38 Concentration-response representation of the effect of Pb on
renal outcomes in animal toxicology studies.
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~
~
~
~
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Table 5-22 Additional characteristics for results of toxicological studies
presented in Figure 5-38
Reference
Species;
Lifestage;
Sex
Pb Dose; Exposure
Duration
Blood Pb Level
with Response
(MS/dL)
Outcome
\Nang et al.
(201 Pen
Rat; Adult;
Female
300 ppm Pb-acetate in drinking
water; 8 weeks
20
Biomarker - Aberrant NAG, GGT, |32-microgobulin expression
Roncal et al.
(2007)
Rat; Adult;
Male
150 ppm Pb-acetate in drinking
water; 16 weeks with remnant
kidney surgery at week 4
26
Inflammation - Elevation in number of macrophages & marker MCP-1
in Pb exposed kidneys with remnant kidney surgery.
Navarro-Moreno
et al. (2009)
Rat; Adult;
Male
500 ppm Pb-acetate in drinking
water; 28 weeks
43
Oxidative stress - Increased kidney lipid peroxidation (i.e., TBARS)
\Nang et al.
(201 Od)
Rat; Adult;
Female
300 ppm Pb-acetate in drinking
water; 8 weeks
20
Oxidative Stress - Pb caused increased lipid peroxidation (i.e., MDA
production), elevated kidney antioxidant enzymes (SOD, GPx, CAT),
and depleted GSH
Masso-Gonzalez
et al. (2009)
Rat;
Weanling
pups
300 ppm Pb-acetate in drinking
water; GD1 to PND21
23
Oxidative stress - Elevated TBARS and catalase activity
Roncal et al.
(2007)
Rat; Adult;
Male
150 ppm Pb-acetate in drinking
water; 16 weeks with remnant
kidney surgery at week 4
26
Morphology - Pb induced pre-glomerular vascular disease of kidney
(i.e., sclerosis, fibrosis, peritubular capillary loss)
Navarro-Moreno
et al. (2009)
Rat; Adult;
Male
500 ppm Pb-acetate in drinking
water; 28 weeks
43
Morphology - Electron micrography showed lumen reduction,
microvilli loss, brush border loss, and mitochondrial damage
\Nang et al.
(201 Od)
Rat; Adult;
Female
300 ppm Pb-acetate in drinking
water; 8 weeks
20
Morphology - Electron micrography showed Pb damages
mitochondria, basement membrane, and brush border in kidney
tissue. Some focal tubal necrosis observed.
Masso-Gonzalez
et al. (2009)
Rat;
Weanling
pups
300 ppm Pb-acetate in drinking
water; GD1 to PND21
23
Morphology - Pb elevated relative kidney weight at PND21
Navarro-Moreno
et al. (2009)
Rat; Adult;
Male
500 ppm Pb-acetate in drinking
water; 28 weeks
43
Kidney function - Pb exposed males had elevated urinary pH and
protein, and glucose and blood in the urine.
Roncal et al.
(2007)
Rat; Adult;
Male
150 ppm Pb-acetate in drinking
water; 16 weeks with remnant
kidney surgery at week 4
26
Kidney function - Remnant kidney surgery and Pb exposure induced
decreased creatinine clearance and proteinuria.
\Nang et al.
(201 Od)
Rat; Adult;
Female
300 ppm Pb-acetate in drinking
water; 8 weeks
20
Kidney function - Elevated urinary total protein, urinary albumin, and
serum urea nitrogen.
Ademuyiwa et al.
(2009)
Rat; Adult
200, 300, and 400 ppm
Pb-acetate in drinking water;
12 weeks
39 and 61
Kidney Function - Renal phospholipidosis and depletion of renal
cholesterol after gestational Pb exposure.
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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. After an initial 3 months of Pb exposure
(in a longitudinal 12-month exposure study to either 100 ppm [lower dose] or 5,000 ppm
[higher dose] Pb-acetate in drinking water, male rats), elevated GFR, consistent with
hyperfiltration, and renal hypertrophy were observed; high dose animals also had
increased NAG and GST (khalil-Manesh et al. 1993a; Khali 1-Manesh et al.. 1992a;
Khalil-Manesh et al.. 1992b). At 6 months of exposure, GFR decreased in the high dose
animals, albuminuria was present, and pathology ensued with focal tubular atrophy and

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interstitial fibrosis formation. This pathology was persistent out to 12 months, and at 12
months glomeruli developed focal and segmental sclerosis. Similarly, GFR remained
decreased after 12 months of exposure in the high dose group. 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, provide biological
plausibility for epidemiological studies that observed a similar phenomenon by age in
adults in association with Pb biomarker levels. These duration-dependent dichotomous
changes in GFR are consistent between the toxicological and epidemiologic literature.
Biomarkers of Pb-induced renal toxicity have been developed including the enzymes
lysosomal NAG, GST, brush border antigens (BB50, BBA, HF5), and Tamm-Horsfall
protein. GST functions as a renal biomarker since renal ALAD is protected by the kidney
antioxidant GSH. Urinary NAG and GST levels were found to increase in rats after 3
months of high dose Pb exposure (resulting in blood Pb level of 125 (ig/dL) in rats
(Khalil-Manesh et al. 1992a). whereas only urinary NAG was increased following low
dose Pb exposure (resulting in blood Pb level of 29 (.ig/dL) in rats (Khalil-Manesh et al.
1993a). An in vitro study also found an increase in NAG with high-level Pb exposure
(50 (ig/dL) (Dehpour et al.. 1999). Occupational studies found that urinary NAG
correlated best with recent blood Pb changes.
The renal effects of chronic Pb exposure as detailed above were partially rescued in rats
with chelation therapy such as DMSA (Khalil-Manesh et al.. 1992b). Improvements
include increased GFR, decreased albuminuria, and decreased inclusion body numbers
but little change in tubulointerstitial scarring. Administration of an Indian herb to Pb-
exposed mice, as is discussed in greater detail in the antioxidant section (Section 5.5.5),
produced similar findings. There was a functional rescue however Pb-induced pathology
remained (Javakumar et al.. 2009). Thus, administration of various compounds
(chelators, antioxidants) to Pb-exposed animals produced hemodynamic rescue.
Recent studies have corroborated the previously observed increase in serum creatinine
following Pb exposure in rats. Abdel Moneim et al. (201 lb) reported Pb-induced (i.p.
20 ppm, 5 days) increased serum creatinine accompanied by histological alterations in rat
kidneys. Berrahal et al. (2011) reported on the effects of age-dependent exposure to Pb on
nephrotoxicity in male rats. 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 (ig/dL) and at age 65 days (post-
puberty; blood Pb level 7.5 (ig/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 <0.001]). Various parameters of Pb-induced renal dysfunction are listed in Table 5-23
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below. The elevated serum creatinine in the Pb-exposed animals compared to controls
suggests that animals exposed to low dosage (i.e., 50 ppm) Pb from birth may develop
renal abnormalities. However, the lack of measurements of GFR or renal pathology
weakens the conclusions. Other investigators have also shown that chronic Pb exposure
has detrimental effects on renal function. Pb-exposed male rats (500 ppm Pb-acetate in
drinking water for 7 months) had elevated urinary pH, proteinuria, as well as glucose and
blood in the urine (Navarro-Moreno et al.. 2009).
Table 5-23 Indicators of renal damage in male rats exposed to 50 ppm Pb for 40
and 65 days, starting at parturition
Biomarker (Mean+SD)
PND40
PND40
PND65
PND65
Control
Pb
Control
Pb
Blood Pb level
1.8+0.33
12.7+1.7
2.1+0.35
7.5+0.78
(MQ/dL)




Plasma Creatinine
4.5+0.21
5.35±0.25a
4.55+0.27
6.04±0.29a
(mg/L)




Plasma Urea
0.37+0.019
0.47±0.021a
0.29+0.009
0.29+0.009
(mg/L)




Plasma Uric Acid
7.51+0.44
7.65+0.32
9.39+0.82
5.91±0.53a
(mg/L)




ap <0.001
Source: Modified with permission of John Wiley & Sons, Berrahal et al. (2011)
Qiao et al. (2006) measured the effect of Pb on the expression of the renal fibrosis-related
nuclear factor-kappa B (NF-kB), transforming growth factor (TGF-(3) and fibronectin in
Sprague-Dawley rat kidney. Pb was administered at a dose of 0.5% Pb-acetate,
continuously for either one, two or three months. All growth 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.
Roncal et al. (2007) found that Pb accelerated arteriolopathy and tubulointerstitial injury
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.
Pb-treated rats had higher systolic BP, 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
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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 low Pb
exposure may exacerbate pre-existing underlying kidney disease.
Histological Changes
Historical 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 (Goveretal.. 1970a). Chelators such as CaNa2EDTA removed these
inclusion bodies from affected nuclei (Goveretal.. 1978V Multiple endpoints indicate
dysfunction in the PT after Pb exposure. Pb-induced formation of intranuclear inclusion
bodies in the PT is protective; Pb is sequestered such that it is not in its bioavailable, free,
toxicologically active form. Intranuclear inclusion bodies are found in the kidney with
acute (i.e., < 4 weeks) Pb exposures but present to a lesser degree with chronic exposures
(See Section 5.2.3 for further discussion). Other PT ultrastructural changes in Pb-induced
nephropathy include changes to the PT epithelium, endoplasmic reticulum dilation,
nuclear membrane blebbing, and autophagosome enlargement (Fowler et al.. 1980; Govcr
etal.. 1970b). Symptoms similar to the PT transport-associated Fanconi syndrome appear
with Pb exposure, albeit often at high doses of Pb, i.e., Pb-poisoning. These symptoms,
which include increased urinary electrolyte excretion (zinc), decreased Na-K-ATPase
activity, mitochondrial aberrations, and aminoaciduria, also have been associated with
blood Pb levels in children.
New studies since the 2006 Pb AQCD are consistent with the historical 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 GDI8, 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
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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. (1992a; 1992b); in particular, the enlarged glomeruli are
consistent with the early hyperfiltration caused by Pb.
Abdel Moneim et al. (201 lb) 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.
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) (Navarro-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 j unctions.
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 D1 (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
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caused by Pb administration may be a result of reduced cell cycle inhibition by p27kip_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.. 2007V
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 (0.2% 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).
El-Nekeety et al. (2009) found an extract of the folk medicine plant Aquilegia vulgaris to
be protective against Pb-acetate-induced kidney injury in Sprague-Dawley rats. Rats were
treated with Pb (20 ppm; 2 weeks) and extract (administered before, during, or after Pb).
Pb treatment resulted in 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 are
presented in more detail in Section 5.5.5 but they consistently show that some
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antioxidants are capable of preventing or rescuing Pb-induced renal histopathological
changes.
Alteration of Renal Vasculature and Reactivity
As discussed in Section 5.5.1, changes in renal vasculature function or induction of
hypertension can contribute to further renal dysfunction. Pb can increase BP through the
promotion of oxidative stress and altered vascular reactivity. 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 exposure. Antioxidants
attenuated Pb-related oxidative/nitrosative stress in the kidney and abrogated the Pb-
induced increased BP (Vaziri et al.. 1999b). 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 historic
data detail the interaction of known modulators of vascular tone with Pb.
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
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.
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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 systolic arterial pressure 60 minutes after treatment without
affecting diastolic arterial pressure or heart rate. With this single injection, serum
angiotensin converting enzyme (ACE) activity was significantly elevated. The Pb-
induced altered systolic BP was under the control of the renin-angiotensin system as
evidenced by attenuation of the effects of Pb in Losartan (Ang II receptor blocker) or
Enalapril (ACE inhibitor) co-treated animals (Simoes et al.. 2011). These data agree with
earlier reports of Pb-related increases in ACE activity in young rats exposed to Pb for 2-8
weeks (Sharif! et al.. 2004) and adult rats exposed to Pb for 10 months (Carmignani et al.
1999).
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
staining in kidney tubule epithelial cells (Abdel Moneim et al. 201 lb). Pb treatment (0.5-
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1 (.iM) of isolated rat proximal tubular cells increased cell death by apoptosis and necrosis
in a concentration- and time-dependent manner (Wang etal.. 201 lb). 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 ofWistar 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 (TBARS 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-Sokkarv et al.. 2005). Histological sections of kidneys from Pb-treated rats
showed tubular degeneration with some necrotic cells. Similarly, El-Neweshy and El-
Sayed 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-Sokkarv 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
membrane potential, and intracellular glutathione levels was observed during Pb
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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.
Table 5-24 presents the acute and chronic renal effects of Pb exposure observed in recent
and past animal toxicology studies.
Table 5-24 Effects of Pb on the kidney/renal system related to exposure
duration- evidence from animal toxicology studies
Effects with less than 3 months of exposure
Mitochondrial dysfunction
Renal cell apoptosis
Nuclear Inclusion Body Formation
Proximal Tubule Cytomegaly
Glomerular Hypertrophy
Increased GFR
Effects with 6 or 12 months of exposure
Mitochondrial dysfunction
Renal cell apoptosis
Oxidant redox imbalance
Altered NO homeostasis
ATPase dysfunction
Aminoaciduria
Increased electrolyte excretion
Elevated blood pressure
Decreased GFR
5.5.2.2 Epidemiology in Adults
A number of advances in research on 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 included research in general and CKD patient populations at much
lower blood Pb levels (5-10 (ig/dL) at the time of evaluation than were previously
studied. These advances contributed to the understanding of the effects of Pb exposure on
kidney dysfunction overall in the population. Pb, 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 for 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.
In the 2006 Pb AQCD (U.S. EPA. 2006b). several key issues could not be completely
resolved based on the Pb-kidney literature published to date. These included
characterizing the lowest Pb dose at which altered kidney function effects occur, the
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impact of higher past exposures on associations with current Pb biomarker levels, 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 cadmium. In the intervening five years, relevant
data addressing several of these challenges have been published.
General Population Studies
The 2006 Pb AQCD reported studies that examined associations between indicators of Pb
exposure and kidney function in general populations. This was a new approach to Pb-
kidney research in the two decade time period covered by the 2006 Pb AQCD. As
illustrated in Figure 5-39 and Table 5-25, studies consistently demonstrate associations
between higher blood Pb level and lower renal function in adults. The studies in this
category provided critical evidence that the effects of Pb on the kidney occur at much
lower doses than previously appreciated based on occupational exposure data. 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 was uncertain. The evidence of Pb-associated
renal effects in general population studies was substantiated by results that were adjusted
for multiple potential confounding factors including age, race, sex, education, household
income, smoking, alcohol use, cadmium exposure, and various health indicators such as
diabetes, SBP, BMI, and history of cardiovascular disease.
The landmark Cadmibel Study was the first large environmental study of this type that
adjusted for multiple kidney risk factors (Staessen et al.. 1992). It included 965 men and
1,016 women recruited from cadmium exposed and control areas in Belgium. Mean
concurrent blood Pb was 11.4 (ig/dL (range 2.3-72.5) and 7.5 (ig/dL (range 1.7-60.3) in
men and women, respectively. After adjustment, log transformed blood Pb 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.
Multiple analyses assessing the kidney impact of Pb exposure have been conducted in the
NAS population (Tsaih et al.. 2004; Wu et al.. 2003a; Kim et al.. 1996; Pavton et al..
1994). Participants in this study were originally recruited in the 1960s in the Greater
Boston area. Inclusion criteria included male sex, age 21 to 80 years, and absence of
chronic medical conditions. Longitudinal data contained in NAS publications remain
essential, particularly in light of the dearth of prospective data on the kidney effects of
Pb. The first of these included 459 men whose blood Pb levels from periodic
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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
etal.. 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. At
baseline, mean (SD) age, blood Pb level, and serum creatinine, at baseline, were 56.9
(8.3) years, 9.9 (6.1) (ig/dL, and 1.2 (0.2) mg/dL, respectively. In the longitudinal
analysis, using random-effects modeling with repeated measures, ln-transformed blood
Pb was 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 (ig/dL
((3= 0.027 mg/dL [95% CI: 0.0, 0.054] per unit increase in In blood Pb); effect estimates
in the entire group and subsets with different peak blood Pb levels (< 10 or 40 (ig/dL)
also were positive (and larger for blood Pb levels <10 (ig/dL) but had p-values between
0.07 and 0.13.
This study made two other key contributions. 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 < 10 (ig/dL. A statistically significant positive
association between blood Pb and concurrent serum creatinine remained in a cross-
sectional analysis. These authors also addressed 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 would not be expected.
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 in a subsequent NAS publication (Tsaih et al.. 2004). At
baseline 6 and 26% of subjects had diabetes and hypertension, respectively. Mean blood
Pb levels and serum creatinine decreased significantly over the follow-up period in the
group. Baseline blood Pb level was not associated with change in creatinine in all
participants. However, diabetes was observed to be an effect modifier of the relations of
blood and tibia Pb with change in serum creatinine. Per unit increase in In 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
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1	were influential there as well. Reverse causality was addressed in a sensitivity analysis of
2	participants whose serum creatinine was <1.5 mg/dL; the authors reported that
3	longitudinal associations did not materially change.
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Study
Population
Blood Pb Median Stood Pb i0th-90th Outcome
mm CwiMF percent])® Imfm
PGSmVF FFFFCT ESTIMATES INDICATE POORFR FUNCTION
LONfi'TUOtMAi RrsiJi TS
K m et ,h	\fbmfR blood pfc l Wufi'dl 8^{W 12*1 4iM>,5
NAS lcr blood Pb 1 22 Hg_'dL
nt-'i blood Ph-' 10 MfJ,
T; , h et V COT4'«
SS-SECTiC
Kim et al.(1996)
Tsaih et at, (2004)
3 9 )2 3 S 6* 2.1-7.6
MAS ntn /•¦"h d db^ter
N \S in»- ^ a fi -t	,
MAS nit.' i**"¦ 5Vpt?f( rviurs
MAS ~5tn a -*Pox t rryof fT-^stan
NAb~iPf» bioadPt-v 4.	iU'-7I?4| 4.0-17.5 Scr (mg/dL)
NASiun hl.inj Pb - 2-. ntij
NAS "i"ti blotfj Pb *¦ It (.i^H
\AS .-irp \K)"U
\AS lite a '."Out JtrioeU -v
NVs nif-n .\s,~ -.yp^oniioo
N}>\% -<>i n	hywt'tTt* m m
Cfransi i * St betv ttn vmN
in r,/dL) * 10
Ch si "t r'm r (>^r ,p,ir impHLJ
x 10
,« <-> Bl«»jii Pb
..1 / a 1! .* i, 11,5 Scr (mg/tU.)
Hi f»«>dt-nJ Tipnr.a lsBIixU Pb
n>H2 8. i
• i / f»
N-Yh «'i» n v- h tin t otcs
\Ai  i
OeBurbure et at. (2006) iromh P, uhr~> 3,9(2.6, 5-7) .1.8-8.1
Si Imn i'i)
4(3 Scr fmjLi 1!
NEGATIVE EFFECT ESTIMATES INDICATE POORER FUNCTION
IQNGiTUDtNAL RESULTS
Yu et at. (2004}	CKO Patients
CROSS-SECTIONAL RESULTS
Akesson et al, (2Q05) Swedish Women
Staessen et at {1392) Belgian Womai
Payttm ct	MAS Me*
Swedish Women
r n w«.kf Ht *i (join) NHANESMAiJo«'\i.t-«s
3.2 (2.5, 4.1}	2.0-5,1
7 2 0 7 3 01	1,3-3,8
7 S !"> / m 91	3.7-15.1
7 i {!> 4 
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32
The impact of Pb on the kidney has been examined in multiple NHANES datasets
obtained over the last few decades (Figure 5-40 and Table 5-25). NHANES data analyses
benefit from a number of strengths including large sample size, ability to adjust for
numerous Pb risk 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 susceptible 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
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 (ig/dL in
14,778 adults aged > 20 years. After adjustment for survey year, sociodemographic
factors, CKD risk factors, and blood cadmium, the odds ratios for albuminuria (>
30 mg/g creatinine), reduced eGFR (<60 mL/min/1.73 m2), and both albuminuria and
reduced eGFR were 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 (ig/dL) to the lowest (<1.1
(ig/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 cadmium was included as a covariate,
blood Pb remained significantly associated. In fact, the most important contribution of
this recent NHANES analysis was the evaluation of joint Pb and cadmium exposure
(discussed in Section 5.5.4.1).
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 during the time periods covered. Additional studies in this category 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	Quartiles of Blood Pb Distribution Used
Muntner et al. (20031
Elevated Serum Creatinine
Hypertensive
Normotensive
Navas-Acien et al. (2009)
Albuminuria >30 mg/g creatinine
eGFR <60 mL/min/1.73m
4

1

A
•
1
-•	
J

Hi



/I







|jg/dL Blood Pb
-40	-10 10	50
% Change per |jg/dL Blood Pb
Note: These articles 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 5-40 Percent change for kidney outcomes associated with blood Pb.
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Table 5-25 Additional characteristics and quantitative data for associations of
blood and bone Pb with kidney outcomes for results presented in
Figure 5-39 and Figure 5-40
Study
„ . Location; .. PbBiomarker „ .
Reference Population ^ N Outcome
Period
Statistical Analysis
Effect Estimate
(95% CI)
FIGURE 5-39: Positive Effect Estimates Indicate Poorer Function

Per 1 [jg/dL increase in blood
Pb within the 10th-90th
percentile interval
Longitudinal Results
Adult males
Boston, MA; 459 Median baseline Change in
Random-effects modeling Peak blood Pb < 40 pg/dL:
Kim et al.
Multiple
examination
s 1979-
1994
blood = 8.6 pg/dL
10th-90th
percentile: 4.0-17.5
serum creatinine
between visits x
10 (mg/dL)
adjusted for baseline age,
time since initial visit, BMI,
smoking status, alcohol
ingestion, education level,
hypertension, baseline
serum creatinine, and time
between visits
0.012 (-0.0001, 0.025)
Peak blood Pb < 25 pg/dL:
0.015(0.0002, 0.03)
Peak blood Pb < 10 pg/dL:
0.021 (-0.005, 0.048)
Adult males Boston, MA; 448 Mean (SD)
8/1991-
1995 with
mean
6 year
follow-up
Tsaih et al.
Baseline
Blood Pb
Mg/dL
10th-90th
percentile: 2.1-7.6
Tibia Pb = 21.5
(13.5) pg/g
Patella Pb = 32.4
(20.5) pg/g
Change in
serum creatinine
6.5(4.2) Per year
x 10 (mg/dL)
Log linear regression
adjusted for age, age
squared, BMI,
hypertension, diabetes,
smoking status, alcohol
consumption, analgesic
use, baseline serum
creatinine, serum creatinine
squared
With diabetes: 0.18(0.07, 0.29)
Without diabetes: 0.014 (-0.009,
0.037)
Wth hypertension: 0.019 (-
0.027, 0.065)
Wthout hypertension: 0.021
(-0.007, 0.049)
Per unit increase in In-
transformed tibia Pb
Wth diabetes: 0.082 (0.03,
0.14)
Wthout diabetes: 0.005 (-0.01,
0.02)
Wth hypertension: 0.023
(0.003, 0.04)
Wthout hypertension: 0.0004
(-0.01, 0.01)
Cross-Sectional Results
Kim et al.
Adult males
Boston, MA; 459
Multiple
examination
s 1979-
1994
Median baseline	Serum
blood = 8.6 pg/dL	creatinine
10th-90th	(mg/dL)
percentile: 4.0-17.5
Random-effects modeling
adjusted for baseline age,
time since initial visit, BMI,
smoking status, alcohol
ingestion, education level,
hypertension.
Peak blood Pb < 40 pg/dL:
0.0017 (0.0005, 0.003)
Peak blood Pb < 25 pg/dL:
0.0021 (0.0007, 0.0035)
Peak blood Pb < 10 pg/dL:
0.0033 (0.0012, 0.0053)
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Study
Reference Population !:ocation; N J* Biomarker	Statistical Analysis ™**limale
^	Time	Data	'	(95% CI)
Period
Tsaih et al.
Adult males
Boston, MA; 448 Mean (SD) baseline Serum
8/1991-
1995 with
mean 6 yr
follow-up
Blood Pb =
6.5 (4.2) |jg/dL
10th-90th
percentile: 2.6-11.5
Repeated
measures 10th-90th
percentile: 2.1-7.6
Tibia Pb = 21.5
(13.5) (jg/g
Patella Pb = 32.4
(20.5) (jg/g
creatinine
(mg/dL)
Log linear regression
adjusted for age, age
squared, BMI,
hypertension, diabetes,
smoking status, alcohol
consumption, analgesic use
Baseline blood Pb
With diabetes: -0.009 (-0.038,
0.020)
Without diabetes: -0.004
(-0.010, 0.003)
Wth hypertension: 0 (-0.013,
0.013)
Wthout hypertension: -0.005
(-0.011,0.002)
Follow-up blood Pb
Wth diabetes: 0.053 (-0.032,
0.138)
Wthout diabetes: 0.034 (0.007,
0.061)
Wth hypertension: 0.083
(0.038, 0.128)
Wthout hypertension: 0.014 (-
0.016, 0.044)
De Burbure et Children, mean France,
al. (2006) age = 10 years, Czech
age range = Republic,
8.5-12.3 years and Poland;
dates not
provided
804 Concurrent Blood	Log-transformed
Pb Median (IQR) =	serum
3.9 (2.6, 5.7) pg/dL	creatinine,
10th-90th	cystatin C, and
percentile: 1.8-8.1	^-microglobulin
Log linear regression
adjusted for cadmium,
urinary creatinine, urinary
mercury
Log serum creatinine (mg/L):
-0.062 (-0.106, -0.017)a
Log Cystatin C: -1.3 (-2.4,
-0.21)a
Log ^-microglobulin: -2.2
(-4.0, -0.54)a
FIGURE 5-39: Negative Effect Estimates Indicate Poorer Function
Per 1 [jg/dL increase in blood
Pb within the 10th-90th
percentile interval
Longitudinal Results:
Yu et al. (2004) Adult CKD
Taipei, 121
Mean (SD)
Change in MDRD
Generalized estimating -0.040 (-0.072, -0.008)3
patients
Taiwan; 48
Baseline blood =
eGFR over 4
equations adjusted for

month
4.2 (2.2) pg/dL
yr/100
age, sex, BMI,

longitudinal
10th-90th
(mL/min/1.73 m2
hyperlipidemia,

study period
percentile: 2.0-5.1
body surface
hypertension, smoking,


area)
use of ACE inhibitor,
baseline serum creatinine,
daily protein excretion,
daily protein intake,
underlying kidney disease
Cross-Sectional Results:
Akesson et al.
WHILA,
adult women
Sweden;
6/1999-
1/2000
820
Median (5-95%)
concurrent blood =
2.2 (1.1, 4.6) pg/dL
10th-90th
percentile: 1.3-3.8
Creatinine
clearance/100
(mL/min)
Cystatin C-based
eGFR (Larsson et
al. 2004)/100
(mL/min)
Linear regression
adjusted for age, BMI,
diabetes, hypertension,
' regular use of nephrotoxic
drug, smoking status
-0.018 (-0.03, -0.006)
-0.02 (-0.03, 0.007)
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Reference Population
Study
Location;
Time
Period
Pb Biomarker
Data
Outcome
Statistical Analysis
Effect Estimate
(95% CI)
Staessen et al. Adults
Belgium; 1,981 Concurrent Blood Creatinine
1985-1989	Pb Mean (SD) clearance/100
Males: 11.4 pg/dL (ml_/min)
Females: 7.5 pg/dL
10th-90th
percentile: 3.7-15.1
Log linear regression
adjusted for age, age
squared, sex, BMI, BP,
ferritin level, smoking
status, alcohol ingestion,
rural/urban residence,
analgesic and diuretic
use, blood and urinary
cadmium, diabetes,
occupational exposure to
heavy metals, and gamma
glutamyl transpeptidase
Females: -0.067
(-0.108, -0.027)a
Males: -0.051 (-0.097, -0.047)a
Payton et al. Adult males
Boston, MA; 744
Mean (SD)
Log-transformed
Log linear regression -0.040 (-0.079, -0.0015)
(1994)
1988-1991
concurrent blood =
creatinine
adjusted forage, BMI,


8.1 (3.9) pg/dL
clearance
analgesic and diuretic


10th-90th
(mL/min)
use, alcohol consumption,


percentile: 4.1-12.9

smoking status, SBP, DBP
Fadrowski et al. NHANES, U.S.;
(2010)	adolescents 1988 1994
769
Median concurrent
blood = 1.5 pg/dL
10th-90th
percentile: 0.4-5.4
Q1
<1.0
Q2
1.0 to 1.5
Q3
1.6 to 2.9
Q4
>2.9
Cystatin C-based
eGFR/100
(mL/min/1.73 m2;
calculated using
the Filler and
Lepage equation)
Log linear regression
adjusted for age, sex,
race/ethnicity, urban/rural
residence, smoking,
obesity, household
income, education level of
family reference person,
BP, lipid levels, glucose
levels
-0.022 (-0.038, -0.0054)
Referent
-1.4 (-7.4, 4.5)
-2.6 (-7.3, 2.2)
-6.6 (-12.6, -0.07)
FIGURE 5-40: Analysis of Blood Pb Quartiles:
% change in kidney
outcome
Muntneretal.
NHANESI
adults
U.S.;
1988-1994
4813
Mean (SD)
concurrent blood
Pb
With Hypertension:
4.2(0.14) pg/dL
Q1
Q2
Q3
Q4
0.7 to 2.4
2.5 to 3.8
3.9 to 5.9
6.0 to 56.0
Without
Hypertension: 3.3
(0.10) pg/dL
Elevated Serum
Creatinine
(99th percentile of
each race-sex
specific
distribution for
healthy young
adults)
CKD
Logistic regression
adjusted for age, race,
sex, diabetes, SBP,
smoking, history ofCVD,
BMI, alcohol consumption,
household income,
education level, marital
status, health insurance
Q1: Referent
Wth hypertension
47% (3,110)
80% (34, 142)
141% (46, 297)
Wthout hypertension
11% (-44,121)
19% (-38, 125)
9% (-47, 122)
Q1
Q2
Q3
Q4
0.7 to 1.6
1.7 to 2.8
2.9 to 4.6
4.7 to 52.9
Wth hypertension
Q2: 44% (0, 109)
Q3: 85% (32, 159)
Q4:160% (52, 345)
Wthout hypertension
Q2: -10% (-63, 116)
Q3: 0% (-55, 122)
Q4: 9% (-59, 189)
Navas-Acien et NHANES III, U.S.;
al. (2009) adults	1999-2006
14,778
Geometric
concurrent blood
mean = 1.58 pg/dL
Q1
Q2
Q3
Q4
<1.1
1.2 to 1.6
1.7 to 2.4
>2.4
eGFR <60
mL/minute/1.73
m2
Albuminuria and
eGFR <60
mL/minute/1.73
m2
Logistic regression
adjusted for survey year,
age, sex, race/ethnicity,
BMI, education, smoking,
. cotinine, alcohol intake,
hypertension, diabetes,
menopausal status
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% CI estimated from given p-value.
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Patient Population Studies
CKD as defined by the National Kidney Foundation - Kidney Disease Outcomes Quality
Initiative (NKF-K/DOQI) workgroup (National Kidnev Foundation. 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 to evaluate the impact of
baseline blood Pb level on CKD progression (kidney function decline) in patient
populations (Table 5-26). 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 jxg/72 h, 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 (ig/dL and 99.1 jxg/72 hours, respectively. In a Cox multivariate
regression analysis, chelatable Pb was significantly associated with overall risk for the
primary endpoint (doubling of serum creatinine over the 4-year study period or need for
hemodialysis). 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 ju.g/72 hours) reached the primary end
point than did those in the lower EDTA chelatable Pb levels (<80 jug 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 et al.. 1999) were
modeled separately using generalized estimating equations. Based on these models, a 10
|ig higher chelatable Pb level or 1 (ig/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. Recent studies
expanded the CKD patient populations in which this effect was observed to those with
diabetic nephropathy (Lin et al.. 2006a) and with the lowest blood Pb levels studied to
date (Lin et al.. 2006b'). Results of these observational studies have been summarized
(Weaver and Jaar. 2010).
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Table 5-26 Patient population studies: kidney function decline
Study
n
Baseline
mean (SD)
blood Pb
(M/dL)
Baseline mean
(SD) chelatable
Pb (ng/72 hours)
Baseline mean
(SD) eGFR
(mL/min/1.73 m2)
Years of
follow-up
Decline in eGFR
per 1 SD higher
Pb dose at
baseline per year
Comments
Lin et al.
(2003)
202
5.3 (2.9)
104.5(106.3)
41.6(14.4)
2
0.16
Largest study to date
Yu et al.
(2004)
121
4.2 (2.2)
99.1 (83.4)
36.0 (9.8)
4
2.7 (chelatable)
2.2 (blood Pb)
Longest follow-up; 1 |ig/dL
higher blood Pb, at baseline,
associated with 4.0
mL/min/1.73 m2 reduction in
eGFR over 4 years
Lin et al.
(2006a)
87
6.5 (3.4)
108.5(53.8)
35.1 (9.0)
1
3.87
Type II diabetics with
nephropathy
Lin et al.
(2006b)
108
2.9 (1.4)a
40.2(21.2)
(all <80)
47.6 (9.8)
2
1.1
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 (20101
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. Eighty-one cases and 95 controls were judged to
have had past occupational Pb exposure. Of those, 23 cases and 32 controls were thought
to have been exposed to Pb levels > 30 (.ig/ni1 (the current U.S. OSHA limit is 50 |ig/m3).
In multivariable logistic regression modeling, the OR for CKD (adjusted for age, sex,
smoking, alcohol consumption, diabetes, education, and BMI) was 0.97 (95% CI: 0.68,
1.38) in Pb-exposed compared to non-exposed participants. In analyses comparing low
(> 3 to <10% of occupational exposure limit), medium (> 10 to < 30% of limit), and high
(> 30% of limit) exposure groups, although ORs were elevated (statistically
nonsignificant) in the medium exposure group compared to the never exposed group for
average and lifetime cumulative exposure metrics, a monotonic increase in OR across
exposure groups was not observed. In the low and high exposure groups, odds of CKD
tended to be lower compared with the never exposed group. 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 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. Using mixed-effects
multivariable models, eGFRs declined by 4.27 and 3.39 mL/min/1.73 m2/y in ever and
most Pb-exposed CKD patients, respectively, compared with 4.55 mL/min/1.73 m2/y in
patients without occupational Pb exposure. Thus, the results overall did not provide
strong evidence that Pb exposure was associated with renal effects.
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Strengths noted by the authors included virtually complete case ascertainment and
minimal loss to follow-up. Exposure assessment was listed 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 (Nuvts et al.. 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 ju.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; Elinderet al.. 1986). although blood Pb levels were lower than those noted during
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 Svstem. 2009). Studies
in patients with CKD requiring chronic hemodialysis (ESRD) have also been published in
the past five years. A study of 271 adult patients on regular thrice weekly dialysis
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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 (ig/dL; 25.5% had levels
>20 (.ig/dL. 59% had values of 10-20 (.ig/dL, and 15.5% were <10 (ig/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.
Given these caveats, a 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 (ig/dL
respectively; p <0.001) (Muntner et al.. 2007). This study was unique in that tibia Pb
levels were assessed. Median tibia Pb was higher in ESRD patients although the
difference did not reach statistical significance (17 and 13 jxg/g bone mineral,
respectively [p = 0.13]). 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
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correlated with logarithmic-transformed blood Pb levels after adjustment (Lin et al..
2010). In the prospective portion of this study, blood Pb levels at baseline were
categorized by tertile (range of 0.1 to 29.9 (ig/dL with cut points of 5.62 and 8.66 (.ig/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 (.ig/dL) and highest (> 8.66 (ig/dL) compared to the lowest (< 5.62 (ig/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 (ig/dL)
compared to the lowest tertile (< 8.51 (ig/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 cadmium (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 kidney impact of Pb. These studies have been
performed by Lin and colleagues in Taiwan and involve similar study designs. Initially,
patients are observed in order to compare CKD progression prior to chelation. Then,
CKD patients whose diagnostic EDTA chelatable Pb levels are within certain ranges
(generally 60-600 ju.g/72 hours and thus below the level commonly considered for
chelation) are randomized. The treated group receives weekly chelation with 1 g EDTA
intravenously for up to 3 months. The control group receives placebo infusions. In the
follow-up period, chelation is repeated for defined indications such as increased serum
creatinine or chelatable Pb levels above specified cut-offs. Placebo infusions are repeated
in the controls as well. The results of the most recent of these trials are summarized in
Table 5-27 below.
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Table 5-27 Clinical randomized chelation trials in chronic kidney disease
patients
Reference
Group
n
Baseline
mean
(SD)
blood Pb
(M/dL)
Baseline mean (SD)
chelatable Pb (ng/72
hr)
Baseline mean (SD)
eGFR (mL/min/1.73
m2)
Months of
treatment/
follow-up
Change in
eGFR peryr
(mL/min/1.73
m2)
Comments
Lin et al.
(2003)
Chelated
32
6.1 (2.5)
150.9 (62.4)
32.0(12.1)
27
+ 1.07


Control
32
5.9 (3.0)
144.5 (87.9)
31.5(9.0)

-2.7

Lin et al.
(2006a)
Chelated
15
7.5 (4.6)
148.0 (88.6)
22.4 (4.4)
15
-3.5
Subjects with
Type II
diabetes and
nephropathy

Control
15
5.9 (2.2)
131.4(77.4)
26.3 (6.2)

-10.6

Lin et al.
(2006b)
Chelated
16
2.6 (1.0)a
43.1 (13.7)
41.2(11.2)
27
+3.0
Lowest Pb
exposed and
treated range
Body Lead
Burden (72 h
urinary Pb
excretion) > 20-
<80 |ig

Control
16
3.0(1.1)
47.1 (15.8)
42.6 (9.7)

-2.0

Lin-Tan et al.
(2007b)
Chelated
58
5.0 (2.2)
164.1 (111.1)
36.8(12.7)
51
-0.3
Subjects
without
diabetes

Control
58
5.1 (2.6)
151.5(92.6)
36.0(11.2)

-2.9

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).
This study design requires replication in larger populations at multiple clinical centers. If
confirmed, the effect may be due to removal of Pb. However, chelation may also have a
direct beneficial effect on kidney function, regardless of Pb exposure. Antioxidant effects
of CaNa2EDTA which may improve kidney function directly via improved blood flow to
the kidneys have been reported (Saxena and Flora. 2004; Jacobsen et al. 2001). EDTA
benefits in a Pb rodent model appeared to occur via reduced oxidation (Saxena and Flora.
2004V EDTA administration reduced kidney damage in a rat model of acute renal failure
induced by ischemia (Foglieni et al. 2006). Similarly DMSA has been reported to
prevent renal damage when co-administered during induction of nephrosclerosis in a
nonPb-exposed rat model (Gonick et al.. 1996). Benefits from chelation reported in
rodent models of Pb-related nephrotoxicity (Sanchez-Fructuoso et al.. 2002a; Sanchez-
Fructiioso et al. 2002b; Khalil-Manesh et al. 1992b) did not appear to occur via reversal
of structural damage (Khalil-Manesh et al. 1992b); again suggesting that improved
hemodynamics from reduction of reactive oxidant species, which could be due to reduced
Pb and/or directly to the chelating agent, may be a mechanism (Gonick et al. 1996).
However, the most parsimonious explanation for the combination of the observational
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and experimental chelation work of Lin and colleagues is that reduced Pb is the
underlying reason.
The unique body of work in patient populations by Lin and co-workers, 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 (SD) higher Pb dose at baseline was much lower
in the 2003 study than in subsequent publications (Table 5-27 above). Small sample sizes
and differences in renal diagnoses between groups may be factors in this variability.
Additional research in large populations at multiple centers with assessment of
neuropsychological as well as kidney outcomes is needed.
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. In a two-year prospective cohort study, 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 (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 decreased 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 (Chia etal.. 2006). Similar positive associations were observed
between blood Pb and a wider range of EBE markers in models that adjusted for age, sex,
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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 et al. 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.
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.. 2008a'). An important additional study that analyzed occupational Pb exposure is
discussed below under patient population studies (Evans et al.. 2010).
A few 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 (.ig/dL (Sun et al.. 2008a'). and 25.3 (.ig/dL (Lin and Tai-vi.
2002).
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 (Alasia et al.. 2010; Khan et al.. 2008; Sun et al.. 2008a; Garcon et al.. 2007; Lin
and Tai-vi. 2007; Alinovi et al.. 2005). 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
(karimoov 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.
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
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threshold for a Pb effect on serum creatinine levels (added variable plot shown in Figure
5-41) (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.
_i
T3
O)
E
CD
C
c
o
CC
£
o
E
=3
0
CO
"O
0
-t—'
CO
D
TJ
<
Adjusted blood Pb level (|jg/dL)
Source: Reprinted with permission of the BM J Publishing Group, Weaver et al. (2003a)
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.
Figure 5-41 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 (Nenov
et al.. 2000). As discussed in the 2006 Pb AQCD, 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.. 1992a). Pb could induce glomerular hypertension
resulting in hyperfiltration by several mechanisms including increased ROS, changes in
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eicosanoid levels, and/or an impact on the renin-angiotensin system (Vaziri. 2008b; Roels
etal.. 1994). Whether hyperfiltration contributes to pathology in humans is unclear;
longitudinal studies are needed.
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 et al..
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.
5.5.2.3 Epidemiology in Children
Lead 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 et al.. 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, insensitivity of the clinical kidney outcome
(i.e., GFR) measures for early kidney damage is a particular problem 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 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 etal.. 1988) and may pose
particular challenges in children due to puberty-related biomarker changes (Sarasua et al..
2003). The few studies included the 2006 Pb AQCD that analyzed clinical kidney
outcomes in children found associations with indicators of Pb exposure that were
inconsistent in direction. Fels 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. In a study of 200 Belgian adolescents aged 17
years, higher concurrent blood Pb level was associated with higher serum cystatin-C in
200 (De Burbiire et al.. 2006); however, among 300-600 European children (n varied by
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outcome), higher concurrent blood Pb level was associated with lower serum creatinine
and 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 j^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 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 (ig/dL and 112.9 mL/min/1.73 m2, respectively. Cystatin C-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 (ig/dL) compared with those in the lowest (<1
(ig/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.
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5.5.2.4 Associations between Lead 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 cadmium 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 cadmium (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
cadmium 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 European children, higher
concurrent blood Pb levels were associated with lower serum cystatin C and creatinine;
these inverse associations were attributed to hyperfiltration (De Burbiire et al.. 2006). A
very 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 (Spector etal.. 2011).
Similar to the NHANES adolescent analysis, associations with the cystatin C outcomes
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.
5.5.2.5 Reverse Causality
As discussed briefly above, reverse causality has been considered as an alternative
hypothesis to explain associations observed between indicators of Pb exposure and renal
dysfunction in adults. The reverse causality hypothesis attributes increased blood and
bone Pb levels to reduced Pb excretion from nonPb-related causes rather then implicating
Pb-related renal dysfunction as a contributing factor to CKD. The 2006 Pb AQCD
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concluded that "reverse causality is not likely to be a major explanatory factor accounting
for observed associations between Pb and renal dysfunction" (U.S. EPA. 2006b). The
strongest evidence against reverse causality was provided by longitudinal studies in
which associations between baseline measurements of Pb biomarkers and subsequent
changes in renal function were demonstrated. 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 after adjustment for baseline renal function (Tsaih et al.. 2004; Kim et
al.. 1996). Similar results were reported in studies of patients with renal disease (Yu et
al.. 2004; Lin et al.. 2003). The longitudinal analyses in NAS studies also indicated that
associations between blood Pb level and serum creatinine persisted in the lower range of
serum creatinine levels, including those in the normal range (Tsaih et al. 2004; Kim et
al.. 1996). Thus, the association was not limited to the segment of the population with
potentially clinically significant renal dysfunction in whom reduced Pb excretion would
be more likely.
Additional support against reverse causality was provided by findings in Swedish women
that both higher blood and urinary Pb were associated with lower creatinine clearance
(Akesson et al. 2006; Akesson et al. 2005). If reverse causality were the more likely
hypothesis for these associations, lower creatinine clearance would be associated with
lower urinary Pb, which it is not. Among adults with chronic kidney disease, renal failure
was not associated with increases in blood or bone Pb levels or chelatable Pb levels (Van
De Vvver et al.. 1988). Batuman et al. (1983) found that chelatable Pb levels were similar
in adults with renal disease of unknown and known nonPb-related causes. Study had bone
Pb levels (group means: 18 and 19 jj.g/g) in the range of those measured in recent
epidemiologic studies. In summary, evidence that higher blood Pb levels are associated
with subsequent declines in renal function from baseline levels, that associations persist
among adults with normal renal function, and that renal failure does not increase Pb
biomarker levels collectively do not support reverse causality and thus demonstrate that
reverse causality does not provide a more suitable alternative hypothesis to explain the
associations consistently observed between Pb biomarker levels and renal dysfunction.
5.5.3 Modes of Action for Lead-Induced Nephrotoxicity
5.5.3.1 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
increase serum uric acid by decreasing its kidney excretion (Emmerson and Ravenscroft.
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1975; Ball and So re n sen. 1969; Emmerson. 1965). 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 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). These mechanistic relations have more than just
theoretical importance. Clinically relevant therapies may be possible since EDTA
chelation has been reported to improve both kidney function and urate clearance in
patients with kidney insufficiency and gout, even when EDTA-chelatable Pb body
burdens were low (Lin et al. 2001).
Alterations in serum uric acid have been studied in animal models exposed to Pb.
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. (2011b) reported increased serum uric acid and urea levels after 5 days of
Pb-acetate treatment (i.p. 20 mg/kg). 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 urea levels increased after 40 days of exposure (puberty blood Pb level of 12.7
(ig/dL) but decreased after 65 days of Pb exposure (post-puberty blood Pb level of 7.5
|ig/dL) (Table 5-23).
5.5.3.2 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 5.2.4). 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; Qberlev 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.. 2011b). Pb exposure to rat proximal tubular cells (0.5-1 (.iM) also increased ROS
production, in a concentration-dependent manner (Wang et al.. 201 lb). Oxidative stress
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was also demonstrated by increased lipid peroxidation (i.e., MDA) in serum and renal
tissue after Pb exposure (Abdel Moneim et al.. 201 lb; Lodi etal.. 2011; Wang et al..
201 lb). Berrahal et al. (2011) reported increased MDA in Pb-exposed (50 ppm
Pb-acetate pre- and post-natally) rat kidney relative to controls at both 40 (puberty; blood
Pb 12.7 (ig/dL) and 65 (post-puberty; blood Pb 7.5 (ig/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 (Table 5-24).
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 lb) 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.. 2011).
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, 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 TBARS, 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-
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ALAD in blood was decreased, whereas urinary 8-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. (201 lb) 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
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 (zinc, 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
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were 1.43 (ig/dL in the control group, 22.8 (ig/dL in the Pb group, 21.2 (ig/dL in the Pb
plus zinc plus vitamins group, and 0.98 (ig/dL in the zinc plus vitamin group. The kidney
TBARS were significantly elevated in Pb exposed pups, while treatment with vitamins
and zinc returned TBARS to control levels. Kidney CAT activity was significantly
increased above control with Pb treatment; however supplementation with zinc and
vitamins reduced CAT activity toward normal. Pb exposure inhibited kidney Mn-
dependent SOD but not Cu-Zn-dependent SOD activity. Thus, supplementation with zinc
and vitamins during gestation and lactation was effective in attenuating the redox
imbalance induced by developmental, chronic low-level Pb exposure, likely through the
alteration of Pb accumulation.
Bravo et al. (2007) reported further that mycophenolate mofetil (an immunosuppressive
agent used in renal transplantation which inhibits T and B cell proliferation)
administration reduces 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 TBARS 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
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 TBARS 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 TBARS 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 TBARS were increased
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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.
5.5.3.3 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, 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 8-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 (9-0-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 9-0-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 GD3-0-acetylation could represent a strategy to
attenuate the normal renal apoptotic process and therefore contribute to cell survival
during Pb exposure.
5.5.3.4 Role of Metallothionein
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
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to exert a dual effect on MT expression in the kidney: enhancement of MT gene
transcription but suppression of MT mRNA translation.
Zuo et al. (2009) examined the potential role of a-Synuclein (Sena) and MT in Pb-
induced inclusion body formation. They used MT-I/II double knockout (MT-null) and
parental wild type (WT) cell lines to explore the formation process of Pb-induced
inclusion bodies. Unlike WT cells, MT-null cells did not form inclusion bodies after Pb
exposure. Western blot of the cytosol showed that soluble MT protein in WT cells was
lost during Pb exposure as inclusion bodies formed. However, transfection of MT-1 into
MT-null cells allowed inclusion body formation after Pb exposure. 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 exposure. Sena protein showed poor basal expression in MT-null
cells, and Pb exposure increased Sena expression only in WT cells. MT transfection
increased Sena transcript to WT levels. In both of these cell lines Pb-induced Sena
expression rapidly increased and then decreased over 48 hours as Pb-induced inclusion
bodies were formed. A direct interaction between Sena and MT was confirmed ex vivo
by an antibody pull down assay, where the proteins co-precipitated with an antibody to
MT. Pb exposure caused increased colocalization of MT and Sena proteins. In archival
kidney samples of renal cortex from WT mice chronically treated with Pb, MT was
localized to the surface of inclusion bodies. Thus, Sena may be a component of Pb-
induced inclusion bodies and, with MT, may play a role in inclusion body formation.
5.5.4 Effects of Exposure to Lead 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 iron deficient diet
significantly increased Pb deposition in adult animals (Hashmi et al. 1989). pregnant
dams, and maternally-exposed fetuses (Singh et al. 1991). Dietary thiamine plus zinc
slightly reduced blood and kidney Pb in exposed animals (Flora et al.. 1989). Selenium, 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,
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exposure to other metals or divalent cations can modulate Pb disposition and its effects in
the body.
5.5.4.1 Lead and Cadmium
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, 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 Pb (Roe Is et al..
1994). In another occupational population where both metals were significantly
associated with NAG, a 0.5 jxg/g creatinine increase in Cd had the same effect on NAG as
did a 66.9 jxg/g bone mineral increase in tibia Pb (Weaver et al. 2003a').
The 2006 Pb AQCD noted that data examining the concentration-response relation
between environmental Cd and the kidney were too scarce to determine the impact of Cd
exposure on relations between Pb exposure and other kidney outcomes. A recent
publication in NHANES data collected from 1999 through 2006 addresses this need;
(results pertaining solely to Pb were discussed in Section 5.5.2.2) (Navas-Acicn et al..
2009). Geometric mean concurrent blood Cd level was 0.41 |ig/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 5-25). Compared with
adults with blood Cd levels < 0.2 |ig/L and blood Pb levels <1.1 (ig/dL, adults with blood
Cd levels > 0.6 (ig/L and blood Pb levels > 2.4 (ig/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
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population, and provide novel evidence of increased risk in those with higher
environmental exposure to both metals.
However, a very 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) ng/dL, 26.6 (28.9) |ag/g, 1.15 (0.66) jj.g/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
increase in ln-transformed urine Cd) and ln-NAG, but lower serum creatinine. These
unexpected paradoxical associations have been reported in a few other publications
(De Burbiire 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.
Wang et al. (2009c) 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 synergistic effects of Pb and Cd on kidney mitochondria.
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5.5.4.2
Lead, Cadmium, and Arsenic
Wang and Fowler (2008) present a general review of the roles of biomarkers in
evaluating interactions among mixtures of Pb, Cd, and arsenic. 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; on the other hand, animals
exposed to Cd in addition to Pb or Pb and As showed an additive increase in the 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 Fow ler. 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.. 2005). Additive
or more than additive interactions occurred among Pb, Cd and As under combined
exposure conditions.
5.5.4.3 Lead and Zinc
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 zinc (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 (Upadhvav et al. 2009).
Jamieson et al. (2008) studied the effect of dietary zinc 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 zinc (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
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intensity, distribution, or relative protein amounts. Western blot analysis confirmed that
MT levels were responsive to dietary zinc but not to Pb exposure.
5.5.4.4 Lead and Mercury
Stacchiotti et al. (2009) studied stress proteins and oxidative damage in a renal-derived
cell line exposed to inorganic mercury and Pb. The time course of the expression of
several 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 (jM) and
Pb (PbCl2, 2-500 (j,M) 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
mercury (20 (jM and 40 (j,M) were shown to markedly inhibit growth of the cell line
while the higher doses of Pb (60 |iM to 500 |iM) inhibited cell growth to a lesser degree.
After 24 hours of exposure at 20 (.iM mercury, 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 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
mercury was biphasic: 10 (j,M significantly enhanced GSH by 600%, while the amount of
GSH detected after 20 (.iM mercury only increased by 50% compared to control levels.
GST activity was enhanced by both Pb and mercury. Hsp25 and Hsp72 were up-regulated
by mercury but there was no effect on Grp78 as compared to control. On the contrary, Pb
treatment only upregulated Grp78. Mercury 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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5.5.5 Impact of Treatment with Antioxidants on Renal Lead Accumulation
and Pathology
5.5.5.1 Treatment with Antioxidants
Wang et al. (2010(1) assessed the protective effect of N-acetylcysteine (NAC) on
experimental chronic Pb nephrotoxicity in immature female rats. NAC is a potent oxygen
free radical scavenger, a metal chelator, and the precursor to the antioxidant glutathione.
Sprague-Dawley rats received Pb-acetate (300 ppm in drinking water) and/or NAC
(100 mg/kg/day, by i.p. injection) for 8 weeks to investigate the protective effect of NAC
on Pb-induced renal damage and oxidative stress. Serum and renal cortical Pb levels were
markedly increased in the Pb-treated animals, but reduced in the Pb plus NAC treated
animals. There were time-related increases in urinary alkaline phosphatase, urinary GGT,
urinary NAG, urinary total protein, urinary (3-2 microglobulin, and urinary microalbumin,
which were all decreased by NAC. Serum urea nitrogen was significantly increased by Pb
administration and reduced toward normal by Pb plus NAC. Alterations in proximal
tubular structures were observed in most kidney samples from Pb-treated rats, but
animals treated with combination Pb plus NAC showed well-preserved cell structures and
organelles. Indices of oxidative stress (MDA, SOD, GSH, GPx, and CAT) were altered
by Pb treatment and restored to or toward normal by Pb plus NAC treatment (MDA
increased and the remainder decreased). Thus NAC was shown to have both an anti-
oxidative and a chelator effect on Pb intoxication.
Saxena et al. (2005) investigated the beneficial role of monoesters of meso-2, 3-
dimercaptosuccinic acid in the mobilization of Pb and recovery of tissue oxidative injury
in rats. Dimercaptosuccinic acid (DMSA) is known as a Pb chelator and as an antioxidant
by virtue of its possession of thiol groups. In this study, DMSA, and two of its analogues,
monomethyl dimercaptosuccinic acid (MmDMSA) and mono-cyclohexyl
dimercaptosuccinic acid (MchDMSA) were assessed for their capability to reduce Pb
concentration in blood and soft tissues and to recover Pb-induced oxidative stress. Male
Wistar rats were exposed to Pb-acetate (0.1% in drinking water) for 20 weeks. Rats were
then treated orally once daily for five days with DMSA or its two analogues at doses up
to 100 mg/kg. Exposure to Pb caused a rise in blood Pb levels to approximately 25 (ig/dL.
Exposure to Pb also caused a significant decrease in blood ALAD activity and GSH
levels, accompanied by inhibition of kidney ALAD and an increase in S-aminolevulinic
acid synthetase (ALAS) activity in liver and kidneys. Pb exposure also resulted in
increased blood and soft tissue (brain, liver, and kidney) Pb and TBARS levels and
decreased GSH levels. These were restored by treatment with DMSA and its analogues,
particularly MchDMSA.
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Abdallah et al. (2010) examined the effect of Pb on coenzyme Q levels in rat tissues.
Coenzyme Q acts as an electron and proton carrier in mitochondria and functions as an
antioxidant in its reduced form (ubiquinol). Both coenzyme Q9 and coenzyme Q10 were
measured in rat tissues as coenzyme Q9 is the predominant form found in the rat. Male
albino rats were injected i.p. with Pb-acetate at a dose of 5 mg/kg daily for 6 weeks. No
blood Pb levels were reported. TBARS were elevated above levels in controls in serum,
liver, kidney and brain while non-protein sulfhydryl groups (indicative of GSH) were
decreased in serum and kidney. Both oxidized and reduced coenzyme Q9 levels were
significantly reduced in kidneys from Pb-treated rats as contrasted to controls (mean
[SE]: respectively, 48.6 [5.6] versus 95.5 [10.1] nmol/g tissue for oxidized form and 35.4
[3.0] versus 61.4 [5.1] nmol/g tissue for reduced form). On the other hand, levels of
oxidized and reduced coenzyme Q10 were unchanged. Thus, the reduced levels of
coenzyme Q attributable to Pb exposure may be a factor in the diminished antioxidant
defense mechanism.
El-Sokkary et al. (2005) evaluated the effect of melatonin against Pb-induced hepatic and
renal toxicity in male rats. Melatonin is known to be efficacious as a free radical
scavenger and indirect antioxidant. Three groups of animals were used: control,
Pb-acetate-treated (100 ppm) and Pb-acetate and melatonin (10 mg/kg) given
subcutaneously for 30 days. Lipid peroxidation was measured as the sum of MDA and 4-
hydroxyalkenals (4-HAD). Pb increased kidney lipid peroxidation products, but these
were reduced toward normal with melatonin co-treatment. Both SOD and GSH levels
were reduced by Pb and were increased by melatonin. Histological section of kidneys of
Pb-treated rats showed tubular degeneration with some apparently necrotic cells, while
melatonin-treated rats demonstrated a near normal 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.
Ozsoy et al. (2010) studied the protective effects of L-carnitine on experimental Pb
toxicity in rats. Female two month-old rats were fed 0.5 mg/kg Pb-acetate alone or with
daily injections of 0.5 mg/kg L-carnitine for 60 days. Control animals were injected with
physiological saline. Pb caused an increase in serum creatinine and histopathological
changes in the kidney, consisting of tubule dilatation, degeneration and necrosis and
interstitial inflammation. In the Pb plus L-carnitine group, serum creatinine was reduced
to control values and the histopathological changes were reversed. Immunological
staining indicated Cu/Zn-SOD stimulation by Pb feeding alone and reduction by L-
carnitine co-treatment. The authors attributed the beneficial effects of L-carnitine to its
antioxidant effect.
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Reddy et al. (2010) treated Sprague-Dawley rats with 10 mg/kg/day of Pb-acetate and/or
thiamine (25 mg/kg/day) for 7 weeks. Thiamine treatment normalized the Pb-induced
alterations in blood ALAD activity and urinary NAG activity.
The effectiveness of various plant or bacterial extracts as antioxidants in the kidney was
examined in a few studies. Kharoubi et al. (2008a) described the prophylactic effects of
Wormwood (Artemisia absinthium L.) plant extracts on kidney function in Pb-exposed
animals. Male Wistar rats were exposed to Pb-acetate (750 ppm in drinking water) for 11
weeks, and then received Wormwood extract (200 mg/kg) for 4 weeks. Significant
differences in blood and urinary Pb concentration were observed between the Pb group
and the Wormwood group (e.g., 55.6 (ig/dL blood Pb versus 22.3 (ig/dL, respectively).
Pb induced lipid peroxidation (TBARS and protein carbonyls in the kidney), but these
levels were reduced by Wormwood extract. Wormwood extract also attenuated the
effects of Pb on renal function. These results indicated that Wormwood extract had
significant antioxidant activity and protected the kidney from Pb-induced toxicity.
Jayakumar et al. (2009) evaluated the effect of a methanolic extract of the Indian herb,
Achyranthes aspera, in preventing Pb-induced nephrotoxicity in rats. Male albino Wistar
rats, received Pb-acetate (0.2% for 6 weeks) or Pb-acetate plus A. aspera (200 mg/kg for
6 weeks) simultaneously. A. aspera partially prevented the increases in kidney weight,
BUN, serum uric acid, and serum creatinine caused by Pb administration. The levels of
urinary marker enzymes, GGT, |3-glucuronidase, NAG, Cathepsin D, and LDH, which
were reduced by Pb administration, were increased to or toward normal by A. aspera.
Kidney histology revealed that Pb-treated animals showed tubular damage, whereas the
Pb plus A. aspera-treated animals showed a reduction in tubular damage.
El-Nekeety et al. (2009) evaluated the protective effect of an extract of the folk medicine
plant Aquilegia vulgaris against Pb-acetate-induced oxidative stress in Sprague-Dawley
rats. The experimental group was treated with 200 ppm Pb-acetate and/or 100 ppm of an
extract of A vulgaris for 2 weeks prior to Pb-acetate. Pb-acetate increased serum urea
and decreased serum total protein and albumin. These changes were reversed by
treatment with the extract. Histological examination of kidneys of rats treated with Pb
showed tubular dilatation, interstitial inflammatory cells, hemorrhage, cellular debris, and
hypercellularity in the glomerulus, with apoptotic nuclei in renal tubular epithelial cells.
The rats treated simultaneously with Pb and the extract showed essentially normal renal
tubules and glomeruli while rats treated with Pb and then the extract showed
improvement in tubular structure, but interstitial fibrosis was still present. This
experiment indicated that exposure to Pb generates free radicals, and that an extract of A.
vulgaris resulted in restoration of the different parameters tested. The second experiment
in this group was by Ponce-Canchihuaman et al. (2010) who evaluated the antioxidant
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activity of the cyanobacterium Spirulina maxima against Pb-acetate-induced
hyperlipidemia and oxidative damage in the liver and kidney of male rats. Male Wistar
rats were treated with Pb-acetate by i.p. injection (25 mg/rat on a weekly basis for 3
weeks), and a 5% supplement of Spirulina was given in food. The findings in the kidney
were similar to those in the liver (see Section 5.9.1). Thus, Pb-induced oxidative stress
and renal damage can be attenuated by treatment with Spirulina extract.
Finally, there is a need to examine whether the chelator, CaNa2EDTA, acts also as an
antioxidant and promotes increased vasodilatation and subsequent increased renal blood
flow by enhancing the delivery of NO. This question arose because of the observations of
Lin et al. (2006b) that repeated injections of CaNa2EDTA lead to improvement in kidney
function in patients with chronic renal failure, even in individuals with very low body Pb
stores as indicated by EDTA mobilization tests (i.e., < 80 |ig 72 hour urinary Pb
excretion). Jacobsen et al. (2001) examined the anti-oxidative effects of Gallic acid,
EDTA, and an emulsifier in mayonnaise enriched with 16% fish oil. EDTA was shown to
be an efficient antioxidant in the fish oil enriched mayonnaise as it strongly inhibited the
formation of free radicals and volatile oxidation compounds. The authors suggested that
the antioxidative effect appears to be due to its ability to chelate free iron in egg yolk at
the oil-water interface.
5.5.5.2 Treatment with Antioxidants plus Chelators
Santos et al. (2006a') assessed the potentiating effects of chelators (2,3-
dimercaptopropanol [BAL], 2,3-dimercaptopropane-l-sulfonic acid [DMPS], and meso-
2,3-dimercaptosuccinic acid [DMSA]) given simultaneously with Pb-acetate on S-ALAD
activity, both in vivo and ex vivo. Ex vivo, human blood was pre-incubated with BAL or
DMSA (10 (j,M) or DMPS (1 (j,M) then Pb-acetate added to the reaction mixture. In vivo,
mice were given daily injections of 50 mg/kg Pb-acetate for 15 days and then injected
with 1/3 of LD50 of the chelating agents. In human blood, the inhibitory effect of
Pb-acetate (1 and 100 (jM) on S-ALAD activity was markedly increased in the presence
of BAL and DMPS, whereas DMSA ameliorated the enzyme inhibition caused by 1 (J.M
Pb-acetate. In vivo, Pb-acetate inhibited S-ALAD activity by 42%. Parallel to the ex vivo
results, BAL and DMPS, but not DMSA, increased the inhibitory potency of Pb in blood.
In the kidney, BAL and DMSA but not DMPS increased inhibitory activity. The authors
conjectured that the chelators may deplete the cells of zinc, an essential element for 5-
ALAD activity. These observed effects with chelators were supported by Bradberry and
Vale (2009). Hamidinia et al. (2006). and Aslani et al. (2010) who found decreased
kidney Pb content post-chelation.
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5.5.6
Summary and Causal Determination
The 2006 Pb AQCD 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 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
consistent evidence of Pb-associated lower renal function in populations with mean
concurrent or baseline blood Pb levels of 2-10 (.ig/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 (ig/dL (Muntner et
al.. 2003). The conclusion from the 2006 Pb AQCD was substantiated by the coherence
of effects observed across epidemiologic and toxicological studies. Both human and
animal studies observed Pb-associated hyperfiltration. In animals during the first 3
months after Pb exposure, effects were characterized by increased GFR and increased
kidney weight due to glomerular hypertrophy. However, 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 Pb exposures to associations observed in adults, the impact in
children, and the implication of hyperfiltration.
Recent epidemiologic studies in adult general and patient populations continue to support
Pb-related nephrotoxicity with consistently observed associations of blood and bone Pb
levels with worse kidney function. These studies benefit from a number of strengths that
vary by study but include comprehensive assessment of Pb dose with measurements of
blood Pb, bioavailable Pb, and bone Pb as a biomarker of cumulative exposure;
prospective study design; and statistical approaches that utilize a range of exposure and
outcome measures, while adjusting for numerous potential confounding factors including
age, race, sex, education, household income, smoking, alcohol use, cadmium exposure,
and various health indicators such as diabetes, SBP, BMI, and history of cardiovascular
disease. Large sample sizes provide strength to the general population studies. Re-
examination of a study from the 2006 Pb AQCD provided data to conclude that a 10-fold
increase in concurrent blood Pb (e.g., from 1 to 10 (ig/dL) 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 (ig/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
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index (Akesson et al.. 2005). At current blood Pb levels in the U.S. adult population, a
downward shift in kidney function of the entire population due to Pb may not result in
CKD in identifiable individuals; however, that segment of the population with the lowest
kidney reserve may be at increased risk for CKD when Pb is combined with other kidney
risk factors. For example, in adults with mean (concurrent or baseline measured 4-6 years
before kidney function tests) blood Pb levels that are comparable to that of the general
U.S. population (1.6 to 4.2 (.ig/dL). higher blood Pb level was found to be associated with
clinically-relevant effects (e.g., eGFR < 60 mL/min/1.73 m2, doubling of serum
creatinine) (Fadrowski et al. 2010; Yu et al.. 2004) and larger magnitudes of effect in
potentially at-risk populations with comorbidities for CKD such as diabetes mellitus
(Tsaih et al.. 2004) and hypertension (Tsaih et al.. 2004; Muntner et al.. 2003) or higher
co-exposure to other environmental nephrotoxicants such as cadmium (Navas-Acien et
al.. 2009).
Recent NHANES analyses added to the evidence for Pb-associated lower renal function
in populations with low concurrent mean blood Pb levels (< 2 (ig/dL) (Fadrowski et al..
2010; Navas-Acien et al.. 2009). However, because of uncertainties concerning the
magnitude, timing, frequency, and duration of Pb exposure that contributed to the
observed associations, it is difficult to assess whether a threshold exists for Pb-related
renal effects.
Research in the occupational setting has traditionally been far less consistent than that in
environmentally exposed populations (Section 5.5.2.2). A number of explanatory factors
for this inconsistency, all due to limitations of the occupational literature, were discussed
in the 2006 Pb AQCD. The observation 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 reflects limitations inherent in the study
design. Irrespective of the mechanism, these associations have risk assessment
implications. If associations are in opposite directions in different subgroups of the
population and the relevant effect modifier is not considered, null associations will be
observed. For these reasons, nonsignificant associations or paradoxical associations in the
occupational setting cannot be used as a rationale for discounting Pb-related
nephrotoxicity at lower environmental levels.
Important data on the effects of Pb on the kidney in children were reported in a recent
NHANES analysis in adolescents, ages 12-20 years, which observed an association
between higher concurrent blood Pb (mean: 1.5 (ig/dL) and lower cystatin C-based eGFR
(Fadrowski et al.. 2010). These findings are consistent with results from a rodent model
study in which a low dose of Pb (50 ppm) administered from birth resulted in renal
impairment (elevated serum creatinine as compared to control rats), but these
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observations require confirmation by measurement of GFR and renal pathology (Berrahal
et al.. 2011). This recent epidemiologic study along with several previous studies that
included children with higher Pb exposures (due to residence near sources, Pb poisoning,
or parental occupational exposure) provide evidence that renal function in children may
be affected by Pb exposure; however, additional research is warranted. It is important to
recognize that the NHANES adolescents from Fadrowski et al. (2010) likely had higher
Pb exposures earlier in childhood, thus, the magnitude, timing, frequency, and duration of
Pb exposure contributing to the observed association is uncertain.
CKD results in substantial morbidity and mortality, and, even at earlier stages than those
requiring kidney dialysis or transplantation, is an important risk factor for cardiac disease.
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 5.4). Interestingly,
animal studies have shown Pb-induced vascular injury in the kidney was associated with
increased glomerular sclerosis, tubulointerstitial injury, increased collagen staining, and
an increase in macrophages associated with higher levels of MCP-1 mRNA. It is possible
that the cardiovascular and renal effects of Pb observed are mechanistically linked and
are contributing to the progression of the diseases.
Recently available animal toxicological studies strengthen the evidence regarding the
modes of action for Pb exposure leading to renal alterations, including the influence of
Pb-induced oxidative stress. The mode of action of Pb in the kidneys has been extended
to the field of immunology, evidenced by observations that Pb exposure resulted in
infiltration of lymphocytes and macrophages associated with increased expression of NF-
kB in proximal tubules and infiltrating cells (Roncal et al.. 2007). Additionally, recent
findings expand on the evidence of acute effects of Pb, including mitochondrial
dysfunction, renal cell apoptosis, and glomerular hypertrophy. These mechanisms are
useful in understanding the occurrence of acute hyperfiltration followed by chronic
kidney dysfunction. As indicated in Figure 5-38 and Table 5-22, studies found
dysfunction in various kidney function measures, including urinary flow, ALP,
microalbumin, and NAG in animals with blood Pb levels > 20 (ig/dL (Wang et al..
2010d). Lower concentration Pb exposures and lower blood Pb levels in animals have not
been examined widely.
In summary, new epidemiologic and toxicological studies evaluated in the current review
support or expand upon the strong body of evidence presented in the 2006 Pb AQCD
indicating that Pb exposure is associated with renal effects. The weight of epidemiologic
evidence demonstrates consistently a relationship between higher blood Pb level and
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kidney dysfunction (e.g., lower creatinine clearance, higher serum creatinine, and lower
GFR) in nonoccupationally-exposed adults with mean concurrent or baseline blood Pb
levels of 2-10 j^ig/dL. A few analyses find higher blood Pb levels to be associated with a
greater longitudinal decrease in kidney function overtime (4-15 years), suggesting that
past Pb exposures may contribute to ongoing renal effects. The epidemiologic evidence is
strengthened by associations between Pb biomarker levels and renal function that were
observed after adjustment for multiple potential confounding factors such as age, sex,
comorbid cardiovascular conditions, BMI, smoking, and alcohol use. Similar findings
among adults with CKD indicate that in populations with pre-existing renal disease, Pb
exposure may be associated with greater progression of disease. 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. Coherence for epidemiologic findings is provided by
observations in animal models that Pb exposure for greater than 6 months decreases GFR
and increases serum creatinine. The weight of evidence in animal studies indicates Pb-
induced histopathological changes, including tubular atrophy and sclerosis. Overall,
reduced renal function and increased kidney damage in animals are observed with
chronic Pb (> 4 weeks) exposure that result in blood Pb levels > 20 (ig/dL. By
demonstrating Pb-induced renal oxidative stress, inflammation, mitochondrial
dysfunction, apoptosis, and glomerular hypertrophy, toxicological studies provide
biological plausibility for the associations observed in epidemiologic studies between
blood Pb levels and kidney dysfunction. Collectively, the evidence integrated across
epidemiologic and toxicological studies as well as across the spectrum of renal outcomes
is sufficient to conclude that there is a causal relationship between Pb exposures and renal
health effects.
5.6 Immune System Effects
5.6.1 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
focusing on 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 in the range of
current U.S. population levels (i.e., <10 (ig/dL), particularly early in life, have been
observed only relatively recently within the last 10-15 years (Dictcrt and McCabe. 2007).
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Over the last 10-15 years, advances in the understanding of Pb-associated changes in
immunological parameters in humans have substantiated the immune effects of Pb.
The 2006 Pb AQCD presented consistent evidence for immune system effects associated
with Pb exposure (U.S. EPA. 2006b). Rather than producing overt cytotoxicity or
pathology, Pb exposure was found to be associated with alterations in several subclinical
parameters related to cellular and humoral immunity (Figure 5-42). These conclusions
were based most heavily on toxicological evidence in animals and in vitro models for the
direct effects of Pb exposure in inducing changes in a wide spectrum of immune
outcomes, but principally, a shift away from T-derived lymphocyte helper (Th)l
cytokines toward Th2 cytokines, suppression of Thl-dependent delayed type
hypersensitivity (DTH), elevation in Th2-driven immunoglobulin E (IgE), and
modulation of macrophages into a hyperinflammatory phenotype. Relatively fewer
studies were available in humans and for fewer immune-related endpoints. A majority of
studies in humans were conducted in Pb-exposed male workers, and while a larger
number of immune outcomes were examined, the most consistent evidence comprised
effects on neutrophil functionality. In a smaller body of studies of humans without
occupational exposures, the weight of evidence indicated associations of higher blood Pb
levels with lower abundance of T lymphocyte cells and, in concordance with
toxicological studies, higher IgE levels in children. Due to limited examination, the
immune effects of Pb exposure in adults without occupational exposures were not well
characterized.
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T Cells
/
Elevated IL-4 and IL-5
Suppressed INF-y
Dendritic Cells
Macrophages and
Other Innate Immune Cells
I
Skewed
Th2-biased
responses
Suppressed
Thl- mediated
anti-tumor
host defense
Elevated IL-10
Suppressed IL-12
Increased lipid
and DNA oxidation
in tissues
Increased tissue
inflammation
(e.g. lung, gut, skin)
Elevated TNF-a
Overproduction of ROS
Depleted antioxidant
defenses
Reduced:
Phagocytosis
Nitric oxide production
Peroxynitrite production
Lysosomal activity
Removal of normal
myelomonocytic
suppression
B Cells
Damaged epithelia
and mucosal barriers
\
Increased tumor
cell formation Increased IgE
production
\
Tissue damage
and de novo
antigen
appearance
Inappropriate
T cell proliferation
activation
Increased risk of
later-life cancer
Figure 5-42
Increased risk of atopy
and allergic disease
Increased risk of tissue
inflammatory diseases
Increased risk of
autoimmunity
Reduced host resistance
to bacterial infection
Immunological pathways by which Pb exposure may increase risk
of immune-related diseases.
The pathways by which Pb exposure may alter immune cell function and consequently
increase the risk of immune-related diseases are presented in Figure 5-42. Both
toxicological and epidemiologic studies of children indicated Pb exposure effects on T
cells. A large body of toxicological evidence demonstrated Pb-induced effects on
macrophages. Neutrophil functionality was found to be reduced with higher Pb exposure,
based on studies of Pb-exposed workers with mean blood Pb levels > 40 (ig/dL.
Alterations in immune cells can lead to changes in cell-to-cell interactions, multiple
signaling pathways, and inflammation that affect both innate and acquired immunity, that
in turn, influence the risk of developing infectious, allergic and autoimmune diseases as
well as exacerbating 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. Among the hallmarks
reported for Pb-induced changes in functional pathways were: (1) a suppression of T-
derived lymphocyte helper (Th)l-driven cell-mediated immunity (as measured by a DTH
response); (2) an increase in Th2-driven IgE antibody and Th2 cytokine production; and
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(3) a pro-inflammatory shift in 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 IL-12 and nitric oxide (NO). While epidemiologic
studies did not provide evidence for this full spectrum of immune effects, a shift to a Th2
phenotype was indicated by associations observed between higher blood Pb level and
higher serum IgE levels in children.
Reflecting an inhibition of Thl activity, toxicological evidence presented in the 2006 Pb
AQCD linked Pb exposure of animals to impaired host resistance to bacteria (U.S. EPA.
2006b). Indicating a hyperinflammatory state and local tissue damage, a few available
toxicological studies found Pb exposure-induced generation of autoantibodies, suggesting
an elevated risk of autoimmune reactions. Additionally, the shift toward a Th2 response
suggested that Pb could elevate the risk of atopy and allergic responses. While
toxicological evidence for Pb-induced Th2 activity, elevated IgE, and inflammation
supported the biological plausibility of such effects, epidemiologic evidence was too
sparse to draw conclusions regarding associations between blood Pb levels and these
broader indicators of immune dysfunction in humans.
Changes in the spectrum of immune endpoints were associated with a wide range of
blood Pb levels. Several toxicological studies found blood Pb levels in the range of 7-
100 (ig/dL to be associated with juvenile and/or adult immune effects (e.g., suppressed
DTH, elevated IgE, 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 quantiles) concurrent blood Pb levels >10 (ig/dL.
With respect to critical lifestages of Pb exposure, animal studies provided strong
evidence for immune effects induced by prenatal Pb exposures and by postnatal
exposures in adult animals. There was uncertainty regarding critical lifestages of Pb
exposure in humans as epidemiologic studies of children primarily were cross-sectional
in design and examined concurrent blood Pb levels. Several other limitations of
epidemiologic studies were noted, including small sample sizes; inconsistent adjustment
for potential confounders such as age, sex, smoking, and comorbid conditions; and
reliance on comparisons of immune endpoints among groups with different blood Pb
levels, which provided limited information on the concentration-response function.
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 affected by Pb exposure as presented in Figure
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5-42. For example, a new toxicological study indicated that Pb may modulate the
function of dendritic cells. Results from new toxicological and epidemiologic studies
strengthened the link between Pb-associated effects on immune cells and immune- and
inflammatory-based diseases by providing evidence for changes in intermediary signaling
and inflammatory pathways (Figure 5-42). Several new epidemiologic studies examined
signaling molecules such as pro-inflammatory cytokines and NO to produce findings
parallel with toxicological studies. New toxicological studies expanded knowledge of the
broader role of Pb-associated immune modulation in mediating Pb effects in
nonlymphoid tissues (e.g., in the nervous, reproductive, and respiratory systems).
Although primarily cross-sectional in design, recent epidemiologic studies improved on
the design of earlier studies through greater examination of children and adults with
blood Pb levels more comparable to contemporary levels in the U.S. population and
greater consideration of confounding by age, sex, smoking, comorbid conditions, and
SES-related variables.
5.6.2 Cell-Mediated Immunity
5.6.2.1 T Cells
A majority of the evidence indicating effects of Pb exposure on T cells was provided by
older toxicological and epidemiologic studies reviewed in the 2006 Pb AQCD (U.S.
EPA. 2006b'). Toxicological studies demonstrated Pb exposure-induced shifts in the
partitioning of CD4+ (T helper) cell populations to favor Th2 cells (10-100 (.iM in vitro)
and the production of Th2 cytokines and to suppress production of Thl cytokines (wide
range of Pb exposures) (Section 5.6.5.4). Previous epidemiologic findings were limited
largely to associations between higher concurrent blood Pb level and lower T cell
abundance in children.
Consistent with previous epidemiologic findings, in a recent study of 7 week-old Wistar
rats administered 200 ppm Pb-acetate, the percentages of CD4+ and CD8+ cytotoxic T
cells were decreased (with CD4-CD8- cells elevated) in the submaxillary lymph nodes,
but only with i.p. Pb dosing (p <0.05) and not oral exposure (200 ppm Pb-acetate both
routes of administration) (Teiion et al.. 2010). Limited recent toxicological investigation
added mechanistic information by indicating that Pb may induce a shift to Th2 responses
via T cell-dependent and -independent pathways. Previously, in cultures of human CD4+
T cells, Pb (1 (.iM. 30 minutes) was shown to activate transcription factor NF-kB
(regulates T cell activation) (Pvatt et al.. 1996) and to increase, in a concentration-
dependent manner (10 and 50 (.iM PbCl2, 24 hours), the expression of MHC class II
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surface antigens (e.g., HLA-DR), which mediate the CD4+ response to exogenous
antigens (Guo et al.. 1996b). Recent studies examined higher Pb exposure concentrations
than did previous studies. Heo et al. (2007) recently showed that Pb (25 (.iM) blocked
production of the Thl cytokine interferon-y (IFN-y) in cultures of stimulated mouse T
cells not by affecting gene expression but by suppressing translation of the protein. This
blockage was rescued with the addition of IL-12, which promotes Thl activity. These
results demonstrated a T cell-dependent pathway to skewing toward Th2 responses.
However, Pb was able to decrease the IFN-y to IL-4 ratio (indicating a shift to Th2) in the
absence of STAT6, the preferential signaling pathway for T cells. Thus, Pb also was
found to skew toward Th2 responses via a T cell-independent pathway. Similar
observations were made in vivo by Kasten-Jolly et al. (2010). In this study,
developmental Pb exposure of mice (100 (.iM Pb-acetate in drinking water of dams from
GD8 to PND21, resulting in pup blood Pb levels of 10-30 (ig/dL) induced gene
expression of IL-4 and suppressed production of IFN-y in splenic cells. These changes
occurred in the absence of STAT4 or STAT6 and occurred with concomitant increases in
adenylate cyclase 8 and phosphatidylinositol 3-kinase, adding to the evidence that Pb
may promote Th2 activity via T cell-independent pathways.
While a few available recent epidemiologic studies found associations of blood Pb levels
with lower levels of Thl cytokines and higher levels of Th2 cytokines in humans
(Section 5.6.5.4), the extant evidence for effects on T cells in humans is derived largely
from older studies describing differences in the abundance of several T cell subtypes that
mediate acquired immunity responses to antigens. In most studies of children, higher
blood Pb levels were associated with lower T cell abundance, primarily CD3+ cells. Such
associations were observed in studies that adjusted for potential confounding factors (as
described below) and studies that compared mean cell abundances among groups with
different blood Pb levels. Blood Pb level was less consistently associated with lower
abundance of other T cell subtypes such as CD4+ and CD8+.
The weight of evidence supported lower T cell abundance in association with concurrent
blood Pb level and in groups of children with levels >10 j^ig/dL (Zhao et al.. 2004;
Sarasua et al. 2000; Lutz et al.. 1999). Associations were less consistent in comparisons
of children with lower blood Pb levels. In analyses of 331 children in Germany, 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 (p <
0.05, t-test) compared with children with blood Pb levels < 2.2 j^ig/dL (lowest quartile).
Unlike other studies in children, Karmaus et al. (2005) considered potential confounding
by adjusting for sex, age, number of infections in the past 12 months, passive smoke
exposure, serum lipids, and serum organochlorine levels. Other studies that examined
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blood Pb levels <10 (ig/dL (population mean < 2 (ig/dL or quantile 5-9 (ig/dL) did not
find associations with T cell abundance (Heaazv etal.. 2011; Belles-Isles et al.. 2002).
Another study of U.S. (multiple unspecified locations) children that considered
confounding also found an association between higher concurrent blood Pb level and
lower T cell abundance; however, it was limited to the youngest subjects (Sarasua et al..
2000). Among 241 children 6-35 months in age, a 1 j^ig/dL higher blood Pb level was
associated with a 0.18% (95% CI: -0.34, -0.02) lower CD3+ cell count, a 0.10% (95% CI:
-0.24, 0.04) lower CD4+ cell count, and a 0.04% (95% CI: -0.15, 0.07) lower CD8+ cell
count, adjusting for city 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 driven by lower T cell abundance (3-6%) among children
6-35 months in age with blood Pb levels >15 (ig/dL. It is important to note that 76% of
subjects lived near a Pb smelting operation. These subjects living near Pb sources likely
had higher blood Pb levels and may have driven 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 viral infections. However, it is difficult to assess the potential for
confounding by SES as neither study reported the SES characteristics of the study
population.
In the limited investigation of nonoccupationally-exposed adults, higher concurrent blood
Pb levels were associated with higher T cell abundance (Boscolo et al.. 2000; Sarasua et
al.. 2000; Boscolo et al.. 1999); however, occupational studies conducted in the U.S. and
Asia did not find Pb-exposed workers consistently to have lower or higher abundance of
T cells (Mishra et al.. 2010; Pinkerton et al.. 1998; Yucesov et al.. 1997b; Undeger et al.
1996; Fischbein et al.. 1993). Some studies found that compared with unexposed
controls, Pb-exposed workers had a lower ratio of CD4+/CD8+ cells (Mishra et al.. 2010;
Fischbein et al. 1993). In particular, Fischbein et al. (1993) found this lower ratio in New
York area firearms instructors with a relatively lower mean blood Pb level of 14.6 (ig/dL
and after adjusting for age and smoking. Changes in the CD4+/CD8+ ratio have not been
examined in populations with lower blood Pb levels.
Although several epidemiologic studies have found lower T cell abundance in association
with higher blood Pb levels, after adjusting for a range of potential confounding factors, it
is not clear what effects these small magnitudes of change may have on the cell-to-cell
interactions that mediate downstream acquired immune responses. In children, groups
with higher blood Pb level had lower CD3+ cell abundance that ranged between 1 and
9%. Larger decreases (20-35%) were observed in studies of occupationally-exposed
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males with higher blood Pb levels than those found in the general U.S. population (means
15 and 75 (.ig/dL) (Undeger et al.. 1996; Fischbein et al.. 1993).
In summary, a majority of the evidence indicating effects of Pb exosure on T cells was
provided by previous toxicological and epidemiologic studies reviewed in the 2006 Pb
AQCD. Previous toxicological studies demonstrated Pb-induced expansion of Th2 cells
and increased Th2 cytokine production. A small number of new toxicological studies
expanded the evidence by describing that Pb-induced Th2 skewing may occur via T cell-
dependent pathways (Heo et al.. 2007) and -independent pathways (Kasten-Jollv et al..
2010; Heo et al. 2007). Previous studies did not provide consistent evidence that higher
blood Pb levels or Pb exposure was associated with lower T cell abundance in
nonoccupationally-exposed or occupationally-exposed adults, respectively. A new study
of children did not find children with higher blood Pb level to have lower T cell
abundance (Hegazv etal.. 2011). Epidemiologic findings were limited largely to
associations between higher concurrent blood Pb level (>10 j^ig/dL) and lower T cell
abundance observed in previous studies of children (Karmaus et al.. 2005; Zhao et al..
2004; Sarasua et al. 2000; Lutz et al.. 1999).
5.6.2.2 Lymphocyte Activation
A majority of the evidence indicating that Pb exposure stimulates lymphocyte activation
is provided by previous toxicological studies in which exposures to high concentrations
of Pb (10-100 (.iM) induced an expansion of alloreactive B and T lymphocytes (U.S.
EPA. 2006b). Lymphocyte activation occurs as a result of reversing the normal
suppression that is mediated by a macrophage-like subpopulation. In the limited recent
investigation of Pb-induced lymphocyte activation, toxicological studies characterized
potential mechanisms underlying this effect. Gao et al. (2007) described a potential role
for dendritic cells. Dendritic cells that matured in the presence of 25 (.iM PbCl2 promoted
enhanced alloreactive T cell proliferation compared to control dendritic cells. In addition,
using the local lymph node assay (LLNA), Carey et al. (2006) found that PbCl2 (injection
doses 25-50 |_ig) was able to provide a costimulatory signal to antigens that could activate
T cells in adult female mice. The exact mechanistic basis for this is not known. As
discussed in Section 5.6.5.2, changes in NO production appear to be involved in Pb-
induced lymphocyte activation (Farrer et al.. 2008).
The available epidemiologic evidence for Pb-associated lymphocyte activation was
provided by a small number of previous studies in children and nonoccupationally-
exposed adults. Instead of directly measuring lymphocyte proliferation, these studies
provided indirect evidence by measuring the abundance of cells that expressed HLA-DR,
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an indicator of activated cells. It is important to note that HLA-DR+ cells also may
indicate the presence of activated monocytes. Further, these studies did not adjust
extensively for potential confounding. In the only study of children (ages 9 months-6
years, Missouri), the mean percentage of HLA-DR+ cells was approximately 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) after adjusting for age (Lutz et al.. 1999). However, activated cells were not
elevated in children with blood Pb levels 20-44 j^ig/dL. Studies of nonoccupationally-
exposed adults 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, median blood Pb level both groups
combined: 11 j^ig/dL) (Boscolo et al.. 1999) but only in women without allergies
(Spearman r = 0.44, p < 0.05, n = 25, median blood Pb level: 5.5 (ig/dL) (Boscolo et al..
2000).
Comparisons of Pb-exposed workers and unexposed controls indicated similar levels of
lymphocyte proliferation (<1% difference) between groups (Oueiroz et al. 1994b:
Cohen et al.. 1989) or lower lymphocyte proliferation among Pb-exposed workers (8-
25%) (Mishra et al.. 2003: Fischbein et al.. 1993: Alomran and Shleamoon. 1988: Kimber
et al.. 1986) Toxicological studies have demonstrated the selective expansion of Th2 cells
and suppression of Thl cells (U.S. EPA. 2006b'). Therefore, the differential activation of
specific subtypes may not be discernable in epidemiologic studies that measure overall
lymphocyte proliferation. Additionally, because occupational studies did not provide
concentration-response information, it is difficult to infer whether findings apply to
populations with lower blood Pb levels. Toxicological studies also did not provide
information on Pb-induced lymphocyte activation at exposure concentrations relevant to
humans without occupational Pb exposures.
5.6.2.3 Delayed-type Hypersensitivity
Although recent investigation was limited, a large body of previous 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 and well-documented immunomodulatory effects of Pb exposure in
animal models. A new study indicated that this effect may be mediated by dendritic cells.
The DTH assay is commonly used to assess the T cell-mediated adaptive immune
response, i.e., induration and erythema resulting from the activation of T cells and
recruitment of monocytes to the site of antigen deposition. The DTH response is largely
Thl-dependent in that Thl cytokines drive the production of antigen-specific T cells
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directed against the antigen (sensitizing phase) and the recruitment of antigen-specific T
cells and monocytes to the site of antigen deposition (elicitation phase).
Previous toxicological studies demonstrated suppressed DTH responses in animals after
gestational (Chen et al.. 2004; Bunn et al.. 2001a; Bunn et al.. 2001b. c; Lee et al.. 2001b;
Chenetal.. 1999; Miller et al.. 1998; Faith et al.. 1979) and postnatal (McC'abe et al..
1999; Laschi-Loquerie et al.. 1984; Mulleretal.. 1977) Pb exposures. Blood Pb level
data were not available in all studies; however, DTH was suppressed in animals with a
wide range of blood Pb levels (11 to > 100 (ig/dL), with the associations of DTH with
lower blood Pb levels occurring as a result of gestational exposure.
In some studies that examined Pb exposures at multiple stages of gestation, exposures
later in gestation suppressed DTH in animals (Bunn et al.. 2001c; Lee et al.. 2001^).
These latter findings may reflect the status of thymus and T cell development. A recent
study contributed to the robust evidence 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 PbCl2 (25 (.iM, 10 days) then the antigen ovalbumin
(OVA) and injected the cells into naive adult mice. Mice treated with Pb-exposed
dendritic cells had a diminished OVA-specific DTH footpad response compared with
mice treated with non Pb-exposed dendritic cells.
The capability of Pb to suppress the DTH response is strongly supported by mechanistic
studies in which Pb suppresses Thl cytokine production (Section 5.6.5.4). In some
animal studies, the suppressed DTH response was accompanied by a decreased
production of IFN-y (Lee et al. 2001b; Chenetal.. 1999). which is the primary cytokine
that stimulates recruitment of macrophages, a key component of the DTH response.
Observations of a concomitant decrease in IFN-y further link Pb-induced inhibition of
Thl functional activities with suppression of the DTH response. Further, coherence is
provided by associations observed between Pb exposure or blood Pb levels and other
responses in animals related to the inhibition of Thl-driven adaptive immune responses,
including decreased host resistance (Section 5.6.4.1).
5.6.2.4 Macrophages and Monocytes
As reported in the 2006 Pb AQCD, based on a large body of toxicological evidence, Pb-
induced promotion of a hyperinflammatory phenotype in macrophages was considered to
be a hallmark of Pb-associated immune effects (U.S. EPA. 2006^. Pb-induced
hyperinflammation was indicated by the enhanced production of ROS, suppressed
production of NO, enhanced production of TNF-a, excessive metabolism of arachidonic
acid into immunosuppressive metabolites (e.g., PGE2), impaired growth and
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differentiation of cells, and potentially altered receptor expression [e.g., toll-like
receptors]). Several of these findings are described in detail in Sections 5.6.5.2 and
5.6.5.4. 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 (Section 5.6.4.5). A small number of
available recent studies continue to support previous findings. A study in Pb-treated
(40 ppm by oral gavage, once daily for 40 days) adult mice reported decreased
macrophage yield and viability, and phagocytic index in the kidney (Lodi et al. 2011). In
addition, several studies reinforced the effect of Pb exposure to alter a broad spectrum of
functional alterations in human monocytes (the blood form of macrophages) or mouse
macrophages in vitro (Khan et al.. 2011; Bussolaro et al.. 2008; Mishra et al.. 2006a'). In
particular, Bussolaro et al. (2008) found reduced phagocytic activity with a relatively low
concentration of Pb exposure (0.2 |_iM Pb nitrate, 72 hours). Khan et al. (2011) found that
exposure of human monocytes to 25-100 (.iM Pb-acetate for 3-6 hours induced increases
in the pro-inflammatory cytokine TNF-a. Mishra et al. (2006a) found that 100 ppm
Pb-acetate suppressed lipopolysaccharide (LPS)-induced NO production.
Epidemiologic studies have not widely examined the effects of Pb exposure on altered
macrophage functional activity in humans, and the indices of macrophage function have
varied among the few available studies. Pineda-Zavaleta et al. (2004) was unique in
examining the hyperinflammatory state specifically in macrophages, albeit in relation to
blood Pb levels higher than those in the current U.S. population (range of blood Pb
levels: 3.5-47.5 (ig/dL). This study included children in Lagunera, Mexico, attending
schools at varying distances from an active Pb smelter. Consistent with the large body of
toxicological evidence, higher concurrent blood Pb level was associated with lower NO
production and higher superoxide anion production in macrophages isolated from child
sera (Section 5.6.5.2). Model covariates included sex, age, and presence of allergies.
Other studies in humans examined occupationally-exposed adults and did not find a clear
association between concurrent measurements of blood Pb level and macrophage
abundance. Adjusting for age, race, smoking, and workshift, Pinkerton et al. (1998) found
a lower abundance of monocytes among 145 U.S. Pb smelter workers with a mean blood
Pb level of 39 (ig/dL (7.8%) than among 84 unexposed controls with a mean blood Pb
level of <2 j^ig/dL (8.5%) (p = 0.03). Conterato et al. (In Press) examined a group of male
Pb-exposed painters with blood Pb levels ranging between 1.4 and 14.0 (ig/dL (mean:
5.4 |_ig/dL). lower than those in other occupational studies. The mean level and
percentage of monocytes in the 50 painters were similar to those in the 36 controls (mean
blood Pb level: 1.5 j^ig/dL) and in the 23 battery workers with much higher blood Pb
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levels (mean: 50 (ig/dL). Fischbein et al. (1993) found lower (p < 0.001) abundance of
HLA-DR+ cells in two groups of New York metropolitan area firearms instructors with
mean blood Pb levels of 14.6 (ig/dL (n = 36, 8.8%) and 31.4 (ig/dL (n = 15, 8.7%) than
among the 36 unexposed controls (15.2%). HLA-DR+ is an indicator of activated
functional state of antigen presenting cells (APCs, e.g., macrophages, dendritic cells, T
cells) and is upregulated in response to cell signaling. While this study did not provide
direct evidence for Pb effects on macrophages, the investigators cited previous
observations that 70% of monocytes express HLA-DR antigen compared with 15% of
lymphocytes to suggest that the reduced expression of HLA-DR antigen in Pb-exposed
workers was due primarily to a reduction in activated monocytes.
In summary, a small body of new in vitro studies (Khan et al.. 2011; Bussolaro et al..
2008; Mishra et al.. 2006a) adds to the extensive base toxicological evidence indicating
that Pb exposure decreases functionality of macrophages and promotes a
hyperinflammatory phenotype. The sparse epidemiologic data are provided primarily by
older studies and are not conclusive. Whereas an association between blood Pb level and
a hyperinflammatory state of macrophages was observed in a previous study of children
(Pineda-Zavaleta et al.. 2004). studies of occupationally-exposed adults did not clearly
indicate Pb exposure effects on macrophages. A new study did not find large differences
in monocyte abundance between Pb-exposed workers and unexposed controls (Conterato
et al.. In Press).
5.6.2.5 Neutrophils
In the 2006 Pb AQCD, Pb exposure was not judged to have strong effects on neutrophils
(U.S. EPA. 2006b). This conclusion was based on the relatively limited available
toxicological evidence as compared with that for effects on other immune cells. However,
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 polymorphonuclear cells (PMNs) (Govema 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 reduced ability to kill ingested antigen. Important limitations to applying
epidemiologic findings broadly include male-only study populations, the relatively high
blood Pb levels of workers (range of mean levels: 33.1-71 (ig/dL), lack of concentration-
response information for associations between blood Pb level and neutrophil function,
and lack of consideration of potential confounding factors.
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Instead of examining neutrophil functional activities, a few available recent studies of
animals and occupationally-exposed adults examined the effects of Pb exposure on
neutrophil counts, an increase in which has been interpreted by some investigators to be a
compensatory response to Pb-induced impairment in neutrophil chemotactic activity and
a hyperinflammatory response. In male rats with Pb spheres implanted in their brains,
neutrophils were a major responding cell (kibavashi et al. 2010). Compared with rats
implanted with glass spheres, rats with Pb spheres had greater neutrophil infiltration with
inflammatory-related damage that included apoptosis and indications of
neurodegeneration. It is important to acknowledge the uncertain relevance of these
findings to those expected from typical routes of Pb exposure in humans.
In a group of 68 ceramic, Pb recycling, or bullet manufacturing workers and 50 controls
selected among food plant workers, DiLorenzo et al. (2006) provided information on the
concentration-response relationship and adjusted for potential confounding variables.
Among all subjects, a 1 (ig/dL higher concurrent blood Pb level was associated with a
21.8 cc 11 s/|llL (95% CI: 11.2, 32.4 cc 11 s/|llL) higher absolute neutrophil count (ANC)
adjusted for age, BMI, and smoking status. The geometric mean (range) of concurrent
blood Pb levels was 20.5 j^ig/dL (range: 3.2 to 120 j^ig/dL) among workers and 3.5 (ig/dL
(range: 1 to 11 j^ig/dL) among controls. Eight workers described to have medium to high
Pb exposures (exact blood Pb levels not reported), but no controls had neutrophilia
(n >7,500 cells/mm3), suggesting that chronic, higher-level Pb exposures can lead to a
biologically meaningful excess of circulating neutrophils. Additionally, in analyses
comparing three blood Pb level groups, controls, workers with blood Pb levels <
30 (ig/dL, and workers with blood Pb levels >30 (ig/dL, ANC was observed to increase
monotonically across increasing blood Pb groups, supporting a blood Pb concentration-
dependent relationship. When the three blood Pb groups were further stratified by current
smoking, two-way ANOVA indicated an interaction between concurrent blood Pb level
and current smoking. Higher blood Pb level was associated with higher ANC only in
current smokers. Among nonsmokers, ANCs were similar across blood Pb groups. In
contrast, Conterato et al. (In Press) found lower neutrophil concentrations among two
groups of Pb-exposed male workers with mean concurrent blood Pb levels of 50.0 and
5.4 (ig/dL than among controls with a mean blood Pb level of 1.5 (ig/dL. Pb-exposed
workers did not consistently have higher levels of other immune cells such as
eosinophils, basophils, monocytes, or total lymphocytes either.
Additional evidence for the effects of Pb exposure on neutrophils is provided by findings
that blood Pb level is associated with mediators of neutrophil proliferation, survival,
maturation, and functional activation. These mediators include cytokines such as TNF-a
(Section 5.6.5.4) and complement. The complement system is a component of the innate
immune system that controls various cell-mediated immune responses such as
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chemotaxis of macrophages and neutrophils and phagocytosis of antigens. The effects of
Pb exposure on complement have not been widely examined; however the limited data
suggest Pb may suppress complement activity. Both Ewers et al. (1982) and Undeger et
al. (1996) measured lower serum complement C3 protein among Pb-exposed workers
compared with unexposed controls, with Ewers et al. (1982) additionally observing an
association between higher concurrent blood Pb level and lower C3 in a regression
analysis combining Pb-exposed workers and controls. However, the broad applicability
of these findings may be limited due to the high blood Pb levels in these occupationally-
exposed groups (range of blood Pb levels: 18.6-85.2 (ig/dL and 38-100 (ig/dL,
respectively) and by the lack of adjustment for potential confounding variables.
In summary, a majority of evidence indicating the effects of Pb exposure on neutrophils
was provided by previous studies that found that compared with unexposed controls, Pb-
exposed workers had lower neutrophil functionality (Oueiroz et al.. 1994a; Oueiroz et al..
1993; Valentino et al.. 1991; Bergeret et al.. 1990) and lower complement, which is a
mediator of phagocyte functionality (Undeger et al.. 1996; Ewers et al.. 1982). In the
limited new epidemiologic investigation that examined neutrophil abundance, studies did
not conclusively find Pb-exposed workers to have higher or lower neutrophil abundance.
Overall, the epidemiologic evidence is limited by the high blood Pb levels (range of mean
levels: 33.1 to 71 j^ig/dL) with which neutrophil functionality was observed to be
decreased and the lack of consideration of potential confounding.
5.6.2.6 Dendritic Cells
Since the publication of the 2006 Pb AQCD, new results (from both an ex vivo and in
vitro models) suggest that the effects of Pb exposure 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). Prior research on the effects of Pb in favoring Th2 over
Thl activity emphasized the direct measurement of Thl versus Th2 T-cell populations
and cytokine profiles. But new research techniques have been developed (Gao and
Lawrence. 2010) that provide an opportunity to look upstream at how dendritic cells may
be involved in mediating the effects of Pb on acquired immunity. Gao et al. (2007) used
bone marrow cultures exposed to Pb to examine the impact of Pb on dendritic cell
maturation and function. PbCl2 (25 (iM, 10 days) was found to alter the course of
dendritic cell maturation by changing the ratio of cell surface markers, such as the
CD86/CD80 ratio, that promote Th2 cell development. 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 in altering the cytokine
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expression profile of dendritic cells, in particular, the lower IL-12/IL-10 ratio, may serve
as an important signal to shift naive T cell populations toward a Th2 phenotype.
Strengthening the role of dendritic cells in mediating Pb immune effects were ex vivo
results from the same study which showed that Pb-nai've adult BALB/c mice implanted
with Pb-exposed dendritic cells were skewed toward Th2 activity as indicated by
inhibited DTH (Section 5.6.2.3) and IgG2a antibody (Section 5.6.3) responses (Gao et al.
2007).
5.6.2.7 Natural Killer Cells
Based mostly on epidemiologic studies, previous and recent evidence does not clearly
indicate that the innate immune natural killer (NK) cells are affected to a large extent by
Pb exposure. In previous studies of nonoccupationally-exposed adults in Italy, concurrent
blood Pb levels were correlated with NK cell abundance among men with and without
allergy (Spearman r = 0.49, p~0.05,n = 17 each, median blood Pb level both groups
combined: 11 j^ig/dL) (Boscolo et al.. 1999) and women without allergies (Spearman r =
0.44, p < 0.05) (Boscolo et al.. 2000) In a study of children from St. Lawrence River
communities in Quebec, Canada, cord blood Pb level was not significantly associated
with cord blood NK cell abundance (quantitative results not reported (Belles-Isles et al..
2002). Similar means of NK cell abundance or functional activity were observed in Pb-
exposed workers and unexposed controls (Garcia-Leston et al. 2011; Mishra et al. 2003;
Pinkerton et al. 1998; Yucesov etal.. 1997b; Undeger et al.. 1996; Fischbein et al. 1993;
Kimber et al.. 1986). The lack of evidence for a strong effect on NK cells is underscored
by epidemiologic observations within individual studies that blood Pb levels are
associated with T cell abundance but are not associated with NK cell abundance or their
level of functional activity (karmaus et al.. 2005; Sarasua et al.. 2000; Pinkerton et al..
1998). Consistent with the epidemiologic evidence, toxicological evidence is not
conclusive. In a recent in vitro study comparing the toxicities of metals for different
populations of immune cells, Fortier et al. (2008) found that PbCl2 (7.5-20.7 (ig/dL) did
not affect NK cytotoxicity compared with the DMSO vehicle. However, PbCl2 was not
found to affect other immune parameters (e.g., monocyte phagocytic activity or
lymphocyte proliferation) either. A recent study did show a decrease in NK cell activity
in mice albeit with high-level Pb exposure (1,300 ppm Pb-acetate in drinking water, 10
days, blood Pb level -100 j^ig/dL) (Queiroz etal.. 2011).
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5.6.3
Humoral Immunity
The 2006 Pb AQCD 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 (U.S. EPA. 2006b). Several previous toxicological and epidemiologic studies
(Table 5-28) demonstrated Pb-associated increases in IgE production, which is strongly
implicated in mediating allergic responses and inflammation in allergic asthma. In studies
of juvenile rodents, Pb exposures (Pb-acetate) through drinking water from gestation
through lactation induced concomitant increases in IgE and IL-4 production (a Th2
cytokine) by T cells (Snvder et al.. 2000; Chen et al.. 1999). consistent with the
hypothesis that Th2-mediated mechanisms can induce class switching of B cells to
produce IgE. Evidence from both earlier toxicological and epidemiologic studies (Table
5-28) did not consistently indicate that Pb exposure was associated with changes in other
classes of Igs including IgG, IgM, and IgA. Previous epidemiologic studies of children
(Table 5-28) and adults [Table 5-28 with group comparisons and (Boscolo et al.. 2000;
Boscolo et al.. 1999) with correlation analyses] did not find a consistent association
between blood Pb level and the abundance of B cells, which produce IgE and other
antibodies.
Although previous evidence for effects on Ig classes other than IgE was inconsistent, a
small number of available new toxicological studies found Pb-induced increases in IgG
and pointed to a role for T cell-mediated mechanisms in Pb-induced activation of B cells
and production of Ig antibodies. Important limitations of this evidence include high Pb
exposures tested and use of injection routes of exposure that may have little relevance to
human routes of exposure. Fernandez-Cabezudo et al. (2007) reported evidence for a
subtle shift toward a Th2 immune response following Salmonella infection in mice
exposed to high concentrations of Pb-acetate (5,000-10,000 (.iM. 16 weeks, resulting in
blood Pb levels of 20.5-106 (ig/dL). Serum levels of Salmonella-specific IgGl antibodies
were increased in the Pb-exposed mice compared to controls, whereas IgG2a levels were
increased in control but not Pb-exposed mice following infection. The impaired Thl and
enhanced Th2 response also was evident by the decreased secretion of IL-12 and
increased production of IL-4 by spleen cells taken from Pb-exposed (blood Pb levels
>20.5 (.ig/dL) mice (Fernandez-Cabezudo et al.. 2007).
In a highly-specialized strain of knockout adult mice lacking the ability to produce IFN-y,
i.p. injection with 50 jj.g PbCl2 increased the IgG2a/IgGl ratio (Gao et al.. 2006). 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. Other animal
studies that administered Pb via the i.p., found Pb-induced humoral responses to
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mediated preferentially via Th2 mechanisms. Carey et al. (2006) treated BALB/c mice
with subsensitizing doses of a T cell-independent (Trinitrophenyl-Ficoll [TNP-Ficoll]) or
T cell-dependent (TNP-ovalbumin [TNP-OVA]) hapten-protein conjugate with or
without co-exposure to PbCl2. Seven days later, investigators examined the effects of Pb
on the LLNA response to TNP-Ficoll or TNP-OVA. A bolus injection of PbCl2
(25-50 (.ig) increased the numbers of T and B cells in the lymph node against both TNP-
Ficoll and TNP-OVA. Further, in a concentration-dependent manner, Pb induced
statistically significant elevations of IgM-, IgG2a-, and IgGl-producing cells in the
lymph node. While the increase in IgM-producing cells against TNP-Ficoll indicated a T-
cell independent mechanism, the increases in IgG2a- and IgGl-producing cells against
both antigens indicated a Thl- and Th2-mediated mechanism, respectively. Despite
finding increases in both IgGl - and IgG2a-producing cells, the authors concluded that Pb
skewed the response toward Th2 and had considerable potential for promoting allergic
sensitization against T-dependent antigens.
In a recent microarray study in BALB/c mice, Kasten-Jolly et al. (2010) found that early-
life Pb exposure (100 |_iM Pb-acetate in drinking water of dams from GD8 to PND21,
resulting in pup blood Pb levels 10-30 (ig/dL) produced statistically significant increases
in the expression of genes encoding Ig antibodies or those involved in B lymphocyte
function and activation. These genes included those 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 A (beta 1), Notch gene
homolog 2, and histone deacetylase 7A.
In epidemiologic studies, several of which are new, associations between biomarkers of
Pb exposure and serum IgE level were demonstrated primarily in children, although a
monotonic dose-dependent increase was not consistently observed (Hegazv etal. 2011;
Hon et al.. 2010; Hon et al.. 2009; Karmaus et al.. 2005; Annesi-Maesano et al.. 2003;
Sun et al.. 2003; Lutz et al.. 1999) (Table 5-28). Karmaus et al. (2005) had the most
extensive adjustment for potential confounding factors and examined differences with
lower blood Pb levels. Compared with children with concurrent blood Pb level
<2.2 (ig/dL (quartile 1), children with blood Pb levels 2.84-3.41 (ig/dL (quartile 3) and
>3.4 (ig/dL (quartile 4) had 28% higher serum IgE levels (p = 0.03, F-test). These
differences were observed after the adjustment for potential confounding by age,
biomarkers of various organochlorine exposures, number of infections in the previous 12
months, serum lipids, and passive smoke exposure. Similar differences in IgE count on
basophils were not observed among the blood Pb quartiles; however, it is important to
note that although serum IgE and basophil-bound IgE have been correlated in adults
(Malveaux et al.. 1978; Conrov et al.. 1977). little data are available in children (Dehlink
et al.. 2010). A recent study in children found that serum IgE levels were not correlated
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with basophil-bound IgE but correlated (Spearman r = -0.003) 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 found to
be highly variable across individuals, particularly children (Hausmann et al.. 2011;
Dehlink et al.. 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, and IgM or B
cell abundance. Lutz et al. (1999) found higher serum IgE in children after adjusting for
age, albeit with concurrent blood Pb levels >10 (ig/dL.
Recent studies in children also reported associations between concurrent blood Pb level
and elevated serum IgE but did not adjust for potential confounding variables (Hegazv et
al.. 2011; Hon et al. 2010; Hon et al.. 2009). The studies by Hon and colleagues (2010;
2009) demonstrated associations in children with low blood Pb levels (range: 1.4-
6.0 (ig/dL) and found that blood Pb level was correlated with both serum IgE and atopic
dermatitis, a condition commonly characterized by elevated IgE levels.
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Table 5-28 Comparison of serum immunoglobulin levels and B cell abundance
among various blood Pb groups
Blood Pb


Level







Mean or







Range





Study
Population/Study Details
(Mg/dL)
igEa
lgGb
lgMb
lgAb
B cells0
Children
Karmaus et al.
331 children, ages 7-10 yr, Hesse, Germany
<2.2
46(1.0)
1210
150
123
418e (1.0)
(2005)
Results were adjusted for age, sex, passive smoke
2.21-2.83
30 (0.65)
1214
143
121
353 (0.84)

exposure, number of infections in the previous 12
2.84-3.41
59(1.28)
1241
153
133
389 (0.93)

months, serum lipid concentration, and organochlorine
>3.41
59 (1.28)"
1201
148
136
393 (0.94)

exposures.




Hegazy et al.
318 children, ages 6 mo-7 yr, Egypt
<5
13.0(1.0)




(2011)

5-9
12.0(0.92)






10-14
20.8 (1.60)






15-19
14.9(1.15)






20-44
20.4 (1.57)






45-69
10.2(0.78)"




Sarasua et al.
382 children, ages 6-30 mo, Multiple U.S. locations
0.6-4.9

609
103
50.1
19.1 (1.0)
(2000)
Results were adjusted for age, sex, and study location.
5-9.9

666"
108
55.0
20(1.05)


10-14.9

680"
105
58.2
20.4(1.07)


>15

630
124"
61.4"
22.2(1.16)
Sarasua et al.
562 children, ages 36-71 mo, Multiple U.S. locations
0.6-4.9

817
120
88.6
18.4(1.0)
(2000)
Results were adjusted for age, sex, and study location.
5-9.9

813
116
90.9
17.6(0.96)


10-14.9

856
125
96.3
19.2(1.04)


>15

835
121
94.1
18.6(1.01)
Sarasua et al.
675 children ages 5-16 yr, Multiple U.S. locations
0.6-4.9

1,031
128
140
16.1 (1.0)
(2000)
Results were adjusted for age, sex, and study location.
5-9.9

1,094"
131
143
15.8(0.98)


10-14.9

1,048
136
140
15.3(0.95)


>15

1,221
106
108
20.1 (1.25)
Lutz et al.
279 children, ages 9 mo-6 yr, Springfield, MO
<10
51.8(1.0)



13.4(1.0)
(1999)
Results were adjusted for age.
10-14
74.0 (1.43)



12.6(0.94)


15-19
210.7 (4.07)



16.9(1.26)


20-44
63.7 (1.23)"



11.1 (0.83)
Zhao et al.
75 children, ages 3-6 yr, Zhejiang Province, China
<10




16.58(1.0)
(2004)

>10




16.82(1.01)
Adults without Occupational Pb Exposures
Sarasua et al.
433 children and adults, ages 16-75 yr, Multiple U.S.
0.6-4.9

1,099
175
252
13.9(1.0)
(2000)
locations
5-9.9

1,085
175
242
13.0(0.94)

Results were adjusted for age, sex, study location, and
10-14.9

1,231
262"
283
12.4(0.89)

smoking.
>15

1,169
139
193
14.8(1.06)
Adults with Occupational Pb Exposures
Fischbein et al.
36 unexposed controls, mean age 47 yr
NR




8.6(1.0)
(1993)
36 firearms instructors, mean age 49 yr
14.6




10.5(1.22)

15 firearms instructors, mean age 48 yr
31.4




11.2(1.30)"

New York metropolitan area






Kimber et al.
21 unexposed male controls, ages 20-60 yr
11.8

1062
1294
2235

(1986)
39 male tetraethyl Pb plant workers, ages 25-61 yr
38.4

1018
1040
2425


U.K.






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Study
Population/Study Details
Blood Pb
Level
Mean or
Range
(|jg/dL)
lgEa
lgGb
lgMb
lgAb
B cells0
Pinkerton et al.
(1998)
84 unexposed controls, mean age 30 yr
145 Pb smelter workers, mean age 33 yr
U.S., exact location NR
Results were adjusted for age, race, current smoking
status, and workshift
<2
39

1,090
1,110
94.5
106.2
180
202
14.6(1.0)
13.2(0.90)
Heo et al.
(2004)
606 Pb battery plant workers
Korea
<10
10-29
>30
112.5(1.0)
223.3 (1.99)
535.8
(4.76)d




Anetor and
Adeniyi (1998)
50 male controls, ages 22-58 yr
80 male Pb-exposed workers, ages 21-66 yr, Nigeria
30.4
56.3

1,997
1,187"
215
191
188
144"

Ewers et al.
(1982)
53 male unexposed controls, ages 21-54 yr
72 male Pb battery/smelter workers, ages 16-58 yr,
Germany
11.7
59.0

193'
171
161'
127
140'
128

Undegeretal.
(1996)
25 unexposed male controls, ages 22-56 yr
25 male Pb battery plant workers, ages 22-55 yr,
Turkey
16.7
74.8

1,202.1
854.6"
140.4
93.3d
210.3
168.1
545.5e (1.0)
635.9(1.17)
Alomran and
Shleamoon
(1988)
18 unexposed age-matched controls
39 Pb battery workers, mean age 35 yr, Iraq
NR
NR

1713
1610

183
170
545.5e (1.0)
635.9(1.17)
algE data are presented as lU/mL. (In parentheses are the ratio 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.
CB 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/|jL serum.
'Data are presented as lU/mL.
Sarasua et al. (2000) found associations of higher concurrent blood Pb level with higher
IgA, IgG, and IgM in U.S. children. Blood Pb level was associated with higher levels of
all three Igs (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, and IgM, respectively, per 1 (ig/dL higher blood Pb level,
adjusted for age, sex, and location) in the youngest age group (6-35 months) but not in
older age groups (36-71 months, 6-15 years, 16-75 years), suggesting elevated risk in
very young children. Among infants aged 6-35 months with concurrent blood Pb levels >
15 (ig/dL, serum levels of all three examined Igs were elevated over levels measured in
infants with blood Pb levels < 5 (ig/dL. In this study, IgE was not examined.
While most epidemiologic studies examined concurrent blood Pb levels, some studies
indicated that prenatal Pb exposure may impact Ig levels in newborns (Annesi-Maesano
et al.. 2003; Belles-Isles et al.. 2002). These studies also pointed to an increased risk in
infants; however, important limitations of these studies are the lack of extensive
consideration of potential confounding variables. Belles-Isles (2002) examined newborns
in St. Lawrence River communities in Quebec, Canada and found an association between
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higher cord blood Pb level and higher cord serum IgG in a regression analysis adjusted
for prenatal smoke exposure. Annesi-Maesano et al. ("2003) found that infant hair Pb level
but not cord or placental Pb level was associated with cord serum IgE in newborns in
Paris. From these findings, the authors inferred a stronger effect of Pb exposure
integrated over the entire gestational period compared to exposures closer to birth.
Cotinine was not associated significantly with Pb biomarker levels or with IgE. The
magnitude of association was larger in the subgroup of infants with mothers without
allergies (Spearman r = 0.21 ,p<0 .01 in infants with mothers without allergies versus r =
0.12 in infants with mothers with allergies, pointing to the possible masking of a blood
Pb association with IgE by the stronger association of family history of allergy.
Associations between blood Pb level and IgE also were reported in studies of adults,
without (Pizent et al.. 2008) and with occupational Pb exposure (Heo et al.. 2004). In a
study of urban adults aged 19 to 67 years and of similar SES (i.e., white-collar office
workers) in Zagreb, Croatia, a statistically significant association between higher
concurrent blood Pb level and higher IgE was found in women but not men (Pizent et al..
2008). Several covariates were considered in a stepwise multiple regression, including
age, smoking intensity, and alcohol consumption. Among women not on hormone
replacement therapy or oral contraceptives, a 1 (ig/dL higher blood Pb level was
associated with a 0.60 higher log of IgE (95% CI: 0.58, 1.18). 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, due to the cross-sectional nature of this study, it is difficult to
characterize the timing, level, frequency, and duration of Pb exposure that contributed to
the observed association. Because of likely higher past Pb exposures of adults and the
mobilization of Pb from bone to blood, the associations may reflect effects of higher past
Pb exposures. Investigators did not report an effect estimate in men because it did not
attain statistical significance. Without quantitative results, it was 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 in Italy also did not report quantitative results but
indicated a lack of statistically significant correlation between blood Pb level and IgE in
men (Boscolo et al.. 1999).
A majority of the epidemiologic evidence for the effects of Pb on IgA, IgG, and IgM
levels is provided by previous studies of Pb-exposed workers (Anetor and Adenivi. 1998;
Pinkerton et al.. 1998; Undeger et al.. 1996; Queiroz et al.. 1994b; Alomran and
Shleamoon. 1988; kimber et al.. 1986; Ewers et al.. 1982). Consistent with the collective
body of toxicological findings for these other Ig classes, epidemiologic evidence is
mixed, with studies reporting higher, lower, and similar Ig levels in Pb-exposed workers
compared with unexposed controls. Some studies reporting lower Ig levels in Pb-exposed
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workers included workers with the highest mean blood Pb levels (> 50 j^ig/dL) (Anetor
and Adenivi. 1998; Undeger et al.. 1996). The lack of rigorous statistical analyses in
occupational studies precludes characterization of the factors that may contribute to
inconsistent associations.
In summary, a majority of the evidence for Pb exposure affecting humoral immunity is
provided by previous toxicological and epidemiologic studies that found Pb-associated
increases in IgE. Previous toxicological studies provided evidence for the independent
effects of Pb exposure increasing IgE levels. A few available recent epidemiologic
studies add to the large body of extant evidence supporting associations between higher
concurrent blood Pb level and higher IgE levels in children; however, they did not adjust
for potential confounding variables (Hegazv et al.. 2011; Hon et al. 2010; Hon et al.
2009). Collectively, epidemiologic evidence indicates higher IgE in children with
concurrent blood Pb levels >10 (ig/dL. The epidemiologic evidence in children equally
comprises studies that did not and did adjust for potential confounding. Studies that
adjusted for potential confounding varied in the number and the specific variables
included as model covariates (Table 5-28) but were consistent in finding an association
between blood Pb and IgE. None of the studies in children adjusted for SES or allergen
exposure. Lower SES, which may indicate poorer housing conditions, is associated with
higher exposures to Pb as well as cockroach, rodent, and other allergens. Allergen
exposure and lower SES are associated with higher IgE and IgE-related conditions such
as allergies and asthma (Bryant-Stephens. 2009; Dowd and Aiello. 2009; Aligne et al.
2000). Studies generally did not provide detailed demographic or residential information
to assess whether SES and/or allergen exposure potentially confounded the observed
associations between blood Pb level and IgE. Thus, uncertainty remains as to the extent
to which blood Pb-IgE associations in children are confounded by unmeasured SES
and/or allergen exposure. Evidence in adults is limited and thus, inconclusive. A new
study of environmentally-exposed adults found an association between concurrent blood
Pb level and IgE in women (Pizent et al.. 2008). Evidence integrated across toxicological
and epidemiologic studies does not consistently indicate that Pb exposure affects other
classes of Igs; however, a few new toxicological studies found Pb-induced increases in
IgG (Femandez-C'abezudo et al.. 2007; Gao et al.. 2006) or increases in the expression of
genes encoding Ig antibodies (Kasten-Jollv et al.. 2010).
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5.6.4
Immune-based Diseases
5.6.4.1 Host Resistance
The capability of Pb to reduce host resistance of animals to bacteria has been known for
almost 40 years and was supported by several animal studies described in the 2006 Pb
AQCD (U.S. EPA. 2006b). A recent animal study provided supporting evidence for such
an effect. Biological plausibility for these findings has been provided by observations that
Pb affects mechanisms underlying diminished host resistance, e.g., suppressed
Thl-driven acquired immune responses and increased inflammatory responses in target
tissue resulting in further damage to 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-Teiero and Pamer. 2004) and these have been found to be inhibited
by Pb exposure. The lack of IFN-y can inhibit appropriate and timely macrophage
activation. Nitric oxide is produced by macrophages during cellular activation and plays a
role in host defense against bacterial infection (U.S. EPA. 2006b). and NO production
also has been found to be suppressed by Pb exposure. A recent animal study
characterized a potential mechanism by which Pb may impact both innate immune cells
and natural host defense barriers. Kasten-Jolly et al. (2010) showed that developmental
exposure of mice to Pb (100 (.iM Pb-acetate in drinking water of dams from GD8 to
PND21, resulting in pup blood Pb levels of 10-30 (ig/dL) resulted in an upregulation of
splenic RNA of caspase-12, a cysteine protease that inhibits the clearance of bacteria both
systemically and in the gut mucosa (Saleh et al. 2006).
In the few available epidemiologic studies, a range of indicators of Pb exposure (i.e., cord
blood Pb, concurrent blood Pb, Pb content in total deposition samples, Pb content in
lichen) were associated with viral and bacterial infections in children. Collectively, study
limitations, including lack of consideration for potential confounding variables (karmaus
et al.. 2005; Rabinowitz et al.. 1990). lack of statistical analysis (Karmaus et al.. 2005).
and ecological design (Carreras et al.. 2009). limit the ability to draw inferences
regarding the effects of Pb exposure on viral or bacterial infections in children. Similar to
these studies in children, an occupational study finding higher frequency of self-reported
colds or influenza among Pb battery or smelter plant workers with higher blood Pb levels
than among unexposed controls is limited by a lack of statistical analysis (Ewers et al..
1982). Studies that address the aforementioned limitations are needed to characterize the
relationship between Pb exposure and resistance to bacterial and viral infection in
humans.
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With limited investigation, the effect of Pb on host resistance to parasitic agents is
unclear. The 2006 Pb ACQD described a study in which high-level Pb-exposed (>
10 (.iM) mouse macrophages had diminished ability to kill Leishmania enrietti parasites
(Maucl et al. 1989). However, given the we 11-characterized effect of Pb in promoting
Th2 activity, it is plausible for Pb to enhance host resistance to parasites that require
robust Th2 responses (e.g., helminths) (U.S. EPA. 2006b'). In a recent study, high-level
Pb exposure (>10 (.iM) enhanced host resistance to malaria (Koka et al.. 2007). However,
this was attributed to the capability of Pb to induce eryptosis and the rapid removal of
malaria-infected erythrocytes and not to Pb-induced alterations in immune function.
Nriagu et al. (2008) also reported that higher blood Pb level was associated with lower
malaria prevalence among children (ages 2-9 years) from three Nigerian cities. A
majority of children (75%) had blood Pb levels below 10 (ig/dL. The association
persisted after adjusting for age, sex, number of siblings, and other comorbidities such as
headaches, depressed mood, and irritability.
5.6.4.2 Asthma and Allergy
Toxicological evidence and to a relatively limited extent, epidemiologic evidence, have
supported the effects of Pb exposure on multiple immune parameters, including elevated
production of Th2 cytokines such as IL-4, increased IgE antibody production (Table
5-24), and increased inflammation. These are well-recognized pathways in the
development and exacerbation of allergy and allergic disease, including asthma. This
mechanistic evidence is coherent with a small body of epidemiologic studies that found
associations of blood Pb levels with asthma or allergy in children (Figure 5-43 and Table
5-29). Children examined in these studies encompassed a wide range of ages (i.e., less
than 1 year to 12 years) and across studies, blood Pb was measured during different
lifestages. This body of evidence in children included large studies with multivariate
analyses and studies with prospective follow-up of subjects in which disease occurrence
was ascertained after the measurement of blood Pb levels.
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Study
Population	Outcome
Blood Pb Level Mean
or Group ((jg/dL)
Rabinowitzet al. (1990) Children
Jedrychowski et al. (2011) Children
Eczema
Positive SPT
>10vs. <10
prenatal: 1.16
concurrent: 2.02
Joseph et al. (2005)
.. , „	Incident Asthma	__ _
Children,Caucasian	„„ ,.	>5vs. <5
Requiring Medical Care
Children, African American
Children, African American
Rabinowitzet al. (1990) Children
Pugh Smith et al. (2011) Children
Asthma
Asthma
>5 vs. <5
>10 vs. <5
>10vs. <10
>10vs. <10
0 1
2 3 4 5
Odds ratio (95% CI)
6 7
Note: Results are organized by endpoint and for asthma in order of increasing blood Pb level. For analyses with blood Pb level as a
continuous variable, odds ratios are standardized to a 1 |jg/dL increase in blood Pb level. SPT = skin prick test, BR = bronchial
responsiveness. 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.
Figure 5-43 Associations of blood Pb levels with asthma- and allergy-related
conditions in children.
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Table 5-29 Additional characteristics and quantitative results for studies
presented in Figure 5-43
Study
Population/Location
Blood Pb Level Data (|jg/dL)
Statistical Analysis
Outcome
Odds Ratio
(96% Cl)a
Rabinowitz et
al. (1990)
1,768 children followed
from birth to
unspecified age
Boston area, MA
Prenatal (cord) > 10 vs. <10a
Logistic regression with no
additional covariates included in
model
Eczema
1.0(0.6, 1.6)
Jedrychowski
et al. (2011)
224 children followed
prenatally to age 5 yr
Krakow, Poland
Prenatal (cord): GM: 1.16 (95% CI: 1.12,1.22)
Concurrent: GM: 2.02 (95% CI: 1.95, 2.12)
Logistic regression adjusted for
sex, parity, maternal age,
maternal education, maternal
atopy, maternal smoking
Positive SPT
2.3(1.1,4.6)"
1.1 (0.7, 1.6)"
Joseph et al.
(2005)
4,634 children, ages 1-3
yr followed prospectively
for 12 months
Southeastern Ml
Measured up to 12 mos before outcome
Caucasian > 5 vs. Caucasian <5C
African American > 5 vs. African American <5C
Logistic regression adjusted for
sex, birth weight, and annual
income
Incident asthma
requiring
medical care
2.7(0.9, 8.1)
1.1 (0.8, 1.7)

African American > 10 vs. African American <5C


1.3(0.6, 2.6)
Rabinowitz et
al. (1990)
1,768 children followed
from birth to
unspecified age
Boston area, MA
Prenatal (cord blood)
> 10 vs. <10a
Logistic regression with no
additional covariates included in
model
Prevalent
asthma
1.3(0.8, 2.0)
Pugh Smith
and Nriagu
(2011)
356 children,
ages 0-12 yr
Saginaw, Ml
Levels ascertained from statewide database,
timing unreported but varied among subjects
>10 vs. <10a
Logistic regression adjusted for
age, sex, number of stories in
unit, cat in home, dog in home,
cockroach problem, number of
persons in home, household
smoking, clutter,
candles/incense, type of cooking
stove, main heating source,
months of residency, housing
tenure, type of air conditioning,
peeling paint, ceiling/wall
damage, age of housing, water
dampness/mold/mildew
Prevalent
asthma
7.5(1.3, 42.9)
GM = geometric mean, SPT = skin prick test.
aOdds ratio in children with blood Pb level > 10 |jg/dL with children with blood Pb level <10 |jg/dL serving as the reference group
bOdds ratio presented per 1 |jg/dL increase in blood Pb level.
"Relative risk in each specified subgroup with children with blood Pb level <5 |jg/dL serving as the reference group.
In a study of 4,634 children in southeastern Michigan, blood Pb levels were measured at
ages 1 to 3 years, up to 12 months prior to asthma assessment. In analyses that controlled
for annual income, birth weight, and sex, an elevated risk of incident asthma requiring a
doctor visit or medication (indicator of severe asthma) was reported in association with
blood Pb levels > 5 (ig/dL among Caucasian children (relative risk [RR]: 2.7 [95% CI:
0.9, 8.1] compared with Caucasian children with blood Pb levels < 5 j^ig/dL) (Joseph et
al.. 2005) (Figure 5-43 and Table 5-29). In analyses restricted to African Americans,
children with blood Pb levels >10 (ig/dL had an elevated risk of asthma requiring
medical care (RR: 1.3 [95% CI: 0.6, 2.6] compared with children with blood Pb level <
5 (ig/dL) (Figure 5-43 and Table 5-29). In analyses that used Caucasian children with
blood Pb level <5 j^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. Collectively, the results indicated a stronger association with
race/ethnicity than with blood Pb level. It is important to recognize the small numbers of
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children with asthma requiring medical care in the high blood Pb level categories, which
likely accounted for the wide 95% CIs (5 Caucasian children with blood Pb > 5 j^ig/dL
and 9 African American children with blood Pb level >10 j^ig/dL).
Consistent with Joseph et al. ("2005). a recent cross-sectional study conducted in Saginaw,
Michigan found a higher prevalence of asthma in children (ages <12 years) with blood
Pb levels >10 (.ig/dL (Pimh Smith and Nriagu. 2011). Of the 356 children randomly
selected from a statewide surveillance database of blood Pb measurements, 78% were
African American. This study had greater power than did Joseph et al. ("2005) due to the
larger percentage of children with blood Pb levels >10 (ig/dL (18.6%). 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]). A strength of this study was the adjustment for a
large number of potential confounding variables such as age, sex, pets in the home,
housing characteristics, and household smoking. However, because children were
identified from a statewide database of initial blood Pb measurements collected at
unspecified ages, the timing of blood Pb varied among children. In a study that did not
consider potential confounding variables, Rabinowitz et al. (1990) reported an increased
risk of asthma in children (age of assessment not reported) in association with cord blood
Pb levels > 10 (ig/dL relative to cord blood Pb levels < 10 j^ig/dL (Figure 5-43 and Table
5-29).
While the aforementioned studies examined blood Pb levels measured at a single point in
time, a recent prospective birth cohort study compared associations of prenatal (cord and
maternal) and concurrent blood Pb levels with risk of allergic sensitization at age 5 years
(Jedrychowski et al. 2011). Cord and prenatal maternal blood Pb level were associated
with greater risk of positive skin prick test (SPT) to dust mite, dog, or cat allergen than
was child concurrent blood Pb level (Figure 5-43 and Table 5-29). For prenatal Pb
biomarkers, similar effect estimates were obtained before and after adjusting for sex,
parity, maternal age, maternal education, maternal atopy, and environmental tobacco
smoke exposure. Cord and concurrent blood Pb levels were weakly correlated (r = 0.29),
providing support for an independent association for prenatal Pb biomarkers. The
independent effects of Pb also were substantiated by observations that indicators of other
exposures, including blood levels of mercury, polycyclic 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. While associations were observed with relatively
low cord blood Pb levels (geometric mean: 1.16 (ig/dL [95% CI: 0.12, 1.22]), it is
uncertain the extent to which higher past Pb exposures of the mothers may have
influenced their pregnancy blood Pb levels and newborn cord blood Pb levels. Among
children approximately age 10 years in Hong Kong, Hon et al. (2010; 2009) found
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correlations between low concurrent blood Pb levels (population means < 2 (ig/dL) and
severity of atopic dermatitis (Spearman r = 0.46, p <0.001 and r = 0.33, p <0.005),
another inflammatory condition commonly related to elevated IgE levels. Neither
analysis considered potential confounding variables.
Among the studies in children that found associations of blood Pb level with asthma and
allergy-related outcomes, several adjusted for potential confounding by SES-related
variables. Joseph et al. ("2005) adjusted for annual income. Pugh Smith et al. (2011). who
examined a primarily low SES population of children in Michigan, adjusted for various
factors associated with SES, including multiple indices of housing condition, and
presence of pets and cockroaches in the home, which are indicators of allergens 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. Thus, while SES and/or
allergen exposure has been associated with both Pb exposure and asthma and allergy in
children (Bryant-Stephens. 2009; Dowd and Aiello. 2009; Aligne et al.. 2000). these
collective findings in different populations do not demonstrate that confounding by SES
and/or allergen exposure fully accounts for the associations observed between blood Pb
levels and asthma and allergy. As allergic sensitization, asthma, and elevated IgE have
been correlated in children, the evidence of association of blood Pb level with asthma and
allergy after adjusting for SES, housing conditions, and allergen exposures, may provide
support for the associations observed between higher blood Pb and higher IgE in children
(Section 5.6.3).
Studies in nonoccupationally-exposed adults did not find associations of biomarkers of
Pb exposure with asthma or allergy (Pizent et al.. 2008; Mendv et al.. In Press). The
largest of these studies was a U.S. NHANES 2007-2008 analysis of adults ages 20 years
and older, in which urinary Pb level was not associated with an increase in asthma (OR:
0.72 [95% CI: 0.46, 1.12] per 1 jj.g/g increase in creatinine-adjusted urine) (Mendv et al.
In Press). The results did not provide strong evidence that urinary Pb level was associated
with other respiratory conditions such as emphysema or chronic bronchitis either. In a
study of 216 adults without occupational Pb exposures, Pizent et al. (2008) found that
among women, the association between concurrent blood Pb level and serum IgE was
statistically significant, whereas the association with positive SPT to common inhaled
allergens was not. Among men, higher concurrent blood Pb level was associated with
lower odds of positive SPT (OR: 0.92 [95% CI: 0.86, 0.98] but was not statistically
significantly associated with IgE. These findings appeared to be discordant because an
increase in IgE commonly mediates the acute inflammatory response to allergens.
However, the interpretation of the findings is difficult because only statistically
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significant effect estimates were reported; thus it is not known whether odds ratios were
in the same direction for SPT and IgE. Bener et al. (2001^ found higher prevalence of
asthma and allergy-related conditions such as rhinitis and dermatitis among Pb industrial
workers than among control subjects; however, the blood Pb levels in both the Pb-
exposed group and the control group (geometric means: 77.5 and 19.8 (ig/dL,
respectively) were higher than those in the current U.S. general adult population.
In summary, an available recent epidemiologic study added evidence for associations
between blood Pb level and asthma in children (Pimh Smith and Nriagu. 2011). and a few
recent studies provided new evidence for associations with allergic conditions in children
(Jedrvchowski et al. 2011; 2010; Hon et al. 2009). This epidemiologic evidence was
strengthened by several observations that associations were robust to the adjustment of
multiple potential confounding factors, in particular, SES and allergen exposures.
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 of blood Pb measurement (prenatal, sometime in childhood prior to
outcome assessment, concurrent), or level of blood Pb. The epidemiologic evidence for
associations of blood Pb level with asthma and allergy is well supported by toxicological
and epidemiologic evidence indicating Pb effects on increasing IgE (Section 5.6.3), Th2
cytokines (Section 5.6.5.4), and inflammation (Section 5.6.5.1). In the few recent studies
that investigated nonoccupationally-exposed adults, Pb biomarker levels were not
associated significantly with greater asthma or allergy (Pizent et al. 2008; Mendv et al..
In Press).
5.6.4.3 Other Respiratory Effects
The respiratory effects of Pb have been examined primarily in a small number of recent
epidemiologic studies of nonoccupationally-exposed adults, and as with evidence for
asthma, evidence for Pb-associated respiratory effects in adults is weak. Increased
bronchial responsiveness (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 a study of 525 middle-aged adults 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]) of 0.018 (95% CI: 0.004, 0.03), adjusting for age, sex, height, smoking, lung
function, and asthma diagnosis (Min et al.. 2008a'). The concurrent blood Pb levels in
these adults were low (mean [SD]: 2.90 [1.59] j^ig/dL); however, it is uncertain what
timing, level, frequency, and duration of Pb exposures contributed to the observed
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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 men (2.4% decrease [95% CI:
-4.2, -0.52%] in percent change FEV, post-histamine challenge per 1 (ig/dL increase in
blood Pb level). Although this finding was not discussed in details by investigators, it was
consistent with the observation that higher blood Pb level was associated with lower odds
of positive SPT among men in this study.
In studies of male Pb workers, Pb-associated respiratory effects were not clearly
indicated. None of the studies directly examined associations between blood Pb levels
and lung function. In bus drivers in Hong Kong, China, (Jones et al.. 2008; Jones et al..
2006). drivers of non-air conditioned buses had lower exposures to PMi0, lower blood Pb
levels (mean 3.7 j^ig/dL versus 5.0 (ig/dL in air conditioned buses) but lower indices of
lung function than did drivers of air conditioned buses (Jones et al.. 2006). In this study,
the authors attributed the slightly higher blood Pb levels of air conditioned bus drivers to
the poor efficiency in the filters and higher PMi0 levels measured on those buses versus
the non-air conditioned buses. In a comparison of roadside vendors and adjacent
shopkeepers, blood Pb levels and various lung function parameters were similar between
groups (Jones et al. 2008). Pb industrial workers in the United Arab Emirates had higher
prevalence of respiratory symptoms such as cough, phlegm, shortness of breath, and
wheeze than did unexposed controls (Bener et al. 2001b). Blood Pb levels in both the Pb-
exposed group and the control group (geometric means: 77.5 and 19.8 (ig/dL,
respectively) were higher than those in the current U.S. adult general population, and the
analysis did not consider potential confounding variables.
Toxicological evidence for Pb-associated respiratory effects is provided by observations
that exposure to Pb-containing PM induces a range of inflammatory-related effects in the
airways of animals and in cultured airway cells (Section 5.6.6). Specific effects on the
lung also were demonstrated in a recent study of rats injected with Pb-acetate (25 mg/kg,
3 consecutive days, resulting in 2.1 (ig/dL blood Pb) (Kaczvnska et al.. 2011). Pb-treated
rats exhibited ultrastructural changes in lung tissue, including substantial pulmonary
fibrosis containing numerous lipofibroblasts, collagens, and elastin filaments in the
interstitium. Mast cells also were present in the interstitium following Pb exposure.
Pulmonary inflammation was observed, evidenced by the increased recruitment of
monocytes and thrombocytes inside the capillary vessels and increased macrophage
accumulation in the alveolar space. Evidence of damaged surfactant lining and
destruction of the laminae inside lamellar bodies of epithelial type II cells also was
observed. While the observed pulmonary histological changes have been linked with
functional pulmonary decrements in other studies (unrelated to Pb exposure), they were
observed with injected Pb. It is not clear whether routes of Pb exposure more relevant to
those in humans would result in similar effects.
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5.6.4.4 Autoimmunity
Evidence for the effects of Pb on increasing the risk of autoimmunity is provided
primarily by a small number of older toxicological studies reviewed in the 2006 Pb
AQCD in which Pb exposure of animals (pre- and postnatal) was associated with the
generation of autoantibodies (Hudson et al. 2003; Bunn et al.. 2000; El-Fawal et al.
1999; Waterman et al.. 1994). Whereas some evidence linked this risk of autoimmunity
to a shift toward Th2 responses, other evidence pointed to a shift toward Thl responses.
While recent studies did not examine Pb-induced production of autoantibodies, some
provided indirect evidence by indicating that the changes induced by Pb had broader
implications for increasing risk of autoimmunity. For example, Kasten-Jolly et al. (2010)
examined the impact of developmental (100 (.iM Pb-acetate in drinking water of dams
from GD8 to PND21, resulting in pup blood Pb levels 10-30 (ig/dL) Pb exposure of mice
on changes in gene expression in the spleen. Investigators found that Pb upregulated
digestive and catabolizing enzymes that could lead to the generation of self-peptides,
which in conjunction with other Pb-induced immunomodulatory effects, had the potential
to induce the generation of autoantibodies. In Carey et al. (2006). the activation of
neoantigen-specific T cells in PbCl2-treated adult mice (25-50 |_ig i.p.) also indicated the
potential for autoantibody generation. Evidence of Pb-associated autoimmune responses
in humans is limited to an older study of male Pb battery workers with blood Pb levels
ranging from 10 to 40 (ig/dL (El-Fawal et al.. 1999). In this study, the Pb-exposed
workers had higher levels of IgM and IgG autoantibodies to neural proteins compared
with unexposed controls (blood Pb levels not reported) (El-Fawal et al.. 1999). Pb also
was found to modify neural proteins in rats (Waterman et al. 1994). The generation of
self-peptides and modification of proteins can result in formation of neoantigens, thereby
increasing the risk of autoimmune reactions.
5.6.4.5 Specialized Cells in Other Tissues
As discussed in Section 5.6.2.4, Pb exposure consistently has been found to alter the
function of macrophages. Nonlymphoid tissues also contain specialized macrophages
whose altered function can contribute to organ/tissue dysfunction, cell death, tissue
pathology and tissue-specific autoimmune reactions. Among the specialized macrophages
are microglia and astrocytes in the brain, Kupffer cells in the liver, alveolar macrophages
in the lung, keratinocytes and Langerhans cells in the skin, osteoclasts in the bone, and
preadipocytes in adipose tissue. The evidence demonstrating the effects of Pb on
specialized macrophages, which was provided by recent toxicological studies, is
important as it demonstrates the contribution of immune dysfunction to the effects of Pb
on dysfunction in nonlymphoid tissues (Figure 5-44). Because these specialized cells are
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not always recognized as macrophages, the resulting diseases and conditions are not
always recognized as being linked with Pb-induced immune dysfunction. It is important
to note that in this small body of available studies, several studies treated animals by Pb
injection, and it is unclear whether the effects would be observed with routes of exposure
that are more relevant for humans.
Residentbone
macrophages
i
Osteoblasts and
Osteoclasts
Periodontitis
and
Chondrogenesis
Kupffer cells
Microglia
and Astrocytes
Alveolar
macrophages
Neurobehavioral and
neurodegenerative
conditions
Altered liver metabolism
Oxidative damage
Glutathionestatus
Bronchial hyper-reactivity
Respiratory allergy
Asthma
Susceptibilityto infections
Figure 5-44 Specialized macrophages in nonlymphoid tissue may serve as a
link between Pb exposure and disease in multiple organ systems.
Fan et al. (2009b) reported that Kupffer cells undergo significant changes in phenotypic
expression (e.g., CD68 and ferritin light chain), organization, and functional activity
connected to Pb-induced apoptosis in the liver. 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 was observed to produce chronic glial
activation, increase pro-inflammatory cytokines, and increase neurodegeneration, as
indicated by an increase in GFAP, increase in IL-1 (3, TNF-a and IL-6 cytokines, and a
decrease in synaptophysin (component of presynaptic vesicles), respectively, in brain
tissue (Struzvnska et al.. 2007). In bone, osteoclasts regulate osteoblast function (Chang
et al.. 2008a). and osteoblasts have been shown to be affected by Pb exposure. Effects on
these cell types can contribute to later life diseases such as arthritis [reviewed in Zoeger
et al. (2006)1. Pb-induced elevation of TGF-(3 production was found to be involved in
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chondrogenesis in bone (Zuscik et al.. 2007). Kaczynska et al. (2011) reported effects on
alveolar macrophages after Pb treatment (i.p. 25 mg/kg, 3 days, resulting in blood Pb
levels of 2.1 (ig/dL) in rats. Macrophage recruitment increased, and this macrophage
infiltration limited 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 Pb
exposure. Pace et al. (2005) reported that Pb exposure in mice contributed to poor
reproductive performance that was concomitant with altered homeostasis of the testicular
macrophage population in that organ. The authors proposed that increased oxidative
damage and apoptosis among these macrophages and reduced potential to maintain organ
homeostasis contributed to the observed pattern of male sterility.
5.6.4.6 Tumors
While toxicological evidence indicates that high concentration Pb exposures directly
promote tumor formation or induce mutagenesis and genotoxicity (Section 5.10),
evidence for involvement of the immune system is limited. Kerkvliet and Baecher-
Steppan (1982) observed that male C57B1/6 mice exposed to 130 and 1,300 ppm of
Pb-acetate in drinking water had enhanced moloney sarcoma virus-induced tumor growth
compared with control animals. The findings indicated that Pb-induced
immunomodulation affecting tumors likely resulted from a combination of suppressed
Thl responses and increased inflammation leading to excessive release of ROS into
tissues. The promotion of cancer is a relatively common outcome in chemical-induced
immunotoxicology, particularly when early life exposures are involved (Dietert. 2011).
5.6.5 Modes of Action for Lead Immune Effects
5.6.5.1 Inflammation
The 2006 Pb AQCD indicated that misregulated inflammation represents one of the
major immune-related effects of Pb and a major mode of action for Pb effects in multiple
organ systems such as the liver, kidney, and vasculature (Section 5.2.5). In the 2006 Pb
AQCD, several lines of toxicological evidence demonstrated Pb-induced inflammation:
Pb-induced increases in PGE2 and ROS (Section 5.6.5.2) and increases in pro-
inflammatory cytokines (Section 5.6.5.4), both of which promote a hyperinflammatory
phenotype in immune cells. As described in sections that follow, these findings are
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corroborated by a small number of available recent toxicological studies. A few available
recent epidemiologic studies provided new evidence of Pb-associated inflammation, as
indicated by associations of blood Pb levels with pro-inflammatory cytokines and other
nonspecific indicators of inflammation.
Enhanced inflammation and tissue damage occurs through the modulation of
inflammatory cell function and production of pro-inflammatory cytokines and
metabolites. It is important to note that while inflammation can mediate immune-related
conditions such as asthma, respiratory infections, and BR, inflammation can exacerbate
disease and damage in almost any organ given the distribution of immune cells as both
permanent residents and infiltrating cell populations. Among the problems presented by
this immunomodulation are the overproduction of ROS and an apparent depletion of
antioxidant protective enzymes and factors (e.g., selenium). For example, toxicological
studies have found that Pb-induced inflammatory damage involves the depletion of
antioxidants such as glutathione and catalase (Lodi et al.. 2011; Chettv et al.. 2005V
While several processes have been proposed to explain the mechanisms of Pb-induced
oxidative damage, the exact combination of processes involved remains to be determined
(Section 5.2.4).
In epidemiologic studies, Pb-associated changes in pro-inflammatory cell function
(Section 5.6.2) and cytokine production (Section 5.6.5.4) have been found in the
cumulative body of evidence. A few recent epidemiologic studies added to this evidence
with observations of associations between blood Pb level and other nonspecific indicators
of inflammation that may be related to multisystemic effects as have been demonstrated
in animal and in vitro studies. Using 1999-2004 NHANES data, Songdej et al. (2010)
examined the relationship between concurrent blood Pb levels and the inflammation
markers, C-reactive protein (CRP), fibrinogen, and white blood cell (WBC) count in
adults 40 years of age or older. Adjusting for age, sex, race/ethnicity, education, income,
BMI, physical activity, smoking status, diabetes status, inflammatory disease status, and
cardiovascular disease status, investigators found larger magnitudes of association
between blood Pb and inflammation in men compared with women. Among women,
most ORs for associations between quintiles of blood Pb level and tertiles of CRP,
fibrinogen, and WBC count were less than 1.0 whereas corresponding ORs in men tended
to be greater than 1.0 but were not always statistically significant. For example, compared
with men with concurrent blood Pb levels less than 1.16 (ig/dL, men with blood Pb levels
of 1.16-<1.63 (ig/dL, 1,63-<2.17, 2.17-<3.09 (ig/dL, and > 3.09 (ig/dL had elevated odds
of elevated 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). For all inflammation markers, although the OR was
highest in the highest quintile of blood Pb level (> 3.09 (ig/dL), monotonic concentration-
dependent increases were not observed.
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Consistent with these findings, among men in Incheon, Korea without occupational Pb
exposures, Kim et al. (2007') reported associations of higher concurrent blood Pb level
with higher levels of WBCs and IL-6. Larger effects were estimated for 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. The findings are consistent with Pb effects on promoting a Th2 phenotype. Th2
cells produce IL-6 which is the primary stimulus for expression of CRP and fibrinogen
(Hage and Szalai. 2007; Fuller and Zhang. 2001). Despite the low concurrent blood Pb
level of adults in these aforementioned studies, it is important to acknowledge that the
relative contributions of recent versus past Pb exposures to the associations observed with
concurrent blood Pb levels are not delineated.
In a genome-wide association study that included 37 children with autism and 15 children
without autism (ages 2-5 years; blood Pb level range: 0.37 to 5.2 (ig/dL) in California, in
models that included age, sex, and autism diagnosis, concurrent blood Pb level was
associated with the expression of several genes related to immune function and
inflammation, including human leukocyte antigen genes (HLA-DRB) and MHC Class II-
associated invariant chain CD74 (involved in antigen presentation) (Tian et al.. 2011).
Although blood Pb levels were similar between children with and without autism and
correlations were observed in both groups, they were in opposite directions (positive
among children with autism and negative among children without autism). Pb has been
shown to increase MHC molecule surface expression in mouse and human HLA antigen
presenting cells (Guo et al.. 1996a; McCabe and Lawrence. 1991); however, additional
larger studies with a priori hypotheses regarding specific indices of inflammation are
warranted to characterize Pb-associated changes in inflammation in children.
5.6.5.2 Increased Prostaglandin E2 and Decreased Nitric Oxide
Consistent with the large body of evidence presented in the 2006 Pb AQCD (U.S. EPA.
2006b). a small number of available recent studies continued to indicate that Pb exposure
alters the levels of signaling molecules such as PGE2 and NO. Collectively, the weight of
evidence was provided by toxicological studies. These signaling molecules are involved
in mediating inflammation and host resistance (Figure 5-42). A recent in vitro study with
human neuroblastoma cells found increases in PGE2 with lower Pb concentrations (0.01-
1 (.iM) than those previously reported (Chettv et al. 2005). A large body of evidence
reviewed in the 2006 Pb AQCD demonstrated a Pb-associated decreased production of
NO by macrophages [see 2006 Annex Table AX5.9.6 (U.S. EPA. 2006h)l.
In adult animal models, decreases in NO were observed with short-term exposures (hours
to days) to a wide range of Pb concentrations. Decreases in NO can impact not only
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innate host defenses, but also, acquired immunity. In a recent study, Farrer et al. (2008)
found that high-level Pb glutamate exposure (5 (.iM) of lymphocyte cultures led to
decreased NO production, decreased inducible NO synthase function in myeloid cells,
but no change in its gene expression. Additionally, Pb abrogated the myeloid cell
(CD1 lb+)-mediated suppression of CD4+ T cell proliferation. 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 signal against pathogenic challenge, evidence indicated that multiple
arms of the host defense against infectious challenge can be compromised. The loss of
NO production in innate immune cells such as macrophages would be expected to affect
other physiological systems (e.g., neurological, cardiovascular, endocrine) that require
NO signaling cascades.
Relative to studies in animal and in vitro models, fewer epidemiologic studies have
examined the effects of Pb on signaling molecules; however the limited data supported
associations of blood Pb level with suppressed NO production (Barbosa et al. 2006a;
Pineda-Zavaleta et al.. 2004) and increased ROS production (Pineda-Zavaleta et al.
2004) in populations living near Pb sources. In a previous study of children in Mexico,
with increasing residential proximity to the Pb smelter, mean concurrent blood Pb levels
increased (7.02 to 20.6 to 30.38 (ig/dL) as did superoxide anion release from
macrophages (directly activated by IFN-y/LPS) isolated from children (Pineda-Zavaleta
et al.. 2004). NO release from macrophages (indirectly activated by phytohemagglutinin,
PHA) was lower with higher blood Pb levels. After adjusting for age and sex, a 1 (ig/dL
higher blood Pb level was associated with a higher level of superoxide anion of 0.00389
(95% CI: 0.00031, 0.00748) (imol/mg protein and a lower level of NO of 0.00089 (95%
CI: -0.0017, -0.00005) nmol/(.ig protein. Because PHA activates macrophages indirectly
through the activation of lymphocytes and IFN-y directly activates macrophages, these
results indicated that Pb suppressed T cell-mediated macrophage activation and
stimulated cytokine-induced macrophage activation. Group-level comparisons indicated
that associations likely were driven by changes observed in the group of children living in
closest proximity to the smelter who had blood Pb levels 10.31-47.49 (ig/dL. Results also
demonstrated a larger magnitude of association between blood Pb levels and superoxide
anion release in males. Although not described in detail, higher blood Pb level was not
associated with lower NO in girls. Barbosa et al. (2006a) also observed an association
between higher blood Pb level and lower plasma NO in a group of adults in Sao Paolo,
Brazil residing near a closed battery plant, particularly among adults with the TC or CC
eNOS genotype (r = 0.23, p = 0.048) which is associated with reduced promoter activity
and potentially reduced gene expression. Quantitative results were not reported for
analysis of all adults combined or adults with the TT genotype, but p-values were greater
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than 0.05. Because NO was measured in plasma, it was not possible to identify immune
cells as the specific sources of NO.
Studies of occupationally-exposed adults provided less clear indication of associations of
blood Pb level with NO and ROS. Despite large differences in blood Pb levels between
30 male Pb recycling plant 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'). In a study of male foundry workers (mean
blood Pb level: 21.7 (ig/dL), pottery workers (mean blood Pb level: 9.7 j^ig/dL). and
unexposed workers (mean blood Pb level: 3.9 (ig/dL), Valentino et al. (2007) found
similar plasma NO levels in controls compared with Pb-exposed workers. Also, although
quantitative results were not reported, blood Pb level was reported not to be correlated
with NO.
5.6.5.3 Cellular Death (Apoptosis, Necrosis)
The 2006 Pb AQCD reviewed a small number of toxicological studies in which Pb
exposure had contrasting effects on the apoptosis of macrophages. Since then, a few
available recent toxicological studies found that Pb exposure induces apoptosis or
mediators of apoptosis in immune cells. In a study in mice, Bishayi and Sengupta (2006)
found that Pb treatment of adult mice (10 mg/kg, i.p.) elevated DNA fragmentation in
splenic macrophages. Using mouse resident peritoneal macrophages, Gargioni et al.
(2006) found that 20 and 40 (.iM Pb nitrate induced both necrosis and apoptosis in vitro.
While the exact pathways involved were not determined, the authors concluded that
activation of the Bax pro-apoptotic protein was not the key effect of Pb on inducing
macrophage apoptosis. In an in vivo study in 3 week-old mice, Xu et al. (2008) found
that a 4-week administration of Pb-acetate (50-100 mg/kg, oral) significantly elevated
both ROS and malondialdehyde (an indicator of ROS-induced peroxidation) levels in
peripheral blood lymphocytes. Pb also induced DNA damage (determined by the comet
assay), which was accompanied by elevations in p53 and Bax expression with no change
in Bel-2 expression (creating a Bax/Bcl-2 imbalance). The authors proposed that
oxidative stress was a likely route to Pb-induced apoptosis and tumorigenesis. Because of
the high Pb concentrations administered and systemic administration of Pb employed,
other studies are warranted to examine whether apoptosis of macrophages represent a
mode of action relevant to human Pb exposures.
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5.6.5.4
Cytokine Production
The 2006 Pb AQCD presented a large body of toxicological evidence that clearly
demonstrated Pb-induced changes in cytokine production in vivo and in vitro (U.S. EPA.
2006b). Previous toxicological studies indicated that Pb affects immune cytokine
production via action on T cells and macrophages. The combination of cytokine changes
induced by Pb in multiple cell types can create a hyperinflammatory state among innate
immune cells to skew acquired immunity responses away from Thl responses and toward
Th2 responses. As illustrated in Figure 5-42, downstream effects include altered IgE
production, ROS production, and inflammation. In support of a Pb-induced skew toward
Th2 responses, several studies in rats, mice, and chickens found that pre- and postnatal
Pb exposure 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.
2006g)]. Recent toxicological studies continued to find such Pb-induced changes in
cytokine production, and a study provided new evidence that Th2 skewing may be
mediated via effects on dendritic cells.
Cheng et al. (2006) found that Pb exposure affected TNF-a production in vitro in A/J
mice macrophages by affecting the mitogen-activated protein kinase (MAPK) signaling
pathway. Pb-acetate (10 (.iM) co-administered with LPS stimulated the phosphorylation
of p42/44 MAPK and TNF-a expression (Cheng et al.. 2006). Blocking protein kinase C
or MAPK reduced TNF-a production of macrophages in vitro, which in turn, protected
against Pb + LPS-induced liver injury in vivo. Thus, Pb exposure may induce local tissue
damage through the modulation of immune responses. These findings were consistent
with those from recent studies. Gao et al. (2007) showed that treatment of mouse
dendritic cells with 25 (.iM PbCl2 produced an increased phosphorylation of the
Erk/MAPK signaling molecule, and Khan et al. (2011) showed that monocytes treated
with 25-50 (.iM Pb-acetate increased TNF-a through ERK1/2 and p38 signaling.
In an in vivo study conducted across a lifetime (developmental through adulthood) in
Swiss mice (females and males) using a broad range of dietary Pb concentrations, Iavicoli
et al. (2006a) found a nonlinear hierarchical cytokine response. At the lowest dietary Pb
concentration (0.11 ppm Pb-acetate, resulting in blood Pb level: 1.6 |_ig/dL). IL-2 and
IFN-y were decreased compared to those in the controls (0.02 ppm Pb-acetate, resulting
in blood Pb level: 0.8 (ig/dL), indicating a suppressed Thl response. As dietary Pb
exposure 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 differentially modulated by low-level versus
high-level Pb exposures. Other studies found variable Pb-induced changes in IL-2, with
no change or elevated production, depending upon the protocol used. Recently, Gao et al.
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(2007) found that Pb-treated dendritic cells (25 |_iM) promoted a slight but statistically
significant increase in IL-2 production (quantitative results not reported) among
lymphocytes.
In vitro studies also reported a Pb-induced shift to production of Th2 cytokines. In
conjunction with other indicators of Th2 skewing described in Sections 5.6.2.3, 5.6.2.6,
and 5.6.3, Gao et al. (2007) observed that 25 (.iM Pb elevated the production of Th2
cytokines such as IL-4, IL-5 and IL-6 in dendritic cells. In cultures of human PMNs
activated with Salmonella enteritidis or with monoclonal antibodies of CD3, CD28, and
CD40, Pb-acetate concentrations 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 (Hemdan et al.. 2005).
Consistent with toxicological studies, a few available 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 (p
= 0.08, Kruskal Wallis) and IgE (Section 5.6.3) than did children with lower blood Pb
levels. However, IL-4 levels in children with blood Pb levels 20-44 j^ig/dL were lower
than those in children with blood Pb levels <15 (ig/dL. The elevated IL-4 and IgE in
children with blood Pb levels 15-19 j^ig/dL 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 of
children in grades 5 and 6 in Taiwan, investigators did not group children by blood Pb
levels but by potential for Pb exposures due to age of home and location of residence
(Hsiao et al.). 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). This latter group of children also had the lowest serum levels
of IFN-y (45-fold) and highest levels of IL-4 (6-fold) (lower p < 0.05 for comparisons
with any subgroup). 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
(p > 0.05 for comparisons with any subgroup). Thus, the differences in cytokine levels
between healthy and allergic children living near the oil refinery may have been
influenced by differences in their blood Pb levels.
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Evidence of association between blood Pb levels and cytokine levels in
nonoccupationally-exposed adults was equivocal. Among 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 males in the upper two quartiles
of blood Pb levels, 2.51-10.47 (ig/dL. Notably, the relative contributions of recent versus
past Pb exposures to these associations are not known. In models that adjusted for age,
BMI, and smoking status, 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 (1.14 [95% CI: 0.20, 2.10] pg/mL higher per 1 (ig/dL higher
blood Pb level) than men who were GSTM1 sufficient (0.12 [95% CI: -0.43, 0.67]
pg/mL). Blood Pb also was associated with a larger increase in TNF-a among men who
had the TNF-a GG genotype (0.80 [95% CI: 0.15, 1.45] pg/mL higher per 1 (ig/dL higher
blood Pb level) than men who had the GA or AA genotype (-0.21 [95% CI: -1.1, 0.71]
pg/mL). 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, it is biologically
plausible that 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 were 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. Among adults in Italy, concurrent blood Pb
levels were not statistically significantly correlated with either Th2 or Thl cytokine levels
in men (Boscolo etal.. 1999) or women (Boscolo et al.. 2000) (quantitative results not
reported).
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, analysis were mostly limited to comparisons of mean cytokine levels among
different blood Pb groups or Pb exposure groups (Di Lorenzo et al.. 2007; Valentino et
al.. 2007; Yucesov et al.. 1997a) that did not consider potential confounding variables.
The exception was a study of male foundry workers, pottery workers, and unexposed
workers (Valentino et al.. 2007). Multiple regression analyses were performed with age,
BMI, smoking, and alcohol consumption included as covariates; however, regression
coefficients describing the concentration-response functions were not reported. Pb-
exposed workers had higher IL-10 and TNF-a (ANOVA, p < 0.05). Levels of 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 lower IL-4 levels
compared with controls (ANOVA, p > 0.05). In a similar analysis, DiLorenzo et al.
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(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 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 level group. Results also indicated a potential interaction
between blood Pb level and smoking. Among current smokers, a 12- to 16-fold difference
in TNF-a levels was observed among blood Pb groups. Among nonsmokers, the
differences were less than two fold. In Yucesoy et al. (1997a). levels of the Thl
cytokines, IL-1(3 and IFN-y, were lower in workers than in controls.
In summary, a large body of previous toxicological evidence clearly demonstrated Pb-
induced increases in Th2 cytokines and decreases in Thl cytokines. Several recent
toxicological studies added to this evidence by showing Pb-induced increases in TNF-a
that were mediated by MAPK signaling pathways (Khan et al.. 2011; Gao et al.. 2007;
Cheng et al.. 2006; lavicoli et al.. 2006a; Hemdan et al.. 2005). and demonstrating Pb-
induced increases in cytokine production in dendritic cells (Gao et al.. 2007). The
collective epidemiologic evidence is sparse; however, in the few available new studies,
higher concurrent blood Pb level was associated with higher levels of various pro-
inflammatory cytokines in children, nonoccupationally-exposed adults, and
occupationally-exposed adults. The recent epidemiologic study of children in Taiwan
found an association between higher IL-4 levels in children with higher blood Pb levels
(Hsiao et al.). similar to a previous study of children in Missouri (Lutz et al.. 1999).
Neither study of children considered potential confounding variables. In contrast to
previous studies, a recent study in nonoccupationally-exposed adults found associations
between higher concurrent blood Pb level and higher TNF-a and IL-6 cytokine levels in
men in Korea (Kim et al. 2007). Recent studies of occupationally-exposed adults added
to the evidence for higher levels of several cytokines among Pb-exposed workers
compared to unexposed controls (Di Lorenzo et al.. 2007; Valentino et al.. 2007). Due to
the limited investigation, it is difficult to draw conclusions about the effects of Pb
exposure on cytokine levels in any particular group in the human population.
5.6.6 Air-Lead Studies
Although comprising a smaller body of evidence than do immune studies of Pb
biomarkers, several recent studies used Pb measured in PMi0 and PM2 5 air samples to
represent Pb exposures. Some studies analyzed the Pb component individually, whereas
others analyzed Pb as part of a group of correlated components using source
apportionment techniques or principal component analysis. In concordance with blood Pb
studies, recent time-series epidemiologic studies that examined ambient air Pb-PM
concentrations found associations with respiratory morbidity in children (Gent et al.
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2009; Hong et al.. 2007b). Adjusting for season, day of week, and date, Gent et al. (2009)
found that an increase in lag 0-2 average Pb-PM2 5 was associated with elevated odds of
wheeze (OR: 1.03, p = 0.13 per 5 ng/m3 increase in Pb PM25), shortness of breath (OR =
1.12, p = 0.01), and fast-acting inhaler use (OR: 1.04, p = 0.10) among children with
asthma (ages 4-12 years), followed for 1 year during 2000-2003. Among children in
Korea (grades 3-6), a 1 log increase in lag 1 Pb PMi0 was associated with a -6.83 L/min
decrease (p < 0.01) in morning peak expiratory flow and a -6.37 L/min decrease
(p < 0.01) in daily average peak expiratory flow adjusting for age, sex, height, weight,
smoking exposure, and meteorological factors (Hong et al.. 2007b). Older and recent
toxicological studies also found Pb-containing CAPs to induce pulmonary inflammation.
Uzu et al. (2011) specifically tested particles emitted at a Pb recycling plant and found
that these Pb-rich particles increased the release of the cytokine granulocyte-macrophage
colony-stimulating factor from human epithelial cells. Other toxicological studies found
pulmonary inflammation induced by CAPs in which Pb was one of numerous
components (Wei et al.. 2011; Duval 1 et al.. 2008; Godleski et al.. 2002; Saldiva et al..
2002). As with blood Pb, epidemiologic studies did not consistently find associations of
ambient air Pb-PM concentrations with respiratory-related hospitalizations or mortality in
older adults. Both studies adjusted for meteorological factors and for temporal trends.
Among adults ages 65 years and older in 6 California counties, a 4 ng/m3 increase in lag
3 Pb-PM2 5 was associated with an increased relative risk of respiratory mortality all year
of 1.01 (95% CI: 0.99, 1.03) and during summer months (quantitative results not
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 lag 0 Pb-PM2 5 was associated with a
decrease in respiratory hospital admissions. The 95% CI was wide, indicating lack of
precision in effect estimate.
Although limited available recent findings suggest a relationship between respiratory
effects in children and short-term (over several days) changes in ambient air Pb-PM
concentrations, it is important to note uncertainties that limit the ability to draw
conclusions regarding airborne Pb exposure. Size distribution data for Pb-PM are
relatively limited, so it is difficult to assess the representativeness of these concentrations
to population exposure (Section 3.5.3). Moreover, data on the relationship between blood
Pb and air Pb are relatively limited (see Section 4.5.1) and do not characterize
relationships of varying Pb-PM size distribution with blood Pb level. In several air-Pb
studies, other PM components such as elemental carbon (EC), copper (Cu), and zinc (Zn)
also were associated with respiratory effects. In the absence of detailed data on
correlations among all PM components, measurements on other 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 particles comprised one component,
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it is not possible to attribute the observed associations or lack of associations specifically
to Pb (Samat et al. 2008; Andersen et al. 2007; Veranth et al. 2006; Macieiczvk and
Chen. 2005).
5.6.7 Immune Effects of Lead within Mixtures
Several toxicological studies published since the 2006 Pb AQCD examined immune
effects resulting from exposures to metal mixtures. Some studies indicated that immune
effects may be observed with lower levels of Pb exposure when they occur in conjunction
with other metals. In a study of mice treated with Pb-acetate (10 mg/kg i.p. injection by
weight, daily for 15 days), As (0.5 mg/kg i.p. injection by weight, daily for 15 days), or
both, Bishayi and Sengupta (2006) reported a greater than additive effect of co-
administered Pb and As on macrophages in decreasing bacterial resistance,
myeloperoxidase (MPO) release, and NO production. Investigators assessed the Pb-As
interaction on MPO release using the multivariate ANOVA and constructing an
isobologram by running an ordinary least squares regression between effects (% MPO
release) and dose levels of metals (single and multimetal) in log-linear form.
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 5.6.5.2) found interactions between Pb and As among children in Mexico
aged 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. In multiple regression analyses, urinary As was associated with lower NO
release from macrophages (similar to blood Pb). An interaction was observed between Pb
and As, which indicated that high internal doses of both metals were associated with a
larger decrease in NO than was either metal alone (p for interaction = 0.037). Urinary As
was associated with lower superoxide anion release (opposite direction of Pb). The
interaction between Pb and As indicated that higher internal doses of both Pb and As
were associated with a larger increase in superoxide anion than was blood Pb level alone
(p for interaction = 0.042). Due to the high blood Pb 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 children with lower blood Pb levels.
Results from Institoris et al. (2006) indicated that metal co-exposures potentiated the
effects of Pb. Lymph node weight decreased with exposure of 4 week-old rats to
20 mg/kg Pb-acetate by drinking water plus a second metal (Cd or Hg) but not with
20 mg/kg of Pb alone. In contrast with the aforementioned studies, Fortier et al. (2008)
did not find Hg co-exposure to increase the effects of Pb. PbCl2-exposed (7.5-20.7 (ig/dL)
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human leukocytes did not have alterations in lymphocyte proliferation, monocytic
phagocytic activity, or NK cell activity. The combination of 20.7 (ig/dL PbCl2 plus
12.0 (ig/dL methylmercuric chloride (MeHgCl) decreased lymphocyte proliferation;
however, these effects were attributed to MeHgCl, which had a stronger suppressive
effect independently. Other toxicological studies found immune effects of multiple metal
mixtures that included Pb (e.g., decreased antibody titers, increased neutrophil counts)
(Jadhav et al.. 2007; Massadch et al.. 2007) but did not test each metal individually. Thus,
it cannot be ascertained whether findings are due to interactions between Pb and other
components within the mixture. Overall, results indicated that exposures to Pb-containing
metal mixtures are associated with immune effects. Moreover, several studies found an
interaction between Pb and metals such as As, Cd, and Hg, suggesting that a threshold for
producing Pb-induced immune effects may be lower if additional metals are present.
5.6.8 Summary and Causal Determination
The collective body of evidence demonstrates that Pb exposure is associated with
immune effects. The evidence indicates that rather than resulting in overt cytotoxicity to
lymphoid tissues, Pb exposure is associated predominantly with subtle changes in a
spectrum of cellular and humoral immune responses, most notably, altered function of T
lymphocytes and macrophages, suppression of the DTH response and resistance to
bacterial infection, increased IgE production, changes in cytokine production, and
inflammation. The strength of evidence for Pb-associated immune effects is derived not
only from the consistency of findings within particular endpoints but also from the
coherence of findings among the spectrum of immune changes operating within the same
pathway. In particular, the evidence integrated across the examined spectrum of immune
outcomes clearly indicates that the prominent immune effect of Pb exposure is to shift
responses from a Thl phenotype toward a Th2 phenotype.
The weight of evidence is provided largely by an extensive body of animal studies
characterizing effects on the broad range of immune endpoints as described above.
Comprising a smaller body of evidence, epidemiologic studies in children are consistent
with toxicological studies in reporting associations of blood Pb levels with indicators of
increased Th2 activity, principally, higher levels of IgE, asthma, and allergy in children.
Across outcomes, because most studies examined concurrent blood Pb levels, there is
uncertainty regarding the critical timing, frequency, duration, and level of Pb exposures
in children that contributed to observed associations. In toxicological studies, the shift to
a Th2 phenotype is well characterized by observations that Pb exposures suppress the
production of Thl cytokines (e.g., IFN-y) and increase production of Th2 cytokines
(e.g., IL-4) (Section 5.6.5.4). In animal studies, the shift to Th2 cytokine production was
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observed in juvenile animals with prenatal Pb exposures and in adult animals with long-
term (> 4 weeks) Pb exposures [Table 5-7 in the 2006 Pb AQCD (U.S. EPA. 20062)1.
Blood Pb levels of animals were reported infrequently, and a wide range of Pb exposures
was found to induce changes in cytokine levels. In a recent study, lifetime Pb exposure
that resulted in blood Pb levels between 2-3 (ig/dL suppressed Thl cytokine production
in adult mice (lavicoli et al.. 2006a'). A few available epidemiologic studies in children
(Hsiao et al.. 2001; Lutz et al.. 1999) found higher levels of IL-4, a Th2 cytokine, in
groups of children with higher concurrent blood Pb levels (e.g., mean 8.8 (ig/dL or range
15-19 |_ig/dL. respectively); however, these studies are limited by their lack of
consideration of potential confounding variables and lack of information on the
concentration-response function.
Coherence for a Pb-associated skewed ratio of cytokine production is provided by
evidence demonstrating Pb-induced activation of T cells in animals, with recent studies
describing mechanisms underlying T cell activation (Sections 5.6.2.1 and 5.6.2.2). A
recent toxicological study expanded the extant evidence by showing in vivo and in vitro
that Pb may promote Th2 responses by directly increasing production of Th2 cytokines in
dendritic cells, the major effector in antigen response (Gao et al.. 2007). Pb-induced T
cell activation also is indicated by a relatively small body of evidence for prenatal and
postnatal Pb exposures of animals resulting in the generation of autoantibodies in
response to new antigens. These findings suggest that Pb exposure may increase the risk
of developing autoimmune conditions (Section 5.6.4.4). In epidemiologic studies,
evidence for the effects of Pb on T cells comprised associations of concurrent blood Pb
levels > 10 (ig/dL in children with lower T cell abundance, in particular CD3+ cells.
However, the functional relevance of these changes is unclear.
In addition to T cell responses, a prominent effect of Pb exposure, as demonstrated in an
extensive historical toxicological evidence base, was the induction of macrophages into a
hyperinflammatory state as characterized by enhanced production of ROS, suppressed
production of NO, enhanced production of TNF-a, and excessive metabolism of
arachidonic acid into immunosuppressive metabolites (e.g., PGE2). Consistent with these
observations, a previous epidemiologic study examined and found greater release of ROS
and lower release of NO from macrophages, primarily in children with concurrent blood
Pb levels 10.31-47.49 (ig/dL (Pineda-Zavaleta et al.. 2004). Misregulated inflammation
represents one of the major modes of action for Pb-induced immune effects.
Toxicological studies provide evidence for the modulation of inflammatory cell function,
production of pro-inflammatory cytokines and metabolites, enhanced inflammatory
chemical messengers, and pro-inflammatory signaling cascades. In addition to the
associations reported with IL-4, epidemiologic evidence for Pb effects on inflammation is
limited to a few recent studies in nonoccupationally-exposed adults in which concurrent
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blood Pb level was associated with other indicators of inflammation such as CRP
(Songdei et al.. 2010) and IL-6 (Kim et al.. 2007). The studies commonly adjusted for
potential confounding by age, sex, BMI, and smoking status. However, because only
concurrent blood Pb levels were examined, there is uncertainty regarding the magnitude,
timing, frequency, and duration of Pb exposures that contributed to the observed
associations.
The toxicological evidence for Pb-induced production of Th2 cytokines provides
biological plausibility for the evidence linking Pb exposure with elevated IgE levels. An
increase in IL-4 from activated Th2 cells induces differentiation of B cells into antibody-
producing cells, thereby amplifying B cell expansion to secrete IgE, IgA, and IgG.
Animal studies describe Pb-induced (resulting in blood Pb levels 10-30 (ig/dL) increases
in IgE (Snvder et al.. 2000; Miller et al.. 1998; Heo et al.. 1996). Additionally,
epidemiologic studies in children consistently demonstrated associations between higher
concurrent blood Pb levels and higher in serum IgE (Section 5.6.3). While most studies
found elevated IgE in groups of children with blood Pb levels >10 (ig/dL, Karmaus et al.
(2005) found higher serum in IgE in children with blood Pb levels 2.8-3.4 (ig/dL
compared with children with lower blood Pb levels. A few studies considered potential
confounding by factors such as age (Karmaus et al.. 2005; Lutz et al.. 1999). smoking
exposure, serum lipids, organochlorine biomarkers, and number of previous infections
(Karmaus et al.. 2005) but not SES or allergen exposure. It is important to acknowledge
the extensive evidence in animals for Pb-induced increases in IgE that is not subject to
confounding by SES. Additional support for effects on IgE is provided by toxicological
evidence for Pb-induced increases in Th2 cytokines. Epidemiologic studies did not find
blood Pb level to be associated consistently with B cell abundance.
The toxicological evidence for Pb-induced suppression of Thl cytokine production is
coherent with historical observations in animals that Pb exposure decreases responses to
antigens and bacterial infection (U.S. EPA. 2006b). Toxicological studies provide clear,
consistent evidence that prenatal and postnatal Pb exposure of animals (resulting in blood
Pb levels 11->100 (ig/dL) suppresses the DTH response to antigens (Section 5.6.2.3), and
a recent in vitro study indicates such effects may be mediated by dendritic cells (Gao et
al.. 2007). Recent toxicological studies provided additional support for Pb-induced
decreased host resistance by demonstrating that Pb exposure impairs phagocytic and
chemotactic activity of macrophages (Lodi et al.. 2011; Bussolaro et al.. 2008). These
observations reflect suppressed Thl activity given the role of Thl-dependent IFN-y in
enhancing the killing capacity of macrophages. Epidemiologic evidence for Pb-associated
diminished response to bacterial infection is limited to previous observations of reduced
neutrophil functionality in Pb-exposed workers with mean blood Pb levels > 30 (ig/dL.
While a few previous epidemiologic studies in children found higher prevalence of
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respiratory infections in children with higher blood Pb levels (karmaus et al. 2005;
Rabinowitz et al. 1990). the findings are limited by the weak statistical methods of
studies and the lack of consideration of potential confounding variables.
The effects of Pb exposure on macrophages also suggest a role for the immune system in
mediating Pb-associated effects in multiple other physiological systems. A small body of
new toxicological studies indicated Pb-induced changes in specialized macrophages in
nonlymphoid tissue such as alveolar macrophages, testicular macrophages, and brain
microglia (Section 5.6.4.5); however, these studies primarily used the i.p. route to
administer Pb. Thus, the relevancy of observations to those expected from typical routes
of human exposure is not clear.
Th2-dependent increases in IgE mediate type 1 hypersensitivity resulting in various
allergic conditions and asthma. Observations of Pb-associated increases in Th2 cytokines
and circulating IgE levels provide biological plausibility for the small body of available
epidemiologic evidence indicating associations of blood Pb levels with asthma and
allergic conditions in children (Jedrvchowski et al. 2011; Pugh Smith and Nriagu. 2011;
Joseph et al.. 2005). Several of these studies considered a larger set of potential
confounding variables than did studies of IgE in children. While the set of particular
factors varied among studies, studies frequently considered SES indicators and/or
residential allergen measurement. Low SES has been associated with higher blood Pb
levels, poorer housing conditions, higher exposures to mouse and cockroach allergen, and
with conditions such as asthma and allergy. Jedrychowski et al. (2011) found similar
magnitudes of association between cord blood Pb level and cord blood IgE in models that
did and did not adjust for potential confounding variables, including maternal education.
Studies examining cognitive effects in children also found that associations with blood Pb
levels with and without adjustment for SES-related variables (Section 5.3). Pugh Smith et
al. (2011) and Jedrychowski et al. (2011) also indicated lack of confounding by allergen
exposures and other indicators of housing condition by their inclusion as model
covariates or analysis of their independent associations with outcomes. The robust
association between blood Pb level and allergic sensitization (which indicates elevated
allergen-specific IgE) observed in Jedrychowski et al. (2011). provides support for the
associations observed between blood Pb and IgE, in children, in which confounding by
SES and allergen was not considered. Collectively, these findings do not indicate that
confounding by SES or allergen exposure alone accounts for associations observed
between blood Pb levels and immune effects in children. This evidence for associations
of blood Pb level with clinical conditions such as asthma, allergic sensitization, and
allergic diseases in children expanded by results from recent studies also supports the
public health significance for Pb-associated immune effects.
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With respect to lifestages of Pb exposure, animal studies found that gestational Pb
exposures, encompassing a wide range of concentrations, affected endpoints such as IgE,
cytokine levels, and DTH but also found postnatal short-term (multiple days) long-term
(multiple weeks) Pb exposures to affect cytokine levels in adult animals. The blood Pb
levels and Pb exposure lifestage, magnitude, frequency, and duration associated with
immune effects are not well characterized in humans. Several epidemiologic studies
examined associations (of IgE, asthma, and allergy) with concurrent blood Pb levels of
children >10 (ig/dL. A few studies found cord blood or newborn hair Pb levels to be
associated with endpoints such as Ig levels (Annesi-Maesano et al. 2003; Belles-Isles et
al.. 2002) and allergic sensitization (Jedrvchowski et al. 2011). The small body of
epidemiologic studies of nonoccupationally-exposed adults examined different endpoints
and found associations with concurrent blood Pb levels, which are influenced by current
Pb exposures as well as cumulative Pb stores in bone.
In summary, recent toxicological and epidemiologic studies support the strong body of
evidence presented in the 2006 Pb AQCD that Pb exposure is associated with a broad
spectrum of changes in both cell-mediated and humoral immunity that cumulatively
promote a Th2 phenotype and hyperinflammatory state. The principal findings are Pb-
induced increased production of Th2 cytokines, suppressed production of Thl cytokines,
increased inflammation, and elevated IgE, with the weight of evidence provided by
toxicological studies. Collectively, these findings are coherent with the observed effects
of Pb exposure on decreasing responses to antigens (e.g., DTH, bacterial resistance) in
animals. Both toxicological and epidemiologic studies in children provide evidence for
Pb-associated increases in IgE. The toxicological and epidemiologic findings for Th2
cytokines, IgE, and inflammation provide biological plausibility for associations
observed for blood Pb levels with asthma and allergic conditions in children.
Associations with asthma and allergy were observed after considering potential
confounding by several factors, including, SES and allergen exposure. Animal studies
found a range of immune effects with prenatal exposure in juvenile animals and long-
term postnatal (> 4 weeks) Pb exposures in adult animals. The blood Pb levels and Pb
exposure lifestage, magnitude, frequency, and duration associated with immune effects
are not well characterized in children or adults. Epidemiologic studies of children and
adults primarily examined concurrent blood Pb levels. Little information was provided on
concentration-response functions. In epidemiologic studies, higher IgE and higher asthma
prevalence were examined and found in children with blood Pb levels >10 (ig/dL. In the
large body of studies of adults (mostly males) with occupational Pb exposures, the most
consistent findings were decreased neutrophil functionality in workers with mean blood
Pb levels 21-71 (ig/dL. Recent epidemiologic studies provided new evidence in adults
without occupational Pb exposures; however, each examined a different immune
endpoint, for example, IgE, eNO, IL-6. These endpoints were associated with concurrent
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blood Pb levels in populations of adults with mean blood Pb levels of 1.9-7 (ig/dL;
however, there is uncertainty regarding the contributions of current Pb exposures and
cumulative Pb stores in bone. The consistency and coherence of findings across the
continuum of related immune parameters that demonstrate a stimulation of Th2 responses
in toxicological studies combined with the supporting epidemiologic evidence in children
are sufficient to conclude that there is a causal relationship between Pb exposures and
immune system effects.
5.7 Effects on Heme Synthesis and Red Blood Cell Function
5.7.1 Summary of Findings from 2006 Pb AQCD
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 blood Pb >
40 (ig/dL. Pb-induced anemia is thought to occur due to decreased RBC life span and
effects on hemoglobin (Hb) synthesis. The exact mechanism for these effects was not
known, although Pb-induced changes on iron uptake or inhibition of enzymes in the heme
synthetic pathway may be responsible.
The 2006 Pb AQCD indicated that Pb crosses RBC membranes through passive
(i.e., energy-independent) carrier-mediated mechanisms including a vanadate-sensitive
Ca2+ pump. Once Pb enters the cells, it is predominantly found in protein-bound form,
with Hb and aminolevulinic acid dehydratase (ALAD) both identified as targets. Pb
poisoning (blood Pb levels >100 j^ig/dL) was found to decrease RBC survival in
laboratory animals, as well as alter RBC mobility and morphology, although the precise
mechanisms by which it does so are not known. Pb exposure has been found to
significantly decrease several hematological parameters including Hb, hematocrit (Hct),
mean corpuscular volume (MCV), mean corpuscular hemoglobin (MCH), and mean
corpuscular hemoglobin concentration (MCHC). Pb has also been observed to exert
multiple 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 potentially lead to 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 phosphotidylserine (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 membrane
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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 affects heme synthesis through the inhibition of
multiple key enzymes, most notably ALAD, the enzyme that catalyzes the second, rate-
limiting step in heme biosynthesis (Figure 5-45 presents a schematic representation of the
heme biosynthetic pathway). The 2006 Pb AQCD 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 was observed to be not
dependent on the method of Pb administration. The inhibition of the ALAD enzyme was
observed in RBCs from multiple species, including birds, Cynomolgous monkeys, and
humans. Pb was also observed to inhibit other enzymes responsible for heme
biosynthesis, including ferrochelatase, porphobilinogen (PBG) deaminase, and
coproporphyrinogen oxidase. Pb also potentially alters heme biosynthesis through
inhibition of transferrin (TF) endocytosis and iron 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 possible 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.
Pb was found to alter the activity of membrane-bound ion pumps. Potassium (K+)
permeability was found to be increased by Pb due to altered sensitivity of the membrane
calcium (Ca2+)-binding site that caused selective efflux of K+ ions from the RBC
membrane. Inhibition of RBC sodium (Na+)-K+ adenosine triphosphate synthase
(ATPase), acetylcholinesterase (ACh), and NADH dehydrogenase was also observed. In
human RBCs, Na+-K+ ATPase activity was more sensitive to Pb exposure than were Ca2+
or magnesium (Mg2+) ATPases.
The 2006 Pb AQCD identified oxidative stress as an important potential mechanism of
action by which Pb exposure induced effects on RBCs. Increased lipid peroxidation and
inhibition of antioxidant enzymes (e.g., superoxide dismutase [SOD], catalase [CAT])
were observed following exposure to Pb.
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5.7.2
Red Blood Cell Functions
As stated in the 2006 Pb AQCD, Pb poisoning has been associated with anemia resulting
from shortened RBC life span and Pb effects on Hb synthesis. As of 2006, the
mechanism for this was not clear, but it was determined not to be due to iron deficiency,
which can be found to occur independently of Pb exposure. However, Zimmerman et al.
(2006) found that blood Pb levels were statistically significantly lower in non- or mildly
anemic, iron-deficient 5- to 9-year old children in India fed an iron-fortified diet for 30
weeks compared to 14 weeks (mean [range]: 8.1 [3.1-219] (ig/dL versus 12.1 [3.7-
26.8] (ig/dL; p < 0.02); blood Pb levels were not lower in children receiving the no iron
diet for 30 weeks compared to 14 weeks (mean [range]: 10.2 [4.4-25.3] (ig/dL versus
12.0 [3.8-25.5] (ig/dL). Although a number of epidemiologic studies find decreases in
RBCs and/or Hct levels associated with blood Pb, it is not known whether this is due to
reduced cell survival or a decrease in RBC cell production. However, 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.
5.7.2.1 Pb Uptake, Binding, and Transport into Red Blood Cells
The 2006 Pb AQCD 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) 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 0 (Section 4.2).
Although no studies were indentified that examined transport of Pb into RBCs, Lind et al.
(2009) recently observed that several zinc (Zn) ionophores (8-hydroxyquinoline
derivatives and Zn and Na pyrithione) were able to effectively transport Pb out of RBCs
into the extracellular space.
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5.7.2.2
Red Blood Cell Survival, Mobility, and Membrane Integrity
Although Pb exposure has been consistently shown to shorten the 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+. There are also newer studies that examine the relationship between Pb
and RBC survival, mobility, and membrane integrity. Occupational studies investigating
severely Pb-intoxicated worker populations (mean blood Pb > 28 (ig/dL) observed
increased intracellular RBC free calcium levels (|Ca2 |,) and decreased RBC membrane
Ca2+-Mg2+-ATPase activity in workers compared with unexposed controls (Abam et al..
2008; Ouintanar-Escorza et al. 2007). |Ca2 |, levels were highly correlated with blood Pb
levels even among unexposed control populations with blood Pb levels <10 j^ig/dL (mean
9.9 ± 2 (.ig/dL) (Ouintanar-Escorza et al.. 2007). Changes in |Ca2 |, were associated with
increased 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). concentration-
dependent effects were observed when RBCs from healthy human volunteers were
incubated in vitro with Pb nitrate at physiologically-relevant concentrations. |Ca2 |
increased in a concentration-dependent manner when RBCs were exposed 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) (Ouintanar-Escorza et al. 2010). |Ca2 |, levels were further increased at higher,
less physiologically-relevant concentrations (i.e., 2-6 (.iM Pb nitrate). The increase in
|Ca2 | levels was observed to be related to increased Ca2+ influx: RBCs exposed to
0.4 (.iM Pb nitrate for 24 hours incorporated twice as much Ca2+ as controls did, whereas
exposure for 120 hours resulted in a fourfold increase. Concomitant to increased influx of
Ca2+, efflux of Ca2+ was observed due to reduced Ca2+-Mg2+ ATPase activity; exposure to
0.4 |_iM Pb nitrate for 24 or 120 hours reduced activity by 33% and 83%, respectively,
compared to controls. As was observed among severely Pb-intoxicated workers, changes
in |Ca2 | were associated with increased fragility of the RBCs (measured as increased
hemolysis) and dramatic morphological alterations, including the increased presence of
echinocytes (cells without normal biconcave shape) and crenocytes (speculated cells)
following exposure to 0.4 (.iM Pb nitrate. 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 at 37 °C. Exposure of RBCs to
higher concentrations (concentrations not stated) of Pb nitrate resulted in cell shrinkage.
Ademuyiwa et al. (2009) observed that the cholesterol content of RBC plasma
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membranes, but not the phospholipid content, was statistically significantly higher in rats
exposed to 200 ppm Pb-acetate (resulting in mean [SD] blood Pb: 40.63 [9.21] j^ig/dL) for
three months through drinking water compared to controls. Further, the
cholesterol/phospholipid ratio was increased in the Pb-exposed rats with increased
cholesterol, indicating that RBC membrane fluidity was decreased.
A number of studies have investigated the effect of occupational exposure to Pb on
various inter-connected and related hematological parameters. Most occupationally-
exposed cohorts represent populations highly exposed to Pb, with mean blood Pb levels
ranging from 26-74 (ig/dL. Although effects observed within these cohorts may not be
generalizable to the population as a whole, they are useful in demonstrating consistent
effects on a number of hematological parameters, including Hb, MCV, MCH, MCHC,
total RBCs, and packed cell volume (PCV) (Khan et al.. 2008; Patil et al.. 2006a; Patil et
al.. 2006b; karita et al.. 2005). Additionally, these studies were cross-sectional in design;
thus, there is uncertainty regarding the magnitude, timing, frequency, and duration of Pb
exposure that contributed to the observed observations. Workers occupationally exposed
to Pb may also have been co-exposed to other contaminants than can affect the
hematological system although the potential for co-exposure was not assessed in most
studies. A few occupational studies did investigate the effect of moderate occupational Pb
exposure on hematological parameters. 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. Two
control groups were used for comparison (30 individuals from the same industries not
involved in handling Pb and 20 individuals from the same locality but not involved in Pb
handling). The exposed individuals had a mean blood Pb level of 7.00 (ig/dL compared to
3 (ig/dL in controls drawn from industries not involved in Pb handling (control group I)
and 2 (ig/dL in controls drawn from the general population (control group II) (p < 0.05).
Pb-exposed workers had significantly reduced Hb and PCV levels and increased
percentage of reticulocytes. Although the differences were statistically significant
between the exposed and control subjects, the study authors state that the levels in the
exposed subjects were at the lower range of normal for Nigerians. The percent cell lysis
did not 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). Conterato et al.
(In Press) 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 hematocrit, hemoglobin concentration, and the number of RBCs were
statistically significantly decreased in painters compared to controls, they were not
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correlated with blood Pb levels. These parameters were correlated with blood Cd levels
which were also statistically significantly elevated in painters compared to controls
(mean [SEM]: 1.606 [0.074] versus 0.003 [0.001] (ig/dL). MCHC and RBC distribution
width (a measure of the variability of RBC widths) were correlated (r = 0.23 and 0.44,
respectively; p < 0.05) with blood Pb although these measures were not statistically
different in painters compared to controls. In an investigation of petrol workers in
Sarajevo, Yugoslavia, examinations on hematological parameters were performed on the
same population five years apart, in 2003 and 2008. Workers (mean [SD] duration of
exposure: 12.1 [9.1] years) had increased blood Pb (mean: 5.96 j^ig/dL) in 2008 compared
to 2003 (mean: 4.07 (ig/dL; mean [SD] duration of exposure: 10.4 [5.5]) (Cabaravdic et
al.. 2010). In 2008, levels of MCH and MCHC were significantly decreased from levels
in 2003, whereas RBC numbers, Hb, Hct, and MCV were elevated. 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). Taken together, the above occupational
studies provide consistent evidence that very high (blood Pb > 26 (ig/dL) and moderate
(blood Pb ~ 5.5-7.0 (ig/dL) occupational exposure to Pb reduces the number of RBCs in
circulation. Although this decrease in RBCs observed in highly exposed worker cohorts
may be explained by both decreased cell 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, calcium, iron, and hemoglobin were
investigated in 55 pregnant Brazilian women (21.9% 14-19 years old, 74.5% 20-34 years
olds, and 3.6% > 35 years old) (Zentner et al.. 2008). The majority of women (across all
age groups) had concurrent blood Pb levels below 5 (ig/dL (58.2%), although the mean
blood Pb level was not reported; only 5.4% of women had blood Pb levels above
10 (ig/dL. The vast majority of women (78.2%) were also observed to have adequate
levels of hemoglobin (>11 g/dL). In a multiple linear regression model, blood Pb was
observed to be negatively correlated with hemoglobin ([3 = -0.359 g/dL).
Studies in children were generally supportive of effects on hematological parameters
(e.g., Hb, MCV, MCH) observed in occupational adult populations. In two cross-
sectional studies of children measuring blood hemoglobin as the independent variable,
blood Pb levels were observed to decrease with increasing blood Hb. Riddell et al. (2007)
found that 21% of children 6 months to 5 years of age living in rural Philippines had
concurrent blood Pb levels greater than 10 (ig/dL (total population mean: 6.9 (ig/dL). Hb
levels were inversely related to blood Pb, with a decrease of 3% blood Pb associated with
every 1 g/dL increase in Hb. Similarly, in children aged 6-36 months (n = 222) living in
Montevideo, Uruguay, 32.9% of children had blood Pb greater than 10 (ig/dL (population
mean [SD]: 9.0 [6.0] j^ig/dL) (Queirolo et al.. 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).
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Blood Pb levels were higher in anemic children compared to non-anemic (mean [SD]:
10.4 [6.8] versus 7.9 [5.1] (ig/dL), and in bivariate regression models, blood Pb decreased
0.65 (ig/dL for every 1 g/dL increase in Hb. In children younger than 18 months of age,
blood Pb was, on average, 3.5 ± 1.1 (ig/dL higher in anemic children. In children
aged 5-9 years (n = 189) living in Cartegena, 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 (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). 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 j^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 body mass index, group 2 had a significantly lower PCV compared to group 1. In a
related study, Ahamed et al. (2007) investigated the relationship between blood Pb,
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 <10 j^ig/dL (mean [SD]: 6.89
[2.44] (ig/dL, n = 19), whereas group 2 had blood Pb >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, and children in group 2 had an
increased odds of anemia (OR: 2.87 [95% CI: 1.60, 2.87]) compared to group 1 after
adjustment for age, sex, and area of residence. Similarly, in a study of 340 children (aged
1-5 years) from Karachi, Pakistan, mildly and severely anemic children (mean [SD] Hb
levels: 8.9 [0.9] and 7.4 [0.5] g/dL, respectively) had 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 nonanemic children were 14.9 [0.81], 21.4 [2.7],
and 7.9 [1.7] (ig/dL, respectively (p < 0.01) (Shah et al.. 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 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 (mild anemia), and -0.685 (severe anemia). In iron-deficient anemic
children (n = 23) from Denizli, Turkey, mean (SD) serum Pb levels were statistically
increased compared to healthy children (n = 179): 0.013 (0.004) versus 0.008
(0.001) (ig/dL, respectively (Turgut et al.. 2007). The iron-deficient children were
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observed to have decreased Hb, MCV, and ferritin compared to controls, but increased
RBC and RDW. 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).
Contrary to the observations in the above studies, 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 j^ig/dL). No difference was detected in the
mean Hb levels of the children in the two towns (127.55 g/L in children from the waste
recycling town versus 123.46 g/L in children from the town with no recycling).
Equivocal findings in studies investigating associations of blood Pb levels with
hematological effects in children may be due to the comparatively shorter time period
and magnitude of exposure versus those seen in occupational studies.
A number of animal toxicology studies support associations of blood Pb levels with
hematological parameters observed in epidemiologic studies. Baranowska-Bosiacka et al.
(2009) examined the effects of Pb on RBC hemolysis both in vitro measuring lysate in
human RBCs incubated with Pb at concentrations ranging from 0.1-100 (.iM for 5-
30 minutes, and in vivo using a rat RBC lysate from rats exposed to Pb-acetate (0.1 %) in
drinking water for 9 months. Rats exposed to Pb in the in vivo portion of the study
achieved a blood Pb concentration of 7.1 (ig/dL. The concentration of Hb in the plasma
of chronically-exposed rats, a marker of RBC hemolysis, was statistically significantly (p
= 0.01) increased compared to that in control rats. The in vitro studies demonstrated a
similar concentration-dependent increase in the amount of hemolysis, with a significant
(fourfold) increase 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 5.7.5.1). Lee et al. (2005) observed that rats orally
administered Pb (25 mg/kg) via gavage once a week for 4 weeks had an average plasma
Pb level of 6.5 (ig/dL (9.6-fold higher than that in controls, p < 0.05), and had significant
decreases in Hct, Hb, and RBCs (p < 0.05). Male mice orally administered 50 mg/kg Pb
nitrate in distilled water via gavage for 40 days had mean [SD] final blood Pb levels of
1.72 [0.02] (ig/dL versus 0.09 [0.011] |_ig/dL in control mice. Pb-exposed mice had
significantly reduced total RBC counts, total leukocyte counts, Hb, lymphocytes, and
monocytes compared to controls (p < 0.001) (Sharma et al.. 2010b). Rats exposed to 2
g/L Pb-acetate in drinking water for 30 days (blood Pb not reported) had significantly
decreased RBCs, Hb, PCV, MCH, and MCHC compared to controls (p < 0.05) (Simsek
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et al.. 2009). and not a disruption of hematopoiesis. Mice exposed to 1 g/L Pb-acetate in
drinking water (blood Pb not reported) 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 50 mg/kg
Pb-acetate subcutaneously (blood Pb not reported) had significantly reduced Hb, MCV,
MCH, and MCHC compared to controls injected with 5% dextrose (Wang et al). In
weanling rats (age = 25 days, n = 10) whose dams were exposed to 2.84 mg/mL
Pb-acetate (approximating mean [SD] daily exposures of 342.57 [28.11] and 744.47
[29.27] mg/kg during gestation and lactation, respectively), blood Pb was significantly
elevated compared to controls (mean [SE]: 698.1 [78.2] versus 5.4 [0.8] ng/g). The only
hematological parameter affected by Pb exposure was Hct, which was decreased in
exposed rats (mean [SE]: 27.3 [0.5] versus 33.4 [0.3] percent) (Molina et al.. 2011).
Some toxicological studies found no evidence of hematological effects in animals
following exposure to Pb. Male rats administered Pb-acetate in the drinking water for 4
weeks at concentrations ranging from 100-1,000 ppm had a concentration-dependent
increase in blood Pb (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 0.5% 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 |_iM.
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
postulate 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 (CI /HC03) exchanger and is the most abundant protein in RBC
membranes. AE1 is integral in carbon dioxide (C02) 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.
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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 (Toller et al.. 2008;
Lang et al. 2008). As previously reported in the 2006 Pb AQCD, Kempe et al. ("2005)
found that exposing human RBCs to Pb at concentrations ranging from 0.3-3 (.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.
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),
observed after incubation with 5 (.iM Pb for four hours. The expression of PS in RBCs is
considered to be regulated through a Ca2+ dependent mechanism and, correspondingly,
|Ca2 | was observed to increase with exposure to Pb (mean [SE]: 0.24 [0.21] |_iM in
controls to 6.88 [1.13] (.iM in RBCs treated with 5 (.iM Pb for one hour). Consistent with
this finding, Shin et al. (2007) also observed that scramblase activity, which is important
for induction of PS exposure and is activated by |Ca2 [, was increased in Pb-exposed
RBCs. Flippase, which translates PS exposure to inner membranes, is inhibited by high
levels of |Ca2 | 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 i.p. to 25, 50, or 100 mg/kg Pb-acetate (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 and in vivo experiments. In a follow-up study, the same lab
observed that exposure of human RBCs to low concentrations of Pb-acetate (0.1-0.5 (.iM)
induced PS expression. Most notably, exposure to 0.1 (.iM Pb for 24 hours increased PS
expression on RBC membranes by approximately 20% (Jang etal.. 2011).
Accompanying the expression of PS were abnormal, echinocytic RBCs following
incubation with Pb. Unlike the above study, incubation with low concentrations of Pb
(0.1 (.iM) induced the generation of microvesicles, which also expressed PS on their
membranes. Flippase was inhibited by 0.1 (.iM Pb following incubation for one hour, but
scramblase activity was not changed at any Pb exposure concentration. The intracellular
concentration of ATP was decreased at Pb concentrations 0.25 (.iM and greater, but
|Ca2 | did not increase following exposure. This decrease in ATP levels, but lack of
affect on |Ca2 | may explain why flippase activity, but not scramblase, was altered
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following Pb incubation. At 0.5 (.iM. Pb-exposed RBCs with externalized PS were
observed to be targeted and engulfed by differentiated macrophages. Similar in vitro
effects were observed in rat erythrocytes, although higher concentrations were generally
required. PS expression on rat erythrocytes was observed ex vivo in rats exposed to 10 or
50 mg/kg Pb. To corroborate these in vitro and ex vivo findings, rats were exposed to 0,
50, 250, or 1,000 ppm Pb-acetate in drinking water for 4 weeks. At 100 ppm, Hb and Hct
were significantly decreased relative to control, and liver and spleen weights were
increased. At doses greater than 50 ppm, iron accumulation was observed in the spleen, a
clear sign of increased RBC clearance via phagocytosis.
Ciubar et al. (2007) also found that exposure to Pb nitrate (0.5-2 (.iM) resulted in an
increase in PS exposure to RBCs and cell shrinkage, which 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.
5.7.2.3 Red Blood Cell Hematopoiesis
Erythropoietin is a glycoprotein hormone excreted by the kidney to promote the
development of RBCs in the bone marrow. Sakata et al. (2007) examined the relationship
between blood Pb level and serum erythropoietin levels in Pb-exposed nonanemic
tricycle taxi drivers (n=27) working in Kathmandu, Nepal (mean [SD] age: 5.6 [2.6]
years). The average blood Pb level in the taxi drivers was 6.4 (ig/dL compared to
2.4 (ig/dL in nondrivers. Drivers had a significantly lower mean level of serum
erythropoietin (12.7 versus 18.8 mU/mL) compared to the nondrivers and there was an
inverse relationship between the level of serum erythropoietin and blood Pb (r = -0.68, p
< 0.001). Blood Pb level was not associated with any other hematological effects. The
Sakata et al. (2007) study demonstrated that serum erythropoietin levels are affected by
Pb even at levels low enough not to cause anemia. While decreased erythropoietin is
generally considered a measure of kidney toxicity, it can also indicate that Pb could
possibly affect the level of RBCs through decreasing levels of serum erythropoietin.
Celik et al. (2005) observed that exposure of female rats to 140, 250, or 500 mg/kg
Pb-acetate 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
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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 1 g/L
Pb-acetate 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). Cyto- and
geno-toxicity in RBC precursor cells are strong indications of altered hematopoiesis in
bone marrow.
5.7.2.4 Membrane Proteins
There have been few studies examining the effects of Pb on membrane proteins since the
2006 Pb AQCD. According to the 2006 Pb AQCD, Pb has been found to affect RBC
membrane polypeptides in exposed workers (Apostoli et al.. 1988; Fukumoto et al.
1983V In Pb-exposed workers, Fukumoto et al. (1983) found decreased levels of
polypeptides in band 3, which Apostoli et al. (1988) suggested may represent an anion
channel protein, and increases in the level of polypeptides in bands 2, 4, 6, and 7.
Fukumoto et al. (1983) suggested that the changes in the RBC membrane polypeptides
may cause changes in membrane permeability. Apostoli et al. (1988) found that the
changes in membrane polypeptides in association with blood Pb levels greater than
50 (ig/dL. 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.
Huel et al. (2008) found that newborn hair and cord blood Pb levels (mean [SD]: 1.22
[1.41] jj.g/g and 3.54 [1.72] (ig/dL) were negatively associated with Ca-ATPase activity
in plasma membranes of RBCs isolated from cord blood; newborn hair Pb levels were
more strongly associated with cord Ca pump activity than were cord blood Pb (p <
0.0001 versus p < 0.05). Maternal blood Pb levels were not correlated with Ca pump
activity in maternal or cord blood. Pb-induced disruptions in Ca homeostasis in RBCs can
lead to cytotoxicity and necrosis, and these effects may be representative of cellular
dysfunction in other organ systems.
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5.7.3
Red Blood Cell Heme Metabolism
Pb has been found to inhibit several enzymes involved in heme synthesis, namely ALAD
(cytoplasmic enzyme catalyzing the second, rate-limiting, step of the heme biosynthesis
pathway), coporphyrinogen oxidase (catalyses the sixth step in heme biosynthesis
converting coporphyrinogen III into protoporphyrinogen IX), and ferrochelatase
(catalyses the terminal step in heme synthesis converting protoporphyrin IX into heme)
(Figure 5-45). The observations of decreased Hb (measured as total Hb, MCH, or
MCHC) in occupationally-exposed adults (Ukaeiiofo 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; Simsek et al.. 2009; Marques et al.. 2006; Lee et al.. 2005) and associations
with blood Pb levels in children (Riddell et al.. 2007). is a direct indicator of decreased
heme synthesis due to Pb exposure.
Mitochondria
Heme
Succinyl CoA
Glycine
¦T erythrocyte Zn
protoporphryiri
Ferrochelatase
ALA synthatase
Protoporphyrin IX
6-aminolevulinic acid (ALA)
Protoporphyrinogen IX
oxidase
Protoporphyrinogen IX
Coproporphyrinogen III
oxidase
Coproporphyrinogen III
Cytosol
2x6-aminolevulinic acid
T urinary 6-ALA
ALA dehydratase
(porphobilinogen
synthase
Porphobilinogen (PBG)
PBG deaminase
hydroxymethylbilane I
Uroporphyrin ogen
HI synthetase
Uroporphyrinogen III
Uroporphyrinogen
decarboxylase
Coproporphyrinogen III
Note: Steps in the pathway potentially affected by Pb are indicated with curved arrows pointing to the affected enzyme, and effects
are represented by f and [ arrows.
Figure 5-45 Schematic representation of the enzymatic steps involved in the
heme synthetic pathway.
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5.7.3.1 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 has
been used 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 magnitude, timing, frequency, and duration of Pb
exposure necessary to contribute to the observed effects.
As was seen with epidemiologic studies investigating Pb-induced deficits in
hematological parameters, most occupationally-exposed cohorts investigated ALAD
levels in association with high blood Pb levels, i.e., > 27 (ig/dL (range: 27.0-74.4 (ig/dL).
Although effects observed within these cohorts regarding ALAD levels may not be
generalizable to the population as a whole, they are useful in demonstrating consistent
and negative effects of Pb on the activity of this enzyme (Quintanar-Escorza et al. 2007;
Patil et al.. 2006a; Patil et al.. 2006b; Ademuviwa et al.. 2005b'). Occupationally-exposed
adults had levels of inhibition of ALAD that were as great as 90% relative to control
(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 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; 133.9 [14.5] months duration of exposure), the ALAD reactivation index was
increased over that in controls; ALAD activity did not differ between groups (Conterato
et al.. In Press). However, ALAD activity was negatively correlated with blood Pb (r = -
0.59, p < 0.05) but not blood Cd, 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, 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 and BMDL values of 2.7
and 2.3 (ig/dL, respectively, were substantially lower than the BMDs (28.7-44.2 j^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.
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Wang et al. (2010f) found that there was also a 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.
(2010f) observed that the relationship between blood Pb and ALAD activity was
nonlinear and exponential, with larger decreases in ALAD activity occurring with blood
Pb levels greater than 10 (ig/dL. No correlation was observed between urinary ALA
levels and blood Pb. Ahamed et al. (2006) studied male urban adolescents in India. The
39 adolescents were separated into 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 group
1 and 2 were similar in age (mean [SD]: 16.59 [0.91] versus 16.76 [0.90] years,
respectively), height, weight, and body mass index, group 2 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 Indian children (aged 4-12 and 1-7
years) with elevated blood Pb (mean [SD]: 11.39 [1.39] and 21.86 [7.58] (ig/dL)
compared to children with lower blood Pb levels (mean [SD]: 3.93 [0.61] and 6.89
[2.44] (.ig/dL) (Ahamed et al.. 2007; Ahamed et al.. 2005). Decreases were also observed
in children 3-6 years of age with >10 j^ig/dL, compared to children <10 (ig/dL (mean
blood Pb concentration for groups not reported) in northeastern China (Jin et al.. 2006V
Decreased ALAD activity in response to Pb exposure is also observed in toxicological
studies. Rats administered 500 ppm Pb-acetate in drinking water for 15 or 30 days had
decreased blood ALAD activity that was related to duration of exposure and blood Pb
(Rendon-Ramirez et al. 2007). 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 0.5% Pb-acetate via drinking
water for three weeks statistically significantly decreased ALAD activity 72% compared
to controls (mean [SD]: 7.35 [0.35] versus 26.14 [2.19] nM/min/mL RBCs) (Gautam and
Flora. 2010).
5.7.4 Other Heme Metabolism Enzymes
The 2006 Pb AQCD indicated that Pb affects RBC PBG synthase (Simons. 1995; Farant
andWigfield. 1990. 1987). PBG deaminase (Tomokuni and Ichiba. 1990). and TF
endocytosis and iron transport across membranes (Qian and Morgan. 1990). all of which
are directly or indirectly involved in heme synthesis. Although there are no new studies
that examine the effect Pb has on the activities of other heme metabolism enzymes, a
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number of studies investigated associations of blood Pb with concentrations of various
intermediate products in the heme biosynthetic pathway.
Pb intoxication is known to inhibit the function of ferrochelatase, the enzyme that
catalyzes the last step in the heme biosynthetic pathway. Under normal conditions,
ferrochelatase incorporates ferrous iron (Fe2+) into protoporphyrin IX, converting it into a
heme molecule (Figure 5-45). 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 is statistically significantly associated
with increased RBC ZPP levels in adults occupationally exposed to high levels of Pb
(blood Pb level > 27-54 (.ig/dL) (Patil et al. 2006b; Ademuviwa et al.. 2005b'). workers
exposed to moderate levels of Pb (blood Pb level = 21.92 j^ig/dL) (Mohammad et al..
2008). children aged 1-21 years (blood Pb 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 (Rendon-Ramirez et al.. 2007V Interestingly, Wang et al. (201 Of) 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 j^ig/dL and were only positively
correlated with blood Pb at higher blood Pb concentrations (i.e., > 10 (ig/dL). The authors
suggest 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 [units not specified]; 95% CI: -0.22, 0.03, p =
0.13).
5.7.5 Other Hematological Parameters
5.7.5.1 Energy Metabolism
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 via either an adenine pathway, which requires adenine
phosphoribosyltransferase (APRT), or an adenosine pathway, which requires adenosine
kinase. The 2006 Pb AQCD 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, 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
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phosphoribosyltransferase (HPRT) due to in vitro and in vivo exposures. For the in vitro
exposure, 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. In vivo tests measured
APRT and HPRT in rat RBC lysate from rats exposed to Pb-acetate (0.1 %) 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 were significantly decreased in vitro after only
5 minutes of exposure to the 0.1 (.iM concentration, but the decrease was also
concentration-dependent. However, the study authors considered the inhibition moderate
(30-35%) even with the highest 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
5.7.5.2 Other Enzymes
The 2006 Pb AQCD reported that K+ permeability was increased by Pb 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 new studies were found
that examined the effects of Pb exposure on other enzymes. Ekinci et al. (2007) tested the
effects of Pb on two carbonic anhydrase isozymes (I and II) isolated from human RBCs.
Carbonic anhydrases are metalloproteins that use Zn to catalyze the equilibrium between
carbon dioxide 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 ^M) 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 manganese for normal activity. Pb was
observed to bind directly to the active site of the enzyme in a different position than the
manganese, thus possibly resulting in improper protein folding and inhibition of activity.
5.7.6 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
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an association between the level of blood Pb and lipid peroxidation, antioxidant levels, or
indicators of ROS production.
5.7.6.1 Oxidative Stress, Lipid Peroxidation, and Antioxidant
Enzymes
Malondialdehyde (MDA) is an end product of lipid peroxidation and is commonly used
as an indicator of lipid peroxidation. Numerous occupational studies demonstrated
increased lipid peroxidation in highly-exposed worker populations (blood Pb levels
ranging from 29 to 74.4 (.ig/dL) (Kasoerczvk et al.. 2009; Khan et al. 2008; Ouintanar-
Escorza et al.. 2007; Patil et al. 2006a; Patil et al.. 2006b'). There was a correlation
between MDA levels and blood Pb 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. (Ouintanar-Escorza et al.. 2007). Increases in C-reactive protein and
decreases in RBC SOD, catalase, and plasma ceruloplasm were also observed in these
workers, further indicating increased RBC oxidative stress due to higher Pb exposure.
Evidence of lipid peroxidation was also observed in occupational cohorts moderately
exposed to Pb. These studies were cross-sectional in design; thus, there is uncertainty
regarding the magnitude, timing, frequency, and duration of Pb exposure that contributed
to the observed associations. In auto repair apprentices in Turkey (mean [SD]: 16.8 [1.2]
years of age, 3.8 [1.8] years duration of exposure) with blood Pb levels as low as
7.9 (ig/dL (Ergurhan-Ilhan et al.. 2008). increases in glutathione peroxidase (GPx) and
MDA, as well as decreases in a-tocopherol and [3-carotene were observed compared with
controls (compared to 2.6 (ig/dL in controls; mean [SD] age: 16.3 [1] years). Decreases
were observed in SOD and CAT, but the results did not attain statistical significance. In
painters in India (mean [SD] duration of exposure: 126.08 [49.53 months]) with a mean
blood Pb level of 21.92 (ig/dL (compared to 3.06 (ig/dL in controls), there was a
significant decrease in SOD, glutathione (GSH), and CAT accompanied by a significant
increase in oxidized GSH (i.e., GSSG) and thiobarbituric acid reactive species (TBARS,
expressed in terms of MDA) measured in plasma and RBC lysate (Mohammad et al.
2008). In automotive painters in Brazil (mean [SE] blood Pb: 5.4 [0.4] (ig/dL),
glutathionc-V-transfcrasc, GPx, and SOD were positively correlated with blood Pb (r =
0.34, 0.38, and 0.32, respectively; p < 0.05) (Conterato et al. In Press). Similar effects on
indices of oxidative stress were observed in in vitro studies: increased MDA and
decreased SOD and catalase 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
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in RBCs from healthy volunteers with no Pb exposure incubated with 0.4 (.iM Pb for 24-
120 hours (Ouintanar-Escorza et al.. 2010).
Evidence of lipid peroxidation was also observed in children moderately exposed to Pb;
as with the occupational studies, their cross-sectional design limits their utility in
assessing the magnitude, timing, frequency, and duration of Pb exposure necessary to
elicit the observed effects. Ahamed et al. ("2005) investigated the relationship between
blood Pb levels and antioxidant enzyme levels and lipid peroxidation in children aged 4-
12 years in Lucknow, India. A total of 62 children, with a mean [SD] blood Pb level of
7.47 [3.06] (ig/dL, were included in the study. Children were separated into three groups
based on their blood Pb levels. The means (SDs) were: 3.93 (0.61) (ig/dL group I; 7.11
(1.25) (ig/dL group II; and 11.39 (1.39) (ig/dL group III. Lipid peroxidation, measured as
blood MDA, was statistically significantly greater in group III, compared to group II and
I, whereas GSH was lower in group III relative to groups II and I. Catalase activity was
the only measure of oxidative stress that was statistically significantly elevated in group
II compared to group I. Additionally, blood Pb levels were found to be statistically
significantly positively correlated with MDA and CAT and negatively correlated with
GSH. Ahamed et al. (2006) additionally studied male urban adolescents in India. The 39
adolescents were separated into groups according to their blood Pb level (group 1:
<10 (ig/dL [mean 7.4 j^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 body mass index, group 2 had significantly higher
levels of CAT and MDA compared to group 1. There were no significant differences in
blood GSH levels. 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. In a similar
study, Ahamed et al. (2008) examined oxidative stress in Indian children (aged 3-12
years) with neurological disorders. There was a significantly higher mean blood Pb level
in the study population compared to the control healthy population (18.60 versus
10.37 (ig/dL). In addition, the following indicators of oxidative stress were observed in
the study population: 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 the 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).
Administration of Pb (25 mg/kg) to rats once a week for 4 weeks, which was related to a
blood Pb level of about 6.5 (ig/dL, caused a significant increase in RBC MDA levels (Lee
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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 750 mg/kg Pb-acetate in drinking water for 11 weeks
resulted in decreased concentrations of plasma vitamin C, vitamin E, nonprotein thiol,
and RBC-reduced glutathione, with simultaneous increased activity of SOD and GPx
(kharoubi et al. 2008b'). CAT activity was also slightly elevated in Pb-exposed rats, but
the increase failed to reach statistical significance. Exposure of male rats to 0.5% 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 15 mg/kg Pb i.p. for seven days, but not rats treated with 5 mg/kg Pb
(Berrahal et al.. 2007). GPx activity and MDA concentrations were slightly elevated in
the exposed groups, but differences with the control group failed to reach statistical
significance.
5.7.6.2 Antioxidant Defense
In addition to the studies listed above that examine lipid peroxidation and oxidative
stress, there have been studies that indicate that the use of antioxidants and free radical
reactions is protective against Pb-induced RBC oxidative stress. Rats treated with
500 ppm Pb-acetate in drinking water for 15 or 30 days had an increase in free RBC
protoporphyrin and TBARS that was related to length of exposure and blood Pb
(Rendon-Ramirez et al. 2007). Vitamin E administration after exposure to Pb
significantly reduced the TBARS 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 and 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 desferoxamine (iron 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
(Cavlak et al. 2008V 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 statistically significantly different from that in
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controls, but statistically smaller than concentrations observed with Pb alone. Exposure to
L-methionine and N-acetylcysteine also reduced Pb-induced depletion of vitamin E.
5.7.7 Summary and Causal Determination
There is consistent toxicological and epidemiologic evidence indicating that exposure to
Pb affects hematological endpoints, including decreased RBC survival and function,
altered heme synthesis, and increased RBC oxidative stress. Pb has been shown to
preferentially partition into RBCs following exposure, with RBC concentrations
approximately 100-fold greater than those observed in the plasma (Jin et al.. 2008;
Timchalk et al.. 2006).
Elaborating on the body of evidence presented in the 2006 Pb AQCD, recent
epidemiologic and toxicological studies continue to demonstrate that Pb exposure is
associated with decreased RBC survival and function, with the largest body of evidence
consisting of populations of adults and children in which Pb is associated with effects on
several inter-connected and related hematological parameters such as Hb, PCV, MCV,
MCH, and MCHC. For adult populations, the largest body of evidence consists of
occupationally-exposed workers in which measures of RBC survival are altered when
compared with unexposed control populations (Cabaravdic et al. 2010; Ukaeiiofo et al..
2009; Khan et al.. 2008; Patil et al.. 2006a; Patil et al.. 2006b; karita et al.. 2005;
Conterato et al.. In Press). Although the mean blood Pb level in most occupationally
exposed populations was in excess of 20 j^ig/dL, decreases in Hb and PCV were observed
in adults with blood Pb levels of 7 (ig/dL (compared to controls with blood Pb levels of
3 (ig/dL), and significant correlations were observed between RBC distribution width and
MCHC and blood Pb levels in adult populations with mean blood Pb levels of 5.4 (ig/dL.
Only one non-occupational study was found investigating the association of Pb with
hematological parameters; in pregnant women, blood Pb levels were found to be
negatively correlated with Hb concentrations. Studies in children measuring concurrent
blood Pb levels are generally in agreement with those investigating occupationally-
exposed adults regarding effects on hematological parameters (i.e., Hb, MCV, (Oueirolo
et al.. 2010; Shah et al.. 2010; Ahamed et al.. 2007; Huo et al. 2007; Olivero-Verbel et
al.. 2007; Riddel 1 et al. 2007; Turgut et al. 2007; Ahamed et al.. 2006; Rondo et al.
2006). Any differences in the effects on specific hematological parameters observed
between study findings for adults and those for children may be due to the comparatively
shorter duration and lower magnitude of Pb exposure experienced by children compared
to adults, although there is uncertainty regarding the timing and duration of exposure
needed to induce effects in adults. In addition, Pb was shown to reduce Ca2+- and Ca2+-
Mg2+-ATPase activity in RBC membranes, which leads to an increase in RBC |Ca2 |,.
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increased membrane fragility, and abnormal morphological changes in studies of
occupationally exposed adults (Ouintanar-Escorza et al.. 2007) and in in vitro studies
(Ouintanar-Escorza et al.. 2010; Ciubaretal.. 2007). Heul et al. (2008) observed a
reduction of RBC Ca2+-Mg2+-ATPase activity in children in association with a concurrent
group mean cord blood Pb level of 3.54 (ig/dL. Toxicological studies have also observed
decreases in Hct and Hb and increases in hemolysis and reticulocyte density in rats and
mice with blood Pb levels as low as 6.6-7.1 (.ig/dL (Sharma et al.. 2010b; Baranowska-
Bosiacka et al.. 2009; Simsek et al. 2009; Lee et al. 2005). 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 et
al.. 2007; Shin et al.. 2007). Suggestive evidence of disrupted hematopoiesis evidenced
by decreased serum erythropoietin was observed in occupationally-exposed adults with a
mean blood Pb level of 6.4 j^ig/dL (Sakata et al.. 2007); toxicological studies in rats also
indicate that Pb is cytotoxic to RBC progenitor cells (Alghazal et al.. 2008b; Celik et al..
2005). Taken together, these studies provide consistent evidence that exposure to Pb
affects RBC function and survival, and leads to the reduction of RBCs in circulation.
Although this decrease in RBCs may be explained by both decreased cell survival and/or
disruption of hematopoiesis, the observation of increased reticulocytes seems to represent
compensation for decreased RBC survival due to Pb exposure.
Confirming effects observed in the 2006 Pb AQCD, a large body of evidence consisting
of cross-sectional epidemiologic studies measuring concurrent blood Pb in adults and
children have found that decreases in RBC ALAD levels and activity are strongly
associated with higher blood Pb levels (Wang et al. 2010f; Mohammad et al.. 2008;
Ahamed et al.. 2007; Ouintanar-Escorza et al.. 2007; Ahamed et al.. 2006; Patil et al..
2006a; Patil et al.. 2006b; Ademuviwa et al.. 2005b; Ahamed et al.. 2005; Conterato et
al.. In Press). Although the body of evidence is smaller than for humans, decreases in
blood ALAD activity were also seen in rats with increased blood Pb levels, compared to
controls (Lee et al.. 2005). In addition to ALAD, recent studies have shown that Pb
exposure inhibits the activity of ferrochelatase, leading to increased RBC ZPP in children
and occupationally-exposed adults (Counter et al. 2009. 2008; Mohammad et al.. 2008;
Counter et al.. 2007; Patil et al.. 2006b; Ademuviwa et al. 2005b) and animals (Rendon-
Ramirez et al.. 2007). Pb has also been shown to inhibit the in vitro activities of other
enzymes in RBCs, including those involved in nucleotide scavenging, energy
metabolism, and acid-base homeostasis (Baranowska-Bosiacka et al.. 2009; Ekinci et al..
2007).
Lastly, Pb exposure induces lipid peroxidation and oxidative stress in RBCs.
Epidemiologic studies have observed increases in MDA in occupationally-exposed adult
populations (Ergurhan-Ilhan et al.. 2008; Khan et al.. 2008; Mohammad et al.. 2008;
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Quintanar-Escorza et al. 2007; Patil et al.. 2006a; Patil et al. 2006b). Other changes in
oxidative stress parameters observed included lowered activities of SOD, GR, and CAT,
and increased CRP. Altered indices of RBC oxidative stress were also seen in adolescents
and children in association with blood Pb levels (Ahamed et al.. 2008; Ahamed et al..
2006; Jin et al.. 2006). In vitro and vivo studies have also demonstrated that prior,
concurrent, or subsequent treatment with various antioxidants has been shown to
ameliorate at least partially Pb-induced oxidative stress in RBCs (Cavlak et al.. 2008;
Casadoa et al.. 2007; Rendon-Ramirez et al.. 2007).
Similar to the epidemiologic and 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 aquatic and
terrestrial invertebrates and vertebrates (Sections 7.4.1.2 and 7.4.2.2). 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 on serum enzyme levels and white blood cell counts in birds and mammals.
In summary, new epidemiologic and toxicological studies included in the current review
provide strong evidence that exposure to Pb is associated with numerous deleterious
effects on the hematological system, including effects on RBC survival and function,
altered heme synthesis, and increased oxidative stress, and continue to confirm previous
conclusions from the 2006 Pb AQCD. The principal finding regarding RBC survival and
function are consistent Pb-induced alterations in several inter-connected and related
hematological parameters such as Hb, Hct, and MCV across multiple studies, with the
weight of evidence provided by epidemiological studies in occupationally-exposed adult
populations and children. In occupationally-exposed adults, these findings are most
substantiated in populations with current blood Pb levels > 20 (ig/dL, although effects on
hematological parameters were observed in some occupationally-exposed populations at
concurrent blood Pb levels in the range of 5-7 (ig/dL. In Pb-exposed children, effects on
hematological parameters were most substantiated in populations with blood Pb levels
less than 15 (ig/dL. The weight of evidence in adult rodents exposed long-term to Pb (i.e,
> 4 weeks), although less than the weight of evidence in humans, is coherent with
epidemiologic studies regarding decrements in hematological parameters at blood Pb
levels as low as 6.6-7.1 (ig/dL. Regarding alterations in heme synthesis, the largest body
of evidence again is provided by decreased ALAD activity observed in epidemiologic
studies in occupationally-exposed adult populations and children. In the occupationally-
exposed adult populations, the observation of decreased ALAD activity was most often
observed in populations with concurrent blood Pb levels >15 (ig/dL. In children,
decreases in ALAD activity were observed in populations with concurrent blood Pb
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levels in the range of 7-22 (ig/dL. Animal toxicological studies also provide to the weight
of evidence regarding altered ALAD activity, with effects seen in adult animals exposed
for 3-4 weeks with blood Pb levels as low as 6.5 (ig/dL. As for the effects listed above,
the weight of evidence for oxidative stress (i.e., increased lipid peroxidation or alterations
in antioxidant enzyme levels) primarily comes from epidemiological studies in
occupationally-exposed adults and children. The majority of evidence for increased
oxidative stress in Pb-exposed adults comes from occupational cohorts with concurrent
blood Pb levels >15 (ig/dL. In children, concurrent blood Pb levels of 7-22 j^ig/dL were
associated with measures of oxidative stress. Due to the cross-sectional nature of the
above epidemiologic studies in adults and children, and the measurement of concurrent
blood Pb, the timing and duration of exposure necessary to alter RBC survival and
function, heme synthesis, or the state of oxidative stress in RBCs is unclear. This
uncertainty is greatest in adults as concurrent blood Pb levels reflecting recent exposures
is likely to be less than blood Pb levels resulting from past exposures. The consistency of
findings in epidemiologic studies investigating effects in adults and children, and the
coherence of findings in the toxicological literature and coherence across the disciplines
is sufficient to conclude that a causal relationship exists between Pb exposures and
effects on heme synthesis and red blood cell function.
5.8 Reproductive and Developmental Effects
The effect of Pb on reproductive outcomes has been of interest for years, starting in
cohorts of occupationally-exposed individuals. More recently, researchers have begun to
focus on reproductive effects in populations without occupational exposures, with
environmentally-relevant levels of Pb exposure. In the toxicological and epidemiologic
literature, research on reproductive effects of Pb include female and male reproductive
function (hormone levels, fertility, puberty, and effects on reproductive organs and
estrus), birth defects, spontaneous abortions, infant mortality, preterm birth, low birth
weight/fetal growth, and other developmental effects. In epidemiologic studies, various
biological measures of Pb are used including Pb measured in blood and bone;
toxicological studies only report exposure using blood Pb. Bone Pb is indicative of
cumulative Pb exposure. Blood Pb can represent more recent exposure, although it can
also represent remobilized Pb occurring during times of bone remodeling and pregnancy
or lactation. More detailed discussion of these measures and Pb transfer via umbilical
cord blood Pb across the placenta, and via lactation is given in Section 4.3.5.2 on Pb
Toxicokinetics. A few studies of pregnancy-induced hypertension and eclampsia have
been conducted and are reported on in the section on hypertension (Section 5.4.2.1).
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Briefly, the relatively small number of studies found consistently positive associations
between blood Pb levels and pregnancy-induced hypertension.
Overall, the recent reproductive literature 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 also indicates that Pb exposure is associated with delayed onset of puberty in
both males and females. The new information from epidemiologic and toxicological
studies is integrated with conclusions from previous Pb AQCDs below.
5.8.1 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. For some
effects, there are inconsistent findings, but for others, such as delayed puberty, there are
clear associations with blood Pb levels. In addition, recent toxicological studies add
further knowledge of Pb-related effects on the female reproductive system.
5.8.1.1 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. Epidemiologic studies (characterized in Table
5-30; all studies included in the table used measures of Pb and hormones that were either
concurrent or close in time) support the toxicological findings, which are the major body
of evidence on endocrine effects of Pb.
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Table 5-30 Summary of recent epidemiologic studies of associations between
Pb levels and hormones for females
Reference8
Study,
Location,
and Years
Outcome
Study population
Pb
Biomarker
Mean Pb (SD)
in [jg/dL
Adjusted Effect Estimates
Krieg et al. 12007)
U.S.
1988-1994
FSH, LH
Women aged 35-60
from the NHANES III
study
Blood Pb
2.8
Linear regression slope (95% CI)
transformed Pb
FSH:
Post-menopausal 22.2 (13.5, 30.E
Pregnant 0.1 (-0.1, 0.3)
Menstruating at time of exam 2.1
6.3)
Both ovaries removed 32.6 (10.1,
Birth control pills being used -6.3
2.5)
Pre-menopausal 8.3 (3.8,12.7)
for log-
(-2.1,
55.1)
1-10.0,
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)
Chang et al. (2006)
Kaohsiung City,
Taiwan
1999
2000-2001
Estradiol Women receiving care Blood Pb
at a infertility clinic in
2000-2001 or delivering
a normal infant at a
nearby medical center in
1999
3.12 (0.19) Linear regression |3(SE) for Pb
1.18(0.60)
p-value: 0.049
Pollack etal. (2011)
Buffalo, NY
2005-2007
FSH, estradiol, Healthy, premenopausal Blood Pb
LH,	women aged 18-44 with
progesterone, menstrual cycle length
and cycle of 21-35 days, BMI of
length	18-35 kg/m , not
recently using birth
control, not planning to
become pregnant, and
not breast feeding
0.93	Mean % Estradiol
IQR: 0.68, 1.20 0.30-0.72 pg/dL: Ref
0.73-1.10 pg/dL: 8.2 (-1.2, 18.6)
1.11-6.20 pg/dL: 4.7 (-4.7, 15.2)
Amplitude Estradiol
0.30-0.72 pg/dL: Ref
0.73-1.10 pg/dL:-0.01 (-0.06, 0.04)
1.11-6.20 pg/dL:-0.02 (-0.7, 0.03)
Phase Shift Estradiol
0.30-0.72 pg/dL: Ref
0.73-1.10 pg/dL:-0.09 (-0.24, 0.05)
1.11-6.20 pg/dL: 0.14 (-0.01, 0.29)
Mean % FSH
0.30-0.72 pg/dL: Ref
0.73-1.10 pg/dL: 8.0 (-0.9, 17.7)
1.11-6.20 pg/dL: 3.6 (-5.3, 13.3)
Amplitude FSH
0.30-0.72 pg/dL: Ref
0.73-1.10 pg/dL:-0.01 (-0.03, 0.02)
1.11-6.20 pg/dL:-0.02 (-0.04, 0.01)
Phase Shift FSH
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Study,
Location,	Pb Mean Pb (SD)
Reference3	and Years Outcome Study population Biomarker in [jg/dL	Adjusted Effect Estimates
0.30-0.72 |jg/dL: Ref
0.73-1.10 |jg/dL: -0.06 (-0.25, 0.12)
1.11-6.20 |jg/dL: -0.02 (-0.21,0.18)
Mean % LH
0.30-0.72 |jg/dL:
0.73-1.10 pg/dL:
1.11-6.20 |jg/dL:
Amplitude LH
0.30-0.72 |jg/dL:
0.73-1.10 pg/dL:
1.11-6.20 |jg/dL:
Phase Shift LH
0.30-0.72 |jg/dL:
0.73-1.10 pg/dL:
1.11-6.20 |jg/dL:
Ref
5.1 (-5.1, 16.4)
-0.5 (-10.5, 10.7)
Ref
-0.01 (-0.03, 0.02)
-0.02 (-0.04, 0.01)
Ref
-0.16 (-0.36, 0.03)
-0.11 (-0.32, 0.10)
Mean % Progesterone
0.30-0.72 |jg/dL: Ref
0.73-1.10 pg/dL: 7.5 (0.1,15.4)
1.11-6.20 pg/dL: 6.8 (-0.8, 14.9)
Amplitude Progesterone
0.30-0.72 pg/dL: Ref
0.73-1.10 pg/dL: 0.07 (0.01,0.15)
1.11-6.20 pg/dL:-0.06 (-0.13, 0.01)
Phase Shift Progesterone
0.30-0.72 pg/dL: Ref
0.73-1.10 pg/dL: 0.04 (-0.06, 0.15)
1.11-6.20 pg/dL: 0.15 (0.05, 0.26)
Linear models p (95% CI)
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% CI) for anovulation per I pg/dL
1.20 (0.62, 2.34)
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Study,
Location,	Pb Mean Pb (SD)
Reference3	and Years Outcome Study population Biomarker in [jg/dL	Adjusted Effect Estimates
Adjusted percent change (95% CI) in
serum hormone level for change 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) per unit Pb
<25 d vs. 25-35 d cycle length:
0.9 (0.4, 2.3)
>35 d vs. 25-35 d cycle length:
0.5(0.1,1.9)
"Studies are presented in order of first appearance in the text of this section.
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 follicle stimulating hormone (FSH) and luteinizing hormone (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. 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) blood Pb in this study was 3.12 (0.19) (ig/dL. The
study reported a positive association between blood Pb levels and serum estradiol
Jackson et al. (2011)
Buffalo, NY
2005-2007
FSH, estradiol,
LH,
progesterone,
and cycle
length
Healthy, pre-
menopausal women
aged 18-44 with
menstrual cycle length
of 21-35 days, BMIof
18-35 kg/m not
recently using birth
control, not planning to
become pregnant, and
not breast feeding
Blood Pb
Median: 0.87
IQR: 0.68, 1.20
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concentrations during the early follicular phase, which reflects ovary activity.
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. Neither study detected an
association between unit change in blood Pb and hormone levels. However, when
examining 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. 201IV 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).
The effect of Pb exposure on the female endocrine system was demonstrated in
toxicological studies reviewed in the 1986 and 2006 Pb AQCD (U.S. EPA. 2006b.
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 previous 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
Fisher 344 rat pups as compared to control, non-Pb exposed pups. Dumitrescu et al.
(2008a) observed alteration of hormone levels in female Wistar rats after ingesting
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 for
comparison of 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 at a concentration that did not
affect reproductive performance, implantation or pregnancy outcome (0.05 mg/kg body
weight) to Charles Foster female rats 5 days before mating and during the gestational
period. 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 exposed gestationally and
lactationally to Pb-acetate (subcutaneous daily injections of 0.05 jj.g/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. The results of the study showed
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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 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 exposure (Nampoothiri and Gupta. 2006).
No recent toxicological studies were found that examined Pb-induced effects on the
estrus cycle.
Overall, toxicological studies report alterations in hormone levels related to blood Pb
concentration. Similarly, epidemiologic studies reported associations between blood Pb
levels and hormone levels in female adults. Although Pb-associated changes in hormone
levels are observed, there are discrepancies about the direction of the hormone changes
related to Pb. One explanation is that the direction of change could vary based on current
hormonal and reproductive status.
5.8.1.2 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 support this finding. The epidemiologic studies are summarized in
Table 5-31. Most of these studies examined biological measures of Pb collected at or
during the period of possible fertilization, although Bloom et al. (201 la) measured blood
Pb at baseline and followed women for at least 12 menstrual cycles (or until pregnancy).
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Table 5-31 Summary of recent epidemiologic studies of associations between
Pb levels and fertility for females
Reference8
Study,
Location,
and Years
Outcome
Study population
Pb Biomarker
Mean Pb(SD)
in [jg/dL
Adjusted Effect Estimates
Chang et al. (2006)
Kaohsiung City, Infertility
Taiwan
1999
2000-2001
Women receiving care Blood Pb
at a infertility clinic in
2000-2001 or delivering
a normal infant at a
nearby medical center in
1999
3.12(0.19)	OR (95% CI)
Infertility
<2.5 |jg/dL: 1.00 (Ref)
>2.5 pg/dl_: 2.94 (1.18, 7.34)
Al-Saleh et al. (2008a) Riyadh, Saudi
Arabia
2002-2003
Achieving
pregnancy
and/or
fertilization
Women aged 19-50
undergoing IVF
Blood Pb
Follicular fluid Pb
Blood Pb
3.34 (2.24)
Blood Pb levels >10
|jg/dL: 1.7%
Follicular fluid
0.68(1.82)
OR (95% CI) (unit not given,
assume results are per 1
pg/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, 3.02)
Note: In a reduced adjusted
model for fertilization, the OR
for blood Pb was 0.38 (0.14,
0.99)
Silberstein et al.
Providence, Rl Achieving
NS	pregnancy
Women undergoing IVF
at the study hospital
Follicular fluid Pb Not given
quantitatively
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, study only included 9
women
Bloom et al. (2010)
California
2007-2008
Oocyte
maturity, oocyte
fertilization
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
Blood Pb	0.82 (0.32)	RR per 1 pg/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)*
'Controlling for Cd
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Study,
Location,	Mean Pb (SD)
Reference3	and Years Outcome Study population Pb Biomarker	in [jg/dL	Adjusted Effect Estimates
Bloom et al. (2011b) California
2007-2008
Embryo cell	Women who were part Blood Pb
number,	of the Study of Metals
embryo	and Assisted
fragmentation	Reproductive
score	Technologies (SMART)
0.83 (0.30)
OR per 1 pg/dL*
Embryo cell number:
0.25 (0.07, 0.86)
(women referred to the
Center for Reproductive
Health ofUCSFfor
Embryo fragmentation score:
1.71 (0.45, 6.56)
infertility treatment and
their first IVF procedure)
and who generated
"Adjusted for Hg and Cd
embryos
Bloom et al. (2011a) New York
1996-1997
Achieving Women who were aged Blood Pb
pregnancy 18-34 years, were
previously part of a
study five years prior
about fish consumption,
and were not currently
pregnant
No positive pregnancy	|3 (95% CI)
test: 1.55(0.16)	.0 031 -|.004) per0.6
Positive pregnancy	pg/dL
test: 1.54 (0.12)
"Studies are presented in order of first appearance in the text of this section.
Epidemiologic studies examined women having difficulty conceiving by performing
studies among patients of fertility clinics or undergoing in vitro fertilization (IVF).
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 j^ig/dL to those with
blood Pb levels < 2.5 (ig/dL. 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 (Silbcrstcin et al. 2006)
found that median follicular Pb levels in women who achieved pregnancy were lower
than the follicular Pb levels among nonpregnant women. One limitation present in these
studies is that the participants, especially in the later two studies, 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. The
Study of Metals and Assisted Reproductive Technologies (SMART) enrolled women
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undergoing their first round of IVF and investigated multiple steps before pregnancy as
the outcomes (Bloom et al.. 201 lb; Bloom et al. 2010). Higher blood Pb levels were
associated with lower oocyte maturity although the lack of power made interpretation of
models controlling for Cd difficult. No association was observed between blood Pb and
oocyte fertilization (Bloom et al.. 2010). When examining 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 et al.. 201 lb). The study was limited by the small number of participants.
Overall, these epidemiologic studies examine a variety of fertility-related endpoints and
although some studies demonstrate an association between higher Pb levels and
fertility/pregnancy, as a whole the results are inconsistent across studies.
A prospective cohort study enrolled women who previously participated in a study of fish
consumption for a length of <1 to >6 menstrual cycles and investigated the relationship
between blood Pb levels and having a positive pregnancy test (Bloom et al.. 201 la). No
association was observed between blood Pb and achieving pregnancy.
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. 2006b; Teiion 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. observed a modification in sex ratio of pups born
to dams exposed to Pb before mating and during 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. Karaca and Simsck (2007) observed an increase in the
number of mast cells in ovary tissue after Pb exposure (2,000 (ig/mL in drinking water)
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
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(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.
Recent epidemiologic and toxicological studies on the effect of Pb on fertility outcomes
have generated inconsistent results. However, the bulk of the evidence including the
current and historical Pb literature (U.S. EPA. 2006^ indicate that increased Pb exposure
may decrease fertility.
5.8.1.3 Effects on Puberty
Recent toxicological studies of rodents have examined the effects of Pb on pubertal and
reproductive organ development and on biomarkers of pubertal development. There have
also been recent epidemiologic studies examining associations between blood Pb levels
and onset of puberty, which are summarized in Table 5-32 and in the text below. All of
the epidemiologic studies examined concurrently measured blood Pb and puberty and are
reported below. Additionally, Naicker et al. (2010) followed girls to determine their age
of menarche; however, blood Pb levels were measured once at 13 years of age.
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Table 5-32 Summary of recent epidemiologic studies of associations between
Pb levels and puberty for females
Reference8
Study Location
and Years
Outcome
Study Population
Mean Pb (SD) in
Pb Biomarker	[jg/dL
Adjusted Effect Estimates
Tomoum et al.
(2010)
Cairo, Egypt
2007
Hormones and
pubertal
development
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)
Blood Pb
NS for girls only
(combined with
boys in the study
the mean was 9.46
[3.08])
Breast Development
<10 |jg/dL:
Stage 2: 36.4%
Stage 3: 63.6%
> 10 pg/dL:
Stage 2:100%
Stage 3: 0%
Chi-square p-value<0.01
Pubic Hair Development
<10 pg/dL:
Stage 2: 36.4%
Stage 3: 63.6%
> 10 pg/dL:
Stage 2: 77.8%
Stage 3: 22.2%
Chi-square p-value>0.05
'Quantitative results for
hormones not provided
Denham et al.
Akwesasne Mohawk Age at
10- to 16.9-yr-old girls Blood Pb
0.49 (0.905)
Coefficients for binary logistic
(2005)
Nation (boundaries of menarche
in the Akwesasne
regression predicting menarche

New York, Ontario,
community
Median: 1.2
with Pb centered at the mean:

and Quebec
log blood Pb -1.29 (p-value 0.01)

NS


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 pg/dL




decreased the odds of menarche




attainment by 72%
OR (95% CI)
Delay in breast development at
age 13
<5 pg/dL: 1.00 (Ref)
>5 pg/dL: 2.34 (1.45, 3.79)
Delay in pubic hair development
at age 13
<5 pg/dL: 1.00 (Ref)
>5 pg/dL: 1.81 (1.15, 2.84)
Delay in attainment of menarche
at age 13
<5 pg/dL: 1.00 (Ref)
>5 pg/dL: 2.01 (1.38, 2.94)
Naicker et al.
(2010)
Johannesburg/Soweto,
South Africa
Born in 1990
Self-reported
Tanner staging
at age 13 and
age at
menarche
Girls of black or 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
Blood Pbat 13 yr
of age
4.9(1.9)
blood Pb levels > 10
pg/dL: 1%
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Reference8
Study Location
and Years
Outcome
Study Population
Mean Pb (SD) in
Pb Biomarker	[jg/dL
Adjusted Effect Estimates
DenHondetal. Flanders	Tanner staging, Girls ages 14 and 15, in Blood Pb
(2011)	2003-2004	a9e a' ^e'r ^ year 0^
menarche, secondary education
regular menses and living in the same
study areas for at least
5 years
Median: 1.81
10th percentile: 0.88
90th percentile: 3.81
OR (95% CI) for pubic hair
development with doubling of
exposure
0.65 (0.45, 0.93)
'Association was no longer
statistically significant when PCB
marker included in the model
No association between Pb and
breast development (results not
given)
Wolff et al. New York City, NY
(2008): Wolfet al. -i996-1997
(2007)	laa° 'aa'
Pubertal stages	9-yr old girls from the
defined using	study hospital and
standard	nearby pediatric offices
drawings
Blood Pb	Median: 2.4	PR (95% CI) (unit not given,
assume results are per 1 pg/dL)
Breast stage: 1.01 (0.79,1.30)
Pubic hair stage: 1.25 (0.83,
1.88)
Wu et al. (2003b) U.S.A.
1988-1994
Tanner staging
and age at
menarche
Girls ages 8-16 from
the NHANES III study
Blood Pb
2.5(2.2)
VNfeighted proportion
of the sample with
blood Pb 5.0-21.7:
5.9%
OR (95% CI)
Breast development
0.7-2.0 pg/dL: 1.00 (Ref)
2.1-4.9 pg/dL: 1.51 (0.90, 2.53)
5.0-21.7 pg/dL: 1.20 (0.51,2.85)
Pubic hair development
0.7-2.0 pg/dL: 1.00 (Ref)
2.1-4.9 pg/dL: 0.48 (0.25, 0.92)
5.0-21.7 pg/dL: 0.27 (0.08, 0.93)
Menarche
0.7-2.0 pg/dL: 1.00 (Ref)
2.1-4.9 pg/dL: 0.42 (0.18, 0.97)
5.0-21.7 pg/dL: 0.19 (0.08, 0.43)
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Reference8
Study Location
and Years
Outcome
Study Population
Mean Pb (SD) in
Pb Biomarker	[jg/dL
Adjusted Effect Estimates
Selevan et al.
U.S.A.
1988-1994
Tanner staging
and age at
menarche
Girls ages 8-18 from
the NHANES III study
Blood Pb
Geometric mean
NHWhites: 1.4
NHBIacks: 2.1
Mexican-Americans:
1.7
Blood Pb
levels>5|jg/dL:
NHWhites: 2.7%
NHBIacks: 11.6%
Mexican-Americans:
12.8%
OR (95% CI)
Breast development
NH Whites:
1 |jg/dL: 1.00 (Ref)
3 |jg/dL: 0.82 (0.47, 1.42)
NH Blacks:
1 |jg/dL: 1.00 (Ref)
3 |jg/dL: 0.64 (0.42, 0.97)
Mexican Americans:
1 |jg/dL: 1.00 (Ref)
3 |jg/dL: 0.76 (0.63, 0.91)
Blood Pb levels >10
Mg/dL:
NHWhites: 0.3%
NHBIacks: 1.6%
Mexican-Americans:
2.3%
Pubic hair development
NH Whites:
1 pg/dL: 1.00 (Ref)
3 pg/dL: 0.75 (0.37, 1.51)
NH Blacks:
1 pg/dL: 1.00 (Ref)
3 pg/dL: 0.62 (0.41,0.96)
Mexican Americans:
1 pg/dL: 1.00 (Ref)
3 pg/dL: 0.70 (0.54, 0.91)
HR (95% CI) Included only girls
8-16
Age at menarche
NH Whites:
1 pg/dL: 1.00 (Ref)
3 pg/dL: 0.74 (0.55, 1.002)
NH Blacks:
1 pg/dL: 1.00 (Ref)
3 pg/dL: 0.78 (0.63, 0.98)
Mexican Americans:
1 pg/dL: 1.00 (Ref)
3 pg/dL: 0.90 (0.73, 1.11)
Gollenberg et al.
(2010)
U.S.A.
1988-1994
Luteinizing
hormone (LH)
and inhibin B
Girls ages 6-11 from
the NHANES III study
Blood Pb
Median 2.5 (range
0.07, 29.4)
blood Pb
>10 pg/dL: 5%
OR (95% CI) for exceeding
pubertal inhibin B cutoff
(>35pg/mL)
<1 pg/dL: 1.00 (Ref)
1-4.9 pg/dL: 0.38 (0.12, 1.15)
>5 pg/dL: 0.26 (0.11,0.60)
OR (95% CI) for exceeding
pubertal LH cutoff (>0.4 mlU/mL)
<1 pg/dL: 1.00 (Ref)
1-4.9 pg/dL: 0.98 (0.48, 1.99)
>5 pg/dL: 0.83 (0.37, 1.87)
*a sensitivity analysis including
only those with blood Pb <10
pg/dL had similar results but ORs
were slightly attenuated
'Studies are presented in order of first appearance in the text of this section.
Several epidemiologic studies investigated the association between blood Pb and
indicators of puberty onset. A study among girls aged 10-13 years (median: 12 years)
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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. A study of girls aged 10-16.9 years of age
in the Akwesasne Mohawk Nation reported a nonlinear positive association between
blood Pb and age at menarche (Denham et al.. 2005). No association was observed below
blood Pb of 0.49 j^ig/dL in a nonlinear model of the Pb-menarche relationship. A study
conducted in South Africa reported a positive association between blood Pb levels and
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
blood Pb and pubic hair 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 performed in NYC among 9 year
old girls reported no association between Pb levels and pubertal development (Wolff et
al.. 2008). but this age group may be too young to study when investigating delayed
puberty as the outcome.
Multiple studies have been performed examining blood Pb levels and puberty using
NHANES III data (Gollenberg et al.. 2010; Selevan et al. 2003; Wu et al.. 2003b). A
study that included girls aged 8-16 years and reported an association for delayed
attainment of menarche and pubic hair development, but not for breast development (Wu
et al.. 2003b). The associations were observed even at blood Pb levels of 2.1-4.9 j^ig/dL
compared to girls with blood Pb levels <2.1 (ig/dL. Another NHANES III study included
girls 8-18 years of age and reported the results stratified by race (Selevan et al. 2003).
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. In a
study of girls ages 6-11 years old from NHANES III data, higher blood Pb levels were
associated with lower inhibin B, a protein that inhibits FSH production, but no
association was observed for LH. (Gollenberg et al. 2010). The inverse association
between blood Pb and inhibin B was greater among girls with iron deficiency compared
to those with high Pb but sufficient iron 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."
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Puberty; Neonate/adult; Mouse; Female
(lavieoli. Carelli. Stanek. Castelllno. Li. et al.
2008)
Neurotransmitter; Adult; Mouse; Both
(Leasureetal.. 2008)
Physical development; Adult; Mouse; Male
(Leasureetal.. 2008)
Eye; Adult; Rat; Both
(P. A. Foxetat 2008)
Redox-oxidative stress; Adult; Rat; Male
(Nava-Hernandez et al.. 2009)
Sperm; Adult; Rabbit; Male
(Moorman et al.. 1998)
Neurobehavioral; Adult; Mouse; Male
(Leasure et al.. 2008)
Hematological parameters; Adult; Rat; Both
(Teiion et al.. 2006)
Histology; Adult; Rat; Both
(Teiion et al.. 2006)
BiomarKers; Adult; Rat; Both
(Teiion et al.. 2006)
Physical development; Adult; Rat; Both
(Teiion et al.. 2006)
«
o -1 ghii* i.:r»snjWDr
~ _o e:*;crc i>h
AHighs-f Cone ..itt '.2 sn«e
er'C.rcsrtia'nr
100
Blood Pb Level (yg/dl)
Figure 5-46 Toxicological exposure-response array for reproductive effects
of Pb.
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Table 5-33 Toxicological concentration-response array summary for
reproductive effects of Pb presented in Figure 5-46
Reference
Blood Pb level with Effect
(Hg/dL)
Altered Outcome
Iavicoli 1. et al. (2006b)
8 & 13
Delayed onset female puberty
Leasure et al. (2008)
10 & 42
10, 24 & 42
10 & 42
Neurotransmitter, Dopamine homeostasis
Physical Development, Adult obesity (males)
Aberrant response to amphetamine
Fox et al. (2008)
12
Retinal aberrations
Nava-Hernandez et al. (2009)
19.5
Sperm affected via redox imbalance
Moorman et al. (1998)
25-130
Semen quality affected
Teijon et al. (2006)
40 & 100
40 & 100
40 & 100
100
Hematology
Histology-Offspring renal & hepatic
Biomarker-Offspring renal function
Physical development: birth weight
Fox et al. (2008)
12
Retinal aberrations
Nava-Hernandez et al. (2009)
19.5
Sperm affected via redox imbalance
Moorman et al. (1998)
25-130
Semen quality affected
Teijon et al. (2006)
40 & 100
40 & 100
40 & 100
100
Hematology
Histology-Offspring renal & hepatic
Biomarker-Offspring renal function
Physical development: birth weight
Fox et al. (2008)
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 effects of Pb exposure as
is demonstrated in the exposure response array (Figure 5-46 and Table 5-33 Figure 5-45;
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 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 another recent study, Iavicoli et al. (2006b) reported a statistically significant delay in
several indicators of sexual maturity in offspring (Swiss mice, Fi generation) born to
dams that ingested 3.5-40 ppm Pb in their daily diet; offspring had continuous dietary
exposure until the termination of the experiment. Maternal ingestion of Pb at the various
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doses resulted in female pup blood Pb levels of 3.5-13 (ig/dL. For all diet groups in this
range (3.5-13 (ig/dL), there was a delay in age at vaginal opening, age of first estrus, age
of vaginal plug formation, and age of first parturition when compared to the group at
background Pb concentration (2 (ig/dL). A novel finding in the Iavicoli study was that
very low dose Pb (blood Pb of 0.7 (ig/dL, food concentration of 0.02 ppm continuous
through gestation, lactation and until the termination of the experiment) induced
statistically significant acceleration of markers of sexual maturation in female offspring
versus background Pb level animals (blood Pb of 2 (ig/dL). There were statistically
significant increases in time of vaginal opening (30% earlier), first estrous, first vaginal
plug formation, and first parturition at the very low Pb exposure versus 2 (ig/dL animals.
Thus, the timing of puberty is delayed in a concentration-dependent fashion with very
low dose Pb having a statistically significant earlier onset of puberty than the background
Pb animals (2 (ig/dL). Also, the animals exposed to the higher dose of Pb (blood Pb up to
13 (ig/dL) had statistically significant delays in onset of puberty when compared to the
other dose groups.
In addition, Pb-induced shifts in sexual maturity were observed in the subsequent
generation (F2 generation) across that dose range. These F2 animals continued to be
exposed to same concentrations of Pb over multiple generations through the diet. Results
in the F2 generation closely resembled those of the Fi generation, as both generations
received Pb exposure. The authors concluded that a modest elevation in blood Pb level
(13 (ig/dL) over background (2-3 (ig/dL) can result in a profound delay in the onset of
puberty (15-20%). In the F2 generation, reduction in blood Pb (0.7 (ig/dL) below
background (2-3 (ig/dL) was associated with an earlier onset of sexual maturity (30%
increase) above background.
In the 2006 Pb AQCD (U.S. EPA. 2006^. 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 in Fisher 344 pups.
Subsequently, Pine et al. (2006) evaluated whether IGF-1 replacement could reverse the
effects of Pb on delayed female 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 j^ig/dL) exhibited a
marked increase in LH and luteinizing hormone releasing hormone (LHRH) secretion
after IGF-1 administration (200 ng7(.iL i.p. injection twice daily from PND23 until the
appearance of vaginal opening which appears in control animals at ~ PND 40) resulting
in restored timing of vaginal to that of control animals. It should be noted that, IGF-1
replacement in Pb-exposed animals did not cause advanced puberty over non-Pb-exposed
controls. The results of this study provide support to the theory that Pb-induced delayed
onset of puberty may be due to disruption of pulsatile release of sex hormones (U.S.
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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 sum, 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. New evidence from the toxicology literature
continues to support 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.
5.8.1.4 Effects on Lactation
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 dietary fatty acids were altered via diet. Diets deficient in
n-3 fatty acids 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,
0.2% Pb-acetate trihydrate/BW) during lactation (PND 0-21) led to a decrement in
non-essential fatty acids in the maternal organs at PND25 (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 on the 20-carbon n-6 PUFAs. In general, Pb exposure caused a decrement in
shorter chain monounsaturated and saturated FAs in maternal organs.
Dietary supplementation with calcium can be an especially important contributor to Pb
mobilization 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 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 during the lactation is less beneficial in
modulating maternal blood Pb levels (mean blood Pb at first sampling was 2.4 j^ig/dL):
the Gulson cohort was limited by power (n=10 women). In a cohort of women from
Mexico City, daily calcium supplementation during lactation reduced maternal blood lead
by 15-20% and lead 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 lead levels during pregnancy (Ettinger et al.. 2009) with the effect
strongest in women with higher biomarkers of Pb exposure (elevated baseline bone Pb or
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>5 (ig/dL blood Pb) or in women with higher Pb exposure (self-reported use of lead-
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.
5.8.1.5 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. At low concurrent blood
Pb levels, associations are observed with delayed puberty. These associations are noted
among girls with mean blood Pb levels of 10 j^ig/dL and lower. Some evidence is also
available regarding blood Pb levels and altered hormone levels in adults, although the
direction of the change varied among studies. Although studies reported inconsistent
findings for the association between Pb and fertility, there is some evidence of a potential
relationship. Most of the epidemiologic studies are cross-sectional, thus there is
uncertainty regarding the lifestages of Pb exposure associated with the greatest risk.
Toxicological studies are often dealing with prenatal or early postnatal exposures, except
for puberty studies which use concurrent exposure. 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.
5.8.2 Effects on Male Reproductive Function
The 2006 Pb AQCD (U.S. EPA. 2006b) reported on male Pb exposure or biomarker
levels and reproductive functions 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 reproductive function in
males. These studies are described in the sections below.
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5.8.2.1 Effects on Sperm/Semen Production, Quality, and
Function
1	Multiple epidemiologic and toxicological studies have examined the relationship between
2	Pb and sperm and semen production, quality, and function. These studies are summarized
3	in the text below. In addition, recent epidemiologic studies are included in Table 5-34.
4	The biological samples in these studies are from concurrent measures. The biological
5	samples measuring Pb levels and outcomes were collected concurrently in the
6	epidemiologic studies.
Table 5-34 Summary of recent epidemiologic studies of associations between
Pb levels and effects on sperm and semen
Reference8
Study
Location	Study
and Years Population
Pb Biomarker or Exposure
Measurement
Mean Pb (SD) in
Hg/dL
Adjusted Effect Estimates
Naha and Manna
(2007)
Bangalore,
India
NS
Non-occupationally
exposed controls
and occupational^
exposed workers
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	p-values for difference across the three groups
measurement	for mean values of semen profiles were <0.01
Controls 10 25	^or: liquefaction time, seminal volume, sperm
(2 26)	count sPerm DNA hyploidy, sperm
morphological abnormality, sperm motility,
Low exposure	Sperm ATPase activity, seminal plasma
50.29 (3.45)	fructose, seminal plasma total protein, seminal
High exposure	plasma free amino acid, seminal plasma
68.26 (2.49)	cholesterol
Semen Pb
measurement
Controls 2.99
(0.76)
Low exposure
15.85 (1.95)
High exposure
25.30 (2.28)
Naha and
Chowdhury (2006)
Kolkata, India Men aged 31-45
N3	that were non-
occupationally
exposed controls
and occupational^
exposed workers)
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)
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
Semen Pb
measurement
Controls 3.99
(1.36)
Low exposure
10.85 (0.75)
High exposure
18.30 (2.08)
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Reference8
Study
Location	Study
and Years Population
Pb Biomarker or Exposure
Measurement
Mean Pb (SD) in
Hg/dL
Adjusted Effect Estimates
Hsu et al. (2009b)
Taiwan
NS
Men working at a
battery plant
Blood Pb
Categorized into 3 groups:
<25 pg/dL,
25-45 pg/dL,
>45 pg/dL
40.2
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,
motility, sperm tail abnormalities, sperm
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)
Kasperczyk et al.
Poland
NS
Healthy, non-
smoking, fertile men
that worked at the
Zn and Pb
Metalworks
Blood Pb; seminal fluid Pb
Categorized as high exposure
workers (blood Pb 40-81 pg/dL),
low exposed workers (blood Pb
25-40 pg/dL), and controls
(office workers with no history of
occupational Pb exposure)
Blood Pb
High exposure
workers: 53.1
(2.05)
Low exposure
workers: 34.7
(0.83)
Controls: 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
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Study
Location	Study	Pb Biomarker or Exposure Mean Pb (SD) in
Reference3	and Years Population	Measurement	[jg/dL	Adjusted Effect Estimates
Telisman et al.
(2007)
Croatia
2002-2005
Men aged 19-55,
never
occupational^
exposed to metals
and going to a clinic
for infertility
examination or for
semen donation to
be used for artificial
insemination
Blood Pb
Median: 4.92
(range 1.13-
14.91)
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
Meeker etal. (2008)
Michigan
NS
Men aged 18-55
going to infertility
clinics (distinction
not made between
clinic visits for male
or female fertility
issues)
Blood Pb
Median: 1.50 (IQR
1.10, 2.00)
OR (95% CI) 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)
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)
Slivkova et al.
NS
Men aged 22-48
undergoing semen
analysis at an
infertility clinic
Semen Pb
1.49 mg/kg (0.40 Correlation between Pb and flagellum ball: -
mg/kg)	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
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Reference8
Study
Location	Study
and Years Population
Pb Biomarker or Exposure
Measurement
Mean Pb (SD) in
Hg/dL
Adjusted Effect Estimates
Mendiola et al.
(2011)
Spain
2005-2007
Men attending
infertility clinics and
classified as either
normal sperm
(controls) or oligo-
astheno-
teratozoospermia
(cases) based on
WHO semen quality
criteria
Seminal plasma Pb
Blood Pb
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)
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)
P (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)
Blood plasma: -0.49 (-1.8, 6.2)
% 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 pg/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
'Studies are presented in order of first appearance in the text of this section.
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, blood Pb levels over 40 (ig/dL
have been reported for individuals occupationally exposed to Pb. For example, studies
performed in India (Naha and Manna. 2007; Naha and Chowdhurv. 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 of 77.22 (ig/dL (Naha and
Chowdhurv. 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
Chowdhurv. 2006). Decreased viability and increased lipid peroxidation were detected
(Naha and Chowdhurv. 2006). A study performed in Taiwan among men with high levels
of blood Pb reported that men with higher blood Pb levels 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 and semen volume, sperm
count, motility, velocity, and reactive oxygen species production. A similar study in
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Poland included employees exposed to Pb and compared them with a group of male
office workers (Kasoerczvk 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 compared to both the control and moderate exposure
groups. Lipid peroxidation, which can induce tissue damage in sperm via reactive oxygen
species, was greater in the highest exposure group compared to the controls.
A study performed in Croatia recruited men who had never been occupationally exposed
to metals and therefore had lower blood Pb levels than the occupational studies (Telisman
et al.. 2007). 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 S-aminolevulinic acid dehydratase
[ALAD]) were used instead.
A few studies examined blood or seminal plasma Pb levels and semen quality of men at
infertility clinics (Mendiola et al.. 2011; Slivkova et al.. 2009; Meeker et al.. 2008). In
general, these men had lower levels of Pb biomarkers than men that were occupationally
exposed. 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
include 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. 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 or plasma were associated with sperm
concentration, percentage of immotile sperm, or percentage of morphologically normal
sperm.
An abundance of evidence in the toxicological literature demonstrates that Pb exposure is
detrimental to the quality and overall health of testicular germ cells. Earlier studies
showed that chronic Pb exposure (15 weeks) in adult male rabbits, resulting in blood Pb
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of 16-24 (ig/dL, induced statistically significant decrements in semen quality and greater
testicular pathology (Moorman etal.. 1998). 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 albino Wistar/NIN rats to Pb-acetate (273
or 819 mg/L in drinking water, 0.05% or 0.15%, respectively) for 45 days. Affected
endpoints included reduced epididymal sperm count, motile sperm, and viable sperm,
indicating decreased sperm production and quality. Anjum did not report blood Pb
values. Wistar/NIN rats (0.15% Pb-acetate in drinking water for 70 days) supplemented
with the herb Centella asiatica had significant attenuation of the Pb-induced changes
observed by Anjum et al. (2010). Pillai et al. (In Press) 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. Oliveira et al. (2009)
observed a negative correlation between Pb dose and intact acrosomes. Rubio et al.
(2006). Biswas and Ghosh (2006). and Salawu et al. (2009) observed a decrease in
absolute testicular weight after Pb exposure. Rubio et al. (2006) and Biswas and Ghosh
(2006) also observed a Pb-induced decrease in seminal vesicle and ventral prostate
weights and Rubio et al. (2006) reported that Pb-acetate, in a exposure 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). Reshma Anjum et al. (2010)
reported decreased testicular and epididymal weights, sperm count, and viable sperm of
male rats exposed to Pb-acetate (273 mg/L or 819 mg/L in drinking water) 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; Massanvi 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 exposure to 0.6% Pb-acetate in drinking water. 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,
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dosage schedule, duration of exposure, and age of animals at the time of the study
(Oliveira et al.. 2009).
Data from recent studies suggest that a component of Pb-induced toxicity is the
generation of reactive oxygen species (ROS) which can then affect antioxidant defense
systems of cells (Pandva et al.. 2010). 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 1%
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 etal. 2009; Madhavi et al.. 2007; Rubio et al.. 2006; Wang et al.. 2006a).
Recent studies also demonstrate that Pb may be directly toxic to mature spermatozoa
(Tapisso et al.. 2009; Hernandez-Ochoa et al.. 2006) as well as primary spermatocytes
(Nava-Hemandez et al.. 2009; Rafique et al. 2009). Nava-Hernandez et al. (2009)
exposed two groups of rodent to Pb via drinking water (LI and L2). 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 67 jj.g/1. 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 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 mice administered 21.5 mg/kg body weight Pb-acetate by i.p.
injection. 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
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drinking water (resulting in mean blood Pb levels between 19.5 and 21.9 j^ig/dL). Rafique
et al. (2009) reported degenerative changes from pyknosis to apoptosis in primary
spermatocytes. 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'). Dong et al. (2009) reported a exposure concentration-related increase in
apoptosis in spermatogonia and spermatocytes of Kunming mice after exposure to
0.15-0.6% 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 concomitant with activation of c-fos and active caspase-3 proteins.
Wang et al. (2006a) observed a 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) also reported a statistically significant increase
in mRNA expression and protein levels of Fas, Fas-L and caspase-3 after Pb exposure.
Supplementation with ascorbic acid inhibited or reduced the Pb-induced apoptosis in
germ cells and protected testicular structure and function (El Shafai et al. 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 high levels of Pb exposure have effects on sperm and
semen. In studies of men exposed to Pb in occupational settings, associations were
observed between blood Pb levels as low as 20-45 j^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. Future studies are warranted to determine whether this
association is observed at lower Pb levels.
5.8.2.2 Effects on Hormone Levels
The 2006 Pb AQCD (U.S. EPA. 2006b') provided evidence that Pb acts as an endocrine
disruptor 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 5-35 (biological samples used for the
measurement of Pb in these studies were measured concurrently with hormone levels).
Table 5-35 Summary of recent epidemiologic studies of associations between
Pb levels and hormones for males
Study
Location and
Reference8	Years
Outcome
Study Population
Pb Biomarker
or Exposure
Measurement
Mean Pb (SD) in
Hg/dL
Adjusted Effect Estimates
Telisman et al.
(2007)
Croatia
2002-2005
FSH, LH,
testosterone,
estradiol, prolactin
Men aged 19-55, never
occupationally exposed to metals
and going to a clinic for infertility
examination or for semen donation
to be used for artificial
insemination
Blood Pb
Median: 4.92
(range 1.13-14.91)
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)
Coefficients and p-values not
given if not statistically
significant (LH, FSH)
Naha and
Manna (2007)
Bangalore,
India
NS
FSH, LH,	Non-occupationally exposed
testosterone controls and occupationally
exposed workers
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 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
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Reference8
Study
Location and
Years
Outcome
Study Population
Pb Biomarker
or Exposure
Measurement
Mean Pb (SD) in
Hg/dL
Adjusted Effect Estimates
Hsieh et al.
(2009a)
Taiwan
1991 -NS
FSH, LH,	Workers at a Pb-acid battery
testosterone, factory with annual blood Pb
inhibin B	measures
Current blood
Pb, cumulative
blood Pb, time-
weighted
cumulative blood
Pb
Current blood Pb:
<10 |jg/dL: 11.6%
>40 |jg/dL: 17.1%
|3 from linear regression
Inhibin B
Current blood Pb: 0.40 (p-
value 0.40)
Cumulative blood Pb: 0.05 (p-
value 0.02)
Time-weighted cumulative
blood Pb: 1.33 (p-value 0.007)
Pearson's correlations
detected no correlations
between current 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.
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Study
Location and
Reference8	Years
Outcome
Study Population
Pb Biomarker
or Exposure
Measurement
Mean Pb (SD) in
Hg/dL
Adjusted Effect Estimates
Meeker et al.
(2010)
Michigan
NS
FSH, LH, inhibin	Men aged 18-55 going to infertility Blood Pb
B, testosterone,	clinics (distinction not made
SHBG, FAI,	between clinic visits for male or
testosterone/LH	female fertility issues)
Median: 1.50 (IQR
1.10, 2.00)
Regression coefficients (95%
CI)
FSH
1st quartiie: 1.00 (ref)
2nd quartile: 0.13 (-0.10, 0.37)
3rd quartiie: 0.10 (-0.15, 0.35)
4th quartile: 0.07 (-0.18, 0.31)
LH
1st quartile: 1.00 (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)
Inhibin B
1st quartile: 1.00 (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)
Testosterone
1st quartile: 1.00 (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: 1.00 (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: 1.00 (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)
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Study
Location and
Reference8	Years
Outcome
Study Population
Pb Biomarker
or Exposure Mean Pb (SD) in
Measurement	[jg/dL
Adjusted Effect Estimates
Mendiola et al.
(2011)
Spain
2005-2007
FSH, LH,	Men attending infertility clinics and
testosterone classified as either normal sperm
(controls) or oligo-astheno-
teratozoospermia (cases) based
on WHO semen quality criteria
Seminal plasma
Pb
Blood Pb
Seminal plasma:
2.90 (IQR 270,
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)
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 p (95% CI)
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
pg/dL)
'Studies are presented in order of first appearance in the text of this section.
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.. 2007V In addition, the analysis of an interaction term for blood Pb and
blood cadmium levels demonstrated a synergistic effect on increasing serum testosterone
levels. No association was observed between blood Pb and FSH or LH. Another 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). 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 when using current blood Pb levels (Hsieh et al.. 2009a'). A correlation
between cumulative blood Pb measures and LH levels was detected but correlations were
not present when examining FSH or testosterone levels. No correlations were apparent
between FSH, LH, or testosterone and current blood Pb levels.
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
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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 (T/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..
2011V
In a recent toxicological study, Rubio et al. (2006) observed a decrease in testosterone
levels in Pb-acetate-treated rats in a 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
(0.05% or 0.15%, 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. (In Press) found gestational and lactational exposure to 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 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 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 (Reshma Anjum et al. 2010). In contrast, Salawu et al. (2009) did not
observe a decrease in serum testosterone between control animals and animals
administered 1% 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 0.05-0.3% Pb-acetate treatment in
drinking water (only statistically significant in animals administered 0.05% 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 on the effects of Pb exposure on
circulating testosterone levels.
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5.8.2.3
Fertility
Epidemiologic studies have been performed comparing Pb and infertility in men. A study
conducted in Turkey reported that blood and seminal plasma Pb levels were different in
fertile (mean [SD] blood Pb: 23.16 [5.59] (ig/dL) and infertile men (mean [SD] blood Pb:
36.82 [12.30] jig/dL)(p <0.001, ANOVA) (kiziler et al.. 2007). The SMART study
examined the success of IVF treatment for women and their partners starting their first
round of treatment (Bloom et al.. 201 lb; Bloom et al.. 2010). A small number of the male
partners participated. Their mean (SD) blood Pb level was 1.50 (0.80) (ig/dL. Higher
blood Pb concentration was associated with greater oocyte fertilization (OR 1.08 [95%
CI: 0.97, 1.21] per 1 j^ig/dL increase in blood Pb when adjusted for Cd), 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 associated 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 Cd and Hg) (Bloom et al.. 201 lb). Another study
examined occupational Pb exposure (determined by self-report of occupational exposure)
and detected no difference in reported exposure for infertile versus fertile men (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. As with
the fertility studies among women, 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.
A couple of 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 who were exposed to 273
or 819 mg/L Pb-acetate in drinking water (0.05%or 0.15%, 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 0.05%
Pb-acetate exposed males, and 53% of the 0.15% 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 (Sainath etal. 2011).
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5.8.2.4 Puberty
1	Research has also been published examining the association between blood Pb and onset
2	of puberty in males. These epidemiological studies are summarized in Table 5-36. The
3	majority of studies used concurrent measures of blood Pb and puberty (Den Hond et al..
4	2011; Tomoum et al.. 2010; Hauser et al.. 2008). but Williams et al. (2010) performed a
5	longitudinal analysis of blood Pb levels measured at ages 8-9 years and pubertal onset.
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Table 5-36 Summary of recent epidemiologic studies of associations between
Pb levels and puberty for males
Reference8
Study
Location
Outcome Study Population Pb Biomarker
Mean Pb(SD)
in [jg/dL
Adjusted Effect Estimates
Hauser et al.
Chapaevsk,
Russia
2003-2005
Pubertal stages
defined using
standard
drawings
Healthy boys
aged 8-9
Blood Pb
Median: 3 (IQR 2-5)
blood Pb >10 pg/dL:
OR (95% CI)
Pubertal onset based on testicular
volume
<5 |jg/dL: 1.00 (Ref)
> 5 |jg/dL: 0.83 (0.43, 1.59)
'after adjustment for macronutrients,
the OR (95% CI) became 0.66 (0.44,
1.00)
Genital development
<5 pg/dL: 1.00 (Ref)
> 5 pg/dL: 0.57 (0.34, 0.95)
'after adjustment for macronutrients,
the OR (95% CI) became 0.52 (0.31,
0.88)
Pubic hair development
<5 pg/dL: 1.00 (Ref)
>5 pg/dL: 0.74 (0.34, 1.60)
Williams et al.
(2010)
Chapaevsk,
Russia
2003-2008
Pubertal stages
defined using
standard
drawings
Healthy boys
aged 8-9 at
enrollment who
had annual follow-
up evaluations
Blood Pb at ages Median: 3 (IQR 2-5)
Blood Pb level
>10 pg/dL: 3%
HR (95% CI)
Pubertal onset based on testicular
volume
<5 pg/dL: 1.00 (Ref)
> 5 pg/dL: 0.73 (0.55, 0.97)
Genital development
<5 pg/dL: 1.00 (Ref)
> 5 pg/dL: 0.76 (0.59, 0.9
Pubic hair development
<5 pg/dL: 1.00 (Ref)
>5 pg/dL: 0.69 (0.44, 1.07)
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Reference8
Study
Location
Outcome Study Population Pb Biomarker
Mean Pb(SD)
in [jg/dL
Adjusted Effect Estimates
Tomoum et al.
(2010)
Cairo, Egypt
2007
Hormones and
pubertal
development
Healthy children Blood Pb
aged 10-13
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)
NS for boys only
(combined with girls in
the study the mean
was 9.46 [3.08])
Testicular size
<10 pg/dL:
Stage 1:0%
Stage 2: 44.4%
Stage 3: 55.6%
> 10 pg/dL:
Stage 1:33.3%
Stage 2: 66.7%
Stage 3: 0%
Chi-square p-value<0.01
Pubic Hair Development
<10 pg/dL:
Stage 1:0%
Stage 2: 55.6%
Stage 3: 44.4%
>10 pg/dL:
Stage 1:33.3%
Stage 2: 66.7%
Stage 3: 0%
Chi-square p-value<0.05
Penile staging
<10 pg/dL:
Stage 1:11.1%
Stage 2: 44.4%
Stage 3: 44.4%
>10 pg/dL:
Stage 1:58.3%
Stage 2: 41.7%
Stage 3: 0%
Chi-square p-value<0.05
Mean testosterone level
<10 pg/dL:
4.72 (SD 1.52)
>10 pg/dL:
1.84 (SD 1.04)
DenHondetal. Flanders Tanner staging Boys ages 14 and Blood Pb
(2011)	2003-2004 anc'	15, in their 3rd
gynecomastia year of secondary
education and
living in the same
study areas for at
least 5 years
'Studies are presented in order of first appearance in the text of this section.
'Quantitative results for LH and FSH
not provided
Median: 2.50	OR (95% CI) for gynecomastia with
10th percentile: 1.20 doublin9 of exposure
90th percentile: 5.12 1.84(1.11,3.05)
No association between Pb and pubic
hair or genital development (results not
given)
1	Studies were performed among a cohort of Russian boys enrolled between ages 8-9 years
2	(Williams et al.. 2010; Hauser et al.. 2008). The area where these studies were performed
3	had various environmental contaminants such as dioxin, polychlorinated biphenyls, and
4	other metals, present. Both the cross-sectional study (Hauser et al.. 2008) and the
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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
development compared to those with lower levels (median age in the high blood Pb group
was 12.5 years compared to 13.0 years in the low blood Pb group) (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. 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 older 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 PND 35. These pups were exposed chronically to 0.15 or 0.45%
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 (ig/dL, respectively (Ronis et al.. 1998b).
Cynomolgus monkeys exposed to Pb over a lifetime (10 years, blood Pb levels ranging
from 30-60 (ig/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; Foster. 1992). 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, 0.1 or 0.3% Pb-acetate in drinking water, respective blood Pb levels
of 34 or 60 (ig/dL) that showed significant decreases in FSH, ventral prostate weight and
serum testosterone but no change in serum LH (Sokol et al.. 1985). These Pb-exposed
adult male rats (0.3% 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 the hypothalamus (Klein et al.. 1994; Sokol. 1987).
Thus, Pb likely interferes with the male HPA axis, contributing to its reproductive
toxicity.
In summary, recent epidemiologic studies have demonstrated an inverse effect of Pb on
pubertal development among boys at low concurrent 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
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earlier findings that Pb exposure may result in delayed onset of male puberty and altered
reproductive function later in life in experimental animals.
5.8.2.5 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 of testes in Pb nitrate-treated animals (a single i.p. dose of 12.5, 25, or 50 mg/kg
of BW and were sacrificed 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 (Massanvi et al.. 2007). Pillai
et al. (In Press) found gestational and lactational exposure to 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, 0.05% or 0.15%, 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. .
5.8.2.6 Summary of Effects on Male Reproductive Function
Evidence of associations between Pb exposure and male reproductive function vary 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. Many of the
epidemiologic studies included occupational cohorts, which had high blood Pb levels.
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. In addition, recent epidemiologic studies
found blood Pb levels to be associated with delayed pubertal development in boys with
low concurrent blood Pb levels (mean/median blood Pb levels <10 (ig/dL). This is
supported by earlier toxicological studies. 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.
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5.8.3
Effects on Ovaries, Embryo Development, Placental function, and
Spontaneous Abortions
The 2006 Pb AQCD (U.S. EPA. 2006b) 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 levels (U.S. EPA. 2006^. and no recent studies have been
performed to examine paternal Pb levels and spontaneous abortion. Since the 2006 Pb
AQCD, multiple epidemiologic studies have been published that examine Pb levels in
women and their possible association with spontaneous abortion. Additionally,
toxicological studies have studied the effects of Pb on fetal loss and the contribution of
the ovaries and placenta to fetal loss.
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Table 5-37 Summary of recent epidemiologic studies of associations between
Pb levels and spontaneous abortions
Reference8
Study
Location
Outcome
Study population
Pb Biomarker
Mean Pb (SD) in
Mg/dL
Adjusted Effect Estimates
Yin et al. (2008)
Shanxi
Province,
China
2004-2006
Anembryonic Wfomen age 25-35 yr Maternal blood Pb Cases: 5.3 (95% CI: Comparisons between log-
pregnancy
old 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
after miscarriage for
cases and at study
enrollment for
controls
5.2, 5.9)
Controls: 4.5 (95%
CI: 3.7, 5.0)
transformed blood Pb levels of
cases and controls performed via
Student's t-test had a p-value of
0.03
Vigeh et al. (2010)
Tehran, Iran
2006-2008
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
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: 0.41
OR: 0.331 (95% CI: 0.011,
10.096) for an increase in log-
transformed blood Pb (units not
given, assume 1 pg/dL)
Lamadrid-Figueroa et
al. 12007)
Mexico City,
Mexico
1997-1999,
2001-2004
Previous
miscarriage
Women who had a
previous pregnancy
and were currently
pregnant with
gestational age of
< 14wks
Maternal and
umbilical cord blood
Pb, maternal bone Pb
Overall:
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)
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)
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)
Gundacker et al.
(2010)
Vienna, Austria
2005
Previous
miscarriage
Women recruited
during the second
trimester of
pregnancy
Whole placentas
shortly after birth
Median (IQR):
25.8(21.0, 36.8)
Median Placenta Pb:
Women who had not previously
miscarried: 27 pg/kg
Women who had previously
miscarried: 39 pg/kg
(p-value for difference: 0.039)
'Studies are presented in order of first appearance in the text of this section.
Table 5-37, above, provides a summary of the recent epidemiologic studies examining
the association between Pb biomarker levels and past and current spontaneous abortion.
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. 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 adjusted for nutrient status. Another study
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examining spontaneous abortions occurring early in the pregnancy was conducted in Iran
(Vigeh et al. 2010). Mean blood Pb concentrations 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. A study in Turkey
reported on groups of women who either had a spontaneous abortion before the
20th week of gestation or who had a viable pregnancy (Faikoglu et al. 2006). No
difference was detected between the blood Pb levels of the two groups (Pb levels not
reported here due to calculation errors discovered in the paper; errors do not appear to
affect conclusions). 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 higher rates of previous
spontaneous abortions than did women in the lowest tertile. The authors state that the
plasma/whole blood ratio represents the bioavailability of Pb, which is capable of
crossing the placental barrier for a given blood concentration. No association was
observed when examining 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 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 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 understand to
Pb-dependent 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. (2009c) observed
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) were 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 a positive
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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, 0.05 or 0.15% Pb Acetate,
respectively) male rats from Reshma Anjum et al. (2010) that had an exposure
concentration-dependent decreases in serum testosterone, decreased male reproductive
organ weight and decreased sperm were mated to untreated females. These untreated
dams had male related exposure 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. 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 (0.05 mg/kg body
weight).
Previous studies demonstrated that Pb accumulates in the ovaries and causes histological
changes, thus contributing to Pb-induced effects on female fertility (U.S. EPA. 2006^. 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). Nampoothiri and Gupta
(2006) reported that Pb exposure caused a decrease in cholesterol and total phospholipid
content in the membranes of granulosa cells which resulted in increased membrane
fluidity. 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).
Overall, the recent studies support the conclusions of the 2006 Pb AQCD that there is
insufficient evidence among epidemiologic studies to suggest an 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 are 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 levels do
lead to increased risk of early spontaneous abortions. Toxicological data provide
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mechanistic understanding 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 mixed evidence on the
role of Pb in spontaneous abortions.
5.8.4 Infant Mortality and Embryogenesis
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 32 to >70 (ig/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 of 54 (ig/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.. 2007). No recent studies have reported on the
relationship between Pb levels and infant mortality.
5.8.5 Birth Defects
The 2006 Pb AQCD (U.S. EPA. 2006b) reported the possibility of small associations
between high Pb exposure and birth defects, but many of the epidemiologic studies used
occupational histories instead of actual measures of blood Pb levels. Among the studies
included in the 2006 Pb AQCD, a couple studies reported possible associations between
parental exposure to Pb and neural tube defects (Imens et al.. 1998; Bound et al.. 1997).
Recent studies also examined indicators of Pb exposure and neural tube defects (Table
5-38). 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.
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Table 5-38 Summary of recent epidemiologic studies of associations between
Pb levels and neural tube defects
Reference8
Study Location
Study Population
Pb Biomarkers and
Exposure Measurement
Mean Pb (SD) in [jg/dL
Adjusted Effect Estimates
Zeyrek et al. (2009)
Turkey
NS
Infants with
gestational age of at
least 20 wks
Maternal and umbilical cord
blood Pb taken 0.5h after
birth
Cases: Maternal: 15.5 (15.0)
Umbilical cord: 18.2 (17.8)
Controls: Maternal: 12.5
(12.7)
Umbilical cord: 16.5 (16.1)
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 blood
Pb
Brender et al. (2006)
Texas
1995-2000
Infants of Mexican-
American women
Maternal blood Pb
taken 5-6 wks post-partum
Cases: 2.4 (1.9)
Controls: 2.5 (1.6)
Blood Pb<6.0 pg/dL: 1.0 (Ref)
Blood Pb> 6.0 pg/dL: 1.5 (95% CI:
0.6, 4.3)
Huang et al. (2011b)
China
2002-2004
Live and still births of
women living in the
study area (villages in
the Lvliang region of
Shanxi province)
2 soil samples from each
village
56.14 pg/g (11.43 pg/g)
N/A
aStudies are presented in order of first appearance in the text of this section.
Among the recent epidemiologic studies (described in Table 5-38), 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
included) (Zevrek 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 lb). 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.
Previous studies included in the 2006 Pb AQCD observed associations between Pb and
neural tube defects but were limited due to the lack of biologically measured Pb [Pb was
measured in drinking water (Bound et al.. 1997) and estimated from occupational reports
(Irgens et al.. 1998)1. 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.
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1
2
2011b). Other recent epidemiologic studies of maternal blood Pb levels and neural tube
defects observed no associations (Zevrek et al.. 2009; Brender et al.. 2006).
5.8.6	Preterm Birth
3	Epidemiologic studies on preterm birth included in the 2006 Pb AQCD (U.S. EPA.
4	2006b) reported inconsistent findings regarding the relationship between Pb and
5	gestational age. Recent studies have examined this potential association and again mixed
6	results were reported (Table 5-39). Of these studies, the ones that categorized births as
7	preterm or term all defined preterm birth as less than 37 weeks of gestation. One
8	limitation to note for these studies is that if Pb affects spontaneous abortion and length of
9	gestation via a similar pathway, then the studies that only collect data at delivery and not
10	at earlier stages of pregnancy would be biased toward the null.
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Table 5-39 Summary of recent epidemiologic studies of associations between
Pb levels and preterm birth
Reference8
Study
Location
Outcome
Study Population
Pb Biomarkers or
Exposure Measurement
Mean Pb (SD)
in [jg/dL
Adjusted Effect estimates
Berkowitz et al.
(2006)
Idaho
1970-1981
Preterm birth
(<37wk)
Singleton births
with 28-45 wk
gestation
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"
NS
OR (90% CI) (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)
Patel and Prabhu
(2009)
Nagpur,
India
NS
Gestational age
Consecutive births at
the study hospital
Umbilical cord blood Pb
Umbilical cord
blood Pb: 4.7
(12.1)
>5 pg/dL: mean gestational age 38
wks
< 5 pg/dL: mean gestational age 39
wks
Linear regression: gestational age
decreased 1 wkwith every 1 pg/dL
increase in umbilical cord blood Pb
(exact values and 95% CI not given)
Jones etal. (2010) Tennessee Gestational Aae:
Singleton births > 27 Umbilical cord blood Pb
2.4 (4.3)
Geometric Mean:
2Q06 preterm (<37wk)
wk gestation from
Geometric
Preterm birth: 1.4
, term (37-40
wk), post-term
mothers aged 16-45
living in the Shelby
mean: 1.3
Term birth: 1.2
(>40 wk)
County area for at

Post-term birth: 1.3

least 5 mo during

p-value for difference: >0.10

pregnancy

VNfells et al. (2011a)
Baltimore,
MD
2004-2005
Gestational age
Singleton births from
the Baltimore
Tracking Health
Related to
Environmental
Exposures (THREE)
study
Umbilical cord Pb
0.84 (95%: CI
0.72, 0.96)
: 5 |jg/dL: 0.7%
Ratio for Pb concentration per 10
days of gestation: 0.99 (0.93,1.06]
Jelliffe-Pawlowski
et al. 12006)
California
1995-2002
Preterm birth
(<37 completed
wk)
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
Maximum maternal blood
Pb during pregnancy
> 10 pg/dL:
30.9%
Odd Ratios:
<5 pg/dL: 1.00 (Ref)
6-9 pg/dL: 0.8 (0.1, 6.4)
10-19 pg/dL: 1.1 (0.2, 5.2)
20-39 pg/dL: 4.5 (1.8, 10.9)
>40 pg/dL: 4.7 (1.1, 19.9)
<10 pg/dL: 1.00 (Ref)
>10 pg/dL: 3.2 (1.2, 7.4)
Vigeh et al. (2011)
Tehran, Iran
2006
Preterm birth
(20-37 wk)
Singleton births from
non-smoking, non-
obese mothers aged
16-35 and referred
for prenatal care
during the 8th-12th
week of gestation
Maternal blood Pb	3.8 (2.0)	Mean blood Pb (SD):
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 pg/dL)
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Reference8
Study
Location
Outcome Study Population
Pb Biomarkers or
Exposure Measurement
Mean Pb (SD)
in [jg/dL
Adjusted Effect estimates
Cantonwine et al.
(2010a)
Mexico City
1997-1999
Preterm birth
(<37wk),
Gestational age
Births to mothers
with at least 1 blood
Pb measurement
during pregnancy
and no chronic
diseases requiring
medication
Maternal blood Pb during
pregnancy
Blood Pb
Visit at <20wks
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 <20wks
pregnant 0.17
(0.16)
Visit at 20-28
weeks pregnant
0.13(0.10)
Visit at >28
weeks pregnant
0.16(0.26)
Linear regression p (95% CI)
Blood Pb
Visit at <20wks-2.76 (-5.21,-0.31)
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 <20wks-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)
Plasma-to-blood Pb ratio
Visit at <20wks-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)
Zhu et al. (2010)
New York
2003-2005
Preterm birth
(<37 completed
wk)
Singleton births to
mothers aged 15-49
with blood Pb
measures before or
on the date of
delivery and blood
Pb measuring <10
Mg/dL
Maternal blood Pb
2.1
Odd Ratios:
< 1.0 pg/dL: 1.00 (Ref)
1.1-2.0 pg/dL: 1.03 (0.93, 1.13)
2.1-3.0 pg/dL: 1.01 (0.92, 1.10)
3.1-9.9 pg/dL: 1.04 (0.89, 1.22)
Chen et al. (2006a) Taiwan Preterm birth
1993-1997 (<37 wk)
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
occupational^
exposed to Pb
Maternal blood Pb during
pregnancy (or if that wasn't
available, 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)
Maternal blood
Pb 10.1 (10.4)
Paternal blood
Pb 12.9 (13.8)
Risk Ratios
Maternal blood Pb
<10 pg/dL: 1.00
10-19 pg/dL: 1.97 (0.92, 3.86)
>20 pg/dL: 1.86 (0.68, 4.28)
Paternal blood Pb
<10 pg/dL: 1.00
10-19 pg/dL: 1.17 (0.53, 2.32)
>20 pg/dL: 0.55 (0.19, 1.28)
Orunetal. (2011) Turkey Preterm birth Births to mothers not	Breast milk 2 months post- Median: 20.6
^3	(<37 wk)	occupational^	partum	pg/L
exposed to toxic
metals and living in a
suburban but non-	>WHO limit (5
industrial area	k,9"-): 87%
Median Pb (IQR)
>37 wk:20.6 (11.2, 29.2) pg/L
<37 wk: 20.4(14.4, 27.9) pg/L
p-value for Mann-Whitney U test:
>0.05
'Studies are presented in order of first appearance in the text of this section.
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 period of high Pb emissions due to damage at a local
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1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
32
33
34
35
36
37
38
factory. Preterm birth rates were examined during three time periods: before, during, and
after the time of higher Pb exposure. No association 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.
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 concentrations of cord blood Pb among study participants
were overall low (99.3% had umbilical cord blood Pb < 5 j^ig/dL) (Wells etal.. 2011a).
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) were matched with the birth certificates of their infants
(Je 11 iffe- Paw 1 owski et al.. 2006). Almost 70% of women had maximum blood Pb
measurements <10 (ig/dL with the majority being <5 (ig/dL. Preterm birth was associated
with higher blood Pb when comparing women with maximum 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
J e 11 i ffe- Paw 1 owsk i et al. study (J e 11 i ffe - Pa w 1 o w sk i et al.. 2006). Higher maternal blood
Pb level was associated with higher odds of preterm birth. Another study examining
blood Pb and gestational age among women with lower blood Pb levels 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, a
study conducted in New York among women with lower blood Pb levels (inclusion
criteria mandated that blood Pb concentration be less than 10 (ig/dL), no association was
observed between blood Pb levels and preterm birth (Zhu et al.. 2010). Similarly, a study
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1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
of maternal and paternal blood Pb concentrations reported no association between
maternal or paternal blood Pb levels and preterm birth (Chen et al.. 2006a').
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 (Oriin et al.. 2011).
In sum, as in the 2006 Pb AQCD, 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.
5.8.7 Low Birth Weight/Fetal Growth
The 2006 Pb AQCD reported inconsistent epidemiologic study results for the
associations between Pb and birth weight/fetal growth and 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 5-40 below. Additionally, there have been a few recent
toxicological studies evaluating the effect of Pb exposure during gestation on birth
weight.
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Table 5-40 Summary of recent epidemiologic studies of associations between
Pb levels and low birth weight and fetal growth
Reference8
Study
Location
Outcome
Pb Biomarkers and Mean Pb (SD) in
Study population Exposure Measurement [jg/dL Adjusted Effect Estimates
Lamb et al. (2008)
Mitrovica
and
Pristina,
Yugoslavia
1985-1986
Height and BMi at birth
Participants of the
Yugoslavia Study
of Environmental
Lead Exposure,
Pregnancy
Outcomes, and
Childhood
Development
Mid-pregnancy blood Pb
Mitrovica: 20.56
(7.38)
Pristina: 5.60
(1.99)
Regression Coefficients (95%
CI) for 1 pg/dL increase in Pb:
BMI
Mitrovica: -0.18 (-0.69, 0.33)
Pristina:-0.14 (-0.69, 0.42)
Height
Mitrovica: 0.43 (-0.83,1.69)
Pristina: 0.35 (-0.64,1.34)
Jelliffe-Pawlowski et
al. (2006)
California
1995-2002
Low birth weight
(<2,500g)
Small for gestational
age (birth weight for
gestational age <10th
percentile of race- and
gender- specific norms
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 and
matched to birth
records
Maximum maternal blood
Pb during pregnancy
> 10 pg/dL:
30.9%
Odd Ratios:
Low birth weight < 5 ug/dL:
1.00 (Ref)
6-9 pg/dL: -
10-19 pg/dL: 2.7 (0.5, 14.8)
20-39 pg/dL: 1.5(0.3, 7.7)
>	40 pg/dL: --
<10 pg/dL: 1.00 (Ref)
>10 pg/dL: 3.6 (0.3, 40.0)
Small for gestational age
<5 pg/dL: 1.00 (Ref)
6-9 pg/dL: -
10-19 pg/dL: 2.3 (0.6, 9.2)
20-39 pg/dL: 2.1 (0.7, 6.7)
>	40 pg/dL: --
<10 pg/dL: 1.00 (Ref)
>10 pg/dL: 4.2 (1.3, 13.9)
Iranpour et al.
(2007)
Isfahan,
Iran
2005
Low birth weight
(<2,500g, >37wk)
Full-term infants
born at a hospital
affiliated with
Isfahan University
Umbilical cord and
maternal blood Pb within
12 h of delivery
Maternal blood
Pb:
Cases: 12.5(2.0)
Controls: 13.5
(2.7)
Umbilical cord
blood Pb:
Cases: 10.7 (1.7)
Controls: 11.3
(1.9)
P-values fort-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
P-values for correlations:
Umbilical cord blood Pb and
birth weight:
Low birth weight: 0.84
normal birth weight: 0.26
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Reference8
Study
Location
Outcome
Pb Biomarkers and Mean Pb (SD) in
Study population Exposure Measurement [jg/dL Adjusted Effect Estimates
Gundacker et al.
(2010)
Vienna,
Austria
2005
Birth length, birth
weight, head
circumference
Infants of women Maternal blood Pb between Median (IQR):
recruited during
their second
trimester
wk 34-38 of gestation,
whole placentas and
umbilical cord Pb shortly
after birth, meconium
samples in first five days
after birth
Maternal blood
Pb: 2.5 (1.8, 3.5)
Umbilical cord
blood Pb: 1.3
(0.8, 2.4)
Placenta Pb: 25.8
|jg/kg (21.0, 36.8
MQ/kg)
Meconium Pb:
15.5 pg/kg (9.8,
27.9 pg/kg)
Regression coefficients (units
not given, assume results are
per 10 pg/dLor 1 pg/kg)
Birth length:
Placenta Pb: 0.599 (SE
0.154, p-value <0.001)
Meconium Pb: -0.385 (SE
0.157, p-value 0.012)
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)
Zhu et al. (2010)
New York
2003-2005
Birth weight, small for
gestational age (birth
weight for gestational
age <10th percentile
based on national birth
weight by gestational
week from weeks 25-42
Singleton births to
mothers aged 15-
49 with blood Pb
measures before
or on the date of
delivery and blood
Pb measuring <10
pg/dL
Maternal blood Pb
2.1
Difference in birthweight in
grams:
: Ref
-27.4 (
-38.8 (
-47.5 (
-54.8 (
-61.3 (
-67.2 (
-72.5 (
-77.6 (
0	pg/dL
1	pg/dL
2	pg/dL
3	pg/dL
4	pg/dL
5	pg/dL
6	pg/dL
7	pg/dL
8	pg/dL
9	pg/dL
-82.3 (-113.3,
10 pg/dL: -86.7 (-119.4, -54.0)
After exclusion of blood Pb <1
pg/dL, a 1 pg/dL increase in
blood Pb was associated with
a 7.0 g decrease in
birthweight
Odd Ratios for small for
gestational age:
<1.0 pg/dL: 1.00 (Ref)
1.1-2.0 pg/dL: 1.07 (0.98,
1.17)
2.1-3.0 pg/dL: 1.06 (0.98,
1.16)
3.1-9.9 pg/dL: 1.07 (0.93,
1.23)
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Reference8
Study
Location
Outcome
Pb Biomarkers and Mean Pb (SD) in
Study population Exposure Measurement [jg/dL Adjusted Effect Estimates
Chen et al. (2006a)
Taiwan
1993-1997
Low birth weight
(<2,500 g), small for
gestational age (birth
weights 10th percentile
of sex- and gestational
wk weights for
singletons in 1993-
1996)
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
Maternal blood Pb during
pregnancy (or if that wasn't
available, 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)
Maternal blood
Pb 10.1 (10.4)
Paternal blood Pb
12.9(13.8)
Risk Ratios
Low birth weight
Maternal blood Pb
<10 pg/dL: 1.00 (Ref)
10-19 |jg/dL: 2.22 (1.06,
4.26)
>20 pg/dL: 1.83 (0.67, 4.20)
Paternal blood Pb
<10 pg/dL: 1.00 (Ref)
10-19 pg/dL: 0.83 (0.34,
1.75)
> 20pg/dL: 0.42 (0.12, 1.06)
SGA
Maternal blood Pb
<10 pg/dL: 1.00 (Ref)
10-19 pg/dL: 1.62 (0.91,
2.75)
>20pg/dL: 2.15 (1.15, 3.83)
Paternal blood Pb
<10 pg/dL: 1.00 (Ref)
10-19 pg/dL: 0.94 (0.49,
1.66)
>20 pg/dL: 0.94 (0.51, 1.62)
Kordas et al. 12009)
Mexico City,
Mexico
1994-1995
Head circumference,
birth weight, birth length
Infants of mothers
receiving antenatal
care at hospitals
serving low-to-
middle income
populations (cross-
sectional study of
baseline info from
Ca
supplementation
trial)
Umbilical cord and
maternal blood Pb within
12 h of delivery; maternal
tibia Pb
Maternal tibia Pb:
9.9 pg/g (9.8
Mg/g)
Maternal blood
Pb> 10pg/dL:
27%
Umbilical cord
blood Pb >
Regression coefficients (SE)
(adjusted for maternal BMI,
maternal height, infant
gestational age, and other
variables) for each 1 pg/g
increase in tibia Pb:
Birth weight: -4.9(1.8)
Birth length:-0.02 (0.01)
Head circumference: -0.01
10pg/dL: 13.7% (0.01; p-value<0.05)
Women with 4th quartile tibia
Pb (15.6-76.5 pg/g) delivered
infants 140 g less than
women with tibia Pb in the
lowest quartile
Afeiche et al. (2011)
Mexico City
1994-2005
Birth weight
Term, singleton
births, at least
2,500 grams
enrolled in one of
three birth cohorts
recruited for other
longitudinal studies
Maternal patella and tibia
Pb measured at 1 month
postpartum
Patella Pb 10.4
(11.8) pg/g
Tibia Pb 8.7 (9.7)
pg/g
p (95% CI) 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,
(results not given)
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Reference8
Study
Location
Outcome
Pb Biomarkers and Mean Pb (SD) in
Study population Exposure Measurement [jg/dL Adjusted Effect Estimates
Cantonwine et al.
(2010b)
Mexico City Birthweight
1994-1995
Infants who were
part of a clinical
trial to assess
maternal calcium
supplementation
on bone lead
mobilization during
lactation
Umbilical cord blood Pb
Maternal tibia and patella
Pb one month after delivery
Umbilical cord
blood Pb varied
by genotype from
6.3 to 6.9
Umbilical cord
blood Pb > 10
pg/dL: 12.6%
Regression models
P (95% CI)
Umbilical cord blood Pb -31.1
(-105.4, 43.3)
Maternal tibia Pb
Overall -4.4 (-7.9, -0.9)
<1-4.1 pg/g: Ref
4.1-9.2	pg/g: 17.2 (-75.6,
110.1)
9.2-15.4	pg/g:-19.1 (-112.1,
73.9)
15.4-43.2 pg/g:-95.4 (-189.9,
-0.8)
Llanos and Ronco
(2009)
Santiago,
Chile
NS
Fetal growth restriction
(1,000-2,500g)
"note normal birth
weights were >3,000g
Term births (37-40
wks) from non-
smoking mothers
Placenta Pb
Fetal growth
restricted: 0.21
pg/g (0.04 pg/g)
Controls: 0.04
pg/g (0.009 pg/g)
P-value for Mann-Whitney
U-test <0.01
Zentner et al. (2006)
Santo
Amaro,
Brazil
2002
Birth weight and length
Singleton births
with maternal
residence within
5 km of Pb smelter
Umbilical cord blood Pb
from delivery
Umbilical cord
blood Pb: 3.9
(3.6)
Linear regression coefficient
with umbilical cord blood Pb
as the dependent variable in
model with only length and
weight (unit not given,
assume per 1 pg/dL): Length
-0.46 (p-value 0.003) and
Weight -0.275 (0.048) (i.e., in
this study, Pb is assessed as
the outcome)
Atabeketal. (2007)
Turkey Birth weight, birth
N3	length, head
circumference, mid-arm
circumference
Term, singleton
infants born to
healthy mothers
living in urban
areas and
assumed to have
high Pb
concentrations
Umbilical cord blood Pb 14.4(8.9)
Umbilical cord
blood Pb > 10
pg/dL: 53.7%
Umbilical cord
blood Pb > 25
pg/dL: 9.2%
Regression models
p (p-value)
Birth weight
-0.81 (0.01)
Birth length
0.41 (0.05)
Mid-arm circumference
0.30 (0.05)
Al-Saleh et al.
(2008b)
Saudi
Arabia
2004
Head circumference
Infants with a
gestational age of
at least 34 weeks
born to healthy
mothers aged 17-
46 years and non-
occupationally
exposed to Pb
Umbilical cord blood Pb 2.210(1.691)
Umbilical cord
blood Pb >10
pg/dL: 1.23%
Regression models for those
above the 75th percentile of
cord blood Pb levels
p (SE) per unit of log-
transformed Pb
-0.158 (0.718), p-value: 0.036
Janjua et al. (2009)
Karachi,
Pakistan
2005
Low birth weight
(<2,500g)
Infants of randomly
selected women
who planned to
deliver between
37-42 wk
Umbilical cord blood Pb
Umbilical cord
blood Pb: 10.8
(0.2)
Prevalence ratio:
<10 pg/dL: 1.00 (Ref)
>10 pg/dL: 0.82 (0.57, 1.17)
Jones et al. (2010)
Tennessee
2006
Low birth weight
(<2.500g)
Singleton births
> 27 wks gestation
from mothers aged
16-45 living in the
Shelby County
area for at least 5
mo during
pregnancy
Umbilical cord blood Pb
2.4 (4.3)
Geometric mean:
1.3
Geometric Mean:
Low birth weight: 1.2
Normal birthweight: 1.3
p-value for difference: >0.10
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Reference8
Study
Location
Outcome
Pb Biomarkers and Mean Pb (SD) in
Study population Exposure Measurement [jg/dL Adjusted Effect Estimates
VNfells et al. (2011a) Baltimore,
MD
2004-2005
Birth weight
Singleton births
from the Baltimore
Tracking Health
Related to
Environmental
Exposures
(THREE) study
Umbilical cord Pb
0.84 (95%: CI
0.72, 0.96)
: 5 |jg/dL: 0.7%
Ratio for Pb concentration per
100g birth weight: 1.01 (0.99,
1.02)
Orun et al. (2011)
Turkey
NS
Birth weight and head
circumference
Births to mothers
not occupationally
exposed to toxic
metals and living in
a suburban but
non-industrial area
Breast milk 2 months post-
partum
Median: 20.6 pg/L
>WHO limit (5
pg/L): 87%
Median (IQR)
<2500g: 20.4
(8.5, 27.1) pg/L
> 2500g: 20.6
(11.8, 29.5) pg/L
Correlations for breast milk
Pb and z-scores of head
circumference
Girls 0.087
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
Williams et al.
(2007)
Tennessee Birth weight
2002
Infants from
singleton births or
the firstborn infant
in a set of multiples
Air Pb levels during first
trimester of pregnancy
0.12 pg/m3 (0.04
pg/m3)
p-value for multilevel
regression of Pb with birth
weight: 0.002
Increase of Pb from 0 to 0.04
relates to a 38g decrease in
birth weight
IncreaseofPbfrom Oto0.13
(maximum) relates to a 124g
decrease in birth weight
Berkowitz et al.
Idaho Low birth weight
1970-1981 (<2,500 g and £ 37 wk)
Small for gestational
age (birth weight < 5th
percentile of sex- and
gestational wk weights
for singletons in Idaho)
Singleton infants Three time periods of two Not specified
with 28-45 wk locations (unexposed and
gestation	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"
Term Low birth weight:
OR (90% CI) (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% CI) (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)
'Studies are presented in order of first appearance in the text of this section.
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. Multiple
studies were conducted that examined the association between maternal blood Pb and
birth weight/fetal growth. In California, blood Pb measurements of women during
pregnancy were matched with the corresponding birth certificates (Jelliffe-Pawlowski et
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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 subsequently wide 95% CI. An
association was detected for high blood Pb and having an infant who was small of his/her
gestational age (SGA). A study of term births in Iran reported no difference in blood Pb
of women giving birth to a normal weight infant and women giving birth to an infant with
low birth weight (I ran pour et al.. 2007). 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, with the strongest
associations observed at the lowest blood Pb level (Zhu et al. 2010). No association was
observed between maternal blood Pb and 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).
A study examining the association between Pb biomarker levels and birth weight used
tibia bone measurements 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
genotypes (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, 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 tibia Pb and birth weight, especially at higher levels. This association
was stronger among those mothers with variants of the hemochromatosis iron gene
(HFE).
Multiple studies examined the relationship between Pb level and birth weight using Pb
measured from the placenta or umbilical cord. Researchers in Chile collected the
placentas from term births and compared the Pb levels for those born with normal birth
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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 jj.g/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. 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 association with other measures of growth, such as birth length and
mid-arm circumference, were detected. 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 and
head circumference. 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 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 et al.. 2010). A study performed in Baltimore, MD
also reported no association between umbilical cord blood Pb concentration and birth
weight (Wells etal.. 2011a). This study had low blood Pb levels, with only 0.7% of
participants having umbilical cord blood Pb measuring > 5 (ig/dL. 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.. 2010b). 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.
A study performed in Turkey examined the relationship between Pb levels in breast milk
two months postpartum and size at birth (Oriin 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. 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
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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 a local factory that lead to high Pb emissions
during one of the time periods under study (Berkowitz et al.. 2006). No levels of Pb are
provided. 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 5 th 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 and impaired postnatal growth (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 0.025% 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 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. This effect did not persist
in the postnatal growth of the rats. 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 j^ig/dL), but no changes in body length
were reported. In contrast, Leasure et al. (2008) reported a statistically significant inverse
relationship between Pb exposure and body weight for male mice exposed to low
(27 ppm)-and high (55 ppm)-levels of Pb during gestation. Male mice exposed to the low
and high Pb concentrations during gestation were 26% and 13% heavier than controls at
1 year of age, respectively. In this study, dams were administered 27 ppm (low), 55 ppm
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(moderate), and 109 ppm (high) Pb in drinking water beginning 2 weeks before mating
and continuing until PND10. Resulting blood Pb levels ranged from 10 (ig/dL or less in
the low-exposure offspring to 42 (ig/dL in the high-exposure offspring at PND10. The
authors also reported that when dams received low or moderate levels of Pb in drinking
water from birth to weaning neither male nor female offspring exposed to Pb postnatally
exhibited a difference in body weight when compared to control offspring.
In summary, associations were observed between Pb and low birth weight in a study of
maternal bone Pb and studies of Pb air exposures and birth weight. However, the
associations were less consistent when using maternal blood Pb or umbilical cord and
placenta Pb as the exposure measurement. Previous toxicological studies observed an
association between gestational Pb exposure and reduced birth weight with moderate to
high dose Pb. More recent findings using low dose Pb exposure reported increased
offspring body weight after developmental Pb exposure.
5.8.8 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. 2006^. 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.
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Table 5-41 Summary of recent epidemiologic studies of associations between
Pb levels and postnatal growth
Reference
Study Location and
Years
Study Population Pb Biomarker	Mean Pb (SD)
Effect Estimates
Afieche et al. (2011) Mexico City, Mexico	n=522 boys
Children born between n=477 girls
1994 and 2005
Maternal bone Pb Patella: 10.4 (11.8) pg/g
Change in weight at 5 years of age
(g) per 1 SD increase in maternal
bone Pb
Girls: 130.9 (95% CI:-227.4 to-
34.4)
Boys: 13.0 (95%CI -73.7, 99.9)
Adjusted for cohort, maternal age,
calf circumference, height,
education, number of pregnancies,
breast feeding for 6 months,
calcium treatment, child's
gestational age at birth, height.
Sanna et al. (2011) Sardinia, Italy
n=825 children 11-14 yrs Pb in hair
1998
5.84 (6.56) pg/g
Height
Data collected in 1998,
old
2002
1.49(1.72) pg/g
1998
p log Pb= -0.121 (p=0.0021)
2002 and 2007

2007
0.78 (0.93) pg/g
2002
p log Pb= -0.115 (p=0.0349)




2007
p log Pb= -0.011 (p=0.8665)




Sitting Height




1998
p log Pb=-0.117 (p=0.0017)




2002
P log Pb=-0.036 (p=0.5149)




2007
P log Pb=0.028 (p=0.6633)




ELL





1998
p log Pb=-0.103 (p=0.0209)




2002
p log Pb=-0.164 (p=0.0057)




2007
P log Pb=-0.008 (p=0.9058)




Adjusted for age and sex
Ignasiak et al.
South-western Poland
1995
(Industrial area with copper
smelters and refineries)
school children 7-15 years Concurrent blood Pb
n=463 boys
n= 436 girls
7.7 (3.5) pg/dL	Estimated decrement per 10 pg/dL
increase in blood Pb
Wfeight:
Boys: 2.8 kg
Girls: 3.5 kg
Height:
Boys: 3.2 cm
Girls: 4.0 cm
Trunk length:
Boys: 1.2 cm
Girls: 1.1 cm
Leg length:
Boys: 2.1 cm
Girls: 2.9 cm
Arm length:
Boys: 1.8 cm
Girls: 1.9 cm
Adjusted for age, age2, and
education level of mother
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Reference3
Study Location and
Years
Study Population Pb Biomarker Mean Pb (SD)
Effect Estimates
Hauser et al.
Chapaevsk, Russia
May 2003 - May 2005
n=489 boys
8-9 yrs old
Concurrent blood Pb
3 (2-5) |jg/dL
Median (25-75 percentile)
Regression coefficient (95% CI)
Height (cm):-1.439 (-2,25, -0.63)
VNfeight (kg): -0.761 (-1.54, 0.02)
BMI:-0.107 (-0.44, 0.23)
Adjusted for birth weight,
gestational age, and age at exam
Little etal. 12009)
Dallas, Texas
1980-1989 and 2002
n=191 (1980s)
n=169 (2002)
2-12 yrs old
Concurrent blood Pb
1980s: 23.6 (1.3 SE)
Mg/dL
2002:1.6 (0.2 SE) pg/dL
Blood Pb effect per 10pg/dL Pb
increase - Mean(95%CI)
Height (cm):-2.1 (-1.9, -2.3)
VNfeight (kg): -1.9 (-1.7, -2.1)
BMI (kg/m2): -0.5 (-0.4, -0.7)
Adjusted for age, age2, sex and
cohort effect
Seoul, South Korea
Date(s) not specified
n=62 boys
n= 46 girls
5-13 yrs
Concurrent blood Pb 2.4 (0.7) pg/dL
Linear model estimate (SE; p)
Height:-1.449 (0.639; p=0.026)
Total arm length:-1.804 (0.702;
p=0.012)
Body weight: -0.646 (0.718;
p=0.370)
BMI: -0.006 (0.272; p=0.982)
Adjusted for age, sex, and father's
education
Schell et al. 12009)
Albany, New York
1986-1992, 1992-1998
n=244
Maternal blood Pb
during second
trimester, third
trimester, and
delivery; Infant blood
Pb at delivery, 6
months, and 12
months
Maternal blood Pb during
second trimester 2.8 (2.6)
pg/dL, maternal blood Pb
during third trimester: 2.6
(2.2) pg/dL, maternal
blood Pb at delivery: 2.8
(2.4) pg/dL
Infant blood Pb at
delivery: 2.3 (2.7) pg/dL,
infant blood Pb at 6
months: 3.2 (3.3) pg/dL,
and infant blood Pbat 12
months: 6.3 (4.8) pg/dL
Linear model for maternal second
trimester Pb: -0.242 (p-value 0.01)
for 6-month head circumference.
When examining second trimester
maternal Pb > 3 pg/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
Adjusted for infant sex, infant birth
weight, infant nutrition, maternal
age, marital status, employment,
race, height, parity, second
trimester smoking, and education.
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Reference3
Study Location and
Years
Study Population Pb Biomarker Mean Pb (SD)
Effect Estimates
Lamb et al. (2008)
Kosovo, Yugoslavia
1985-1986
n=309 mother child pairs Maternal blood Pb
Pristina: 5.60 (1.99) pg/dL
Mitrovica: 20.56 (7.38)
Mg/dL
Regression coefficients relating
maternal blood Pb to height (95%
CI):
Pristina
Birth: 0.35 (-0.64, 1.34)
1 yr: -0.61 (-2.24,1.03)
4 yr: 0.79 (-1.71,3.29)
6.5 yr: 0.15 (-2.43, 2.74)
10 yr: -0.09 (-3.69, 3.52)
Mitrovica
Birth: 0.43 (-0.83, 1.69)
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)
10 yr: -2.87 (-6.21,0.47)
To BMI (95% CI):
Pristina
Birth:-0.14 (-0.69, 0.42)
1 yr: 0.61 (-0.28,1.50)
4 yr: 0.17 (-0.67,1.00)
6.5 yr: 0.61 (-0.09,1.30)
10 yr: -0.49 (-1.45, 0.46)
Mitrovica
Birth:-0.18 (-0.69, 0.33)
1 yr: 0.23 (-0.84,1.30)
4 yr: 0.16 (-0.66, 0.98)
6.5 yr: -0.12 (-0.90, 0.66)
10 yr: 1.31 (-0.95, 3.57)
Adjusted for sex, ethnicity, parity,
maternal height or maternal BMI,
maternal education, gestational age
at delivery, gestational age at blood
sample, and HOMES score
Zalina et al. (2008) Kuala Lumpur, Malaysia
n=269 children 6.5-8.5 yrs Concurrent blood Pb
old
n=169 urban
n=100 industrial
Industrial: 3.75 pg/dL
Urban: 3.56 pg/dL
Correlation with blood lead:
Height for age:
Urban:-0.095 (p=0.219)
Industrial:-0.037 (p=0.716)
VNfeight for age:
Urban: 0.019 (p=0.806)
Industrial: -0.063 (p=0.535)
VNfeight for height:
Urban: 0.136 (p=0.079)
Industrial: -0.069 (p=0.493)
Left arm circumference:
Urban: 0.041 (p=0.595)
Industrial: -0.055 (p=0.587)
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Reference3
Study Location and
Years
Study Population Pb Biomarker Mean Pb (SD)
Effect Estimates
Tomoum et al.
(2010)
Cairo, Egypt
Jan-Jun 2007
n=45 boys and girls
10-13 yrs old
Concurrent blood Pb 9.46 (3.08) pg/dL
Percentage of the median (SD):
Pb<10 |jg/dL
Wfeight:
Boys: 127.56 (16.26)%
Girls: 114.8(10.8)%
Height:
Boys: 98.06 (3.19)%
Girls: 96.75 (2.91)%
Pb> 10 |jg/dL
Wfeight:
Boys: 122.0(16.71)%
Girls: 123.11 (12.52)%
Height:
Boys: 99.5 (5.04)%
Girls: 100.33 (4.53)%
Adjusted for age and sex
Olivero-Verbel et al. Cartegena, Columbia
(2M)	Jun-Aug 2004
n=189 children 5-9 yrs old Concurrent blood Pb 5.49 (0.23) pg/dL
Spearman correlation coefficient (p-
value) between blood Pb and body
size: -0.224 (0.002)
*no significance in partial
correlation between blood Pb and
size when controlled for age:
-0.096 (0.189)
Liu et al. (2011a)
Guiyu, China
Chendian, China
Jan-Feb 2008
n=303
3-7 yrs old
Concurrent blood Pb
Guiyu: 13.2 (4.0-48.5)
Mg/dL
Chendian: 8.2 (0-21.3)
Mg/dL
Median (range)
Mean chest circumference in girls
was lower among those with higher
blood Pb levels (>10|jg/dL)
Mean chest and head
circumference in children > 6 years
old greater among those with
higher blood Pb levels (>10 pg/dL)
No multivariate adjustment,
stratification by age/sex
Mahram et al.
Zanjan province,, Iran
Date(s) not specified
n=42 boys
n= 39 girls
7-11 yrs
Concurrent blood Pb
Area with lead smelters:
37.0 (24.7) pg/dL
Area without lead
smelters: 15.6 (13.4)
Mg/dL
Comparison of control and study
groups Height, standardized for
age: p-value 0.52
VNfeight, standardized for age: p-
value 0.8
'Estimated Lower Limb Length
aStudies are presented in order of first appearance in the text of this section.
1	Results from recent epidemiologic studies of postnatal growth are summarized in Table
2	5-41. Afeiche et al. (2011) conducted a longitudinal study of children in Mexico City,
3	born between 1994 and 2005. Maternal bone Pb during pregnancy was associated with a
4	significant decrease in BW at age 5 years in girls but not in boys. The findings were
5	robust to additional adjustment for child's blood Pb level. In a study of children in
6	Sardinia Italy, Sanna et al. (2011) measured Pb in hair at three points in time (1997, 2002,
7	and 2007) and reported cross-sectional results from regression analyses for each of these
8	time periods. Pb in hair decreased over time and significant associations of Pb in hair
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with height were observed only in earlier time periods when hair Pb levels were relatively
high.
Ignasiak et al. (2006) studied school children aged 7-15 years living close to copper
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 et al. 1990). Hauser et al. (2008)
investigated the relationship between blood Pb and height in boys living in Chapaevsk,
Russia. 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 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 reported. A study in New York reported an 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, 6 month blood Pb levels were not
associated with 12 month measures of growth. In comparisons of changes in blood Pb
levels overtime, 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 maternal and infant blood Pb measures.
Contrary to the results summarized above, several studies do not observe associations
between blood Pb levels and impaired growth. In a prospective study of 309 mother-child
pairs from Yugoslavia, the relationship between maternal blood Pb and attained height in
children was investigated in those living in an 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 age) nor rate of height change per month (at birth-1 year, 1-4 years, 4-6.5 years,
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6.5-10 years age) was associated 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. 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 (> 1 Oj^ig/dL) and low concurrent blood Pb level groups; however, the
direction of the difference was not consistent (Liu etal. 201 la). Among girls, comparing
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).
Recent toxicological studies report significant changes in postnatal or adult body weight
after Pb 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. Blood Pb level in the
control group was 1.43 (ig/dL, in the Pb group it was 22.8 (ig/dL. Dong et al. (2009)
reported decreased body weight in adult Kunming mice after exposure to 0.6% Pb-acetate
in drinking water for 8 weeks. In contrast, Leasure et al. (2008) reported a statistically
significant inverse relationship between Pb exposure 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 (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 the lower and higher 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 (low), 55 ppm
(moderate), and 109 ppm (high) Pb in drinking water beginning which resulted in
respective blood Pb levels from 10 (ig/dL or less in the low-exposure offspring to 42
(ig/dL in the high-exposure offspring at PND10. Leasure et al. (2008) also exposed a
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separate group of mice to Pb only during the postnatal period (PND0-PND21, lactation
only exposure) and mice exposed to the same aforementioned low or high dose of Pb did
not exhibit a difference in body weight when compared to control offspring. Wang et al.
(20096) observed a statistically significant decrease in fetal body weight and body length
of Wistar rats at GD20 after maternal exposure to 0.025% 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 birthweight of litters administered 200 ppm or 400 ppm Pb-acetate
in drinking water, 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 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 0.6% 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 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
0.45% Pb-acetate (resulting in blood Pb of 263 (ig/dL at PND85) yielded the same results
(Ronis et al. 1996) with mechanistic insight showing decrements in insulin-like growth
factor 1 (IGF1) accompanying the decreases in growth rates.
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. 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.
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5.8.9
Toxicological Studies of Developmental Effects
5.8.9.1 Developmental Effects on Blood and Liver
The 1986 and 2006 Pb AQCDs 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 (TBARS) production did not change in pups at
postnatal 0, but increased statistically significantly by PND21 indicating reactive oxygen
generation. No change in acid phosphatase (ACP) activity was observed in the livers of
pups at PND0 or 21.
Masso-Gonzalez and Antonia-Garcia (2009) reported normochromic and microcytic
anemia and a significant decrease in hematocrit values and blood 8-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 TBARS and CAT activity in
the liver after Pb exposure. Intoxication with Pb also resulted in decreased liver protein
concentrations and manganese-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 during
gestation for 1 month (mean daily consumption 27 j^ig/kg). In contrast, long-term
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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 (S-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. 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 content.
5.8.9.2 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.
5.8.9.3 Developmental Effects on the Retina
The 2006 Pb AQCD 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 low (27 ppm), moderate (55 ppm), and high (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 (low), 21-24 (ig/dL (moderate), and 40-46 (ig/dL (high). Results of the study
demonstrated supernormal persistent rod photoreceptor-mediated (scotopic)
electroretinograms (ERGs) (Figure 5-46, Table 5-33) in adult rats similar to ERG
findings in male and female children after gestational exposure to low- and moderate-
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levels of Pb. Low- 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 (Figure 5-46,
and Table 5-33). High-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.
5.8.9.4 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. 2005V 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. The authors concluded that additional research
is needed on the intracellular uptake of Pb during tooth development to fully understand
the spatial distribution of Pb in teeth.
5.8.10 Summary and Causal Determination
Many epidemiologic and toxicological studies of the effects of Pb on reproductive
outcomes have been conducted since the 2006 Pb AQCD. These studies covered
outcomes such as female and male reproductive function, birth defects, spontaneous
abortions, infant mortality, preterm birth, low birth weight, and developmental effects.
There is an abundance of evidence in the literature demonstrating that Pb induces
reproductive and developmental effects in laboratory animals exposed to Pb during
gestation and/or lactation. Many of the Pb-induced effects occur in a concentration-
dependent manner and have been observed at maternal blood Pb levels that do not result
in clinical toxicity in the dams. Additionally, epidemiologic studies have demonstrated
strong evidence of an association between Pb and delayed puberty as well as decrements
to sperm/semen quality and function.
Many of the animal toxicology studies included in the 2006 Pb AQCD examined the
effect of Pb on reproduction and development at blood Pb levels greater than 40 (ig/dL, a
dose where maternal toxicity can develop during pregnancy. Data from the 2006 Pb
AQCD on male fertility showed Pb exposure or biomarkers of Pb-exposure were
associated with decrements in semen quality. Recent studies have shown the effects of Pb
exposure during early development to include disruption of endocrine function; delay in
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the onset of puberty and alteration in reproductive function later in life; and changes in
morphology or histology in sex organs and placenta. Additionally, recent epidemiologic
studies of reproductive factors among males and females investigated whether Pb
biomarker levels were associated with hormone levels, fertility, and onset of puberty.
Epidemiologic studies showed associations between blood Pb and hormone levels for
females. Studies of Pb and fertility are limited and inconsistent for females and males.
Strong and consistent associations were observed between Pb levels in males in
occupational settings with blood Pb levels as low as 20-45 j^ig/dL and sperm count and
quality.
Delayed puberty has been linked to decreased peak bone mass and increased risk of
osteoporotic fractures (Gilsanz et al.. 2011; Naves et al.. 2005). Multiple epidemiologic
studies of Pb and puberty have shown associations between concurrent blood Pb levels
and delayed pubertal development for girls and boys (evidence is stronger among girls).
These associations are consistently observed in multiple epidemiologic studies in
populations with blood Pb levels < 10(.ig/dL. Confounders considered in the
epidemiologic studies that performed regression analyses varied. Most studies controlled
for age and BMI. Other variables, such as measures of diet and SES and race/ethnicity,
were included in some of the studies. Many of the studies demonstrating positive
associations adjusted for many of the potential confounders. No patterns were detected in
the associations between Pb and puberty based on inclusion of specific confounders.
Pb-mediated changes in levels or function of reproductive and growth hormones have
been demonstrated in past and more recent toxicological studies; however the findings
are inconsistent. More data are needed to determine whether Pb exerts its toxic effects on
the reproductive system by affecting the responsiveness of the hypothalamic-pituitary-
gonad axis or by suppressing circulating hormone levels. More recent toxicological
studies suggest that oxidative stress is a major contributor to the toxic effects of Pb on
male and female reproductive systems. 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 an association between
increased generation of ROS and germ cell injury as evidenced by destruction of 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.
Overall, results of pregnancy outcomes were similar to those of the 2006 Pb AQCD;
inconsistent evidence of a relationship with Pb was available for preterm birth and little
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evidence was available to study the associations with spontaneous abortions. The 2006
Pb AQCD included a few studies that reported potential associations between Pb and
neural tube defects, but the recent epidemiologic studies found no association. Some
associations were observed between Pb and low birth weight when epidemiologic studies
used measures of maternal bone Pb or air exposures, but the associations were less
consistent when using maternal blood Pb 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). Effects of Pb exposure during early development on
toxicological studies included reduction in litter size, implantation, birth weight and
postnatal growth. Findings from epidemiologic studies of postnatal growth are
inconsistent.
Toxicological studies demonstrated that the effects of Pb exposure during early
development include impairment of retinal development and alterations in the developing
hematopoietic and hepatic systems. Negative developmental outcomes were also noted
including effects on the eyes and teeth.
Similar to toxicological and epidemiologic studies that observed Pb to be associated with
delayed puberty, delays of dynamic changes in the HPT axis are seen in the ecological
literature, i.e., delayed metamorphosis in Pb exposed frogs. Additionally, Pb exposure
has been shown to have detrimental effects on sperm, albeit often at higher blood Pb
levels in epidemiology studies but in lower doses in the toxicology literature. Again,
these findings agree with the ecological literature where Pb-dependent sperm effects are
seen in rotifers, earthworms, and trout (Sections 7.3.5.2, 7.2.4.2, and 7.3.5.3).
In conclusion, the recent toxicological and epidemiologic literature provides strong
evidence that Pb exposure is associated with effects on reproduction and development.
The weight of the evidence supports the association of Pb exposure with delayed onset of
puberty in both males and females and detrimental effects on sperm and semen quality in
occupationally-exposed males and in laboratory animals. In cross-sectional
epidemiologic studies of girls (ages 6-18 years) with mean and/or median concurrent
blood Pb levels less than 5 (ig/dL consistent associations with delayed pubertal
development (measured by age at menarche, pubic hair development, and breast
development) were observed. Toxicological studies indicate that prenatal and lactational
exposures to Pb can cause a delay in the onset of female puberty at blood Pb levels as low
as 8 (.ig/dL (lavicoli et al.. 2006b; lavicoli et al.. 2004). Recent studies show that pubertal
onset is one of the more sensitive markers of Pb exposure with effects observed after
maternal exposures leading to blood Pb levels in the pup of 3.5-13 j^ig/dL (lavicoli and
Carelli. 2007). In boys (ages 8-15 years), fewer studies were conducted but associations
were observed in most. Male animal toxicology studies have reported delayed sexual
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maturity as measured with prostate weight, among other outcomes, seeing significant
decrements at blood Pb levels of 34 (ig/dL (Sokol et al.. 1985). Additionally, Pb exposure
has been shown to have detrimental effects on sperm. These were observed in
epidemiologic studies at population mean blood Pb levels of 30 j^ig/dL and greater among
men occupationally exposed (mean blood Pb levels in study controls around 10 (ig/dL)
and in animal toxicological studies with rabbits exposed to subcutaneous Pb 3 times
per week for 15 weeks with blood Pb levels of 20 (ig/dL (Moorman et al.. 1998). The
data on preterm birth, low birth weight, spontaneous abortions, birth defects, hormonal
influences, and fecundity are less consistent between the toxicological and epidemiologic
literature. The collective body of evidence integrated across epidemiologic and
toxicological studies with a focus on the strong relationship observed with detrimental
effects on sperm and delayed pubertal onset is sufficient to conclude that there is a causal
relationship between Pb exposures and reproductive and developmental effects.
5.9 Effects on Other Organ Systems
5.9.1 Effects on the Hepatic System
Hepatotoxic effects of Pb indicated in various animal models and human populations
include alterations in hepatic metabolism, hepatic cell proliferation, changes in
cholesterol metabolism, as well as oxidative stress-related injury. Animal studies have
also shown that exposure to Pb causes a decrease in Phase I along with a simultaneous
increase in Phase II enzymes following exposure to Pb. Induction of oxidative stress by
Pb exposure is well supported by an increase in lipid peroxidation along with a decrease
in glutathione (GSH) levels and catalase (CAT), superoxide dismutase (SOD) and
glutathione peroxidase (GPx) activities.
5.9.1.1 Summary of Key Findings of the Effects on the Hepatic
System from the 2006 Lead AQCD
The 2006 Pb AQCD stated that the large 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 (CYP) P450 levels decreased following single doses of
Pb nitrate. Inhibition of induced and constitutive expression of microsomal CYP 1A1 and
1A2 was observed among various P450 isozymes. Inhibition of Phase I enzymes was
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accompanied by an increase in Phase II enzymes following exposure to Pb nitrate and
other Pb compounds, suggesting that Pb is capable of causing a biochemical phenotype
similar to hepatic nodules. Studies relating to Pb-induced hepatic hyperplasia suggested
alterations in the gluconeogenic mechanism, DNA hypomethylation along with changes
in proto-oncogene expression as well as cholesterol synthesis. Cholesterol metabolism
changes following exposure to Pb were reportedly mediated as a result of 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 in studies using inhibitors to block TNF-a activity. Other Pb-related
effects presented in the 2006 Pb AQCD include liver cell apoptosis mediated by Kupffer
cell derived signals and Pb-induced oxidative stress in vitro cell cultures. The 2006 Pb
AQCD further suggested that alterations in liver heme metabolism may involve changes
in 5-aminolevulinic acid dehydrogenase (ALAD) activity, porphyrin metabolism,
transferrin (TF) gene expression and changes in iron metabolism.
With regard to human studies, the 2006 Pb AQCD stated that nonspecific liver injury
generally observed as increases in liver enzymes in the serum was reported in adults with
occupational Pb exposure, although associations specifically with Pb exposures have not
been well established. In addition, similar to effects noted in animal studies, cytochrome
P450 activity was also suppressed in children and adults (drawn from the general
population) following exposure to Pb. The 2006 Pb AQCD reported that hepatic effects
occurred only at high Pb exposure levels (blood Pb levels >30 (ig/dL).
5.9.1.2 New Epidemiologic Studies
A few occupational epidemiologic studies examined liver biochemical parameters effects
on antioxidant status and oxidative stress resulting from exposure to Pb. However all of
these occupationally-exposed cohorts represented populations highly exposed to Pb, with
blood Pb levels ranging from 29 to 53 (ig/dL. Although the hepatotoxicity observed
within these cohorts may not be generalizable to 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).
Additionally, these studies were cross-sectional in design; thus, there is uncertainty
regarding the magnitude, timing, frequency, and duration of Pb exposure that contributed
to the observed associations.
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In one of the occupational cohorts (spray painters in western Maharashtra, India, exposed
to Pb for > 6 hours/day for 2 to 20 years) examined by Patil et al. (2007). mean (SD)
blood Pb levels in workers were 22.32 (8.87) (ig/dL. The blood Pb levels in workers were
statistically significantly higher compared to those in the concurrent control group (mean
[SD]: 12.52 [4.08] (ig/dL), who had no history of Pb exposure. Liver function enzymes,
including serum glutamic oxaloacetic transaminase (SGOT)/AST, and serum glutamic
pyruvic transaminase (SGPT)/ALT levels were statistically significantly increased in
spray painters compared to those in controls, whereas total serum protein levels were
decreased. In another occupational epidemiolgic study, Conterato et al. (In Press)
investigated liver function parameters in automotive painters exposed to Pb in Brazil.
Exposed painters had a mean (SD) blood Pb concentration of 5.4 (0.4) (ig/dL compared
to 1.5 (0.1) (ig/dL in controls. The mean (SD) duration of exposure to Pb in painters was
133.9 (14.5) months, whereas the controls were not occupationally exposed to Pb. In
exposed workers, the levels of aspartate aminotransferase (AST), but not y-
glutamyltranseferase, were increased approximately 2-fold compared to levels in controls
(p < 0.05). The activity of AST was positively correlated with blood Pb levels (r = 0.26, p
< 0.05). The authors suggested that confounding exposures to 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). Co-exposure to other environmental contaminants may also
explain the effects seen in occupationally-exposed spray-painters in Patil et al. (2007).
5.9.1.3 New Toxicological Studies
Hepatic Metabolism
As stated in the 2006 Pb AQCD, acute exposures of rodents to Pb nitrate and other Pb
compounds cause a decrease in Phase I enzymes accompanied by a simultaneous increase
in Phase II enzymes. The conclusions presented in the 2006 Pb AQCD were also
reviewed by Mudipalli (2007).
Changes in biochemical parameters, suggestive of liver damage, in undernourished male
Wistar rats (fed low-protein diet without mineral supplements) treated with 500 ppm
Pb-acetate in drinking water over a 10 month period, included decreases in serum protein
and albumin levels as well as increases in aspartate aminotransferase (AST), alanine
aminotransferase (ALT), serum alkaline phosphatase (ALP), and gamma glutamyl
transpeptidase (GGT) levels (PS et al.. 2009). In 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
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increased to approximately 110 j^ig/dL (greatly in excess of blood levels in the general
human population) by six months time and remained at this level until the termination of
exposure at 10 months. The study authors reported that similar biochemical changes were
not observed in animals treated with Pb-acetate maintained on protein-adequate, mineral
rich diet and concluded that nutritional management is important in managing Pb-related
poisoning. Mice gavaged with 50 mg/kg Pb nitrate for 40 days demonstrated similarly
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 35 mg/kg Pb via i.p. injection for 3 days (blood Pb not
reported) significantly increased the activities of ALT, AST, serum ALP, and acid
phosphatase over those in controls, whereas alkaline phosphatase activity was decreased
in Pb-treated animals. Concomitant treatment with zinc and varying levels of ascorbic
acid were observed to ameliorate the 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 injected with
50 mg/kg Pb-acetate daily for 15 days (blood Pb not reported) (Wang et al.). Swarup et
al. (2007) investigated serum biochemical changes in cows living in Pb-contaminated
environments. Serum levels of ALT, AST, alkaline phosphatase, total protein, albumin,
globulin, and A/G ratio were statistically significantly altered in cows living near Pb-Zn
smelters (mean [SD] blood Pb: 86 [6] |_ig/dL. greatly in excess of blood levels in general
human populations) compared to control cows (mean [SD] blood Pb: 7 [1] (ig/dL).
Significant positive correlations were found between blood Pb and ALT and AST,
whereas a negative correlation was observed between blood Pb and total lipids, protein,
and albumin.
Pillai et al. (2009) investigated effects on hepatic phase I and II enzymes in male and
female rats born to dams that were treated with 50 j^ig/kg Pb-acetate via subcutaneous
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 (PND55-56) to assess continuing
exposure to bioaccumulated Pb. The activities of hepatic phase I enzymes NADPH- and
NADH-cytochrome c reductase were statistically significantly reduced in Pb-exposed
male and female rats on PND56 (blood Pb not reported), compared to controls. In rats
treated with 25 j^ig/kg Pb and Cd, 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 in Pb-
treated rats on serum glutamate pyruvate dehydrogenase or ALP activities in males or
females. Histological observations in both male and female rats demonstrated fatty
degeneration, vacuolization, and pycnotic nuclei, indicating general hepatotoxicity
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following Pb treatment. In a similar study, Teijon et al. (2006) exposed Wistar rats to 200
or 400 ppm throughout gestation, lactation, and 3 months postweaning, or only 1 month
postweaning. In the animals exposed continuously throughout gestation and lactation, the
concentrations of Pb in the liver were elevated in the 200- and 400-ppm groups 1 and 3
months postweaning. 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 (ig Pb/g
tissue [0.30] versus 0.76 [0.06], respectively), but were similar between the 2 dosing
regimens at 3 months postweaning (mean [SE]: 0.054 [0.06] versus 0.55 [0.07],
respectively). ALP activity was increased at 2 weeks postweaning in animals
continuously exposed to Pb throughout gestation and lactation, whereas ALT activity was
decreased only at 2 and 3 months postweaning. In animals exposed only for 1 month
postweaning, serum ALP activity was significantly increased, although not in a
concentration-dependent manner. ALT and AST activities were not affected.
Cheng et al. (2006) studied the mechanism of Pb effects on bacterial lipopolysaccharide
(LPS)-induced TNF-a expression. A/J mice were injected via i.p with 100 (imol/kg Pb,
with or without 5 mg/kg LPS. Pb alone did not affect liver function (measured as AST or
ALT activity) or the level of TNF-a in the serum. 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 an in vitro study, the authors reported that 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) and concluded that monocytes/macrophages
(rather than hepatocytes) were primarily responsible for Pb increasing LPS- induced
TNF-a levels via the protein kinase C (PKC)/MAPK pathway. Similarly, Pb chloride
potentiated bovine serum albumin (BSA)-induced inflammation in the livers of mice
subcutaneously injected with Pb (Sa et al.. In Press).
Lipid Metabolism
In a lipid metabolism study, Ademuyiwa et al. (2009) reported that male albino Sprague
Dawley rats exposed to 200, 300 and 400 ppm Pb 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) jj.g/g,
compared to 3.24 (1.19) and 2.41 (0.31) in animals exposed to 300 or 400 ppm Pb,
respectively. Animals exposed to Pb exhibited increased hepatic cholesterogenesis at all
doses tested compared to controls. Additionally, a decrease in triglyceride was observed
at 300 and 400 ppm Pb; a decrease in phospholipid levels was observed at 400 ppm Pb.
The authors also reported positive correlations between tissue cholesterol and
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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
100 mg/kg Pb-acetate for as little as 14 days (blood Pb not reported) exhibited disorders
in lipid metabolism that were supported by increased levels of total cholesterol and
triglyceride levels in the liver tissue, whereas Sharma et al. (2010a) reported increased
liver cholesterol in mice gavaged with 50 mg/kg Pb nitrate for 40 days. Pillai et al. (2009)
observed decreases in total liver cholesterol in PND56 male and female rats that had been
exposed to 50 j^ig/kg Pb-acetate continuously throughout gestation and lactation. These
results suggest that induction of cholesterogenesis and phospholipidosis in the liver by Pb
may cause subtle effects at the cellular level that may lead to hepatotoxicity. Kojima and
Degawa (2006) examined the sex-related differences in the hepatic sterol regulatory
element binding protein-2 (SREBP-2) and 3-hydroxy-3-methylglutaryl-CoA reductase
(HMGR) gene expressions in male and female Sprague Dawley rats injected with
100 (imol/kg body weight of Pb nitrate intravenously (blood Pb not reported). The
SREBP-2 expression, which is a transcription factor for the HMGR gene, was
significantly increased in males and females with the increase occurring earlier in male
rats (6-12 hours, compared to 24-36 hours in females). In contrast, expression of the
HMGR gene, a rate limiting enzyme in cholesterol biosynthesis, was significantly
increased in both Pb-exposed males and females at earlier time frames (3-48 hours in
males; 12-48 hours in females) compared to the SREBP-2 gene expression. 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 expressions and increase in total cholesterol levels in the liver in response to Pb
occur earlier in males compared to females and also suggest that the HMGR gene
expression and increase in total cholesterol levels in the liver occur before an increase in
the SREBP-2 gene expression in either sex.
Hepatic Oxidative Stress
A number of studies pertaining to hepatic oxidative stress as a result of exposure to
various Pb compounds were identified. Adegbesan and Adenuga (2007) reported that
protein undernourished male Wistar rats injected with 100 (imol/kg Pb nitrate (blood Pb
not reported) exhibited increased lipid peroxidation, increased CAT activity, decreased
SOD activity, and increased GSH levels, compared to undernourished rats not exposed to
Pb. Increased lipid peroxidation and decreased CAT and SOD activity were also
observed when comparing undernourished Pb-exposed rats to well-nourished control rats.
Study authors concluded that malnutrition exacerbated Pb exposure effects on liver lipid
peroxidation and the involvement of free radicals in Pb toxicity. Male Foster rats treated
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with 0.025 mg/kg Pb via i.p. injection (blood Pb not reported) also exhibited statistically
significant increases in lipid peroxidation levels and decreases in SOD, CAT, and
glucose-6-phosphatase dehydrogenase (G6PD) levels in liver mitochondrial and
postmitochondrial fractions (Pandva et al.. 2010). Statistically nonsignificant decreases
were also observed in GSH levels and GPx and GR activities in Pb-treated animals. In
mice gavaged with 50 mg/kg Pb nitrate for 40 days, lipid peroxidation was increased and
SOD, CAT, and GSH were decreased compared to controls (Sharma et al.. 2010a').
Additionally, exposure to Pb nitrate resulted in histopathological changes in the structure
of the liver: hepatocytes were damaged and were marked by cytoplasmic vacuolization
and pycnotic nuclei. Yu et al. (2008) reported similar concentration-dependent increases
in lipid peroxide levels and decreases in GSH levels and CAT, SOD and GPx activities in
castrated boars that received a supplemental diet with 0, 5, 10, or 20 mg/kg Pb. 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 free radical scavenging capability, along with increased
lipid peroxidation, potentially causing serious damage to hepatic function and structure.
Khotimchenko and Kolinchenko (2007) also reported an increase in lipid peroxidation
and development of hepatitis in male albino rat liver parenchyma following treatment
with 100 mg/kg Pb-acetate for as little as 14 days. Lipid peroxidation was demonstrated
by increases in malondialdehyde (MDA) levels along with decreases in GSH and thiol
groups indicating injury in the liver antioxidant system. In another experiment, Jurczuk et
al. (2007) reported that male Wistar rats treated with 500 mg/L Pb in drinking water
(blood Pb not reported) exhibited decreases in liver vitamin E and GSH levels along with
an increase in lipid peroxidation. The study authors hypothesized that vitamin E is
involved in the mechanism of peroxidative action of Pb in the liver, and concluded that
the suggested protective role of vitamin E in the potential toxicity by Pb may be related to
scavenging of free radicals that are generated either directly or indirectly by Pb. In a
study examining the role of low molecular weight thiols on peroxidative mechanisms,
Jurczuk et al. (2006) stated that male Wistar rats treated with 500 mg/L Pb-acetate in
drinking water exhibited a decrease in blood ALAD as well as decreases in GSH and
nonprotein sulfhydryl (NPSH) levels in the liver. Metallthionein levels were also reported
to be higher in the liver following exposure to Pb. Levels of hepatic lipid peroxidation
were observed to be significantly increased in rats treated with 35 mg/kg Pb via i.p.
injection (blood Pb not reported), whereas hepatic GSH was significantly decreased
(Uoadhvav et al.. 2009). In vitro exposure of human embryonic hepatocytes (WRL-68) to
5 (.iM Pb-acetate for 30 days resulted in increase production of reactive oxygen species
(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
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lipid peroxidative damage also increased throughout the first 24 days of exposure, but
returned to normal levels by the end of the incubation period.
In a study examining the effects of Pb exposure to fetuses, Masso et al. (2007') exposed
pregnant Wistar rats with 300 mg/L Pb in drinking water starting at day 1 of pregnancy to
parturition or until weaning to determine the effects of Pb exposure in the fetal liver.
Blood Pb levels were higher at parturition (mean [SD]: 31.5 [0.80] (ig/dL) than at
weaning (mean [SD]: 22.8 [0.50] j^ig/dL). Pups exhibited liver damage that was
accompanied by an increase in thiobarbituric acid-reactive species (TBARS) production
and increased CAT activity compared to controls. In addition, increased ALP and acid
phosphatase activity was also observed. Uzbekov et al. (2007) exposed female Wistar rats
to 0.3 and 3.0 mg/L Pb nitrate for 1 and 5 months prior to, and continuing during
pregnancy, and measured fetal hepatic SOD activity on GD20. Control rats had a mean
(SD) blood Pb level of 16.1 (0.63) (ig/dL, whereas rats exposed to 0.3 and 3.0 mg/L Pb
had mean blood Pb levels that were 26.5% (20.4 j^ig/dL) and 51.8% (24.4 j^ig/dL) higher,
respectively. In the 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 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 mg/L Pb nitrate for one month suggests that 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. No effects on GSH
or MDA levels were observed in PND56 male and female rats following continuous
exposure to 50 j^ig/kg Pb-acetate throughout gestation and lactation (blood Pb not
reported) (Pillai et al.. 2009).
The studies presented above all support the possible oxidative stress impacts following
exposure to various doses of Pb administered in various forms and the potential for
hepatotoxicity as a result of oxidative stress.
Hepatic Apoptosis
Fan et al. (2009b) reported that a single i.v. injection (tail vein) of 200 (imol/kg Pb nitrate
resulted in an increase in the expression of ferritin light-chain (FLT) in rats (mean [SD]:
3.5 [1.0]-fold increase) over that in controls. Immunohistochemical analysis revealed that
hepatocytes around the central vein were heavily stained by anti-FTL antibodies, as were
nonparenchymal cells identified as Kupffer cells. The authors hypothesized that the
expression of FTL in Kupffer cells may be the result of phagocytosis of apoptotic cells:
in Pb-treated rats, apoptotic hepatocytes represented a mean (SD) 2.5 (1.4)% of total cells
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whereas only a mean (SD) 0.31 (0.31)% of hepatocytes were apoptotic in control
animals. FTL expression in Kupffer cells was not increased in rats treated with clofibrate,
which induced hepatocellular proliferation, but not apoptosis.
5.9.2 Effects on the Gastrointestinal System
Gastrointestinal effects of Pb exposure resulting in blood Pb levels ranging from 30 up to
80 (ig/dL in humans primarily include abdominal pain, constipation, and internal
paralysis. In animals, degeneration of the intestinal epithelial mucosa and a decrease in
duodenal motility has been reported following Pb exposure.
5.9.2.1 Summary of Key Findings on the Effects on the
Gastrointestinal System from the 2006 Lead AQCD
The 2006 Pb AQCD stated that a number of factors influence the gastrointestinal
absorption of Pb including the chemical and physical form of Pb, the age at Pb intake, as
well as various nutritional factors. Potential malabsorption of Pb as a result of
degeneration of the intestinal epithelial mucosa has been observed in rats exposed to Pb.
In suckling rat pups, casein micelles incidences were reported as a result of Pb present in
bovine and rat milk and in infant milk formula. Pb ingestion through water was more
toxic compared to Pb ingestion via milk. Pb ingested in milk was reported to be taken up
by the ileal tissue, whereas Pb administered intragastrically as a soluble salt was
primarily accumulated in the duodenum irrespective of vehicle used for administration.
Decreases in duodenal motility and the amplitude of contractility in the intestinal tract
were observed in rats following Pb exposure. Nutritional studies examining different
dietary levels of Pb in rats, calcium, and vitamin D indicated competition in absorption
between Pb and calcium. Dietary supplement with vitamin D led to an increase in
intestinal absorption of Pb and calcium. In instances where severe calcium 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
reported evidence that symptoms associated with gastrointestinal colic (abdominal pain,
constipation, intestinal paralysis) were prevalent in occupationally exposed adults with
blood Pb levels > 50 (ig/dL.
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5.9.2.2
New Epidemiologic Studies
The 2006 Pb AQCD reported that in humans, gastrointestinal effects generally include
abdominal pain, constipation, and internal paralysis. Kuruvilla et al. (2006) reported
gastrointestinal effects including stomach pain and gastritis along with other Pb-related
clinical manifestations in painters (mean [SD] blood Pb: 8.04 [5.04] (ig/dL)
occupationally exposed to Pb in India. Case reports involving children and individuals
occupationally exposed to Pb with very high blood Pb levels (>15 (ig/dL) were
consistent with these observations of GI pain and gastritis (Fonte et al.. 2007).
5.9.2.3 New Toxicological Studies
A few new studies pertaining to gastrointestinal effects of Pb exposure were identified.
Santos et al. (2006b) examined the impact of Pb exposure on nonadrenergic
noncholinergic (NANC) relaxations in rat gastric fundus. Male Wistar rats treated with
0.008% Pb-acetate (80 ppm) via drinking water for 15, 30, and 120 days (blood Pb not
reported) exhibited a significant difference in NANC relaxations in the gastric fundus
following electrical field stimulus (EFS). While frequency-dependent relaxations were
observed in all groups, including the control group, the relaxations were significantly
inhibited in rats treated with Pb-acetate for all three durations. When gastric fundus strips
from rats were incubated with L-nitroarginine (L-NOARG), a nitric oxide synthase
(NOS) 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 in the control group and group exposed to Pb-acetate for 120 days. Study authors
concluded that chronic 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 another study examining Pb-induced oxidative stress in the gastric mucosa, Olaleye et
al. (2007) treated Albino Wistar rats with 100 or 5,000 mg/L of Pb-acetate for 15 weeks
(blood Pb not reported). Exposure to Pb-acetate caused a significant increase in gastric
mucosal damage caused by pretreatment with acidified ethanol. Study authors reported
that though the basal gastric acid secretory rate was not altered, stomach response to
histamine was significantly higher in animals treated with 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 mg/L dose levels. In contrast, CAT, and SOD activities and nitrite
levels were significantly decreased in the gastric mucosa. Study authors concluded that
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exposure to Pb may increase the formation of gastric ulcers as a result of changes in the
oxidative metabolism in the stomach.
5.9.3 Effects on the Endocrine System
Endocrine processes that are most commonly found to be impacted by Pb exposure
include changes in the thyroid, such as changes in the thyroid stimulating hormone
(TSH), triiodothyronine (T3), and thyroxine (T4).
5.9.3.1 Summary of Key Findings of the Effects on the Endocrine
System from the 2006 Lead AQCD
The 2006 Pb AQCD reported that 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 l,25-(OH)2D3 levels. However, these effects were reported
to be observed only with blood Pb levels exceeding 30-40 (ig/dL. In addition, alterations
in calcitropic hormones were affected in children with blood Pb levels ranging from 10-
120 (ig/dL. A summary of key findings pertaining to reproductive hormones in males and
females in the current document is presented in the section on reproductive and
developmental effects (Sections 5.8.1 and 5.8.2).
5.9.3.2 New Epidemiologic Studies
Recent epidemiologic studies have reported associations between indicators of exposure
to Pb and thyroid hormone levels. In workers highly exposed to Pb (blood Pb =
71.1 (ig/dL) thyroid stimulating hormone (TSH) and free T4, but not free T3, were
increased over levels in controls (mean blood Pb level of 0.2 |_ig/dL). although these
results are most likely not generalizable to the general public due to the high blood Pb
levels of exposed workers (Pekcici et al.. 2010). Abdelouahab et al. (2008) performed a
cross-sectional study in a Canadian population characterized by high consumption of
freshwater fish. The median concurrent blood Pb level was 3.1 (ig/dL for men and
1.7 (ig/dL for women. It is important to note that the median blood Pb level for women
was lower than the limit of detection for Pb in the blood (2.1 (ig/dL), effectively meaning
that greater than 50% of women in the study had nondetectable levels of Pb in their
blood. The study authors conducted a stratified analysis and concluded that TSH levels
were negatively correlated with blood Pb in women who consumed fish contaminated
with Pb and other environmental pollutants. No associations T3 and T4 levels were
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reported in women. TSH, T3 and T4 levels were not observed to be correlated with blood
Pb in males. However, study authors stated that occupational exposure to Pb in men can
affect pituitary thyroid axis homeostasis and the relation between low-level Pb exposure
thyroid hormone homeostasis in men and women needs to be investigated further. The
authors also concluded that environmental contaminants not investigated (e.g., As) may
be influencing TSH levels. Dundar et al. (2006) examined associations of blood Pb with
thyroid function in 42 male adolescent auto repair workers exposed long term to Pb. A
control group comprising 55 healthy subjects was also used for comparison purposes.
Mean blood Pb levels were reported to be higher in the auto repair workers compared to
the control subjects (mean [SD]: 7.3 [2.92] versus 2.08 [1.24] (ig/dL). Free T4 (FT4)
levels were significantly lower in the study group compared to the control group, which
had no abnormal FT4 levels reported. In contrast, free T3 (FT3) and TSH levels were
comparable between the study and control group. Blood Pb level was reported to be
negatively correlated with FT4 levels. Based on the study outcome, the study authors
reported that long-term Pb exposures that result in the studied blood Pb levels may lead
to lower FT4 levels without impact on T3 and TSH levels in adolescents. The study
authors stated that this effect is likely secondary to the toxic effects of Pb on the pituitary-
thyroid axis and to the earlier findings of primary hypothyroidism as a result of impaired
production of peripheral thyroid hormones. Similar findings were reported by Croes et al.
(2009) in a study conducted in Belgium. Croes et al. (2009) examined the hormone levels
in 1,679 adolescents residing in nine study areas with varying exposures to multiple
industrial pollutants including Pb. The median concurrent blood Pb of the participants
from the nine different regions ranged from 1.6 to 2.8 (ig/dL. The study authors reported
that, after adjustment for potential confounding, significant interregional differences were
observed FT3 hormone levels. When individual neighborhoods were analyzed within the
larger study areas, altered levels of FT3 levels were also observed. Though varying levels
of FT3 levels were observed, the study authors reported that these changes were not
wholly due to exposure to various pollutants, including Pb that were measured in the
study and stated that other pollutants and environmental factors may also have
contributed to the effects noted. In a prospective study of 309 mother-child pairs from
Yugoslavia, the relationship between maternal TSH and T4 and blood Pb was
investigated in those living in a highly exposed town with a smelter and battery plant (n =
156 mother-child pairs) and those living in a relatively unexposed town (n = 153 mother-
child pairs) (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). Mid-pregnancy maternal free T4 levels were observed to be inversely
related with maternal blood Pb levels, but this association was not observed in the
unexposed town. In 24 newborns delivered in Tokyo, Japan, neither TSH nor free T4
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(sampled 4-6 days postpartum) was associated with cord blood Pb sampled at delivery
(mean: 0.67 (.ig/dL) (liiima et al.. 2007).
Gump et al. (2008) examined Cortisol response to acute stress in children (aged 9.5 years)
whose prenatal and postnatal blood Pb levels had been determined prior to the study at
birth (from cord blood) and at a mean (SD) age of 2.62 (1.2) years, respectively. 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. The study authors reported that blood Pb was not associated
with initial salivary Cortisol levels. However following an acute stressor, which
comprised submerging the dominant arm for a minute in a gallon of one part ice to one
part water, increasing prenatal and postnatal blood Pb levels were statistically
significantly associated with 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 significantly correlated to salivary Cortisol reactivity.
Based on these results, the study authors reported that relatively low prenatal and
postnatal blood Pb levels, notably those well below 10 j^ig/dL, can alter adrenocortical
responses of children following acute stress and the health impact and behavioral aspects
of this Pb-induced HPA deregulation in children needs to be further examined.
In another study on the impact of Pb in children, Kemp et al. (2007) examined the blood
Pb levels in 142 young, U.S. urban African-American and Hispanic children in winter
and summer to determine the seasonal increase in blood Pb and its association with
vitamin D (l,25-(OH)2D3), age and race. There was a winter/summer (W/S) increase in
blood Pb levels in children aged between 1 and 3 years (mean [SE]: 4.94 [0.45] (ig/dL in
winter, 6.54 [0.82] (ig/dL in summer), with a smaller W/S increase observed in children
aged between 4 and 8 years (mean [SE]: 3.68 [0.31] (ig/dL in winter, 4.16 [0.36] (ig/dL
in summer). Additionally, the winter and summer blood Pb levels were highly correlated
with one another. The percentage of African-American children with blood Pb levels >
10 (ig/dL increased from 12.2% in winter to 22.5% in summer. In children aged 4-8
years, the concentrations of l,25-(OH)2D3 were greater in the summer compared to the
winter (mean [SE]: 33.8 [1.1] (ig/L in summer versus 25.3 [1.2] (ig/L in winter). No
difference in seasonal l,25-(OH)2D3 was observed in children 1-3 years old.
Additionally, winter and summer concentrations of l,25-(OH)2D3 were highly correlated
(r = 0.635, p < 0.0001). There was a significant correlation between seasonal differences
in blood Pb and serum l,25-(OH)2D3 in all children and African-American children
between 4 and 8 years. Based on these results, the study authors concluded that higher
summertime increase in serum l,25-(OH)2D3 levels in children between 4 and 8 years is
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most likely due to increased sunlight-induced vitamin D synthesis and may be a
contributing factor to seasonal changes in blood Pb levels.
5.9.3.3 New Toxicological Studies
. In a study examining the effects of Pb and cadmium in adult cows reared in a polluted
environment in India, Swarup et al. (2007) stated that the mean plasma T3 and T4 levels
were significantly higher in cows near Pb and zinc smelters (mean [SD] blood Pb: 86
[6] |_ig/dL) and near closed Pb and operational zinc 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 from 269 cows examined in the study showed a
significant positive correlation between blood Pb and plasma T3 and T4 levels, whereas
the correlation between blood Pb and plasma Cortisol was nonsignificant. Mean plasma
estradiol level was significantly higher in cows near closed Pb and operational zinc
smelter industries compared to the control group. Based on these results, the study
authors concluded that endocrine profile in animals can be impacted following exposure
to Pb in polluted environments.
Biswas and Ghosh (2006) investigated the effect of Pb treatment on adrenal and male
gonadal functions in Wistar rats treated with 8.0 mg/kg Pb-acetate via i.p. injection for 21
days (blood Pb not reported). Pb treatment was observed to significantly increase 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. Spermatogenesis was decreased and the percent of
spermatid degeneration was increased in animals treated with Pb. Lastly, serum
concentrations of testosterone, FSH, and LH, were decreased in Pb-treated animals.
Supplementation with testosterone during the last 14 days of Pb treatment was observed
to ameliorate these effects.
5.9.4 Effects on Bone and Teeth
Primary effects on bone as a result of Pb exposure 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.
Similar to bone, calcium in the teeth can be easily substituted by Pb following Pb
exposure. Exposure of animals to high levels of Pb (30 mg/kg body weight) may result in
the formation of "Pb line" and Pb can also cause a decrease in cell proliferation,
procollagen type I production, intracellular protein, and osteocalcin in human dental pulp
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cell cultures. Accumulation of Pb was also associated with tooth loss and higher
incidence of periodontitis.
5.9.4.1 Summary of Key Findings of the Effects on Bone and
Teeth from the 2006 Lead AQCD
The 2006 Pb AQCD reported many effects on bone and some in teeth in animals and
following exposure to Pb. Pb easily substituted for calcium in bone and was taken up by
the bone causing changes in bone cell function. Exposure of animals to Pb during
gestation and immediate postnatal period was reported to significantly depress early bone
growth with the effects showing a concentration-dependent trends,. In mature animals,
long-term exposure (up to one year) to Pb, along with poor nutrition (low calcium)
affected bone growth as well 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 l,2-[OH]2D3 and impact on calcium binding proteins and increases in
calcium 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 (EGF), transforming
growth factor-beta 1 (TGF-(31), and parathyroid hormone-related protein (PTHrP).
As in bone, Pb can easily substitute for calcium in the teeth and is taken and incorporated
into developing teeth in experimental animals. Since teeth do not undergo remodeling
like the bone does during growth, most of the Pb in the teeth remains in a state of
permanent storage. High dose exposure of Pb to animals (30 mg/kg body weight) has
lead to 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 are easily
evident in the enamel and the dentin within these "Pb lines." Pb has also been shown to
decrease cell proliferation, procollagen type I production, intracellular protein, and
osteocalcin in human dental pulp cell cultures. Adult rats exposed to Pb have exhibited an
inhibition of the posteruptive enamel proteinases, delayed teeth eruption times, as well as
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 enamel
areas that were calcium deficient. The 2006 Pb AQCD also reported that a number of
epidemiologic and animal studies have both separately suggested that Pb is a caries-
promoting element.
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5.9.4.2 New Toxicological and Epidemiologic Studies
As reported in the 2006 Pb AQCD, Pb appears to be capable of causing effects in bones
of humans and animals following exposure. The association between blood Pb levels and
osteoporosis was examined in several epidemiologic studies. Most studies were cross-
sectional in design; thus, there is uncertainty regarding the magnitude, timing, frequency,
and duration of Pb exposure that contributed to the observed associations. Campbell and
Auigner (2007) examined subjects >50 years of age using NHANES III for association
between concurrent blood Pb level and osteoporosis. The study authors used the bone
mineral density in the hip as the primary outcome in groups comprising non-Hispanic
white men (mean blood Pb: 4.9, range: 0.7 to 48.1 (ig/dL), non-Hispanic white women
(mean blood Pb: 3.6, range: 0.7 to 28.7 (ig/dL), African-American men (mean blood Pb:
7.7, range: 0.7 to 52.9 (ig/dL), and African-American women (mean blood Pb: 4.5, range:
0.7 to 23.3 (ig/dL). The results indicated that the adjusted mean total hip bone mineral
density in the non-Hispanic white males who had the lowest blood Pb levels (actual
concentration not reported) was statistically significantly higher than that in the males
with higher blood Pb levels. Similar associations, although not statistically significant,
were reported among white females. Likely due to the small sample size, similar results
were not observed among African-American men and women. No association was
observed between blood Pb and osteoporotic fractures in any sex/race group. Since the
NHANES study comprised a cross-sectional design, no inferences could be made
regarding the temporal sequence of the observed association. The study authors
concluded that further inquiry was needed to study the possible causal association
between Pb exposure and osteoporosis. In a similar study, Sun et al. (2008b) examined
the association between concurrent blood Pb levels and osteoporosis in 155 males and 37
females in China occupationally-exposed to Pb (mean blood Pb: 20.22 and 15.5 (ig/dL,
respectively). Bone mineral density was reported to be statistically significantly lower in
exposed females compared to exposed males. When all participants (including 36 male
and 21 female unexposed controls) were divided into groups according to blood Pb and
urinary Pb levels, the study authors reported that there were significant decreases in bone
mineral density in groups that had high urinary Pb levels (> 5 jj.g/g creatinine) compared
to groups with low urinary Pb in both sexes. In contrast, a significant difference was
observed between blood Pb and bone mineral density only in males with blood Pb
>30 (ig/dL. Prevalence of osteoporosis was reported to increase significantly with
increasing blood Pb in a linear manner. Khalil et al. (2008) reported similar associations
between blood Pb level and osteoporosis in older women. The study authors conducted a
prospective study using 533 women aged 65-87 years with a mean (SD) blood Pb of 5.3
(2.3) (ig/dL to determine the association between blood Pb and recurring fractures.
Analysis of bone mineral density was conducted in 1986-1990 and 1993-1994, while
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blood Pb was measured from 1990-1991. The bone mineral density was 7% lower in the
total hip (p <0.02) and 5% lower in the femoral neck (p <0.03) in the highest blood Pb
group (> 8 (ig/dL) compared to the lowest blood Pb group (< 3 (ig/dL). The trend across
all dose groups was also observed to be statistically significant for hip and femoral neck
bone mineral density. In addition, hip, femoral neck, and calcaneus bone loss was
observed to be greater in the medium (blood Pb: 4-7 j^ig/dL) and high Pb groups
compared to the low Pb group, but the observed trend was only significant for calcaneus
bone loss. Multivariate analysis indicated that women with high blood Pb levels had an
increased risk of non-spine fracture and women with medium or high blood Pb levels had
a higher risk of falls compared to the low blood Pb level group. Based on these results,
the study authors concluded that blood Pb is associated with an increased risk of falls and
fractures leading to osteoporosis-related fractures.
To examine the association between biomarkers of joint tissue metabolism and blood Pb
levels, Nelson et al. (In Press) performed a cross-sectional analysis of 329 male and 342
female participants in the Johnson County Osteoarthritis Project Metals Exposure Sub-
study. In women (mean age = 62 years), the median concurrent blood Pb level was
1.9 |_ig/dL (range: 0.5-25.4 (ig/dL). Blood Pb levels were higher in African-American
compared to Caucasian women. Unadjusted correlation analyses demonstrated significant
positive correlations between blood Pb and uNTX-I (a marker of bone
resorption/turnover), uCTX-II (a marker associated with the progression of radiographic
knee and hip osteoarthritis), and COMP (a cartilage biomarker related to osteroarthritis);
however, significant associations only remained for uNTX-I and uCTX-II after adjusting
for age, BMI, race, and smoking status. In men (median age = 65), the median blood Pb
level was 2.2 (ig/dL (range: 0.5-25.1 (ig/dL). As with women, African-American men
had higher blood Pb levels than did Caucasian men. In unadjusted correlation analyses,
blood Pb levels in men were positively associated with uCTX-II, COMP, and C2C:CPII
ratio (an indication of the balance between cartilage collagen degradation and synthesis),
and negatively associated with CPU (a marker of collagen synthesis). After controlling
for age, BMI, race, and smoking, only the positive association between blood Pb and
COMP remained borderline statistically significant in men. The authors concluded that
blood Pb is associated with bone turnover and mineralized cartilage turnover in women
and non-mineralized cartilage turnover in men; however, as this study was cross-sectional
in nature, it is impossible to conclude whether increased cartilage turnover is a product of
increased blood Pb, or whether cartilage turnover itself results in increased Pb. Similarly,
Machida et al. (2009) investigated bone matrix turnover rates in Japanese women related
to menopause status and blood Pb. Perimenopausal women (n = 319) were observed to
have significantly higher geometric mean blood Pb (2.0 (ig/dL), than did premenopausal
women (n = 261, blood Pb = 1.6 (ig/dL), younger postmenopausal women (n = 397,
blood Pb = 1.8 (ig/dL), or older postmenopausal women (n = 248, blood Pb = 1.7 (ig/dL).
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In all subjects and perimenopausal women, markers of bone matrix turnover (osteocalcin
[OC], bone-specific alkaline phosphatase [BALP], and N-telopeptide cross-linked
collagen type I [NTx]) were significantly positively associated with blood Pb in
unadjusted correlation analyses. In multivariate regression models controlling for bone-
mineral density, NTx, and age, OC was additionally statistically positively associated
with blood Pb in all subjects and perimenopausal women. Bone mineral density and NTx
were also positively associated with blood Pb in these models for all subjects. As with
Nelson et al. (In Press), the cross-sectional nature of this study precludes a determination
whether higher blood Pb is cause of increased bone matrix turnover biomarkers, or a
consequence of increased bone turn over.
To understand the importance of bone as a target tissue of Pb toxicity, Jang et al. (2008)
studied the effect of Pb on calcium release activated calcium influx (CRACI) using
primary cultures of human osteoblast-like cells (OLC). When cells were incubated with
1,000 or 3,000 (.iM Pb, a concentration-dependent impact on the CRACI was observed, as
was a concentration-dependent increase in the influx of Pb into human OLC. These
results suggest that Pb interferes with CRACI in human OLCs by initiating the CRACI
(i.e., the measurable influx of calcium upon re-addition of calcium is partially inhibited
by Pb) and the influx of Pb is enhanced after CRACI is induced. Since studies have found
associations between higher blood Pb level and reduced skeletal growth in children,
Zuscik et al. (2007) conducted a study using murine limb bud mesenchymal cells (MSCs)
to test the hypothesis that Pb alters chondrogenic commitment of mesenchymal cells and
also to assess the effects of Pb on various signaling pathways. Exposure to 1 (.iM Pb
caused increased basal and TGF-(3/BMP induction of chondrogenesis in MSCs which was
supported by nodule formation and upregulation of Sox-9, type 2 collagen, and aggrecan
which are all key markers of chondrogenesis. The study authors also observed enhanced
chondrogenesis during ectopic bone formation in mice that had been pre-exposed to Pb in
drinking water (55 or 233 ppm, corresponding to 14 or 40 (ig/dL blood Pb). MSCs
exposed to Pb exhibited an increase in TGF-(3, but BMP-2 signaling was inhibited. Pb
was also reported to induce NF-kB and inhibit AP-1 signaling. Based on these results, the
study authors concluded that chondrogenesis following exposure to Pb 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 epidemiologic studies. Since
individuals may be impacted by the release of Pb stored in their skeletal compartments,
Arora et al. (2009) examined the association between bone Pb concentrations and loss of
natural teeth in 333 male participants of the NAS. Tooth loss in men was categorized as
0, 1-8 or > 9. Individuals with > 9 teeth missing had significantly higher tibia and patella
Pb concentrations (measured within 3 years of dental assessment) compared to those with
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no tooth loss; no significant difference in blood Pb levels (measured within 3 years of
dental assessment lYHuetal. 1996a) I) was observed between the categories of teeth loss.
Following adjustment for age, education, smoking status, pack-years of smoking, and
diabetes, men with the highest tibia Pb concentrations (>23 jj.g/g) had higher odds of
tooth loss (OR: 3.03 [95% CI: 1.60, 5.75]) compared to men with tibia Pb < 15 jj.g/g.
Men with the highest patellar Pb (>36 jj.g/g) also had higher odds of tooth loss (OR: 2.41
[95% CI: 1.30, 4.49]) compared to men with patellar Pb < 22.0 jj.g/g. Tooth loss was not
statistically associated with blood Pb levels. Based on these results, the study authors
concluded that long-term cumulative exposure to Pb is associated with increased odds of
tooth loss. In a study examining the effects of Pb exposure on periodontitis in the U.S,
Saraiva et al. (2007) analyzed data for 2,500 men and 2,399 women aged between 30 and
56 years from NHANES III. The analysis took into account various covariates including
age, NHANESIII phase, cotinine levels, poverty ration, race/ethnicity, education, bone
mineral density, diabetes, calcium intake, dental visits, and menopause in women. After
adjusting for these covariates and comparing individuals with a concurrent blood Pb level
of >7 (ig/dL to those with a blood Pb level of <3 (ig/dL, the prevalence ratios of
periodontitis was 1.70 (95% CI: 1.02, 2.85) for men and 3.80 (95% CI: 1.66, 8.73) for
women. Based on these results, the study authors concluded that there was a positive
association between periodontitis and blood Pb levels for both men and women. In a
similar study, Yetkin et al. (2007) recruited 60 male subjects (30 apprentices with Pb
exposure, 30 controls), to examine the impact of occupational exposure to Pb on
periodontal status and association between periodontitis and blood Pb or oxidative stress.
The results of their analysis indicated that blood Pb was significantly higher in
apprentices exposed to Pb compared to controls (mean [SD]: 7.38 [4.41] versus 2.27
[1.49] (ig/dL, respectively). No clinical periodontal or oxidative stress parameters were
significantly different between apprentices and controls. While the correlation between
blood Pb and periodontal parameters was not reported, significant correlations between
plaque index and CAT, probing depth and SOD, clinical attachment level and SOD, and
clinical attachment level and malondialdehyde in Pb-exposed apprentices were observed.
These results demonstrate that there is significant association between clinical
periodontal parameters and oxidative stress/damage indices in Pb-exposed apprentices. In
a multiple regression analysis, a statistically significant association between gingival
index and working status, family income and either probing depth or clinical attachment
level was noted.
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5.9.5
Effects on Ocular Health
Ocular effects most commonly indicated to be associated with exposure to Pb include
formation of cataract, impaired vision, edema and retinal stippling.
5.9.5.1 Summary of Key Findings of the Effects on Ocular Health
from the 2006 Lead AQCD
The 2006 Pb AQCD stated that various changes in the visual system were observed with
Pb poisoning including retinal stippling and edema, cataract, ocular muscle paralysis and
impaired vision. The 2006 Pb AQCD reported that retinal responses were observed in
children of mothers with a blood Pb range of 10.5 to 32.5 (ig/dL during pregnancy, while
cataracts were noted in middle-aged male with tibia bone Pb levels of 31-126 jxg/g.
5.9.5.2 New Toxicological and Epidemiologic Studies
A small number of human studies pertaining to ocular effects of Pb were identified.
Mosad et al. (2010) studied the association between subcapsular cataract and Pb,
cadmium, vitamin C, vitamin E, and beta carotene blood levels in middle-aged male
smokers compared to nonsmokers. Blood Pb was statistically significantly elevated in
light (mean [SD]: 14.5 [0.41] (ig/dL), moderate (14.5 [0.41] (ig/dL), and heavy smokers
(18.7 [1.24] (ig/dL) compared to nonsmokers (12.2 [0.21] (ig/dL). Blood Pb
concentrations were also observed to be statistically higher in the cataracts of smokers
versus nonsmokers. Similar associations were also observed for cadmium blood and lens
levels, while vitamins C, E, and beta carotene levels were significantly decreased in
smokers. Based on these results, the study authors concluded that the Pb and cadmium
present in high concentration in smokers were associated with cataracts due to oxidative
stress which was indicated by reduced levels of antioxidants such as vitamins C, and E
and beta carotene. Erie et al. (2009) investigated the association between age-related
macular degeneration and Pb and Cd in retinal tissue of human eye donors. The authors
observed that Pb, but not Cd, was significantly elevated in the neural retina tissue of
donors with age-related macular degeneration (n = 36 donors, 72 eyes; median [IQR]:
12.0 [8-18] ng/g Pb) versus normal control donors (n = 25 donors, 50 eyes; median
[IQR]: 8.0 [0-11] ng/g Pb). Neither heavy metal was significantly elevated in the retinal
pigment epithelium-choroid complex in donors with age-related macular degeneration
and normal controls.
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New animal studies pertaining to the ocular effects of Pb have investigated endpoints
such as retinal progenitor cell proliferation and neurogenesis,(Section 5.3.4.3). An in
vitro study was found that investigated whether exposure of cultured lenses from rats (4-6
weeks age) to 1 (.iM Pb nitrate increased opacity with or without secondary oxidative
challenge (Neal et al.. 2010b). Lenses incubated for 3 days in the presence of 1 (.iM Pb
were transparent, with no difference in amino acid incorporation compared to control
lenses. The authors concluded this indicated that short-term Pb exposure does not induce
osmotic swelling or lens shrinkage. However, exposure to lenses to Pb for 5 days
dramatically decreased the percentage of transparent lenses (30%) compared to controls
(80%). In Pb-exposed lenses, 30% displayed "definite cataracts" compared to only 2.5%
in control lenses. By culture day 8, all exposed lenses were described either as clearly
opaque or definite cataracts, while only 7% 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 opacity in Pb-
exposed lenses at culture day 7, compared to less than 20% in control cells. Exposure to
Pb additionally altered epithelial nutrient transport and lens histology relative to that in
controls.
5.9.6 Effects on the Respiratory System
The collective body of toxicological and epidemiologic studies demonstrates Pb-
associated effects on multiple immunological pathways, including a shift from a Thl to a
Th2 phenotype, increased IgE antibody production, and increased inflammatory
responses (Sections 5.2.5.1. and 5.6). These are well recognized pathways that contribute
to increased susceptibility to infections and also to the development of respiratory
diseases such as asthma. Recent investigation of the respiratory effects of Pb exposure
has been limited; however, cross-sectional studies have indicated an association of
increasing blood Pb level with increased prevalence of asthma in children
(Section 5.6.4.2). As described in Section 5.2.4, Pb 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.
Studies investigating these airway responses also are limited in number and collectively
do not provide strong evidence of an association with blood Pb (Section 5.6.6).
Collectively, panel and time-series epidemiologic studies demonstrate associations
between 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
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(Section 5.6.4.3). Toxicological studies have found pulmonary inflammation induced by
CAPs in which Pb was one of numerous components (Wei et al.. 2011; Duval 1 et al..
2008; Godleski et al.. 2002; Saldiva et al. 2002). Despite this evidence for respiratory
effects related to air-Pb concentrations, it is important to note the limitations of air-Pb
studies, including the limited data on the size distribution of Pb-PM (Section 3.5.3), the
uncertain relationships of Pb-PMi0 and Pb-PM2 5 with blood Pb levels, and the lack of
adjustment for other correlated PM chemical components.
5.9.7 Summary
There is evidence from epidemiologic and toxicological studies that substantial exposure
to Pb can result in altered liver function and hepatic toxicity. Biochemical changes
indicative of liver injury, including decreases in serum protein and albumin levels and
increased AST, ALT, ALP, and GGT activities, have been observed in occupationally-
exposed humans (blood Pb > 22 j^ig/dL) (Can et al.. 2008; Khan et al.. 2008; Patil et al..
2007)	mature animals exposed to high levels of Pb during adulthood (Sharma etal..
2010a; Wang et al.; PS et al.. 2009; Cheng et al.. 2006). and animals exposed during
gestation and lactation (Pillai et al.. 2009; Teiion et al.. 2006). In humans with mean
blood Pb levels 5.4 |_ig/dL. altered AST levels were observed, but increases may be
related to exposure to contaminants other than Pb (Conterato et al.. In Press). Increased
hepatic cholesterogenesis, altered triglyceride and phospholipid levels, and disorders in
lipid metabolism accompanied by increased levels of total cholesterol and triglycerides
have been reported in the animal literature (Ademuviwa et al.. 2009; Khotimchenko and
Kolenchenko. 2007). These results suggest that induction of cholesterogenesis and
phospholipids in the liver may cause subtle effects at the cellular level, leading to hepatic
injury. Multiple studies in humans and animals have observed hepatic oxidative stress,
generally indicated by an increase in lipid peroxidation along with a decrease in GSH
levels and CAT, SOD, and GPx activities following exposure to Pb (Pandva et al.. 2010;
Sharma et al.. 2010a; Khan et al.. 2008; Yu et al. 2008; Adegbesan and Adenuga. 2007;
Jurcziik et al.. 2007; Khotimchenko and Kolenchenko. 2007; Jurcziik et al.. 2006).
Indices of increased oxidative stress were additionally observed in the livers of fetuses
exposed to Pb throughout gestation (Masso et al.. 2007).
Relatively few human studies have been conducted on the gastrointestinal toxicity of Pb
since the completion of the 2006 Pb AQCD. A case study reporting on GI symptoms in a
child reported that elevated blood Pb was associated with nonlocalized abdominal pain,
vomiting, nausea, constipation, lack of appetite, fatigue, and headaches (C'abb et al..
2008).	Symptoms were reported to diminish following cessation of exposure. Similar GI
symptoms (stomach pain and gastritis) were observed in battery works and painters
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exposed to Pb in India with blood Pb levels ranging from 0.4-116.6 (ig/dL (Kuruvilla et
al.. 2006). Toxicological evidence for Pb-induced GI effects in rats includes inhibition of
NANC relaxations in the gastric fundus and the observation of oxidative stress (lipid
peroxidation, decreased SOD and CAT) in the gastric mucosa (Olaleve et al.. 2007;
Santos et al. 2006b'). The observation of oxidative stress was accompanied gastric
mucosal damage.
The endocrine processes most impacted by exposure to Pb include changes in thyroid
function. FT4, but not FT3, was decreased in adolescent male auto repair workers
(Dundar et al.. 2006) and mid-pregnancy T4 levels were negatively associated in women
living in a highly contaminated town in Yugoslavia, whereas this association was not
observed in a non-contaminated town (Lamb et al. 2008). Toxicological evidence for
similar effects was found in adult cows reared in an environment contaminated with Pb.
A positive correlation was reported between blood Pb and plasma T3 and T4 levels
(Swarup et al. 2007). In children challenged with an acute stressor, increasing prenatal
maternal and age 2-year blood Pb levels were associated with significant increases in
salivary Cortisol responses at age 9 years, even with blood Pb levels less than 10 (ig/dL
(Gump et al.. 2008). A summary of key findings pertaining to reproductive hormones in
males and females in the current document is presented in the section on reproductive and
developmental effects (Sections 5.8.1 and 5.8.2).
Numerous epidemiologic studies investigated the association between Pb biomarkers and
osteoporosis in adults. Higher blood Pb was observed to be associated with decreased
bone mineral density in non-Hispanic white males (Campbell and Auinger. 2007).
whereas urinary Pb, but not blood Pb, was associated with decreased bone mineral
density in Chinese individuals occupationally exposed to Pb (Sun et al.. 2008b). In
elderly women, blood Pb levels were positively associated with risk of falls and
osteoporosis-related fractures (khalil et al. 2008). Linear skeletal growth was reduced in
children living near copper smelters and refiners (concurrent mean blood Pb level =
7.7 (ig/dL) (Ignasiak et al.. 2006). In vitro studies indicate that Pb interferes with CARCI
in human OLCs and that Pb perturbs multiple signaling pathways during murine limb bud
growth, potentially resulting in altered skeletal development (Jang et al.. 2008; Zuscik et
al.. 2007). Blood Pb levels have also been shown to be related to biomarkers of joint
tissue metabolism in elderly populations, but the cross-sectional nature of these analyses
prevents conclusions being drawn on whether Pb increases bone and joint turnover
biomarkers, or whether increased turnover releases Pb into the bloodstream (Machida et
al.. 2009; Nelson et al. In Press). Epidemiologic studies investigating Pb exposure and
tooth loss indicate that long-term, cumulative exposure to Pb is associated with increased
odds of tooth loss, periodontitis in men and women, and that periodontitis is associated
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with oxidative stress/damage in individuals exposed in an occupational setting (Arora et
al.. 2009; Saraiva et al.. 2007; Yetkin-Av et al.. 2007V
New toxicology studies have reported ocular effects (i.e., retinal progenitor cell
proliferation) due to Pb exposure (Section 5.3.4.3). and studies in humans report
associations between heavy smoking, increased blood Pb levels, and cataracts (Mosad et
al.. 2010) and retinal Pb concentrations and age-related macular degeneration (Erie et al..
2009). Investigation of the respiratory effects of Pb exposure has been limited; however,
cross-sectional studies have indicated an association of higher blood Pb levels with
increased prevalence asthma in children (Section 5.6.4.2).
In summary, recent toxicological and epidemiologic evidence regarding the effects of Pb
exposure on the liver, GI tract, endocrine system, bone and teeth, eyes, and respiratory
tract largely are supportive of those effects noted in the 2006 Pb AQCD. However, recent
evidence of these effects is relatively limited, and therefore no causal determinations are
made regarding Pb-induced effects in these organ systems.
5.10 Cancer
The previous epidemiologic studies included in the 2006 Pb AQCD (U.S. EPA. 2006^
"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) recently classified inorganic
Pb compounds as probable human carcinogens (Group 2A of IARC classifications) based
on stronger evidence in animal studies than human studies, and organic Pb compounds as
not classifiable (Group 3 of IARC classifications) (IARC. 2006a; Rousseau et al. 2005V
Additionally, the National Toxicology Program has listed Pb and Pb compounds as
"reasonably anticipated to be human carcinogens" (NTP. 2004). The typical cancer
bioassays used 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
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addition, recent studies of Pb and DNA and cellular damage, as well as epigenetics
studies, 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 measures of Pb are used including Pb measured in blood and bone. Bone Pb is
indicative of cumulative Pb exposure. Blood Pb can represent more recent exposure,
although it can also represent remobilized Pb occurring during times of bone remodeling.
Toxicological studies only report exposure by blood Pb or exposure dose. More detailed
discussion of these measures is given in Section 4.3.5.
5.10.1 Cancer Incidence and Mortality
Recent studies have included epidemiologic evaluations of the associations between Pb
and both specific cancers, such as lung cancer and brain cancer, and overall cancer. Table
5-42 provides an overview of the study characteristics and results for the epidemiologic
studies that reported effect estimates. This section also evaluates toxicological evidence
on the potential carcinogenicity of Pb.
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Table 5-42
Summary of recent epidemiologic studies9 of cancer incidence and
mortality
Reference8
Study
Location
Cancer Measure of Pb
Outcome Study Population Exposure
Mean Pb(SD)
Adjusted Effect Estimates
Cancer Mortality:
Menke et al.
(2006)
Multiple U.S. Overall
locations cancer
mortality
NHANES III cohort Blood Pb at
with Blood Pb baseline
measures in 1988-
1994
At least 12 years of
follow-up
Blood Pb <10 pg/dL
2.58 pg/dL
(geometric mean)
Tertile 1
Tertile 2
Tertile 3
Tertile 1: <1.93 pg/dL
Tertile 2:
1.94-3.62 pg/dL
Tertile 3: > 3.63
pg/dL
1.00
0.72 (95% CI: 0.46, 1.12)
1.10 (95% CI: 0.82, 1.47)
Schober et al.
Multiple U.S. Overall
locations cancer
mortality
NHANES III cohort
At least 40 years of
age
Blood Pb at
baseline
Blood Pb<5 pg/dL:
67.7%
Blood Pb 5-9 pg/dL:
26.0%
Blood Pb>=10 pg/dL:
6.3%
Blood Pb<5 pg/dL: 1.00
Blood Pb 5-9 pg/dL: 1.44 (95% CI: 1.12,1.86)
Blood Pb> 10 pg/dL: 1.69 (95% CI: 1.14, 2.52)
Note: Modification by age assessed and
associations varied slightly
VNfeisskopf et al. Boston, MA
Overall
cancer
mortality
NAS
Included men only
Mean follow-up
period for this study:
8.9 yr
Blood Pb at
baseline,
Patella Pb at
baseline
Blood Pb: 5.6 pg/dL
(3.4)
Blood Pb Tertile 1
Blood Pb Tertile 2
Blood Pb Tertile 3
Tertile 1 of Blood
Pb:



<4 pg/dL




Pb:
Patella
Pb Tertile 1
1.00
Tertile 2 of Blood
4-6 pg/dL

Patella
Pb Tertile 2
0.82
Tertile 3 of Blood
Pb:
Patella
Pb Tertile 3
0.32
>6 pg/dL




1.00
1.03 (95% CI: 0.42, 2.55)
0.53 (95% CI: 0.20, 1.39)
Tertile 1 of patella
Pb: <22pg/g
Tertile 2 of patella
Pb: 22-35pg/g
Tertile 3 of patella
Pb: >35 pg/g
Khalil et al.
(2009a)
Baltimore,
MD, and
Monongahela
Valley, PA
Overall
cancer
mortality
Subgroup of the
Study of
Osteoporotic
Fractures cohort
Included white
women aged 65-87;
12 yr (+/- 3 yr)
follow-up
Blood Pb at
baseline
Blood Pb Level
5.3 (2.3) pg/dL
Blood Pb<8 pg/dL: 1.00
Blood Pb> 8 pg/dL: 1.64 (95% CI: 0.73, 3.71)
Lung Cancer:
Lundstrom et al. Sweden
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
Median peak
blood Pb level
Median number
of yr with at least
one blood sample
obtained
Median
cumulative blood
Pb index (sum of
annual blood Pb
Level)
Median peak blood
Pb Level: cases 2.4
pmol/L, controls 2.7
pmol/L
Median number of yr
with at least one
blood sample
obtained: cases 4.5
yr, controls 6.0 yr
Median cumulative
blood Pb index:
cases 9.0 pmol/Pb,
controls 11.9
pmol/Pb
Median peak blood Pb Level: 1.00 (95% CI:
0.71, 1.42)
Median number of yr with at least one blood
sample obtained: 0.96 (95% CI: 0.91,1.02)
per pmol/L
Median cumulative Blood Pb index: 0.99 (95%
CI: 0.96,1.02) per pmol/L
Note: similar results were observed when
restricted to smokers only
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Reference8
Study
Location
Cancer
Outcome
Study Population
Measure of Pb
Exposure
Mean Pb(SD)
Adjusted Effect Estimates
Jones etal.
(2007)
Humberside, Lung cancer Male tin smelter
UK	mortality employees
Personnel record NA
cards and air
sampling
conducted from
1972-1991
Three exposure
scenarios
determined for
working lifetime
cumulative
exposure - all
have similar
medians of
approximately 2
mg yr/m3
RR for Pb exposure weighted age and time
since exposure: 1.54 (90% CI: 1.14, 2.08)
Note: Similar results for other exposure
determination scenarios.
Rousseau et al. Montreal, Lung cancer	Men aged 35-79 Interview of job
(2007) Canada and other	history and
cancer	exposure matrix
incidence
Ever exposed to:
Organic Pb 3.0%
Inorganic Pb 17.0C
Pb in gasoline
emissions 38.6%
Organic Pb exposure compared to no
exposure:
Lung 1.3 (95% CI: 0.5, 3.1)
Inorganic Pb exposure compared to no
exposure:
Lung 1.1 (95% CI: 0.7,1.7)
Pb in gasoline emissions exposure compared
to no exposure:
Lung 0.8 (95% CI: 0.6,1.1)
Note: results are for comparisons using
population-based controls; results for controls
with other types of cancers were similar
Brain Cancer:
van Wijngaarden
and Dosemeci
Multiple U.S
locations
Brain cancer
mortality
National
Longitudinal
Mortality Study -
included individuals
with occupational
information
-included follow-up
from 1970-1989
Interview about NA
current or most
recent job within
the past 5 years
and a job
exposure matrix
Any Pb exposure compared to no exposure
1.56 (95% CI: 1.00, 2.43)
Note: HRs were greatest among those with
high probabilities of exposure and
medium/high exposure intensity
Rajaraman et al.
Phoenix, AZ,
Boston, MA,
and
Pittsburgh,
PA
Brain cancer
incidence
NCI Brain Tumor
Study
- included
individuals >=18 yr
diagnosed with brain
cancer less than 8
wk before
hospitalization;
frequency-matched
controls were
individuals admitted
to the same
hospitals for non-
neoplastic
conditions
Interviews of
lifetime work
history and
exposure
databases
NA
Meningioma:
Ever exposure to Pb 0.8 (95% CI: 0.5,1.3)
Glioma: Ever exposure to Pb 0.8 (95% CI: 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
genotypes; these associations were not
observed for glioma incidence
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Reference8
Study
Location
Cancer
Outcome
Measure of Pb
Study Population Exposure
Mean Pb(SD)
Adjusted Effect Estimates
Bhatti et al.
Phoenix, AZ,
Boston, MA,
and
Pittsburgh,
PA
Brain cancer
incidence
NCI Brain Tumor
Study
- included non-
Hispanic whites >
18 yr diagnosed
with brain cancer
less than 8 wk
before
hospitalization;
frequency-matched
controls were
individuals admitted
to the same
hospitals for non-
neoplastic
conditions
Interviews of
lifetime work
history and
exposure
databases
Glioma: 70.5 pg/m y Per 100 pg/m y increase in cumulative Pb
(193.8 pg/m3y) exposure
Glioblastoma
multiform: 97.5
pg/m3y (233.9
pg/m3y)
Glioma: 1.0 (95% CI: 0.9,1.1)
Glioblastoma multiform: 1.0 (95% CI 0.9,1.1)
Meningioma: 101.1 Meningioma: 1.1 (95% CI: 1.0,1.2)
pg/m3y (408.7
pg/m y)
Controls: 69.7
pg/m3y (248.8
pg/m3y)
Note: modification by SNPswas conducted
and associations varied by SNP
Breast Cancer:
Pan et al.
(2011)
Canada
Breast
cancer
incidence
National Enhanced
Cancer Surveillance
System (NECSS) -
population-based
sample of cancer
cases and controls
with information
collected from 1994-
1997
Self-reported
previous
addresses and
their proximity to
Pb smelters
(determined
using
Environmental
Quality Database
[EQDB])
NA
Residing >3.2 km from Pb smelter or no
nearby smelter: 1.00
Residing 0.8-3.2 km from Pb smelter: 0.41
(95% CI: 0.11,1.51)
Residing <0.8 km from Pb smelter: 0.61 (95%
CI: 0.11, 3.42)
Multiple and Other Cancers:
Absalon and
Slesak (2010)
Silesia
province,
Poland
Overall
cancer
incidence
Children living in this
province at least five
years
Pb-related air
pollution
measures
NA
Reported correlations between changes in Pb
and cancer incidence - no/low correlations
observed (correlation coefficients between -0.3
and 0.2)
Obhodas et al.
(2007)
Island of Krk,
Croatia
Incidence
rates for
neoplasms
Individuals living in
the Island of Krk
from 1997-2001
Soil and
vegetation
samples,
household
potable water
samples,
children's hair
samples
NA
No association observed between Pb in the
samples and incidence of neoplasm
(numerical results not provided)
Mendey et al.
(In Press!
Multiple U.S.
locations
Incidence of
cancer or
"malignancy
of any kind"
2007-2008
NHANES cohort -at
least 20 years of
age
Concurrently
measured
creatinine-
adjusted urinary
Pb
Geometric mean for
creatinine-adjusted
urinary Pb marker:
0.59 pg/g (95% CI:
0.57, 0.61)
Greater than log-transformed mean creatinine-
adjusted urinary Pb level compared to less
than log-transformed mean creatinine-
adjusted urinary Pb level: 0.76 (95% CI: 0.44,
1.33)
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Reference8
Study Cancer	Measure of Pb
Location Outcome Study Population Exposure Mean Pb (SD)
Adjusted Effect Estimates
Rousseau et al. Montreal, Lung cancer	Men aged 35-79 Interview of job
(2007) Canada and other	history and
cancer	exposure matrix
incidence
Ever exposed to:
Organic Pb 3.0%
Inorganic Pb 17.0C
Pb in gasoline
emissions 38.6%
Never exposed is referent group
Organic Pb:
Esophageal 1.7 (95% CI: 0.5, 6.4)
Stomach 3.0 (95% CI: 1.2, 7.3)
Colon 1.5 (95% CI: 0.7, 3.6)
Rectum 3.0 (95% CI: 1.2, 7.5)
Pancreas 0.9 (95% CI: 0.1, 5.2)
Prostate 1.9 (95% CI: 0.8, 4.6)
Bladder 1.7 (95% CI: 0.7, 4.2)
Kidney 2.3 (95% CI: 0.8, 6.7)
Non-Hodgkin's lymphoma 0.4 (95% CI: 0.1,
2.2)
Inorganic Pb:
Esophageal 0.6 (95% CI: 0.3,1.2)
Stomach 0.9 (95% CI: 0.6,1.5)
Colon 0.8 (95% CI: 0.5,1.1)
Rectum 0.8 (95% CI: 0.5,1.3)
Pancreas 0.9 (95% CI: 0.4,1.8)
Prostate 1.1 (95% CI: 0.7,1.6)
Bladder 1.1 (95% CI: 0.7,1.5)
Kidney 1.0 (95% CI: 0.6,1.7)
Melanoma 0.4 (95% CI: 0.2,1.0)
Non-Hodgkin's lymphoma 0.7 (95% CI: 0.4,
1.2)
Pb in gasoline emissions:
Esophageal 0.6 (95% CI: 0.4,1.1)
Stomach 1.0 (95% CI: 0.7,1.4)
Colon 0.8 (95% CI: 0.6,1.1)
Rectum 1.0 (95% CI: 0.7,1.4)
Pancreas 0.9 (95% CI: 0.5,1.4)
Prostate 0.9 (95% CI: 0.7,1.2)
Bladder 0.8 (95% CI: 0.6,1.1)
Kidney 1.0 (95% CI: 0.7,1.5)
Melanoma 0.8 (95% CI: 0.5,1.4)
Non-Hodgkin's lymphoma 0.7 (95% CI: 0.5,
1.0)
Note: results are for comparisons using
population-based controls; results for controls
with other types of cancers were similar
except no association was present between
organic Pb and rectal cancer
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Reference8
Study Cancer	Measure of Pb
Location Outcome Study Population Exposure Mean Pb (SD)
Adjusted Effect Estimates
All esophageal cancers:
Unexposed: 1.00
Low workplace Pb exposure (< 0.237 pmol/L):
0.79 (95% CI: 0.43, 1.46)
High workplace Pb exposure (>0.237 pmol/L):
1.69 (95% CI: 0.57, 5.03)
Esophageal squamous cell carcinoma:
Unexposed: 1.00
Low workplace Pb exposure (< 0.237 pmol/L):
0.70 (95% CI: 0.34, 1.43)
High workplace Pb exposure (>0.237 pmol/L):
0.91 (95% CI: 0.22, 3.75)
Adenocarcinoma:
Unexposed: 1.00
Low workplace Pb exposure (< 0.237 pmol/L):
0.95 (95% CI: 0.32, 2.82)
High workplace Pb exposure (>0.237 pmol/L):
5.30 (95% CI: 1.39, 20.22)
'associations not changed or slightly
increased when restricted to occupational
exposures> 15yr
aStudies listed in order of appearance in the text.
5.10.1.1 Overall Cancer Mortality
Several recent epidemiologic studies examined the association between Pb levels and
cancer mortality, including multiple analyses of the NHANES III population. In one
NHANES III analysis, the cohort 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 individuals with all blood Pb levels (including those greater than
10 (ig/dL), reported associations between blood Pb and cancer mortality (Schoberet 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 j^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 odds 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 from the greater Boston area enrolled in the NAS found no association between
Santibanez et al.
Valencia and
Alicante,
Spain
Esophageal
cancer
incidence
PANESOES study
included 30-80 yr
old men hospitalized
in any of the
participating study
hospitals
Interviews to
determine
occupational
history and a job
exposure matrix
NA
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blood or bone Pb and cancer mortality in adjusted analyses. 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 , white women aged 65-87 were included in a sub-study of blood Pb level and
cancer mortality and were followed for approximately 12 years (khalil et al. 2009a). The
mean (SD) blood Pb levels were 5.3 (2.3) (ig/dL and no association was detected between
blood Pb and cancer mortality in the study population.
Overall, epidemiologic studies of blood Pb levels and cancer mortality reported
inconsistent results. One epidemiologic study using NHANES III data demonstrated an
association between blood Pb and increased cancer mortality; however, other studies
reported weak or no associations.
5.10.1.2 Lung Cancer
Most of the evidence regarding lung cancer incidence is provided by studies of
occupationally-exposed adults. In a study of smelter workers, 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 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 j^ig/dL for controls. In a study of tin smelters 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, Pb exposure was calculated
by combining historical air sampling data and personnel record cards, which specified
work histories. The median Pb exposure was estimated to be approximately 2 mg-
year/m3. It is important to note that the smelter workers were exposed to other metals as
well, such as arsenic and antimony. A population-based case-control study
performed among men in Montreal, Canada assessed Pb exposure via interviews
regarding job histories and calculated 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.
Studies of Pb and lung cancer that compared the lung tissue of individuals with lung
cancer to those without lung cancer were also conducted. The controls for these studies
were individuals with metastases in the lung from other primary cancers (De Palma et al..
2008) and individuals with non-cancerous lung diseases (De Palma et al.. 2008; kuo et
al.. 2006). Findings are mixed among the studies. De Palma et al. (De Palma et al.. 2008)
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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. (kuo et al.. 2006)
found no statistical difference in Pb levels for lung tissue of individuals with lung cancer
compared to controls .
Some studies in the 2006 Pb AQCD reported associations between Pb exposure and lung
cancer; however, more 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 levels. In addition, studies
limited to occupational exposures may be confounded by other workplace exposures.
5.10.1.3 Brain Cancer
A few studies of brain cancer examined the association between cancer and Pb using
exposures determined via exposure databases and patient interviews about past jobs and
known exposures. Interpretation of these results is limited due to the lack of biological Pb
measures and the potential confounding by other occupational exposures. The National
Longitudinal Mortality Study, a study that included a national sample of the U.S.
population, estimated Pb exposure based on current/most recent employment among
individuals (Van Wijngaarden and Dosemeci. 2006). Although not all estimates 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, glioma was reported to have no association with any Pb exposure metric;
however, positive associations were observed between high cumulative Pb exposure and
meningioma among individuals w ith AIA1)2 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/nrV to those
not exposed to Pb) (Raiaraman et al.. 2006). This association was not present among
individuals with the ALAD1 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 |ag/m\ to those not exposed to Pb).
Another study of the association between 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) (Bhatti
et al.. 2009). After control for multiple comparisons, individuals with GPX1 variants
(rs 105 0450) had positive associations between cumulative Pb exposure and glioblastoma
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multiforme and meningioma. Individuals without RAC2 variants (rs2239774) showed a
positive association between Pb and glioblastoma multiforme. Also, individuals without
XDH variants (rs7574920) displayed a positive association between Pb and meningioma.
Overall, associations between Pb exposure and brain cancer incidence and mortality were
found to vary according to several genetic variants. These studies were limited in their
methods because they do not have individual level biological measures and the potential
for confounding by other workplace exposures exist. Future research will be important in
confirming these associations and the modification by various genetic variants.
5.10.1.4 Breast Cancer
A population-based case-control study in Canada examined the proximity to a Pb smelter
based on residential addresses (Pan etal. 2011). 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).
A few studies examined Pb levels and breast tumors among individuals with and without
breast tumor and/or cancer present. A study of newly diagnosed breast cancer patients
and controls examined Pb levels in blood and hair samples and reported higher levels of
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 also reported
no statistical difference in Pb levels (Pasha et al. 2008b). However, one 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 bisphostphonates (a combination responsible for bone loss) were excluded from
the analysis (McElrov et al.. 2008).
Overall, these studies demonstrate the possibility that women with breast cancer may
have higher Pb levels in blood measurement, whereas the results for actual breast tissue
are mixed. However, these studies are limited by their study design. The samples are
taken after cancer is already present in the cases, leading to issues of temporality for the
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Pb levels. Additionally, the sample sizes are often small and the studies may be
underpowered.
5.10.1.5 Other Cancers
Studies of all cancers combined, multiple different cancers, or cancers not listed above
have also been performed. An ecologic analysis compared levels of Pb in the air from
1990 to 2005 with incidence rates of cancer (cancer sites not specified) among children
during this time period (Absalon and Slesak. 2010). The highest Pb levels were measured
in 1990 when over 50% of the study area exceeded the limit of 1 |_ig/m2-ycar. No
correlation was observed both overall and in sex-specific analyses. A similar study
examined correlations between Pb concentrations in soil, water, vegetation, and hair
samples with incidence of neoplasms (Obhodas. 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
having ever had cancer or a malignancy (Mendv et al. In Press). The timing of cancer
diagnosis in relation to the urine sample collection was not identified.
A study performed among men evaluated multiple cancer outcomes and determined
exposures to organic Pb, inorganic Pb, and Pb from gasoline emissions via interviews
regarding job histories and then subsequent exposure approximations by chemists and
hygienists (Rousseau et al. 2007). Adults exposed to organic Pb exposure had greater
odds of stomach cancer compared to adults never exposed 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 exposure to inorganic Pb. When occupational exposure to Pb in
gasoline was categorized as unexposed, nonsubstantial level, and 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, reported no association
between Pb exposure and esophageal squamous cell carcinomas, but an association was
present between high Pb exposure and adenocarcinoma of the esophagus (Santibanez et
al.. 2008). However, neither of these studies was able to quantify Pb levels using
biological measurements.
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Several studies compared Pb levels in blood, tissue, and urine of individuals who have
cancer with 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 (Olszew ski et al.. 2006). the blood of
individuals with gastric cancer (Khorasani et al.. 2008). the plasma and hair of
individuals with gastrointestinal cancer (Pasha et al.. 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
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.. 2006V All of these comparison studies are limited by the inability to examine
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 who have not
undergone certain treatments. Additionally, studies are limited by their small sample size
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 sum, epidemiologic studies reported no associations between various measures of Pb
exposure and overall cancer incidence. Studies examining specific cancers reported
varying associations. Associations were null for Pb and most cancer sites examined;
however a positive association was observed between Pb exposure and adenocarcinoma
of the esophagus as well as exposure to organic Pb and stomach cancer. Associations
between organic Pb exposure and rectal cancer and exposure to Pb in gasoline and
stomach cancer were inconsistent. These conclusions are limited by the small number of
studies and a lack of biological measurements of Pb.
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5.10.1.6 Toxicological Models of Carcinogenicity
Carcinogenicity in Animal Models
Previous AQCDs have established that Pb has been shown to act as a carcinogen in
animal toxicology models, albeit at relatively high concentrations. The 2006 Pb AQCD
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. The kidneys are the most common
target of Pb-dependent carcinogenicity (kasprzak et al.. 1985; Roller et al.. 1985; Azar et
al.. 1973; Van Esch and kroes. 1969) but the testes, brain, adrenals, prostate, pituitary,
and mammary gland have also been 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., 0.1 or 1% Pb-acetate) or drinking water
(i.e., 26 or 2,600 ppm Pb-acetate) for 18 months to two years in duration (Kasprzak et al..
1985; Roller et al.. 1985; Azar et al.. 1973; Van Esch and Kroes. 1969). These two year
cancer bioassays and the doses employed are typical of cancer bioassays employed by
other chemicals. Recognition of the importance of windows of exposure in Pb-induced
cancer bioassays is a focus of more recent studies. In one study, early life gestational and
lactational exposure of laboratory rodents to inorganic Pb induced carcinogenicity in
adult offspring (Waalkes et al.. 1995). Another recent study considered Pb-dependent
carcinogenesis in laboratory animals with early life Pb exposure. Tokar et al. (2010)
considered tumorigenesis in homozygous metallothionein I/II knockout mice and their
corresponding wild type controls (groups of ten mice each) that were exposed by
drinking water to 2,000 or 4,000 ppm Pb-acetate 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. 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-dependent carcinogenesis in humans. Thus, 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 (i.e., 2 years), which is typical of cancer bioassays. Newer studies are showing
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early life maternal Pb exposure can contribute to carcinogenicity in offspring and have
shown that metallothionein is protective against cancer in this pathway.
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 |ag/cm2
PbCr04 for 120 h. PbCr04 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 PbCr04 for 120 h. PbCr04 induced foci formation in a
concentration-dependent manner in the Mrel 1 deficient cells. Mrel 1 was required to
prevent PbCr04-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 due to the intersection 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 anti-oxidant defenses, can increase the likelihood of mutagenesis,
cellular instability, and tumor cell formation. For example, Xu et al. (2008) found
toxicological evidence that supports the association with Pb exposure and DNA damage
and 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.
5.10.2 Cancer Biomarkers
A study of men aged 21-40 years without occupational history of metals exposure
examined prostate specific antigen (PSA), a biomarker for prostate cancer. This study
reported a positive association between Pb levels and PSA levels (measured in the same
blood samples) in regression models adjusted for confounders, including other metals
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(Cd, Zn, Se, and Cu) (Pizent et al.. 2009). The median 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.
5.10.3 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.
5.10.3.1 Epidemiologic Evidence for DNA and Cellular Damage
Multiple studies examined the relationship between Pb and sister chromatid exchange
(SCE). SCEs are exchanges of homologous DNA material between chromatids on a
chromosome and is a test for mutagenicity or DNA damage. A study of male policemen
reported mean blood Pb levels for the study population of 43.5 j^ig/dL ("Wiwanitkit et al..
2008V When dichotomized as having 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
levels of 40.14 (ig/dL) and office workers (mean blood Pb levels of 9.77 (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 levels of 7.69
[4.29] (ig/dL) reported no correlation between blood Pb levels and SCE (Mielzviiska 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 (children attending a school far from a Pb
smelter: median blood Pb levels 4.6 (ig/dL; children attending a school near a Pb smelter:
median blood Pb levels 28.6 j^ig/dL for) reported no association between blood Pb and
baseline DNA damage or repair ability after a peroxide challenge (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
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(although this was not statistically significant in linear regression models controlling for
age) (Grover et al. 2010V 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 j^ig/dL) compared to the office workers
(blood Pb level <10 j^ig/dL for) (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.
5.10.3.2 Toxicological Evidence for DNA and Cellular Damage
Sister Chromatid Exchanges
Tapisso et al. (2009). considered sister chromatid exchanges (SCE) in Algerian mice
(groups of six mice each) that were exposed by i.p. injection to 5 or 10 doses of
0.46 mg/kg Pb-acetate. The SCE in bone marrow were elevated after Pb exposure alone,
which increased with time. Co-exposure with cadmium or zinc 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 animals. In another study,
Turkez et al. (In Press) considered the ability of boron compounds 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.
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Micronuclei Formation
The 2006 Pb AQCD stated "studies of genotoxicity consistently find associations of Pb
exposure with DNA damage and MN formation" and the current document continues to
report these associations. Alghazal et al. (2008b). considered the ability of Pb-acetate
trihydrate to induce MN in bone marrow of Wistar rats. Animals were given a daily dose
of 100 mg/1 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 alone, Pb plus zinc and Pb plus cadmium-
induced MN in rodents. Algerian mice were exposed by i.p. injection to 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. Co-exposure with cadmium or
zinc did not further increase MN levels.
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. (In Press) 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 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 was 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.
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(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) considered
Pb-acetate induced HPRT mutations in Chinese hamster ovary AA8 cells and AA8 cells
overexpressing human Apel. 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, an 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 0.5% w/w to 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.
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 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, 1.2 or 2 mM
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
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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. Wise et al. (2006b) treated BEP2D cells with 0, 0.5, 1, 5,
or 10 lag/cm2 Pb chromate for 24 hours. 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. (2006a'). 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
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 h. Specific
CHO lines used included AA8 (wildtype) EM9 (XRCC1-deficient), and H9T3 (EM9
complemented with human XRCC1 gene). Pb chromate induced statistically significant,
concentration-dependent increases in chromosomal aberrations that were statistically
significantly increased by XRCC1 deficiency. Nestmann and Zhang (2007) treated
Chinese hamster ovary cells (clone WB(L)) with 0, 0.1, 0.5, 1, 5, or 10 (ig/cm2 Pb
chromate (as pigment yellow) for 18 h. 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 h. Specific CHO lines used included AA8 (wildtype), KO40
(i
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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 (XRCC3-deficient), and
lSFwt8 (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 which 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/cm2 Pb 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
specific DNA repair enzymes helped to elucidate which pathways may be most sensitive
to Pb-dependent chromosomal aberration.
COMET Assay
Multiple studies considered the ability of Pb to induce DNA single strand breaks in
laboratory animals 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) considered the ability of Pb-acetate to induce DNA
damage measured by the comet assay in lymphocytes of male ICR mice. 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 dose-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
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(2011) considered the ability of Pb nitrate to induce DNA damage in liver tissue of adult
male Wistar rats. Animals (8 per group) were treated for 60 days with doses of 0, 0.5 or
1% Pb(N03)2 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 (.iM) did not cause DNA
damage with the comet assay (Carmona et al.).
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, 2.1, 2.4, 2.7, 3.0, 3.3 Pb nitrate for 2 hours and found dose-dependent
increases in Comet tail length. Concerns with the study are that apparently no untreated
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. 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. Mrel 1 is a
component of the MRN complex and plays a role in telomere maintenance and double-
strand break repair. Pb chromate induced a concentration-dependent increase in DNA
double strand breaks measured by the comet assay. In another study, Pb nitrate exposure
(30 (ig/mL) induced statistically significant increased DNA damage in human liver
HepG2 cells that was statistically significantly attenuated with co-exposure with the
antioxidant NAC (500 (.iM) (Yediou 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.
Other Assays
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.
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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/cnr 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 other three
compounds. Two studies demonstrated the ability of Pb to destabilize DNA by forming
DNA-histone cross links, which can lead to histone aggregation. Extracts of rat liver co-
incubated with Pb nitrate (<300 |_iM) were shown to react with chromatin components
and induce chromatin aggregation via histone-DNA cross links (Rabbani-Chadegani et
al.. 2011; Rabbani-Chadegani et al.. 2009).
Genotoxicity testing of Drosophila 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.). Further, wing spot
assays employing Pb co-exposure with gamma radiation showed no effect of Pb on
gamma radiation induced spotting frequency. 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 Wiirgler. 1986).
Multiple studies considered Pb and DNA repair. Most were conducted in cultured cells,
and one was done in an animal model. El-Ghor et al. (In Press) 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 male albino rats exposed to Pb-acetate (acute: single oral dose of
467 mg/kg BW or sub-chronic: 47 mg/kg BW six d/wk for 4 wk) showed increased MSI
at three microsatellite loci (D6mit3, D9mit2, and D15Mghl). 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 (NHEJ) 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
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required to prevent Pb chromate-induced DNA double strand breaks. McNeill et al,
(2007) considered Pb-acetate effects on Apel. Chinese hamster ovary cells (AA8) 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. An earlier study in rats showed Pb nitrate-induced increased proliferation of
liver cells after a partial hepatectomy, effects that were more prominent in males than
females (sexual dimorphism) (Tessitore et al.. 1995). The newer studies showed similar
trends in males. Fortoul et al. ("2005) exposed adult male CD1 mice (24 animals per
group) to 0.01 M Pb-acetate, 0.006 M cadmium chloride or a mixture of the two
chemicals for 1 h twice a week for 4 weeks by inhalation. The lungs were then examined
by electron microscopy for changes. Pb induced cellular proliferation in the lungs.
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.
5.10.4 Effects of Lead within Mixtures
Several studies considered the impact of mixtures with Pb. All considered genotoxicity.
Mendez-Gomez et al., (2008). evaluated 65 children from Mexico exposed to both
arsenic and Pb. DNA damage and decreased DNA repair were seen using the comet assay
and other assays, but did not correlate with urinary arsenic or blood Pb levels. Tapisso et
al., (2009). considered Pb alone, Pb plus zinc and Pb plus cadmium-induced MN in
rodents. Algerian mice (groups of six mice each) were exposed (i.p.) to 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
cadmium or zinc 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 550 (ig/L Pb chloride, 15 (ig/L cadmium sulfate,
25 (ig/L cobalt chloride or all three combined for 72 hours. There was a clear interaction
of all three metals impacting RNA expression.
No recent studies of the protective role of calcium in Pb-dependent carcinogenesis or
genotoxicity were found. There were some data suggesting that boron, melatonin,
N-acetylcysteine, turmeric and myrrh protect cells against Pb-induced genotoxicity
(Sections 5.10.3.2 and 5.10.6).
A new study details Pb and selenium interactions in virus-dependent carcinogenesis in
laboratory animals. Schrauzer (2008) considered the impact of selenium on
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carcinogenesis by studying 4 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 selenium
and then were exposed via drinking water to acetic acid (control group) or 0.5 ppm
Pb-acetate (treated group). The second set of two groups were fed a diet containing
0.65 ppm selenium 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 selenium diet. The data
suggest that selenium is anticarcinogenic as in the groups without Pb exposure, the
animals exposed to the higher selenium levels had fewer mammary tumors and these
tumors had a delayed onset of appearance. Pb exposure with low selenium caused the
same delayed onset as did the higher dose of selenium and also caused some reduction in
the tumor frequency. Pb exposure with higher selenium 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 selenium, but they
suggest that Pb mimics or antagonizes selenium. They do not suggest that selenium is
protective of Pb-induced toxicity or carcinogenesis.
5.10.5 Modes of Action
The carcinogenic mechanism of action of Pb is poorly understood. It is unclear whether
the mechanism 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 chromosomal mutation and
DNA damage, i.e., clastogenicity. However, the ability of Pb to alter gene expression
(epigenetic effects) and to interact with proteins may be a means by which Pb induces its
carcinogenicity. It is known that Pb can replace zinc in zinc-binding (zinc-finger)
proteins, which include hormone receptors, cell-cycle regulatory proteins, the Ah
receptor, estrogen receptor, p53, DNA repair proteins, protamines, and histones. These
zinc-finger proteins all bind to specific recognition elements in DNA. Thus, Pb may act at
a post-translational stage to alter protein structure of Zn-finger 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(s). These
modifications then provide the cancer cells with a selective growth advantage. In this
schematic, Pb appears to contribute to epigenetic changes, and chromosomal aberrations.
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The genomic instability paradigm requires a cascade of genome-wide changes caused by
interfering with 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 can lead to cancer by altering cellular functions without altering the genetic
material. The most commonly studied epigenetic change is methylation alterations. A
small number of studies show that Pb can induce epigenetic changes, but studies are still
missing to clearly tie these effects to Pb-induced carcinogenesis and genotoxicity. Thus,
the mechanism is difficult to define but, if Pb is a human carcinogen, the mechanism
likely involves either genomic instability or epigenetic modification paradigms or some
combination of the two. However, Pb-dependent carcinogenicity is it not likely to occur
by a multistage paradigm. More work is needed to determine the mechanism.
Exposure to mixtures can also contribute to understanding of modes of action. No recent
studies of the protective role of calcium or zinc in Pb-dependent carcinogenesis or
genotoxicity were found. There were some data suggesting that metallothionein protects
rodents from Pb-induced cancers. There were some data suggesting that boron,
melatonin, N-acetylcysteine, turmeric and myrrh protected cells against Pb-induced
genotoxicity. There were some data suggesting that Pb mimics or antagonizes selenium
in rodents. These data are discussed in more detail elsewhere in the cancer section and
point to the relevance of mixtures in assessing toxicity.
5.10.5.1 Epigenetics
Air pollution exposure is being linked increasingly with epigenetic changes (Pavanello et
al.. 2010; Baccarelli and Bollati. 2009; Tarantini et al.. 2009; Bollati et al.. 2007).
Epigenetic changes involve changes in DNA expression 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 et al.. 2010; Pilsner et al.. 2009). Wright
et al. (2010) utilized a sample of participants from the NAS with mean (SD) Pb levels of
20.5(14.8)g/g for tibia measures, 27.4 (19.7)g/g for patella measures, and 4.1 (2.4) (ig/dL
for blood measures. In both crude and adjusted analyses, patella Pb levels were inversely
associated with LINE-1 methylation but not with Alu, both of which are indicators of
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global methylation. When examining 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 j-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 and measured
LINE-1 and Alu methylation in umbilical cord blood samples (Pilsner et al. 2009). In
unadjusted models, maternal pregnancy tibia Pb levels [mean (SD) 10.5 (8.4) jj.g/g] were
inversely associated with Alu methylation; maternal patella Pb levels [mean (SD) 12.9
(14.3) jJ.g/g] were inversely associated with LINE-1 methylation. The associations
persisted in adjusted models although the association between patella Pb and LINE-1 was
only apparent when the adjusted models also included umbilical cord blood Pb levels. No
association was detected between umbilical cord Pb levels and the DNA methylation
markers. Overall, the studies consistently demonstrate an association between patella Pb
levels and LINE-1 methylation.
Toxicological studies have been performed examining Pb-dependent 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 550 (ig/L Pb chloride, 15 (ig/L cadmium sulfate, 25 (ig/L
cobalt 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 full interpretations of the data as epigenetic changes are complicated by the
absence of a measure to determine if these changes were a result of genotoxic effects. A
recent publication using HepG2 cells in tissue culture showed that cells exposed to a high
dose of Pb (100 (.iM Pb-acetate) experienced 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 et al.. 2011).
5.10.6 Summary and Causal Determination
In summary, the toxicological literature on the genotoxic, mutagenic, and carcinogenic
potential of Pb provide strong evidence of effects in laboratory animals. In laboratory
studies, high-dose Pb has been demonstrated to be an animal carcinogen. There are data
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to suggest Pb is a human carcinogen among toxicological studies, but they are not
definitive. Multiple toxicological studies showed neoplastic transformation in cultured
cells, but all focused on Pb chromate and the positive response to the chromate ion and
not the Pb likely contributed to these findings. Multiple 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 blood Pb and cancer mortality, but the other
studies reported null results. Although the 2006 Pb AQCD reported some studies that
found an association between Pb exposure indicators and lung cancer, current studies
mostly included occupationally-exposed adults and observed no associations. Most
studies of Pb and brain cancer were null among the overall study population, but positive
associations were observed among individuals with certain genetic variants. However the
studies of Pb and brain cancer were all performed among occupational cohorts using
interviews instead of biological measurements to represent Pb exposure. 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 recent studies of stomach cancer are lacking. One study examining Pb and
stomach cancer has been performed since the last AQCD and the results of the study are
mixed
Among epidemiologic studies, high Pb levels were associated with SCEs among adults
but not children. 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, but
continues to not produce chromosomal aberrations except for Pb chromate; this again is
likely due to the chromate. Pb does not appear to be very mutagenic as the HPRT assays
were typically negative unless a cell signaling pathway was disturbed.
Mechanistic understanding of the carcinogenicity of Pb is expanding with work on the
antioxidant selenium and metallothionein, a protein that binds Pb and reduces its
bioavailability. 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 zinc, Pb replaces zinc at many zinc binding sites or zinc finger
proteins. This substitution has the potential to induce indirect effects that can contribute
to carcinogenicity via interactions at hormone receptors, at cell-cycle regulatory proteins,
with tumor suppressor genes like p53, with DNA repair enzymes, with histones, etc.
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These indirect effects may act at a post-translational level to negatively alter protein
structure and DNA repair.
Epigenetic changes associated with Pb exposure, particularly with respect to methylation
and effects on DNA repair, are beginning to appear in the literature. These modifications
may further alter DNA repair or change the expression of a tumor suppressor gene or
oncogene. A small number of epidemiologic studies examining Pb and global epigenetic
changes including LINE-1 and Alu demonstrated an inverse association between patella
Pb and LINE-1 methylation, an 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 either genomic instability,
epigenetic modifications, or both.
In conclusion, the toxicological literature provides the strongest evidence for Pb exposure
and cancer with supporting evidence provided by the epidemiologic literature. 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 toxicological studies demonstrates an association
between Pb and cancer, genotoxicity/clastogenicity or epigenetic modification.
Carcinogenicity in historical animal toxicology studies with 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, Pb exposure was found to be associated with
stomach cancer, but there was only one recent study on stomach cancer and Pb exposure,
which reported mixed findings. Similarly, some studies in the 2006 Pb AQCD reported
associations between Pb exposure and lung cancer. More recent epidemiologic studies of
lung cancer focused on occupational exposures and reported no associations. The
majority of epidemiologic studies of brain cancer had null results overall, but positive
associations between Pb exposure and brain cancer were observed among individuals
with certain genotypes. In toxicological studies, chromosomal aberrations after Pb
exposure are most often reported with Pb chromate exposure, which is likely due to
toxicity of the chromate moiety. Mechanistic understanding of Pb and its effect on cancer
and genotoxicity is expanding through toxicological work focusing on antioxidants and
other proteins that sequester Pb or reduce its bioavailability. The collective body of
evidence integrated across toxicological and epidemiologic studies is sufficient to
conclude that there is a likely causal relationship between Pb exposure and cancer.
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CHAPTER 6 POTENTIALLY AT-RISK
POPULATIONS
The NAAQS are intended to provide an adequate margin of safety for both the population
as a whole and those groups with unique factors that make them potentially at increased
risk for health effects in response to ambient air pollutants (Preface to this ISA).
Interindividual variation in human responses to air pollution exposure suggests that some
populations are at increased risk for detrimental effects of ambient exposure to an air
pollutant. To facilitate the identification of populations at greater 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. 2010a. 2009). 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.
To examine whether Pb differentially affects certain populations, epidemiologic studies
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 greater risk for Pb-related health effects. Toxicological
studies, using animal disease models, also provide support and biological plausibility for
factors that may lead to increased risk for Pb-related health effects. These epidemiologic
and toxicological studies provide the scientific basis for an overall weight of the evidence
evaluation for the increased risk of specific populations to Pb-related health effects.
The first section of this chapter summarizes physiological factors that possibly influence
Pb levels in the body. The second section of this chapter summarizes information on
factors potentially related to differential Pb exposure. The studies presented in this
section supplement the material provided in Chapters 3 and 4 by examining how factors
such as age, sex, race and ethnicity, SES, proximity to Pb sources, and residential factors
may affect Pb exposure. The third section of this chapter discusses the epidemiologic and
toxicological studies evaluated in Chapter 5 that provide information on potential factors
related to increased risk of Pb-induced health effects. Highlighted studies include only
those where the population was stratified into subgroups (e.g., males versus females or
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smokers versus nonsmokers) for comparative analysis. In the case of many biomarker
studies and the epidemiologic studies considered, this approach allowed for a comparison
between populations exposed to similar Pb concentrations and within the same study
design. Numerous studies that focused on only one potentially at-risk population were
described in previous chapters (Chapter 5) 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 5 but are not included here. Included toxicological studies may have
categorized the study populations by age, sex, diet/nutrition status, genetics, etc. or are
those that examined animal models of disease.
Additionally, it is understood that some of the stratified variables/factors discussed in this
third section 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
and health outcomes, whereas effect measure modifiers are factors that result in changes
in the measured associations between Pb 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 6-4.
6.1 Physiological Factors that Influence the Internal Distribution
of Lead
Blood and bone Pb measures are influenced to varying degrees by biokinetic processes
(absorption, distribution, metabolism, etc.), which are discussed in detail in 0. 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 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). 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'Flahertv. 1995; Leggett.
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1993). Older adults are more 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 (Mivaki 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, iron, and zinc 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 (Liu et al..
2011b). Diets designed to limit or reduce caloric intake and induce weight loss have been
associated with increased blood Pb levels in adult animals (Han et al. 1999). A
toxicological study reported negative effects of Pb on osmotic fragility, TBARS
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.
A zinc-supplemented diet attenuated bone and renal Pb content. Toxicological studies
have shown that dietary deficiency of calcium induces increased Pb absorption and
retention (Fullmer. 1992; Mvkkanen and Wasserman. 1981; Six and Gover. 1970).
Increased calcium intake reduces accumulation of Pb in bone and mobilization of Pb
during pregnancy and lactation (Bogden et al.. 1995). Additionally, studies have reported
that iron deficiencies may result in higher Pb absorption or altered biokinetics (Schell et
al.. 2004; Marcus and Schwartz. 1987; Mahaffev and Annest. 1986).
Finally, co-exposures with other metals and non-metals have also been studied to assess
how they affected the uptake and absorption of Pb. Recent toxicological studies
examined the addition of arsenic (As) to Pb and cadmium (Cd) mixtures, and reported
increased bioavailability of Pb (Wang and Fowler. 2008). A toxicological study by
Sawan et al. (2010) reported co-exposure with fluoride increased Pb deposition in
calcified tissues.
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In summary, age, genetics, diet, and other exposures affect the biokinetics of Pb, which in
turn affects the internal distribution of Pb. These factors were discussed in greater detail
in 0 where more information on overall biokinetics and physiological factors affecting Pb
distribution was provided.
6.2 Population Characteristics Potentially Related to Differential
Lead 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 population characteristics on exposure, absorption, and biokinetics varies
on an individual basis 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 our previous understanding of the differential influence of population
characteristics on Pb exposure. Differential response to given Pb exposures is discussed
in Section 6.3.
6.2.1 Age
6.2.1.1 Early Childhood
Typically, children have increased exposure to Pb compared with adults because
children's behaviors and activities include increased hand-to-mouth contact, crawling,
and poor hand-washing that typically result in increased ingestion compared with adults
(U.S. EP A. 2006b). Children can also be susceptible to Pb exposure because outdoor play
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 2007-2008 NHANES data are presented in Table 6-1 by age
and sex. Among children, highest blood Pb levels occurred in the 1-5 year age group
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(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: 16.9 (ig/dL). 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
were located near sources. 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
overtime (see Table 6-2). 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. These distribution differences may be attributable to differences in exposure,
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 younger than 73 months living in proximity to a Turkish coal mine. They
observed a low but statistically significant negative correlation between blood Pb and age
(r =-0.38, p <0.001).
Table 6-1 Blood Pb levels by age and sex, 2007-2008 NHANES
Age
Sex
N
Avg.
Std. Dev.
5%
25%
50%
75%
95%
99%
1-6 yr
Total
811
2.03
2.01
0.69
1.08
1.64
2.34
4.60
10.60

Male
440
2.01
2.14
0.71
1.10
1.50
2.40
4.21
8.56

Female
371
2.05
1.85
0.66
1.02
1.60
2.28
4.65
10.70
6-11 yr
Total
1,002
1.27
0.87
0.49
0.76
1.06
1.60
2.84
4.80

Male
500
1.29
0.92
0.48
0.74
1.08
1.50
2.90
4.76

Female
502
1.25
0.83
0.50
0.78
1.03
1.49
2.80
4.80
12-19 yr
Total
1,089
0.99
0.73
0.40
0.69
0.81
1.13
2.11
4.00

Male
586
1.13
0.82
0.44
0.69
0.94
1.30
2.41
4.11

Female
503
0.83
0.57
0.37
0.52
0.69
0.91
1.70
3.36
20-64 yr
Total
4,278
1.76
1.67
0.63
0.91
1.40
2.10
4.20
7.42

Male
2,079
2.13
1.93
0.68
1.12
1.65
2.49
5.22
9.07

Female
2,199
1.41
1.01
0.47
0.78
1.13
1.73
3.13
5.10
66+yr
Total
1,086
2.31
1.64
0.80
1.30
1.90
2.70
6.22
8.64

Male
542
2.63
1.66
0.99
1.52
2.19
3.20
5.86
9.03

Female
544
1.98
1.56
0.75
1.17
1.62
2.40
4.14
6.95
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Age
Sex
N
Avg. Std. Dev. 5%
25%
50% 75%
95%
99%
Overall Total
8,266
1.70 1.63 0.61
0.86
1.30 2.04
4.10
7.20

Male
4,147
1.94 1.77 0.58
0.99
1.5 2.3
4.72
8.69

Female
4,119
1.45 1.19 0.46
0.76
1.14 1.78
3.35
5.77
Source: (NCHS. 2010V
Table 6-2
Percentage of children within six categories/brackets of blood Pb
levels, 1999-2004 NHANES
Pb Units:
Mg/dL
(95% CI)
n
Geometric
mean
<1 Mg/dL, %
1 to <2.5
[jg/dL, %
2.5 to <5
|jg/dl_, %
5 to <7.5
[jg/dL, %
7.5 to <10
[jg/dL, %
> 10
|jg/dl_, %
Overall
2,532
1.9(1.8-2.0)
14.0(11.6-16.6)
55.0(52.1-57.9)
23.6(21.1-26.1)
4.5 (3.3-5.9)
1.5(1.0-2.1)
1.4(1.0-2.0)
Sex
Female
1,211
1.9(1.7-2.0)
14.1 (10.8-17.7)
54.5(51.1-57.8)
23.9 (20.3-27.8)
4.5 (3.3-5.8)
1.4(0.8-2.3)
1.7(0.9-2.6)
Male
1,321
1.9(1.7-2.0)
14.0(11.4-16.7)
55.5(51.4-59.5)
23.2 20.3-26.3)
4.6 (3.0-6.5)
1.5(0.9-2.3)
1.3(0.7-2.6)
Age
1-2 yr
1,231
2.1 (2.0-2.2)
10.6(7.7-13.9)
51.0(46.7-55.3)
27.9(24.9-31.0)
6.7 (5.0-8.6)
1.4(0.8-2.2)
2.4(1.4-3.5)
3-5 yr
1,301
1.7 (1.6-1.9)
16.2 (12.9-19.9)
57.6(53.8-61.4)
20.7 (17.9-23.7)
3.1 (1.9-4.6)
1.5(0.8-2.3)
0.9(0.4-1.5)
Race/Ethnicityc
Non-Hispanic
Black
755
2.8 (2.5-3.0)
4.0 (2.5-5.7)
42.5 (37.8-47.2)
36.2 (33.1-39.3)
9.4(6.9-12.2)
4.6 (3.0-6.5)
3.4(1.8-5.5)
Mexican
American
812
1.9(1.7-2.0)
10.9(8.6-13.4)
61.0(56.9-65.1)
22.1 (18.0-26.5)
3.4 (2.2-5.0)
1.3(0.6-2.2)
1.2(0.4-2.6)
Non-Hispanic
White
731
1.7(1.6-1.8)
17.6(14.0-21.5)
57.1 (52.4-61.7)
19.7(16.1-23.5)
3.6(1.9-5.8)
0.8(0.3-1.6)
1.2(0.6-2.0)
Poverty-income Ratio (PIR)c
<1.3
1,302
2.4 (2.2-2.5)
6.7 (4.6-9.2)
49.3 (44.9-53.7)
32.5 (28.6-36.4)
6.9 (2.2-8.8)
2.8(1.7-4.1)
1.8(1.1-2.7)
>1.3
1,070
1.5(1.4-1.6)
19.9(16.3-23.8)
60.4 (56.9-63.8)
16.0(12.9-19.3)
2.3(1.2-3.7)
0.6(0.1-1.4)
0.8(0.3-1.6)
Source: Reprinted with permission of the American Acedemy of Pediatrics; Jones et al. f2009a1
Fetal and child Pb biomarkers have been demonstrated to relate to maternal Pb
biomarkers; several older studies in the literature were presented 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). Miranda et al. (2010) observed that 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 in the (25- to 29-year-olds) 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 4.2 and summarized in Section 6.1. Elevated blood Pb levels among mothers
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present a potential exposure route to their children in utero or through breast milk.
Additionally, maternal pica presents a potential Pb source to a developing fetus
(Hamilton et al.. 2001).
6.2.1.2 Adulthood
Blood Pb levels tend to be higher in older adults compared with the general adult
population (U.S. EPA. 2006b). Table 6-1 presents 2007-2008 NHANES data broken
down by age group and shows that blood Pb levels were highest in the 65+years age
group in comparison with adults aged 20-64 years. 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 < 0.001). Average blood Pb
levels among 14- to 24-year-old subjects was 2.60 ± 0.16 (ig/dL compared with 4.29 ±
0.56 (ig/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 4.2).
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 increased with age, although the data quality for patella Pb was not as
high, so 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.
6.2.2 Sex
Several studies have suggested that sex influences levels of Pb biomarkers because
differences in behavior between sexes may cause a differential increase in exposure. The
2007-2008 NHANES showed that overall, males have significantly higher blood Pb
levels (average: 1.94 (ig/dL) than females (average: 1.45 (ig/dL) (p < 0.0001). Among
adults aged 20-64 years, average blood Pb levels were 51% higher for males compared
with females (p < 0.0001). Among adults 65 years or older, average blood Pb levels were
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33% higher for males compared with females (p < 0.0001). 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 jj.g/g) and women (19.4 jj.g/g).
Among U.S. children, the 2007-2008 NHANES data showed that blood Pb levels were
higher among girls than boys for the 1- to 5-years age group (Table 6-1). 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 in the 12- to 19-years age
group. At the same time, blood Pb levels among both adolescent males and females were
lower than blood Pb levels for the other age groups. The 2007-2008 NHANES data
suggest that sex-based differences in blood Pb levels are not substantial until
adolescence.
6.2.3 Race and Ethnicity
Higher blood Pb and bone Pb levels among African Americans has been well
documented (U.S. EPA. 2006^. 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. 2010) from 1988-1991 and
1999-2004 are shown in Figure 6-1. The authors found that differences among children
from different racial/ethnic groups with regard to the percentage with blood Pb levels >
2.5 (ig/dL overthe 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 ng/dL
compared with non-Hispanic whites and Mexican Americans, with large observable
differences for blood Pb levels between 2.5 and <10 ng/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 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 jj.g/g, C: 16.7 jj.g/g, p <0.01) but not patella Pb (AA: 7.1 jj.g/g, C: 7.1 jj.g/g, p =
0.46) or blood Pb levels (AA: 3.6 j^ig/dL, C: 3.6 (ig/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).
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70 -i
60
c 50
QJ
5 40
H—
o
c 30
0J
aj
CL 20
10
1988-1991
<1 1 - < 2.5 2.5-<5 5-<7.5 7.5-<10 >10
1999 - 2004
<1 1 - < 2.5 2.5 - < 5 5 - < 7.5 7.5 - <10 >10
Blood Pb Level (ug/dL)
Non-Hispanic black M Mexican American —Non-Hispanic white
Data used with permission of the American Academy of Pediatrics, Jones et al. (2009a)
Note: from the NHANES survey, 1988-1991 (top) and 1999-2004 (bottom).
Figure 6-1 Percent distribution of blood Pb levels by race/ethnicity among
U.S. children (1-5 years).
Differences in exposure among ethnic and racial groups have also been noted. In a study
of three parishes in the greater metropolitan New Orleans area, Campanella and Mielke
(2008) found that, where 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,
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1	and 4% Hispanic (Figure 6-2). As described in Section 6.2.4, the differences observed by
2	Campanella and Mielke (2008) may also be attributable to SES factors, or SES may be a
3	confounding factor in the relationship between Pb soil levels and race/ethnicity of nearby
4	residents.
50,000
45,000
40,000
35,000
30,000
C
O
| 25,000
Q.
O
20,000
15,000
10,000
5,000
0
Source: Data used with permission of Springer Science; Campanella and Mielke (20081.
Note: By Census 2000 race/ethnicity demographic groups.
Figure 6-2 Soil Pb concentration exposure among the population of three
parishes within greater metropolitan New Orleans.
6.2.4	Socioeconomic Status (SES)
5	Socioeconomic factors have sometimes been associated with Pb exposure biomarkers,
6	although these relationships have not always been consistent (U.S. EPA. 2006b). Nriagu
7	et al. (2006) performed a multiple regression analysis of blood Pb and saliva Pb levels on
8	various socioeconomic, demographic, and exposure variables among an adult population
m
Ln
Ln
¦	Black
¦	White
~	Asian
~	Hispanic


$


#
V
c/
Soil Pb Concentration (mg/kg)
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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 gap between income levels with respect to blood Pb 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 (ig/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 (ig/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 (ig/dL) compared to children not enrolled in Medicaid (1.7
(ig/dL). When adding data for 2003-2004 to the analysis (i.e., for 1999-2004), the gap
between Medicaid enrolled and non-enrolled children widened for blood Pb levels >10
(ig/dL (1.9 % versus 1.1%), but the difference was still not statistically significant (p >
0.05). Median blood Pb levels with respect to Medicaid status did not change when
adding the 2003-2004 data (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 (ig/dL was higher
for PIR <1.3 (1.8 versus 0.8); however, this difference was not statistically significant.
Additionally, Campanella and Mielke (2008) observed a linear increase in soil Pb content
outside a home with respect to decreasing average household income, with soil Pb
between 2.5 and 20 mg/kg associated with a Census block-averaged median household
income of $40,000 per year, while soil Pb between 5,000 and 20,000 mg/kg was
associated with a Census block-averaged median household income of $24,000 per year.
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6.2.5
Proximity to Lead Sources
Airborne and soil Pb concentrations are higher in some industrialized and urbanized
areas, as described in Sections 3.2, 3.3, 3.5 and 4.1, as a result of historical and
contemporaneous 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 (Mielke et al.. 2010a; Laid law and
Filippelli. 2008; Weiss et al.. 2006). Several studies describe mechanisms by which larger
particles present in soil can become resuspended to the air by perturbation from traffic-
and wind-generated turbulence (Harris and Davidson. 2008; Lough et al. 2005; Zereini et
al.. 2005; Nicholson. 1988; Gillette et al.. 1974). However, there are no recent monitoring
studies of this process, and the current routine monitoring network is not designed for the
assessment of the specific contribution of resuspension to ambient air concentrations. Air
Pb concentrations exhibit high spatial variability even at low concentrations (-0.01
Hg/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 2007-2009 data presented in Section 3.5.
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 (ig/dL [95% CI: 1.5,
3.3 |ig/dL|) compared with a reference group of neonates not living near a potentially
contaminated site (median: 1.1 (ig/dL [95% CI: 0.8, 1.3 |ig/dL|). 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, the Jones et al. (2010) study 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 soil Pb exposure is related to land use type and historical
exposures of the soil, as described in Section 3.6.1. For instance, Wu et al., (2010)
observed that bioavailable Pb concentrations in Los Angeles soils were significantly
associated with traffic-related variables and parcel age (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. (2011) observed that soil
Pb levels dropped following Hurricanes Katrina and Rita, from 329 mg/kg to 203 mg/kg
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when comparing samples from 2000 with those from 2006. The reduction in soil Pb was
thought to occur because sediments with lower Pb concentrations were distributed by the
storms and coated existing soils having higher Pb concentrations. At the same time, blood
Pb levels obtained from children ages 0-6 years at the same time periods declined by
1.55 (ig/dL for each 1% reduction in soil Pb (p < 0.05). Following this observation, that
lower Pb concentration material can be used to contain older soils with higher
concentrations of Pb, Mielke et al. (201 lb) initiated an experiment where a playground
was covered with geotextile material topped with 15 cm of river alluvium. Median soil
Pb concentrations decreased from 558 mg/kg to 4.1 mg/kg after the mitigation.
6.2.6 Residential Factors
A recent study of the association between blood Pb and housing factors by Dixon et al.
(2009), which drew inference from the NHANES national survey for 1999-2004, is
consistent with previous studies presented in the 2006 Pb AQCD that drew associations
between blood Pb and house dust (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 construction, floor surface condition, floor dust Pb level, windowsill dust Pb
level, and renovation in homes built before 1978. They found that blood Pb (log
transformed) was significantly associated with homes built after 1950 (p = 0.014),
windowsill Pb level (p = 0.002), dust Pb level (p < 0.001), and renovation in pre-1978
homes (p = 0.045). Detailed results of this regression are shown in Table 6-3. As part of
the same study, Gaitens et al. (2009) performed a regression analysis of floor dust Pb
(PbD) and windowsill dusts Pb on several factors. Floor dust Pb (log transformed) was
significantly associated with the following housing-related factors: floor surface
condition (p < 0.001), windowsill dust Pb (log transformed) (p < 0.001), year of
construction (p < 0.001), and renovation in apre-1950 home (p < 0.001). Windowsill
dust Pb (log transformed) was significantly associated with the following housing-related
factors: year of construction (p < 0.001), window surface condition (0.001), and
deteriorated indoor paint (p = 0.028).
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Table 6-3 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
Levels
Estimate (SE)
p-Value
Intercept
0.172

-0.517 (0.373)
0.172
Age (in years)
< 0.001
Age
2.620 (0.628)
< 0.001


Age'
-1.353 (0.354)
<0.001


Age3
0.273 (0.083)
0.002


Age4
-0.019 (0.007)
0.008
Year of construction
0.014
Intercept for missing
-0.121 (0.052)
0.024


1990—present
-0.198 (0.058)
0.001


1978-1989
-0.196 (0.060)
0.002


1960-1977
-0.174 (0.056)
0.003


1950-1959
-0.207 (0.065)
0.003


1940-1949
-0.012 (0.072)
0.870


Before 1940
0.000
—
PIR
< 0.001
Intercept for missing
0.053 (0.065)
0.420


Slope
-0.053 (0.012)
<0.001
Race/ethnicity
< 0.001
Non-Hispanic white
0.000
—


Non-Hispanic black
0.247 (0.035)
<0.001


Hispanic
-0.035 (0.030)
0.251


Other
0.128 (0.070)
0.073
Country of birth
0.002
Missing
-0.077 (0.219)
0.728


U.S.b
0.000
—


Mexico
0.353 (0.097)
<0.001


Elsewhere
0.154 (0.121)
0.209
Floor surface/condition * log floor PbD
< 0.001
Intercept for missing
0.178 (0.094)
0.065


Not smooth and cleanable
0.386 (0.089)
<0.001


Smooth and cleanable or carpeted
0.205 (0.032)
<0.001
Floor surface/condition * (log floor PbD)"1

Not smooth and cleanable
0.023 (0.015)
0.124


Smooth and cleanable or carpeted
0.027 (0.008)
0.001
Floor surface/condition * (log floor PbD)J

Uncarpeted not smooth and cleanable
-0.020 (0.014)
0.159


Smooth and cleanable or carpeted
-0.009 (0.004)
0.012
Log windowsill PbD
0.002
Intercept for missing
0.053 (0.040)
0.186


Slope
0.041 (0.011)
<0.001
Home-apartment type
< 0.001
Intercept for missing
-0.064 (0.097)
0.511


Mobile home or trailer
0.127 (0.067)
0.066


One family house detached
-0.025 (0.046)
0.596


One family house attached
0.000
—


Apartment (1-9 units)
0.069 (0.060)
0.256


Apartment (> 10 units)
-0.133 (0.056)
0.022
Anyone smoke inside the home
0.015
Missing
0.138 (0.140)
0.331


Yes
0.100 (0.040)
0.015


No
0.000
—
Log cotinine concentration (ng/dL) in blood
0.004
Intercept for missing
-0.150 (0.063)
0.023


Slope
0.039 (0.012)
0.002
Window cabinet or wall renovation in a
0.045
Missing
-0.008 (0.061)
0.896
pre-1978 home

Yes
0.097 (0.047)
0.045


No
0.000
-
/? = 2,155;/? = 40%.
Source: (Dixon etal.. 20091
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
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significantly associated with floor dust Pb concentration (p < 0.001). Paint scraping
within the last twelve months was nearly significantly 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 significantly associated with blood Pb concentration (p = 0.045). Mielke et al.
(2001) tested dust produced roughly six months after a home (having paint containing
-130,000 mg Pb/kg) was power sanded. Dust concentrations in various locations around
the house ranged from 390-27,600 mg/kg after the first sampling. Following several
rounds of cleaning, Pb concentrations dropped. Eight months later, most interior Pb
concentrations were < 3 mg/kg, but five sites ranged from 17-1,340 mg/kg. Two children
had blood Pb tested as part of this study. A 33 month-old child exhibited no change in
blood Pb before or after renovation (blood Pb = 4 j^ig/dL), while a twelve-month old child
had an increase from 4 (ig/dL to 12 (ig/dL. Note that blood was acquired in different
manners from the children because the younger child had smaller veins; the older child
had a venous sample while the younger child had a finger-stick. It is unclear if either of
these methods is more prone to contamination than the other. 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 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).
6.3 Factors Potentially Related to Increased Risk of Lead Induced
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. Table 6-4 provides an
overview of the factors examined and populations identified as "at-risk" of Pb-related
health effects based on the recent evidence integrated across disciplines. Each
characteristic is described in greater detail in the following subsections.
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Table 6-4 Summary of evidence for factors that potentially increase the risk of
lead-related health effects
Factor Evaluated
Potentially Related to Increased Risk
Age (Section 6.3.1)
Children
Sex (Section 6.3.2)
Males,® Females®
Genetics (Section 6.3.3)
ALADa, VDRa* DRD4a* GSTMf, TNF-aa, eNOSa, APOEa, HFEa
Pre-existing Disease (Section 6.3.4)
Autism3, Atopya,b, Hypertensionb
Smoking (Section 6.3.5)
Smokers'
Race/Ethnicity (Section 6.3.6)
Non-Hispanic Blacks', Hispanics3
Socioeconomic Status (SES) (Section 6.3.7)
Low SESa
Nutrition (Section 6.3.10)
Iron deficiency8
Stress (Section 6.3.11)
High stress®
Cognitive Reserve (Section 6.3.12)
Low cognitive reservea,b
Other Metals (Section 6.3.13)
High/co-exposure to Cda, Asa, Mna
'Evidence for this factor was limited.
'Possible mediator
6.3.1 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.
6.3.1.1 Children
According to the 2000 Census, 28.6% of individuals living in the U.S. were under the age
of 20, with 6.8% aged 0-4 years, 7.3% aged 5-9 years, 7.3% aged 10-14 years, and 7.2%
aged 15-19 years (SSDAN CensusScope. 2010a'). It is recognized that Pb can cross the
placenta and affect the developing nervous system of the fetus (Sections 4.2.2.4 and
5.3.9) 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 5.3.2.1). 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 5.
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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
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 5.6.2.1 and 5.6.3).
A study among Lebanese children examined the association between blood Pb levels and
transferrin saturation (TS) less than 12% and iron-deficiency anemia (IDA) (Muwakkit et
al.. 2008). A positive association was detected for blood Pb levels >10 j^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 iron deficiency with Pb toxicity.
Overall evidence indicates early childhood as a lifestage of increased risk for Pb-related
health effects. Both recent 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 (Fullmer et al.. 1985).
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6.3.1.2
Older Adults
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
2030, which is approximately 35 million and 71.5 million individuals, respectively)
(SSDAN CensusScope. 2010a; U.S. Census Bureau. 2010). As of the 2000 Census, 7.2%
of the U.S. population were ages 60-69, 5.8% were 70-79, and 3.3% were age 80 and
older (SSDAN CensusScope. 2010a).
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 interactions 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 j^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 j^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. 2003b). 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 multiple species
(mice and monkeys, for more details see Section 5.3.6). These pathologies include the
development of neurofibrillary tangles and increased amyloid precursor protein and its
product beta-amyloid (Basha et al.. 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.
6.3.2 Sex
The distribution of males and females in the U.S. is similar. In 2000, 49.1% of the U.S.
population was male and 50.9% was female. The distribution of sex varied by age with a
greater prevalence of females > 65 years old compared to males (SSDAN CensusScope.
2010a). 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).
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. (1987b)
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 attributable risks were greater among males than 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
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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 (Brubakeret al.. 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 (Jedrvchowski 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 (Songdei 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
exposure and meningiomas for males but not females (Raiaraman et al.. 2006). An
ecologic analysis of Pb pollution levels and cancer incidence among children reported
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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 (dc 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 etal. 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 (Lcasure 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 (SDN). 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
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
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exposure (pup blood Pb level 64 j^ig/dL at birth) induced reductions in SDN volume in
35% of Pb-exposed male rats (McGivem 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 (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 (Lcasure 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 (Lcasure 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.. 2009).
Multiple associations between Pb and various health endpoints have been examined for
effect measure modification by sex. Although not observed in all endpoints, some studies
reported differences between the associations for males and females, especially in
neurological studies. However, 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 in
which sex the associations are greater.
6.3.3 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.
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6.3.3.1
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 (Raj an 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 (Raian 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.
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
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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).
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).
6.3.3.2 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).
6.3.3.3 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.
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6.3.3.4 Apolipoprotein E
Apolipoprotein E (APOE) is a transport protein for cholesterol and lipoproteins. The gene
appears to regulate synapse formation (connections between neurons), which may be
particularly critical in early childhood. A genetic variant, called the APOE4 allele is a
haplotype between two exonic SNPs and is perhaps the most widely studied genetic
variant with respect to increasing risk of neurologic disease. A study of occupationally-
exposed adults observed tibia Pb levels to be associated with greater cognitive
decrements in some, but not all, neurobehavioral tests (such as digit symbol, pegboard
assembly, and complex reaction time) among adults with at least one APOE-s4 allele
(Stewart et al.. 2002). Conversely, in a study of children in Mexico, children without the
APOE-s4 allele had a greater inverse association between umbilical cord blood Pb and
Bayley's MDI than children with this allele, although the interaction term was not
statistically significant (Wright et al.. 2003a).
6.3.3.5 Hemochromatosis
The hemochromatosis (HFE) gene encodes a protein believed to be involved in iron
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.
6.3.3.6 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.
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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
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..
2006a). 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, APOE, 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.
6.3.4 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, atopy, diabetes, and hypertension.
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6.3.4.1
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).
6.3.4.2 Atopy
Atopy, a type of allergic hypersensitivity, was evaluated as a factor affecting risk in a
study of Pb and IgE (Annesi-Maesano et al.. 2003). The study examined hair Pb levels in
infants and IgE and reported a positive correlation overall. However, in stratified
analyses, this association remained only among infants of mothers without atopy. Among
mothers with atopy, the correlation was positive, although smaller, and was not
statistically significant.
6.3.4.3 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 was 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
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).
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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.
6.3.4.4 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).
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).
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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.
6.3.5 Smoking
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.
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 non-smokers (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 (Pi Lorenzo et al.. 2007). A
recent study of fertile and infertile men examined blood and seminal plasma Pb levels for
smokers and non-smokers (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 non-smoking fertile men and highest among smoking infertile men.
Prenatal smoking exposure was examined in a study of children's concurrent blood Pb
levels and prevalence of attention-deficit/hyperactivity disorder (APHP) 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 APHP (Froehlich et al.. 2009).
Overall, the studies have mixed findings on whether smoking modifies the relationship
between Pb levels and health effects. Future studies of Pb-related health effects and
current, former, and prenatal smoking exposures among various health endpoints will aid
in determining changes in risk by this factor.
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6.3.6 Race/Ethnicity
Based on the 2000 Census, 69.1% of the U.S. population is comprised of non-Hispanic
whites. Approximately 12.1% of people reported their race/ethnicity as non-Hispanic
black and 12.6% reported being Hispanic (SSDAN CensusScope. 2010b). Studies of
multiple Pb-related health outcomes examined effect measure modification 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
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.. 2007b). 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
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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 suggest that there may be race/ethnicity-
related increased risk for some 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.
6.3.7 Socioeconomic Status
Based on the 2000 Census data, 12.4% of Americans live in poverty (poverty threshold
for family of 4 was $17,463) (SSDAN CensusScope. 2010c). Ris et al. (2004) examined
modification of the associations between early-life Pb levels and Learning/IQ among
adolescents in the Cincinnati Lead Study. In models examining the association between
Pb and Learning/IQ, the prenatal and 78-month blood Pb levels were associated with
larger decrements in Learning/IQ in the lower two quintiles of socioeconomic status
(SES) (measured based on family SES levels).
6.3.8 Body Mass Index
In the U.S. self-reported rates of obesity were 26.7% in 2009, up from 19.8% in 2000
(Sherry et al.. 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.
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6.3.9
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).
6.3.10 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.
6.3.10.1 Calcium
Using the Normative Aging Study cohort, researchers examined the association between
Pb levels and hypertension, modified by calcium intake (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).
6.3.10.2 Iron
The 2006 Pb AQCD included studies that indicated individuals with iron-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, but this association was modified
by iron deficiency; girls with iron deficiency had a stronger inverse association between
Pb and inhibin B than those who were iron sufficient (Gollenberg et al.. 2010).
Toxicological studies also reported that iron-deficient diets exacerbate or potentiate the
effect of Pb. A study of pregnant rats given an iron-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
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higher damblood Pb level in the highest exposure groups (Singh et al. 1993; Saxcna et
al.. 1991). Thus, in this model, iron deficiency makes rat dams more at risk for Pb-
dependent embryo and fetotoxicity (Singh et al.. 1993).
6.3.10.3 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.
6.3.10.4 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).
6.3.11 Stress
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; Corv-
Slechta et al.. 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
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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 (Dvatlov and Lawrence. 2002).
Similar to studies of stress in animals, maternal self-esteem has also 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. The
investigators indicated that high maternal self-esteem may serve as a buffer against stress
by improving mother-child interactions and care giving practices, and maternal self
esteem may also be a surrogate of biological stress responses in the child.
Although examined in a limited number of studies, 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.
6.3.12 Cognitive Reserve
Cognitive reserve has been defined as "the maintenance of cognitive performance in spite
of ongoing underlying brain pathology" (Bleecker et al. 2007b). 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.. 2007b). 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.
6.3.13 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
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exposure and co-exposures with other metals. These studies, as well as toxicological
studies of these metals, are described below.
6.3.13.1 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
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
Fow ler. 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.
6.3.13.2 Arsenic
In a study of immune function among children living at varying distances from a Pb
smelter in Mexico, higher levels of both Pb and arsenic (As) were associated with greater
decreases in NO and greater increases in superoxide anion (Pineda-Zavaleta et al.. 2004).
6.3.13.3 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). Compared to children
with low blood Mn levels, those with high blood Mn levels had greater reductions in full
scale IQ and verbal IQ associated with increased blood Pb 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
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months, compared to decreases in neurodevelopment observed for increased Pb levels
with low levels of Mn (Claus Henn et al.). No interaction was observed between the two
metals and neurodevelopment at 24 months.
6.4 Summary
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. Recent
epidemiologic studies of infants/children detected increased risk of Pb-related health
effects, and this was supported by toxicological studies. However, this is based on a
limited number of epidemiologic studies, and more studies are needed for comparing
various age groups and examining adolescents.
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. 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.
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.
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. Similarly, the gap
between SES groups with respect to Pb body burden appears to be diminishing, with Pb
body burden being higher, but not appreciably higher, among lower income subjects.
Studies of race/ethnicity as a factor affecting risk indicate that some modification of
associations between Pb levels and health effects may be present. Compared to whites,
non-white populations were observed to be more at risk of Pb-related health effects;
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. Although limited
by the number of studies, individuals with lower SES appear to represent an at-risk
population. A study of Pb and IQ reported greater inverse associations among those in the
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lowest SES groups. Additionally, there is evidence associating proximity to areas with Pb
sources, including urban areas with large industrial sources, with increased Pb body
burden and risk of Pb exposure.
Various genes were examined as 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, APOE, and
HFE, although the number of studies examining effect measure modification by these
genes was small.
Evidence for other factors (pre-existing diseases/conditions, smoking, BMI and/or
obesity, alcohol consumption, nutritional factors, stress, cognitive reserve, and co-
exposure with other metals) was limited regarding their effect on Pb-related health
outcomes. 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
and atopy as potential factors affecting Pb-related health effects; differences were
observed but few studies were available to examine these factors. Recent 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.
BMI, alcohol consumption, and nutritional factors were also examined in recent
epidemiologic and toxicological 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. Among nutritional factors, those with
iron deficiencies were observed to be an at-risk population for Pb-related health effects in
both epidemiologic and toxicological studies. Other nutritional factors, such as calcium,
zinc, and protein intake, demonstrated the potential to modify associations between Pb
and health effects in toxicological studies. Recent epidemiologic studies of these factors
were either not performed or observed no effect modification. Folate was also examined
in an epidemiologic study of birth size but no interaction was reported between Pb and
folate. Stress was evaluated as a factor that potentially increases the risk of Pb-related
health outcomes and although there were a small number of recent epidemiologic studies,
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1	increased stress was observed to negatively impact the association between Pb and health
2	endpoints. Toxicological studies supported this finding. An epidemiologic study
3	evaluated cognitive reserve as a modifier of the associations between Pb and cognitive
4	and motor functions. Cognitive reserve was an effect measure modifier for the
5	association between Pb and cognitive function but not motor function. Finally,
6	interactions between Pb and co-exposure with other metals were evaluated in recent
7	epidemiologic and toxicological studies of health effects. High levels of other metals,
8	such as Cd, As, and Mn, were observed to result in greater effects for the associations
9	between Pb and various health endpoints.
10
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Tong. S: McMichael. AJ: Baghurst. PA. (2000). Interactions between environmental lead exposure and
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CHAPTER 7
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 four sections. The
introduction (Section 7.1) presents the organizing principles of this chapter and several
important general ecology concepts. Section 7.2 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 7.3,
including water-only exposures and sediment related effects. Both the terrestrial and
aquatic sections conclude with a discussion of alterations in ecosystem service functions
as a consequence of Pb deposition. Finally, an integrative synthesis of effects of Pb
across biota and causal determinations for effects of Pb in both terrestrial and aquatic
systems are presented in Section 7.4. Although terrestrial and aquatic ecosystems are
considered separately in this chapter, the deposition of Pb to land and water and the
subsequent flux of Pb through watersheds are interconnected by fate and transport
processes discussed in Section 3.3. 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.
7.1 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 (Fairbrother et al.. 2007). This section introduces
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critical concepts for understanding how Pb from atmospheric deposition may affect
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.
7.1.1 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 (IPC'C. 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
(IPC'C. 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. 1998).
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 form populations, and populations of
different species form communities. 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 etal.. 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
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behavioral changes that alter breeding behaviors or predator avoidance. However, effects
on organisms must result in changes to their survival or reproductive output to have any
effect on the population. 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 (Bamthouse. 2007). 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 7.1.2) and critical loads (Section 7.1.3)
serve as benchmarks or measures of the impacts of pollutants on ecosystems.
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7.1.2 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 (Dailv. 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 ("WRI. 2000; Dailv. 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; Avensu et al.. 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, composition and function can be affected (Dailv and
Ehrlich. 1999; Wall. 1999; Chapin et al.. 1998; Lcvlin. 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).
Hassan (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 02, soil formation and retention, nutrient cycling, water cycling and
provisioning of habitat. Biodiversity is a supporting service in that it is
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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 (Ciitav 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.
7.1.3 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;
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¦	Bioavailability and bioaccumulation rates;
¦	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. 20086). 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.
7.1.4 Ecosystem Exposure, Lag Time and Re-entrainment of Historically
Deposited Lead
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 7.2.2.
and 7.3.2). However, there is frequently a lag time between when metals are emitted and
when an effect is seen, particularly in terrestrial ecosystems and, to a lesser extent, in
aquatic sediments; water exposures result in more immediate system responses. This is
because the buffering capacity of soils and sediments permits Pb to become sequestered
into organic matter, making it less available 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 from the time when the Pb was initially added to the system. Excluding erosion
processes, the time required to achieve 95% of steady state is about 4 half-lives (t12)1
(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 is dependent 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:
1 Time required to reduce the initial concentration by 50% if metal input is zero.
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0.69 xdx 10,000
Equation 7-1
1	where:
2	d is the soil depth in meters (m)
3	y is the annual crop yield (tons/ha-yr)
4	TF is the ratio of the metal concentration in plant to that in soil
5	R is the net drainage loss out of the soil depth of concern (m3/ha-yr)
6	P is the bulk density of soil [kg(dry weight)/L]
7	Kd is the ratio of the metal concentration in soil to that in soil pore solution (L/kg)
8	Metals removed by crops (or plants in general) comprise a very small fraction of the total
9	soil metal and can be ignored for the purpose of estimating time to steady state. Thus,
10	equation 7-1 is simplified to:
^ _ 0.69 xdx 10,000
fl/2 —	jT
pKd
Equation 7-2
11	and becomes a function of soil depth, the amount of rainfall, soil density, and soil
12	properties that affect Kd. Pb has a relatively long time to steady state compared to other
13	metals, as shown in Table 7-1.
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Table 7-1
Comparison among several metals: Time to achieve 95% of steady
state metal concentration in soil; example in a temperate system
Metal
Loading rate (g/ha/yr)
Kd (L/kg)
Time (yr)
Se
100
0.3
1.3
Cu
100
480a
1,860a
Cd
100
690a
2,670a
Pb
100
19,000a
73,300a
Cr
100
16,700a
64,400a
aMean Kd (ratio of total metal concentrations in soils to that in soil pore water); and Time to achieve 95% of steady-state concentration in soil. (49
Dutch soils) fde Groot et al.. 19981.
Note: Based on a soil depth of 23 cm, a rain infiltration rate of 3,000 m3/hayr, and the assumption that background was zero at the start of loading.
Source: Smolders, Fairbrother et al. (2007)
1	In aquatic systems, t1/2 for Pb in the water column depends on the ratio of the magnitudes
2	of the fluxes coming from and going into the sediment, the ratio of the depths of the
3	water column and sediment, and the sediment half-life. Sediment t12 is dependent upon
4	the particulate and dissolved fractions and is calculated as for soils (Equation 7-2).
5	Re-entrainment of Pb particles via windblown dust from surface soils or dry sediments
6	may occur. Amount and distance of re-entrained particles and deposition rates are
7	dependent upon wind velocity and frequency; size, density, shape, and roughness of the
8	particle; soil or sediment moisture; and terrain features including openness (including
9	amount of vegetation), aspect relative to wind direction, and surface roughness.
10	Resuspension is defined in terms of a resuspension factor, K, with units of m1, or a
11	resuspension rate (A), with units of sec"1 (Equation 7-3). The resuspension rate, A, is the
12	fraction of a surface contaminant that is released per time and is defined by:
A = *
c
Equation 7-3
13	where:
14	R is the upward resuspension flux (|_ig/m2/scc)
15	C is the soil (or dry sediment) Pb concentration (|ag/nr)
16	Such emissions may have local impacts, but are not likely to have long-range effects, as
17	particles generally remain low to the ground and are not lifted into the upper atmosphere.
18	Although re-entrainment may alter the particle size distribution in a local area, it
19	generally does not alter the bioavailable fraction, and deposited particles will be subject
20	to the same biogeochemical forces affecting bioavailability. Therefore, exposure via re-
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entrainment should be considered additive to exposure from atmospheric particulate
deposition in terrestrial and aquatic ecosystems.
7.2 Terrestrial Ecosystem Effects
7.2.1 Introduction to Terrestrial Ecosystem Effects
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 this ISA is on the effects of Pb on terrestrial organisms including plants, invertebrates
and vertebrates, with particular focus on current ambient level. 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. Data on levels of Pb in U.S. soils are reviewed in Section 3.6.1 and summarized in
Table 2-1. 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 7.2 is organized to consider
uptake of Pb and effects at the species level, followed by community and ecosystem level
effects. Soil biogeochemistry of Pb is reviewed in Section 7.2.2 Section 7.2.3 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 7.2.4) are followed by data on
exposure and response of terrestrial species (Section 7.2.5). Effects of Pb at the
ecosystem level of biological organization are discussed in Section 7.2.6 Section 7.2
concludes with a discussion of critical loads in terrestrial systems (Section 7.2.7), soil
screening levels (Section 7.2.8), characterization of sensitivity and vulnerability of
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ecosystem components (Section 7.2.9), and effects on ecosystem services
(Section 7.2.10).
7.2.2 Soil Biogeochemistry and Chemical Effects
According to data presented in the 2006 Pb AQCD, 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 small relative to bulk soil concentration. The
amount of Pb dissolved in soil solution is controlled by at least six variables: (1)
solubility equilibria; (2) adsorption-de sorption 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.
7.2.2.1 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 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. 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.
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 (Douav 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').
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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 (Acosta et al.). 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 relative bioavailability 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.
7.2.2.2 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).
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 fiilvic
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
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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. Dayton et al. (2006) established
significant negative correlations between log-transformed Pb content of lettuce plants
(Lactuca sativa), soil organic content, and CEC, and similar negative relationships were
also confirmed between soil pH and amorphous Fe and Al oxide content. 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 of Lolium 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 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 etal.. 2011).
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7.2.2.3
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 7.2.3), 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 individual simulated soil components freshly spiked with 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 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
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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 et al. 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.
7.2.3 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
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" (Fairbrother et al.. 2007). 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
bioaccessible fraction of metal. As stated in the Framework for Metals Risk Assessment
(U.S. EPA. 2007c). the bioaccessible fraction of a metal is the portion (fraction or
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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 McGeer et al. (2004) and the
Framework for Metals Risk Assessment summarizes metals bioavailability and
bioaccumulation in aquatic, sediment and soil media (Figure 7-1).
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Bioaccessible Fraction (BF)a:
Percent soluble metal ion
concentration relative to total
metal concentration (measured in
solution near biomembrane)
Relative Bioavailability (RBA)b:
Percent adsorbed or absorbed
compared to reference material
(measure of membrane dynamics)
Absolute Bioavailability (ABA)0
Percent of metal mass absorbed
internally compared to external
exposure (measures systemic
uptake/accumulation)
Bioaccessibility
ioavailability
Environmental availability
Exp
Bioaccumulation df metal
==========-» Effects
Ylembrane
Cationic
Toxicoloical
accumulation
Benign
accumulation
uptake
competition
Uptake
¦~I Soluble speciesA
Soluble species B
Soluble species C |
TM.:u n ¦
Internal
Transport
and
Distribution
Uptake
Site of
Toxic
Action
Other dissolved
species
Uptake
Nonavailable
metal
Dietary intake
Essentiality
Detoxification
and Storage
Particulate fraction |
Soluble fraction
Excretion
Physiological
membrane
Predation
Foraging
Bioaccumulated Metal
Total Metal Concentration
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).
CABA 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 7-1 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 7-2), 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 7.3.4).
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Competing Cations
M-Biotic
Ligand
M-DOC
Free
Metal
Ion
Site of Action
Organic Matter
Complexation
MOHh
MHCO.
MCI+
Inorganic Ligand
Complexation
Source: Reprinted with permission of John Wiley and Sons ; from Di Toro et al. (20011
Figure 7-2 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 3, while new information on the
influence of soil biogeochemistry on speciation and chemical lability was presented in
Section 7.2.2. This section summarizes new literature on uptake and subsequent presence
of Pb in tissues. The 2006 Pb AQCD 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) for some terrestrial and aquatic biota. 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).
7.2.3.1 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
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AQCD where it was stated that most plants cannot survive in soil containing 10,000 jj.g/g
(mg Pb/kg) dry weight if the pH is below 4.5 and the organic content is below 5% (UJL
EPA. 1986b). Pb can be absorbed across the leaf surface into internal plant tissues
although the vast majority of uptake is via roots (U.S. EPA. 1986b). The 2006 Pb AQCD
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 3.6.6.
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).
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 apoint source (0.1, 0.2, 5.0 km). Pb isotope
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
shoot:root 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, and Rostellariaprocumbens, ratio = 1.7).
However, the two species studied at Site II were not studied at Site I or Site III. In the
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control site (Site III), no plant was found with a Pb shoot:root 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 root:shoot 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. (2011) 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 0.41(0.0007-
0.17) for leafy vegetables, such as spinach, Chinese cabbage and celery (Zhang et al.
2011). 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 PbC03,
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 dactylifera) 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 |_ig Pb/g dry
weight, with the highest Pb date concentrations detected near freeways and industrial
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areas (Aldiain et al.. 2011). Likewise, Pb concentrations in rosemary (Rosmarinus
officinalis) flowers, stems, and leaves were significantly higher in the urban areas of Al-
Mafiraq 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 lead (El-
Rioob 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 lead
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. (2009b) 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. 2009b). 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.. 2009a). Similarly, Kuang et al. (2007)
noted that the Pb concentration in the inner bark of Pinus 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
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. EP A. 1986b). The advent of laser ablation inductively coupled plasma mass
spectrometry has made measurement of Pb concentration in individual tree rings possible
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(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. (2008b) 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 tridentata). In addition, species such as bristlecone pine (Pinus
aristata), old-growth redwood (Sequoia sempervirens), and giant sequoia (S. gigantea)
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 (Guvette 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)
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
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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
dactylifera) 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) (Fuiiwara et al.. 201IV However, decreases in tissue
Pb concentration with increasing distance from point 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
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 (1.12 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
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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, and Macrolepiota procera, 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 delica) 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
(Macrolepiota procera) to 0.45 (Lepista nuda) (Campos and Teiera. 201IV
Translocation and Sequestration of Lead 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. 1977). The 2006 Pb AQCD 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. 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 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 (Vicia
faba) cultivated in mine tailings (average Pb concentration of 7,772 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.
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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 [Glycine 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, (C'ho et al.. 2009) showed that green onion (Allium
fistulosum) 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. (2011c) 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.. 2011c). Sonmez et al.
(2008) reported that Pb accumulated by three weed species (Avena sterilis, Isatis
tinctoria, 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 incarnatum)
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 etal.. 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 et al.. 2011).
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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) (Tarnura 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 etal.. 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)
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
hieraciodides (Pteridaceae) tissues collected from Jinding. These results indicate
significant interspecies differences in Pb uptake, as well as potential soil-specific
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differences in Pb bioavailability. S. vestita, in particular, was determined 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
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).
7.2.3.2 Terrestrial Invertebrates
At the time of publication of the 2006 Pb AQCD, 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
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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 Qloghobo. 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
I,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
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 |_ig Pb/g 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 aspera snails to standardized
(International Organization for Standardization methodology [ISO 11267:1999])
artificial-substrate soils 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 |ag Pb/g 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
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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 physical properties that
were freshly spiked with Pb to give a standard concentration of 2,000 mg Pb/kg dry
weight. Both internal earthworm Pb concentration and mortality rates increased with
decreasing pH and CEC although the apparent role of CEC may only have been due to its
correlation with other soil characteristics. These data corroborate that Pb bioavailability
and toxicity are increased in acidic soils and in soils with a low CEC (Section 7.2.2). 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 iron 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 reach depths
of 6 inches). 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
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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..
2009V
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
(Viiver 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 toxicologically 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. 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
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content, although abundance was unaffected (Schippcr et al.. 2008). By contrast, two
grasshopper species inhabiting Pb and Cd-contaminated areas near zinc 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 et al.. In Press). 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 0. 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.
7.2.3.3 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)I.
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 nigra) 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
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59 fj.g 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.
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 (Nakavama et al.. 2011). Angelova et al. (2010) reared rabbits on a fodder
mixture containing lead-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 (Sits scrofa) and Sprague-Dawley rats (Rattus norvegicus) reared on Pb-
contaminated soil (5% of 1,000 |_ig Pb/g soil as dietary component) consumed
significantly different amounts of Pb. Over a 30-day period, rats consumed an average of
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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 (.ig Pb/g
for micropigs, and 0.2, 7, 0.5, and 1.5 |_ig Pb/g 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 cerussite
and manganese-Pb oxides and poor absorption of 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 Ca deficiency (0.45 mg Ca 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
were roughly doubled by a Ca-deficient diet. However, it is not known whether this level
of dietary Ca deficiency is common in wild populations of birds.
7.2.3.4 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).
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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. litura-ied
chickens (Gallus gallus domesticus) accumulated 0.58 mg Pb/kg and 3.6 mg Pb/kg in
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, an insectivorous bird species, the black-tailed godwit (Limosa limosa) 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 robur) 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 deer tissues with attenuation (i.e., decreasing concentration) of Pb up the
food chain. Interestingly, soil geochemistry likely was affected by mining activity as
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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 of fence
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
available to predators, but returns to background level with each shedding, while granular
Pb is likely unavailable, at least to other invertebrates (Viiver et al. 2004).
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.
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7.2.4 Biological Effects
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. In this section, the observed biological effects caused by
exposure to atmosphere-derived Pb will be discussed, while the results of dose-response
experimentation will be addressed in Section 7.2.5. 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
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 (ATS DR. 2004V
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.
7.2.4.1 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 Lorcnc-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. EP A. 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
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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. ("2008c') demonstrated that Pb exposure interfered with and decreased light
absorption by spinach (Spinacia oleracea) plants. Spinach seeds were soaked in 5, 12, or
25 mM PbCl2 for 48 hours prior to germination, and following 42 days of growth, plants
were sprayed with PbCl2 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 0.1 to 0.3 mg Pb/g dry weight (Ahmad et al.). 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 exposures of 0.1 and 0.5 mM Pb 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 amblyoclada lichen was exposed to aqueous Pb solutions of 0.5, 1,
5, and 10 mM Pb nitrate; 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 lead at the same rate as slower growing mosses, indicating that
moss growth allowed for further lead uptake, rather than a "dilution" effect (Remoter 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 (Carre ras 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
(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 Hypnum
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plumaeforme mosses was decreased by 5.8% following exposure to 10 mM Pb, while
lower exposures slightly elevated chlorophyll content.
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 (Vicia faba) (Wang et al..
2010c; Wang et al.. 2010b; 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 etal. 2011a). 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
•02~ and H202, although no single moss species could be identified as most sensitive to
Pb exposure (Sun et al.. In Press). 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 exposed to Pb solutions of 200, 500, and 800 mg Pb/kg exhibited
increased antioxidant activity: at exposures of 800 mg Pb/kg, root and shoot SOD activity
increased to 2-3 times that of controls, and induction was slightly higher inM 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 in M. 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 of Sedum alfredii (an Asian
perennial herb), one an accumulator of Pb and the other not. Glutathione level was
increased in both, and root and shoot lengths were decreased following long-term
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exposures to Pb up to 200 (.iM. 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, 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 tremula) trees in eastern Slovakia were chemically analyzed for metal
concentration arising from exposure to smelter pollution (Dzubai 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, C02 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).
Growth
Evidence of effects of Pb on higher growth processes in terrestrial plants was reported in
early NAAQS reviews. 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 photosynthetic rate, inhibition of respiration,
cell elongation, root development or premature senescence (U.S. EPA. 1986b). All of
these effects were observed to occur in isolated cells or in plants grown hydroponically in
solutions comparable to 1 to 2 |ag Pb/g soil or in soils with 10,000 mg Pb/kg or greater
(U.S. EPA. 1986b). Pb effects on 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.
Both growth and carotenoid and chlorophyll content of Brassica juncea (mustard) plants
were negatively affected by Pb exposure (John et al.. 2009). Pb treatments of 1,500 (.iM
(as Pb-acetate solution) decreased root lengths and stem heights by 50% after 60 days.
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Exposure to 600 (.iM Pb and greater decreased carotenoid content, while chlorophyll a
was decreased at Pb exposures of 450 (.iM 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), 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 686 (.ig 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).
Further, 14-day growth bioassays conducted with lettuce seedlings (Latuca sativa) and
field-collected Pb-arsenate contaminated soil produced an unbounded NOEC value of
390 mg Pb/kg (and 128 mg As/kg) (Delistratv and Yokel. In Press).
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, 4, and 8 mM Pb/kg dry weight total Pb and collected leaf samples for 11 days. At
exposures of 4 and 8 mM Pb/kg, leaf nitrite content was decreased by 29% and 20%,
while nitrate content was affected only at the highest Pb exposure (70% of control
levels). 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
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).
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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 nM of Pb nitrate 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 10,
100 and 1,000 (.ig Pb/L as Pb sulfate (Ahmad et al.). 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 (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 and iron and decreased sulfur concentration in radish tops.
Interestingly, as in zebra finch (Section 7.2.3.3) Ca 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 Ca concentration of 1.2 mM
and higher (Antosiewicz. 2005).
7.2.4.2 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.
In addition to these endpoints, recent literature also indicates that Pb exposure can cause
significant neurobehavioral aberrations, and in some cases, endocrine-level impacts.
Second-generation effects have been observed in some invertebrate species.
The morphology of y-aminobutyric acid (GABA) 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 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. Younger
C. elegans larvae were more likely to exhibit neurobehavioral toxicity symptoms in
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response to Pb exposure (2.5 (.iM) (Xing et al.. 2009c). 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..
2009b). C. elegans nematodes exposed to Pb concentration as low as 2.5 (.iM for 24 hours
also exhibited significantly altered behavior characterized by decreased head thrashes and
body bends. Exposures of 50 (.iM Pb 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
(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 \\M. Vigneshkumar et al. (In Press) 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 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 to 2.5, 75, and 200 (.iM Pb nitrate
exhibited significant indications of multi-generational toxicity (Wang and Peng. 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 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 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. 2009a).
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
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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 lead-arsenate pesticides,
exhibited no change in survival, behavior or morphology (Dclistratv and Yokel. In Press).
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
(Maitv etal.. 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
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 (50 to
15,000 |_ig Pb/g diet dry weight) (El-Gendv et al.. 2011). A five week dietary exposure to
1,000 |_ig Pb/g and greater resulted in reduced snail growth. Decreased food consumption,
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growth, and shell thickness were also observed in juvenile A. achatina snails exposed to
Pb-contaminated (concentration greater than 134 mg Pb/kg) diet for 12 weeks (Ebenso
and Qloghobo. 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
Qloghobo. 200%). 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) (Owoiori et al.. 2011).
Consequently, it is unlikely that oribatid mites will avoid soils containing toxic Pb
concentrations.
Lock et al. (2006) compared the toxicity of both laboratory-spiked soils and field-
collected Pb-contaminated soils to springtails (/¦'. 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 had no significant effect on Sinella
curviseta springtail survival or reproduction (Xu et al.. 2009b).
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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 F1 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 (Gorur. 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 were demonstrated to decrease with increasing Pb
exposure. Conversely, in a study by Zheng and Li (2009). vitellogenin mRNA was
detected at higher levels in males exposed to 12 and 25 mg Pb/kg, although vitellogenin
levels were not affected. Similarly, the sperm morphology of the Asian earthworm
(Pheretima guillelmi) was found to be altered significantly following 2-week exposure to
soils containing 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 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).
7.2.4.3 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
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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 lead 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 lead 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
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, 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
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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
(Chiesa et al.. 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 tissue concentration in field-collected urban
blackbirds (Turdus merula) were determined to be 3.2 mg Pb/kg, 4.9 mg Pb/kg, and
0.2 mg Pb/kg wet mass in breast feathers, washed tail feathers, and blood, respectively
(Scheifler et al.. 2006a). Although these levels were significantly higher than those
measured in rural blackbirds, elevated Pb tissue concentration was not significantly
correlated to any index of body condition. On the other hand, Hargreaves et al. (2010)
showed that Pb tissue concentration of female arctic shorebirds was negatively correlated
with reproductive success. Maternal blood Pb levels were negatively associated with
hatching success in black bellied plovers (Pluvialis squatarola) and ruddy turnstones
(Arenaria interpres), and with nest duration in all species tested. There was no significant
correlation between adult whole-blood or feather Pb concentration and Pb levels in
produced eggs.
The long-term effect of atmospheric Pb deposition on pied flycatcher (Ficedula
hypoleuca) nestlings was determined in native communities residing in the Laisvall
mining region of Sweden (Berglund et al.. 2010). Moss samples indicated that Pb
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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 (Beralund 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 j^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. (201 la)
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 Swanson'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).
Blood Pb was significantly elevated in waterfowl in the Lake Coeur dAlene 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 |_ig Pb/g); 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 |_ig Pb/g) (Spears et al.. 2007). The level
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of corticosteroid hormones in field populations of white stork nestlings (Ciconia ciconia)
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 91 jj.g Pb/L (±51), which was higher than blood Pb levels from the
mining area (44 ± 34 (.ig Pb/L). 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 (Larus 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 coturnix japonica) to 5,000 and
50,000 (.ig 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 E.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 5.6.4.1.
Again, dietary or other health deficiencies unrelated to Pb exposure are likely to
exacerbate the effects of Pb. Ca-deficient female zebra finches (T. guttata) had a
suppressed secondary humoral immune response following 28-day exposures to
20,000 (.ig Pb/L in drinking water (Snoeiis et al.. 2005). This response, however, was not
observed in birds fed sufficient Ca. 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
function of gilts (S. domestica). 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
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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 (Sits scrofa) inhabiting a Pb-contaminated mining area in
Spain exhibited increased liver and bone Pb concentrations (geometric means of 0.35 and
0.46 (.ig Pb/g for red deer, and 0.81 and 7.36 (.ig Pb/g 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.
Pb-treated oocytes of buffalo (Bubalus bubalis) assessed in vitro at concentrations
ranging from 0.5 to 1,000 (ig/dL in one-day cultures indicated a significant decline in
viability of oocytes at 100 (ig/dL (Nandi et al.. 2010). 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 250 (ig/dL
and 100% oocyte death at 3,200 (ig/dL. Similarly, the reproductive viability of the 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.. 2009b).
7.2.5 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 biota at the
organism, population, or community level, 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. No new exposure-
response information was available for plants, birds, or mammals.
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) (Maitv
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
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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 of E. fetida earthworms showed concentration-
dependent relationships with soil Pb concentration during the course of standard 14- and
56-day toxicity tests (Jones et al. 200%). 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. 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.
Other studies have shown no correlation between Pb concentration in either earthworm
tissue or soil, and earthworm survival rate. Although the Pb content of E. 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
fNahmani 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 for E. fetida, another standard
laboratory test species. In the more recent study of E. fetida sensitivity summarized
above, Jones (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
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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, Sinella coeca and F.
Candida. Pb concentration 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 of F. 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.. 2009d). 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 7.2.2, 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 of 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
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root phytochelatins) or sequestration of Pb (such as granule formation in root tissues and
earthworms, or substitution of Pb for calcium in bones).
7.2.6 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 (.ig Pb/g dry weight [mg Pb/kg]) can eliminate populations of bacteria and fungi
on leaf surfaces and in soil. At soil concentrations of 500 to 1,000 |_ig Pb/g (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, natural terrestrial ecosystems near significant Pb point
sources (such as smelters and mines) 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 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
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communities may prevent these ameliorating effects. R.Y. Yang et al. (2008b) found that
both the mycorrhizal colonization and the growth of Solidago 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 etal.. 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,
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-
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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 (-4.5 mg/g dw).
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, (Khan et al.. 2010c) reported that it is possible for indicators of microbial
activity to recover after an initial period depression. (Khan et 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. Basal respiration was decreased in two soil types
tested at the highest Pb treatment (32 mM), and in a third at the two highest Pb treatments
(8 and 32 mM). 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 type of soils; concentrations between 5 and 50 nM Pb
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significantly decreased phenol oxidase activity in all soils tested, while 400 nM and
greater completely arrested phenol oxidase activity in one soil tested (a high pH sandy
loam) (Carine et al.. 2009). Carine (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 et al.. 2010b). Further, the weekly soil carbon dioxide evolution rate was
significantly reduced by concentrations of 5, 10, and 50 mg Pb/g, 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 |_ig Pb/g),
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 etal.. 2011). Authors also
noted a strong correlation between microcalorimetry estimates and the number of colony
forming units isolated from soil samples.
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 (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.
(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.
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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
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. 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,
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and biomass of testate amoebae communities grown on Sphagnum fallax mosses were
significantly decreased following moss incubation in Pb solutions of either 625 or
2,500 (.ig Pb/L (Nmiven-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 cyanobacteriabiomass (Meyer et al.. 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 limosa) 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
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.
7.2.7 Critical Loads in Terrestrial Systems
The general concept and definition of critical loads is introduced in Section 7.1.3 of this
chapter [also see Section 7.3 of the 2006 Pb AQCD (U.S. EPA. 2006c)l. 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
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quantified varies. Complexation, sorption, ion exchange and precipitation are well
understood in the lab, 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 section, 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 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 jj.g Pb/dL blood; HC5 values for birds ranged from
65 to 71 |ag Pb/dL. The authors proposed the use of 18 and 71 jj.g 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 physiochemical 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
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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.
There are few critical loads for Pb reported for terrestrial ecosystems in the U.S.;
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, the Pennines, and southern Wales), 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 et al. (2009) similarly observed that the uncertainty inherent in a critical
load approach to Pb risk 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).
7.2.8 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
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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.
7.2.9 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 7.2.2), 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.
7.2.9.1 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. Similarly, 28-day EC50 values derived for !•'. Candida
collembola (springtails) were between 2,060 and 3,210 mg Pb/kg in Pb-spiked soils
(Lock et al.. 2006). while the springtail species S. curviseta exhibited no response to a 28-
day exposure to 3,200 mg Pb/kg Pb-spiked soil (Xu et al.. 2009b). 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.. 2009c'). and also demonstrated more pronounced neural degeneration than
older larvae and adults (Xing et al.. 2009b).
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7.2.9.2
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 describes how Ca, 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, Ca deficiencies may increase the
susceptibility of different terrestrial species to Pb, including plant (Antosiewicz. 2005).
avian (Dauwe et al.. 2006; Snoeiis et al.. 2005) and invertebrate species. Antosiewicz
determined that, for plants, Ca deficiency decreased the sequestration capacity of several
species (tomato, mustard, rye, and maize), and that this likely resulted in an increased
proportion of Pb at sites of toxic action. Because Pb ions can interact with plant Ca
channel pores, in the presence of low Ca 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 Ca [Simkiss et al. (1982). as cited in Jordaens et al. (2006)1. Hence,
in environments with disproportionately high Pb versus Ca concentration, accumulation
of Pb may be accelerated, as in plants. Ca deficiency in birds was demonstrated to
stimulate the production of Ca-binding proteins in the intestinal tract, which extract more
Ca 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)1.
7.2.9.3 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, the proportionality of toxicity to
Pb content disappeared following experimental aging of freshly spiked soils through
leaching. The leaching-aging factor for Pb was determined to be 4.2, and represented the
ratio of EDio values derived in aged soils to freshly spiked soils (factors greater than one
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indicate decreased toxicity in aged field soils relative to laboratory spiked soils).
Consequently, the sensitivity of terrestrial vertebrates to environmental Pb exposures will
be heavily dependent on the relative rate of aging and site-specific bioavailability.
7.2.9.4 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.
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. (2007V 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 S02 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.
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7.2.10
Ecosystem Services
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. The
evidence reviewed in this ISA illustrates that Pb can cause ecological effects in each of
the four main categories of ecosystem services (Section 7.1.2) 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 cadmium and Pb concentrations
(Mimila 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 (kwartimikov et al. 1999). while no effects were found in ingestion studies
of Pb with willow leaf beetle, Lochmaea caprae (Rokvtova et al.. 2004) mottled water
hyacinth weevil, Neochetina eichhorniae (Kav and Haller. 1986) and hairy springtail,
Orchesella cincta (Van Capelleveen et al.. 1986).
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
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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.
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.
7.2.11 Summary of Effects in Terrestrial Systems
This summary of the effects of Pb on terrestrial ecosystems covers information from the
publication of the 2006 Pb AQCD to present. Refer to Section 7.4: Causal
Determinations for Ecological Effects of Lead for a synthesis of all evidence dating back
to the 1977 Pb AQCD considered in determining causality.
7.2.11.1 Biogeochemistry and Chemical Effects
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.
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(2009) demonstrated that the two most important determinants of both solubility and
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.
7.2.11.2 Bioavailability and Uptake
Plants
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 new 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. 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.
Invertebrates
Since the 2006 Pb AQCD, various species of terrestrial snails have been found to
accumulate Pb from both diet and soil. New 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 may result from layer differences in interacting factors such as pH and CEC.
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
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arthropods collected at contaminated sites, recent studies found gradients in accumulated
Pb that corresponded to gradients in soil with increasing distance from point sources.
Vertebrates
There were few new studies of Pb bioavailability and uptake in birds 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 point sources,
and none was able to identify the origin of the Pb. 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
showed that the form of Pb determined accumulation.
Food web
New 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.
7.2.11.3 Biological Effects
Plants
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,
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reduced growth was observed in some experiments, as well as genotoxicity, decreased
germination, and pollen sterility.
Invertebrates
Recently published studies have shown neuronal damage in nematodes exposed to low
concentrations of Pb (2.5 |iM). 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, 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, 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 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. Evidence suggested that aging and pH are important
modifiers.
Vertebrates
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. 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. Effects of
dietary exposure were studied in several mammalian species, and cognitive, endocrine,
immunological, and growth effects were observed.
7.2.11.4 Exposure Response
Evidence reviewed in previous sections demonstrates clearly that increased exposure to
Pb is generally associated with increases in observed effects in terrestrial ecosystems. It
also demonstrates that many factors, including species and various soil physiochemical
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properties, interact strongly with 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, a few studies of exposure-
response have been conducted with earthworms, and results have been inconsistent.
7.2.11.5 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
effects of 6 metals individually (Akcrblom 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.
7.2.11.6 Critical Loads, Sensitivity and Vulnerability
Over the longer term, terrestrial systems may be more affected particularly by those
metals with a long soil residence time, such as Pb. Exploratory studies of critical load
approaches for risk assessment for Pb have been recently conducted in the U.K., the
Netherlands, and Italy. Their authors suggested that the main limitations of critical loads
approaches in those countries were gaps and uncertainty in both ecotoxicological and Pb
deposition data. The most visible indication of the need for improvement was that critical
load values were often below background values. Smolders (2009) suggested that
correcting for aging and other interacting factors would likely raise predicted-no-effect
concentrations, and others proposed basing risk management on tissue residue in
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organisms, or creating extraction methods that more closely mimic uptake and
accumulation.
Given the heterogeneity of ecosystems affected by 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 within the U.S.
Recent studies have addressed differences in sensitivity explicitly, and clearly
demonstrated high variability between related species, as well as within larger taxonomic
groupings. 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).
Protective effects of dietary Ca have been found in plants, birds, and invertebrates.
7.3 Aquatic Ecosystem Effects
7.3.1 Introduction to Aquatic Ecosystem Effects
This section of the Pb ISA reviews the new literature since the 2006 Pb AQCD (U.S.
EPA. 2006c') on the effects of Pb on freshwater and marine ecosystems. Freshwater and
marine/estuarine systems are considered separately due to differences in Pb speciation,
bioavailability of Pb, and salinity as modifying factors for Pb toxicity. The focus is on the
effects of Pb, with particular focus on ambient level, to aquatic organisms including
algae, aquatic plants, invertebrates, vertebrates, and other biota with an aquatic lifestage
(e.g., amphibians). Pb from atmospheric sources can be directly deposited over a water
surface or enter aquatic systems through runoff from terrestrial systems. Pb loadings to
aquatic ecosystems, especially freshwater systems, are primarily derived from the runoff
and erosional transport from terrestrial systems ("U.S. EPA. 2006c). Wet and dry
deposition of Pb to land and water (Section 3.3.1.2), Pb in runoff (Section 3.3.2.4) and Pb
in water and sediment (Section 3.3.2.3) are considered in Section 3.3 Fate and Transport
of Pb. The flux of Pb in aquatic ecosystems is therefore influenced by the dynamic
physical and chemical interactions within a watershed. Data on ambient Pb
concentrations in rain, snow, natural waters, and sediment are summarized in Section 3.6.
7.3.2 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 and discussed in
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detail in Section 3.3 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 Ca concentration, pH, alkalinity, 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 PbS04,
PbCl4, Pb2+, cationic forms of Pb hydroxide, and ordinary hydroxide [Pb(OH)2], while in
alkaline waters, common forms of Pb include Pb carbonates [Pb(C03)] and hydroxides
[Pb(OH)2]. In freshwater systems, Pb complexes with inorganic OH" and C032 and
forms weak complexes with CF; conversely, Pb speciation in seawater is a function of
chloride concentration and the primary species are PbCl3, PbC03, PbCl2, and PbCl+. 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.
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 Ca and Mg ions) of the water
in a specific aquatic system. Models such as the BLM (Paquin et al.. 2002) 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.
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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 iron, SHM predicted Pb speciation predominantly
characterized by both iron 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 concentrations in Furbach stream (both in the Coeur
D'Alene River Basin in Idaho), indicating that such models may not be able to accurately
describe metal speciation under all environmental conditions (Balistrieri and Blank.
2008).
Quantification of different sediment metal-binding phases, including sulfide, organic C,
Fe, and Mn phases, is important to fully understand the bioaccessible fraction of Pb and
the toxicity to benthic organisms (Simpson and Batlev. 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
02) (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
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macrophyte roots resulted in the oxidation of sediment-bound Pb, leading to the release
of bioaccessible Pb fractions (Sundbv et al. 2005).
7.3.2.1 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 new evidence in this ISA 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 concentrations (0.25 (.iM) 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, Ca 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 Ca (2.5 mM). Ca 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 bivalves Macomona liliana 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).
7.3.2.2 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
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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
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.
7.3.2.3 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 lb).
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7.3.2.4
Dissolved Organic Matter (DOM)
Uptake of Pb by water-column organisms is affected by the concentration of DOM
(Manor et al.. 201 la; Mager et al.. 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 fiilvic 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
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.. 2010b).
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.. 2010a). 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..
2010a). Formation of a ternary complex that is better absorbed by biological membranes
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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 et al. (2011) subsequently have shown that different components of DOM
have different effects on Pb bioavailability in marine systems. Their initial research using
commercially-derived 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. 2007V 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.. 201IV
As little as 1 (imol of humic acid introduced into surface waters was sufficient to reduce
Pb uptake by perennial ryegrass, Lolium perenne, grown in nutrient solution. This
resulted from a decrease in the concentration of the free Pb fraction by several orders of
magnitude following complexation with the OM. Pb content on the root surface was
reduced to 8 (imol/g from 20 (imol/g following humic acid addition, and relative Pb
absorption (absorption in the presence of humic acid divided by absorption in the absence
of humic acid) was determined to be approximately 0.2 (Kalis et al.. 2006). Conversely,
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 Slavevkova. 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 Ca-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, the attempted field verification of freshwater
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bioaccessibility models was conducted at sites with distinct point-sources of Pb
contamination, and only one model (SHM) adequately predicted Pb bioaccessibility.
7.3.2.5 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 (Anklev et al. 1996; Di
Toro et al. 1992) and in the development of sediment quality criteria (Section 7.3.3). The
role of sulfides in the flux of Pb from sediments is discussed further in Section 3.3.2.3.
7.3.3 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. 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 (.ig Pb/g dry weight, range 0.5 to 12,000 |_ig Pb/g dry weight) and
fish tissues (median 0.59 (.ig Pb/g dry weight, range 0.08 to 23 (.ig Pb/g dry weight [whole
body]) in the U.S. based on a synthesis of National Water Quality Assessment
(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 3.6.5 and Table 2-1 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 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
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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 (AWQC) 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. 2010b'). In freshwater, the CMC is 65 j.ig 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 7.2.1, Table 7-2 and
Section AX7.2.1.4). The first approach is based on empirical correlations between metal
concentrations in bulk sediment and associated biological effects to derive threshold
effect concentrations (TEC) and probable effects concentrations (PEC) (MacDonald et
al.. 2000). The TEC/PEC approach 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, they 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 7.2.3 and 7.3.4. 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
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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, new information since the 2006 Pb AQCD on Pb in aquatic
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. Freshwater ecosystem effects
are considered first, followed by corresponding sections on effects in saltwater
ecosystems. New research on the bioavailability and uptake of Pb into freshwater
organisms including plants, invertebrates and vertebrates is presented in Section 7.3.4
Effects of Pb on the physiology of freshwater fauna and biota (Section 7.3.5) are
followed with data on exposure and response of freshwater organisms (Section 7.3.6).
Responses at the community and ecosystem levels of biological organization are
reviewed in Section 7.3.7 followed by a brief consideration of critical loads in freshwater
systems (Section 7.3.8), and characterization of sensitivity and vulnerability of ecosystem
components (Section 7.3.9). Corresponding sections for saltwater ecosystems include
bioavailability (Section 7.3.11), biological effects of Pb in saltwater (Section 7.3.12) and
exposure and response of saltwater species (Section 7.3.13). Community and ecosystem
level effects in saltwater are considered in Section 7.3.14, followed by characterization of
sensitivity and vulnerability in saltwater species (Section 7.3.15). Finally, the effects of
Pb on ecosystem services associated with aquatic environments are discussed in
Section 7.3.16 and an overall summary of aquatic effects is presented in Section 7.3.17.
7.3.4 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" (Fairbrother et al.. 2007).
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
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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 7-1 and in Section 7.2.3. 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 3. 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/iron-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.
Routes of Exposure
According to the 2006 Pb AQCD, 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 7.3.2. 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
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(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 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. This is followed by a discussion of
3)	Bioavailability in algae, plants, invertebrates and vertebrates. 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.
The bioaccumulation and toxicity of Pb to aquatic organisms are closely linked to the
environmental fate of the metal under variable environmental conditions (Section 3.3) 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
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bioavailability and uptake of the metal by aquatic organisms. The BLM (Figure 7-2)
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.. 2009V 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-
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 (Di 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
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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
(AhIf et al. 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 (Pemet-C'oudrier 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 (Lam el as 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
can 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.
7.3.4.1 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). Aquatic macrophytes and algae can accumulate Pb from
either the water column or sediments, based on their specific microhabitats. For instance,
rooted macrophytes may be more likely to accumulate Pb from sediment sources, while
floating macrophytes or algae will take up Pb suspended or dissolved in the water
column. However, significant species-dependent differences in bioaccumulation rates, as
well as concentrations of sequestered metals within different parts of the plants (shoots
versus roots), have also been observed and some authors have concluded that the plant
species is a more important determinant of Pb uptake than is habitat type. 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, however, this chapter focuses on
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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. 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.
The microalgae Spirulina platensis was demonstrated to accumulate Pb from Zarrouk
culture medium, with 2.7, 6.9, 19, 45 and 145 (.ig Pb/mg accumulated at high Pb
concentrations of 5,000, 10,000, 30,000, 50,000 and 100,000 (ig Pb/L, 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 3 |_imol Pb/g dry weight over a 7-day
exposure to 100 |_imol Pb, despite saturation of intracellular Pb concentrations after 5
days of exposure (Ran et al.. 2007). 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. When exposed to water concentrations of
up to 100 (.iniol 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 100 |_imol 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 et al.. 2010). Pb bioaccumulation in curly
pondweed (Potamogeton crispus) was observed to be dose-dependent, with plants
accumulating 3.3, 5.5, 15.4, and 23.6 |_ig Pb/g, at aquatic concentrations of 10, 20, 30, 40,
and 50 mM Pb (Xu et al.. 2011).
Given that atmospherically-derived Pb is likely to become sequestered in sediments,
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
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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 jj.g/g dry weight),
while Potamogeton malainus tissue contained the least amount of Pb, 2.4 jj.g/g 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 and Alternantheraphiloxeroides 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 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 and Mars ilia
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).
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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 5 and 10 mg Pb/L, plant
biomass was found to contain an average of 2.9 and 6.6 mg Pb, respectively, versus 0.2
and 0.3 mg in sediment (Hurd and Sternberg. 2008).
Young Typha latifolia, another rooted macrophyte, were grown in 5 and 7.5 mg/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 environmental Pb
exposure (Alonso-C'astro et al.. 2009).
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). Environmental Pb concentrations in water and sediment averaged
0.4 |ag 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 beccabunga) 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
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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 measured 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.
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.
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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.
7.3.4.2 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 new uptake study in freshwater mussels available since the 2006 Pb AQCD, the
Eastern elliptio mussel (Elliptio complanata) was shown to accumulate Pb rapidly from
water and then reach an equilibrium with exposure level and tissue concentration by two
weeks following exposures to 1, 4, 14, 57 or 245 |_ig Pb/L as Pb-nitrate (Mosher et al.. In
Press). 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 |_ig Pb/g. The authors concluded that the mussels
were likely eliminating Pb via pseudo feces and through storage of Pb in shell.
The 2006 Pb AQCD 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
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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
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
(0.05 (.iniol Pb 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
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 (Borgmann 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 to 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-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.
Measured tissue concentrations ranged from approximately 15 to 35 (.ig Pb/g.
Hydropsychid Pb BCFs ranged from 41 to 65, and averaged 54, indicating a relatively
high accumulation rate 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 Pb distribution in freshwater organisms generally support the findings
of the 2006 Pb AQCD that Pb is primarily sequestered in the gills, hepatopancreas, and
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muscle. Uptake of Pb by the crayfish Cherax destructor exposed to 5,000 (.ig Pb/L 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 exposure to 50, 100 and 500 |_ig Pb/L 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
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 etal. 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 ofPb fori?, virgo mayflies, and
although there was no statistical difference in Pb tissue concentrations between different
lifestages, Pb bioaccumulation did change as mayflies aged (Cid et 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)l. 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 (C'asas 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
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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).
7.3.4.3 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. 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 vertebrates, 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). New 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.
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, accumulation rates of Pb are
influenced by both environmental factors, such as water pH, DOC, and Ca
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 (Grose 11 et al.. 2006b). 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 |_ig Pb/g), followed by whole fish and
carcass (whole fish minus gill, liver, muscle and intestine) Pb concentrations
(approximately 70 to 80 (ig Pb/g). 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 200 nmol Pb/g tissue, although this number
was decreased from initial body burdens of greater than 500 nmol Pb/g at test initiation
(Manor et al.. 2010). Tissue distribution at 300 days was consistent with Grosell et al.
(2006b) 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 of fish to 157 nM Pb in base water for 150
days resulted in fathead minnow whole body concentrations of approximately 150 nmol
Pb/g tissue, 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: 177 nM Pb in hard water (Ca2+
500 (.iM) or 187 nM Pb in humic acid supplemented water (4 mg/L). While the addition
of humic acid significantly reduced Pb bioaccumulation in minnows (to approximately 50
nmol Pb/g on a whole body basis), Ca sulfate did not alter uptake. Despite the fact that
Ca-mediated Pb toxicity occurred in larval fathead minnow, there was no concurrent
effect on whole body Pb accumulation.
Uptake studies in other teleosts of Pb from freshwater have generally followed the pattern
of uptake described above for fathead minnow. In the cichlid, Nile tilapia (Oreochromis
niloticus) Pb accumulated significantly in gill (45.9 ±34.4 jj.g/g dry weight at 10 (.iM, 57.4
±26.1 jj.g/g dry weight at 20 (.iM) and liver (14.3 jj.g/g dry weight at 10 |_iM and 10.2 jj.g/g
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dry weight at 20 |_iM) during a 14-day exposure (Atli and Canli. 2008). In rainbow trout
exposed to 100 (.ig Pb/L for 72 hours, the accumulation in tissues was gill > kidney >
liver and this same pattern was observed in all concentrations tested (100-10,000 (.ig
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/L 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 trutta) exposed to aqueous Pb concentrations ranging from 15 to
46 |ag Pb/L for 24 days accumulated 6 |ag Pb/g dry weight in gill tissue and Pb
concentrations in liver tissue reached 14 jj.g Pb/g 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 from media containing 0.025 (imol Pb was
significantly increased by neutral pH (versus a pH of 6 or 8) and by Ca concentrations of
0.5 mmol; uptake rate of Pb was increased from 10 L/kg-h to 35 L/kg-h by increasing pH
from 6 to 7, and from 20 L/kg-h to 35 L/kg-h by increasing Ca concentration from 0.1
mmol to 0.5 mmol (komiarova and Blust. 2009c). This study also 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.
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 Pb body burdens ranging
from 12 to 68 mg Pb/kg dry weight (Kliozhina 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 jj.g 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 (ig Pb/g tissue and 1.3 and 1.0 |_ig Pb/g
tissue for C. idella and L. rohita, respectively), they accumulated significantly different
amounts of Pb in gill and muscle tissues. C. idella accumulated more than twice the Pb in
these tissues (1.6 and 1.3 |ag Pb/g) than did L. rohita (0.5 and 0.4 |_ig Pb/g).
The studies reviewed above generally support the conclusions of the 2006 Pb AQCD that
the gill is a major site of Pb uptake in fish and that there are species-dependent
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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 fingerlings 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) fingerlings, 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.. 2006b) and 300 days (Manor 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.. 2006b). 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 (Manor et al..
2008). African catfish (Clarias gariepinus) exposed to aqueous Pb concentrations of 50
to 1,000 |ag 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 (Mever et al.. 2005; Rozon-Ramilo et al.. In Press). Alves et al. (2006)
administered a diet of three concentrations of Pb (7, 77 and 520 |ag Pb/g dry weight) to
rainbow trout for 21 days. Doses were calculated to be 0.02 |_ig Pb/day (control),
3.7 |ag Pb/day (low concentration), 39.6 |_ig Pb/day (intermediate concentration) and
221.5 |ag 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)
sequestered Pb in a dose-dependent manner. The gills had the greatest concentration of
Pb on day 7(8.0 (ig Pb/g tissue wet weight) and this accumulation decreased to
2.2 |ag Pb/g 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
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study red blood cells were identified as a reservoir for dietary Pb. Plasma did not
accumulate significant Pb (0.012 (ig Pb g wet weight in the high dose), however, Pb was
elevated in blood cells (1.5 (.ig Pb g 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 (100, 400, and 800 |_ig Pb/g dry weight)
accumulated the greatest concentration of Pb in the intestine, followed by the stomach
and then the liver (Dai et al.. 200%). 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 50 or 500 (ig Pb/g, accumulated Pb preferentially in anterior
intestine (Alves and Wood. 2006V 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 et al.. 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 (ig Pb/g wet weight. Fish skin accumulated the greatest Pb concentrations (0.02 to
0.05 |_ig Pb/g wet weight), followed by kidney, gills, liver, and muscle. Pb accumulation
in muscles (5 ng Pb/g) 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).
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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 (Oio and
Wood. 2007V
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 of fish 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 new 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 |ag Pb/g dry weight in Rana clamitans, 0.65 |ag Pb/g dry weight in Rana
catesbeiana and 1.32 |ag Pb/g dry weight in Hyla cinerea (Albrecht et al.. 2007). Blood-
Pb levels in Ozark hellbenders (Cryptobranchus alleganiensis alleganiensis), a candidate
species for the Endangered Species Act, ranged from 0.04 to 0.06 jj.g/g 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 (Cryptobranchus alleganiensis bishopi), a
species of concern, collected from four rivers and ranged from 0.075 to 0.088 |_ig Pb/g
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
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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
concentrations ranging from 0.001 to 30 mg Pb/L (2.91 to 114.5 Pb |_ig/g 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 {liana 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
New 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 |_ig Pb/g 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 |_ig Pb/g wet weight) followed by kidney (0.343 |_ig
Pb/g wet weight) blood (0.108 |_ig Pb/g wet weight), muscle (0.103 |_ig Pb/g wet weight)
and liver (0.063 |_ig Pb/g wet weight). No interspecies differences were observed in blood
Pb (range 0.04 to 0.1 |_ig Pb/g) from field-collected banded water snakes (Nerodia
fasciata), brown water snakes (N. taxsipilota) 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 Savannah River site in South
(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.
Mammals
Pb bone levels in Eurasian otters (Lutra lutra) measured in dead individuals collected in
southwest England fell by 73% between 1992 and 2004 (Chadwick et al.. 2011). Annual
mean bone Pb levels were 446 (ig Pb/kg in 1992 and 65 (ig Pb/kg in 2004. The 73%
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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 Ca in streams.
7.3.4.4 Food Web
In the 2006 Pb AQCD, trophic transfer of Pb through aquatic food chains was considered
to be negligible ("U.S. EPA. 2006c). 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). New 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 measured in sediment, water, 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 (Croisetiere 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
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.
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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 Neris 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 ofN.
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 (Bovle 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, U.S. Generally, Pb
concentrations decreased with increasing trophic level: detritus contained 20 to 60 (.ig
Pb/g dry weight, while periphyton and algae contained 1 to 30 (ig Pb/g dry weight;
invertebrates and fish collected from the same areas exhibited Pb tissue concentrations of
0.1 to 8 (ig Pb/g 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
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..
2010V
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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 polymorpha) have created a new benthic
pathway for transfer of Pb and other metals (Southward Hogan et al.. 2007V 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.
7.3.5 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 toxicity varying depending upon
the species and lifestage tested, duration of exposure, the form of Pb tested, and water
quality characteristics. The 2006 Pb AQCD noted that the physiological effects of Pb in
aquatic organisms can occur at the biochemical, cellular, and tissue levels of 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.
New 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. In this ISA, effects on reproduction, growth and survival are reported in
additional species with some effects occurring in sensitive biota at or near ambient levels
of Pb. 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. 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.
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7.3.5.1
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.
The effect of Pb exposure on the structure and function of plant photosystem II was
studied in giant duckweed, S. polyrrhiza (Ling and Hong. 2009). 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.
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 (Arunakumara and Zhang. 2008). Arunakumara et al. (2008)
determined the effect of aqueous Pb on the algal species S. platensis using solutions of
Pb-nitrate. Exposures at 5,000 (.ig Pb/L stimulated 10-day algal growth, growth was
inhibited at higher concentrations of 10,000 30,000 50,000 and 100,000 (.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 environmentally relevant concentrations of Pb (0.1 to 250 nmol/L) with the
greatest response at 3.15 times more stained cells compared to the control sample (Szivak
et al.. 2009).
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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 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,
algae, and moss species exposed to Pb. An aquatic moss, F. antipyretica, exhibited
increased SOD and ascorbate levels following a 2-day exposure to Pb-chloride solutions
of concentrations of 1, 10, 100, and 1,000 (.iniol. 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 exposures of 1,000 10,000 and 100,000 (.ig Pb/L (Doganet
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, 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 exposures to aqueous Pb (1 to
100 (imol), 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,
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activities of these antioxidant enzymes decreased (Mishra et al.. 2006b). Similarly, curly
pondweed (P. crispus) exposed to lead (10 to 50 mM) exhibited increased generation of
reactive oxygen species (Xu etal. 2011).
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 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). exposure to 0.3 mmol of
Pb for 21 days was sufficient to induce a gradual reduction of both chlorophyll and
protein content in the macrophyte Eichhornia crassipes. Decreased proteins were
theorized to be related to inefficient protein formation following disruption of nitrogen
metabolism after Pb exposure (Odiegba and Fasidi. 2006). Foliar proline (which is
thought to act as an antioxidant) concentrations were found to increase in a
concentration-dependent manner as Pb concentrations increase from 0.1 to 5.0 mmol.
Following 72-hour aqueous exposure to 41 (.iniol Pb-nitrate, phytochelatin and
glutathione concentrations in the freshwater algae Scenedesmus vacuolatus were
significantly increased over that of non-exposed algal cultures (Le Faiicheur 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. Root elongation was
significantly reduced in a number of wetland plant species (Beckmannia syzigachne,
Juncus effusus, Oenanthe javanica, Cyperus flabelliformis, Cyperus malaccensis, and
Neyraudia reynaudiana) following Pb exposures of 20,000 |ag Pb/L (Deng et al.. 2009).
Further, while both Zn and Fe exposures exerted some selective pressure on plants, the
authors did not observe the same with Pb, leading them to theorize that concentrations of
bioavailable Pb were not present in high enough quantities to have such an effect.
However, while Lemna sp. aquatic plants were determined to effectively sequester
aqueous Pb, the plant growth rate was not significantly different from zero following
exposures of 5,000 and 10,000 |_ig Pb/L, while exposure to 15,000 |_ig Pb/L was
associated with notable plant mortality (Hurd and Sternberg. 2008). In fact, Paczkowska
et al. (2007) observed that low Pb exposures (0.1 to 1.0 mmol 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,
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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 Wolffla arrhiza plants. Fresh weight of plants was
reduced following both 7- and 14-day exposures to Pb concentrations greater than 10
mmol, while chlorophyll a content was decreased at concentrations greater than 1 mmol
Pb (Piotrowska et al. 2010).
All of the observed effects on aquatic macrophytes and algae occur at concentrations not
typically encountered in surface waters.
7.3.5.2 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 (.ig Pb/L (U.S. EPA. 1977). In the 1986 Pb AQCD, 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. Endpoints of
effects of Pb on aquatic invertebrates reviewed in the 2006 Pb AQCD included
metabolism, 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. New
evidence that supports previous findings of Pb effects on reproduction and growth in
invertebrates is reviewed here as well as limited studies on changes in gene expression
and behavioral effects associated with Pb exposure.
Recent literature strengthens the evidence indicating that Pb affects enzymes and
antioxidant activity in aquatic invertebrates. These alterations at the suborganismal 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 mussel 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 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 elliptio mussels was
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significantly reduced following a 28-day exposure to 57 |_ig Pb/L and 245 (.ig Pb/L
(Mosher et al. In Press). A significant reduction in Na2' and significant increase in Ca 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. Induction of
heat shock proteins in zebra mussel exposed to 500 |_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 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 the osmoregulatory response has been studied since the 2006 Pb
AQCD. The combined effects 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. 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.
Recently, genomic-level responses to Pb-exposure have been observed in Daphnia. The
effects of Pb on D. magna hemoglobin gene expression was measured in daphnids
exposed for 24 hours to 25, 250, or 2,500 |_ig Pb/L as Pb-nitrate (Ha and Choi. 2009). A
significant induction in all four genes tested was observed at the highest concentration
with a two-fold change in two of the genes. Hemoglobin expression was also
significantly elevated above the controls in one gene at 25 |_ig Pb/L and in two genes at
250 |ag Pb/L although this expression was less than two-fold when compared to the
controls. Changes in gene expression can lead to changes at the whole organism level
although these links are frequently not established.
Behavioral responses of aquatic invertebrates to Pb reviewed in the 2006 Pb AQCD
included avoidance. A limited number of new studies have considered additional
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behavioral endpoints. Feeding rate of the blackwormZ. 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 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).
New 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. Sublethal concentrations of Pb negatively affected the
total number of eggs, hatching success and embryonic survival of the freshwater snail B.
glabrata exposed to 50, 100, or 500 |_ig Pb/L (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 Daphniapulex at >500 (.ig Pb/L in 21 day exposures. In a 21-
day reproductive test in D. magna the number of neonates born per female was
significantly reduced at 25, 250, and 2,500 j.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 (Manor et al..
2011a).
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, 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
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(kitvatanachai et al.. 2005). Larval emergence rates of 78% (F0), 86% (Fl) and 86% (F2)
were observed in the control group while emergence rates decreased in each generation
46% (F0), 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.
As noted in the 2006 Pb AQCD, 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 new literature, 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 EC2o <4 (.ig Pb/L. (Grosell and Brix. 2009; Grosell et al.. 2006a). In L.
stagnalis exposed to 18.9 (ig/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 (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), Pb concentration (50,
100 and 200 |_ig Pb/L) and presence of a competitor, the population growth rate of two
freshwater rotifer species, Brachionus havanaensis and B. rubens, 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. In freshwater habitats, odonates are highly tolerant of
Pb with no significant differences in survival time of dragonfly larvae Pachydiplax
longipennis and Erythemis simplicicollis exposed to concentrations as high as 185 mg
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Pb/L (185,000 (.ig Pb/L) (Tollett et al.. 2009). Other species are more sensitive to Pb in
the environment and these responses are reviewed in Section 7.3.6.
7.3.5.3 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. Studies on aquatic vertebrates reviewed in the 1986 Pb AQCD
were limited to hematological, neurological and developmental responses in fish. In the
2006 Pb AQCD, effects on freshwater vertebrates included new data for fish specifically
considering the effects of water quality parameters on toxicity, as well as limited
information on sensitivity of aquatic stages of frogs and turtles to Pb. Biological effects
of Pb on freshwater fish that have been studied since the 2006 Pb AQCD report are
reviewed here, and limited new evidence of Pb effects on amphibians are considered.
This section also presents new 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 of fish
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. Finally,
limited new information since the 2006 Pb AQCD on physiological effects of Pb on
amphibians and marine mammals is presented.
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 mucous secretion, anemia, 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 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
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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.
Evidence of Pb effects on fish available since the 2006 Pb AQCD generally support the
findings in previous Pb reviews and elucidate 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 organism-level endpoints
including behavior and reproduction. 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. 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.
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 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. In fathead minnow, three genes, glucoses-
phosphate dehydrogenase, glutathione-S-transferase and ferritin were upregulated, in
microarray analysis, during 30 day exposures to Pb in base water (33jj.g Pb/L), or (37jj.g
Pb/L [hard]-water supplemented with 500 (.iM Ca2+) or (39 (.ig Pb/L [DOC]-water
supplemented with 4 mg/L humic acid). However, no changes in whole body ion
concentrations were observed (Mager et al.. 2008).
In the freshwater fish Nile tilapia, liver catalase, liver alkaline phosphatase, NA, K-
ATPase and muscle Ca-ATPase activities were quantified in various tissues following a
14 day exposure to 5, 10, and 20 (.iM concentrations of Pb nitrate (Atli and Canli. 2007).
Liver catalase activity significantly increased in the 5 and 20 (.iM concentrations while
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liver alkaline phosphatase activity was significantly increased only at the 20 (.iM
concentration. No significant change in alkaline phosphatase activity was observed in
intestine or serum. Ca-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 (Firat et al.. In
Press). 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 5, 10, and 20 (.iM concentrations of 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 50 to
500 |ag Pb/g 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 |_ig Pb/g indicated that alanine transaminase, aspartate transaminase,
and lactate dehydrogenase activities significantly decreased in kidney in a concentration-
dependent manner (Dai et al. 2009a') 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.. 200%). 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.
Pb was shown to inhibit hepatic cytochrome P450 in carp (C. carpio), silver carp
(Hypothalmichtys molitrix) and wels catfish (Silurus glanis) in a concentration-dependent
manner from 0-4.0 (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 (ig Pb/L) while carp EROD activity
was inhibited at 0.76 |ag Pb/L. Interaction of Pb with cytochrome P450 was verified by
spectral changes using Fourier Transform Infrared (FTIR) spectroscopy. Liver damage to
African catfish exposed to Pb (50-1,000 |_ig Pb/L) for 4 or 8 weeks included hepatic
vacuolar degeneration followed by necrosis of hepatocytes (Adevemo. 2008b). The
severity of observed histopathological effects in the liver was proportional to the duration
of exposure and concentration of Pb.
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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). Costa et al. (2007) observed inhibition
of ALAD in hepatocytes of the neotropical traira (Hoplias malabaricus) following
dietary dosing of 21 |_ig Pb/g 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), CaS04 (37 |_ig
Pb/L) or (39 |_ig Pb/L) humic acid-supplemented water 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 (Manor et al. 2008).
In fish, changes in blood chemistry associated with Pb exposure were noted in the 2006
Pb AQCD and limited new literature since the last Pb review support these findings. In
the African catfish, packed cell volume decreased with increasing 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. In traira exposed to dietary doses
(21 |ag Pb/g via prey [Astyanax sp.]) for five days, 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.
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 Ca 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 CI" influx (Rogers et al..
2005). Alves et al. (2006) administered a diet of three concentrations of Pb (7, 77 and
520 |ag Pb/g dry weight) to rainbow trout for 21 days, and measured physiological
parameters including Na+ and Ca+ influx rate from water. Dietary Pb had no effect on
brachial Na+ and Ca+ rates except on day 8 where Na+ influx rates were significantly
elevated. These studies suggest that Pb is intermediate between purely Ca antagonists
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such as Zn and Cd and disruptors of Na and CI balance such as Ag and Cu. This finding
has implications for BLM modeling since it suggests that both Ca and Na need to be
considered as protective cations for Pb toxicity. Indeed, protection from Pb toxicity by
both Na and Ca have been documented in freshwater fish (komiarova and Blust. 200%).
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 Pb nitrate revealed focal areas of epithelial hyperplasia and necrosis at the
lower exposure concentrations (50 (.ig Pb/L and 100 (.ig Pb/L) (Adevemo. 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 2.5, 12.5
and 25 mg Pb/L 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).
In addition to recent evidence of Pb interruption of Na+ and CI" at the gill (Rogers et al..
2005).	Pb can interfere with the ionoregulation of Na+ and CI" and tubular reabsorption of
Ca+, Mg2+, glucose, and water in the teleost kidney (Pate 1 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 1,200 |_ig Pb/L as Pb nitrate.
Rates of Na+ and CI" excretion decreased by 30% by 48 hours while Mg excretion
increased two-to-three fold by 96 hours. Urine flow 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 Ca+ 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 CI" were decreased and K+ levels were elevated following a 21 day
exposure to 50 (ig Pb/L (Firat et al.. In Press).
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 p38VIAPK were identified for the first time as
possible molecular targets for Pb neurotoxicity in a teleost (Leal et al.. 2006). The
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phosphorylation of ERK1/2 and p38v,APK by Pb was determined in vitro and in vivo in the
catfish (Rhamdia queleri). R. quelen exposed to 1,000 (.ig Pb/L acetate for two days
showed a significant increase in phosphorylation of ERK1/2 and p38VIAPK m the nervous
system. Incubation of cerebellar slices for 3 hours in 5 and 10(iM 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 (.ig Pb/g wet weight (Rabitto et al.. 2005).
Histopathological observations of brains of African catfish exposed to 500 |_ig Pb/L or
1,000 |ag Pb/L Pb for 4 weeks included perivascular edema, focal areas of malacia, and
diffuse areas of neuronal degeneration (Adevemo. 2008b').
Evidence from the 2006 Pb AQCD 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
exposure to either 800 or 1,600 |_ig Pb/L for 112 days (Hou et al. 2011). These
chondrostei 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. 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 |ag Pb/L for 300 days, then subsequently
tested in a breeding assay for 21 days. The time interval between 1st and 5th ingestion of
10 prey items (Artemia nauplii) 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
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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 (Manor and
Grosell. 2011).
Zebrafish embryos exposed to low concentrations of Pb (lOnM or 30 nM) 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 etal.. 2011). 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 adult fish hatched from Pb-exposed embryos (30 nM),
visual response to a rotating black bar against a white background (ability to detect
contrast) was significantly degraded. In another study with zebrafish embryos exposed to
higher concentration of Pb (200,000 nM) from 0 to 6 days post hatch, swim movements
and escape action were also significantly slower than the control group (Dou and Zhang.
2011).
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
(scoliosis) (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).
Limited new studies on reproductive effects of Pb in fish from oocyte formation to
spawning are available. The effects of metals on embryonic stage of fish development in
C. carpio and other species were reviewed in Jezierska et al. (2009) and included
developmental 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. A significant concentration-dependent increase in
morphological malformations was observed in African catfish embryos exposed to
100 |_ig Pb/L, 300 |ag 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
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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 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 g Pb/kg-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 (120 (ig Pb/L) reduced total reproductive output, while a significant
increase in average egg mass was observed in the high Pb HC03" and DOC treatments as
compared to egg mass size in controls and in low HC03" and DOC treatments with Pb.
No significant differences were present between treatments in egg hatchability.
Reproductive and endocrine effects of Pb have also been observed at the cellular level in
fish, including alterations in gonadal tissue and hormone secretions that are associated
with Pb-exposure. 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) (Adevemo. 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 Pb-nitrate (0, 001, 0.1, 1, 3, and 10 |_ig Pb/mL).
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 disruption of
steroid production and altered hormone secretion patterns observed at the low
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 of fish. 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 significantly increased throughout the study period while plasma
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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.. 2009V Cortisol levels were
significantly decreased in Nile tilapia exposed to 50 (ig Pb/L for 4 days but were
followed by a return to control levels at 21 days of exposure (Firat et al.. In Press).
Vitellogenin was significantly induced in juvenile goldfish (Carassius auratus) following
96-hour exposure to 1 x 10"9 and 1 xlO"10 M Pb 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.
Reduction of growth in fish was noted as an effect of Pb exposure in the 2006 Pb AQCD.
New studies available since the 2006 Pb AQCD do not present consistent evidence of
growth reduction in fish associated with Pb. In a series of exposures in which Ca+2, DOC
and pH were varied to assess effects on Pb toxicity to fathead minnows, Grosell et al.
(2006b) 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.. 2006b'). Fathead minnows
exposed to 33 |ag 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 |_ig Pb/L),
intermediate (82-156 |_ig Pb/L) and high (297-453 |_ig Pb/L) concentrations of Pb, Mager
et al. (2011b) did not observe growth impairment in fathead minnows at environmentally
relevant concentrations of Pb.
No effects on growth rates were observed in rainbow trout administered a diet containing
three concentrations of Pb (7, 77 and 520 |_ig Pb/g dry weight) for 21 days (Alves et al..
2006) or in Nile tilapia fed diets with 100, 400, or 800 jj.g/g Pb dry weight for 60 days
(Dai et al. 2009b). 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 contaminated with Pb (850-1,000 |_ig Pb/L-g dry mass for 30 days.
(Erickson et al. 2010). 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 |ag 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. 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 et al.. 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.
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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 7.2.3.3 and 7.2.4.3. 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 at environmentally relevant
concentrations of Pb 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.
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. 2002c). 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
clinodactyly, brachymely and spinal curvature and these effects persisted after
metamorphosis. At the highest concentration with surviving tadpoles (2,360 mg Pb/kg)
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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
2,360 mg Pb/kg (8,121 |_ig Pb/L sediment pore water) treatment. Conversely, no effects
were observed on organogenesis in X. laevis embryos exposed to a range of Pb
concentrations from 8,600 to 220,500 (.ig Pb/L using the Frog Embryo Teratogenesis
Assay (Gungordu et al.. 2010).
Endpoints of oxidative damage were measured in testes of the black-spotted frog {liana
nigromaculata) treated with 100 |ag Pb/L, 200 |ag Pb/L, 400 |ag Pb/L, 800 |_ig Pb/L or
1,600 |ag 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 |_ig 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 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 (Mouchet et al.. 2007). The lowest Pb concentrations w ith X. laevis (10
and 100 |ag 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.. 2007b). The authors noted that the
erythrocytic micronuclei and erythrocytic nuclear abnormalities frequencies generally
decreased with increasing exposure time and that this may be indicative of regulation of
genotoxic factors by tadpoles.
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 atra) from a lagoon in Spain impacted by Pb shot (Martinez-Haro et al.. 201IV
ALAD ratio in mallards decreased linearly with blood Pb levels between 6 and 40 |_ig
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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-Harp 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 100 |ag 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.
7.3.6 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
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). Overall, new data for
freshwater invertebrates generally supports the previous finding of sensitivity of juvenile
lifestages and indicates some effects of Pb observed in some species at environmentally
relevant concentrations.
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
4,000 and 24,500 |_ig Pb/L for growth as measured by cell division rate were derived
(Atici et al.. 2008). 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 10,000 |_ig Pb/L (Dirilgen. 2011).
Experiments with the blue-green algae Spirulina platensis produced a LC50 value of
75,300 (.ig Pb/L (95% CI: 58.5, 97.0) (Arunakumara et al.. 2008).
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In the 2006 Pb AQCD, effects of Pb-exposure in amphipods (H. azteca) and water fleas
(D. magna) 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; Grose 11 et
al.. 2006a). Growth of juvenile L. stagnalis was inhibited below the lowest concentration
tested resulting in an EC2o of < 4 (.ig Pb/L. In the same study, the NOEC was 12 (.ig 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 j.ig Pb/L (Sawasdee and Kohler. 2010).
Additional studies on Pb effects in aquatic invertebrates published since the 2006 Pb
AQCD have indicated differences in sensitivity of different lifestages of aquatic
organisms to Pb. In the freshwater mussel, Lampsilis 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.. 2010e). The 96-hour EC50 values for the
5-day-old L. siliquoidea in two separate toxicity tests were 142 and 298 |_ig Pb/L (mean
EC50220 (.ig Pb/L) in contrast to older juveniles (2 months old) with an EC50 >426 j^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 et al.. 2010e). A 96-hour test on newly
transformed juveniles was also conducted on Lampsilis rafmesqueana (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
CaC03 (Bechard et ai.. 2008). The 24-hour LC50 of 610 (ig 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.. 2006a). 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 |ag Pb/L (Mebane et al.. 2008). In fourth instars of the freshwater
midge larvae Chironomus javanus the 24, 48, 72 and 96 hour LC50's were 20,490, 6,530,
1,690 and 720 j.ig Pb/L, respectively (Shuhaimi-Othman et al.. In Press). This was
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comparable to the 96-hour LC50 (420 (ig Pb/L) in Chironomus plumosus (Vedamanikam
and Shazilli. 2008a,).
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 180 (ig Pb/L and were correlated with DOC (Esbaimh et al. 201IV
Median lethal concentrations for Moina micrura (LC50 690 (ig Pb/L), Diaphanosoma
birgei (LC50 3,160 (ig Pb/L), and Alona rectangular (LC50 7,000 (ig Pb/L) indicate
differences in sensitivity to Pb in these freshwater cladocerans from Mexico (Garcia-
Garcia et al.. 2006). An acute study of Pb with D. pulex identified a 48-hour LC50 of
4,000 (ig/L for this species (Theegala et al.. 2007). 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 (ig Pb/L), a LOEC (284 (ig Pb/L), and an EC2o of 125 (ig Pb/L was established for
this species (Grosell et al.. 2006a). 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 contrast, for rotifer Brachionus patulus neonates, the 24-hour LC50
was 6,150 (ig Pb/L (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 (ig Pb/L (kitvatanachai et al. 2005). A 48-hour LC50 of 5,200 (ig Pb/L was observed
in water-only exposures of the blackworm Lumbriculus variegatus (Penttinen 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).
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In the freshwater ostracod Stenocypris major, the 96-hour LC50 was 526.2 |_ig Pb/L
(Shuhaimi-Othman et al.. In Press). In another freshwater crustacean, the prawn
Macrobrachium lancesteri, the 96-hour LC50 was 35 (.ig Pb/L in soft water (<75 mg/L as
CaC03) (Shuhaimi-Othman et al.. 201IV
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. (2006a')
conducted a series of 30-day exposures with larval fathead minnow in which varying
concentrations ofCa2+ (as CaS04) 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, CaS04 and pH influenced Pb toxicity considerably over the range of
water parameters tested. The 30-day LC50 for low hardness (19 mg CaSO/L) in basic test
water was 39 |_ig dissolved Pb/L and the highest LC50 value (obtained from the protection
from increased concentrations of DOC and CaS04) was 1,903 |_ig dissolved Pb/L (Grosell
et al.. 2006b).
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. (2011b) 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 NaHC03 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 H with Pb2+ (Mager et al.. 201 lb). 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.. 201IV
In the 2006 Pb AQCD, fish lifestage was recognized as an important variable in
determining the sensitivity of these organisms to Pb. New data available since the 2006
Pb AQCD 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 ACR's for this species (Mebane et al..
2008). Two early-lifestage chronic tests were conducted, the first with an exposure range
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of 12-384 |_ig Pb/L (69 days) at 20 mg CaC03/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 CaC03/L. In
the 69-day test, the following chronic values were observed for survival: NOEC=24 (.ig
Pb/L, maximum acceptable toxicant concentration=36 (.ig Pb/L, ECi0=26 (.ig Pb/L,
EC2o=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, ECi:)=7(.ig Pb/L, EC2o=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 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 new 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). A
96-hour LC50 of 96,100 |_ig Pb/L was determined for X. laevis embryos exposed to a
range of Pb concentrations from 8,600 to 220,500 |ag Pb/L using the Frog Embryo
Teratogenesis Assay (Gungordu et al.. 2010).
7.3.7 Freshwater Community and Ecosystem Effects
As discussed in the 1986 Pb AQCD and the 2006 Pb AQCD, 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
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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 this ISA, 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,
the 1986 Pb AQCD, the 2006 Pb AQCD and this ISA on 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. 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.
Microcosm evaluations of the effects of Pb on aquatic cyanobacteria communities
indicated that exposure to 25 mM Pb reduced bacterial biomass and diversity. After one
week of Pb exposure, total bacteria biomass was reduced from an average of 9.3 mg
carbon/cm3 sediment to 1.3 mg carbon/cm3 (Burnat et al.. 2009). Pb exposure impacted
individual cyanobacteria species differently, with Microcoleus sp. experiencing a greater
decrease in abundance than Halomicronema-like cyanobacteria.
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 /<'. 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 m and 1,500 m downstream from the mining point-source (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
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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 et al.. 2008).
Exposure to three levels of sediment Pb contamination (322, 1,225, and 1,465 |_ig Pb/g
dry weight) had variable effects on different species within an aquatic 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%) and Marylynnia 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 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 (Aimclo 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.
Rhea et al. (2006) examined the effects of multiple heavy metals in the Boulder River,
MT, U.S., 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
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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
ecosystems 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-
Frnnchini et al.. 2008). They showed that M. micrura had twice the Pb body burden of I),
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 (Komjarova
and Blust. 2009cJ.
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. Although this may be a result of differing bioaccumulation tendencies, the
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authors suggested that other factors, including the potential for sediment exposures, may
be responsible for Pb uptake in lentic invertebrates (Larsson et al.. 2007).
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.. 200%).
Density of Ozark sculpin (Cottus hypselurus) and banded sculpin (Cottus 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
infestations in fish, increasing both microbial density and resilience, and potentially
increasing the likelihood of serious disease outbreak.
7.3.8 Critical Loads in Freshwater Aquatic Systems
The general concept and definition of critical loads is introduced in Section 7.1.3 of this
chapter [also see Section 7.3 of the 2006 Pb AQCD (U.S. EPA. 2006c)l. Critical load
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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 lead 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. (In Press) apply the UWM 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%) vs. the oligotrophic lakes (2%),
owing 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-yr 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 an order of magnitude 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 PbC03 in sediment. Refer to Section 7.3.6 of the 2006 Pb AQCD for
additional discussion of critical loads of Pb in aquatic systems.
7.3.9 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 7.3.3 and 7.3.4) may also alter Pb toxicity.
The 2006 Pb AQCD 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
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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.
7.3.9.1 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 (Jothinavagi 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 high molecular weight (>10 kDaltons) cationic Pb species 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.
7.3.9.2	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 et al.. 2010). Therefore, it is likely that such plant species would be
significantly less susceptible to Pb exposure than those species without those abilities.
7.3.9.3	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
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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.. 2006V
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 concentrations of Pb (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 (Uvsal and Taner. 2009). Decreased pH
was also demonstrated to increase the uptake of environmental Pb in aquatic plants
(Wang et al.. 2010a; Uvsal and Taner. 2009). 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.
7.3.9.4 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. (2010e)
demonstrated that the newly transformed juvenile mussels, L. siliquoidea and L.
rajinesqueana, 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 et al. 2010e).
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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 (Osmanet
al.. 2007a).
7.3.9.5 Species Sensitivity
Species-specific Ca requirements have been shown to affect the vulnerability of aquatic
organisms to Pb. The snail, L. stagnalis, exhibits an unusually high Ca demand due to
CaC03 formation required for shell production and growth, and exposure to Pb prevents
the uptake of needed Ca, leading to toxicity. Consequently, aquatic species that require
high assimilation rates of environmental Ca 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
Ca uptake (Grosell et al.. 2006a).
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 lb;
Mohapatra et al.. 2009; Tollett et al.. 2009; Bergev and 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 Gooal. 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
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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 jj.g 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.. 2006b').
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).
7.3.9.6 Ecosystem Vulnerability
Relative vulnerability of different aquatic 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 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
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invertebrate species endemic to such systems may be more at risk from Pb contamination
than corresponding species in other habitats.
7.3.10 Introduction to Bioavailability and Biological Effects of Pb in Saltwater
Ecosystems
Saltwater ecosystems include salt marsh, estuaries, beaches and other coastal areas that
may receive Pb contributions from direct atmospheric deposition and/or via runoff from
terrestrial systems. A range of 0.005-0.4 (.ig Pb/L for saltwater was reported in by Leland
and Kuwabara (1985) and 0.01 to 27 |_ig Pb/L by Sadiq (1992) with the higher values
associated with sites involving human activity. Levels of Pb in the North Atlantic and
North Pacific surface waters ranged from 0.005 to 0.05 (.ig Pb/L but the range of values in
coastal waters and estuaries were approximately equal to the range of Pb in freshwater.
Additional information on Pb levels in water is available in Section 3.6. 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 (AWQC) 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. 2010b). 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 this ISA in Section 7.3.2.
In the following sections, new 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, the 1986 Pb AQCD and the 2006 Pb
AQCD are included where appropriate. The sections are organized to consider uptake of
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Pb and effects at the species level, followed by community and ecosystem level effects.
New research on the bioavailability and uptake of Pb into saltwater organisms including
plants, invertebrates and vertebrates is presented in Section 7.3.11. Effects of Pb on the
physiology of marine fauna and biota (Section 7.3.12) are followed with data on exposure
and response of saltwater organisms (Section 7.3.13). Responses at the ecosystem level
of biological organization are reviewed in Section 7.3.14 followed by characterization of
sensitivity and vulnerability of ecosystem components (Section 7.3.15).
7.3.11 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" (Fairbrother et al.. 2007).
Factors affecting bioavailability of Pb to marine organisms are the same as those in
freshwater systems. 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 (Nivogi 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 than in
freshwater systems, varying mostly with amount and type of dissolved organic matter.
BLMs (Figure 7-2) now being developed for marine organisms are functionally similar to
those applied to freshwater organisms (Section 7.3.4).
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 7.3.2.4 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.. 2010a). This also was observed in
the marine diatom Thalassiosira weissflogii, where humic acids absorbed to cell surfaces
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increased metal uptake (Sanchez-Marm et al.. 2010a'). Formation of a ternary complex
that is better absorbed by biological membranes was another proposed mechanism that
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
commercially-derived 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 lead
concentration measurements, potentially as a result of adsorption of DOM-Pb complexes.
Based on the above, BLMs used to predict bioavailability of Pb to aquatic organisms
(Di 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 Ca channel permeability. Veltman et al.
(2010) proposed an 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 Ca, 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.
7.3.11.1 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
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evidence for Pb uptake. New data available since the 2006 Pb AQCD includes Pb
bioaccumulation studies conducted with five species of marine algae, (Tetraselmis chuii,
Rhodomonas salina, Chaetoceros sp., Isochrysis galbana and Nannochloropsis
gaditana). In this study it was demonstrated that bioaccumulation rates varied with
species. I. 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 1.4 and
6,348 (.ig Pb/L (Debelius et al. 2009).
New 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 trans locators of Pb (MacFarlane et al.. 2007).
7.3.11.2 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.
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 estuary in New
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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 (Einspom et al.. 2009; Einspom 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 (Einspom 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 (C'asas 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 in M. 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 (Mi ram and
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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7.3.11.3 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 |_ig Pb/g in a 42-day
feeding study (Soto-Jimenez et al.. 201 la). 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 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 (.ig Pb/g. 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.
7.3.11.4 Marine Food Web
As discussed in Section 7.3.4.4 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). New
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 aquatic marine 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 la; Soto-Jimenez et al.. 201 lb). 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 Artemia franciscana (crustacean, brine shrimp)
which were then fed to Litopenaeus vannamei (crustacean white 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 la).
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 (Ruclas-lnzunza 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 white
shrimp (Litopenaeus vannimei) and mullet (Mugil cephalus) to pelican (Pelecanus
occidentalis).
Other studies have traced Pb in aquatic 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; Re in folder et al.. 1998).
7.3.12 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. 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.
7.3.12.1 Saltwater Algae and Plants
New evidence in this ISA on toxicity of Pb to marine algae indicates that species exhibit
varying sensitivities to Pb in saltwater. Pb tested at concentrations up to 10 |_imol/L did
not affect photosynthetic activity in seven species of marine macroalgae (Ascophyllum
nodosum, Fucus vesiculosus, Ulva intestinalis, Cladophora rupestris, Chondrus crispus,
Palmaria palmate, Polysiphonia lanosa) as measured by pulse amplitude modulation
chlorophyll fluorescence yield although Pb was readily accumulated by these species
(Baumann et al.. 2009). The lowest 72-hour EC50 for growth inhibition reported for
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marine algae was 105 (.ig Pb/L in Chaetoceros sp (Debelius 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.
7.3.12.2 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. As in freshwater invertebrates, the antioxidant
system, survival, growth and reproduction are affected by Pb in marine organisms. In
aquatic invertebrates, Pb has also been shown to affect stress responses and
osmoregulation. New evidence on reproduction and growth in invertebrates is also
reviewed here.
Recent literature strengthens the evidence indicating that Pb affects enzymes and
antioxidant activity in marine invertebrates. Activity of enzymes associated with the
immune defense system in the mantle of pearl oyster were measured at 0, 24, 48 and 72
hour exposure to 104 |_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 digestive gland of the bivalve Chlamys farreri (Zhang et al.. 2010c). In
contrast, Einsporn et al. (2009) observed no change in catalase activity in the digestive
gland and gill of blue mussel M edulis following exposures 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 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
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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-ethoxyresorufin-o-deethylase (EROD) activity
when compared to controls (Zhang et al.. 2010c). 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. 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 etal.. 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 new 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 Pb
(40 (ig/L to 400 (ig/L) for 20 days. Damage to 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 1 18 |_ig Pb/g compared to
controls. In the higher concentrations, start of offspring production was delayed further; 4
days in 234 |_ig Pb/g and 8 days in 424 |_ig Pb/g (Ringenarv et al.. 2007). Fecundity and
time of first offspring production was also reduced with increasing Pb concentration in
sediment above 118 |_ig Pb/g. The authors indicate that this concentration is below the
current marine sediment regulatory guideline for Pb (218 |_ig Pb/g sediment) (NQAA.
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1999) and that reproductive endpoints are more sensitive than survival. 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 100 (.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 50 (.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 new studies reviewed in this ISA, Pb
exposure negatively affects the growth of aquatic 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 with increasing concentrations of Pb associated with sediment
(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), 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 sediment up to 0.41 (imol/g had no effect on
survival (Horng et al.. 2009). Other species are more sensitive to Pb in the environment
and these responses are reviewed in Section 7.3.13.
7.3.12.3 Saltwater Vertebrates
Saltwater Fish
There is a dearth of information in previous Pb AQCDs on Pb effects in saltwater fish.
New data available since the 2006 Pb AQCD includes 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, Ca and CI, in plasma. No effects were observed in Pb-
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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 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. 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.. 201 la). The fish were fed white shrimp exposed to Pb via brine shrimp that were in
turn fed microalgae cultured at an environmentally-relevant concentration of 20 |_ig Pb/L.
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.
Limited new studies on reproductive effects of Pb in marine fish are available. Decreased
oocyte diameter and density in the toadfish (Tetractenos glaber) were associated with
elevated levels of Pb in the gonad of field-collected fish 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 (Tursiops truncates) (Camara Pellisso et al.. 2008). No effects on viability of
immune cells, apoptosis, or phagocytosis were observed in 72 hour exposure to
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concentrations of 1, 10, 20 and 50 mg Pb/L. Proliferative response of bottlenose dolphin
leukocytes was significantly reduced at 50 mg Pb/L, albeit by only 10% in comparison to
the control.
7.3.13 Exposure and Response of Saltwater 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)
(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.
A series of 72-hour Pb toxicity tests were conducted with five marine microalgae species
(T. chuii, R. salina, Chaetoceros sp., I. galbana and TV. gaditana) 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
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.
Other studies of marine invertebrates published since the 2006 Pb AQCD 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 for adults was 3,000 (.ig Pb/L and 1,520 (.ig Pb/L for
juveniles. Ten-day exposures of juveniles in seawater resulted in an LC50 of 1,270 (.ig
Pb/L, an NOEC of 190 |_ig Pb/L and a LOEC of 390 |_ig Pb/L. In comparison, the LC50,
NOEC, and LOEC value for the adults exposed in sediment was 3,560 (.ig Pb/L. Juvenile
sediment tests results were LC50 1,980, NOEC 580 and LOEC 1,020 (.ig Pb/L. 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 (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 aLC50 of 830 |_ig
Pb/L in a 96-hour exposure (Sobrino-Figueroa et al.. 2007). Morphological deformities
were observed in 50% of veliger larvae of blacklip abalone (Haliotis rubra) at 4,100 |_ig
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Pb/L following a 48-hour exposure to Pb, suggesting this species is not as sensitive to Pb
as other marine invertebrate larvae (Gorski and Nugegoda. 2006). In the marine
polychaete H. elegans, EC50 values of gametes, embryos, larvae (blastula to 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. The EC50 for
developmental defects in this species was 1,150 (.ig Pb/L with aNOEL of 250 (.ig Pb/L.
7.3.14 Community and Ecosystem Effects in Saltwater Systems
As discussed in the 1986 Pb AQCD and the 2006 Pb AQCD, 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 in saltwater ecosystems 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).
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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
natural mineralogical sources dominated. The roots of S. 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.
In a laboratory microcosm experiment with spiked sediments collected from the
Swartkop River estuary, South Africa, the effects of Pb (3 to 6,710 (ig/Pb g sediment dry
weight) on meiobenthos was tested alone and in combination with Cu, Fe, and Zn
(Gvedu-Ababio and Baird. 2006). Total meiofauna density decreased after 32 days in Pb-
treated sediments (range 3 to 5 taxa) compared to control (9 taxa). 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, however, 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.
7.3.15 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
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AQCD 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.
7.3.15.1 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
re suspension of Pb triggered by high river runoff events was a key factor affecting
bioaccumulation of Pb inM edulis.
In another biota 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 siliqua). 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
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) measured 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
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upon tissue composition, corrections for mussel physiological condition are need to
compare results from different seasons and different locations.
7.3.15.2 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 100 (.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 50 j.ig Pb/L and greater (Gopalakrishnan et al.. 2008).
7.3.15.3 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 lb;
Mohapatra et al.. 2009; Tollett et al.. 2009; Bergev and Weis. 2007). Crab species U.
pugnax (Bergev and Weis. 2007) and Scylla serrata (Mohapatra et al.. 2009), and white
shrimp L. vannamei (Soto-Jimenez et al.. 201 lb) 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.
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Some tolerant species of fish (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.
7.3.16 Ecosystem Services Associated with Freshwater and Marine Systems
Pb deposited on the surface of, or taken up by organisms has the potential to alter the
services provided by aquatic biota to humans. At this time, a few publications address Pb
impacts on ecosystem services associated with aquatic systems, mostly estuaries, salt
marsh and freshwater wetlands rather than lakes and streams. The evidence reviewed in
the ISA illustrates that Pb can affect the ecological effects in each of the four main
categories of ecosystem services (Section 7.1.2) 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 of fish 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
that Pb poses an ecologically relevant risk to dunlin, Calidris alpina (Smith et al.. 2009b).
However, the authors noted that a major source of uncertainty in this study was the
NOAEL values for Pb.
Ecological services provided by freshwater wetlands are similar to those provided by
estuaries and 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 (Qdum. 2000; Gambrell. 1994). Recent reviews of pollution
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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. 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
(Ouan et al.. 2009). Other aquatic organisms have been considered for their role in
remediation of Pb in the environment. 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.
7.3.17 Summary of Aquatic Effects
This summary of the effects of Pb on aquatic ecosystems covers information from the
publication of the 2006 Pb AQCD to present. Refer to Section 7.4: Causality
determinations for Pb in Terrestrial and Aquatic Systems for a synthesis of all evidence
dating back to the 1977 Pb AQCD considered in determining causality.
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7.3.17.1 Biogeochemistry and Chemical Effects
Atmospherically-derived Pb can enter aquatic ecosystems by direct deposition or via
watershed processes. Once the Pb enters surface waters its fate and bioavailability are
influenced by Ca concentration, pH, alkalinity, and total suspended solids, and DOC,
including humic acids. Once in sediments, Pb bioavailability may be influenced by the
presence of sulfides and Fe and Mn oxides, physical disturbance, the presence of other
metals, biofilm and organisms. In many, but not all aquatic organisms, Pb dissolved in
the water can be the primary exposure route to gills or other biotic ligands. 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. As recognized in the 2006 Pb AQCD and further supported in this review,
chronic exposures to Pb may also include dietary uptake, and there is an increasing body
of evidence showing that differences in uptake and elimination of Pb vary with species.
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.
7.3.17.2 Bioavailability
There is evidence over several decades of research previously reviewed in Pb AQCDs
and in recent studies reviewed in this ISA that Pb bioaccumulates in plants, invertebrates
and vertebrates in aquatic systems, just as it does in terrestrial systems. According to the
2006 Pb AQCD, and further supported in this review, Pb adsorption, complexation,
chelation, etc., are processes that alter bioavailability to aquatic biota. Given the low
solubility of Pb 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.
As reviewed by Wang and Rainbow (2008) and supported by additional studies reviewed
in this ISA, 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 toxicity of Pb to aquatic organisms are closely linked to the
environmental fate of the metal under variable environmental conditions (Section 3.3) as
they are highly dependent upon the proportion of free metal ions in the water column.
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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. Trees that inhabit semi-aquatic environments have also been shown to
absorb Pb from contaminated sediments.
In the case of invertebrates, Pb can be bioaccumulated from multiple sources, including
the water column, sediment, and dietary exposures, and factors such as amount of
bioavailable Pb, lifestage, age, and metabolism can alter the accumulation rate.
Additional studies have considered the relative importance of water versus dietary uptake
of Pb in aquatic invertebrates. Use of stable isotopes has enabled simultaneous
measurement of uptake and elimination in several aquatic species 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.
Tissue accumulation of Pb via gill and dietary uptake has been further characterized since
the 2006 Pb AQCD in aquatic vertebrates and stable isotope techniques have been
applied to further elucidate bioaccumulation of Pb. The conclusions of the 2006 Pb
AQCD (that the gill is a major site of Pb uptake in fish and that there are species
differences in the Pb accumulation and distribution of Pb within the organism) are
supported in this review. In general, the accumulation of Pb in fish tissues is observed to
be gill>kidney>liver. The anterior intestine has been newly identified as a site of uptake
of Pb through dietary exposure studies (Alves et al.. 2006). Additional detoxification
strategies for Pb have been elucidated since the 2006 Pb AQCD. 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 (Goto and Wallace. 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), scallop and cuttlefish and was found to be
bound principally to organelles (Einsoom et al.. 2009; Einsporn and Koehler. 2008).
There are few new studies on Pb uptake by amphibians and mammals. In the 2006 Pb
AQCD, trophic transfer of Pb through aquatic food chains was considered to be
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negligible. Measured 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 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).
7.3.17.3 Biological Effects
Evidence in this review further supports the findings of the previous Pb AQCDs that
waterborne Pb is highly toxic to aquatic organisms, with toxicity varying with species
and lifestage, duration of exposure, form of Pb, and water quality characteristics.
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. Effects on plants supported by
additional evidence in this review and evidence from previous reviews include oxidative
damage, decreased photosynthesis and reduced growth. The mechanism of Pb toxicity in
plants is likely mediated by damage to photosystem II through alteration of chlorophyll
structure. Elevated levels of antioxidant enzymes are commonly observed in aquatic
plant, algae, and moss species exposed to Pb.
As observed in terrestrial invertebrates, upregulation of antioxidant enzymes is a common
biomarker of Pb exposure in aquatic invertebrates. Since the 2006 Pb AQCD, there is
additional evidence for Pb effects on antioxidant enzymes, lipid peroxidation, stress
response and osmoregulation. Studies of reproductive and developmental effects of Pb in
aquatic invertebrates in this review provide further support for findings in the 2006 Pb
AQCD. These new studies include reproductive endpoints for rotifers and freshwater
snails as well as multigenerational effects of Pb in mosquito larvae. Growth effects are
observed at lower concentrations in some aquatic invertebrates since the 2006 Pb AQCD,
including juveniles of the freshwater snail L. stagnalis where growth is affected at <4 (.ig
Pb/L (Grosell et al.. 2006a'). Behavioral effects of Pb in aquatic invertebrates reviewed in
this ISA include decreased valve closing speed in scallops and slower feeding rate in
blackworms.
Additional mechanisms of Pb toxicity have been elucidated in the gill and the renal
system of fish since the 2006 Pb AQCD. Further supporting evidence of reproductive,
behavioral, growth effects and effects on blood parameters have become available since
the 2006 Pb AQCD. The mitogen-activated protein kinases, ERK1/2 and p38MAPK were
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identified for the first time as possible molecular targets for Pb neurotoxicity in a teleost
(Leal et al.. 2006). Pb toxicity at the fish gill primarily involves disruption of Ca
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 CI" influx. Recent studies on fish behavior associated with Pb
exposure include decreased prey capture rate (Mager et al.. 2010) and swim speed
(Mager and Grosell. 2011) in fathead minnow and decline in startle response and visual
contrast in zebrafish (Rice etal. 2011).
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 of the 2006 Pb
AQCD ("U.S. EPA. 2006c). Since the 2006 Pb AQCD, studies conducted at
concentrations approaching environmental levels of Pb have indicated sublethal effects
on tadpole endpoints including growth, deformity, and swimming ability.
7.3.17.4 Exposure and Response
Concentration-response data on plants, invertebrates and vertebrates is consistent with
findings in previous reviews of species differences in sensitivity to Pb in aquatic systems.
Growth in plants continues to be a commonly measured endpoint affected by Pb
exposure. The lowest EC50 for growth observed in marine microalgae and freshwater
microalgae was in the range of 100 |_ig Pb/L.
In the 2006 Pb AQCD, concentrations at which effects were observed in aquatic
invertebrates ranged from 5 to 8,000 |_ig Pb/L. Several studies in this review have
provided evidence of effects at lower concentrations. 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. 2006a). A chronic value of 10 (ig Pb/L obtained in
28-day exposures of 2-month-old L. siliquoidea juveniles was the lowest genus mean
chronic value ever reported for Pb (Wang et al.. 2010e). 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 180 |_ig Pb/L and were most significantly
influenced by DOC and water ionic strength (Esbaugh et al.. 2011).
In the 2006 Pb AQCD, effects were reported 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,
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DOC, and pH has been conducted since the last review. However, effects in fish observed
in recent studies fall within the range of concentrations observed in the previous Pb
AQCD.
7.3.17.5 Community and Ecosystem Effects
Since the 2006 Pb AQCD, additional evidence for community and ecosystem level
effects of Pb have been observed primarily in microcosm studies or field studies with
other metals present. One ecological effect reported in previous Pb AQCDs is a shift in
community composition in Pb-impacted habitats toward more Pb-tolerant species. New
studies in this ISA provide evidence in additional habitats for community composition
shifts associated with Pb. Alteration of aquatic plant community composition was
demonstrated in the presence of elevated surface water Pb concentrations at three lake
sites impacted by mining effluents. Lakes with the highest levels of Pb had the lowest
number of aquatic plant species when compared to sites with lower Pb concentrations. In
an aquatic macrophyte community, both plant species and type of habitat were
determined to be factors affecting the rate of Pb accumulation from contaminated
sediments. While the rooted macrophyte 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 most susceptible to
atmospherically-deposited Pb, resulting in shifts in plant community composition as a
result of Pb pollution.
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.
Sensitive species may become locally extinct from habitats where Pb toxicity is greater.
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. A series of freshwater
species sensitivity distributions constructed by Brix et al. (2005) indicated that sensitivity
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to Pb was greatest in crustacean species, followed by coldwater fish, and warmwater fish
and aquatic insects, which exhibited a similar sensitivity.
7.3.17.6 Critical Loads, Sensitivity and Vulnerability
In the short term, metal emissions generally have greater effects on biota in aquatic
systems than in terrestrial systems because metals are more readily immobilized in soils
than in sediment. However, over the longer term, terrestrial systems may be more
affected particularly by those metals with a long soil residence time, such as Pb. Thus, for
a particular locale, either the terrestrial or the aquatic ecosystem at that site may have the
lower critical load. Given the heterogeneity of ecosystems affected by 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.
Recent studies have identified seasonally-affected physiological changes and lifestage as
important determinants of differential sensitivity to Pb in aquatic organisms. These
factors are in addition to species differences in metabolism, sequestration, and
elimination rates, diet, lifestage, genetics, age, and body size that were considered in the
2006 Pb AQCD. Although evidence is available to support Pb impacts to supporting,
provisioning, regulating and cultural ecosystem services, there is insufficient data
available to adequately quantify these effects.
7.4 Causal Determinations for Ecological Effects of Lead
This section presents key conclusions regarding causality for welfare effects of Pb (Table
7-2). Evidence considered in establishing causality was drawn from recent studies
summarized in this ISA, integrated with findings presented in the 1977 (U.S. EPA. 1977).
1986 (U.S. EPA. 1986a) and 2006 Pb AQCDs (U.S. EPA. 2006c) for Pb. The causal
statements for terrestrial and aquatic effects are arranged according to ecologically
meaningful levels of biological organization (organism, population, community,
ecosystem). As recognized in EPA's Framework for Ecological Risk Assessment (U.S.
EPA. 1992). and in the adverse outcome pathway (AOP) framework (Anklev et al.. 2010)
endpoints that are measured at one level of 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 survival and to developmental and reproductive endpoints that can be used
in ecological risk assessment, i.e., at the population level and higher. Fecundity, growth,
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and survival are organism-level attributes that lead to population-level (e.g., abundance,
production, extirpation), community-level (taxa richness, relative abundance) and
ecosystem-level responses (Anklev et al.. 2010; Suteret al.. 2005). In the case of Pb,
physiological stress, hematological effects and neurobehavioral alterations are likely to
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 chronic toxicity of Pb in
ecosystems is thus likely attained through multiple modes of action.
Relevant sources of information on 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 also include potentially
confounding factors (e.g., other metals, environmental conditions), thus increasing the
uncertainty in associating effects with exposure to Pb specifically. The vast majority of
available Pb studies 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. Potential effects of Pb at the population, community and ecosystem level are
often, of necessity, inferred from organism-level studies (U.S. EPA. 2006c). Considering
the various types of ecological evidence available, effects determined to be causal at the
organism or species level of biological organization contribute to the body of evidence
for causal effects at the community and ecosystem levels of biological organization. For
some ecological endpoints, support for causality 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.
For aquatic biota, generally, the number of studies available on freshwater organisms is
greater than on saltwater organisms, covering more taxa as well as more endpoints. The
causal determinations for aquatic biota are, therefore, based primarily on evidence from
freshwater species. When available, data from saltwater ecosystems are discussed
separately under the appropriate causal endpoint. For most of the endpoints under
consideration, evidence is inadequate to establish causality in saltwater species. When
sufficient evidence is available for marine organisms, data on concentrations at which
effects are observed are presented.
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Table 7-2 Summary of Pb causal determinations for plants, invertebrates and
vertebrates
Effect
Terrestrial
Aquatic3
Physiological Stress-All organisms
Causal
Causal
Hematological Effects-Invertebrates
Inadequate
Causal
Hematological Effects-Vertebrates
Causal
Causal
Neurobehavioral Effects-Invertebrates and Vertebrates
Likely Causal
Likely Causal
Developmental and Reproductive Effects-Plants
Inadequate
Inadequate
Developmental and Reproductive Effects-Invertebrates and Vertebrates
Causal
Causal
Growth-Plants
Causal
Causal
Growth-Invertebrates
Inadequate
Causal
Growth-Vertebrates
Inadequate
Inadequate
Survival-Plants
Inadequate
Inadequate
Survival- Invertebrates and Vertebrates
Causal
Causal
Community and Ecosystem Level Effects
Likely Causal
Likely Causal
'Causal determinations for aquatic biota are based primarily on evidence from freshwater organisms.
Characterization of environmental concentrations of Pb in terrestrial and aquatic systems
is important for identification of levels of exposure to organisms. Information on ambient
Pb concentrations in non-air media and biota is reported in Section 3.6 and Table 2-1. A
survey of 1,319 locations in the conterminous U.S. found an average background soil
concentration of 19.3 mg Pb/kg , with a 95th percentile of 50 mg Pb/kg, and 16 locations
between 100 and 700 mg Pb/kg (Shacklette and Boerngen. 1984). The U.S.EPA's Eco-
SSL Guidance documents (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 national average of state-level soil averages was 18.9 mg Pb/kg. The range of
the state mean soil Pb concentrations was from 5 mg Pb/kg to 38.6 mg Pb/kg.
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 (.ig Pb/g dry weight, range 0.5
to 12,000 (.ig Pb/g 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). The range of Pb levels in
saltwater 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 salt water was reported by Leland and Kuwabara
(1985) and 0.01 to 27 (.ig Pb/L by 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).
The level at which Pb elicits a specific effect is difficult to establish in terrestrial and
aquatic systems due the influence of other environmental variables on Pb bioavailability
and toxicity and substantial species differences in Pb sensitivity. In the 1977 Pb AQCD,
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no correlation could be established between toxic effects in invertebrates, fish, birds and
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 levels of Pb that can cause toxic effects in small mammals or birds (U.S. EPA.
1986b). In the 2006 Pb AQCD several studies on effects of Pb exposure on natural
ecosystem structure and function advanced the characterization of Pb levels in the
environment near point sources (i.e., smelters, mining). According to the 2006 Pb AQCD,
natural terrestrial ecosystems near significant Pb point 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 are 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 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 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.
The key conclusions and causal determinations for welfare effects of Pb are presented in
ascending order of the level of ecological organization to which they apply, moving from
responses in organisms (physiological stress, hematological effects, neurobehavioral
effects) to endpoints relevant to population, community, and ecosystem-level processes
(development and reproduction, growth, survival). They 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 conditions were also considered in
the body of evidence. Studies at higher concentrations were used to the extent they
informed our understanding of modes of action and illustrated the wide range of
sensitivity to Pb across taxa.
7.4.1 Causal Determinations for Lead in Terrestrial Systems
7.4.1.1 Physiological Stress-Terrestrial Biota
Induction of enzymes associated with oxidative stress response in terrestrial biota 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,
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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 (Vicia faba) (Wang et al.. 2010c; Wang et al..
2010b; 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 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.. 2011a). 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.
The oxidative stress responses associated with Pb exposure in terrestrial biota are
consistent with responses in aquatic organisms (Section 7.4.2.1), in experimental animal
studies, and in humans (Section 5.2.4). 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,
invertebrates and vertebrates.
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7.4.1.2
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 (Beveret 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 aquatic biota (Section 7.4.2.2) 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 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 exposure to Pb and
hematological effects in humans and in laboratory animals (Section 5.7). 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.
7.4.1.3 Neurobehavioral Effects-Terrestrial Biota
The central nervous system of animals was recognized as a target of Pb toxicity in the
1977 Pb AQCD and subsequent Pb reviews have provided additional supporting evidence
of Pb as a neurotoxicant in terrestrial biota (U.S. EPA. 1977). Effects of Pb on
neurological endpoints in terrestrial animal taxa include changes in behaviors that may
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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, growth, and shell thickness were 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-Gendv et
al.. 2011). while other snails exposed to Pb at similar concentrations have shown no
effects on feeding rate (Beebv and Richmond. 2010). Consumption of 10 mg/Pb kg diet
resulted in lower body weight and food intake after 120 days of dietary exposure in pigs
(S. domestica) (Yu et al.. 2005).
In limited studies available on worms there is evidence that Pb may affect the ability to
escape or avoid predation. For example, the motility of nematodes was disrupted in Pb-
contaminated soils (Wang and Xing. 2008). Additionally, there is new evidence for
neurotoxic action of Pb in invertebrates with exposure to Pb observed to cause changes in
the morphology of GABA motor neurons in nematodes (C. elegans) (Du and Wang.
2009).
There is limited evidence for additional behavioral endpoints altered by Pb exposure. 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 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 20 mg Pb/kg per day
for 60 days in the laboratory exhibited abnormal coloration and posturing behaviors.
These findings in terrestrial biota are coherent with findings from studies in aquatic biota
that showed neurobehavioral alterations in various species of fish, and also in some
aquatic invertebrates (Section 7.4.2.3). They are also coherent with findings in laboratory
animals (Section 5.3) that show that Pb induces changes in learning, memory, attention
and motor skills. New behaviors induced by exposure to Pb are detailed in that section,
including 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. Human epidemiological evidence, based on several cohort and cross-
sectional studies in diverse populations provides clear and consistent evidence of
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association between blood Pb concentrations and decrements in neurodevelopmental
outcomes in young children (Section 5.3). In addition to the consistency of findings in
children, the evidence is strengthened by the coherence of findings with toxicological
studies and by coherence of association of blood Pb with a spectrum of related endpoints
including IQ, verbal and reading skills, motor coordination, mood and attention
problems, and behavioral problems. The evidence presented in the health chapter is
sufficient to conclude that there is a causal relationship between Pb exposure and
neurobehavioral effects (Section 5.3). Data from ecological studies are highly coherent
with these human health-related data, especially neurobehavioral findings and evidence
of structural changes. Overall, the evidence from aquatic and terrestrial systems is
sufficient to conclude that there is a likely causal relationship between Pb exposures and
neurobehavioral effects in invertebrates and vertebrates in terrestrial ecosystems.
7.4.1.4 Developmental and Reproductive Effects-Terrestrial Biota
In terrestrial biota, evidence from Chapter 7 of this ISA and in previous Pb AQCDs
indicates an association between reproductive effects and Pb exposure. 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). Evaluation of the literature on
Pb effects in terrestrial species indicates that exposure to Pb is associated with
reproductive effects. Various endpoints have been measured in various taxa of terrestrial
organisms to assess the effect of Pb on development, fecundity and hormone homeostasis
and although reproductive effects were demonstrated, no single endpoint in a single taxon
has been extensively studied.
In terrestrial plants, few studies are available 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 10, 100 or 1,000 (.ig Pb/L as Pb sulfate (Ahmad et al.).
In terrestrial invertebrates, Pb can alter developmental timing, hatching success, sperm
morphology and hormone homeostasis. However, effects are observed in a limited
number of species and at concentrations that typically exceed Pb levels in U.S. soils. In
fruit flies, Pb exposure increased time to pupation and decreased pre-adult development
at high Pb concentrations (Hirsch et al. 2010). An increase in development time
(approximately two days) and a reduction in relative fecundity was observed in aphids
feeding on plants exposed to high concentrations of Pb (Gorur. 2007). Hatching success
of the collembolan F. Candida was decreased following 10 day exposure to Pb-spiked
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soils (EC50 2,361 mg Pb/kg dry soils) (Xu et al.. 2009a). Sperm morphology in Asian
earthworms was significantly altered following 2-week exposures to soils containing
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. No significant changes to the number of juveniles
were noted in the collembolan Proisotoma minuta exposed for 42 days to 300, 750, 1,500
or 3,000 mg Pb/kg in acidic sandy loam soil (Nursita et al.. 2005) in agreement with a
previous study with collembolans in which aNOEC of 2,000 mg Pb/kg and LOEC of
5,000 mg Pb/kg was determined for reproduction (Sandifer and Hopkin. 1996).
In terrestrial vertebrates, there is evidence for reproductive effects associated with Pb
exposure in this ISA and previous Pb AQCDs. Reproductive effects observed in birds
near areas of Pb-contamination or where Pb tissue concentration has been correlated with
effects include declines in clutch size, number of young hatched, number of young
fledged, decreased fertility, and decreased eggshell thickness (U.S. EPA. 2006c).
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. 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.. 2009b).
Evidence from aquatic invertebrates, amphibians, and fish is sufficient to conclude that
there is a causal relationship between Pb exposure and reproductive effects in aquatic
invertebrates and vertebrates (Section 7.4.2.4). In Chapter 5, evidence obtained using
mammals in the context of human health research demonstrates a consistency of adverse
effects of Pb on sperm (Section 5.8.2.1) and the onset of puberty in males and females
(Sections 5.8.1.3 and 5.8.2.4) 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 5.8).
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The evidence presented in Section 7.4.1.4 is thus highly coherent with evidence from
aquatic ecosystems and laboratory animal studies, and sufficient to conclude that there is
a causal relationship between Pb exposures and developmental and reproductive effects
in terrestrial invertebrates and vertebrates.
In terrestrial plants, the evidence is inadequate to conclude a causal relationship between
Pb exposures and plant developmental and reproductive effects.
7.4.1.5 Growth Effects-Terrestrial Biota
Alterations in growth which is an organism-level variable 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 point sources. Growth effects observed in both invertebrates and vertebrates,
where juvenile organisms are generally more sensitive than adults, underscore the
importance of lifestage to overall Pb susceptibility.
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). Specifically, Pb has been shown to affect photosystem II with
the hypothesized mechanism being that Pb may replace either Mg or Ca in chlorophyll,
altering the pigment structure and decreasing the efficiency of visible light absorption in
exposed plants. Decreases in chlorophyll a and b 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 heavily
on the Ecological Soil Screening Levels for Lead (U.S. EPA. 2005b) document regarding
ecotoxicological effects of Pb for which growth (biomass) was considered 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 F1 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. Reduced growth of garden snail T.
pisana, was observed following a five week dietary exposure of 1,000 (.ig Pb/g and
greater (El-Gendv et al. 2011). In previous AQCDs, growth effects of Pb have been
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reported in birds (changes in juvenile weight gain), at concentrations typically higher
than currently found in the environment.
Evidence is sufficient to conclude that there is a causal relationship between Pb
exposures and growth effects in terrestrial plants. Evidence is inadequate to establish
causal relationship between Pb exposures and growth effects in terrestrial invertebrates
and vertebrates.
7.4.1.6 Survival-Terrestrial Biota
The relationship between Pb exposure and survival has been well demonstrated in
terrestrial species as presented in Sections 7.2.5 of this ISA and in the previous Pb
AQCDs. Pb exposure can either be 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).
Additional studies in the 1986 and 2006 Pb AQCDs and current ISA provide evidence for
an exposure-response relationship between Pb exposure and mortality in terrestrial biota.
Survival is a biologically important response that may have direct impact on population
size. However, the typically reported concentration where there is 50% mortality of test
organisms (LC50) is a poor measure for comparing environmental concentrations. LC50 is
applicable to acute toxicity whereas Pb effects on ecosystem receptors is likely
characterized as a chronic, cumulative exposure rather than an acute exposure. The
scenario where half of test animals die in a laboratory situation would be unacceptable
from an ecological perspective. 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 well below concentrations that result in mortality. When available,
LCio,NOEC or LOEC is reported.
Pb is generally not phytotoxic to plants at concentrations found in the environment away
from point sources, 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.
Survival of soil-associated organisms is adversely affected by Pb exposure. 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. Alterations to
bacterial and fungal communities can lead to inhibited or delayed decomposition of
organic matter and disruption of energy flow through the food chain. As reviewed in the
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2006 Pb AQCD, the LC50 values for soil nematodes vary from 10-1,550 mg Pb/kg dry
weight dependent upon soil OM content and soil pH ("U.S. EPA. 20060). 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.
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-day) as reviewed for the
development of Eco-SSL's (U.S. EPA. 2005b'). The NOAELs for survival ranged from
3.5 to 3,200 mg Pb/kg-day. The association of Pb exposure with mortality in vertebrates
is coherent with observations in aquatic biota and with the observation of consistently
positive associations between Pb exposure and mortality observed in human
epidemiologic studies (Section 5.4.5).
The evidence is inadequate to conclude that there is a causal relationship between Pb
exposures and survival in terrestrial plants.
The evidence is sufficient to conclude that there is a causal relationship between Pb
exposures and survival in terrestrial invertebrates and vertebrates.
7.4.1.7 Community and Ecosystem Level Effects-Terrestrial Biota
Most direct evidence of community and ecosystem level effects is from near point
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 S02 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 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,
physiological stress, blood, neurobehavioral and developmental and reproductive
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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 (.ig Pb/g 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 (U.S. EPA. 1986b) and altering community structure. At soil concentrations of
500 to 1,000 (.ig Pb/g (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-point
sources.
New evidence since the 2006 Pb AQCD 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, which may provide a better estimate of ecological
impacts than 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 significantly decreased phenol oxidase activity in all soils
tested, while 400 nM 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 etal. 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 etal. 2011V 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
(Boiarczuk and Kieliszewska-Rokicka. 2010; Sudova and Vosatka. 2007). Invertebrates
affected by Pb in terrestrial systems may be altering community structure. New 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 available earthworm biomass.
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Furthermore, the presence of earthworms increased Pb uptake by plants (Ruiz et al..
2011; Sizmur et al.. 201IV
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 biota is from single-species
assays in laboratory studies. Although the evidence is strong for effects of Pb on growth,
reproduction and survival 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 there is a likely causal relationship
between Pb exposures and the alteration of species richness, species composition and
biodiversity in terrestrial ecosystems.
7.4.2 Causal Determinations for Lead in Aquatic Systems
7.4.2.1 Physiological Stress-Aquatic 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 aquatic
plants, invertebrates and vertebrates. A few studies provide evidence of effects at current
concentrations of Pb measured in the environment. In plants, increases of antioxidant
enzymes with Pb exposure occur in algae, aquatic mosses, and floating and rooted
aquatic macrophytes. Most available evidence in aquatic plants for antioxidant responses
to Pb are from laboratory studies conducted 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. The macrophyte Myriophyllum quitense exhibited
elevated antioxidant enzyme activity (glutathione-S-transferase, glutathione reductase,
peroxidase) following transplantation in anthropogenically polluted areas containing
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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 considerable evidence for antioxidant activity in response to Pb exposure in
aquatic invertebrates, including gastropods, mussels, and crustaceans. Some of the recent
evidence for invertebrate antioxidant response indicates effects at Pb levels detected in
the 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.. 2010c'). Markers of oxidative damage
are also observed in fish, amphibians and mammals in laboratory studies. Across all
biota, 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. Heat shock
protein induction has been observed in zebra mussels exposed to 500 (.ig Pb/L for 10
weeks (Singer et al.. 2005). 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). 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 near environmentally
relevant concentrations of Pb (50 (ig/L and higher) (Ansaldo et al.. 2006).
Evidence for stress responses observed in aquatic biota is supported by findings in
terrestrial species (Section 7.4.1.1) as well as in humans and experimental animal studies
of oxidative stress following impairment of normal metal ion functions (Section 5.2.4).
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 there is a causal
relationship between Pb exposures and physiological stress in aquatic plants,
invertebrates and vertebrates.
7.4.2.2 Hematological Effects-Aquatic 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
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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 to fish included a decrease in red blood cells and inhibition of ALAD (U.S.
EPA. 1986b). 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 bivalves, and lower ALAD activity has been
correlated with elevated blood Pb levels in 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 fish, confirms that the
decreased activity in this enzyme is an indicator for Pb exposure across a wide range of
taxa. 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 organisms (Section 7.4.1.2). It is also
coherent with observations from human epidemiologic and animal toxicology studies
(Section 5.7) 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. Taken together, the overall weight
of epidemiologic and toxicological evidence is sufficient to conclude that a causal
relationship exists between exposure to Pb and hematological effects in humans
(Section 5.7). Based on observations in both terrestrial and aquatic organisms and
additionally supported by findings in terrestrial systems 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
invertebrates and vertebrates in aquatic ecosystems.
7.4.2.3 Neurobehavioral Effects-Aquatic Biota
Evidence from laboratory studies and limited data from field studies reviewed in this
chapter 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.
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 Hinc.
1970). Some organisms exhibit behavioral avoidance while others do not seem to detect
the presence of Pb (U.S. EPA. 2006c). Evidence of alteration in behaviors at the level of
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the organism is a potential endpoint at the ecological level when its effects on population
or community are taken into consideration (U.S. EPA. 2003a').
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 in juvenile fathead
minnows following 2-week pre-exposure to 500 (.ig Pb/L (Weber. 1996V 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).
New information since the 2006 Pb AQCD provides further evidence for Pb impacts on
behaviors that may affect feeding and predator avoidance in freshwater environments.
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). Zebrafish embryos exposed to low concentrations of Pb
(lOnM or 30 nM) until 24 hours post-fertilization and then subsequently tested as larvae
or adult fish exhibited behavioral deficits in response to mechanosensory and visual
stimuli (Rice etal.. 2011). 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 non-exposed embryos. In adult fish
hatched from Pb-exposed embryos (30 nM), visual response to a rotating black bar
against a white background (ability to detect contrast) was significantly degraded.
In marine organisms there is evidence for neurobehavioral effects of Pb in a few species.
In a study reviewed in the 2006 Pb AQCD, prey capture rate and predator avoidance was
affected in mummichogs starting at 300 (.ig Pb/L (Weis and Weis. 1998). New studies on
neurobehavioral endpoints in marine species 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 j^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 Cace res-
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 la). These fish were fed white shrimp exposed to
Pb via brine shrimp that were initially fed microalgae cultured 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. The ornate wrasse, T. pavo, was exposed
to sublethal (400 (.ig Pb/L) or a maximum acceptable toxicant concentration (1,600 (.ig
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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.
Findings in laboratory animals and humans support the limited evidence for
neurobehavioral effects of Pb in aquatic biota. In studies reviewed in Section 5.3, effects
observed in laboratory animals following exposure to Pb include hyperactivity and mood
disorders, visual and auditory sensory alterations, and changes in structure and function
of neurons and supporting cells in the brain, including 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 Section 5.3. 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). Human epidemiologic evidence, based on several cohort and cross-sectional
studies in diverse populations provides clear and consistent evidence of association
between blood Pb concentrations and decrements in neurodevelopmental outcomes in
young children (Section 5.3). In addition to the consistency of findings in children, the
evidence is strengthened by the coherence of findings with toxicological studies and by
coherence of association of blood Pb with a spectrum of behaviorally-related endpoints.
The evidence presented in sections of this ISA dealing with human health is sufficient to
conclude that there is a causal relationship between Pb exposure and neurobehavioral
effects (Section 5.3). Evidence in terrestrial ecosystems is not as extensive, but it is
highly coherent with findings in aquatic ecosystems. Data related to human health are
also highly coherent with aquatic ecosystem data. Overall, the evidence from available
studies on neurobehavioral effects of Pb in aquatic systems is limited, but sufficient to
conclude that there is a likely causal relationship between Pb exposures and
neurobehavioral effects in aquatic invertebrates and vertebrates.
7.4.2.4 Developmental and Reproductive Effects-Aquatic Biota
Evaluation of the findings in previous Pb AQCDs and new 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 in some 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
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aquatic organisms to assess the effect of Pb on development, fecundity and hormone
homeostasis. However, there are typically only limited studies available from different
taxa.
There are no studies reviewed in the ISA or previous Pb AQCDs on development and
reproductive effects of Pb in freshwater or saltwater aquatic algae or macrophytes.
Reproductive effects of Pb in freshwater aquatic invertebrates are well-characterized in
previous Pb AQCDs and in the current ISA and have been observed at or near current
ambient concentrations in some species. 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 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 jj.g Pb/L while in amphipods
receiving both water-borne and dietary Pb the LOEC for reproduction was 3.5 |_ig Pb/L.
Reproductive output in chronic testing with the freshwater rotifer B. calyciflorus 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 LOEC of 194 (.ig Pb/L (Grosell et al.. 2006a). In a 7 day
exposure of the cladoceran C. dubia to 50 to 500 |_ig Pb/L , increased DOC lead to an
increase in mean EC50 for reproduction ranging from approximately 25 |_ig Pb/L to
>500 (.ig Pb/L (Manor et al.. 201 la).
Reproductive effects of Pb have also been reported in marine invertebrate species. In the
2006 Pb AQCD 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 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). 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
1 hour LC50 value for Pb toxicity to eggs in the same species was 692 |_ig Pb/L while 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 690 |_ig Pb/L,
respectively. In a multigenerational study 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 (ig Pb/g sediment and higher
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(Ringenarv et al. 2007). The authors indicate that this concentration is below the current
sediment regulatory guideline for Pb (218 (.ig Pb/g sediment) (NO A A. 1999) and that
reproductive effects are a more sensitive endpoint than lethality. The LOEC for
embryogenesis in the marine bivalve M. galloprovincialis was 50 fj.g Pb/L with an EC50
for embryogenesis of 221 (.ig Pb/L (Beiras and Albentosa. 2003).
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). Previous Pb
AQCDs have reported developmental effects in fish, specifically spinal deformities in
brook trout (Salvelinus fontinalis) exposed to 119 (ig Pb/L for three generations (U.S.
EPA. 1977). and in rainbow trout as low as 120 j.ig 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 (.ig Pb/L (Weber. 1993). 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 fish, there is
new evidence of Pb in this ISA on alteration of steroid profiles and additional
reproductive parameters. Reproduction in fathead minnows was affected in breeding
exposures following 300 day chronic toxicity testing. However, reproductive
performance was unaffected in zebrafish exposed to Pb-via consumption of contaminated
prey. 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.
Observations of Pb toxicity to reproductive and developmental endpoints in aquatic fauna
are further supported by evidence terrestrial invertebrates and vertebrates
(Section 7.4.1.4) and from laboratory animals (Section 5.8). Pb appears to affect multiple
endpoints associated with reproduction and development in aquatic biota. A few sensitive
species have been identified where effects are observed in laboratory studies at
concentrations of Pb currently measured 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 invertebrates and vertebrates.
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In aquatic plants, the evidence is inadequate to conclude a causal relationship between Pb
exposures and plant developmental and reproductive effects.
7.4.2.5 Growth Effects-Aquatic 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.
Evidence for Pb effects on growth in invertebrates has been observed most extensively in
freshwater aquatic taxa, with inhibition in sensitive species occurring at or near the
current range of Pb in surface waters. Growth effects observed in both invertebrates and
vertebrates underscore the importance of lifestage to overall Pb sensitivity. In general,
juvenile organisms are more sensitive than adults.
In the 2006 Pb AQCD, 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 (Gaur et al.. 1994V
Growth effects of Pb on aquatic invertebrates are reviewed in the draft Ambient Aquatic
Life Water Quality Criteria for Pb (U.S. EPA. 2008b) and the 2006 Pb AQCD. In the
2006 Pb AQCD, the LOEC for growth of amphipods H. azteca in 42-day chronic
exposure to Pb was 16 jj.g Pb/L (Besser et al.. 2005). Recent studies provide additional
evidence for effects on growth of freshwater aquatic invertebrates at < 10 (ig Pb/L.
Growth of juvenile freshwater snails L. stagnalis was inhibited below the lowest
concentration tested resulting in an EC2o < 4 j.ig Pb/L (Grosell and Brix. 2009; Grosell et
al.. 2006a'). 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 (/.. siliquoidea) juveniles
were the most sensitive lifestage (Wang et al.. 2010e). 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. The EC2o for reduced growth and emergence of the midge C.
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) The ECio for molting in the mayfly B.
tricaudatus was 37 jj.g Pb/L (Mebane et al. 2008). These effects are occurring at or near
the current range of Pb concentrations in surface waters (median 0.50 |_ig Pb/L, range
0.04 to 30 (.ig Pb/L) (U.S. EPA. 2006c).
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Evidence for growth effects of Pb in 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 new study
reviewed in this ISA 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 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 and are limited to
freshwater species (Manor. 2012). In a review cited in the 2006 Pb AQCD general
symptoms of Pb toxicity in fish include growth inhibition (Eisler. 2000) however, other
studies with Pb show no growth effect (Manor. 2012). In the studies reviewed for the
current ISA no growth effects were observed in fish exposed to Pb via dietary intake.
Recent 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 (Manor and
Grosoll. 2011; Grosoll et al.. 2006b). Other studies with fathead minnows have shown
growth reductions with Pb exposure, however, concentrations of observed effects
typically exceeded the 96-hour LC50 value (Manor. 2012; Manor et al. 2010; Grosoll et
al.. 2006b). 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 jj.g Pb/L with
an EC2o of >87 (ig Pb/L .
Growth effects associated with exposure to Pb have also been reported in marine species.
New studies in marine algae provide evidence for effects of Pb on growth within one to
two orders of magnitude of measured levels of Pb in the environment. The lowest 72-
hour EC50 for growth inhibition reported for marine algae was 105 |_ig Pb/L in
Chaetoceros sp (Dcbolius et al. 2009). Growth of microalgae T. suecica exposed to
environmentally relevant concentrations of Pb (20 |_ig Pb/L) was 40% lower than control
cultures (Soto-Jimonoz et al.. 201 la). 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. Wang et al., (2009d) 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-Finuoroa et al.. 2007). Rate of growth of the deposit
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feeding polychaete Capitella sp. decreased significantly with increasing concentrations of
Pb associated with sediment from polluted estuaries (Horng et al.. 2009V
Evidence of effects of Pb exposure on growth in terrestrial plants is highly coherent with
evidence from aquatic plants. Evidence is sufficient to conclude that there is a causal
relationship between Pb exposures and growth effects in aquatic plants and aquatic
invertebrates. Evidence is inadequate for a causal relationship between Pb exposures and
growth effects in aquatic vertebrates.
7.4.2.6 Survival-Aquatic Biota
The relationship between Pb exposure and survival has been well demonstrated in aquatic
species as presented in Section 7.3.5 of this ISA 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 may have a direct
impact on population size. However, the concentration typically reported at which there
is 50% mortality of test organisms (LC50) is a poor measure for comparing environmental
concentrations. LC50 is a relevant measure for acute toxicity whereas Pb effects on
ecosystem receptors are likely characterized as a chronic, cumulative exposure rather
than acute exposure. 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, LCi0,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 or
saltwater environments.
There is considerable data available on toxicity of Pb to aquatic invertebrates as reviewed
in the previous Pb AQCDs and Ambient Water Quality Criteria for Lead (U.S. EPA.
1985). 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'). Freshwater
invertebrates are generally more sensitive to Pb exposure than other taxa, with survival
impacted in a few species at concentrations occurring near point sources or at
concentrations that approach ambient levels. These 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
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movement (a measure of viability) for newly transformed juveniles of this species was
188 (.ig Pb/L. (Wang et al.. 20106).
Freshwater biota that exhibit sensitivity to Pb in the 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. 2006c).
include some species of gastropods, amphipods, cladocerans and rotifers although the
toxicity of Pb is highly dependent upon water quality variables such as DOC, hardness
and pH. In the 1986 Pb AQCD, increased mortality was observed in the freshwater
gastropod Lymnaea palutris 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 this ISA). A 7 day LC50 of 1 (.ig Pb/L was observed
in soft water with the amphipod H. azteca (Borgmann et al.. 2005). In this same species,
the 96 hour LC50 for Pb at pH of 5 was 10 jj.g Pb/L (Mackie. 1989). In 42-day chronic
exposures of H. azteca exposed to Pb via water and diet, the LC50 was 16 (.ig Pb/L
(Besser et al. 2005). At higher pH and water hardness, amphipods are less sensitive to Pb
(U.S. EP A. 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 180 (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. EP A. 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). Other freshwater invertebrates
such as odonates may be tolerant of Pb concentrations that greatly exceed environmental
levels. Some invertebrates are able to detoxify Pb such as through sequestration of Pb in
the exoskeleton which is subsequently molted.
In aquatic vertebrates, 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 of fish from streams (U.S. EP A. 1977). The lowest LC50
for fish reported in the 1977 Pb AQCD was 1,000 (ig Pb/L in soft water for rainbow trout
(Salmo gairdneri) following 96-hour exposure to Pb (U.S. EPA. 1977). More recently
reviewed studies using fish have considered the role of water quality variables on Pb
toxicity. 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 fathead minnow conducted
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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 (Esbaimh et al..
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 39 to 1,903 (.ig Pb/L in varying concentrations of DOC, CaS04 and pH
(Grose 11 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 CaC03, a value much
lower than in testing with older fish (Mebane et al.. 2008).
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 the marine amphipod Melita
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 |ag Pb/L for juveniles in comparison to adults (3,000 |_ig Pb/L). Ten-day exposures
of juveniles in seawater resulted in an LC50 of 1,270 |_ig Pb/L, an NOEC of 190 |_ig Pb/L
and a LOEC of 390 |_ig Pb/L compared to the adult values of 3,560 |_ig Pb/L. 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 to 20 |_ig Pb/L and
subsequently fed to brine shrimp A. franciscana, (mean Pb content 12 to 15 |_ig Pb/g)
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.. 201 la).
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
marine fish is limited. Acute toxicity of Pb to plaice (Pleuronectes platessa) ranged from
50 |ag Pb/L to 300,000 |_ig Pb/L depending on the form of Pb (Eisler. 2000). Evidence of
effects of Pb on survival in aquatic animals is coherent with evidence in terrestrial ones.
The evidence is inadequate to conclude that there is a causal relationship between Pb
exposures and survival in aquatic plants.
The evidence is sufficient to conclude that there is a causal relationship between Pb
exposures and survival in aquatic invertebrates and vertebrates.
7.4.2.7 Community and Ecosystem Level Effects-Aquatic Biota
Most direct evidence of community- and ecosystem-level effects in aquatic systems is
from near point sources where Pb concentrations are higher than typically observed
environmental concentrations for this metal. Impacts of Pb on aquatic habitats that
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receive runoff from point sources 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,
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 aquatic ecosystems, numerous field studies reviewed in the 2006 Pb AQCD
(summarized in Table AX7-2.5.2) and this ISA 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
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 Ephemeroptera, Plecoptera and
Trichoptera (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 j.ig Pb/L (Femandez-Lcborans and Antonio-Garcia.
1988).
Since the 2006 Pb AQCD, there is further evidence for effects of Pb in sediment-
associated communities and aquatic cyanobacteria communities. Exposure to three levels
of sediment Pb contamination (322, 1,225, and 1,465 (.ig Pb/g dry weight) in a microcosm
experiment significantly reduced nematode diversity and resulted in profound
restructuring of the community structure (Mahmoudi et al.. 2007). Sediment-bound Pb
contamination appears to differentially affect members of the benthic invertebrate
community, potentially altering ecosystems 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
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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. Bacterial biomass and diversity of aquatic
cyanobacteria communities was reduced in a microcosm following one week exposure to
25 mM Pb (Bumat et al.. 2009). Pb exposure impacted individual cyanobacteria species
differently, with Microcoleus sp. experiencing a greater decrease in abundance than
Halomicronema-like cyanobacteria.
In a new 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) (Mishra et 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 biota in this heavily
contaminated ecosystem. Mean Pb concentrations in 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
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endpoints in waterfowl from Lake Coeur d'Alene that ingest Pb contaminated sediments
and vegetation during feeding.
Fewer studies are available for community and ecosystem effects in saltwater systems.
No studies on marine systems were reviewed in the 1977 Pb AQCD or the 1986 Pb
AQCD. 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 (300 to 1,000 (.ig Pb/L). 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. In a
recent study, significant differences in macroinvertebrate communities associated with
seagrass beds were observed 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) (Gvedu-Ababio and Baird. 2006).
Most evidence for Pb toxicity to biota is from single-species assays in laboratory studies.
Although the evidence is strong for effects of Pb on growth, reproduction and survival 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 presence of multiple stressors, variability in field
conditions and to differences in Pb bioavailability at that level of organization.
Nevertheless, evidence of ecosystem effects in aquatic ecosystems is coherent with
similar evidence in terrestrial ecosystems, and based on the cumulative evidence from
laboratory studies and field data, there is a likely causal relationship between Pb
exposures and the alteration of species richness, species composition and biodiversity in
aquatic ecosystems.
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